Biosorbents for Metal Ions

Biosorbents for Metal Ions To Romana For reading every word in this book Biosorbents for Metal Ions Edited by DR JO...

Author: D A John Wase | John Wase

292 downloads 1487 Views 3MB Size Report

This content was uploaded by our users and we assume good faith they have the permission to share this book. If you own the copyright to this book and it is wrongfully on our website, we offer a simple DMCA procedure to remove your content from our site. Start by pressing the button below!
Report copyright / DMCA form

DOWNLOAD PDF

Biosorbents for Metal Ions

To Romana For reading every word in this book

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 system, or transmitted, in any form or by any means, electronic, electrostatic, magnetic tape, mechanical, photocopying, recording or otherwise, without the prior permission of the 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

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

List of Contributors

D.ADERHOLD Dortmund University Dortmund Germany S.J.ALLEN Department of Chemical Engineering The Queen’s University of Belfast Belfast UK C.J.BANKS Department of Civil Engineering University of Southampton Southampton UK R.G.J.EDYVEAN Department of Chemical and Process Engineering The University of Sheffield Sheffield UK C.F.FORSTER School of Civil Engineering The University of Birmingham Birmingham UK

G.W.GARNHAM BNFL Springfield Works Salwick Preston UK Y.S.HO School of Chemical Engineering The University of Birmingham Birmingham UK A.KAPOOR Faculty of Engineering University of Regina Regina Saskatchewan Canada G.MCKAY Department of Chemical Engineering Hong Kong University of Science and Technology Clear Water Bay Kowloon Hong Kong

ix

List of Contributors

M.TSEZOS Environmental Engineering Unit Materials Science and Engineering National Technical University Athens Greece M.M.URRUTIA Department of Biological Sciences The University of Alabama Tuscaloosa Alabama USA T.VIRARAGHAVA Faculty of Engineering University of Regina Regina Saskatchewan Canada

x

D.A.J.WASE School of Chemical Engineering The University of Birmingham Birmingham UK C.J.WILLIAMS Department of Chemical and Process Engineering The University of Sheffield Sheffield UK M.M.WILSON Smith Kline Beecham Worthing UK

1

Biosorption of Heavy Metals: An Introduction C.F.FORSTER and D.A.J.WASE

Introduction Two of the ages which define elements of man’s existence on earth have ‘metal names’. It might, therefore, be argued that heavy metal pollution began in these periods. However, a more realistic scenario is that man’s use of metals seriously began to affect the environment during the Industrial Revolution. Today, 200 years later, we could be said to be in the Metal Removal Age and we are all too aware of the risks inherent in the uncontrolled dissemination of heavy metals into the environment. From the environmental point of view, the metals that are of greatest concern are those which, either by their presence or their accumulation, can have a toxic or an inhibitory effect on living things. Metals can be dispersed, both naturally and by man’s activities, into any of the earth’s elements: soil, water or air. However, the aquatic environment will be the main focus of this book. The metals which are of greatest environmental concern are listed in Table 1.1 and it is appropriate, at this stage, to make some comment on their effect on the environment. Some of these effects are summarised in Table 1.2, although it must be remembered that environmental conditions—temperature, pH, water hardness—will affect the toxicities.

Table 1.1 Heavy metals which have an environmental effect

1

Biosorbents for Metal lons Table 1.2 Typical toxicities of heavy metals in fresh water

‘LC50’ means that the given concentration will be lethal to 50% of organisms in the given time.

Toxic metals

Cadmium Cadmium is a highly toxic element which, in humans, can cause serious damage to the kidneys and bones, and is probably best known for its association with itai-itai disease. For this reason, the World Health Organisation has recommended a maximum intake of 0.4–0.5 mg/week, and the maximum admissible concentration specified by the UK Water Supply (Water Quality) Regulations (1989) is 5 µg/l.

Chromium The toxicity of chromium is such that it is on the list of priority pollutants defined by the US Environmental Protection Agency. In the aqueous environment it can exist in either a hexavalent or a trivalent form. Of the two, the former has the greater chronic toxicity for freshwater fish, particularly in soft water (Mance et al., 1984). The same is true for livestock. Once inside the body, chromium becomes more or less immobilised as trivalent chromium and, therefore, tends to accumulate. Bioconcentration factors of 116 and 192 have been reported for rainbow trout and mussels respectively (Elwood et al., 1980; Anon., 1980).

Cobalt Cobalt is not generally included in lists of heavy metals which are harmful to the environment. Indeed, it has been reported that the US Environmental Protection Agency has no water quality criterion for cobalt because of the lack of toxicological data (Diamond et al., 1992). However, it is known that cobalt, although essential for microbial growth, can have adverse effects at high concentrations. Concentrations of 280 mg/l (as total soluble cobalt) have been shown to cause the complete inhibition of methanogenesis (Bhattacharya et al., 1995).

Copper For humans, copper is an essential element and the body can regulate its level homeostatically, although large, acute doses can have harmful, even fatal, effects. 2

Biosorption of Heavy Metals: An Introduction

There is some evidence to suggest that copper may be carcinogenic (Luckey and Venugopal, 1977). Copper can cause damage to a variety of aquatic fauna (fish and invertebrates). For example, the LC50 values for freshwater invertebrates are generally less than 100 µg/l, although there is evidence to suggest that these values may be affected by the hardness of the water. In addition, copper is phytotoxic and, indeed, has been used as an algicide to control algal blooms. It can, therefore, cause plant damage if, for example, it is present at too high a concentration in sewage sludge which is applied to agricultural land.

Lead Lead is a metal which is toxic to humans, aquatic fauna and livestock. In the aquatic environment it will exist mainly in the inorganic, divalent state. It can also exist in a tetravalent state (e.g. alkyl leads), and there is evidence that bacteria can methylate inorganic lead (Wong et al., 1975; Harrison and Laxen, 1978). This means that lead which has been partially immobilised in soils or sediments can be redispersed into the environment in a form which, in some cases, is more toxic than inorganic lead. It is another metal that will accumulate in body tissue, although it is mainly inorganic lead which has these characteristics. This is of some significance as lead is a cumulative poison. The effects of lead toxicity in humans include hypertension and brain damage.

Mercury Mercury can enter the aquatic environment either through natural processes—the solublisation of material bound in sediments or ores—or by anthropogenic routes. In the aquatic environment it will exist, essentially, in one of two states: as an inorganic, divalent ion, which may be hydrated or complexed, or as an organic ion, monomethyl mercury. Although there has to be concern about the presence of the inorganic ion in water, it is the monomethyl mercury that must be thought of as the main problem in relation to human health. Most of the mercury found in fish is in the monomethyl form and 90% of any monomethyl mercury eaten by humans is absorbed by the body. Another point of concern is that monomethyl mercury can cross the placenta more easily than the inorganic species and cause damage to the foetus. The results of mercury poisoning—and there have been catastrophes—have been well documented and there is a high awareness of the risks associated with this metal and its compounds. The World Health Organisation recommends a maximum uptake of 0.3 mg/week and the maximum acceptable concentration in drinking water is 1 µg/l.

Nickel Nickel is rather similar to copper in that it is a metal which, in solution, does not have serious effects on humans but has an appreciable phytotoxicity. As such, its concentration in sludges to be applied to agricultural land has to be restricted. It can also be damaging to some fish species, particularly in soft waters. 3

Biosorbents for Metal lons

Silver The impact of silver is essentially as a bactericide. Indeed, in earlier times this property was used to prevent the spoilage of beverages, in surgery and as a disinfectant. Discharges of silver need to be restricted, not merely because of the price of silver but because of the possible effects it could have on biological treatment processes (Linkson, 1987). However, the work by Cooley et al. (1988) suggests that the silver-thiosulphate complex, which is the predominant form of silver in photographic wastewaters, will be converted to insoluble silver sulphide in the secondary treatment stage and immobilised in the sludge.

Tin Tin is not usually considered as one of the toxic heavy metals. Nevertheless, it is a metal which can have harmful effects on humans, although these effects are not as drastic as those caused by some of the other metals discussed in this chapter. Tin has also been found to be dispersed into the aquatic environment, including sediments, by the natural weathering of ores, from mines or the use of organotin compounds in antifouling paints (Peterson et al., 1979; Seidel et al., 1980). Of the various forms in which tin can be found in the environment, the organotins will have the greatest toxicity to aquatic life and it has been reported that inorganic tin can be methylated by bacteria (Shugui et al., 1989).

Zinc The use of zinc is widespread, with major inputs to the aquatic environment being likely from the manufacture of alloys, from galvanising and paper production (Mance and Yates, 1984). It is one of the more ubiquitous heavy metals, being present in most natural waters, even rainwater, but is not perceived to pose a serious risk to human health; the limit in drinking water is 5 mg/l. However, it is phytotoxic and limits are imposed on its concentration in sewage sludges applied to agricultural land.

Control Pollution of the aqueous environment is controlled within the European Union under the ‘framework’ Dangerous Substances Directive (76/464/EEC). This defines two lists (Table 1.3). List 1, the so-called Black List, contains those compounds considered so toxic, persistent or bio-accumulative within the environment that priority should be given to eliminating pollution by them. List 2, the Grey List, contains compounds which are environmentally harmful but less so than those in List 1. The limits for discharges of Black List metals are currently controlled by ‘daughter’ Directives concerning: • mercury for discharges by the chloralkali electrolysis industry (82/176/EEC) 4

Biosorption of Heavy Metals: An Introduction Table 1.3 Black and Grey List metals

• mercury for discharges other than from the chloralkali electrolysis industry (84/ 156/EEC) • cadmium and its compounds (85/513/EEC). Within the UK, the laws which are critical to the control of heavy metals are the Environmental Protection Act, Part 1 (1990), the Environmental Protection (Control of Dangerous Substances) Regulations and the Water Industry Act (1991), together with the Environmental Protection (Prescribed Processes and Substances) Regulations (1991). This legislation: • introduces the concept of Integrated Pollution Control (IPC), which is enforced by the Environment Agency (EA) • defines special category wastes whose discharge may not be permitted by water service companies without reference to the Secretary of State for the Environment, for whom the EA acts • empowers the Secretary of State (under Section 62 of the Environmental Protection Act) to make regulations for the treatment, keeping or disposal of particularly difficult or dangerous waste • defines prescribed processes. This means that a metal plating effluent discharged to a sewer might be ‘consented’ in three different ways, depending on the metal it contained. A metal which was not prescribed (e.g. nickel) would receive its consent in the normal way from the water service company. An effluent from a prescribed process, such as cadmium plating, would be assessed by the EA and would require IPC documentation. In the case of zinc plating, where some cadmium could be expected as an impurity of the zinc, because a prescribed substance was involved the case would be referred to the Secretary of State and the EA, on his behalf, would issue a notice of determination.

Treatment Because of these regulations, effluents or waters contaminated with heavy metals must be treated. The degree of treatment may range from a main process stream for a seriously polluted industrial waste to al polishing process to remove the trace concentrations which can remain after the main treatment. Thus, the type of process or combination of processes used will depend on the metal(s) involved and the ultimate concentration allowed. There are a variety of treatment processes for waste 5

Biosorbents for Metal lons Table 1.4 The composition of a typical untreated metal finishing effluent (Wild, 1987)

waters contaminated by heavy metals, and it is important to recognise that there may be primary pollutants (metals directly associated with the industry) and secondary pollutants (which arise from impurities in the primary metals). A good example of this is cadmium, which is frequently present when zinc is the primary metal. Also, it must be remembered that the primary pollutants themselves may be a mixture of

Figure 1.1 Schematic diagram of a continuous flow treatment for plating wastes

6


metals, together with a non-metallic pollutant such as cyanide. This can be seen from the data in Table 1.4, which gives the typical composition of an untreated metal finishing effluent. Taken together, these two points mean that the selection of a process stream for treating a metal-contaminated effluent will depend on the metals which are present and their chemistry. Types of process which have been used in process streams for the treatment of this type of effluent include: • precipitation, as either hydroxides or sulphides • oxidation/reduction • solid/liquid separation; settlement or flotation • membrane technology • ion exchange • adsorption. Of these, the precipitation, oxidation/reduction and solid/liquid separation route has traditionally been the most commonly used. Depending on the volumes involved, the treatment is operated as either a batch or a continuous process. Whichever process is used, it is essential to control the addition of chemicals, either by using metering pumps or, more precisely, by using pH and redox control electrodes (Figure 1.1). pH control is particularly important as both oxidation and reduction reactions may have pH optima. In addition, metal hydroxides, particularly those of zinc, tin and lead, can redissolve at pH values higher than those quoted in Table 1.5. One of the disadvantages of this technology is that it produces sludge. In other words, it transforms an aquatic pollution problem to one associated with solid waste disposal. The amount of sludge, as well as its quality, will vary with the chemicals being used and the metals involved. For example, sodium hydroxide often produces a bulky, poor settling sludge whereas lime gives a dense sludge with good settling properties. It must also be remembered that sludge dewatering is advisable before disposal, either to reduce transportation costs or to comply with disposal requirements. This is usually achieved by filtration. This type of technology provides for an effective removal of the polluting metals and, if a coagulation/flocculation stage is also used, removals of up to 99% can be

Table 1.5 pH values for hydroxide precipitation

7


Figure 1.2 Schematic diagram of the Chemelec cell

achieved. However, as increasingly more stringent standards are being required, the disposal of the solid residue (the sludge) may pose problems and the ethos of recycling is more prevalent than in earlier decades. Thus, alternative technologies must be considered. For the more valuable metals, recovery/recycling can be an attractive option and electrolysis is an effective technology for doing this. One of the more successful systems is the Chemelec cell. It consists of a fluidised bed of glass beads surrounded by an expanded mesh electrode (Figure 1.2). The beads are maintained in an expanded state by the recirculated rinse water. The metal is plated onto the cathode and, if the metal is reasonably uncontaminated, it may be used directly as the anode in the main plating process. Adsorption is also a technology which has been examined for removing heavy metals. The adsorbent which has probably received most attention is granular activated carbon (GAC), and the adsorptive characteristics of both commercial carbons and activated carbons produced from waste materials have been examined. For example, Bowers and Huang (1981) and Kim and Zoltek (1977) reported on the use of Filtrasorb 400 for the removal of hexavalent chromium, while Alaerts et al. (1989) used a carbon produced from coconut shells. Metals other than chromium have also been tested (Netzer and Hughes, 1984; Wilczak and Keinath, 1993). However, GAC cannot be thought of as a panacea; poor adsorption has been demonstrated for some metals. In these cases, an organic complexing agent is needed

Table 1.6 Typical biosorbents

8


(Bhattacharya and Cheng, 1987). GAC filters are also being evaluated for the ‘point of use’ removal of lead from drinking water (Kuennen et al., 1992). However, carbon adsorption is an expensive treatment process and, with environmental concern growing in developing countries, there has been considerable research interest into the use of alternative low-cost adsorbents which would be considered as single-use materials (Bhattacharya and Venkobachar, 1984). These have included carbons produced from wastes or natural products as well as the direct use of this type of material (Pollard et al., 1992). Since the range of materials which have been tested is extensive, it is not sensible to attempt to provide a comprehensive list of potential biosorbents. Table 1.6, however, shows a typical selection of the types of material which have been examined.

References ALAERTS, G.J., JITJATURANT, V. and KELDERMAN, P., 1989, Use of coconut shell based activated carbon for chromium(VI) removal, Water Science and Technology, 21, 1701– 1704. ANON., 1980, Ambient water quality criteria for chromium. Ref. Pb 81–117467, US Environmental Protection Agency. BHATTACHARYA, D. and CHENG, C.Y.R., 1987, Activated carbon adsorption of heavy metal chelates from single and multicomponent systems, Environmental Progress, 6, 110–117. BHATTACHARYA, S.K., UBEROI, V., MADURE, R.L. and HAGHIGHIPODESH, M.R., 1995, Effect of cobalt on methanogenesis, Environmental Technology, 16, 271–278. BHATTACHARYA, A.K. and VENKOBACHAR, C., 1984, Removal of cadmium(II) by lowcost adsorbents, Journal of Environmental Engineering, 110, 110–122. BOWERS, A.R. and HUANG, C.P., 1981, Activated carbon process for the treatment of chromium(VI)-containing wastewaters, Water Science and Technology, 13, 629–650. COOLEY, A.C., DAGON, T.J., JENKINS, P.W. and ROBILLARD, K.A., 1988, Silver and the environment, Journal of Imaging Technology, 14, 183–189. DIAMOND, J.M., WINCHESTER, E.L., MACKLER, D.G., RASNAKE, W.J., FANELLI, J.K. and GRUBER, D., 1992, Toxicity of cobalt to fresh-water indicator species as a function of water hardness, Aquatic Toxicology, 22, 163–179. ELWOOD, J.W., BEAUCHAMP, J.J. and ALLEN, C.P., 1980, Chromium levels in fish from a lake chronically contaminated with chromium from cooling towers, Journal of Environmental Studies, 14, 289–290. HARRISON, R.M. and LAXEN, D.P.H., 1978, Natural sources of tetra-alkyl lead in air, Nature (London), 275, 738–740. KIM, J.I. and ZOLTEK, J., 1977, Chromium removal with activated carbon, Progress in Water Technology, 9, 143–155. KUENNEN, R.W., TAYLOR, R.M., VANDYKE, K. and GROENEVELT, K., 1992, Removing lead from drinking-water with a point-of-use GAC fixed-bed adsorber, Journal of the American Water Works Association, 84, 91–101. LINKSON, P.B., 1987, The treatment of silver-bearing wastewaters, In: BARNES, D., FORSTER, C.F. and HRUDEY, S.E. (Eds) Surveys in Industrial Wastewater Treatment, Vol. 3, Longman Scientific and Technical, Harlow, pp. 65–90. LUCKEY, T.D. and VENUGOPAL, B., 1977, Metal toxicity in mammals, In: Physiological and Chemical Basis for Metal Toxicity, Volume 1. Plenum Press, New York and London. MANCE, G., BROWN, V.M., GARDINER, J. and YATES, J., 1984, Proposed environmental quality standards for List II substances in water: chromium, Technical Report TR 207, Medmenham: Water Research Centre. 9

Biosorbents for Metal lons MANCE, G. and YATES, J., 1984, Proposed environmental quality standards for List II substances in water: lead, Technical Report TR 209, Medmenham: Water Research Centre. NETZER, A. and HUGHES, D.E., 1984, Adsorption of copper, lead and cobalt by activated carbon , Water Research, 18, 927–933. PETERSON, P.J., BURTON, M.A.S., GREGSON, M., NYE, S.M. and PORTER, E.K., 1979, Accumulation of tin by mangrove species in west Malaysia, Science of the Total Environment, 11, 213–221. POLLARD, S.J.T., FOWLER, G.D., SOLLARS, C.J. and PERRY, R., 1992, Low-cost adsorbents for waste and wastewater treatment: a review, Science of the Total Environment, 116, 31–52. SEIDEL, S.L., HODGE, V.F. and GOLDBERG, E.D., 1980, Tin as an environmental pollutant, Thalassia Jugoslavica, 16, 209–223. SHUGUI, D., GUOLAN, H. and YONG, C., 1989, The methylation of inorganic tin by humic materials in an aquatic environment, Applied Organometallic Chemistry, 3, 437–441. WILCZAK, A. and KEINATH, T.M., 1993, Kinetics of sorption and desorption of copper(II) and lead(II) on activated carbon, Water and Environment Research, 65, 238–244. WILD, J., 1987, Liquid wastes from the metal finishing industry, In: BARNES, D., FORSTER, C.F. and HRUDEY, S.E. (Eds), Surveys in Industrial Wastewater Treatment, Vol. 3, Longman Scientific and Technical, Harlow, pp. 21–64. WONG, P.T.S., CHAU, Y.K. and LUXON, P.L., 1975, Methylation of lead in the environment, Nature (London), 253, 263–264.

10

2

The Use of Algae as Metal Biosorbents G.W.GARNHAM

Introduction In this chapter the use of various types of algae and living, dead, immobilised and algal derived products which can be used as metal biosorbents will be described. This includes metal biosorption using photosynthetic microorganisms, both unicellular microalgae and cyanobacteria (formerly known as the blue-green algae), as well as multicellular macroalgae. Biosorption can be defined as the removal of metal or metalloid species, compounds and particulates from solution by biological material (Gadd, 1990). Large quantities of metals can be accumulated by a variety of processes dependent and independent on metabolism. Both living and dead biomass as well as cellular products such as polysaccharides can be used for metal removal. Such algal biomass can be used in a ‘natural form’ free in solution or immobilised by various techniques on to a solid support or to produce granules for a metal removal/recovery process. Algal biosorption techniques/processes can be used to remove toxic metals and/or radionuclides from liquid effluents before their safe discharge in addition to their use as a recovery process for metals of value. In the majority of cases biosorption of metals can be considered as rapid metabolism-independent ionic and covalent binding to a particular structure of the cell surfaces, although initial binding to the cell surface may be followed by further inorganic deposition of increased amounts of metal, possibly up to 50% of dry weight (Gadd, 1990). In addition, precipitation or crystallisation of metals may occur within and around cell walls as well as the production by algae of metal binding polysaccharides and siderophores: these are processes which could be considered as biosorption, but may in some cases be dependent on cell metabolism and are thus better termed ‘bioaccumulation’. Active metal uptake by energy-dependent transport systems should not, and will not, be considered as biosorption. An important feature of biosorption is that dead cells can be responsible for binding and accumulation of metals. Remaining cell debris such as cell walls can still represent a potent biosorbent (Volesky, 1990). So far there has been little commercial exploitation of algal biosorption for metal removal or recovery processes, despite the wealth of information and studies on such algal-based processes and, in many cases, the availability of abundant and relatively inexpensive algal biomass for use in future metal biosorption processes (Kuyucak, 11


1990). The filamentous algae, e.g. Cladophora, have been shown to contribute significantly to the removal of heavy metals (Shumate and Strandberg, 1985). Their use in artificial stream-meander systems, located at mine-mill operations to encourage biological removal of metals prior to discharge to the receiving water course, stands out as one of the few examples of the commercial exploitation of algal biosorption of metals. This chapter will describe and review metal biosorption by algae. It will try to offer explanations for the lack of exploitation of this technology and identify areas where future research should concentrate to make it an attractive alternative to existing metal removal and recovery processes.

Biosorption by algae and the mechanisms involved Accumulation of metals/radionuclides by algae can be described as consisting of two phases: a rapid phase of metabolism-independent binding to the cell surface (biosorption) followed by a slower phase due to simultaneous effects of growth and surface adsorption, active or intracellular uptake by passive diffusion (Khummongkol et al., 1982; Garnham et al., 1991). This can be illustrated by cobalt accumulation by the microalga Chlorella salina (Figure 2.1). Here there is an initial phase of biosorption in the first five minutes of accumulation, independent of light or the presence of metabolic inhibitors, e.g. KCN, DNP or CCCP. In contrast, the second phase is clearly dependent on light and could be inhibited by metabolic inhibitors, indicating that it is an active process and should not be considered as biosorption as previously discussed. This bi-phasic uptake has been described in many studies with microalgae, macroalgae and cyanobacteria with the metals zinc, cobalt, copper, manganese, cadmium, mercury, nickel, caesium, gold and magnesium (Gadd, 1988; Kuyucak and Volesky, 1990; Garnham et al., 1991; Karamushka et al., 1995). Equally, there are metals which show only one phase of uptake with algae due to biosorption, and in which no active uptake is observed, e.g. uranium, technetium, molybdenum, arsenic, copper, plutonium, polonium, zirconium, platinum, lead, silver, thorium and radium (Nakajima et al., 1981; Sakaguchi et al., 1981; Kuyucak and Volesky, 1990; Garnham et al., 1992a, 1993a, 1993b). This first rapid phase of metal uptake by algae independent of metabolism is the process which will be considered as biosorption in this chapter. The variety of different algal-based systems which biosorb metals and the amount of different metals which can be biosorbed by algal systems are summarised in Tables 2.1–2.3. This biosorption phase can be evaluated quantitatively and comparisons made of different biosorption systems using various adsorption isotherms. Two adsorption isotherm models are commonly used to describe the biosorption of metal species by algae. These are the Langmuir and Freundlich models, which can describe biosorption of metals by algae equally well. The Langmuir model can be represented by the following equation:

where qe is the uptake of metal at equilibrium (mg/g), Ce is the concentration of metal species in solution at equilibrium (mg/l), aL and KL are Langmuir constants and Q0 is the maximum metal uptake (=KL/aL). 12

The Use of Algae as Metal Biosorbents

Figure 2.1 Accumulation of cobalt by Chlorella salina at a cell density of 106 cells/ml in 10 mM TAPS buffer, pH 8, with 25 µM cobalt at 23°C: () light incubation; () dark incubation or in the presence of a metabolic inhibitor. Adapted from Garnham et al. (1994)

The Langmuir model assumes the following: the surface consists of adsorption sites, all adsorbed species interact only with a site and not with each other, adsorption is limited to a monolayer, and adsorption energy of all the sites is identical and independent of the presence of adsorbed species on neighbouring sites. The Freundlich adsorption model can be described by the following equation:

He re, K gives a measure of the adsorbent capacity and the slope 1/n gives the intensity of adsorption. Although the biosorption of metals by algae often fits these two models which have been developed to describe ‘pure’ monolayer adsorption mechanisms, it should not be assumed that the biosorption mechanism is just ‘pure adsorption’: other mechanisms will probably be involved. The Scatchard plot used to describe protein-ligand binding can also be used to describe the binding of metals to algal surfaces. The types of plots which can be obtained and their interpretation are shown in Figure 2.2. Kads, the binding affinity constant, and R tads, the maximal binding capacity of the algal biomass, can be determined from a Scatchard plot. The intercept on the y-axis will give the binding capacity and the slope of the line the binding affinity. With this model, as with the other two, a good fit between the data recorded and the plots does not necessarily prove the 13

Biosorbents for Metal lons Table 2.1a Metal biosorption by ‘living’ micr roalgae free in solution

14

The Use of Algae as Metal Biosorbents Table 2.1a (Continued)

Figure 2.2 The Scatchard plot: r =number of moles of metal bound to set quantity of algal biomass, i.e. gram dry wt; [C]=concentration of unbound metal; Rtads=(intercept) maximal binding capacity; K ads =(gradient) binding affinity constant; 1=single set of binding sites with no cooperativity; 2=two sets of binding sites with no cooperativity; 3=single set of binding sites with cooperativity; 4=single set of binding sites with negative cooperativity

15

Biosorbents for Metal lons Table 2.1b Metal biosorption by ‘living’ cyanobacteria free in solution

16

The Use of Algae as Metal Biosorbents Table 2.2 Biosorption by macroalgal biomass

binding mechanism involved, but the binding constants obtained from the models may be valid for the specified conditions, and may be used in order to evaluate the relative extent of metal binding with cell surfaces (Xue and Sigg, 1990). All the models shown have been used to described the biosorption of a single metal by algal biomass with some success (Volesky, 1994; Crist et al., 1981; Garnham et al., 1994). Little work has investigated the biosorption of more than one metal by algal biomass; most studies represent any results concerning this in a simplistic manner with no model to allow a real performance assessment of the biosorption system or comparisons of different sets of results (Carvalho et al., 1995). However, studies by Carvalho et al. (1995) and Chong and Volesky (1995) have addressed this problem and modelled the biosorption of two metals in a two-metal system, e.g. (Zn/Cd, Cu/Cd, Cu/Zn) by biomass derived from the brown macroalgae Ascophyllum nodosum. Usually studies of two-metal biosorption evaluation have concentrated on the biosorptive uptake of one primary metal and how it is affected by the presence of a secondary metal, but Carvalho et al. (1995) and Chong and Volesky (1995) looked at the total metal uptake for both metals together on an x-axis and the 17

Biosorbents for Metal lons Table 2.3 Biosorption of metals by treated or immobilised algal biomass systems

18


final concentration of each metal on the y-axis. This approach led to the construction of a three-dimensional plot of a ‘sorption surface’ instead of a two-dimensional adsorption isotherm plot to evaluate the biosorption process. In general, the mechanism of biosorption is based on a number of metal-binding processes taking place with components of the algal cell wall. The algal cell wall can reversibly biosorb metals and thus functions in a similar way to an ion-exchange resin (Darnall et al., 1986). Thus, the biosorption mechanism can be considered as being dependent on the composition of the algal cell wall. The cell wall structure and chemical composition of most algal species have not yet been described in detail. There is a wide variation in the chemical composition of eukaryotic algal species (both macroalgae and microalgae), the only cell wall component common to all the eukaryotic algal divisions being cellulose (Greene and Darnall, 1990). Algal cell walls can be made up with further polysaccharides: mannan, xylan, alginic acid, chitin, etc. These components, along with the proteins present, can provide acid binding sites such as amino, amine, hydroxyl, imidiazole, phosphate and sulphate groups (Crist et al., 1981). The proteins in the cell wall provide functional groups for binding (Crist et al., 1981) as illustrated in Figure 2.3. The biosorption mechanism has been described as not involving van der Waals’ forces at the cellulose network of the cell wall, thus both ionic charge and covalent bonding are involved in the metal biosorption process (Greene and Darnall, 1990). It is thought that the proteins and polysaccharides are the major components responsible for the biosorption. Covalent bonding could be expected with amino and carboxyl groups and ionic charge bonding with carboxyl and sulphate groups associated with these components. Studies by Stary and Kratzer (1984) with the microalga Scenedesmus obliquus indicated that the cell wall behaved like a weak acidic cation exchanger containing various cell wall ligands with different exchange capacities. Further information regarding metal biosorption by algae has been obtained through NMR studies which indicated that carboxyl groups were responsible for the biosorption of cadmium by the microalga Stichococcus bacillaris (Majidi et al., 1990). Because of the wide morphological and chemical variation in algal cell walls within algal species, there is also variation in metal biosorption properties between algal species as illustrated in Tables 2.1 and 2.2. The cyanobacteria (blue-green algae), which here are considered with the eukaryotic algae, have a cell wall similar to that of Gram-negative eubacteria. The major components of cyanobacterial walls are murein (peptidoglycans or mucopeptides associated with as many as eight amino acids) containing diaminopimelic acid, muramic acid and N-acetyl glucosamine (Bold et al., 1980).

Figure 2.3 Protein functional groups involved in biosorption of metals by algae

19


Some cyanobacteria also have external sheaths composed of a matrix of pectic acids and mucopolysaccharides. Most metal deposition occurs at polar regions of the constituent membranes or within the peptidoglycan layer (Shumate and Strandberg, 1985). Reed and Gadd (1990) described the metal binding mechanism as possibly being a two-stage process, first involving interactions between metal ions and reactive groups as previously described with the eukaryotic algae, followed by inorganic deposition of increased quantities of metal. This leads to large amounts of metal being accumulated which cannot be interpreted solely in terms of an ion-exchange phenomenon. As with eukaryotic macroalgae and microalgae, variation in metal biosorption is seen between species due to differences in cell wall structure and composition (Table 2.1b).

Factors affecting the biosorption of metals by algae Previously biosorption of metals by algae has been described as being dependent on the algal species used and differences in the cell wall composition of the species. However, other factors affecting the biosorption of metals by algal biomass should be considered, such as cell size and morphology, pH of external media, cation and anion concentration of external media, metal speciation, temperature and physiology of the algal biomass used.

Cell size and morphology Differences observed in the amounts of metal biosorbed by different algal species can be explained by differences in cell size and shape between species, in addition to cell wall structure differences. The greater the cell surface area to dry weight ratio, the greater the amount of metal biosorbed by a cell surface per unit weight. Large metal concentration factors observed by Fisher (1985) with a Synechococcus sp., a small spherical cyanobacterium, were explained as being due to their small size and thus great surface area to weight ratio.

pH of external medium The biosorption of metals by algae is dependent on the pH of external medium. Generally, there is increased biosorption of cationic metal species with increasing pH values, as illustrated for cobalt in Figure 2.4. Greene and Darnall (1990) classified metal ions into three classes based on their pH dependence of biosorption to algae. The first class are tightly bound at pH>5 and can be desorbed at pHCu 2+ >Cd 2+ >Ni 2+>Pb 2+>Zn 2+ =Co 2+ >Cr 3+ with the microalga Chlorella vulgaris (Greene et al., 1987) and Cu2+>Sr2+> Zn2+>Mg2+>Na + with the macroalga Vaucheria (Crist et al., 1981). This selectivity in biosorption can be explained by binding sites on algae having preferences for ‘hard’ or ‘soft’ metal ions. Biosorptions of metals occurring at ‘hard’ and ‘soft’ sites would not be expected to compete with each other. However, borderline metals would compete with each other because of similarities in their co-ordination chemistry (Greene and Darnall, 1990). Studies with biomass from the macroalga Ascophyllum nodosum by Carvalho et al. (1995) described competition between zinc, copper and cadmium for binding sites. The biomass showed a preference for copper over zinc and cadmium. However, the presence of zinc tends to lower the total metal uptake in a two-metal system, but the uptake of zinc could not compensate for the inhibition in biosorption of the other metals. At low concentrations independent binding sites were indicated for copper and cadmium, but at higher concentration>1 mM interference of the biosorption of the metals was observed, indicating a degree of overlap in the sorption site function. In some situations cation concentrations increase the biosorption of anionic metal species, as illustrated by the increased biosorption of pertechnetate by microalgae at increased external concentrations of Na+, K+, Mg2 + and Ca2+ (Garnham et al., 1992b). This is explained by cationic binding to binding sites on the algae increasing subsequent anionic binding as seen with a pH decrease and proton binding. Anionic concentrations rarely affect the biosorption of metals by algae. One of the few exceptions is a reduction in uranium biosorption by Chlorella in the presence of concentrations of either (Sakaguchi et al., 1978).

Metal speciation This is of importance since, as discussed previously, the charge of a metal species will affect how much metal is biosorbed and how pH and competing ions affect biosorption. An example of the effect speciation can have on the biosorption of a metal is the reduction in biosorption of uranium by Chlorella in marine conditions due to the formation of the following and complexes instead of and UO2OH+ (Sakaguchi et al., 1978). In marine conditions cadmium biosorption by algae is also decreased, since cadmium exists as the chloride complexes and CdCl2 instead of Cd2+ (Trevors et al., 1986). Under various pH conditions and redox potentials, mercury can be found in the forms , Hg(OH)2, Hg0, HgCl(OH) and 22


HgCl2 as well as Hg2+ (Farrell et al., 1993), and biosorption of mercury is dependent on which species are present.

Temperature In most metal biosorption studies with algae, temperature is usually kept close to that of the temperature of the laboratory during experiments. In studies where temperature varied between 0 and 30ºC little effect was seen on the biosorption of manganese or molybdenum by Chlorella cells (Nakajima et al., 1979; Sakaguchi et al., 1981).

Physiological state of algal biomass This greatly affects the amount of metal biosorbed by algal biomass. There is a large difference between the metal biosorbed by ‘dead’ and ‘living’ algal biomass. In most studies biosorption of metals is greater or equal with ‘dead’ biomass as compared to an equal amount of ‘living’ biomass. When algal biomass is in a ‘dead’ state the cells are permeable and allow metals to enter and bind on internal components and surfaces of the cell as well as the external surface, thus increasing the metal uptake/ biosorption. Horikoshi et al. (1979) showed that killed cells biosorbed five times more uranium than living cells. The nutritional regime under which algal cells are cultivated and their resulting physiology may also affect metal biosorption. Biosorption of cobalt, zinc and manganese by Chlorella emersonii and Scenedesmus obliquus was reduced when the cells were grown under photoheterotrophic and chemoheterotrophic as compared with photoautotrophic conditions (Garnham et al., 1992b). This was explained by changes in cell wall structure and/or polysaccharides as described by Martinez et al. (1991) due to growth conditions affecting the biosorption. There is also evidence that when cells are immobilised and still ‘living’, e.g. in alginate beads, there is an increase in metal biosorption which can be related to changes in cell physiology due to immobilisation (Garnham et al., 1992d). This effect will be discussed further below. Obviously, all the factors described above should be considered when developing an industrial algal biosorption system for the separation of metals. The metal to be separated, the species in which it is present and the pH of the effluent stream should be considered. Unless the metal is present as a negatively charged species it is unlikely that an algal system would be efficient in removing the metal from an effluent or liquor of low pH (pH 9 or 99% removal with metal loadings>150 mg/g biomass for economic success, and that many examples of biosorption (including algal-based systems) satisfy this criterion (Gadd, 1990). Although a large amount of data is available regarding algal biosorption, as yet few algal-based systems have been commercialised. The first algal-based system used commercially was based on benthic algal mats contained in a meander system. These were successfully used by New Lead Belt, Missouri (USA) to remove lead and other metals from its mining and milling effluents. Gale and Wixon (1979) report levels of lead in the effluent decreasing from 3 ppm to 50 ppb within 3 hours. The advantage of this system was that, after the initial cost of construction, there was no further cost 31


as long as the algae continued growing. The disadvantage of the system is that the metals are removed from the effluent but remain in the environment, immobilised in the algal mats. In addition to this application, three other companies in North America are producing algal-based biosorption systems, as follows: • Bio-recovery Systems Inc. (Las Cruces, New Mexico) produces a biosorbent based on immobilised Chlorella in silica or polyacrylamide gels. This is called AlgaSORB™ and efficiently removes metal from concentrations of 100 mg/l to levels below 1 mg/ml. The biosorbent resembles an ion-exchange resin and can undergo more than 100 biosorption-desorption cycles. In addition, the system is not affected by concentrations of elements such as calcium and magnesium. The cost of producing such a biosorbent has not been revealed, apart from the cost of the Chlorella used, which was US$0.25 per lb (1 lb=0.45 kg) (Kuyucak, 1990) • B.V. Sorbex Inc. (Montreal, Canada) has developed different biosorbents for specific metal recovery using all types of biomass, including the algae Sargassum natans, Ascophyllum nodosum, Halimeda opuntia, Palmyra pamata, Chondrus crispus and Chlorella vulgaris. These can work over a range of pH values and solution conditions, and can biosorb a wide range of metals without suffering interference from concentrations of calcium or magnesium. They can be used with a wide range of metal concentrations, are not affected by organics and can be regenerated easily (Volesky, 1990) • U.S. Bureau of Mines (Golden, Colorado) has produced Bio-fix, a granular biosorbent consisting of a variety of biomasses including algae immobilised in porous polypropylene beads. The biosorbent is selective for heavy metals over alkaline earth metals, and the beads have been tested extensively for the treatment of acid mine waste (Eccles and Holroyd, 1993). In all these systems the biosorbent is either used as a fixed bed using a normal downflow stream or fluidised by an upflow of liquid. Pre-treatment of a liquor may be required in some cases, and the type of system employed is dependent on the amount of liquor to be processed, its composition and the regeneration conditions required for the biosorbent. Both Bio-recovery Systems and B.V. Sorbex offer small canisters as flow-through fixed-bed systems, as well as large-scale fluidised-bed or pulsed-bed systems capable of handling flows in excess of 100 m3/day (Kuyucak, 1990). The full costs of an algal-based biosorption metal recovery system are not well documented, although Volesky (1990) compared the costs of algal-based biosorbents with those of other types of metal adsorbents. He described the cost of a marine macroalgal biosorbent as ranging from US$1 to US$2/kg; specifically cultured microalgal biosorbent cost US$7–8/lb and an ion-exchange resin cost US$7–14/lb (1 lb=0.45 kg). Darnall described the procurement of the algal biomass for a metal removal process as being the most important part of the final cost of the process (Kuyucak, 1990). Eccles and Holroyd (1993) made a cost assessment of two biosorption processes as compared to alkaline precipitation or ion-exchange for treating electroplating waste (nickel, cadmium, chromium and zinc giving a total metal concentration of 60 mg/l) with a flow rate of 50000 gallons/day, five days a week (1 gallon=4.5 l). In the assessment, capital equipment price included metal 32


removal, biosorbent regeneration and metal recovery systems. A five year depreciation was allowed for the capital equipment. Use of the bacteria-based (Bacillus subtilis) AMT-Bioclaim process (Visa Tech Ltd) was shown to give a 50% saving over alkaline precipitation and a 28% saving over ion-exchange. The cost of the Bio-fix process (partially based on algal biomass) was compared with lime precipitation for treating a waste containing zinc, manganese and cadmium in mg/l quantities at a flow rate of 1700 gallons/min for 365 days/year with a pH of 6.9. A three-column packed-bed system of Bio-fix beads was envisaged. The first column would be for removal of most metals, the second would scavenge for metals and the third would be eluted and the metals recovered. At the end of each loading cycle, column one would be eluted, column two would become the new lead column and the previously eluted column would become the scavenger. When the costs of this system were compared with lime precipitation it was found that the costs of the two systems were similar per 1000 gallons of waste treated. However, the income from the possible sale of recovered metals from the biofix and the increased cost of lime precipitation when the waste liquor contains iron were not considered. To summarise, algal biosorptive processes have not yet been significantly commercialised even though they possess the following advantages over currently available technologies for metal separation: • versatility and flexibility for a wide range of applications • robustness • selectivity for heavy metals over alkaline earth metals • ability, in some cases, to reduce metal concentrations to drinking water standards • cost-effectiveness against alternative processes. Perhaps one reason why algal biosorption systems have not been widely used in industry is the lack of knowledge about the engineering of such systems: engineers may feel more comfortable with tried systems such as ion-exchange which they understand and can relate to better than a biological-based system. Thus, it is felt that in order for such algal-based metal separation systems to be fully exploited, it will be necessary to adopt a multidisciplinary approach in which chemists, biologists and engineers work together.

References AKSU, Z. and KUTSAL, T., 1991, A bioseparation process for removing lead (II) ions from waste water by using C. vulgaris. Journal of Chemical Technology and Biotechnology, 52, 109–118. ASHLEY, N.V. and ROACH, D.J.W., 1989, Review of biotechnology applications to nuclear waste treatment, Journal of Chemical Technology and Biotechnology, 26, 381–394. ASTHANA, R.K., CHATTERJEE, S. and SINGH, S.P., 1995, Investigations on nickel biosorption and its remobilisation, Process Biochemistry, 30, 729–734. BADOUR, S.S., GODAVARI, H.R. and WAYGOOD, E.R., 1970, A review of mass cultivation of microalgae, Biomass Energy Institute, Winnipeg, Manitoba. BEDELL, G.W., 1990, Propagation of freshwater algal biosorbents, In: Biosorption of Heavy Metals, VOLESKY, B. (Ed.), CRC Press, Boca Raton, Florida, pp. 360–368. 33

Biosorbents for Metal lons BIRCH, L. and BACHOFEN, R., 1990, Complexing agents from microorganisms, Experimentia, 46, 827–834. BOLD, H.C., ALEXOPOULUS, C.S. and DELEVORYUS, T., 1980, The Morphology of Plants and Fungi, 4th Edn, Harpers and Row, New York. BOROWITZKA, M.A. and BOROWITZKA, L.J., 1988, Microalgal Biotechnology, Cambridge University Press, Cambridge. CAIN, J.R., PASCHAL, D.C. and HAYDEN, C.M., 1980, Toxicity and bioaccumulation of cadmium in the colonial green alga Scenedesmus obliquus, Archives of Environmental Contamination and Toxicology, 9, 9–16. CARVALHO DE, R.P., CHONG, K.-H. and VOLESKY, B., 1995, Evaluation of Cd, Cu and Zn biosorption in two-metal system using algal biosorbent, Biotechnology Progress, 11, 39–44. CHONG, K.H. and VOLESKY, B., 1995, Description of two-metal biosorption equilibria by Langmuir-type models, Biotechnology and Bioengineering, 47, 451–460. CHRISMADHA, T. and BOROWITZKA, M., 1994, Effect of cell density and irradiance on growth, proximate composition and eicosapentaeoic acid production of Phaedactylum tricornutum grown in a tubular photobioreactor , Journal of Applied Phycology, 6, 67– 74. CONWAY, H.L. and WILLIAMS, S.C., 1979, Sorption of cadmium and its effect on growth and the utilization of inorganic carbon and phosphorus of two freshwater diatoms, Journal of Fish Reserves Board of Canada, 36, 579–586. CRIST, R.H., MARTIN, J.R., GUPTILL, P.W., ESLINGER, J.M. and CRIST, D.R., 1990, Interaction of metals and protons with algae. 2. Ion exchange in adsorption and metal displacement by protons, Environmental Science and Technology, 24, 337–342. CRIST, R.H., OBERHOLSER, K. and NGUYEN, M., 1981, Nature of bonding between metallic ions and algal cell walls, Environmental Science and Technology, 15, 1212–1217. DA COSTA, A.C.A. and LEITE, S.G.F., 1992, Cadmium and zinc biosorption by Chlorella homosphaera, Biotechnology Letters, 12, 941–944. DARNALL, D.W., GREENE, B., HOSEA, M., M C PHERSON, R.A., HENZL, M. and ALEXANDER, M.D., 1986, Recovery of metals by immobilised algae, In: Trace Metal Removal from Aqueous Solutions, THOMPSON, R. (Ed.), Litho Ltd, Whitstable, Kent, pp. 1–24. DONGMANN, G. and NURNBERG, H.W., 1982, Observations with Thalassariosira rotula on the toxicity and accumulation of cadmium and nickel, Ecotoxicology and Environmental Safety, 6, 535–544. ECCLES, H. and HOLROYD, C., 1993, Why select a metal biosorption process?, Proceedings of SCI conference on Biological Removal of Toxic Metals, Preston, pp. 1–2. FARRELL, R.E., GERMIDA, J.J. and MING HUANG, P., 1993, Effects of chemical speciation in growth media on the toxicity of mercury (II), Journal of Applied and Environmental Microbiology, 59, 1507–1514. FERNANDEZ-PINAS, F., MATEO, P. and BONILLA, I., 1991, Binding of cadmium by cyanobacterial growth media: free ion concentration as a toxicity index to the cyanobacteria Nostoc UAM 208, Archives of Environmental Contamination and Toxicology, 21, 425–431. FISHER, N.S., 1985, Bioaccumulation of metals by marine picoplankton, Marine Biology, 87, 137–142. FISHER, N.S., BOHE, M. and TEYSSIE, J.-L., 1984, Accumulation and toxicity of Cd, Zn, Ag and Hg in four marine phytoplankters, Marine Ecology—Progress Series, 18, 201–213. FLETCHER, R.L., 1993, Green tides: an overview, The Phycologist, 34, 26–27. FOLSOM, B.R., POPESCU, N.A. and WOOD, J.M., 1986, Comparative study of aluminium and copper transport and toxicity in an acid-tolerant freshwater green alga, Environmental Science and Technology, 20, 616–620.

34

The Use of Algae as Metal Biosorbents GADD, G.M., 1988, Accumulation of metals by microorganisms and algae, In: Biotechnologyspecial microbial processes, Vol. 6B, REHM, H.-J. (Ed.), Weinheim: VCH Verlagsgellschaft, pp. 401–433. GADD, G.M., 1990, Biosorption, Chemistry and Industry, 13, 421–426. GALE, N.L. and WIXON, B.G., 1979, Control of heavy metals in lead industry effluents by algae and other aquatic vegetation, Proceedings of the International Conference on Management and Control of Heavy Metals in the Environment, Edinburgh, CEP Consultants Ltd, pp. 580–583. GARNHAM, G.W., AVERY, S.V., CODD, G.A. and GADD, G.M., 1994, Interactions of microalgae and cyanobacteria with toxic metals and radionuclides: physiology and environmental implications, In: Fluxes in Estuaries, DYER, K.R. and ORTH, R.J. (Eds), Olsen and Olsen, Denmark, pp. 289–293. GARNHAM, G.W., CODD, G.A. and GADD, G.M., 1991, Effect of salinity and pH on cobalt biosorption by the estuarine microalga Chlorella salina, Biology of Metals, 4, 151–157. GARNHAM, G.W., CODD, G.A. and GADD, G.M., 1992a, Kinetics of uptake and intracellular location of cobalt, manganese and zinc in the estuarine microalga Chlorella salina, Journal of Applied Microbiology and Biotechnology, 37, 270–276. GARNHAM, G.W., CODD, G.A. and GADD, G.M., 1992b, Uptake of technetium by freshwater green microalgae, Journal of Applied Microbiology and Biotechnology, 37, 679–684. GARNHAM, G.W., CODD, G.A. and GADD, G.M., 1992c, Effect of nutritional regime on accumulation of cobalt, manganese and zinc by green microalgae, FEMS Microbiology Letters, 98, 45–50. GARNHAM, G.W., CODD, G.A. and GADD, G.M., 1992d, Accumulation of cobalt, zinc and manganese by the estuarine green microalgal Chlorella salina immobilised in alginate microbeads, Environmental Science and Technology, 26, 1764–1769. GARNHAM, G.W., CODD, G.A. and GADD, G.M., 1993a, Accumulation of zirconium by microalgae and cyanobacteria, Journal of Applied Microbiology and Biotechnology, 39, 666–672. GARNHAM, G.W., CODD, G.A. and GADD, G.M., 1993b, Accumulation of technetium by cyanobacteria, Journal of Applied Phycology, 5, 307–315. GARNHAM, G.W. and GREEN, M., 1995, Chromate (VI) uptake and interactions with cyanobacteria, Journal of Industrial Microbiology, 14, 247–251. GLADSTONE, J. and LOPEZ-REAL, J.M., 1993, Recycling and composting of seaweed wastes, The Phycologist, 34, 27–28. GNASSIA-BARELLI, M. and ROMEO, M., 1982, Short-term time series study of copper and cadmium uptake by Cricosphaera elongata, Journal of Experimental Marine Biology and Ecology, 61, 287–298. GNASSIA-BARELLI, M. and ROMEO, M., 1987, Uptake of zinc by cultured phytoplankters Hymenomonas elongata, Diseases of Aquatic Organisms, 3, 45–49. GREENE, B. and DARNALL, D.W., 1990, Microbial oxygenic photoautotrophes (cyanobacteria and algae) for metal-ion binding , In: Microbial Mineral Recovery, EHRLICH, H.L. and BRIERLY, C.L. (Eds), Mc Graw-Hill, New York, pp. 227–302. GREENE, B., MCPHERSON, R. and DARNALL, D.W., 1987, Algal sorbents for selective metal ion recovery, In: Metal Speciation, Separation and Recovery, PATTERSON, J. and PASINO, R. (Eds), Lewis, Michigan, pp. 315–332. HAMMOUDA, O., GABER, A. and ABDEL-RAOUF, N., 1995, Microalgae and wastewater treatment, Ecotoxicology and Environmental Safety, 31, 205–210. HARRIS, P.O. and RAMELOW, G.J., 1990, Binding of metal ions by particulate biomass derived from Chlorella vulgaris and Scenedesmus quadricauda, Environmental Science and Technology, 24, 220–234.

35

Biosorbents for Metal lons HASSETT, J.M., JENNET, J.V. and SMITH, J.E., 1981, Microplate technique for determining accumulation of metals by algae, Applied and Environmental Microbiology, 41, 1097–1106. HORIKOSHI, T., NAKAJIMA, A. and SAKAGUCHI, T., 1979, Studies on the accumulation of heavy metal elements in biological systems IV. Uptake of uranium by Chlorella regularis, Agriculture, Biology, Chemistry, 43, 617–623. HOSEA, M., GREENE, B., M C PHERSON, R., HENZL, M., ALEXANDER, M.D. and DARNALL, D.W., 1986, Accumulation of elemental gold on the alga Chlorella vulgaris, Inorganic Chimica Acta, 123, 161–165. JAKUBOWSKI, M. and SKOWRONSKI, T., 1991, Cadmium binding by some cyanobacteria and lake sediments, Proceedings 8th Conference on Heavy Metals in the Environment, CEP Consultants, Edinburgh, pp. 432–435. KARAMUSHKA, V.I., GRUZINA, T.G. and UL’BERG, Z.R., 1995, Accumulation of gold (III) by the cells of cyanobacterium Spirulina platensis, Microbiology, 64, 157–160. KHUMMONGKOL, D., CANTERFORD, G.S. and FRYER, C., 1982, Accumulation of heavy metals in unicellular algae, Biotechnology and Bioengineering, 4, 22643–22660. KUYUCAK, N., 1990, Feasibility of biosorbents application, In: Biosorption of Heavy Metals, VOLESKY, B. (Ed.), CRC Press, Boca Raton, Florida, pp. 372–377. KUYUCAK, N. and VOLESKY, B., 1988, Biosorbents for the recovery of metals from industrial solutions, Biotechnology Letters, 10, 137–142. KUYUCAK, N. and VOLESKY, B., 1990, Biosorption of algal biomass, In: Biosorption of Heavy Metals, VOLESKY, B. (Ed.), CRC Press, Boca Raton, pp. 173–198. MAHAN, C.A., MAJIDI, V. and HOLCOMBE, J.A., 1989, Evaluation of the metal uptake of several algae strains in a multicomponent matrix utilizing inductively coupled plasma emission spectrometry, Analytical Chemistry, 61, 624–627. MAJIDI, V., LAUDE, D.A. and HOLCOMBE, J.A., 1990, Investigation of the metal binding sites with 113Cd nuclear magnetic resonance, Environmental Science and Technology, 24, 1309–1312. MANN, H., FYFE, W.S. and KERRICH, R., 1988, The chemical content of algae and waters: bioconcentration, Toxicity Assessment, 3, 1–16. MARTINEZ, F., ASCASO, C. and ORUS, M.I., 1991, Morphometric and stereological analysis of Chlorella vulgaris under heterotrophic growth conditions, Annals of Botany, 67, 239–245. MCKNIGHT, D.M. and MOREL, F.M.M., 1980, Copper complexation by siderophores from filamentous blue-green algae, Limnology and Oceanography, 25, 62–71. NAKAJIMA, A., HORIKOSHI, T. and SAKAGUCHI, T., 1979, Uptake of manganese ion by Chlorella regularis, Agriculture, Biology and Chemistry, 43, 1461–1466. NAKAJIMA, A., HORIKOSHI, T. and SAKAGUCHI, T., 1981, Studies on the accumulation of heavy metal elements in biological systems XVII. Selective accumulation of heavy metal ions by Chlorella regularis, European Journal of Applied Microbiology and Biotechnology, 12, 76–83. NAKAJIMA, A., HORIKOSHI, T. and SAKAGUCHI, T. 1982, Recovery of uranium by immobilised microorganisms, European Journal of Applied Microbiology and Biotechnology, 16, 88–91. ORON, G., SHELEF, G. and LEVI, A., 1979, Growth of Spirulina maxima on cow-manure wastes, Biotechnology and Bioengineering, 21, 2169–2175. PAYER, H.D., CHIEMVICHAK, Y., HOSAKUL, K., KONGPANICHKUL, C., KRAIDEJ, L., NGUITRAGOOL, M., REUNGMANIPYTOON, S. and BURI, P., 1980,Temperature as an important climatic factor during mass production of microscopic algae, In: Algae Biomass, SHELEF, G. and SOEDER, C.J. (Eds), Elsevier/North Holland Biomedical Press, Amsterdam, pp. 389–400. REED, R.H. and GADD, G.M., 1990, Metal tolerance in eukaryotic and prokaryotic algae, In: Heavy Metal Tolerance in Plants, SHAW, J. (Ed.), CRC Press, Boca Raton, Florida, pp. 105–118. 36

The Use of Algae as Metal Biosorbents ROBINSON, N.J., 1989, Algal metallothioneins: secondary metabolites and proteins, Journal of Applied Phycology, 1, 5–18. ROBINSON, P.K., MAK, A.L. and TREVAN, M.P., 1986, Immobilised algae: a review, Process Biochemistry, 122–127. SAKAGUCHI, T., HORIKOSHI, T. and NAKAJIMA, A. 1978, Uptake of uranium from sea water by microalgae, Journal of Fermentation Technology, 56, 561–565. SAKAGUCHI, T., NAKAJIMA, A. and HORIKOSHI, T., 1981, Studies on the accumulation of heavy metal elements in biological systems XVII. Accumulation of molybdenum by green microalgae , European Journal of Applied Microbiology and Biotechnology, 12, 84–89. SAKAGUCHI, T., TSUJI, T., NAKAJIMA, A. and HORIKOSHI, T., 1979, Accumulation of cadmium by green microalgae, European Journal of Applied Microbiology and Biotechnology, 8, 207–212. SHUMATE, S.E., and STRANDBERG, G.W., 1985, Accumulation of metals by microbial cells, In: Comprehensive Biotechnology, Vol. 4. ROBINSON, C.W. and HOWELL, J.A. (Eds), Pergamon Press, London, pp. 235–247. SINGH, S.P., VERMA, S.K., SINGH, R.K., and PANDY, P.K., 1989, Copper uptake by free and immobilized cyanobacteria, FEMS Microbiology Letters, 11, 193–196. SKILL, S., Biotechnica Ltd UK, 1995, personal communication. SOEDER, C.J. and STENGEL, E., 1974, Physico-chemical factors affecting metabolism and growth rate, In: Algal Physiology and Biochemistry, STEWART, W.D. P. (Ed.), Blackwell, Oxford, pp. 714–740. STARY, J. and KRATZER, K., 1984, Mechanisms of the uptake of metal cations by algal cell walls, Toxicology and Environmental Chemistry, 12, 67–71. TING, Y.P., LAWSON, F. and PRINCE, I.G., 1989, Uptake of cadmium and zinc by the alga Chlorella vulgaris: part 1. Individual ion species, Biotechnology and Bioengineering, 34, 990–999. TREVORS, J.T., STRATTON, G.W. and GADD, G.M., 1986, Cadmium transport, resistance and toxicity in bacteria, algae and fungi, Canadian Journal of Microbiology, 32, 447–464. TROLLOPE, D.R. and EVANS, B., 1976, Concentrations of copper, iron, lead, nickel and zinc in freshwater algal blooms, Environmental Pollution, 11, 109–116. TSEZOS, M., 1984, Recovery of uranium from biological adsorbents. Desorption equilibrium. Biotechnology and Bioengineering, 26, 973–981. TSEZOS, M., 1986, Adsorption by microbial biomass as a process for removal of ions from process or waste solutions, In: Immobilisation of Ions by Biosorption, ECCLES, H. and HUNT, S. (Eds), Ellis Horwood, Chichester, pp. 201–218. VOLESKY, B., 1990, Removal and recovery of heavy metals by biosorption, In: Biosorption of Heavy Metals, VOLESKY, B. (Ed.), CRC Press, Boca Raton, Florida, pp. 7–44. VOLESKY, B., 1994, Advances in biosorption of metals: selection of biomass type, FEMS Microbiology Reviews, 14, 291–302. WILKINSON, S.C., GOULDING, K.H., and ROBINSON, P.K., 1990, Mercury removal by immobilized algae in a batch culture system, Journal of Applied Phycology, 2, 223–229. WONG, P.K. and CHAN, K.-Y., 1990, Growth and value of Chlorella salina grown on highly saline sewage effluent, Agriculture, Ecosystems and Environment, 30, 235–250. WONG, K.H., CHAN, K.-Y. and NG, S.L., 1979, Cadmium uptake by the unicellular green alga Chlorella salina CU-1 from culture media with high salinity, Chemosphere, 11, 887– 891. XUE, H.-B. and SIGG, L., 1990, Binding of Cu(II) to algae in a metal buffer, Water Research, 24, 1129–1136.

37

3

General Bacterial Sorption Processes M.M.URRUTIA

Introduction The recognition of bacteria as avid adsorbents of a great variety of metals has opened a broad line of research in recent years for two main reasons: the importance of this type of process in natural environments, and the potential applicability in biotechnology. The use of bacteria as biosorbents is a fast growing field in remediation because of their small size, their ubiquity, their ability to grow under controlled conditions, and their resilience to a wide range of environmental situations. In this chapter, some basic principles underlying the ability of bacterial cells to adsorb or, more generically, ‘sorb’ metals, and the implications for biotechnological applications, are reviewed. The term ‘sorption’ covers several different processes including adsorption, binding or surface complexation, surface precipitation and/or mineral nucleation. The interaction of bacterial surfaces with soluble metals in the aqueous environments where microorganisms live is inevitable. Bacterial exchange of nutrients and wastes with the surrounding medium occurs through diffusion, both internally and externally. Bacteria have adopted shapes which provide them with greater surface area to volume ratios (i.e. rods versus spheres) so as to optimise diffusion. However, at the bacterial size range (~ 1 µm or less), viscous forces predominate and bacterial cells experience an enormous drag which prevents them from outswimming their local aqueous environment (Figure 3.1). Therefore, bacteria carry an envelope of water around their periphery so any substance or element in solution in their environment comes first into contact with the bacterial cell envelope.

Bacterial surface The ability of bacterial cells to bind metals is associated with the characteristic and indispensable cell envelope. This is formed by the external layers which separate the cell protoplast from the aqueous media, and represents the external limit of the bacterial cell. Its importance goes further than a role merely as a physical or chemical barrier. It maintains cellular shape, and participates in growth, division, and the 39


Figure 3.1 Diffusion processes in bacteria. Dimensions are in µm for both types of cell. Sa: v=surface area to volume ratio. The local environment outside the cells is due to the cells’ low Reynolds number. Modified from Beveridge (1988) with the permission of Canadian Journal of Microbiology

diffusion of nutrients and wastes. It also makes up a boundary through which the cells ‘relate’ to their environment. Several types of structures may form the cell envelope. The most important one is the cell wall, but capsules, S-layers and sheaths are commonly found superimposed on the wall (Beveridge, 1993; Thompson and Beveridge, 1993).

Bacterial cell walls There are essentially three different types of bacterial cell walls, the Archaeobacterial cell walls and two types of eubacterial cell walls (now denominated Archeae and Bacteria, respectively) (Woese et al., 1990). Eubacterial cell walls are defined by their response to the staining technique developed by Christian Gram in 1884 into Gram-positive and Gram-negative walls. The staining response to the Gram procedure is the result of fundamental differences in chemical composition and structural organisation (Salton, 1963; Beveridge, 1993). The Gram-positive wall is the simpler of the two, and is made up by a framework of peptidoglycan, which is a polymer of repeating dimers of acetylglucosamine and N-acetylmuramic acid. Attached to the acetyl-muramic acid residues are short peptide chains of usually four to five amino acids organised in sequences which vary among bacterial groups. This peptidoglycan core is usually between 20 and 40 layers thick, and adjacent glycan chains are cross40

General Bacterial Sorption Processes

linked through the aminoacid stems forming a highly resilient, three-dimensional macromolecule that surrounds the cells. One of several types of secondary polymer (teichoic acid, teichuronic acid, proteins or, sometimes, lipids) is also generally intermeshed with the peptidoglycan strands. Both the phosphoryl groups of the secondary polymers and the carboxyl groups of the peptide chains provide negatively charged sites in the Gram-positive wall (Beveridge and Murray, 1980). The cell surface does not have an even distribution of charge density; pole-cylinder junctions and septa show a greater charge density (Graham and Beveridge, 1994) indicative of regions of active hydrolysis/accretion processes during wall turnover and cell (and wall) growth and division (Graham and Beveridge, 1994). The Gram-negative wall is chemically and structurally more complex. Generally, it is a multilayered structure formed (from the inside to the outside) by an inner bilayered plasma membrane, external to which there is a thin peptidoglycan layer within the periplasmic space, and an outer membrane in the periphery. The periplasmic space is the translucent region seen under the electron microscope in conventional embedding preparations, and it houses the periplasm, a complex mixture of ions, enzymes, carbohydrates, wall precursors and peptidoglycan (Graham et al., 1991). The peptidoglycan is covalently bound to the outer membrane by lipoproteins. The outer membrane is the most characteristic feature of the Gram-negative wall; it is a bilayered structure like the plasma membrane and has an asymmetric lipid distribution. The outer leaflet of the outer membrane contains lipopolysaccharide (LPS), a characteristic component of eubacteria. There are also various types of proteins embedded within the outer membrane, some of which (porins) serve as channels for diffusion of hydrophilic solutes through the outer membrane (Ferris, 1989). The lipopolysaccharide has phosphoryl or carboxyl group-rich zones in its molecule, which, with the phosphoryl groups of the phospholipids, provide most of the negative charge of the Gram-negative walls (Hoyle and Beveridge, 1984). Experiments were conducted on isolated outer membrane and LPS from Escherichia coli K12 using 31P nuclear magnetic resonance, the paramagnetic metal ion europium, and manganese. This work conclusively proved that the phosphate groups within LPS and phospolipids were the primary sites for metal interaction (Ferris and Beveridge, 1984, 1986). Only one of the three carboxyl groups in the LPS is actually free to interact with metals (Ferris and Beveridge, 1986). Finally, archaeobacterial cell walls show a great diversity in structure and composition, probably in response to the diversity of environments where most of these organisms live, many of them under extreme conditions of pH, temperature or salinity. Four types of archaeobacterial wall have been defined (Konig, 1988): (i) those made of pseudomurein, a heteropolymer that resembles peptidoglycan (e.g. order Methanobacteriales); (ii) those made of methanochondroitin, a sulphated heteropolysaccharide similar to that found in animal connective tissue (e.g. Methanosarcina and Halococcus); (iii) those made of glycoprotein (S-layer) (e.g. Halobacterium); and (iv) those that completely lack a wall and are surrounded exclusively by the plasma membrane (e.g. Thermoplasma). Little information is available about archaeobacterial wall interactions with metal ions. However, studies with the sheath of Methanospirillum hungatei GP1 have suggested that selected surface sites have exposed carboxyl groups which provide anionic charge (Beveridge et al., 1988). Also, protein subunits of the S-layer in Halobacterium contain covalently bonded sulphate groups which are highly anionic (Wieland et al., 1980). 41


Capsules Capsules are the most common superficial layers above the bacterial cell wall. They are highly hydrated (>95% water), loosely arranged homopolymers or heteropolymers of carbohydrate and protein. These polymers are very flexible and extend up to several hundred nanometres from the cell surface, entirely bathed in water and possibly alternating between gel and liquid states (Geesey and Jang, 1989). Several different terms can be found in the literature to describe capsular materials. In general, exopolymers which are tightly linked to the cell surface are called capsules or extracellular polysaccharide (EPS), whereas those that slough off into the surrounding media are called slimes (Geesey and Jang, 1989). However, all these terms refer to substances of equal composition and properties. Capsules can consist of neutral polysaccharide (e.g. Alcaligenes faecalis), charged polysaccharide (e.g. E.coli K-30), or charged polypeptide (Bacillus licheniformis ATCC 9945A) (Beveridge, 1989a). Capsular exopolymers are usually acidic in nature, although their net electric charge varies in relation to the number of negative charges per polysaccharide subunit, and the number of capsular fibres per unit of cell surface area (Bayer and Bayer, 1994). Electron microscopic evidence suggests that the capsule fibres are arranged with their long axes almost perpendicular to the cell surface, although not all capsules exhibit identical structure (Graham et al., 1991). This arrangement is of great significance for metal binding because the negative charge of the capsulated cell reaches far beyond the charged double layer of the outer membrane of a non-capsulated cell (Bayer and Bayer, 1994).

S-layers S-layers are paracrystalline arrays of either protein or glycoprotein in p1, p2, p4 or p6 symmetry (Koval, 1988; Sleytr and Messner, 1988): hexagonal (p6) and tetragonal (p4) are most common. Non-covalent interactions, such as hydrogen bonding, electrostatic attraction, and salt-bridging, are involved in the attachment between neighbouring subunits and the underlying wall (Koval, 1988). Commonly, divalent metal cations contribute to the correct assembly of the structure (Koval, 1988; Sleytr and Messner, 1988). Metals can also be bound after assembly. For example, the Slayer of the unicellular cyanobacterium Synechococcus GL24 binds Ca2+ on the polar residues of the protein which surround the holes in the hexagonal array (SchultzeLam et al., 1992).

Sheaths These are a relatively rare structure in the bacterial world. Sheaths are composed of loose arrangements of homopolymers and heteropolymers which encase chains of cells and are in close contact with the underlying cell envelope components. They can appear as highly ordered proteinaceous structures with small proportions of lipid and carbohydrate (i.e. Archaeobacteria such as Methanothrix and Methanospirillum), or as fibrillar composites of acidic polysaccharides (i.e. bacteria such as Leptothrix or Sphaerotilus spp.). Probing of the sheath of Methanospirillum hungatei with cytochrome c showed that carboxylate groups were periodically exposed within the 42


holes between the protein subunits (Beveridge, 1989b). The sheaths of both M.hungatei and Methanothrix concilii show certain selectivity for metals, the former preferring Ca2+ and Mg2+ whereas the latter favours Zn2+ and Fe2+ (Patel et al., 1986). The sheaths of Leptothrix or Sphaerotilus participate in oxidation and precipitation of iron and manganese from solution (Ghiorse, 1984), and accumulate conspicuous amounts of these metals around the cell surfaces.

Biofilms The ubiquity of bacteria in nature reflects their great metabolic and physiological versatility. Because of their dependence on water, bacteria are particularly abundant and diverse in aquatic environments, in which they can exist in either planktonic (free-living) or attached form. Attached microorganisms generally develop into communities which form microbial films referred to as biofilms. Virtually all submerged substrata where bacteria are present are likely physical substrates for development of biofilms, from rocks and pebbles in river beds to submerged glass surfaces, medical devices or teeth (Costerton et al., 1987; Wimpenny et al., 1994). Natural biofilms are generally mixed communities of both prokaryotic (bacteria) and eukaryotic (algae) microorganisms, with bacteria being the predominant component in most instances (Costerton et al., 1987; Lewandowski et al., 1995; Flora et al., 1995a). Biofilms are bound together by polysaccharidic polymers which the cells produce when approaching surfaces (Turlear and Characklis, 1982). They are dynamic structures in which a continuous combination of growth and sloughing processes occurs (Santos and Callow, 1991). As such, they exhibit variable thicknesses and local cell and polymer densities (Stewart et al., 1995). Recent studies show that they are three-dimensional porous structures formed by basic units of bacterial clusters or microcolonies with water-filled interstitial voids; these voids form a network of channels which connect with each other and with the biofilm surface (Massol-Deya et al., 1995). The EPS component is very important in biofilm formation and critical for bacterial persistence and survival on a surface. The charge and physical properties of EPS provide biofilms with a great affinity for various types of molecules, ions and substances, including metals (Freeman and Lock, 1995). Several studies in natural environments have shown that biofilms are capable of binding significant quantities of metal ions under natural conditions, and they serve as templates for the precipitation of insoluble mineral phases (Brown et al., 1994; Southam et al., 1995). Biofilm metal binding probably plays a very significant role in metal ion budgets in natural ecosystems. However, the importance of this phenomenon in natural systems is largely unexplored and open to quantitative evaluation.

Charge of bacterial cell surfaces Electrokinetic studies of bacterial cell mobility in suspensions of known pH and ionic strength allow determination of cell surface charge density, which is an averaged charge over all the surface. These measurements, however, give no indication as to the location of specific charged groups (James, 1982). The nature and number of charged surface groups on the cells, and the nature of the suspension medium dictate 43


the charge carried by cells in suspension. If the ionic strength and pH of the suspension medium are kept constant, any changes in electrokinetic mobility are due to variations in the nature and/or number of surface charged groups. Studies of this type have shown that the surface charge of some bacterial species can vary with the age of the cells and with the nature and composition of the growth medium (James, 1982). For example, Klebsiella aerogenes exhibited slightly higher negative charge during logarithmic growth phase. The electrokinetic mobility of E.coli cells changes throughout the cell division cycle: new daughter cells are especially electro-negative and their electrophoretic mobility declines as the cells age (Gilbert et al., 1991). In contrast, no change in electrokinetic mobility with changes in growth rate was observed in chemostat populations of this bacterium (Gilbert et al., 1991). Charge is a function of capsular polysaccharide production during the growth cycle in Streptococcus pyogenes (Gilbert et al., 1991) and Zoogloea ramigera (Norberg and Enfors, 1982). In the latter organism, increased metal binding results from increased EPS production (Norberg and Persson, 1984). On the contrary, Gram-positive organisms such as Bacillus subtilis and Micrococcus luteus do not show significant changes in surface charge during the cell growth cycle, at least as indicated by their metal-binding capacity (Cotoras et al., 1992). However, growth medium composition affects the surface charge of B. subtilis cells, which form teichoic acids when grown in the presence of phosphate and magnesium (Beveridge and Murray, 1980). The phosphate groups in this polymer ionise above pH 2.5, so that a teichoic acid of 40 repeating units would contain 40 negatively charged phosphates per molecule (Hancock, 1991); evidently these phosphate groups provide the walls with further negative charge (Beveridge and Murray, 1980). Bacterial cell charge is strongly affected by pH, because the different ionogenic groups at the cell surface are susceptible to protonation/deprotonation reactions. James (1982) showed that the response of a bacterial surface to changes in the pH of the suspension media reflects the nature of its surface charges. Experimental data on different types of bacteria corresponded to four different pH response types. Type 1 showed a linear increase in negative charge with increased pH; Type 2 showed an exponential increase in negative charge with increasing pH; in Type 3 positive charges predominated from acidic to neutral pH, but the surface rapidly changed to a negative charge at alkaline pH; Type 4 showed positive charge at very acidic pH, which changed linearly to negative as pH increased to 5, above which the negative charge stabilised to an almost constant value up to pH 8, followed by a sharp increase in negativity at pH>9. Type 1 behaviour represents the adsorption of anions and/or desorption of protons from an uncharged surface with increased pH; Type 2 reflects an anionically charged surface with carboxyl, sulphate or phosphate groups (e.g. K. aerogenes, E.coli). This behaviour agrees with that expected for dissociation of protons at a surface-electrolyte interface, e.g. carboxyl deprotonation at pH values above the pKa value of the groups. Type 3 typifies an amino-group dominated surface, whereas Type 4 reflects mixed amino-carboxyl surfaces such as those of B. cereus, B. megaterium, B. subtilis, Micrococcus lysodeikticus or Streptococcus pyogenes. Experimental evidence shows that carboxyl groups are present in excess over amino groups in these bacterial surfaces, and thus they dominate electrokinetic behaviour. Collectively, these studies illustrate an important influence of environmental pH on the charge properties of the bacterial surface, which, as a consequence, will greatly affect their metal sorption capacity. pH will also affect other interactions in which charge is involved, such as adhesion of bacteria to 44


colloidal components in water columns or to fine-grained mineral particles and/or organic matter. Surface charge is not exclusively determined by external factors or growth cycles. Internal factors also contribute to the overall charge, and provide bacterial surfaces with intrinsic characteristics which differ from the behaviour expected, based on the individual components of the cell envelope (Hancock, 1991). Most important are the interactions established between charged groups within the wall macromolecules, which depend on location and packing densities. This effect may occur in two ways. On the one hand, intramolecular charge repulsion has a major effect in the conformation of polyelectrolytes, e.g. teichoic acids or capsular polysaccharides such as streptococcal hyaluronic acid. On the other hand, intramolecular charge neutralisation is also possible. The positively charged amino groups of esterified alanine residues in teichoic acids may partially neutralise adjacent negatively charged phosphodiester groups in the same molecule. Furthermore, separate polymeric components may interact with one another, e.g. phosphate groups in teichoic acids interact directly with amino groups in proteins and peptidoglycan in the wall (Hancock, 1991). Therefore, the charge properties of polymers integrated into the cell wall may be appreciably different from those of the isolated molecules. This complication makes estimation of the overall charge properties based on characterisation of individual components problematic and, therefore, methods which measure total charge (such as those explained above) are probably the most valid approach to determining bacterial surface charge. Intrinsic charge differences within the cell wall may also result from the presence of cell appendages (e.g. flagella, fibrils, fimbriae or conjugative pili). Such appendages can alter physical properties of the surface since they penetrate the cell wall and thus may perturb wall organisation (Hancock, 1991). In Gram-negative bacteria perturbation of the cell wall could be caused by exposure of outer membrane hydrophobic groups; in the Gram-positive B. subtilis evidence suggests that some reduction of peptidoglycan cross-linking may be necessary at sites of flagellar penetration of the thick wall (Hancock, 1991), which again may enhance the local charge at those points. In a similar way, fimbriated cells of E. coli grown on nutrient agar showed increased electrophoretic mobility as they sequentially lost the fimbriae when grown in broth (James, 1982). A bacterium could also regulate cation-wall interactions to a certain extent through alteration of cell wall charge characteristics. Although lacking clear experimental evidence, Doyle (1989) has suggested several mechanisms which a bacterium could use for this purpose. These included the ratio of ammonium to carboxylate or phosphate groups in the wall, N-acetylation, the presence of D-alanine, and the degree of cross-linking between peptidoglycan chains. Results with chemically modified walls showed that the higher the ratio of ammonium to carboxylate or phosphate, the lower the extent of complex formation with cations (Doyle et al., 1980). This suggests that it is possible for cells to modulate surface charge and metal binding by regulating the insertion of ammonium groups. The possible control of surface charge through N-acetylation would occur during wall precursor synthesis because the glucosamine residue of the peptidoglycan is approximately 71% Nacetylated and the muramic acid is 67% N-acetylated; a small change in the degree of N-substitution could result in a major change in metal binding to the wall (Doyle, 1989). When D-alanine is esterified to the teichoic acid in Gram-positive bacteria, an ammonium group is introduced into the chain, which suggests a possible mechanism by which D-alanine could contribute to a decrease in metal ion binding by bacteria. In 45


support of this idea, cells grown in high sodium chloride concentrations have walls with reduced D-alanine contents in their teichoic acids, possibly to enhance Mg2+ binding by the wall. Finally, increased cross-linking between diaminopimelic acid and a terminal D-alanine results in a loss of both a positive and a negative charge; the loss of a positive charge may preferentially enhance metal binding (Doyle, 1989). To date, there are no studies of metal binding to bacterial walls of equal chemotype and variable degrees of cross-linking, or with variable N-acetylation in metal-sufficient and metal-limiting media which could confirm Doyle’s proposed mechanisms. Another important aspect of bacterial wall charge is its asymmetric distribution. Although all portions of the cell wall contain anionic groups (Beveridge and Murray, 1976; Beveridge et al., 1982, 1995), there is evidence to suggest that the external face of the wall, directed toward the external environment, has a greater density of charge than the inner face, directed toward the plasma membrane (Umeda et al., 1987; Graham and Beveridge, 1994). Furthermore, certain topographic areas within the wall appear to have a greater charge density. In particular, cell pole-cylinder junctions and areas next to septa show a greater density of negative charge (Graham and Beveridge, 1994). In B. subtilis these areas are related to fast turnover and/or accretion of components within the expanding wall (Graham and Beveridge, 1994). Recent work using small-angle X-ray scattering (SAXS) and transmission electron microscopy (TEM) showed that certain metals (Ga and U) were spread over the entire Pseudomonas fluorescens cell envelope, whereas metals such as Pt were associated only with the lipopolysaccharide-rich external face of the outer membrane (Krueger et al., 1993). These data show that for some metals the entire wall matrix can participate in metal binding, whereas for other metals the outer leaflet of the wall will be the most reactive area for metal sorption. In general terms, the surface of bacterial cells behaves like any other particle submitted to a boundary between two phases in that an electrical double layer is established at the bacterial cell surface. However, the bacterial wall represents an active, flexible and dynamic boundary, which may respond to the changes in environmental conditions in ways in which inorganic particles cannot (Bayer and Bayer, 1994).

Sorption of metal cations and mechanisms Much information has been collected in the past 20 years on the binding or sorption of metals by bacterial surfaces. A mechanistic analysis of this process will be presented together with a compilation of relevant quantitative data from the literature. Starting with the work of Marquis and collaborators (1976) it became apparent that bacterial walls are flexible and dynamic cell components whose composition reflects that of the external environment rather than that of the cytoplasm. In the words of these authors, ‘Cell walls are polyelectrolytes and so they are always accompanied by counterions that are required to maintain electroneutrality’. It is difficult to express more clearly the reason for the interaction between bacterial surfaces and metal ions. Many studies conducted on metal sorption by bacteria have shown that isolated walls from Gram-positive organisms bind quantitatively more metal than isolated envelopes of non-capsulated Gram-negative bacteria. Typically 10 times more metal is bound to Gram-positive walls than to Gram-negative envelopes (Beveridge and 46


Fyfe, 1985). However, working with whole cells, Mullen and coworkers (1989) did not find significant differences in the ability of two Gram-negative and two Grampositive bacteria to bind Cd, Cu, Ag and La. These authors claimed that intact cells are much more chemically complex than purified walls or envelopes, and thus may show differences in metal-binding capacity, particularly in the case of the Gramnegative variety. An explanation may lay in the fact that gamma-irradiated or chemically uncoupled cells bind more metal than active metabolising cells with energised plasma membranes (Urrutia et al., 1992), because of the competition between the H+ generated by an energised membrane (in metabolising cells) and the metal cations in solution (Figure 3.2). In a similar way, isolated walls, obtained after several chemical treatments, are probably free of their own counterions (e.g. protons) and may subsequently bind more metal than intact walls of live cells. Greater metal binding by dead cells in relation to their living counterparts has been observed experimentally on other occasions (Doyle et al., 1980; Kong et al., 1993a, 1993b). Detailed studies of the metal-binding capacity of the outer membrane (OM) of the Gram-negative E. coli showed that the OM exhibited certain selectivity which depended on the metal suitability for various physicochemical roles. For instance, Ca2+ stabilised the LPS and Mg2+ formed part of complexes in LPS and protein (Ferris and Beveridge, 1986). The interactions between soluble metallic ions and the OM followed three physical principles which regulate cation selectivity of biological membranes: (i) binding depended on the free energy difference between the site

Figure 3.2 Influence of a protonated cell wall on metal binding by Bacillus subtilis. PM=plasma membrane; PG=peptidoglycan layer. For more details, refer to Urrutia et al. (1992)

47


bound cation and cation-water interactions (Na/K binding); (ii) the free energy of interaction originated from electrostatic forces; (iii) the principal electrostatic forces were Coulomb forces. Ferris and Beveridge (1986) also concluded that the salt condition of the OM would partially determine how its surface interacts with its aqueous environment in terms of the diffuse-double layer theory. Apparently, metal binding occurs through a passive mechanism which involves electrostatic interaction between the negatively charged groups in the wall and the metallic cations. Most metal binding occurs after initial metal complexation and neutralisation of the chemically active site (Rudd et al., 1984). Binding to the cell walls might proceed through at least a two-step mechanism, as postulated by Beveridge and Murray (1980) for Bacillus subtilis: the first step is the stoichiometric interaction of metal with reactive chemical groups, followed by a second stage in which those same sites nucleate the deposition of more metal as a chemical precipitate, which results in the development of fine-grained minerals. The initial interaction is the ‘adsorption’ phase, but metal retention ability by bacterial walls goes further than their adsorption capacity, since bacterial surfaces are favourable interfaces for mineral nucleation. The latter stems from lowering of the activation energy barrier which normally retards spontaneous crystal nucleation and growth, with the result that microbial cells are particularly efficient agents of heterogeneous mineral nucleation (Schultze-Lam et al., 1992). This effect is possibly quite pronounced in metastable solutions which are at saturation or oversaturated with respect to certain minerals, i.e. carbonates. Some metals have a greater tendency to form precipitates than others. In systems which include living cells it is possible that some active uptake or nonspecific cation transport system could take place. While the overall metal removal process can be considered as a ‘sorption’ process, metals may in fact be retained by one or more of the above mechanisms. This complication may contribute to the frequent inability of adsorption isotherms to describe experimental data. Adsorption equations may be useful for describing bacterium-metal interactions with certain metals (such as Cd and Cu), but this approach may not be adequate when precipitation of metals occurs (Mullen et al., 1989). This is an important point which will later be discussed further, since adsorption equations are the most commonly used method to describe these processes, to both organic and inorganic adsorbents. When EPS is the binding surface, mechanisms of metal binding may be different from those mentioned for cell walls. A good compilation of data of metal binding by EPS is presented by Geesey and Jang (1990). These authors show that the most effective electron donor group in acidic polysaccharides is the carboxyl residue. Lone-pair electrons on carboxyl groups interact with the charge-compensating metal ions. However, oxygen atoms associated with the ether bond and hydroxyl residues on the sugar subunits are also weak electron donors which may also contribute to binding in both acidic and neutral polysaccharides (Martell, 1971). In uncharged polysaccharides, metal binding occurs because of coordination between the metal cation and oxyanion and hydroxyl groups on the donor molecule. Therefore, the affinity of uncharged polysaccharides for metal ions generally decreases with increasing ionic radius (Rendleman, 1978). Metal interactions with charged polysaccharides depend on charge density. Experimental evidence supports the concept of ‘salt bridging’ as the binding mechanism in this case, in which a polyvalent metal ion bonds to two anionic groups on separate polymer chains. Steric fit thus becomes important and is reflected by the fact that the ionic radius of a metal 48


frequently decides whether complexation to EPS will occur (Steiner et al., 1976; Manzini et al., 1984). In general, carboxylated polysaccharides exhibit preferential binding of cations with large ionic radii, i.e. they prefer transition metals to alkaline earth metals. Although metal-binding affinities by different types of EPS are complex and show intra-metal and intra-specific differences, these studies suggest that acidic capsular and slime polysaccharides bind metal ions through the oxygen atoms of carboxyl and hydroxyl groups on uronic acid subunits. Oxygen atoms of hydroxyl groups on neighbouring neutral sugars also contribute to the coordinate binding of metal ions, which promotes the formation of stable complexes (Geesey and Jang, 1990). The presence of the metal salt-bridging may induce a conformational change within the capsule such that its physical and chemical properties may be altered, i.e. other metal-binding sites may become inaccessible (Geesey and Jang, 1990). Therefore, although the binding sites in walls and capsular polymers are chemically identical, the metal-binding characteristics of capsule and wall are different, due to the conformation of those groups in the two types of polymer. Differences in metal-binding characteristics of the EPS and the cell wall were found for Bacillus licheniformis capsule (McLean et al., 1990). In the peptidoglycan, anionic carboxylate residues are held in an almost rigid orientation due to the turgor of the wall (in live cells) and to the extensive cross-linking between peptidoglycan strands (Beveridge, 1989a). Capsule polymers, on the contrary, are more flexible since they are not cross-linked by covalent bonds, they are highly hydrated, they have greater charge density per unit area, and presumably they are better able to adapt to the coordination sphere of a particular metal though folding into a fit-binding conformation. These differences would benefit some organisms, since some metals are necessary within the cell wall, whereas others (such as Cr3+ or other heavy metals) can be toxic to the cells (McLean et al., 1990). Binding of metals by isolated EPS is more easily described by adsorption isotherms than metal binding by walls or whole cells, which further suggests that there are differences in the mechanisms of metal binding by walls and EPS. Freundlich isotherms adequately describe the binding of uranium to the EPS of Pseudomonas sp., showing that the process occurs as single layer adsorption. Here, the acyl residues were responsible for approximately 50% of the uranium uptake (Marques et al., 1990). Working with B. subtilis, Doyle and coworkers (1980) showed that Na+, Ca 2+, Mn2+ , Ni+, Sr2+, Zn 2+ and Mg 2+ all bind to the same sites on the walls and that a direct competition for sites occurs when two or more metals are present. Their data suggest that, when the negative sites are close to saturation, the phenomenon of negative cooperation takes place, i.e. occupation of one site may decrease binding of metals to other sites nearby. Competition for the same sorption sites was also found for binding of Ni and Zn during growth of the filamentous bacterium Thiothrix strain Al. The primary mechanism of metal sorption by Thiothrix was apparently ion exchange, because 66% to 75% of the Ni or Zn bound could be desorbed by Ca. A biphasic Freundlich isotherm described adsorption of Ni and Zn, which indicates that there are at least two types of binding site for these metals on Thiotrix surfaces. Binding increased with cell age but was not proportional to the amount of protein (Shuttleworth and Unz, 1993), so the increase was not due simply to an increase in biomass. Work with Mycobacterium smegmatis showed that there was intra-metal competition for the same binding sites (Andres et al., 1993). Scatchard model plots showed that there were at least two types of cation complexation sites (strong and weak) for several actinide and lanthanide cations on 49


the surface of M. smegmatis (Andres et al., 1993). Negative cooperation phenomena, i.e. bonding priority of one type of site with respect to another, were also observed. Adsorption data fitted BET (Brunauer-Emmett-Teller) isotherms, which assume a multi-layer adsorption process in which one layer need not necessarily be completely filled before another is commenced; the authors suggested that cation adsorption occurred in this bacterium through exchange reactions, since partial release of magnesium from the cell wall accompanied the adsorption of the test metals (Andres et al., 1993). Binding was independent of bacterial viability (Andres et al., 1993), a result which contradicts other experimental data that show greater binding ability by dead or uncoupled cells than by live ones (Doyle et al., 1980; Urrutia et al., 1992; Kong et al., 1993a, 1993b). Studies conducted by Falla and Block (1993) showed clearly that at least two types of binding sites are present in the envelopes of Pseudomonas fluorescens. These two sites showed different affinities for Cd2+, Ni +, Cu2+ and Zn2+. One group is a high-affinity binding site and the other a low-affinity site. The Ka values for the high-affinity binding site show that Cd has the highest affinity for the outer membrane, Ni and Cu for peptidoglycan, and Zn for the cytoplasmic membrane. The Ka values for the low-affinity binding sites suggested that Zn had the highest affinity for the outer membrane, Cd and Cu for peptidoglycan and Ni for the cytoplasmic membrane. The authors concluded that carboxylic groups on the cell envelope of this bacterium play an important role in binding of Ni, Cu and Zn but not of Cd (Falla and Block, 1993). In other studies, voltammetric determination of Cu, Pb, Zn and Ca binding by chemically-killed cells of Klebsiella pneumoniae showed that metal ions bind through surface complexation to the highest affinity surface ligand first (Simoes Goncalves et al., 1987). Subsequently, binding occurs on those groups with lesser reactivity. Metal adsorption decreased at increasing metal loading of the bacterial surfaces (Simoes Goncalves et al., 1987). Two different sites for metal binding were also found on Rhizopus arrhizus biomass at subsaturating ion concentrations (Brady and Tobin, 1995). In this case, there was an excellent correlation between the potential of a metal ion to form covalent bonds (described by the covalent index of Nieboer and Richardson (1980)) and the metal uptake affinity by Rhizopus biomass. This finding agrees with the hard and soft principle of metals. For example, some metals such as Sr 2+ exhibit only ionic binding, in agreement with the hard/soft acid classification and with the experimental results of binding to the Rhizopus biomass. This work suggests that metals form inner-sphere complexes with the bacterial surface sites. This intriguing idea is worthy of further consideration and research. The important conclusion to be drawn from the experimental data described above is that there is a great deal of heterogeneity among different bacterial species in relation to (1) number of metal-binding sites on the bacterial surfaces, (2) binding strength in relation to different cations and (3) binding mechanisms involved, such as electrostatic interactions, covalent bonding, ion exchange, or surface complexation. This heterogeneity reflects diversity of species studied and environmental conditions in which the experiments were conducted, and thereby points to the need for careful consideration of experimental set-up used to assess bacterial metal-binding mechanisms. In general, it appears that the number of binding sites, strength of binding and binding mechanisms by bacteria are not amenable to generalisation among different types of bacteria. In contrast to the above conclusion, it should be mentioned that some authors 50


believe that well-established mechanisms of adsorption of metals to inorganic particles and colloids can be extended to cellular surfaces (Collins and Stotzky, 1992). The reason is that the speciation form of the metal, rather than the type of cell or particle, may be the factor which decides adsorption. They base their conclusion on studies showing that the adsorption of hydrolysable metal ions on microbial cells may occur through a similar mechanism to that for the adsorption of these metal ions to mineral oxide surfaces. However, certain intrinsic characteristics of biological entities, such as their ability to adapt to environmental changes, their metabolical pathways, and their ability to grow and divide, already add major complications to Collins and Stotzky’s point of view. However, biological entities are, of course, subject to the same laws of physics and chemistry as inorganic particles, and therefore it is possible that application of some or maybe much of the knowledge accumulated by studying metal adsorption to inorganic sorbents could enhance our understanding of metal binding to cell surfaces. Let us now examine the controversy regarding the use of adsorption isotherms to describe metal-binding processes to bacterial surfaces. Some authors think that adsorption equations may accurately describe certain bacterium-metal interactions (such as those of Cd or Cu), but stress that this approach may not be adequate when precipitation of metals occurs (Mullen et al., 1989). Others, however, maintain that the Langmuir equation explains their data adequately (Cotoras et al., 1992) because the Langmuir model can describe surface precipitation reactions as well as adsorption reactions under certain experimental conditions (Veith and Sposito, 1977). From the author’s own experience, binding of Fe 2+ by the bacterium Shewanella alga strain BrY, an EPS-forming bacterium, could be described only by Freundlich-type isotherms, which suggest a greater degree of heterogeneity in binding sites (Urrutia, unpublished data). In view of the previous discussion of bacterial surfaces and, particularly, EPS charge and binding properties, heterogeneity in metal binding by bacteria would certainly be expected. Part of the difficulty is that at high surface coverage it is difficult to distinguish between surface complex formation and precipitation. However, models which consider a continuum between surface complex formation and bulk solution precipitation of the sorbing ion may conceptually explain this behaviour (Stumm et al., 1994). Recent modelling approaches allow better a description of adsorption to heterogeneous systems (Hinz et al., 1994; Pedit and Miller, 1994). As the ability to describe the adsorptive behaviour of heterogeneous surfaces advances, it should be possible for microbiologists and geochemists to use these techniques to model biosorption of metals. Such improvement in our ability to quantitatively describe metal binding to bacterial surfaces could allow a better understanding and prediction of their behaviour as metal sorbents. This is important if the role of bacterial surfaces on global metal budgets is to be accurately assessed. This obviously represents a complex but very interesting line for future research in the field of metal binding by microorganisms.

Sorption of metal anions and mechanisms There have not been many studies of anionic metal species binding to bacterial surfaces. One example is Cr(VI), which forms highly soluble (and toxic) chromates (i.e. ) in aqueous solution. Chromates are relatively ubiquitous 51


contaminants due to human activity. Most of the work conducted with chromates and microorganisms seeks the reduction of Cr(VI) to the less soluble and then less toxic Cr(III). Several studies have centred on adsorption of chromate to inorganic adsorbents, such as Fe(III) or Al oxides (Hsia et al., 1992, 1993; Fendorf, 1995), but few have examined chromate adsorption to bacteria. Kong and coworkers (1993a, 1993b, 1994) have studied the adsorption of chromate to starved and ‘fresh’ cells from several groundwater bacterial isolates of Pseudomonas stutzeri. These authors found sorption of chromate to the bacterial cell wall and internal uptake of the anion. Adsorption to the cell wall increased after a 50-day starvation period in the absence of other anions such as sulphate, which effectively outcompeted chromate in the adsorption process (Kong et al., 1993a, 1993b). The authors explain this increase in Cr binding by the cell wall during the longer starvation period in terms of possible reduction of the chromate to the chromic ion by the cells, since chromate may replace nitrate as a substitute electron acceptor under long starvation conditions for this bacterium (Kong et al., 1994). An energydependent intracellular transport of the Cr was also observed, since Cr was found in the membrane and ribosome cellular fractions. Intracellular Cr transport drastically decreased (65%) in killed cells experiments. Unfortunately, there are no details in Kong et al. (1994) about the redox state of the intracellular Cr or the mechanism through which the metal may have entered the killed cells. Cell binding of non-metallic anions, such as silicates (Urrutia and Beveridge, 1993a, 1995), phosphates and sulphides (Beveridge et al., 1983), or carbonates (Thompson and Ferris, 1990; Schultze-Lam et al., 1992) has also been studied. Silicate binding to the walls of B. subtilis occurred preferentially through the formation of ternary wall-metal cation-silicate complexes, although it might also result from direct electrostatic interaction with native residues in the wall (Urrutia and Beveridge, 1993a). Fine-grained carbonates in the S-layer of a unicellular cyanobacterium, Synechococcus, formed by the interaction of soluble carbonate anions with Ca2+ cations bound to the S-layer protein units (Schultze-Lam et al., 1992). It is possible that binding of other anions to the cell walls could occur through similar processes, including metallic anions such as chromate, arsenite or arsenate, without requirement for reduction processes previous to binding. Reduction, however, could occur quite easily with metals with lower reduction potentials; for example, Au3+ was reduced to Au0 by the wall components (possibly through nonenzymatic mechanisms) before or simultaneously with binding to wall structures (Beveridge and Murray, 1976). Many questions about the mechanisms of anion sorption to bacterial cell surfaces remain unanswered.

Binding constants Tables 3.1 and 3.2 compile some current data and classic papers on stability constants and maximum sorption capacities (respectively) for the binding of metals to different bacterial cell components. Quantitative information on metal binding to cell walls or whole cells is quite scarce; much more detailed work has been done with exopolysaccharidic components and siderophores. Table 3.3 presents stability constants for metallic cation complexes with organic compounds such as humic acids, fulvic acids, nitriloacetic acid (NTA) or ethylenediamine tetraacetic acid (EDTA). There is great scatter of data, in part 52

General Bacterial Sorption Processes Table 3.1 Stability constants (Ki) for metal binding by bacteria or bacterial components

due to the innate variability of natural organic matter, which results in its polyfunctional character and its poly-electrolytic properties (Buffle et al., 1984). This scatter may also arise because different studies use different analytical and computational approaches, a problem which several authors have discussed extensively (Buffle et al., 1984; Kramer and Duinker, 1984; Buffle, 1990). Despite these limitations, it appears that binding constants for heavy metal binding to bacterial components are close to, and sometimes greater than, those determined for metal complexation by organic ligands, particularly natural fulvic or humic acids (Table 3.1 versus Table 3.3). In contrast, bacteria show considerably greater affinities for metals than inorganic adsorbents such as Fe oxides, Al oxides, clays or calcite (Table 3.1 versus Table 3.4). For instance, EPS from Klebsiella aerogenes, Xantomonas campestris or a freshwater sediment bacterium will effectively compete for Cu with fulvic and humic acids from soils at pH values around 6 to 8 (log Ki values for EPS were between 8 and 8.75 and for the humic and fulvic materials between 1.2 and 8.8, Tables 3.1 and 3.3). Similarly, the affinity of EPS for Ni, Cd or Co is greater than that of natural fulvic 53

Biosorbents for Metal lons Table 3.2 Maximum complexation capacities [CC] (mol/g dry weight) of bacteria or bacterial components

acids from soil or peat, and equals the affinity of humic acids for Cd (Tables 3.1 and 3.3). These comparisons suggest that bacterial exopolymers may be able to compete with natural humic and fulvic acids for metals such as Cu, Ni, Cd or Co under natural conditions. Particularly interesting are the affinity constants which have been defined for P. putida siderophore (Chen et al., 1994), since they equal or exceed those for EDTA and NT A (Peters and Shem, 1995). Specifically, the siderophore shows greater affinity than EDTA or NTA for Cu, Mn 2+ and Fe 3+ , and similar affinity to EDTA for Zn. When comparing metal complexation data for natural organics with data obtained 54

Table 3.3 Binding constants for some other common organic complexing agents (Site 1=carboxylic or strong binding site; Site 2=phenolic or weak binding site)

Biosorbents for Metal lons Table 3.4 Binding constants for some common inorganic adsorbents (ionic strength=0.1 M; pH=7.0 to 9.5)

* Ionic strength=0.01 M.

with purified organic materials, it is commonly observed that the K values obtained in laboratory studies are often lower than those obtained in situ (Buffle, 1990). This may result from the higher metal/FA ratios used in laboratory studies, but it may also be due to the presence in natural materials of bacterial siderophores, peptides, pigments, amino acids and nucleic bases with greater binding strengths (Buffle, 1990). The influence of microbial biomass in speciation of metallic cations might be very important because of the generally high affinity of bacterial surfaces for metals and because of the considerable contribution of bacterial biomass to total biomass in natural systems. Maximum complexation capacities of several bacterial species or cell components for different metals (Table 3.2) are also of interest in this discussion. The highest complexation capacities reported in the literature are, in general, for Cu; values range between 5 mmol/g dry weight for Zoogloea ramigera 115 EPS (Norberg and Persson, 56


1984) to 0.8 mmol/g for Pseudomonas aeruginosa (Philip et al., 1995) to 4 µmol/g for K. aerogenes capsule (Rudd et al., 1984). For other metals, including , Fe2+, 3+ Fe , Cd, Co and Ni, maximum sorption capacities ranged from 6 µmol/g for Co binding by K. aerogenes capsule (Rudd et al., 1984) to 0.7 mmol/g for binding by B. subtilis cells (Cotoras et al., 1992). Comparison of the Fe2+ binding ability of the bacterium Shewanella alga strain BrY with that of a synthetic goethite (crystalline Fe(III) oxide) showed that they were of comparable magnitude at neutral pH (0.15 mmol/g for BrY versus 0.25 mmol/g for goethite) (Urrutia, unpublished data). The data in Table 3.4 are remarkably consistent, even though many of these results were obtained at different pH ranges. From these data it appears that maximum complexation capacities by bacterial cells and/or components might be up to several mmol metal per gram (dry weight) of bacterial mass. These results indicate that bacterial components might be quantitatively important contributors to metal binding in sediments and soils. Another important aspect of metal binding to bacterial biomass is the strength with which bacterial components retain the metals. Several studies have shown that metals bound to bacterial walls are quite resilient to remobilisation (Flemming et al., 1990; Urrutia and Beveridge, 1993b); sometimes, as with Pb, even to mobilisation by EDTA (Urrutia and Beveridge, 1993b). This property increases the potential for long-term immobilisation of metals once they are sorbed to bacterial structures, which in natural environments may often be a favourable circumstance. This might decrease or alleviate metal toxicity. However, strength of binding causes difficulty in the recovery of the metal and the possibility for reutilisation of bacterial adsorptive matrixes in biotechnological applications. Falla et al. (1995) were able to remove 90% of the Cd, Cu and Zn accumulated by Pseudomonas fluorescens if the metal was provided after cell growth had ceased, only 50% of the metal accumulated during growth and just 10% of the metal from activated sludges. Other techniques such as elutriation, acidification, ion exchange, or complexation were not capable of removing all the accumulated metal from the biomass either (Alibhai et al., 1985; Legret et al., 1987). If, in fact, some metals form inner-sphere complexes with bacterial surface sites (as discussed in a previous section), we could speculate that these may account for the fraction of metal which is not removable from the surface of the cells. In certain cases, the binding of metals by exopolymeric substances may enhance metal mobility. Chen et al. (1995) found that bacterial extracellular polymers from Acinetobacter calcaceticus and several Arthrobacter, Pseudomonas, and Zoogloea species could release Cd and Pb adsorbed to aquifer sand in batch experiments. However, in most cases, exopolymers will immobilise the metals (Southam et al., 1995). For example, EPS from Klebsiella aerogenes retained Cu and Cd within the biomass due to metal adsorption to floc-associated polymer (Rudd et al., 1984; Corzo et al., 1994). Addition of Cu, Al, Cr(III) or Fe(III) to Bacillus licheniformis ATCC 9945 capsule induced flocculation (McLean et al., 1990), and cyanobacteria in a mixed microbial mat produced metal-binding extracellular polysaccharides which completely flocculated added metals (Bender et al., 1994).

Modelling Different approaches to modelling the influence of bacteria on metal speciation have been taken in recent years, particularly in relation to the role of biofilms in 57


bio-reactors. Chen and coworkers (1995) modelled the effect of a bacterial polymer on trace metal reactions with an inorganic surface. They used a three-component retardation model based on experimental data for Cd and Pb binding by extracellular polymers from 13 bacterial isolates. Hsieh and coworkers (1994a, 1994b) developed a mechanistic model of growth of a biofilm-forming bacterium which adheres to a substratum through production of extracellular polymers. The model describes the dynamics of cell growth and polymer production, all of which may influence trace metal behaviour. Each bacterial cell was modelled as a twocomponent structure formed by active cell mass and biopolymer. The biopolymer component was further divided into cell-associated and soluble (Hsieh et al., 1994a). Subsequently, the model was used to predict the behaviour of transient and steady-state biofilm systems in the presence of Pb (Hsieh et al., 1994b). To the author’s knowledge, this represents the most comprehensive approach to modelling the influence of biofilms on metal binding and speciation. Other published modelling approaches have focused on understanding the development of the biofilm under different conditions relevant to bioremediation processes, e.g. in an aerobic, constant flow reactor (Wanner et al., 1995), in anaerobic systems (Flora et al., 1995b, 1995c) or in anaerobic granules within upflow sludge beds (Arcand et al., 1994). Such models are important for quantitative evaluation of bacterial biomass contribution to metal sorption in natural and biotechnological systems, and should provide useful tools for future evaluation of biofilms’ contribution to metal cycling in the environment.

Applications in biotechnology Chapters 4–6 will examine the application of microbes in general and bacteria in particular to biotechnology. Therefore, only a brief account is offered here of some remarkable technologies which use bacteria as biosorbents for metal ions. Although metal binding by bacterial surfaces is a major mechanism involved in metal removal in bioremediation applications, it is not the only one. Bacteria have other capabilities to remove soluble metals. Brierley et al. (1989) and Brierley (1990) summarise these as follows: extracellular precipitation reactions, intracellular accumulation, oxidation and reduction reactions, methylation and demethylation, and, as we have seen, extracellular complexation and surface binding. Examples of some of these alternative processes are the precipitation of metal sulphides by sulphate-reducing bacteria, bacterial reduction of chromium, and methylation of elements such as Hg or Se. Metal removal can be accomplished by using either living or non-living biological systems. Living systems can be employed by using natural ecosystems (e.g. wetlands) or man-made ponds, bogs or marshes as purifying agents in which bacterial communities develop while treating the waste. A review of some possibilities in this field is given by Brierley et al. (1989). A good example of the diverse capabilities of living bacterial communities in bioremediation is illustrated by a mixed microbial mat system developed by Bender and coworkers. Cyanobacteria, purple autotrophic bacteria, mixed populations of heterotrophs such as Pseudomonas, and sulphur-reducing bacteria primarily form this mat (Bender et al., 1989, 1994), whose relative composition changes in response to the environment (i.e. composition of the waste water). Such mats tolerate low 58


temperatures and effectively remove all the heavy metals they have been presented with, including Pb, Cd, Cu, Zn, Co, Cr, Fe, Mn and U 238 (Bender et al., 1995). Another approach is the use of immobilised bacterial biomass or exopolymeric material as biosorbent, many examples of which can be found in the literature (Brierley et al., 1989; Brierley, 1990; Mellor et al., 1992). There are also some more specific applications of bacterial structures in biotechnology, such as the use of S-layers for industrial purposes. S-layers can be used as industrial filters with very fine threshold exclusions (Sara and Sleytr, 1987), as immobilisation matrices for enzymes or antigens (Sara et al., 1993) or as biosensors (Pum et al., 1992). It is clear that the capabilities of bacterial structures of different kinds and of the great diversity of bacterial metabolism will play a definitive role in future endeavours in biotechnology.

Summary The intrinsic composition and structural organisation of the cell envelope (in all its variations, such as cell walls, capsules, S-layers and sheaths) provide bacterial cell surfaces with a high density of negative charge, and result in great metal-binding capacity. Cell surfaces also provide favourable interfaces for mineral formation by facilitating heterogeneous nucleation processes. The few quantitative studies that have been made to date, show that bacterial components have metal-binding affinities and capacities which are comparable to those of natural humic and fulvic acids and are greater than various inorganic solid-phase sorbents. Therefore, bacterial components are likely to be important competitive matrixes for metal binding in the natural environment, particularly exopolysaccharidic materials which form capsules or slimes. In certain cases (i.e. siderophores) metal affinity is remarkably high, with affinity constants which approach those of synthetic organic chelators. Application of bacterial metal binding to biotechnological processes, particularly bioremediation, has been in place for several years. However, the quantitative role of bacterial communities as metal-binding matrixes in natural environments still awaits detailed consideration. Thus, there now exists a challenge to integrate the fields of microbiology, geochemistry and modelling in order to quantify the effect of bacterial surfaces as sorptive matrixes on metal balances in the environment.

References ALIBHAI, K.R.K., MEHORATRA, I. and FORSTER, C.F., 1985, Heavy metal binding to digested sludge, Water Research, 19, 1483–1488. ANDRES, Y., MACCORDICK, H.J. and HUBERT, J.C., 1993, Adsorption of several actinide (Th, U) and lanthanide (La, Eu, Yb) ions by Mycobacterium smegmatis, Applied Microbiology and Biotechnology, 39, 413–417. ARCAND, Y., CHAVARIE, C. and GUIOT, S.R., 1994, Dynamic modelling of the population distribution in the anaerobic granular biofilm, Water Science and Technology, 30, 63–73.

59

Biosorbents for Metal lons BAYER, M.E. and BAYER, M.H., 1994, Biophysical and structural aspects of the bacterial capsule, American Society for Microbiology News, 60, 192–198. BENDER, J., ARCHIBOLD, E.R., IBEANUSI, V. and GOULD, J.P., 1989, Lead removal from contaminated water by a mixed microbial ecosystem, Water Science and Technology, 21, 1661–1664. BENDER, J., GOULD, J.P., VATCHARAPIKARN, Y., YOUNG, J.S. and PHILLIPS, P., 1994, Removal of zinc and manganese from contaminated water with cyanobacteria mats, Water and Environmental Federation, 66, 679–683. BENDER, J., LEE, R.F. and PHILLIPS, P., 1995, Uptake and transformation of metals and metalloids by microbial mats and their use in bioremediation, Journal of Industrial Microbiology, 14, 113–118. BENEDETTI, M.F., MILNE, C.J., KINNIBURGH, D.G., VAN RIEMSDIJK, W.H. and KOOPAL, L.K., 1995, Metal ion binding to humic substances: application of the non-ideal competitive adsorption model, Environmental Science and Technology, 29, 446–457. BEVERIDGE, T.J., 1988, The bacterial surface: general considerations towards design and function, Canadian Journal of Microbiology, 34, 363–372. BEVERIDGE, T.J., 1989a, Role of cellular design in bacterial metal accumulation and mineralization, Annual Review on Microbiology, 43, 147–171. BEVERIDGE, T.J., 1989b, Interactions of metal ions with components of bacterial cell walls and their biomineralization, In: Metal-microbe Interactions, POOLE, R.K. and GADD, G.M. (Eds), IRL Press, Oxford, pp. 65–83. BEVERIDGE, T.J., 1993, Current trends and future prospects in prokaryotic envelope research: a microscopist’s view, Journal of Applied Bacteriology, Symposium Supplement, 74, 143S–153S. BEVERIDGE, T.J., FORSBERG, C.W. and DOYLE, R.J., 1982, Major sites of metal binding in Bacillus licheniformis walls, Journal of Bacteriology, 150, 1438–1448. BEVERIDGE, T.J. and FYFE, W.S., 1985, Metal fixation by bacterial cell walls, Canadian Journal of Earth Science, 22, 1893–1898. BEVERIDGE, T.J., MELOCHE, J.D., FYFE, W.S. and MURRAY, R.G. W., 1983, Diagenesis of metals chemically complexed to bacteria: laboratory formation of metal phosphates, sulfides and organic condensates in artificial sediments, Applied and Environmental Microbiology, 45, 1094–1108. BEVERIDGE, T.J. and MURRAY, R.G. E., 1976, Uptake and retention of metals by cell walls of Bacillus subtilis, Journal of Bacteriology, 127, 1502–1518. BEVERIDGE, T.J. and MURRAY, R.G. E., 1980, Sites of metal deposition in the cell wall of Bacillus subtilis, Journal of Bacteriology, 141, 876–887. BEVERIDGE, T.J., SARA, M., PUM, D., SPROTT, G.D., STEWART, M. and SLEYTR, U.B., 1988, The structure, chemistry and physicochemistry of the Methanospirillum hungatei GP1 sheath, In: Crystalline Bacterial Cell Surface Layers, SLEYTR, U.B., MESSNER, P., PUM, D. and SARA, M. (Eds), Springer-Verlag, Berlin, pp. 26–30. BEVERIDGE, T.J., SCHULTZE-LAM, S. and THOMPSON, J.B., 1995, Detection of anionic sites on bacterial walls, their ability to bind toxic heavy metals and form sedimentable flocs and their contribution to mineralization in natural freshwater environments, In: Metal Speciation and Contamination of Soil, ALLEN, H.E., HUANG, C.P., BAILEY, G.W. and BOWERS, A.R. (Eds), CRC Press, Boca Raton, Florida, pp. 183–206. BRADY, J.M. and TOBIN, J.M., 1995, Binding of hard and soft metal ions to Rhizopus arrhizus biomass, Enzyme Microbial Technology, 17, 791–796. BRIERLEY, J.A., 1990, Production and application of a Bacillus-based product for use in metals biosorption, In: Biosorption of Heavy Metals, VOLESKY, B. (Ed.), CRC Press, Boca Raton, Florida, pp. 305–311. BRIERLEY, C.L., BRIERLEY, J.A. and DAVIDSON, M.S., 1989, Applied microbial processes for metals recovery and removal from wastewater, In: Metal Ions and Bacteria, 60

General Bacterial Sorption Processes BEVERIDGE, T.J. and DOYLE, R.J. (Eds), John Wiley and Sons, New York, pp. 359– 382. BROWN, D.A., KADIMENI, D.C., SAWICKI, J.A. and BEVERIDGE, T.J., 1994, Minerals associated with biofilm occurring on exposed rock in a granitic underground research laboratory, Applied and Environmental Microbiology, 60, 3182–3191. BROWN, M.J. and LESTER, J.N., 1982, Role of bacterial extracellular polymers in metal uptake in pure bacterial cultures and activated sludge—I. Effects of metal concentration , Water Research, 16, 1539–1548. BUFFLE, J., 1990, Complexation Reactions in Aquatic Systems. An Analytical Approach, Ellis Horwood, Chichester. BUFFLE, J., TESSIER, A. and HAERDI, W., 1984, Interpretation of trace metal complexation by aquatic organic matter, In: Complexation of Trace Metals in Natural Waters, KRAMER, C.J.M. and DUINKER, J.C. (Eds), Martinus Nijhoff/Dr W.Junk, The Hague, The Netherlands, pp. 301–316. CHEN, Y., JURKEVITCH, E., BARNESS, E. and HADAR, Y., 1994, Stability constants of pseudobactin complexes with transition metals, Soil Science Society of America Journal, 58, 390–396. CHEN, J.H., LION, L.W., GHIORSE, W.C. and SHULER, M.L., 1995, Mobilization of adsorbed cadmium and lead in aquifer material by bacterial extracellular polymers, Water Research, 29, 421–430. COLLINS, Y.E. and STOTZKY, G., 1992, Heavy metals alter the electrokinetic properties of bacteria, yeasts and clay minerals, Applied and Environmental Microbiology, 58, 1592– 1600. CORZO, J., LEON-BARRIOS, M., HERNANDO-RICO, V. and GUTIERREZ-NAVARRO, A.M., 1994, Precipitation of metallic cations by the acidic exopolysaccharides from Bradyrhizobium japonicum and Bradyrhizobium (Chamaecytisus) strain BGA-1, Applied and Environmental Microbiology, 60, 4531–4536. COSTERTON, J.W., CHENG, K.J., GEESEY, G.G., LADD, T.I., NICKEL, J.C. and DASGUPTA, M., 1987, Bacterial biofilms in nature and disease, Annual Review of Microbiology, 41, 435–464. COTORAS, D., VIEDMA, P., CIFUENTES, L. and MESTRE, A., 1992, Sorption of metal ions by whole cells of Bacillus and Micrococus, Environmental Technology, 13, 551– 559. COUGHLIN, B.R. and STONE, A.T., 1995, Nonreversible adsorption of divalent metal ions (Mn-II, Co-II, Ni-II, Cu-II and Pb-II) onto goethite: effects of acidification, Fe-II addition and picolinic acid addition, Environmental Science and Technology, 29, 2445– 2455. COWAN, C.E., ZACHARA, J.M. and RESCH, C.T., 1991, Cadmium adsorption on iron oxides in the presence of alkaline earth elements, Environmental Science and Technology, 25, 437–446. DAVIS, J.A. and LECKIE, J.O., 1978, Surface ionization and Complexation at the oxide/ water interface. II. Surface properties of amorphous iron oxyhydroxide and adsorption of metal ions, Journal of Colloid and Interface Science, 67, 90–107. DOYLE, R.J., 1989, How cell walls of Gram-positive bacteria interact with metal ions, In: BEVERIDGE, T.J. and DOYLE, R.J. (Eds) Metal Ions and Bacteria, New York, John Wiley and Sons, pp. 275–293. DOYLE, R., MATTHEWS, T.H. and STREIPS, U.N., 1980, Chemical basis for selectivity of metal ions by the Bacillus subtilis cell wall, Journal of Bacteriology, 143, 471–480. FALLA, J. and BLOCK, J.C., 1993, Binding of Cd2+, Ni2+, Cu2+ and Zn2+ by isolated envelopes of Pseudomonas fluorescens, FEMS Microbiological Letters, 108, 347–352. FALLA, J.A., PETIT, E. and BLOCK, J.C., 1995, Extractability of cadmium, copper and zinc from contaminated biomass using NTA (nitriloacetic acid), Environmental Technology, 16, 685–691. 61

Biosorbents for Metal lons FENDORF, S.E., 1995, Surface reactions of chromium in soils and waters, Geoderma, 67, 55– 71. FERRIS, F.G., 1989, Metallic ion interactions with the outer membrane of Gram-negative bacteria, In: Metal Ions and Bacteria, BEVERIDGE, T.J. and DOYLE, R.J. (Eds), John Wiley and Sons, New York, pp. 295–323. FERRIS, F.G. and BEVERIDGE, T.J., 1984, Binding of a paramagnetic metal cation to Escherichia coli K-12 outer membrane vesicles, FEMS Microbiological Letters, 24, 43. FERRIS, F.G. and BEVERIDGE, T.J., 1986, Physicochemical roles of soluble metal cations in the outer membrane of Escherichia coli K-12, Canadian Journal of Microbiology, 32, 594–601. FLEMMING, C.A., FERRIS, F.G., BEVERIDGE, T.J. and BAILEY, G.W., 1990, Remobilization of toxic heavy metals adsorbed to bacterial wall-clay composites, Applied and Environmental Microbiology, 56, 3191–3203. FLORA, J.R. V., SUIDAN, M.T., BISWAS, P. and SAYLES, G.D., 1995a, Modeling algal biofilms: role of carbon, light, cell surface charge and ionic species, Water and Environmental Research, 67, 87–94. FLORA, J.R. V., SUIDAN, M.T., BISWAS, P. and SAYLES, G.D., 1995b, A modeling study of anaerobic biofilm systems. 1. Detailed biofilm modeling, Biotechnology and Bioengineering, 46, 43–53. FLORA, J.R. V., SUIDAN, M.T., BISWAS, P. and SAYLES, G.D., 1995c, A modeling study of anaerobic biofilm systems. 2. Reactor modeling, Biotechnology and Bioengineering, 46, 54–61. FREEMAN, C. and LOCK, M.A., 1995, The biofilm polysaccharide matrix: a buffer against changing organic substrate supply? Limnology and Oceanography, 40, 273–278. GEESEY, G.G. and JANG, L., 1989, Interactions between metal ions and capsular polymers, In: Metal Ions and Bacteria, BEVERIDGE, T.J. and DOYLE, R.J. (Eds), John Wiley and Sons, New York, pp. 325–357. GEESEY, G.G. and JANG, L., 1990, Extracellular polymers for metal binding, In: Microbial Mineral Recovery, EHRLICH, H.L. and BRIERLEY, C.L. (Eds), McGraw-Hill Book Co., New York, pp. 223–247. GHIORSE, W.C., 1984, Biology of iron and manganese-depositing bacteria, Annual Review of Microbiology, 38, 515–550. GILBERT, P., EVANS, D.J., DUGUID, I.G., EVANS, E. and BROWN, M.R. W., 1991, Cell surface properties of Escherichia coli and Spathylococcus epidermidis, In: Microbial Cell Surface Analysis. Structural and Physicochemical Methods, MOZES, N., HANDLEY, P.S., BUSSCHER, H.J. and ROUXHET, P.G. (Eds), VCH, New York, pp. 340–356. GRAHAM, L.L. and BEVERIDGE, T.J., 1994, Structural differentiation of the Bacillus subtilis-168 cell-wall, Journal of Bacteriology, 176, 1413–1421. GRAHAM, L.L., HARRIS, R., VILLIGER, W. and BEVERIDGE, T.J., 1991, Freezesubstitution of gram-negative eubacteria: general cell morphology and envelope profiles, Journal of Bacteriology, 173, 1623–1633. HANCOCK, I.C., 1991, Microbial cell surface architecture, In: MOZES, N., Microbial Cell Surface Analysis. Structural and Physicochemical Methods, HANDLEY, P.S., BUSSCHER, H.J. and ROUXHET, P.G. (Eds), VCH, New York, pp. 21–62. HINZ, C., GASTON, L. and SELIM, H., 1994, Effect of sorption isotherm type on predictions of solute mobility in soil, Water Resource Research, 30, 3013–3021. HOLM, T.R. and ZHU, X.F., 1994, Sorption by kaolinite of Cd 2+, Pb 2+ and Cu 2+ from landfill leachate-contaminated ground water, Journal of Contaminant Hydrology, 16, 271–287. HOYLE, B.D. and BEVERIDGE, T.J., 1984, Metal binding by the peptidoglycan sacculus of Escherichia coli K-12, Canadian Journal of Microbiology, 30, 204–211. 62

General Bacterial Sorption Processes HSI, C.-K. and LANGMUIR, D., 1985, Adsorption of uranyl onto ferric oxyhydroxides: application of the surface complexation site-binding model, Geochimica et Cosmochimica Acta, 49, 1931–1941. HSIA, T.H., Lo, S.L. and LIN, C.F., 1992, Interaction of Cr(VI) with amorphous iron oxide—adsorption density and surface charge, Water Science and Technology, 26, 181– 188. HSIA, T.H., Lo, S.L., LIN, C.F. and LEE, D.Y., 1993, Chemical and spectroscopic evidence for specific adsorption of chromate on hydrous iron oxide, Chemosphere, 26, 1897–1904. HSIEH, K.M., MURGEL, G.A., LION, L.W. and SHULER, M.L., 1994a, Interactions of microbial biofilms with toxic trace metals: 1. Observation and modeling of cell growth, attachment and production of extracellular polymer , Biotechnology and Bioengineering, 44, 219–231. HSIEH, K.M., MURGEL, G.A., LION, L.W. and SHULER, M.L., 1994b, Interactions of microbial biofilms with toxic trace metals: 2. Prediction and verification of an integrated computer model of lead (II) distribution in the presence of microbial activity, Biotechnology and Bioengineering, 44, 232–239. IKEDA, F., SHUTO, H., SAITO, T., FUKUI, T. and TOMITA, K., 1982, An extracellular polysaccharide produced by Zooglea ramigera 115, European Journal of Biochemistry, 123, 437–445. JAMES, A.M., 1982, The electrical properties and topochemistry of bacterial cells, Advances in Colloid and Interface Science, 15, 171–221. JARDIM, W.F. and ALLEN, H.E., 1984, Measurement of copper complexation by naturally occurring ligands, In: Complexation of Trace Metals in Natural Waters, KRAMER, C.J.M.and DUINKER, J.C.E. (Eds), Martinus Nijhoff/Dr W.Junk, The Hague, The Netherlands, pp. 1–15. KONG, S., JOHNSTONE, D.L., YONGE, D.R., PETERSEN, J.N. and BROUNS, T.N., 1993a, Remobilization of chromium from starved and fresh subsurface bacterial consortium, Biotechnology Letters, 15, 1081–1084. KONG, S., JOHNSTONE, D.L., YONGE, D.R., PETERSEN, J.N. and BROUNS, T.M., 1994, Long-term intracellular chromium partitioning with subsurface bacteria, Applied Microbiology and Biotechnology, 42, 403–407. KONG, S., YONGE, D.R., JOHNSTONE, D.L., PETERSEN, J.N. and BROUNS, T.M., 1993b, Competing ion effect on chromium adsorption with fresh and starved subsurface bacterial consortium, Biotechnology Letters, 15, 87–92. KONIG, H., 1988, Archaeobacterial cell envelopes, Canadian Journal of Microbiology, 32, 395–406. KOVAL, S.F., 1988, Paracrystalline protein surface arrays on bacteria, Canadian Journal of Microbiology, 34, 407–414. KRAMER, C.J.M. and DUINKER, J.C., 1984, Complexation of Trace Metals in Natural Waters, Martinus Nijhoff/Dr W. Junk Publishers, The Hague, The Netherlands. KRUEGER, S., OLSON, G.J., JOHNSONBAUGH, D. and BEVERIDGE, T.J., 1993, Characterization of the binding of gallium, platinum and uranium to Pseudomonas fluorescens by small-angle X-ray scattering and transmission electron microscopy, Applied and Environmental Microbiology, 59, 4056–4064. LEGRET, M., DIVET, L. and JUSTE, C., 1987, Mobilité et extraction des metaux lourds associés aux boues de station d’épuration, Water Research, 21, 541–547. LEWANDOWSKI, Z., STOODLEY, P. and ALTOBELLI, S., 1995, Experimental and conceptual studies on mass transport in biofilms, Water Science and Technology, 31, 153–162. MANTOURA, R.F.C. and RILEY, J.P., 1975, The use of gel filtration in the study of metal binding by humic acids and related compounds, Analytica Chimica Acta, 78, 193–200. MANZINI, G., CESARO, A., DELBEN, F., PAOLETTI, S. and REISENHOFER, E., 1984, 63

Biosorbents for Metal lons Copper (II) binding by natural ionic polysaccharides Part 1. Potentiometric and spectroscopic data, Bioelectrochemistry and Bioenergetics, 12, 443. MARQUES, A.M., BONET, R., SIMON-PUJOL, M.D., FUSTE, M.C. and CONGREGADO, F., 1990, Removal of uranium by an exopolysaccharide from Pseudomonas sp., Applied Microbiology and Biotechnology, 34, 429–431. MARQUES, A.M., ROCA, X., SIMON-PUJOL, M.D., FUSTE, M.C. and CONGREGADO, F., 1991, Uranium accumulation by Pseudomonas sp. EPS—5028, Applied Microbiology and Biotechnology, 35, 406–410. MARQUIS, R.E., MAYZEL, K. and CARSTENSEN, E.L., 1976, Cation exchange in cell walls of gram-positive bacteria, Canandian Journal of Microbiology, 22, 975–982. MARTELL, A.E., 1971, Principles of complex formation, In: Organic Compounds in Aquatic Environments, FAUST, S.J. and HUNTER, J.V. (Eds), Marcel Dekker, New York, pp. 239–263. MASSOL-DEYA, A.A., WHALLON, J., HICKEY, R.F. and TIEDJE, J.M., 1995,Channel structures in aerobic biofilms of fixed-film reactors treating contaminated groundwater, Applied and Environmental Microbiology, 61, 769–777. MCLEAN, R.J. C., BEAUCHEMIN, D., CLAPHAM, L. and BEVERIDGE, T.J., 1990, Metalbinding characteristics of the gamma-glutamyl capsular polymer of Bacillus licheniformis ATCC 9945, Applied and Environmental Microbiology, 56, 3671–3677. MELLOR, R.B., RONNENBERG, J., CAMPBELL, W.H. and DIEKMANN, S., 1992, Reduction of nitrate and nitrite in water by immobilized enzymes, Nature, 355, 717–719. MITTLEMAN, M.W. and GEESSEY, G.G., 1985, Copper-binding characteristics of exopolymers from a freshwater sediment bacterium, Applied and Environmental Microbiology, 49, 846–851. MULLEN, M.D., WOLF, D.C., FERRIS, F.G., BEVERIDGE, T.J., FLEMMING, C.A. and BAILEY, G.W., 1989, Bacterial sorption of heavy metals, Applied and Environmental Microbiology, 55, 3143–3149. NIEBOER, E. and RICHARDSON, D.H. S., 1980, The replacement of the non-descript term ‘heavy metals’ by a biologically and chemically significant classification of metal ions, Environmental Pollution Series B, 1, 3–26. NORBERG, A.B. and ENFORS, S.O., 1982, Production of extracellular polysaccharide by Zooglea ramigera, Applied and Environmental Microbiology, 44, 1231–1237. NORBERG, A.B. and PERSSON, H., 1984, Accumulation of heavy metal ions by Zooglea ramigera, Biotechnology and Bioengeneering, 26, 239–246. PATEL, G.B., SPROTT, G.D., HUMPHREY, R.W. and BEVERIDGE, T.J., 1986, Comparative analyses of the sheath structures of Methanothrix concilii GP6 and Methanospirillum hungatei strains GP1 and JF1, Canadian Journal of Microbiology, 32, 623–631. PAYNE, T.E. and WAITE, T.D., 1991, Surface complexation modelling of uranium sorption data obtained by isotope exchange techniques, Radiochimica Acta, 52–53, 487–493. PEDIT, J. and MILLER, C.T., 1994, Heterogeneous sorption processes in subsurface systems. 1. Model formulations and applications, Environmental Science and Technology, 28, 2094–2104. PETERS, R.W. and SHEM, L., 1995, Treatment of soils contaminated with heavy metals, In: ALLEN, H.E., HUANG, C.P., BAILEY, G.W. and BOWERS, A.R. (Eds), Metal Speciation and Contamination of Soil, CRC Press, Boca Raton, Florida, pp. 255–274. PHILIP, L., IYENGAR, L. and VENKOBACHAR, C., 1995, Biosorption of copper (II) by Pseudomonas aeruginosa, International Journal of Environment and Pollution, 5, 92– 99. PUM, D., SARA, M. and SLEYTR, U.B., 1992, Two-dimensional (glyco)protein crystals as patterning elements and immobilization matrices for the development of biosensors, In: Immobilized Macromolecules: Application Potential, SLEYTR, U.B., MESSNER, P., SARA, M. and PUM, D. (Eds), Springer-Verlag, London, pp. 141–160. 64

General Bacterial Sorption Processes RENDLEMAN, J.A., 1978, Metal-polysaccharide complexes—Part II, Food Chemistry, 3, 127–162. RUDD, T., STERRIT, R.M. and LESTER, J.N., 1984, Formation and conditional stability constants of complexes formed between heavy metals and bacterial extracellular polymers, Water Research, 18, 379–384. SALTON, M.R. J., 1963, The relationship beween the nature of the cell wall and the Gram stain, Journal of General Microbiology, 30, 223–225. SANTOS, R. and CALLOW, M.E., 1991, The structure of Pseudomonas fluorescens biofilms in contact with flowing systems, Biofouling, 4, 319–336. SARA, M., KUPCU, S., WEINER, C., WEIGERT, S. and SLEYTR, U.B., 1993, S-layers as immobilization and affinity matrices, In: Advances in Bacterial Paracrystalline Surface Layers, BEVERIDGE, T.J. and KOVAL, S.F. (Eds), Plenum Publishing Corporation, New York, pp. 195–204. SARA, M. and SLEYTR, U.B., 1987, Production and characteristics of ultrafiltration membranes with uniform pores from two-dimensional arrays of proteins, Journal of Membrane Science, 33, 27–49. SCHULTZE-LAM, S., HARAUZ, G. and BEVERIDGE, T.J., 1992, Participation of a cyanobacterial S-layer in fine-grain mineral formation, Journal of Bacteriology, 174, 7971–7981. SHUTTLEWORTH, K.L. and UNZ, R.F., 1993, Sorption of heavy metals to the filamentous bacterium Thiothrix strain A1, Applied and Environmental Microbiology, 59, 1274–1282. SIMOES GONCALVES, M.L., SIGG, L., REUTLINGER, M. and STUMM, W., 1987, Metal ion binding by biological surfaces: voltammetric assessment in the presence of bacteria, Science of the Total Environment, 60, 105–119. SLEYTR, U.B. and MESSNER, P., 1988, Crystalline surface layers in procaryotes, Journal of Microbiology, 170, 2891–2897. SOUTHAM, G., FERRIS, F.G. and BEVERIDGE, T.J., 1995, Mineralized bacterial biofilms in sulphide tailings and in acid mine drainage systems, In: LAPPIN-SCOTT, H.M. and COSTERTON, J.M. (Eds) Microbial Biofilms, 5, Cambridge University Press, Cambridge, pp. 148–170. STEINER, A.E., MCLAREN, D.A. and FORSTER, C.F., 1976, The nature of activated sludge flocs, Water Research, 10, 25–30. STEVENSON, F.J. and FITCH, A., 1986, Chemistry of complexation of metal ions with soil solution organics, In: Interactions of Soil Minerals with Natural Organics and Microbes, Soil Science Society of America, Madison, Wisconsin, pp. 29–58. STEWART, P.S., MURGA, R., SRINIVASAN, R. and DE BEER, D., 1995, Biofilmstructural heterogeneity visualized by three microscopic methods, Water Research, 29, 2006–2009. STUMM, W., SIGG, L. and SULZBERGER, B., 1994, The role of coordination at the surface of aquatic particles, In: BUFFLE, J. and DE VITRE, R.R. (Eds) Chemical and Biological Regulation of Aquatic Systems, CRC Press, Boca Raton, Florida, pp. 45–90. THOMPSON, J.B. and BEVERIDGE, T.J., 1993, Interactions of metal ions with bacterial surfaces and the ensuing development of minerals, In: Paniculate Matter and Aquatic Contaminants, RAO, S.S. (Ed.), Lewis Publishers Inc., Chelsea, Michigan, pp. 65– 104. THOMPSON, J.B. and FERRIS, F.G., 1990, Cyanobacterial precipitation of gypsum, calcite and magnesite from natural alkaline lake water, Geology, 18, 995–998. TIPPING, E. and HURLEY, M.A., 1992, A unifying model of cation binding by humic substances, Geochimica et Cosmochimica Acta, 56, 3627–3641. TURLEAR, M.G. and CHARACKLIS, W.G., 1982, Dynamics of biofilm processes, Journal of the Water Pollution Control Federation, 54, 1288–1381. UMEDA, A., UEKI, Y. and AMAKO, K., 1987, Structure of the Staphylococcus aureus cell wall determined by the freeze-substitution method, Journal of Bacteriology, 169, 2482– 2487. 65

Biosorbents for Metal lons URRUTIA, M.M. and BEVERIDGE, T.J., 1993a, Mechanism of silicate binding to the bacterial cell wall in Bacillus subtilis, Journal of Bacteriology, 175, 1936–1945. URRUTIA, M.M. and BEVERIDGE, T.J., 1993b, Remobilization of heavy metals retained as oxyhydroxides or silicates by Bacillus subtilis cells, Applied and Environmental Microbiology, 59, 4323–4329. URRUTIA, M.M. and BEVERIDGE, T.J., 1995, Formation of short-range ordered aluminosilicates in the presence of a bacterial surface (Bacillus subtilis) and organic acids, Geoderma, 65, 149–165. URRUTIA, M.M., KEMPER, M., DOYLE, R.J. and BEVERIDGE, T.J., 1992, The proton motive force influences metal binding ability in Bacillus subtilis, Applied and Environmental Microbiology, 58, 3837–3844. VEITH, J.A. and SPOSITO, G., 1977, On the use of the Langmuir equation in the interpretation of ‘adsorption’ phenomena, Soil Science Society of America Journal, 41, 497–502. WANNER, O., CUNNINGHAM, A.B. and LUNDMAN, R., 1995, Modeling biofilm accumulation and mass transport in a porous medium under high substrate loading, Biotechnology and Bioengineering, 47, 703–712. WIELAND, F., DOMPERT, W., BERNHARDT, G. and SUMPER, M., 1980, Halobacterial glycoproteins saccharides contain covalently linked sulfate, FEBS Letters, 120, 110–114. WIMPENNY, J., NICHOLS, W., STICKLER, D. and LAPPIN-SCOTT, H., 1994, Bacterial Biofilms and their Control in Medicine and Industry, Bioline, London. WOESE, C.R., KANDLER, O. and WHEELIS, M.L., 1990, Towards a natural system of organisms: proposal for the domains Archaea, Bacteria and Eucarya, Proceedings of the National Academy of Science of the U.S.A., 87, 4576. ZACHARA, J.M., COWAN, C.E. and RESCH, C.T., 1991, Sorption of divalent metallic cations on calcite, Geochimica et Cosmochimica Acta, 55, 1549–1562. ZACHARA, J.M., SMITH, S.C., MCKINLEY, J.P. and RESCH, C.T., 1993, Cadmium sorption on specimen and soil smectites in sodium and calcium electrolytes, Soil Science Society of America Journal, 57, 1491–1501.

66

4

Fungi as Biosorbents A.KAPOOR and T.VIRARAGHAVAN

Introduction Metals such as copper, zinc, manganese, nickel and cobalt in trace amounts serve as micro-nutrients for growth of microorganisms such as fungi and yeast. Fungi and yeast can also accumulate non-nutrient metals such as cadmium, mercury, lead, uranium, silver and gold in substantial amounts. Both living and dead fungal cells possess a remarkable ability for taking up toxic and precious metals. The uptake of metals by fungi and yeast has generated an interest in using them for removal of toxic metals from wastewater and recovery of precious metals (such as gold and silver) from process waters. The potential of fungal biomass in removing metals ions from wastewater was recognised by Zajic and Chiu (1972), Jilek et al. (1975) and Shumate et al. (1978). The fungi and yeast are used in a variety of industrial fermentation processes (Volesky, 1991a). Such industrial fermentation processes can serve as an economical and constant supply source of biomass for biosorption of metal ions. The fungi and yeast can also be easily grown in substantial amounts using unsophisticated fermentation techniques and inexpensive growth media (Kuyucak, 1990). Thus, fungal biosorption can serve as an economical means for removal/recovery of metal ions from aqueous solutions.

Modes of metal ion uptake The metal ion uptake by living and dead cells can consist of two differing modes. The first uptake mode is independent of cell metabolic activity, and is referred to as biosorption or passive uptake. It involves the surface binding of metal ions to cell walls and extracellular material. The second mode of metal uptake into the cell across the cell membrane is dependent on the cell metabolism, and is referred to as intracellular uptake, active uptake or bio-accumulation (Gadd, 1986; Volesky, 1991b). The first mode is common to metal uptake by both living and dead cells; the second mode, which is metabolism-dependent, occurs in living cells. For living cells the metal uptake is also facilitated by the production of metal-binding proteins. 67


Therefore, metal uptake may take place by different modes, depending on whether the cells are dead or living. Both living and dead cells are capable of metal uptake. The use of dead biomass seems to be a preferred alternative for the majority of metal uptake studies reported. The wider acceptability of dead cells is due to the absence of toxicity limitations, absence of requirements of growth media and nutrients in the feed solution, and the fact that biosorbed metals can be easily absorbed and recovered, the regenerated biomass can be re-used, and the metal uptake reactors can be easily modelled mathematically. The dead fungal cells have been shown to bind metals at levels greater than live cells depending on the methods used to kill (pre-treat) the live cells (Galun et al., 1987; Paknikar et al., 1993; Ross and Townsley, 1986; Kapoor and Viraraghavan, 1996). The pretreatment of biomass (which will be discussed later in detail) changes the characteristics of the cell surface and usually results in increased metal uptake.

Modelling of biosorption The equilibria of biosorption of metal ions can be described by adsorption-type isotherm models developed for adsorption of gases. The adsorption isotherm models of Langmuir, Freundlich and Brunauer-Emmett-Teller (BET) have been used to describe the biosorption equilibria. The Langmuir and Freundlich equations have been discussed in Chapter 2. The BET isotherm has the form:

where Cs is the saturation concentration of the metal ion, Q0 is the amount adsorbed per unit weight of biomass for monolayer biosorption, and B is a constant relating to the energy of interaction with the surface. The Langmuir model is developed on the assumptions that maximum adsorption occurs when a saturated monolayer of solute molecules is present on the adsorbent surface, the energy of adsorption is constant and there is no migration of adsorbate molecules in the surface plane. The Freundlich model is empirical in nature, and was developed for heterogeneous surfaces. The BET isotherm describes the multi-layer adsorption at the adsorbent surface and assumes that the Langmuir isotherm applies to each layer (Weber, 1972). The adsorption models are a useful means for describing the degree of biosorption as a function of equilibrium metal ion concentration at constant pH and temperature conditions. These models are mathematical representations of biosorption equilibria and, by providing metal uptake capacities for various fungal strains, serve as a means of comparing different species in terms of metal removal. Gadd (1990) pointed out that the mechanistic conclusions from the good fit of the model should be carefully interpreted. Scatchard (1949) developed a model to describe the attraction of proteins for small molecules and ions (see also Chapter 2). The Scatchard model has also been used to describe the biosorption equilibria (Huang et al., 1990; Tobin et al., 1990). The 68

Fungi as Biosorbents

interaction of a metal ion with the cell surface binding sites can be described by the following equations:

where K is the association constant, M is the metal ion under consideration, X is a binding site on the biomass surface and MX is the metal ion biosorbed on biomass. The plot of the amount of bound metal ion on biomass divided by equilibrium metal ion concentration versus the amount of bound metal ion gives a typical Scatchard plot (see Figure 2.2). The equation can be represented as follows:

where M e and X0 are the equilibrium metal ion concentration and the concentration of potential binding sites. The total number of binding sites may be extrapolated from data in which the complete saturation of the biomass binding sites is not observed. The Scatchard plots have been used to represent the binding of ligands to macromolecules (Dahlquist, 1978). As indicated earlier, the biosorption of metals by cell surface binding can take place for both living and dead cells, and is of particular interest in the removal and recovery of metal ions. In the next two sections the metal uptake by living and dead cells is discussed. Table 4.1 gives the biosorptive capacities of various fungi and yeast for different metals. The biosorptive capacities presented in Table 4.1 may not be the maximum values, but they do reflect the degree to which a particular fungus is effective in metal removal.

Biosorption by living cells The living cells of Penicillium, Aspergillus, Trichoderma, Rhizopus, Mucor, Saccharomyces and Fusarium have been shown to biosorb metal ions (Galun et al., 1983c; Kiff and Little, 1986; Strandberg et al., 1981; Townsley and Ross, 1985, 1986; Townsley et al., 1986a, 1986b). Even the spores of Penicillium italicum can accumulate copper (Somers, 1963). Penicillium biomass has been found to biosorb heavy metals such as Cu, Zn, Cd, Au, Pb (Siegel et al., 1987). The living fungal cells of A. niger, M. rouxii and R. arrhizus have been shown to take up precious metals such as gold and silver (Gee and Dudeney, 1988; Kuyucak and Volesky, 1988; Mullen et al., 1992). The living cells of Penicillium, Rhizopus and Saccharomyces can biosorb radionuclides such as uranium, strontium and caesium (Galun et al., 1983a, 1983b; Rome and Gadd, 1991; Strandberg et al., 1981). The metal ion uptake by living cells is a function of the cell age, composition of growth media, contact time, pH of metal solution, and temperature. The biosorption of metals such as copper, zinc, cadmium, lead and uranium by non-growing cells of Penicillium, Aspergillus, Saccharomyces, Rhizopus and Mucor has been observed to reach equilibrium in 1 to 4.0 h (Gadd et al., 1988; Galun et al., 1983c; Huang et al., 1990; Kiff 69

Biosorbents for Metal lons Table 4.1 The biosorptive capacity of various fungal organisms

70

Fungi as Biosorbents Table 4.1— (Continued)

and Little, 1986; Mullen et al., 1992; Townsley and Ross, 1985; Townsley et al., 1986a). The kinetics of biosorption of metals is usually biphasic in nature, consisting of an initial rapid phase, contributing up to 90% of biosorption, and lasting for 10 minutes. The second phase is slower and has been found to last for up to 4 hours. Cell age has also been found to effect the biosorption of metal ions. Increased biosorption has been observed during the lag period or early stages of growth and declined as cultures reached stationary phase. A. niger, P. spinulosum and T. viride showed a similar uptake pattern (Townsley and Ross, 1985, 1986; Townsley et al., 1986b). Zajic and Chiu (1972) observed that uranium biosorption by five-day-old mycelia of Penicillum was approximately twice that of 15-day-old mycelium. Volesky and May (1995) observed that 12 h old cultures of baker’s yeast were able to biosorb 2.6 times more uranium than 24 h cultures. The biosorption also depends on the culture conditions and growth media composition. Higher copper uptake has been observed by live and freeze-dried aerobically grown S. cerevisiae than by anaerobically grown cells (May, 1984). The higher uptake in aerobic conditions could be due to the copper-binding protein metallothionin. Strandberg et al. (1981) observed that S. cerevisiae cells grown on a synthetic medium have a faster rate of biosorption (2.5 times) than cultures grown on a rich organic medium, while no difference was observed for aerobically and anaerobically grown yeast cells. Volesky (1994) indicated that R. nigricans cultures grown on glucose-peptone, sucrose and potato-dextrose medium exhibited a uranium uptake capacity of 0.475, 0.450 and 0.34 mmol U per g of biomass respectively. The A. oryzae grown using glucosepeptone medium also exhibited higher uranium uptake. Treen-Sears (1986) found that biosorption of Cu by non-living R. javanicus improved by the addition of divalent cations such as Mn, Zn, Ca and Mg to the growth medium. The growth medium 71


controls the composition and structure of cell wall, which in turn affects biosorption. Research has indicated that it is possible to manipulate fungal cells to increase the metal biosorptive capacities and, therefore, produce efficient fungal biosorbents. The biosorption of metal ions strongly depends on pH. The biosorption of Ni, Zn, Cd, and Pb by P. digitatum was observed to be inhibited below pH 3.0 (Galun et al., 1987). Cadmium biosorption was found to increase with pH from acidic (2.0) to basic range (7.0) (Huang et al., 1988a, 1988b; Mullen et al., 1992; Townsley et al., 1986b). The degree of biosorption of cadmium was explained using proton-competitive adsorption reactions. The surface charge of the fungal cells is predominantly negative over the pH range 3 to 10. The magnitude of negative surface charge for various fungal cells follows the order Fusarium solani>T. viride>A. nidullans>A. niger+A. oryzae>P. notadum (Huang et al., 1988a, 1988b). Cadmium removal capacity followed the order F. solani>T. viride>A. nidullans+A. oryzae>P. notadum>A. niger. Thus, biosorption appeared to be proportional to negative surface charge, the only exception being A. niger. Biosorption of molybdenum on S. cerevisiae was maximum at pH 1.5 and reduced as pH increased to 5.0 (Pryfogle et al., 1989). The rate of uranium biosorption was observed to be greatest at pH 5.5 (Shumate et al., 1978). Strandberg et al. (1981) noted that the initial rate of uranium biosorption on S. cerevisiae increased as pH increased from 2.5 to 5.5, but maximum equilibrium uptake occurred between pH 3 and 4.0. Biosorption is also affected by biomass concentrations (Itoh et al., 1975). Lower cadmium uptake was observed at higher concentrations of A. oryzae (Kiff and Little, 1986). The amount of copper biosorbed per unit weight of biomass decreased with an increase in concentration of R. arrhizus, Cladosporium resinae and P. italicum (Rome and Gadd, 1987). Itoh et al. (1975) attributed lower uptake of metals at higher biomass concentrations to electrostatic interactions between cells. At higher concentrations, the cells in suspension can attach to surfaces or each other, thus lowering the cell surface area in contact with the solution. Biosorption of metals by living cells has been modelled using Langmuir, Freundlich, BET and Scatchard models. Biosorption of copper on P. italicum has been described by the Freundlich and Langmuir isotherms, and for R. arrhizus copper biosorption was described by the BET isotherm (Rome and Gadd, 1987). Biosorption of cadmium by resting cells of S. cerevisiae and A. oryzae has been described by both the Langmuir and Freundlich isotherms. Mullen et al. (1992) modelled the biosorption of cadmium, copper, lanthanum and silver on A. niger and M. roxii by the Freundlich isotherm. The uptake of metals discussed so far was mainly due to the passive mode. The biosorption of heavy metal ions on the cell surface occurs by ion-exchange and complexation reactions with functional groups. The various functional groups believed to be involved in metal binding can include carboxyl, amine, amides, hydroxyl, phosphate and sulfhydryl groups (Strandberg et al., 1981; Remacle, 1990). Metals such as silver, gold and uranium have also been removed as a result of precipitation/crystallisation on fungal and algal surfaces (Mullen et al., 1992; Gee and Dudeney, 1988). The biosorption contributed by precipitation and crystallisation can possibly be explained by BET isotherms. The metal uptake can also take place by an active mode, which is dependent on the cell metabolic cycle, and metal ions are transported into the cell material across the cell wall. The active mode can contribute significantly to metal removal for yeast (Norris and Kelly, 1977, 1979). The metal uptake by active mode has been observed 72


for metals such as Cu, Cd, Ni, Zn, Co, Mn, Sr, Mg and Ca (Avery and Tobin, 1992; Borst-Pauwels, 1981; Gadd and Mowll, 1983; Norris and Kelly, 1977; Parkin and Ross, 1986a, 1986b; White and Gadd, 1987). Gadd (1990) suggested that at high metal concentrations active mode may not contribute significantly to metal uptake, especially for filamentous fungi. The mechanism of intracellular uptake is complex and not fully understood (Gadd, 1986).

Biosorption of metal ions by non-living cells The efficiency of dead cells in biosorbing metal ions may be greater than, equivalent to or less than that of the living cells. The use of dead cells offers the following advantages over living cells: • the metal removal system is not subject to toxicity limitations • no requirements for growth media and nutrients • biosorbed metal ions can be easily desorbed and biomass can be re-used • raw biomass can be obtained from the industrial fermentation industries • biomass can be pretreated to enhance the metal biosorptive capacity • much simpler process control • no concern over the disposal of metabolic products or surplus nutrients which would be required for a system using living cells • the biomass can be stored for a period of time. The cells can be killed for biosorption by physical or chemical methods. Physical methods include vacuum and freeze drying, boiling, autoclaving and mechanical disruption. The chemical methods include contacting the biomass with various organic and inorganic compounds (Siegel et al, 1990; Kapoor and Viraraghavan, 1995). One of the main aims of the pretreatment of biomass is to produce biosorbents with enhanced metal removal capacities. Table 4.2 gives the methods employed to pretreat biomass and their effect on metal removal. The alkali treatment of fungal biomass has been shown to increase significantly the metal sorption capacity of Aspergillus, Mucor and Penicillium (Kapoor and Viraraghavan, 1996; Muzzarelli et al., 1980a, 1980b, 1981). The increased removal capacity results from deacetylation of chitin in the cell wall to form chitosan-glucan complexes with higher affinity for metal ions. The biosorption of metal ions is a rapid process and often reaches equilibrium within four to six hours (Akthar et al., 1995; Brady et al., 1994; Gadd et al., 1988; Guibal et al., 1992). Tobin et al (1984) investigated the biosorption of a variety of heavy metals by heat-killed R. arrhizus at pH 4.0. The biosorption capacity was observed to be related to ionic radii of the metal ions. The biosorption was higher for metals with a larger ionic radius, the exceptions being chromium and the alkali metal ions. The alkali metal ions were not biosorbed, as they lack the ability to form complexes with the various ligand groups present on the fungal surface. The biosorption of metal ions such as Sr, Mn, Zn, Cd, Cu and Pb has been observed to ), where Xm is electronegativity and r is be proportional to their covalent index ( the ionic radius (Brady and Tobin, 1995). The fungal biomass possesses 73

Table 4.2 The various physical and chemical methods used in pretreatment of biomass

*1, Decrease in metal uptake in comparison to living cells; 2, increase in metal uptake as compared to living cells; 3, no change in metal uptake; 4, no data available for comparison to control.


exceptionally high biosorptive capacity for radionuclides, such as uranium and thorium (in excess of 180 mg/g) (Guibal et al., 1992; Treen-Sears et al., 1984; Tsezos and Volesky, 1981; Yakubu and Dudeney, 1986; White and Gadd, 1990). Similarly for silver, the uptake by alkali-treated A. niger has been observed to be as high as 100 mg/g (Akthar et al., 1995). Removal of these metal ions by fungi was higher than for activated carbon and ion exchange resins. The metal removal capacity of cadmium by S. cerevisiae, A. oryzae, Mortierella ramannjian, R. sexualis, R. stolonfier, Zygorhynchhus heterogamus and Z. moelleria was also higher in comparison with adsorbents such as aluminium oxide, activated carbon, and activated charcoal (Azab and Peterson, 1989; Azab et al., 1990; Huang et al., 1988a). The biosorption of metal ions has been observed to be pH-dependent. Optimum removals are usually achieved in the pH range of 4–5, and biosorption is substantially reduced at a pH of 2.5. Uranium biosorption on Penicillium was pHinsensitive in the range 2.5–9.5 (Galun et al., 1983a). The biosorption of lead, cadmium, nickel and zinc was severely inhibited at pH below 4.0 (Brady et al., 1994; Kapoor and Viraraghavan, 1996; Niu et al., 1993). The pH variation changes the speciation and availability of the metallic elements in solution, and also the chemical state of the functional groups responsible for biosorption. The competitive inhibition by protons has been shown by Huang et al. (1991). Fourest et al. (1994) observed that zinc biosorption on R. arrhizus, M. miehei and P. chrysogenum was less than 4.0 and for R. arrhizus, which exhibited al higher zinc uptake, it was 5.8. The metal uptake for R. arrhizus, M. miehei and P. chrysogenum increased from 16 to 35, 3 to 32 and 4.5 to 22 mg/g respectively, when pH of the reaction mixture was controlled at 7.0. Under uncontrolled conditions of pH, the drop in pH may create an undesired competition for metal ions from protons, thus lowering the metal uptake capacity. Based on this observation, the waste with buffering capacity will experience increased metal removal as the decrease in equilibrium sorption of the pH will be prevented. Fourest et al. (1994) observed that the calcium saturation of the biomass also prevented the equilibrium of the pH from dropping below 5.5 and thus enhanced the biosorption of zinc. The pH also affects the rate of metal biosorption (Guibal et al., 1992). The optimum pH for the removal of a particular metal ion will depend not only on the fungal strain being used but also on the extent of pretreatment the biomass receives. The biosorption of metal on certain fungal strains has been found to be both selective and, in some cases, competitive (Galun et al., 1984; Nakajima and Sakaguchi, 1986; Niu et al., 1993; Siegel et al., 1987; Tobin et al., 1988; Tsezos and Volesky, 1982a). The biosorption of uranium, mercury and lead is selective on fungal biomass. The uranium uptake by R. arrhizus was reduced in the presence of iron and zinc (Tsezos and Volesky, 1982a) and for Penicillium iron alone inhibited uranium biosorption and zinc had no significant effect (Galun et al., 1984; Siegel et al., 1987). This shows the variability in the metal biosorption which occurs in fungal species. The nature of biosorption on fungi is complex and extreme caution is warranted in drawing any general conclusion on fungal biosorption. Limited information is available on the competitive sorption of metal ions on fungal biomass. P. chrysogenum exhibits selectivity for Pb over metals such as Cd, Cu, Zn and As. The biosorptive uptake for metals decreased in the following order: Pb>Cd>Cu>Zn>As. Yakubu and Dudeney (1986) showed that biosorption of uranium on A. niger was substantially reduced in the presence of Cu, Zn, and Fe, and the preferential order for 76


biosorption was FeⰇU>Cu>Zn. Zhou and Kiff (1991) indicated that Mn, Zn, Cd, Mg and Ca inhibited Cu biosorption by R. arrhizus. The biosorption of radium is also reduced in the presence of Ca, Mg, Ba, Cu and Fe, and the effect was most pronounced in the presence of Ba (Tsezos et al., 1986). The metal binding capacity of Penicillium was observed to follow the series Fe>Cu, Zn, Ni>Cd, Pb>U (Galun et al., 1987). The biosorptive capacity series for Rhizopus was U>Pb>Cd>Zn>Cu (Tobin et al., 1984) and the Aspergillus series was Fe>U>Cu>Zn Yakubu and Dudeney, 1986). Brady et al. (1994) observed that biosorption of metals on aerobically grown yeast treated with hot alkali was selective. The metal uptake in two three-metal systems followed the order: Cu>Cr>Cd and Cu>Pb>Ni. The presence of anions also affects the biosorption of metal ions. Biosorption is reduced in the presence of ethylenediamine tetraacetate (EDTA), sulphate, chloride, phosphate, carbonate, glutumate, citrate and pyrophosphate (Zhou and Kiff, 1991; Tobin et al., 1987; Tsezos and Noh, 1984; Rao et al., 1993). The presence of EDTA has been found to severely affect the biosorption of copper, lanthanum, uranium, silver, cadmium and lead. The uranium biosorption has been found to be severely inhibited by carbonate. The anions present in wastewater can form complexes with the metal ions. The stability constants of metal-anion complexes can be high and, if these constants are greater than stability constants for metal biosorption sites on the cell surface, the biosorption can be expected to be considerably reduced. Therefore, anion complexation can decrease the biosorption of metal ions. Temperature also affects the biosorption of metal ions. Adsorption reactions are normally exothermic and the extent of adsorption increases with decreasing temperature (Weber, 1972). Uranium biosorption was found to increase with temperature in the range of 20 to 50ºC (Tsezos and Volesky, 1981), indicating that the biosorption process, in this case, was endothermic. Similar findings were obtained by Strandberg et al. (1981) and Shumate et al. (1978).

Regeneration of fungal biomass and elution of biosorbed metals The application of fungi and yeast as biosorbents depends not only on the biosorptive capacity, but also on the ease with which biomass can be regenerated and re-used. The cell surface-bound metal might be easily removed by the use of elutants, but the intracellular metal can only be released by destructive treatment. Various elutants have been screened for recovery of uranium from Rhizopus and Penicillium biomass (Galun et al., 1983b; Tsezos, 1984). Mineral acids, such as sulphuric and hydrochloric acid, are effective in recovery of sorbed uranium, but can cause damage to the biomass and reduce uranium biosorptive capacity in subsequent uses. Sodium bicarbonate was found to be an effective elutant. Galun et al. (1983b) observed that uranium was desorbed by ammonium carbonate and sodium bicarbonate. EDTA was less effective as an elutant, but it did increase uranium biosorption two-fold for subsequent use. The elution of Ni, Cu, Zn and Cd biosorbed on Penicillium biomass can be achieved by using dilute hydrochloric acid or dilute sodium hydroxide solution. Elution with 0.1 N NaOH and washing to pH 5.5 to 6.0 resulted in a 2–6 fold increase in mycelial uptake in subsequent use. No significant changes in biosorption of Ni, Cu, and Zn were observed when using HCl alone. The use of mineral acids as an elutant has been widely studied (Akthar et al., 1995; Huang et 77


al., 1990; Kapoor and Viraraghavan, 1996; Luef et al., 1991; Matheickal et al., 1991; Rome and Gadd, 1991; Townsley et al., 1986b; White and Gadd, 1990; Zhou and Kiff, 1991). Zhou and Kiff (1991) found no decrease in biosorptive capacity in subsequent uses when 0.1 N HCl was used as an elutant. Concerns over the damage to biosorbent structure from using mineral acids as elutants prompted us to investigate less aggressive elutants. 0.1 M calcium chloride and magnesium sulphate solutions were able to elute cadmium and nickel biosorbed by A. niger (Kapoor and Viraraghavan, 1996).

Use of immobilised fungal biomass in biosorption The biomass in contact with water becomes soft and can have low mechanical strength. The small particle size, low density and strength can cause difficulties in column application and separation of biomass and treated effluent. Therefore, industrial applications of fungal biosorption prefer immobilised or pelletised biomass. The living mycelial pellets of A. niger and R. arrhizus have been used for removal of thorium using air-lift column bio-reactors (White and Gadd, 1990). Pellets of A. niger have also been used for biosorption of uranium in a fluidised-bed reactor, and the system was found to be more efficient than ion exchange resin IRA-400 (Yakubu and Dudeney, 1986). For immobilisation of biomass, sand, paper or textile making fibres, foam biomass support particles, polysulphone, alginate, and an inorganic compound (identity of compound not disclosed) have been successfully used for biosorption (Ferguson et al., 1989; Huang et al., 1990; Khalid et al., 1993; Kiff and Little, 1986; Lakshmanan et al, 1989; Rome and Gadd, 1991; Tsezos and Deutschmann, 1990; Zhou and Kiff, 1991). Tobin et al. (1993) developed protocols for immobilisation of dried R. arrhizus biomass using alginate, polyacrylamide, epoxy resin and polyvinyl-formal materials. The biomass immobilised using polyvinyl-formal could retain up to 85% of the adsorption capacity of the free biomass. Muraleedharan et al. (1994b) used Ganoderma lucidum, a macro-fungus, without any immobilisation to treat effluents from monazite processing industry, and indicated the suitability of the biosorbent in a packed-bed reactor. The need to use immobilised biomass would not be of concern if an efficient method for separation of metal-loaded biomass from aqueous suspensions could be developed. The flotation technique for separation of fungal biosorbents is an alternative, but at the present time only limited data are available (Zouboulis and Matis, 1993).

Biosorption mechanism The biosorption of metal by biomass has been attributed to many different mechanisms which include metabolism-dependent transport, ion-exchange or complexation, adsorption of simple ionic species and hydrolysis products of metal ions. The metabolism-dependent uptake involves mechanisms which may be part of the cell metabolic cycle. Tobin et al. (1984) reported that the biosorption by R. arrhizus was independent of the ionic charge or electrostatic strength and was linearly influenced by the ionic radius, while for Penicillium biomass the biosorption was inversely related to the ionic radius and also on the ionic charge for ions other 78


than uranium (Galun et al., 1987). The biosorption of oxy-anions (molybdate, vandate) was explained on the basis of electostatic attraction to the positively charged functional groups (nitrogen-containing). The metal ion-uncharged molecule interactions can also play an important role in biosorption processes (Siegel et al., 1990). Functional groups such as phosphate, carboxyl, amine, amide and sulfhydryl groups can complex the metal ions (Remacle, 1990; Shumate and Strandberg, 1985). Guibal et al. (1995) indicated that uranyl sorption on A. niger, P. chrysogenum and M. miehei changed the relative band intensities of amine and amide groups in infra-red spectra. The carboxylate and phosphate groups were proposed by Tobin et al. (1984) to be actively involved along with hydroxyl groups. Treen-Sears (1986) reported that uptake of metals depended on the ratio of phosphate to carboxyl groups in the biomass. Tobin et al. (1990), through a series of chemical treatments to R. arrhizus, indicated that phosphate and carboxyl groups were important metal-binding sites. Ion-exchange was believed to be the principal mechanism for metal sequestering. Treen-Sears et al. (1984) also indicated that uranium biosorption was due to ionexchange. Chitin and chiosan present in fungal cells can also sequester metal ions (Muzzarelli, 1972; Tsezos, 1983). Tsezos and Volesky (1982a, 1982b) indicated that biosorption of radionuclides initiated from their association with the nitrogen of the chitin monomer N-actyl-glucosamine. The hydrolysis of the uranium-chitin complex and its deposition causes substantial uranium uptake. Yakubu and Dudeney (1986) indicated that ion-exchange predominated in the uranium biosorption process. Uranyl cations were proposed to replace protons reversibly on the amino acid groups of proteins and glycoproteins, although Muraleedharan and Venkobachar (1990) found that proteins in Ganoderma lucidum, a macro-fungus, do not play any substantial role by themselves in copper uptake. The role of chitin in biosorption was also ruled out, as the chitin fraction of G. lucidum and the fraction devoid of chitin indicated the presence of a free radical. Tsezos and Mattar (1986) and Muzzarelli et al. (1979) showed the presence of this free radical to be associated with the chitin nitrogen, in contrast to the observations of Muraleedharan and Venkobachar (1990). The groups having the free radical were believed to be coordinating with metal and the cellular matrix embedding the free radical opened upon metal uptake. Muraleedharan et al. (1994a) further proposed that structural polysaccharides of the cell wall were probably the main site of interaction and the complexing ligand is rich in oxygen. Most of the metal uptake was due to ion-exchange. Copper ions were observed to be exchanged with calcium and hydrogen ions. The preliminary results of research conducted showed that heavy metal ions are exchanged, not only with calcium and hydrogen ions as observed by Muraleedharan et al. (1994a), but also with magnesium and potassium ions (unpublished results). Avery and Tobin (1993) also observed the displacement of calcium, magnesium and hydrogen ions by biosorption of hard (e.g. Sr, Mn) and soft (e.g. Cd) metals on S. cerevisiae.

General considerations in the use of fungi as biosorbents A number of factors affect the application of a biosorbent in practice: (a) the biosorptive capacity; (b) the availability of the biosorbent; (c) the cost of the biosorbent; (d) the ease of regeneration and subsequent use of the biosorbent; and (e) the ease with which the biosorbent can be used in various reactor configurations. In 79


many instances, the metal removal capacity of fungal biomass has been observed to be greater than that of conventional adsorbents, such as activated carbon, charcoal, and ion-exchange resins and oxides (Azab and Peterson, 1989; Huang et al., 1988a; Niu et al., 1993; Tsezos and Volesky, 1981; Yakubu and Dudeney, 1986). The fungal biomass for producing biosorbent can be obtained from fermentation industries which waste the biomass. For example, the food and beverage industry can serve as a source of yeast (S. cerevisiae); the chemical and pharmaceutical industries can serve as sources of A. niger, T. reessii, R. arrhizus and R. nigricans. The waste biomass of M. meihie, R. arrhizus, A. niger and P. chrysogenum procured from fermentation industries has been found to be effective in biosorption (Guibal et al, 1992; Luef et al., 1991; Mattuschka et al., 1993; Niu et al., 1993; Rao et al., 1993; Tobin et al., 1993). The regeneration and subsequent use of biosorbent impact on the economics of a biosorbent. The fungal biomass can be regenerated using conventional elutants such as dilute acid and carbonates. The biomass of Aspergillus, Penicillium and Saccharomyces after 10 cycles of biosorption/regeneration did not experience significant change in biosorptive capacity (Galun et al., 1983b; Huang et al., 1988b). The cost of procuring/growing the biomass is an important factor in determining the overall economics of the process. The cost of a biosorbent will be significantly lower if it can be obtained from an industrial fermentation process wasting the biomass. Even the cost of culturing the biomass for biosorption has been shown to be competitive with the cost of conventional adsorbent such as activated carbon and ionexchange resins (Kuyucak, 1990). The use of a pelletised or immobilised biomass provides flexibility by enabling the biomass to be used in different reactor configurations. The cost of immobilisation need not be high, as biomass can be immobilised using smaller amounts of support material (Tzezos and Deutschmann, 1990). Thus, biosorption offers an economical and technically feasible option for treatment of wastewater. The biosorption is technically competitive with conventional metal removal technologies, such as precipitation, ion-exchange, reverse osmosis and activated carbon adsorption (Brierley et al., 1986; Kuyucak, 1990). Fungal biosorption is also of natural geochemical importance in the concentration of metals in soils rich in fungi.

References AKTHAR, M.N., SASTRY, K.S. and MOHAN, P.M., 1995, Biosorption of silver ions by processed Aspergillus niger biomass, Biotechnology Letters, 17, 551–556. AKTHAR, MD., N., SASTRY, K.S. and MOHAN, P.M., 1996, Mechanism of metal ion biosorption by fungal biomass, BioMetals, 9, 21–28. AVERY, S.V. and TOBIN, J.M., 1992, Mechanism of strontium uptake by laboratory and brewing strains of Saccharomyces cerevisiae, Applied and Environmental Microbiology, 58, 3883–3889. AVERY, S.V. and TOBIN, J.M., 1993, Mechanism of adsorption of hard and soft metal ions to Saccharomyces cerevisiae and influence of hard and soft ions, Applied and Environmental Microbiology, 59, 2851–2856. AZAB, M.S. and PETERSON, P.J., 1989, The removal of cadmium from water by the use of biological sorbents, Water Science and Technology, 21, 1705–1706. AZAB, M.S., PETERSON, P.J. and YOUNG, T.W.K., 1990, Uptake of cadmium by fungal biomass. Microbios, 62, 23–28. 80

Fungi as Biosorbents BORST-PAUWELS, G.W.F.H., 1981, Ion transport in yeast, Biochimica et Biophysica Acta, 650, 88–127. BRADY, D., STOLL, A., and DUCAN, J.R., 1994, Biosorption of heavy metal cations by nonviable yeast biomass, Environmental Technology, 15, 429–438. BRADY, J.M. and TOBIN, J.M., 1995, Binding of hard and soft metal ions to Rhizopus arrhizus biomass, Enzyme and Microbial Technology, 17, 791–796. BRIERLEY, J.A., BRIERLEY, C.L. and GOYAK, G.M., 1986, AMT-BIOCLAIM™: a new wastewater treatment and metal recovery technology, In: Fundamental and Applied Biohydrometallurgy, LAWRENCE, R.W., BRANION, R.M. R. and EBNER, H.G. (Eds), Elsevier, Amsterdam, The Netherlands, pp. 291–304. DAHLQUIST, F.W., 1978, The meaning of Scatchard and Hill plots. In: Methods in Enzymology XLVIII, HIRS,C.H.W. and TIMASHEFF,S.N. (Eds), Academic Press, New York, pp. 270–299. FERGUSON, C.R., PETERSON, M.R. and JEFFERS, T.H., 1989, Removal of metal contaminants from waste using biomass immobilised in polysulfone beads, Biotechnology in Minerals and Metal Processing, SCHEINER, B.J., DOYLE, F.M. and KAWATRA, S.K. (Eds), Society of Mining Engineers, Littleton, Colorado, pp. 193–199. FOUREST, E., CANAL, C. and Roux, J., 1994, Improvement of heavy metal biosorption b y m y c e l i a l d e a d b i o m a s s ( R h i z o p u s a r r h i z u s , M u c o r m i e h e i , Pe n i c i l l i u m chrysogenum): pH control and cationic activation. FEMS Microbiology Reviews, 14, 325–332. GADD, G.M., 1986, The uptake of heavy metals by fungi and yeasts: the chemistry and physiology of the process and applications for biotechnology, In: Immobilization of Ions by Biosorption, ECCLES, H.H. and HUNT, S. (Eds), Ellis Horwood, Chichester, pp. 135–147. GADD, G.M., 1990, Fungi and yeasts for metal accumulation, In: Microbial Mineral Recovery, EHRLICH, H.L. and BRIERLEY, C.L. (Eds), McGraw-Hill Publishing Company, New York, pp. 249–276. GADD, G.M. and MOWLL, J.L., 1983, The relationship between cadmium uptake, potassium release and viability in Saccharomyces cerevisiae, FEMS Microbiology Letters, 16, 45–48. GADD, G.M., WHITE, C. and DE ROME, L., 1988, Heavy metal and radionuclide uptake by fungi and yeasts, In: Biohydrometallurgy, NORRIS, P.R. and KELLY, D.P. (Eds), Science and Technology Letters, Kew, pp. 421–435. GALUN, M., GALUN, E., SIEGEL, B.Z., KELLER, P., LEHR, H. and SIEGEL, S.M., 1987, Removal of metal ions from aqueous solutions by Pencillium biomass: kinetic and uptake parameters, Water, Air and Soil Pollution, 33, 359–371. GALUN, M., KELLER, P., FELDSTEIN, H., GALUN E., SIEGEL, S. and SIEGEL, B., 1983a, Recovery of uranium (VI) from solution using fungi II. Release from uranium loaded Penicillium biomass, Water, Air and Soil Pollution, 20, 277–285. GALUN, M., KELLER, P. and MALKI, D., 1983b, Removal of uranium (VI) from solution by fungal biomass and fungal wall related biopolymers, Science, 219, 285–286. GALUN, M., KELLER, P., MALKI, D., FELDSTEIN, H., GALUN, E., SIEGEL S. and SIEGEL, B., 1983c, Recovery of uranium (VI) from solution using precultured Pencillium biomass, Water, Air and Soil Pollution, 20, 221–232. GALUN, M., KELLER, P., MALKI, D., FELDSTEIN, H., GALUN, E., SIEGEL, S. and SIEGEL, B., 1984, Removal of uranium (VI) from solution by fungal biomass: inhibition by iron, Water, Air and Soil Pollution, 21, 411–414. GEE, A.R. and DUDENEY, A.W.L., 1988, Adsorption and crystallization of gold at biological surfaces, In: Biohydrometallurgy, NORRIS, P.R. and KELLY, D.P. (Eds), Science and Technology Letters, Kew, pp. 437–451. GUIBAL, E., ROULPH, C. and LE CLOIREC, P.L., 1992, Uranium biosorption by a filamentous fungus Mucor miehei pH effect on the mechanisms and performances of uptake, Water Research, 26, 1139–1145. GUIBAL, E., ROULPH, C. and LE CLOIREC, P.L., 1995, Infrared spectroscopic study of 81

Biosorbents for Metal lons uranyl biosorption by fungal biomass and material of biological origin, Environmental Science and Technology, 29, 2496–2503. HOLAN, Z.R. and VOLESKY, B., 1995, Accumulation of cadmium, lead, and nickel by fungal and wood biosorbents, Applied Biochemistry and Biotechnology, 53, 133–146. HORIKOSHI, T., NAKAJIMA, A. and SAKAGUCHI, T, 1981, Studies on the accumulation of heavy metals in biological system XIX: accumulation of uranium of microorganisms, European Journal of Applied Microbiology Biotechnology, 12, 90–96. HUANG, C., HUANG, C.P. and MOREHART, A.L., 1990, The removal of Cu (II) from dilute aqueous solutions by Saccharomyces cerevisiae, Water Research, 24, 433–439. HUANG, C., HUANG, C.P. and MOREHART, A.L., 1991, Proton competition in Cu (II) adsorption by fungal mycelia, Water Research, 25, 1365–1375. HUANG, C.P., WESTMAN, D., QUIRK, K. and HUANG, J.P., 1988a, The removal of cadmium (II) from dilute aqueous solutions by fungal adsorbent, Water Science and Technology, 20, 369–376. HUANG, C.P., WESTMAN, D., QUIRK, K., HUANG, J.P. and MOREHART, A.L., 1988b, Removal of cadmium (II) from dilute solutions by fungal biomass, Paniculate Science and Technology, 6, 405–419. ITOH, M., YUASA, M. and KOBAYASHI, T., 1975, Adsorption of metal ions on yeast cells at varied cell concentration, Plant Cell Physiology, 16, 1167–1169. JILEK, R., PROCHAZKA, H., KATZER, J. and NEMEC, P., 1975, Properties and development of cultivated biosorbent, Rduy, 23, 282–286. KAPOOR, A. and VIRARAGHAVAN, T., 1995, Fungal biosorption—an alternative treatment option for heavy metal bearing wastewater: a review, Bioresource Technology, 53, 195–206. KAPOOR, A. and VIRARAGHAVAN, T., 1996, Biosorption of heavy metals on fungal biomass, Proceedings of the 4th CSCE Environmental Engineering Speciality Conference, Edmonton, Alberta, 333–343. KHALID, A.M., SHEMSI, A.M., AKTHAR, K. and ANWAR, M.A., 1993, Uranium biosorption by Trichocerma Harzianum entrapped in polyester foam beads, In: Biohydrometallurgical Technologies Vol. II, TORMA, A.E., APEL, M.L. and BRIERLEY, C.L. (Eds), The Minerals, Metals and Materials Society, Warrendale, Pennsylvania, pp. 309–317. KIFF, R.J. and LITTLE, D.R., 1986, Biosorption of heavy metals by immobilized fungal biomass, In: Immobilization of Ions by Biosorption, ECCLES, H.H. and HUNT, S. (Eds), Ellis Horwood, Chichester, West Sussex, England, pp. 71–80. KUREK, E., CZABAN, J. and BOLLAG, J., 1982, Sorption of cadmium by microorganisms in competition with other soil constituents, Applied Environmental Microbiology, 43, 1011– 1015. KUYUCAK, N., 1990, Feasibility of biosorbents application, In: Biosorption of Heavy Metals, VOLESKY, B. (Ed.), CRC Press, Boca Raton, Florida, pp. 372–377. KUYUCAK, N. and VOLESKY, B., 1988, Biosorbents for the recovery of metals from industrial solutions, Biotechnology Letters, 10, 137–142. LAKSHMANAN, V.I., HEINRICH, G., and M C CREADY, R.G.L., 1989, Stabilization of biomass for removal of uranium from dilute solutions by biosorption, In: Biotechnology in Minerals and Metal Processing, SCHEINER, B.J., DOYLE, F.M. and KAWATRA, S.K., (Eds), Society of Mining Engineers Inc., Littleton, Colorado, pp. 187–191. LUEF, E., PREY, T. and KUBICECK, C.P., 1991, Biosorption of zinc by fungal mycelial waste, Applied Microbiology and Biotechnology, 34, 688–692. MATHEICKAL, J.T., IYENGAR, L. and VENKOBACHAR, C., 1991, Sorption and desorption of Cu (II) by Ganoderma Lucidum, Water Pollution Research Journal of Canada, 26, 187–200. MATTUSCHKA, B., JUNGHANS, K. and STRAUBE, G., 1993, Biosorption of metals by waste biomass, In: Biohydrometallurgical Technologies Vol. II, TORMA, A.E., APEL, M.L. and BRIERLEY, C.L. (Eds), The Minerals, Metals and Materials Society, Warrendale, Pennsylvania, pp. 125–132. 82

Fungi as Biosorbents MAY, H., 1984, Biosorption by Industrial Biomass, M. Eng. Thesis, McGill University. Montreal, Quebec. MULLEN, M.D., WOLF, D.C., BEVERIDGE, T.J. and BAILEY, G.W., 1992, Sorption of heavy metals by the soil fungi Aspergillus niger and Mucor rouxxi. Soil Biology and Biochemistry, 24, 129–135. MURALEEDHARAN, T.R. and VENKOBACHAR, C., 1990, Mechanism of biosorption of copper (II) by Ganoderma lucidum, Biotechnology and Bioengineering, 35, 320–325. MURALEEDHARAN, T.R., IYENGAR, L. and VENKOBACHAR, C., 1994a, Further insight into the mechanism of biosorption of heavy metals by Ganoderma lucidum, Environmental Technology, 15, 1015–1027. MURALEEDHARAN, T.R., PHILLIP, L., IYENGAR, L. and VENKOBACHAR, C., 1994b, Application studies of biosorption for monazite processing industry effluents, Bioresource Technology, 49, 179–186. MUZZARELLI, R.A.A., 1972, Chitin, Pergamon Press, London. MUZZARELLI, R.A.A., TANFANI, F. and EMANUELLI, M., 1981, The chelating ability of chitinous materials from Streptomyces, Mucor rouxii, Phycomcyes blakesleeanus, and Choanephora cucurbitarum, Journal of Applied Biochemistry, 3, 322–327. MUZZARELLI, R.A.A., TANFANI, F. and SCARPINI, G., 1980a, Chelating,film-forming, and coagulating ability of the chitosan-glucan complex from Aspergillus niger industrial wastes, Biotechnology and Bioengineering, 22, 885–896. MUZZARELLI, R.A.A., TANFANI, F., SCARPINI, G. and MUZZARELLI, M.G., 1979, ESR characterization of chitin and chitosans, Biophysical Research Communications, 89, 706–712. MUZZARELLI, R.A.A., TANFANI, F., SCARPINI, G. and TUCCI, E., 1980b, Removal and recovery of cupric and mercuric ions from solutions using chitosan-glucan from Aspergillus niger, Journal of Applied Biochemistry, 2, 54–59. NAKAJIMA, A. and SAKAGUCHI, T., 1986, Selective accumulation of heavy metals by microorganisms, Applied Microbiology and Biotechnology, 24, 59–64. NIU, H., Xu, S.X., WANG, J.H. and VOLESKY, B., 1993, Removal of lead from aqueous solutions by Pencillium biomass, Biotechnology and Bioengineering, 42, 785–787. NORRIS, P.R. and KELLY, D.P., 1977, Accumulation of cadmium and cobalt by Saccharomyces cerevisiae, Journal of General Microbiology, 99, 317–324. NORRIS, P.R. and KELLY, D.P., 1979, Accummulation of metals by bacteria and yeast, Developments in Industrial Microbiology, 20, 299–308. PAKNIKAR, K.M., PALNITKAR, U.S. and PUANIK, P.R., 1993, Biosorption of metals from solution by mycelial waste Penicillium chrysogenum , In: Biohydrometallurgical Technologies Vol. II, TORMA, A.E., APEL, M.L. and BRIERLEY, C.L. (Eds), The Minerals, Metals and Materials Society, Warrendale, Pennsylvania, pp. 125–132. PARKIN, M.J. and Ross, I.S., 1986a, The specific uptake of manganese in yeast Canadis utilis, Journal of General Microbiology, 132, 2155–2160. PARKIN, M.J. and Ross, I.S., 1986b, The regulation of Mn2+ and Cu2+ uptake in cells of the yeast Canadida utilis grown in continuous culture, FEMS Microbiology Letters, 37, 59–62. PRYFOGLE, P.A., MAIERS, D.T., WICHLUCZ, P.L. and WOLFRAM, J.H., 1989, Biosorption of Molybdenum and chromium, In: Biotechnology in Minerals and Metal Processing, SCHEINER, B.J. DOYLE, F.M. and KAWATRA, S.K. (Eds), Society of Mining Engineers, Littleton, Colorado, pp. 201–207. RAO, C.R.N., IYENGAR, L. and VENKOBACHAR, C., 1993, Sorption of copper (II) from aqueous phase by waste biomass, Journal of Environmental Engineering, 119, 369–377. REMACLE, J. 1990, The cell wall and heavy metals, In: Biosorption of Heavy Metals, VOLESKY, B. (Ed.), CRC Press, Boca Raton, Florida, pp. 83–92 ROME, L. DE and GADD, G.M., 1987, Copper adsorption by Rhizopus arrhizus, Claddosporium resina, and Pencillium italicum, Applied Microbiology and Biotechnology, 26, 84–90. 83

Biosorbents for Metal lons ROME, L. DE and GADD, G.M., 1991, Use of pelleted and immobilised yeast and fungal biomass for heavy metal and radionuclide recovery, Journal of Industrial Microbiology, 7, 97–104. Ross, I.S. and TOWNSLEY, C.C., 1986, The uptake of heavy metals by filamentous fungi, In: Immobilization of Ions by Biosorption, ECCLES, H. and HUNT, S. (Eds), Ellis Horwood, Chichester, pp. 49–57. SAG, Y.,OZER, D. and KUTSAL, T., 1995, A comparative study of the biosorption of lead (II) ions to Z. ramigera and R. arrihizus, Process Biochemistry, 30, 169–174. SCATCHARD, G., 1949, The attractions of proteins for small molecules and ions. Annals of New York Academy of Sciences, 51, 660–672. SHUMATE, S.E., STRANDBERG, G.W. and PARROTT, J.R., Jr., 1978, Biological removal of metal ions from aqueous process streams, Biotechnology and Bioengineering Symposium, 8, 13–20. SHUMATE, S.E. and STRANDBERG, G.W., 1985, Accumulation of metals by microbial cells, In: Comprehensive Biotechnology Vol. 4, ROBINSON, C.W. and HOWELL, J.A. (Eds), Pergamon Press, New York, 235–247. SIEGEL, S., KELLER, P., GALUN, M., LEHR, H., SIEGEL, B. and GALUN, E., 1986, Biosorption of lead and chromium by Pencillium preparations, Water, Air and Soil Pollution, 27, 69–75. SIEGEL, S.M., GALUN, M., KELLER, P., SIEGEL, B.Z. and GALUN, E., 1987, Fungal biosorption: a comparative study of metal uptake by Penicillium and Cladosporium, In: Metal Speciation, Separation and Recovery, PATTERSON, J.W. and PASSING, R. (Eds), Lewis Publishers, Chelsea, Michigan, pp. 339–361. SIEGEL, S.M., GALUN, M. and SIEGEL, B.Z., 1990, Filamentous fungi as metal biosorbents: a review, Water, Air and Soil Pollution, 53, 335–344. SOMERS, E., 1963, The uptake of copper in fungal spores, Annals of Applied Biology, 5, 425–437. STRANDBERG, G.W., SHUMATE, S.E. and PARROTT, J.R., 1981, Microbial cells as biosorbents for heavy metals: accumulation of uranium by Saccharomyces cerevisiae and Pseudomonas aeruginosa, Applied and Environmental Microbiology, 41, 237–245. TING, Y.P. and TEO, W.K., 1994, Uptake of cadmium and zinc by yeast: effects of co-metal ion and physical/chemical treatments, Bioresource Technology, 50, 113–117. TOBIN, J.M., COOPER, D.G. and NEUFIELD, R., 1984, Uptake of metal ions by Rhizopus arrhizus biomass, Applied and Environmental Microbiology, 47, 821–824. TOBIN, J.M., COOPER, D.G. and NEUFIELD, R.J., 1987, Influence of anions on metal adsorption by R.. arrhizus biomass, Biotechnology and Bioengineering, 30, 882–886. TOBIN, J.M., COOPER, D.G. and NEUFIELD, R.J., 1988, The effects of cation competition on metal adsorption by Rhizopus arrhizus biomass, Biotechnology and Bioengineering, 31, 282–286. TOBIN, J.M., COOPER, D.G. and NEUFIELD, R.J., 1990, Investigation of the mechanism of metal uptake by denatured Rhizopus arrhizus biomass, Enzyme and Microbial Technology, 12, 591–595. TOBIN, J.M., L’HOMME, B. and Roux, J.C., 1993, Immobilisation protocols and effects on cadmium uptake by Rhizopus arrhizus biosorbents, Biotechnology Techniques, 7, 739–744. TOWNSLEY, C.C. and Ross, I.S., 1985, Copper uptake by Penicillium spinulosum, Microbios, 44, 125–132. TOWNSLEY, C.C. and Ross, I.S., 1986, Copper uptake in Aspergillus niger during batch growth and in non-growing mycelial suspensions, Experimental Mycology, 10, 281–288. TOWNSLEY, C.C., Ross, I.S. and ATKINS, A.S., 1986a, Biorecovery of metallic residues from various industrial effluents using filamentous fungi, In: Fundamental and Applied Biohydrometallury, LAWRENCE, R.W., BRANION, R.M. R. and EBNER, H.G. (Eds), Elsevier, Amsterdam, The Netherlands, pp. 279–289. TOWNSLEY, C.C., Ross, I.S. and ATKINS, A.S., 1986b, Copper removal from a simulated leach effluent using filamentous fungus Trichoderma virdie, In: Immobilization of Ions by Biosorption, EDDLES, H. and HUNT, S. (Eds), Ellis Horwood, Chichester, pp. 159–170. 84

Fungi as Biosorbents TREEN-SEARS, M.E., 1986, Propagation and characterization of Rhizopus biosorbents, PhD Thesis, McGill University, Montreal, Quebec. TREEN-SEARS, M.E., VOLESKY, B. and NEUFELD, R.J., 1984, Ion exchange/ complexion of uranyl ion by Rhizopus biosorbent, Biotechnology and Bioengineering, 26, 1323–1329. TSEZOS, M., 1983, The role of chitin in uranium adsorption by R. arrhizus, Biotechnology and Bioengineering, 25, 2025–2040. TSEZOS, M., 1984, Recovery of uranium from biological adsorbents. Desorption equilibrium, Biotechnology and Bioengineering, 26, 973–981. TSEZOS, M., BAIRD, M.H. I. and SCHEMILT, L.W., 1986, Adsorptive treatment with microbial biomass of 226Ra containing wastewaters, Chemical Engineering Journal, 32, B29-B38. TSEZOS, M. and DEUTSCHMANN, A.A., 1990, An investigation of engineering parameters for the use of immobilized biomass particles in biosorption , Journal of Chemical Technology and Biotechnology, 48, 29–39. TSEZOS, M. and MATTAR, S., 1986, A further insight into the mechanism of biosorption of metals, by examining chitin EPR spectra, Talanta, 33, 225–232. TSEZOS, M. and NOH, S.H., 1984, Extraction of uranium form sea water using biological origin adsorbents, Canadian Journal of Chemical Engineering, 62, 559–561. TSEZOS, M. and VOLESKY, B., 1981, Biosorption of uranium and thorium, Biotechnology and Bioengineering, 23, 583–604. TSEZOS, M. and VOLESKY, B., 1982a, The mechanism of uranium biosorption by Rhizopus arrhizus, Biotechnology and Bioengineering, 24, 385–401. TSEZOS, M. and VOLESKY, B., 1982b, The mechanism of thorium biosorption by Rhizopus arrhizus, Biotechnology and Bioengineering, 24, 955–969. VOLESKY, B., 1991a, Biosorption by fungal biomass, In: Biosorption of Heavy Metals, VOLESKY, B. (Ed.), CRC Press, Boca Raton, Florida, pp. 139–172. VOLESKY, B., 1991b, Biosorption and biosorbent, In: Biosorption of Heavy Metals, VOLESKY, B. (Ed.), CRC Press, Boca Raton, Florida, pp. 3–5. VOLESKY, B., 1994, Advances in biosorption of metals: selection of biomass types, FEMS Microbiology Reviews, 14, 291–302. VOLESKY, B. and MAY, H., 1995, Biosorption of heavy metals by Saccharomyces cerevisiae, Applied Microbiology and Biotechnology, 42, 797–806. VOLESKY, B., MAY, H. and HOLAN, Z.R., 1993, Cadmium biosorption by Saccharomyces cerevisiae, Biotechnology and Bioengineering, 41, 826–829. WALES, D.S. and SAGAR, B.F., 1990, Recovery of metal ions by microfungal filters, Journal of Chemical Technology and Biotechnology, 49, 345–355. WEBER, W.J., Jr., 1972, Adsorption, In: Physicochemical Processes for Water Quality Control, WEBER, W.J., Jr., (Ed.), John Wiley and Sons, New York, pp. 199–260. WHITE, C. and GADD, G.M., 1987, The uptake and cellular distribution of zinc in Saccharomyces cerevisiae, Journal of General Microbiology, 133, 727–737. WHITE, C. and GADD, M., 1990, Biosorption of radionuclides by fungal biomass, Journal of Chemical Technology and Biotechnology, 49, 331–343. YAKUBU, N.A. and DUDENEY, A.W. L., 1986, Biosorption of uranium with Aspergillus niger, In: Immobilization of Ions by Biosorption, EDDLES, H.H. and HUNT, S. (Eds), Ellis Horwood, Chichester, pp. 183–200. ZAJIC, J.E. and CHIU, Y.S., 1972, Recovery of heavy metals by microbes, Developments in Industrial Microbiology, 13, 91–100. ZHOU, J.L. and KIFF, R.J., 1991, The uptake of copper from aqueous solution by immobilized fungal biomass, Journal of Chemical Technology and Biotechnology, 52, 317–330. ZOUBOULIS, A.I. and MATIS, K.A., 1993, Flotation as a bioseparation process for fungi removal, Biotechnology Techniques, 7, 867–872.

85

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). 87


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 88

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 89


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


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 91


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)


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 96


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. 97


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

98

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 99


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


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

101


Figure 5.10 Typical uranium biosorption breakthrough curve

Figure 5.11 Typical uranium elution curve

102


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 103

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

104


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


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


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


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).

References AVERY, S.V. and TOBIN, J.M., 1992, Mechanisms of strontium uptake by laboratory and brewing strains of S. cerevisiae, Applied and Environmental Microbiology, 58, 3883–3889. AVERY, S.V. and TOBIN, J.M., 1993, Mechanism of adsorption of hard and soft metal ions to Saccharomyces cerevisiae and influence of hard and soft ions, Applied and Environmental Microbiology, 59, 2851–2856. BEVERIDGE, T.J. and MURRAY, R.G. E., 1980, Sites of metal deposition in the cell wall of Bacillus subtilis, Journal of Bacteriology, 141, 876–883. BRADY, J.M. and TOBIN, J.M., 1994, Adsorption of metal ions by Rhizopus arrhizus biomass characterisation studies, Enzyme Microbial Technology, 19, 671–675. BRIERLEY, J.A., GOYAK, G.M. and BRIERLEY, C.L., 1986, Considerations for commercial use of natural products for metal recovery, In: Immobilisation of Ions by Biosorption, Ellis Horwood, Chichester, pp. 105–117. BRIERLEY, J.A. and VANCE, D.B., 1988, Recovery of precious metals by microbial biomass, In: Biohydrometallurgy, Proceedings of the International Symposium Warwick 1987, NORRIS, D.R. and KELLY, D.P. (Eds), Science and Technology Letters, Kew, pp. 477–485. 110

Lanthanides, Actinides and Related Materials DARNALL, D.W., M C PHERSON, R.A. and GARDEA-TORVESDEY, J., 1989, Metal recovery from geothermal waters and ground waters using immobilised algae, In: Biohydrometallurgy, Proceedings of the International Conference at Jackson Hole, Wyoming, CANMET, Ottawa, Ontario, Canada, pp. 341–348. DIELS, L., 1989, Accumulation and precipitation of Cd and Zn ions by Alcaligenes eutrophus strains, In: Biohydrometallurgy, Proceedings of the International Conference at Jackson Hole, Wyoming, CANMET, Ottawa, Ontario, Canada, pp. 369–377. DIELS, L., VAN ROY, S., CORBISIER, P., NUYTS, S, MACASKIE, L., BONTHRONE, K., TSEZOS, M., REMOUDAKI, E., ANGELATOU, V., PUMPEL, T., PERNFUSS, B., SCHINNER, F., HUMMEL, A., ECKARD, L., GLOMBITZA, F., BOYD, D. and ESPRIT, M., 1995, Modelling of Generic, Biochemical, Cellular and Microenvironmental Parameters determining Bacterial Sorption and Mineralisation Processes for Recuperation of Heavy or Precious Metals. Summary Report for publication of BriteEuram 5350 Project, European Commission BRITE-EURAM Programme, DG XII . DIELS, L., WOUTERS, H., MACASKIE, L., HUMPHREYS, P., WOEBKING, H., KESZTHELY, Z., TSEZOS, M., REMOUDAKI, E., PUMPEL, T., PERNFUSS, B., SCHINNER, F., ECKARD, L., GLOMBITZA, F. and ESPRIT, M., 1996, Removal and Recovery of Heavy Metals from WasteWater by Sand Filters Inoculated with Metal Biosorbing or Bioprecipitating Bacteria, RTD—Project for Industrial and Materials Technologies, BRITE-EURAM 95–1610 CEC, DG—XII. EHRLICH, H.L. and BRIERLEY, C.L., 1990, Microbial Mineral Recovery, McGraw-Hill, New York. GADD, G.M., 1992, Biosorption, Journal of Chemical Technology and Biotechnology, 55, 302–304. GEORGOUSIS, Z., 1990, The Loss of Uranium Uptake Capacity of Rhizopus arrhizus Immobilised Biomass on Recycle, M. Eng. Thesis, McMaster University, Canada. GLOMBITZA, F., ISKE, U., GWENNER, CHR. and KRISHNAS, M.V., 1984, Biosorption of mercury, Acta Biotechnology, 4, 281–284. GREENWOOD, N.N. and EARNSHAW, A., 1993, Chemistry of the Elements, Pergamon Press, Cambridge. HAREL, P., MIGNOT, L. and JUNTER, G.A., 1995, Cd removal from dilute aqueous solutions by beads of polysaccharide gels usable for microbial cell immobilization, International Seminar on Biosorption and Bioremediation, Merin, Czech Republic, October 1–4. HUANG, C., HUANG, C.P. and MOREHART, A.L., 1991, Proton competition in Cu(II) adsorption by fungal mycelia, Water Research, 25, 1365–1375. KIFF, R.S. and LITTLE, D.R., 1986, Biosorption of heavy metals by immobilised fungal biomass, In: Immobilisation of Ions by Biosorption, Ellis Horwood, Chichester, pp. 71–80. LOVLEY, D.R. and PHILLIPS, E.J. P., 1992, Reduction of uranium by Desulfovibrio desulfuricans, Applied and Environmental Microbiology, 58, 850–856. LOVLEY, D.R., PHILLIPS, E.J. P., GORBY, Y.A. and LANDA, Y., 1991, Microbial uranium reduction, Nature, 350, 413–416. MACASKIE, L.E., 1991, Application biotechnology to the treatment of wastes from the nuclear fuel industry, CRC Critical Review in Biotechnology, 11, 41–112. MACASKIE, L.E. and DEAN, A.C. R., 1984, Cadmium accumulation by immobilised cells of Citrobacter sp., Environmental Technology Letters, 5, 177–186. MACASKIE, L.E. and DEAN, A.C. R., 1985, Strontium accumulation by immobilised cells of a Citrobacter sp., Biotechnology Letters, 7, 457–462. PHILLIPS, E.J.P., LANDA, E.R. and LOVLEY, D.R., 1995, Remediation of uraniumcontaminated soils with bicarbonate extraction and microbial U (VI) reduction, Journal of Industrial Microbiology, 4, 203–207. PUMPEL, T. and SCHINNER, F., 1986, Silver tolerance and silver accumulation of microorganisms from soil materials of a silver mine, Applied Microbiology and Biotechnology, 24, 244–247. 111

Biosorbents for Metal lons RYON, A.D., BOND, W.D., HURST, F.J., SCHEITLIN, F.M. and SEELEY, F.G., 1982, Investigation of nitric acid for removal of noxious radionuclides from uranium ore or mill tailings, In: Uranium Mill Tailing Management, Proceedings of two NEA Workshops on geomorphological evaluation of the long-term stability of uranium tailings disposal sites, OCDE Press, Paris, France. SENES CONSULTANTS, 1985, Evaluation of Biomass for Uranium Recovery from Process Strains, Energy Mines and Resources Canada, CANMET Project Report 4–9173. STAMBERG, J., JILEK, K. and SLAMBERG, K., 1975, Proceedings of the Czechoslovak Atomic Energy Symposium, PraComiku Banskeho, Prumyslu, Czechoslovakia. TOBIN, J.M., WHITE, C. and GADD, G.M., 1994, Metal accumulation by fungi: application in environmental biotechnology, Journal of Industrial Microbiology, 13, 126–130. TOLLEY, M.R., MACASKIE, L.E., MOODY, J.C. and STRANDLING, G.N., 1991, Actinide and lanthanum accumulation by immobilised cells of Citrobacter sp., Proceedings of the 201st Meeting American Chemical Society, 31, 213–216. TSEZOS, M., 1984, Recovery of uranium from biological adsorbents. Desorption equilibrium, Biotechnology and Bioengineering, 26, 973–981. TSEZOS, M., 1985, The selective extraction of metals from solution by microorganisms. A brief overview, Canadian Metallurgical Quarterly, 24, 141–144. TSEZOS, M., 1990, Engineering aspects of metal binding by biomass, In: Microbial Mineral Recovery, EHRLICH, H.L. and BRIERLEY, C.L. (Eds), McGraw Hill, New York, pp. 325–329. TSEZOS, M., BAIRD, M.H. I. and SHEMILT, L.W., 1986a, Adsorptive treatment with microbial biomass of 226Ra containing waste-waters, Chemical Engineering Journal, 32, B29–38. TSEZOS, M., BAIRD, M.H. I. and SHEMILT, L.W., 1986b, The elution of radium adsorbed by microbial biomass, Chemical Engineering Journal, 34, 3, B57–64. TSEZOS, M., BAIRD, M.H. I. and SHEMILT, L.W., 1986c, The kinetics of radium biosorption, Chemical Engineering Journal, 33, 3, B35–41. TSEZOS, M., BAIRD, M.H. I. and SHEMILT, L.W., 1987, The use of immobilised biomass to remove and recover radium from Elliot Lake uranium tailings streams, Hydrometallurgy, 17, 357–368. TSEZOS, M. and DEUTSCHMANN, A.A., 1990, An investigation of the engineering parameters for the use of immobilised biomass particles in biosorption, Journal of Chemical Technology and Biotechnology, 48, 29–39. TSEZOS, M. and DEUTSCHMANN, A.A., 1992, The use of a mathematical model for the study of important parameters in immobilised biomass biosorption , Journal of Chemical Technology and Biotechnology, 53, 1–12. TSEZOS, M., GEORGOUSIS, Z. and REMOUDAKI, E., 1996c, The mechanism of aluminium interference on uranium biosorption by R. arrhizus, Biotechnology and Bioengineering, in press. TSEZOS, M. and KELLER, D., 1983, Adsorption of radium-226 by biological origin adsorbents, Biotechnology and Bioengineering, 25, 201–215. TSEZOS, M. and MC CREADY, R.G. L., 1989, The continuous recovery of uranium from biologically leached solutions using immobilised biomass, Biotechnology and Bioengineering, 34, 10–17. TSEZOS, M. and M CCREADY, R.G. L., 1991, The pilot plant testing of the continuous extraction of radionuclides using immobilised biomass, In: Environmental Biotechnology for Waste Treatment, SAYLER, G., Fox, R. and BLACKBURN, J. (Eds), Plenum Press, New York, pp. 249–260. TSEZOS, M., NOH, S.H. and BAIRD, M.H. I., 1988, A batch reactor kinetic model for uranium biosorption using immobilized biomass, Biotechnology and Bioengineering, 32, 545–553. TSEZOS, M., REMOUDAKI, E. and ANGELATOU, V., 1995, A systematic study on 112

Lanthanides, Actinides and Related Materials equilibrium and kinetics of biosorptive accumulation, International Biodeterioration and Biodegradation, 35,1–3,129–153. TSEZOS, M., REMOUDAKI, E. and ANGELATOU, V., 1996a, A study of the competing ions effects in the biosorption of metals, International Biodeterioration and Biodegradation, 36, 19–28. TSEZOS, M., REMOUDAKI, E. and ANGELATOU, V., 1996b, Biosorption sites of selected metals using electron microscopy, Comparative Biochemistry and Physiology. in press. TSEZOS, M. and VOLESKY, B., 1981, Biosorption of uranium and thorium, Biotechnology and Bioengineering, 23, 583–604. TSEZOS, M. and VOLESKY, B., 1982a, The mechanism of thorium biosorption by R. arrhizus, Biotechnology and Bioengineering, 24, 955–969. TSEZOS, M. and VOLESKY, B., 1982b, The mechanism of uranium biosorption by R. arrhizus, Biotechnology and Bioengineering, 24, 385–401. TSEZOS, M. and WANG, X., 1991, A study on lindane biosorption and biodegradation interactions , Biotechnology Forum Europe, 8, 120–125. WATKINS, J.W., ELDER, R.C., GREEN, B. and DARNALL, W., 1987, Determination of gold binding in an algal biomass using EXAFS and XANES spectroscopies, Inorganic Chemistry, 26, 1147–1151. WEST, R., 1976, National nuclear reactors. The Oklo phenomenon, Journal of Chemical Education, 53, 336–340.

113

6

Scavenging Trace Concentrations of Metals C.J.BANKS

Introduction As noted in the preceding chapters, the process of metal removal, by a range of microorganisms, has been intensively studied at a cellular level. Indeed, several potential mechanisms could be exploited, depending on the metal, its concentration, the presence of other materials, the type of microbe and its metabolic state. There are, however, two points which are important to consider in the development of a removal system capable of scavenging metals at low concentration: that the microbe must be in intimate contact with the metal ions in solution for effective sorption to take place, and that the biomass does not have an infinite capacity for the sorption of metals. A system aimed at scavenging metals from solution must, therefore: • present the biomass to the metal-laden solution in an optimal way so as to ensure ideal physical conditions for the maintenance of the biomass and the sorption process • permit the biomass to be harvested once its sorptive capacity has been reached, or allow regeneration of the biomass with the concurrent generation of a concentrated metal-laden stream. Unfortunately, these are not the only problems which need to be solved for exploitation of the biosorption phenomena as a useful weapon in the arsenal of the environmental engineer. Effluent streams are rarely of uniform strength; they frequently show wide, and often unpredictable, variations in flow and their composition may be highly variable. For a treatment system to be successful it must be robust and capable of withstanding hydraulic, concentration and potential shock loads of interfering materials. Ideally, the system should also offer potential for recovery of the pollutant metal and not be just a means of concentrating the pollutant for disposal by another route. The aim of this chapter is to look at possible practical solutions that satisfy the above conditions. 115


Coincidental sorption systems For the purposes of this chapter, these can be considered as treatment processes which have been designed primarily for the removal of organic material by biological oxidation, but coincidentally remove metal. The most common examples of such systems are: suspended biomass processes, such as activated sludge; fixed film processes, such as biological filters; and symbiotic processes, such as stabilisation ponds and wetlands. In all such systems, metals are removed from aqueous solution and become associated with the biomass, either by uptake into the cellular material (absorption), which is usually an active metabolically linked process, or by extracellular passive association with exposed biomass surfaces (adsorption). The ability of these systems to remove metals from a continuous aqueous stream depends on their overall sorptive capacity and replenishment of this through new biomass growth. In none of these systems are active control measures taken to recover metals from the biomass and to return that same biomass to the treatment system in an unsaturated (with respect to metals) form. Hence, if the metals loading to the system per unit of time is greater than the new sorptive capacity generated, in the same unit of time through growth processes, then breakthrough into the effluent from the plant will occur. All these processes will also produce surplus biomass which is metal-laden, this being the only exit route other than through the aqueous effluent stream.

Activated sludge process Studies on the metal mass balance through the activated sludge process have reported substantial metal removal (Brown et al., 1973; Oliver and Cosgrove, 1974; Stoveland et al., 1979; Lester, 1981), and the process is probably the best documented example of a biological heavy metals removal system. In the process, a retained biomass population, grown under aerobic conditions, oxidises the dissolved and/or suspended organic matter present in the wastewater. Metal removal takes place by adsorption and absorption of the dissolved metals onto the activated sludge floc (Neufield and Hermann, 1975; Sterritt and Lester, 1981). There is also potential for some retention of precipitated metals by entrapment within the extracellular matrix excreted by the flocculant microorganisms and for volatilisation of metal to the atmosphere (Brown and Lester, 1979). Of the proposed mechanisms, the role of surface microbial polysaccharides has been documented best (Brown and Lester, 1979, 1982). Indeed, specific metal ion removal processes have been suggested in which polysaccharide polymer production is stimulated and then used in a separate stage for heavy metal removal from solution (Dugan, 1970; Dugan and Pickrum, 1972; Norberg and Rydin, 1984). As early as 1968 it was concluded (Friedman and Dugan, 1968) that observed metal uptake by activated sludge could not account for all the metal removal exhibited, and that intracellular accumulation also played a role in the immobilisation of metal ions in the process. A number of operating parameters are cited in the literature as having an influence on metals removal in the process, i.e. sludge age, sludge volume index (SVI), suspended solids removal efficiency, dissolved oxygen concentration, mixed liquor suspended solids (MLSS), effluent suspended solids (ESS), and effluent chemical 116

Scavenging Trace Concentrations of Metals

oxygen demand. Because of the different metals removal mechanisms inherent to the system, it is likely that all of the above will influence the overall metals removal efficiency of the plant, as will both the volumetric and biomass metals loading rate. If it is assumed that the system has a finite metals solid phase retention capacity, then the sludge age must be the single most important factor governing overall metals removal. This must be the case when the influent wastewater metals concentration is such that it will cause saturation of biological removal mechanisms and/or the saturation coefficient of any insoluble precipitated metal to be exceeded. This can be demonstrated by reference to Figures 6.1 and 6.2, in which four laboratory scale activated sludge plants were acclimatised at sludge ages of 5, 10, 15 and 20 days on a feed containing only background concentrations of zinc for a period of 60 days before being exposed to an influent zinc concentration of 5 mg/l. Over the next 5 days, the activated sludge mixed liquor zinc concentration (mg/kg dry weight) increased until saturation values were reached (Figure 6.1). Over the same period the

Figure 6.1 Activated sludge zinc concentration in response to an influent zinc concentration of 5 mg/l introduced on day 61 to a previously unexposed sludge. Reactors 1–4 were operated at sludge ages of 5, 10, 15 and 20 days respectively

117


Figure 6.2 Effluent zinc concentration in response to an influent zinc concentration of 5 mg/l introduced on day 61 to a previously unexposed sludge. Reactors 1–4 were operated at sludge ages of 5, 10, 15 and 20 days respectively

effluent zinc concentration remained low (less than 2 mg/l) and then increased sharply to 5 mg/l between days 66 and 68 (Figure 6.2). The steady state sludge zinc concentrations corresponded to 11.4, 14.0, 17.1 and 16.8 mg/kg at sludge ages of 5, 10, 15 and 20 days respectively with MLSS values of 2.2, 2.98, 4.14 and 5.29 g/l. Simple calculation reveals the following daily removal capacities, based on maintaining steady state MLSS concentrations under saturation influent zinc concentrations: 5 day sludge age 10 day sludge age 15 day sludge age 20 day sludge age

5.016 4.172 5.302 4.430

mg/m3 mg/m3 mg/m3 mg/m3

aeration aeration aeration aeration

capacity capacity capacity capacity

The fact that the values obtained are of the same order is due to the lower sorption capacity of the young sludge being compensated by the higher sludge take-off rate. 118


The calculation reveals that a typical activated sludge plant, working at a sludge age of 15 days with a sewage retention time of 8 h, could tolerate an influent zinc concentration of 1.76 mg/l without reaching its sludge sorptive saturation capacity. The results presented are of course from a simple system in which only one metal was present in appreciable concentrations. They do, however, illustrate the capability of suspended growth biological sewage treatment systems to perform well as scavengers of metals in a coincidental manner.

Biological filters Metals removal in fixed film systems, as typified by biological filters, is less well documented than in the more easily studied suspended growth systems. In principle, the same limitations apply in that the biofilm will have a saturation capacity per unit of dry weight and, under steady state conditions, that capacity can only be replenished by removal of biomass from the system. As conventional biological filters usually have a much lower overall growth yield, as a result of endogenism and secondary feeding activities, it would be anticipated that their metals removal capabilities would be impaired as the biofilm became saturated, and that this would occur at influent metal concentrations lower than those calculated above for activated sludge. In high rate systems, which exhibit high shearing rates of biomass from its support medium, the metals removal capability would be expected to equal that of a suspended growth system. Again, a number of potential removal mechanisms may exist, including precipitation of insoluble metal hydroxides within the biofilm, and this may influence the overall removal capability. This has been demonstrated in relation to metals removal within a denitrifying biofilm. Under conditions of reduced dissolved oxygen concentrations, nitrate is used as a terminal electron acceptor in oxidative processes. This leads to an increase in alkalinity together with a rise in pH and, under these conditions, the solubility of many metals is reduced, enhancing the possibility for metal entrapment within the anoxic region of the biofilm. Liehr (1995) reported this mechanism and derived a predictive model for it. The same phenomena may exist in flocs as well as films and reinforce the concept that metals removal within mixed microbial populations is by a number of different mechanisms, of which sorption is but one.

Wetlands Over recent years interest in the use of wetlands for the treatment of wastewater has been stimulated by the growing number of successful applications to a wide variety of wastewater types (Hammer, 1989). These include wetlands for the removal of trace metals through both aerobic and anaerobic biological processes. In principle, aerobic processes rely on sorption phenomena, whereas anaerobic systems work through the microbial mediated sulphate reduction process which neutralises pH and precipitates metals. Because of the complex nature of the substratum, drainage pattern, degree of aeration, and nutrient availability within a wetlands system, it is difficult to estimate the metals removal capacity which may be mediated by biological means. In effect, the entire system acts as a sink for metals, but the capacity of that sink is dependent on its mineralogy and carbon content, the reactive surface area, and hydraulic 119


conductivity, as well as the type and valency state of the metals in the influent stream. Iron removal primarily occurs through the oxidation of ferrous iron to the ferric form and its subsequent removal as a ferric oxyhydroxide with concurrent acid production (Weider, 1989; Eger, 1994). In a neutral pH trace-metal-containing wastewater, removal occurs primarily in the aerobic zone through a variety of reactions, including adsorption, chelation and ion exchange (Eger, 1994), involving a range of organic materials such as algae, bacteria, detritus and organic layer coatings on silicate clay or other mineral surfaces. Ultimately, sedimentation and coagulation of particulate and colloidal matter result in metal fixation in bottom sediments. Surface vegetation, which is important in maintaining oxygenation of the root zone and hydraulic conductivity of the substratum, is not a major sink for metals through absorptive processes. Total uptake by plant species has not been shown to correlate with total metal concentrations in the sediments, and the order of concentrations in plant tissues was roots>rhizomes>non-green leaves>green leaves; most of the metal taken up by the plant being retained within the roots (Dunbabin and Bowmer, 1992). The harvesting of shoots to remove metals from wetlands is not usually considered worthwhile because of the low concentrations of metals found in green tissues. In the peat bed systems used by Eger (1994) for nickel removal from a neutral mine drainage, the level of nickel increased 10-fold over a two year period to a concentration of 1000 mg/kg, while the concentration in the leaves of rooted vegetation was only 15 mg/kg. With the use of non-living biosorbents (peat, spent mushroom compost, rice hulls, composted municipal wastes and composted garden wastes) in wetland systems, there again must be a finite adsorptive capacity. These organic materials are, however, slowly broken down and, if an anaerobic sediment is allowed to develop, insoluble metal sulphides will be formed if the aqueous phase is rich in sulphates. In practice, it could be envisaged that the surface layers of such wetland systems are replenished with fresh adsorbents at a rate equal to that of their natural decay and sulphide immobilisation in bottom sediments. The use of wetlands as metal scavenging systems is still an infant technology: the systems are complex in their mechanism but biosorption undoubtedly plays a major role. The disadvantage appears to lie in the fact that they are only sinks for metals, offering little opportunity for recovery and, unless properly constructed, operated and maintained, a potential source of future groundwater contamination.

Waste stabilisation ponds As detailed in Chapter 2, the biosorptive capacity of algae for metals is well known. Algae play an important role in low-cost treatment systems in the form of facultative wastewater stabilisation ponds in which they are, in part, responsible for supplying oxygen to the surface waters. The metals removal potential of these ponds is very high and a number of mechanisms are likely to play an important role, as was the case in the wetland systems. Within the pond, a number of micro-environments exist in relation to the availability of oxygen and pH. These are likely to result in a high degree of precipitation of metals within bottom sediments which are subsequently not bio-available. As such, the overall metal removal capacity is likely to be greater than with activated sludge and trickling filters, but less than in a wetland system. There is, however, potential for bioaccumulation of metals through the food chain within a 120


stabilisation pond system which may have chronic toxicity effects on algal planktonic feeding fish species in the longer term.

Biosorption systems specifically for metal removal

Preliminary considerations The systems described above, in which metals are removed coincidentally, have limitations in relation to the metals loading which can be effectively handled. This is due, primarily, to the saturation capacity and the ability to replenish active sorption sites. To overcome these limitations and to develop efficient metal scavenging biosorptive systems, the potential of each biosorbent must be fully exploited so as to minimise reactor size and operating costs and produce a quality assured final effluent. Exploration of the surrounding chapters in this book will reveal a vast array of potential biosorbents with representatives from the Protista, plant and animal kingdoms. Some have a high specificity and are cultured with the sole aim of selective metal removal; others are waste products of biological origin and may display a broad spectrum of biosorptive ability. In many cases, exploitation of their potential on a large scale has not been attempted and research has ended on a laboratory scale once the biosorptive mechanism has been elucidated. It is often the case that such fundamental studies are performed under idealised conditions using solutions of Analar or laboratory grade reagents. Thus, potentially competing species are excluded, and the potential for recovery or disposal routes for metal-laden biomass is not considered. The following pages consider some of these problems, together with potential and real solutions. These are not exhaustive but illustrate the types of system available to the environmental engineer which may be adapted to exploit the huge potential for metallic species biosorption. One of the prime concerns in any design should be the ease with which metals can be recovered from the metal-laden biomass, or surplus biomass, removed from the system and contained or detoxified before disposal. Systems which offer nondestructive recovery and allow the biomass to be regenerated for further use are, at first sight, the most appealing. However, the cost of desorbent and regenerative chemicals has to be considered, as does the expected life span of the biosorbent and any loss of efficiency through repetitive regeneration cycles. In general, where adsorption is a surface phenomenon, as is the case with most non-viable biomass systems, the process of desorption of metals is usually a simple matter. Zinc and copper, for example, can usually be desorbed using dilute acids at pH values around 2. Gold, silver and mercury may bond more securely and it may not be possible to remove them using an acidic treatment without destroying the biomass. In such cases, other desorbents, such as mercapto-ethanol (Greene and Darnall, 1990), have been used. Also of importance is the volume of any desorbent as a fraction of the volume of the waste water treated. The smaller this fraction, the more concentrated the elutant will be, yielding more potential for recovery and re-use. Engineering design should be kept as simple as possible, bearing in mind that wastewater treatment processes are often poorly understood by their operators. For many situations, such as ground water remediation and leachate treatment, the 121


equipment may be located at a remote site and may be left unattended for long periods of time. In summary, Muraleedharan et al. (1991) have noted that an ideal biosorbent should fulfil the following criteria: • it should have a low cost and be re-usable • it should have a particle size, shape and mechanical strength suitable for use in bio-reactors under continuous flow conditions • the uptake of metal should be efficient and rapid • separation of the biosorbent from the solution should be rapid, efficient and cheap and should have a high selectivity • regeneration should be metal-selective and economically feasible.

Choice of a biosorbent material Some heavy metals are essential for the metabolism of living organisms. These include copper, zinc and manganese. Others have no biological function. All heavy metals are toxic to life, although different groups of organisms, species and even strains show varying degrees of tolerance. In any biosorption system, where living cells are used, this has to be considered as one of the limiting factors as the organisms can only be exposed to non-toxic concentrations of the metals. On the other hand, live biomass can exhibit a much wider range of metal accumulation mechanisms, as already noted, and some species have a high resistance to toxicity, even to the point of being able to live under saturation conditions. However, cells which have been killed, rendering them non-viable, are potentially more useful as biosorbents than their living counterparts; this view has been expressed by a number of authors (Brady et al., 1994) whose comments are summarised as follows. • Dead cells are not prone to metabolic toxins, but the adsorption sites can be ‘poisoned’, rendering them ineffective in removing the target metallic species. Many biosorbents are particularly prone to interference by light metal cations occupying sorption sites. This is exemplified by the work of Corder and Reeves (1994), who found that selected species of Cyanobacter, screened for their ability to remove nickel from solution, were ineffective in removing nickel from an electroplating wastewater due to the interference of sodium ions present. • They can be stored for prolonged periods, often by drying or by immersion in a strong preservative solution such as formaldehyde. Such treatment may also improve the biosorptive properties as well as cross-linking cells, which effectively immobilises them. • Killed organisms pose less of a threat to public health. • They do not require a constant supply of growth nutrients and, as new biomass is not produced as a result of growth processes, disposal problems are not aggravated. • Because the removal is by surface adsorption, metals can usually be eluted off the biomass by acid treatment; non-viable biomass is not usually affected by this type of treatment. 122


The use of a non-living biomass also offers a greater potential for re-use, with the following benefits (Ashley and Roach, 1990): • a concentrated eluate can be produced • only a small volume of concentrate is produced • there are no organic residues • biomass volume for ultimate disposal is minimised. Biomass is not the only biosorbent, and the potential for using other material of biological origin, e.g. extracellular polymers and other extractable polymers, has been explored in some detail. Such polymers are, of course, devoid of life and the advantages proposed for non-viable biomass generally apply also to these materials. The extraction of a polymer represents an additional processing step, or steps, in the production of the biosorbent and the economics of their use needs careful evaluation as a function of any enhancement the material may exhibit over the native biomass from which it was extracted.

Exposure of biosorbent surfaces to metal-laden wastewaters Reactor types often considered for use with biosorbents are shown schematically in Figure 6.3, and their properties and potential applications are discussed below.

Conventional stirred tank reactors A conventional stirred tank reactor (CSTR) is regarded as the traditional way in which microbes are maintained in intimate contact with their growth medium in a uniform and controlled manner. As such, this reactor type has found widespread application in both the fermentation and pharmaceutical bio-industries, as well as being adapted to suit the needs of the wastewater treatment industry in processes such as activated sludge. The reactor may be operated in a batch mode or in a continuous flow mode. The advantage of this type of reactor lies in the fact that the microorganisms are maintained in a homogeneous suspension, under well mixed conditions, thus optimising mass transfer of both nutrients and gases between cells and the surrounding medium. In addition, a supply of air, where necessary, can be introduced and carefully regulated so as to maintain a uniform dissolved oxygen concentration throughout the bulk of the medium. In continuous flow mode, the growth rate of the microbial population can also be manipulated by control of the reactor dilution rate, or, where cell recycle is practised, by the biomass wastage rate. Nutrients are replenished continuously by means of a constant influent substrate feed and substrate degradation products are diluted out of the reactor in the effluent stream. The reactor is, thus, ideally suited to applications where a microorganism has to be maintained in a living, or metabolically active, state and the mechanism of removal is by accumulation across the cell membrane. For the majority of applications involving non-viable biomass or polymeric biosorbents the CSTR is not an ideal reactor, especially when used in a continuous mode. This can be explained in terms of the 123


Figure 6.3 Reactor designs commonly used in biosorption studies

sorption kinetics, which indicate a first-order rate of reaction. Hence, for a rapid removal to take place, a high concentration of the metal ion must be maintained across the adsorption front. In a continuously operated CSTR, the effluent concentration of a material must be equal to that in the bulk liquid of the reactor (assuming complete mixing); low effluent concentrations necessarily mean low rates of reaction. The problem can be partly overcome by the use of a series of reactors or by adopting a plug flow regime in which the biomass flows in tandem with the wastewater over a prescribed period of time, which is equal to the retention time of 124


the reactor. This method is commonly used in the activated sludge system to ensure a high quality of final effluent with respect to biodegradable carbon. Stirred tank reactors also exert a shearing effect on the material contained within them, causing particles to be broken and destroyed. In living systems, this destruction of the biomass is compensated for by fresh growth. In a non-viable biomass, or in the case of a polymer, such repair is not possible and material is gradually lost by the attrition process. Biomass in its native form shows varying degrees of mechanical strength, but even the strongest will eventually be ground down within this type of reactor. As discussed later, various attempts have been made to strengthen biological materials through the process of immobilisation, and such immobilised materials may find application in stirred tank reactors. In the continuous mode of operation, therefore, the biosorbent will pass out of the reactor with the effluent stream as a function of the influent flow rate (assuming homogeneous mixing occurs). In a single pass reactor, the concentration of the biosorbent is, therefore, being continuously diminished unless the biomass is viable and has a growth rate potential greater than the reactor dilution rate. If the dilution rate exceeds the maximum growth rate, washout occurs and it is necessary to recycle the biomass in order to retain it within the system. In a similar way, non-viable biomass and polymers would need to be recycled. In both cases, the problem of phase separation arises. In most applications the biological phase, typically, is separated from the aqueous phase either by filtration or centrifugation. In some cases, where the biomass pelletises or flocculates, separation may be possible by sedimentation, although this rarely produces a sludge of better than 1–2% solids and a subsequent dewatering stage must be included. Where metal accumulation is intracellular, as is the case with some living biosorbents, it is usually necessary to destroy the biomass, and its future biosorptive potential, to recover the metal. CSTRs operated in batch mode are commonly used in laboratory studies (see earlier chapters) to determine the sorption parameters used in the design of other reactor systems.

Packed bed reactors Packed bed reactors have traditionally been used in both the fermentation and wastewater treatment industries for the maintenance of biofilms on an immobilisation matrix of an inert medium; typical examples are the biological (percolating) filter used for sewage treatment with its rock or slag medium and the vinegar tower in the rapid vinegar process which uses hardwood shavings as the support. More recently, packed beds have been used with a variety of manufactured supports including porous ceramics and glass, high surface area and void volume plastics, and reticulated packings (Table 6.1). In addition, microbes have been immobilised by cross-linking with a variety of polymers to produce semi-rigid self-supporting column packings. The basic principle of a packed bed reactor relies on the irrigation of the bed, either in an upflow or a downflow mode, so as to effect a contact between the immobilised biomass and the aqueous phase. Under ideal conditions, in a continuous mode of operation, the passage of the liquid follows a plug flow regime with a molecular contact time equal to the retention time of liquid within the bed. The biomass is retained within the reactor and, thus, the problems of washout are overcome, alleviating the requirements for a downstream separation and biomass recycle in low 125

Biosorbents for Metal lons Table 6.1 Surface area and voidage characteristics of some commonly used support media for biofilm growth in packed bed reactors

growth rate systems having a high hydraulic throughput. When used for growth mediated reactions, such as in the treatment of wastewater, then biomass growth has to be balanced by biomass removal from the bed for steady state operation. This is usually achieved by inducing a hydraulic shear which results in a scouring of the biofilm. The problems of biomass growth are less easily dealt with in polymer matrices, where the integral structure can be destroyed by an expansion of the matrix due to biomass growth. The reactor is, therefore, not ideally suited to applications where a microorganism has to be maintained in a living or metabolically active state, with the mechanism of removal by accumulation across the cell membrane, unless a method of removing the metal-laden biomass can be accomplished. The sorption kinetics of metals onto non-viable biomass is ideally suited to a packed bed reactor, as the mode of irrigation and the static position of the biomass effectively creates an adsorption front which will move through the reactor as a plug, ensuring a high concentration gradient between the metal ion and a proportion of the biomass at all times until column saturation occurs; at this point there will be a breakthrough of metal ion in the effluent stream. A packed bed reactor utilising an immobilised adsorbing biomass behaves in much the same way as an adsorption column of a material such as activated carbon or an ion-exchange column. To have it act in this manner, attention must be paid to the design of the column and its contents so as to avoid short circuiting and the disruption of the adsorption front by forward and back mixing. In this respect, the aspect ratio of the column, the inlet distribution system, and the nature of the packing material all need careful consideration, the general aim being to maximise the degree of plug flow within the reactor. For this reason, recirculation of effluent, which is common in waste treatment applications, would tend to be detrimental to the process as it effectively increases the degree of mixing. The major problem with packed bed reactors is their tendency to become blocked; this can be caused by the presence of insoluble particulate material in the influent stream, by the development of biofilm by a contaminating (living) microorganism, or 126


by the formation of precipitates within the medium. All of these will lead to an increasing back-pressure when operated in an upflow mode or a restriction in drainage of the bed when operated in a downflow mode. Such problems can usually be overcome from an operational point of view, but should be considered in the design of the column and by the inclusion of pretreatment stages where necessary. A second type of problem which may be faced is the irreversible binding of the metal, or other contaminant, with the immobilised biomass rendering it useless for further application; at this point the column would have to be repacked with fresh sorbent material. The spent biomass would then need to be separated from its immobilisation matrix, which may or may not be practicable or economical, and the biomass replenished. The life expectancy of the column packing, its amenability to regeneration, and disposal of an increased volume of the biomass/support medium matrix are primary factors in the selection of this type of system. Unless regeneration of the attached/entrained biomass is possible, then the use of a packed column is usually feasible only using robust self-supporting biomass generated as a by-product from another process, e.g. crustacean shells. Microbial biomass cultured onto an immobilisation matrix and then rendered nonviable, or the polymerisation of biomass or cellular fractions which are capable of regeneration, is ideally suited to biosorption of metals in packed columns for the following reasons: • maximum sorption efficiency will be maintained until breakthrough occurs • on breakthrough, the sorptive capacity of the column will have been reached, thus maximising efficiency and minimising reactor size • the biomass is retained within the reactor and, thus, no downstream separation of it from the effluent is required • immobilisation, coupled with low hydraulic flows, offers some protection from attrition of the biomass as a result of hydraulic shear. Because the regeneration capability of the biomass is likely to have a critical impact on the economics of packed bed reactor operation, an evaluation of the loss of sorptive capacity, in relation to the number of regenerative cycles, should be undertaken. Equally, the volume of regenerated liquor as a function of the bed volumes passed is useful in calculating the factor by which the pollutant metal can be concentrated in the recovery process. To date, packed beds have shown the greatest promise for practical application and commercialisation of the biosorption phenomena.

Expanded bed, fluidised bed and airlift reactors The unifying feature of these three reactor types is the fact that a motive force is applied which causes the biomass to resist a natural tendency to settle into a packed bed of material. In an expanded bed the motive force is usually supplied by the upflow velocity of the liquid passing through the material, being roughly equal to the sedimentation velocity of the particles. In this way, the particles have a limited movement relative to one another and the void volume within the bed is increased over that of a static bed of the same material. In a fluidised bed the motive force is 127


greater and unrestricted movement of particles relative to one another is observed, leading to an almost random motion of particles within a defined volume. In both expanded and fluidised bed reactors the biomass is in constant motion and, thus, channelling and clogging should be prevented. This allows these designs to be used with an influent feed stream containing suspended material, although, in most cases, this is avoided. In addition, because of the constant motion of the biomass, efficient mass transfer is achieved through an accelerated flow of liquid across the biomass surface, minimising the potential for localised concentration gradients. Both expanded bed and fluidised bed reactors have found application in the water and wastewater industries, examples being the UASB anaerobic digestion process (expanded bed) and anoxic fluidised bed denitrification reactors. Of critical importance in reactors of this type is the density of material requiring expansion or fluidisation, as this will not only have a bearing on the process energy requirements but will also determine the manner in which the motive force is applied (the aspect ratio of the reactor is also critical in this respect). For the purposes of passive adsorption onto biomass, as was the case with packed bed reactors, a concentration gradient between the adsorbent and the liquor is required which is best achieved by a plug flow hydraulic regime. Consequently, if the motive force required for expansion or fluidisation of the bed requires recirculation of the liquor in order to achieve it, then this can only be detrimental to the process. In terms of expansion or fluidisation the motive force can usually be applied with less energy expenditure by utilising a gas, and for aerobic microbial reactions air can be used simultaneously to provide fluidisation and aeration. The airlift fermenter differs slightly in that the air is applied in such a way as to induce a liquid circulation within a centrally baffled reactor. The design, therefore, offers a degree of plug flow in which there is a finite initial contact time between the influent flow and the microorganism equal to the time passage of the liquid up the riser section of the reactor. The internal recirculation of the liquor, caused by the airlift effect, reduces this initial contact time proportionally in the ratio of inlet flow to induced flow. Both fluidised beds and airlift designs may be considered useful in applications where the biomass is in a metabolically active state and removal of the metal pollutant is across the cell membrane; they are, however, less useful when considering passive adsorption where maximisation of the concentration gradient between free liquid and the cell biomass surface is important in determining the rate. The expanded bed presents greater opportunities, provided expansion of the bed can be achieved without recirculation, while at the same time maintaining the required retention time; the reactor kinetics are likely to be similar to those of a packed bed.

Immobilisation matrices For commercial application (see Chapter 2), it is generally necessary to consider immobilisation of the biomass so as to increase its mechanical strength and produce a material of more uniform particle size. Immobilised cells are usually easier to handle, require less complex separation systems, allow a high biomass density to be maintained, and provide a greater opportunity for re-use and recovery. Immobilisation of cells has also permitted the successful use of the reactor designs described above, and can be used to overcome limitations such as porosity in static beds, biomass retention in expanded and fluidised beds, and attrition of the 128


biosorbent; most importantly, it allows simple separation of biomass from the aqueous phase. Both living and dead biomass can be immobilised, although some of the immobilisation processes themselves may render the biomass non-viable, making them unsuitable for metabolic dependent metals removal processes.

Polymeric beads and granules Immobilisation of biomass by its entrapment into polymeric beads, or using polymers to form robust granules, has found the widest application in commercial biosorption processes. The technique is particularly useful as the biosorbent biomass can be prepared independently prior to the immobilisation process and thus optimised to realise its maximum sorptive potential. The technique also offers the opportunity to incorporate waste biomass from other processes; most commonly used are those from the fermentation industries where fungal species are used extensively for the production of a wide range of products including alcoholic beverages (Saccharomyces cerevisiae); certain enzymes (Mucor miehei, Aspergillus niger), penicillin antibiotics (Penicillium chrysogenum), and the fermented oriental food ‘tempeh’ (Rhizopus arrhizus). Immobilisation, to produce robust granules, overcomes the problems associated with small microbial cells, parts of cells and tangled filaments. Granules of uniform size and consistency are likely to offer more uniform diffusion characteristics and better overall performance. A novel method of overcoming the problems of fully colonising a porous biomass support particle has been devised by the U.S. Bureau of Mines’ Salt Lake City Research Centre (Bennett et al., 1991). The method involved in preparing the material, known as Bio-fix beads, allows the polymer to form around the biomass to be immobilised; this is achieved by dissolving polysulphone pellets in dimethylformide and then blending the dried biomass into the solution. The resulting slurry is then sprayed into water through an atomiser and beads of 0.5–2.5 mm are formed: this can be varied depending on the conditions used. The Bio-fix beads have been used in studies on over 50 wastewaters, including mine waters and contaminated ground waters, with mixed metal ion concentrations ranging from 0.5 mg/l to over 100 mg/l. The beads, reported to be able to be used in all conventional reactor types, are stable to acid washing, and can be regenerated by this process. The stability of the beads has been tested over 170 regeneration cycles using cheap sources of biomass (peat, algae, etc.). In later work, Jeffers et al. (1991, 1993) showed Bio-fix beads to work over a wide range of pH (3–8) and to have a metal affinity series Al>Cd>Cu>Zn>Fe> Mn>Ca>Mg. In effect this means that the beads will initially take up the water hardness minerals but these will be displaced progressively by heavy metal ions. Bio-fix beads have been extensively field tested, both conventionally and with two low-cost/maintenance process systems. In all cases, the engineering design is worth detailing as a model for other potential immobilised biomasses. For high flow rate applications, a three-column system was utilised consisting of a lead column, a scavenger column and an elutant column (Figure 6.4). During operation, the lead and scavenger columns are loaded with the wastewater in series, while the third column is eluted of metals and regenerated. At the end of a loading cycle the scavenger column becomes the lead column, the regenerated column the scavenger column, and the lead 129


Figure 6.4 Three-column biosorption system utilising a lead (i) and scavenger column (ii) in series operation while the third column (iii) is being regenerated

column is by-passed and regenerated. The arrangement is ideal for dealing with waste streams containing multiple metal ions, in that ions which are lower on the affinity scale, and which become displaced from the lead column as sites become saturated, will be captured on the scavenger column. Since the scavenger column cannot be saturated prior to the lead column, the chances of breakthrough are minimal. The third column allows the regeneration stage to be conducted in isolation without impairment of the process. Columns can be operated either in a downflow mode or, where the levels of suspended solids are high and likely to cause blinding of the bed, in an upflow mode resulting in some expansion of the bed. In two pilot scale studies this process design was successful in treating acid mine drainage (AMD) water from a techonite operation and a zinc mining operation. In both cases the elutant concentrated liquor was precipitated and the supernatant returned to the process stream. The low-cost/maintenance processes both comprised the Bio-fix beads contained in porous polypropylene bags. The first of these systems employed a weired trough in which the water was forced consecutively through a series of bags in alternating upflow and downflow directions as it passed down the series by the force of its own hydrostatic head. At regular intervals the bag at the head of the series was removed, the subsequent bags moved upstream one position, and a bag of regenerated beads inserted at the end of the series. Such a system successfully treated, over an eleven month period, the adit waters from an abandoned silver mine. During the period of operation, bags containing 2.3 m 3 of beads were removed and regenerated, corresponding to a total flow through the system of 450 m 3. The second system, described as a bucket system, comprised a series of sealed canisters into which bags of the beads were placed. Each canister was served by an influent pipe in its base and 130


a top overflow. The latter subsequently fed the next canister in the series which was placed at a lower level, allowing hydrostatic head to move the water through the system. A similar maintenance regime was adopted to that used in the trough system, i.e. removing the bag from the first canister in the series, moving the bags, and inserting a regenerated bag in the final canister. This second system was used to treat the ADM from an abandoned zinc mine alongside a second trough system. Both performed well when beads were removed, regenerated and replaced in line with design expectations, but if the cycle was delayed or extended, then the system performance was adversely affected. Both of these low-cost systems illustrate how sound engineering principles can be applied to provide a reliable, robust, low maintenance solution to treatment in remote locations where reprocessing facilities for spent biomass are not available. Indeed, a number of such plants could be served on a regional basis by a single reprocessing facility. The AMT-Bioclaim process (Brierley et al., 1986a) is a second commercial process for metal scavenging which uses a metal recovery agent (MRA) in a granular form. Column tests, used to determine the maximum practical loading of the MRA with reference to metal salt solutions, showed removal of Ag, Cd, Cu, Pb, and Zn with the following efficiencies: 86, 214, 152, 601, and 137 mg/l. The MRA granules are reported to be useable in fixed bed reactors and fluidised beds. In the latter, the heavier particles sink and the lighter particles are buoyed up and this feature is used, in a process known as slugging, to remove granules from the base of the unit. Different biomasses can be incorporated into the MRA, but the one found to be most efficient and used in the AMT-Bioclaim process is not revealed. The process has been tested for gold recovery from a cyanide complexed waste stream from jewellery manufacture. Platinum and palladium and a gold mine process leach solution were tested later (Brierly and Vance, 1988). It is claimed that the MRA can be recovered and re-used, but the method is not given. Economics of the Bio-claim system are considered in a separate publication (Brierley et al., 1986b). A second immobilised cell biosorbent has been prepared by the same group (Brierley and Brierley, 1993). This has the form of stable granules in which a Bacillus species was immobilised using polyethyleneimine (PEI) and the polymerisation initiator gluteraldehyde. The process involved mixing the biomass with the PEI to a smooth paste, adding the gluteraldehyde and mixing until a dough-like paste was formed. This was then extruded, dried and broken up before being sieved to a range of mesh sizes. It was found that the performance of the granules as metal sorbents could be improved by a caustic treatment after the immobilisation. Advanced Mineral Technologies has developed a non-viable biomass immobilisation technique which uses a commercial strain of yeast grown on molasses and involves a caustic treatment at high temperatures. The treated material is then dried and ground to a granular consistency (Brady et al., 1994). The finished product is claimed to be 99% efficient in metal cation removal. The microscopic granules can be retained by a simple mesh and it has proved successful in treating electroplating waste, where it was shown that copper bound in preference to zinc and zinc in preference to chromium. No attempt at re-use of the biosorbent was reported, as this was not thought to be necessary because of the low-cost of waste yeast. AlgaSORB (Darnall, 1991) is a product in which the biosorbent is algal cells which have been killed in the manufacturing process (see Chapter 2). The polymeric matrix in which the cells are immobilised is claimed to be robust enough to be packed into columns. The polymer/biomass matrix is reported to be durable for long periods 131


of immersion (2 years), not to lose efficiency in its metal binding capacity, to be capable of regeneration, and not to be interfered with significantly by Ca2+, Mg2+, Na+ or K + ions. It can, therefore, be used with hard or saline waters. The product is claimed to have use in metals recovery from electroplating wastewaters and precious metal recovery from mining processes. In remediation it can be used for lead removal from drinking water, leachate treatment from contaminated ground water and contaminated land sites. Other industrial applications include wastewaters from printed circuit board manufacture. Generally, the volume of the metal elutant stream constitutes about 4–5% of the treated wastewater stream. Polyacrylamide gel beads also have application as an immobilisation matrix (Nakajima and Sakaguchi, 1986; Darnall et al., 1986) and have been shown to function over a wide range of pH values (4–9). It has been suggested that this may be due to the buffering capacity of the micro-environment generated within the beads by the microbes. The procedure adopted by Nakajiima and Sakaguchi (1986) immobilised the precultured cells by suspending them in 4.5 ml of isotonic sodium chloride solution to which was added 680 mg of acrylamide monomer, 34 mg of N,Nmethylene-bis (acrylamide), 0.3 ml of propionitrile solution (3%), and 0.34 ml of potassium persulphate solution (2.5%) to produce a solidified gel which could be crushed into small pieces. Using Streptomyces albus in free cell suspension for the sorption of uranium, almost 50% of the dry weight of biomass was lost after 5 sorption/desorption cycles. Using the same culture in the immobilised state, less than 2% losses were observed. The losses using free cells are similar to those observed by Gadd and White (1992) in thorium uptake experiments using freely suspended bacteria and serving to emphasis the attrition which can take place in continuously stirred cultures of non-viable microbes. Similar gels have also been used in the immobilisation of Chlorella regularis and Streptomyces viridochromogenes for batch adsorption/desorption studies using uranium (Sakaguchi and Nakajima, 1991). It is also claimed that the gel beads are suitable for column use. Harris and Ramelow (1990) have, however, criticised the use of polyacrylamide, used in some early experiments (Darnall et al., 1986), as an immobilisation matrix due to its low strength, high water content (80–95%) and leakage which occurs through the open network. To overcome these problems they developed an ethyl acrylate-ethylene glycol dimethylacrylate co-polymer, with a cross-linked network, for the immobilisation of Chlorella vulgaris and Scenedesmus quadricauda. The entrapped biomass was studied in column experiments using solutions of Cu, Cd, and Zn with the result that copper was bound by both the immobilised biomasses to a greater extent than had been observed with dried and sieved samples of the biomass in contact experiments. Desorption of the column containing S. quadricauda was tested using sodium acetate buffers at pH values down to 2.0, with maximum desorption, approximately 70%, occurring at around pH 3.0. Bio-polymers, such as alginate derived from seaweeds, have also been used for biomass immobilisation (Gee and Dudeney, 1988) but are unlikely to produce a material of sufficient strength for commercial application. Alginate itself is a biosorbent and has been used in a number of trial applications (Hartmeier et al., 1992; Apel et al., 1991). These showed that alginate beads had the capability of binding as much as 453 mg of lead per gram of dry matter if, after formation, they were hot air dried (80ºC) for 24 h to a dense (450 g/1) granular form. In this condition, they showed only a small potential for rehydration on immersion in water and could be regenerated for further use by using a 0.1 M solution of CaCl2 acidified to pH 0.3 with HCl. 132


The potential of the immobilisation matrix to act as an adsorbent has also been expressed by Tobin et al. (1993), who have developed protocols for the immobilisation of Rhizopus arrhizus using alginate, epoxy resin and polyvinylformal materials. The cell loadings possible with each of these matrices were investigated and the PVF was found to give the highest cell-immobiliser ratio for use in a practical form. At an 80% cell loading, it was found to form a pliable matrix which was stable in the testing procedures developed. Cell free polyacrylamide beads were found to have no biosorptive capacity, while those of alginate tested positive, as did the epoxy resin and PVF preparations. The epoxy resins also resulted in an almost total loss of biomass uptake efficiency. Obviously, the adsorption characteristics of the immobilisation material are important to the overall adsorptive capacity of the matrix and should be taken into account in any design calculations based on free cell suspension adsorptive capacities. Similarly, the size and diffusional characteristics of the matrix will affect performance, and the only real way of deriving the necessary design information is by trials with the product itself. An immobilisation technique has also been used by Tsezos and Deutschmann (1990) using an undisclosed immobilisation protocol; the use of this biosorbent is described elsewhere (Tsezos, 1988, 1990). The development of microfungal filters in which fungal mycelia are laid down with various conventional paper making and textile fabrics is also of interest (Wales and Sagar, 1990). These filters can have a high wet strength, exhibit good tensile properties and achieve flow rates through individual filters as high as 0.5 ml/cm2 with up to 90% reduction in metal ion concentration. The filters are, of course, limited by their adsorption capacity, which is proportional to the weight of mycelia incorporated and, hence, the filter thickness. However, they appear to offer more potential as a laboratory aid than as a commercial treatment technology. In general, the process of immobilisation of biomass offers advantages to its large scale or commercial application as a biosorbent. Even with biomass such as fungi, which have a tendency to grow in a pelleted form under liquid culture conditions, the strength of these pellets is rarely adequate to tolerate the forces exerted upon them. Pelleted biomass is a useful tool for carrying out laboratory sorption studies in both batch and column experiments of short duration. Most studies over longer periods of time, and especially those using regenerative steps, have shown high rates of attrition and compression of the biomass when used in columns and other reactor designs.

Biofilm systems The development of a biofilm relies on the natural colonisation of the surface of an inert material which may be either planar or porous. A wide variety of materials have been used as support media in fixed film biological systems, ranging from the traditional rock and blast furnace slag of percolating filters to specifically designed plastic media. Of importance in the selection of a medium is its surface area and percentage void volume, which can range from 100 m2/m3 for 250 mm rock (with 50% void volume) to 1000 m 2 /m 3 (with 90% void volume) for a reticulated polyurethane biomass support particle (BSP). Some of the more common biofilm support media and their properties are shown in Table 6.1. Cotoras et al. (1993) used a species of Bacillus grown as a biofilm on glass Rashig rings in a small laboratory column reactor. The biomass was grown in situ (12 h) after 133


sterilisation of the reactor under conditions of aeration, after which time the reactor was drained and the biomass rinsed with distilled water. The column was used for the removal of uranium from solution (0.1 mM) and later desorbed using a hydrochloric acid solution at pH 1. The authors state that the reactor was used repetitively, but do not say whether the cells remained viable after the acid washing or whether fresh nutrient was added between (or with) the metal ion solution. The desorbent volume was 400 ml for a feed volume of 1500 ml, i.e. 27%: a figure unlikely to find commercial application as the elutant would be at only 4 times the concentration of the wastewater. Huang et al. (1989, 1990) used acid-washed, live Saccharomyces cerevisiae which was immobilised onto white sand (particle size 0.42–0.5 mm) in a packed column by pumping the yeast through the bed. The amount of biomass trapped in this way was estimated by the difference in optical density (420 nm) between the influent and effluent streams. The packed bed reactor was tested on its ability to remove copper, lead and cadmium from solution; the results showed the reactor to behave as an idealised adsorption column with no detectable metal in the effluent before breakthrough occurred. Metal removal was attributed to surface adsorption, but some intracellular uptake was also shown to be occurring with copper solutions. The bed could be regenerated using a 5×10-3 M solution of HClO4 and then used repetitively. The volume of eluant required for each desorbent cycle was 5% of the volume of metal-containing solution treated. It is not clear whether or not the yeast was maintained in a viable state during extended regenerative runs by intermittent feeding with a nutrient solution, or if the yeast attached to the sand particles or was simply entrapped within the sand matrix of the bed. Cultures of Rhizopus arrhizus can be made to colonise the interstitial spaces of both reticulated foam sheet and stainless steel knitted meshes to give biomass concentrations as high as 4.33 g dry weight per litre of immobilisation matrix (Ileri et al., 1990). Once grown, the biomass can be sterilised by treatment with a formaldehyde solution and stored for many months. The immobilised cell matrix on the inert support can then be used in a number of different reactor configurations. In this particular study, copper removal was demonstrated to a saturation level of 50–65 mg Cu2+/g dry weight of immobilised cells. This did not compare favourably with the saturation capacity of 175 mg Cu2+/g dry weight of freely suspended biomass in shake flask experiments, and the difference was attributed to diffusional problems in the immobilisation matrix. Reticulated polyurethane foam has been used for the immobilisation of other strains of fungi such as Aspergillus oryzae, Penicillium sp. (Kiff and Little, 1986) and Rhizopus arrhizus (Lewis and Kiff, 1988) for the removal of Cd, Cu, Fe, Mn, Pd and Zn in upflow packed columns, although it is claimed that the method of immobilisation is suitable for a wide range of contactor configurations. The system was reported to be unaffected by the presence of low concentrations of organic matter and alkali metals, but was sensitive to low pH, temperature and the presence of competing cations. The biomass, once entrapped in its immobilisation matrix, is very stable and can be used repetitively without apparent loss in its effect. A pilot scale biosorption column containing Rhizopus arrhizus immobilised in this manner was used over an experimental period of 6 months, with repetitive backwashing, to remove humic materials from a raw water supply (Banks and Parkinson, 1992). The volume of elutant was 2% of the volume of water treatment, and a similar performance would be anticipated if using the system for metals removal. The major 134


disadvantage of this system is the necessity to cultivate the fungus in association with the support particle: an adequate technique for achieving this at a large scale has still to be developed. Equally, the technique excludes the use of spent waste biomass, and the cost of the BSPs is high. The only large scale biofilm process (to which reference can be found) that has been specifically designed with one of the purposes being the removal of metals is the treatment plant located at the Homestake Mine (Whitlock and Mudder, 1985). This treatment facility deals with a flow of 21000 m3 /day of wastewater containing cyanide, thiocyanate, ammonia, iron, copper, nickel, lead and zinc. The treatment process utilises a mutant bacterium capable of oxidising the organic components of the waste while, at the same time, accumulating metals through adsorption, ingestion and precipitation. The biomass is immobilised as a film on the discs of a series of rotating biological contactors (RBCs), each of which is automated to provide the required degree of oxygenation and biomass accumulation. The plant is reported to remove 95% of toxic heavy metals, but no data are given on the mass balance of metals through the system or the quantities of metal-laden sludge requiring disposal. It is interesting to note that a patent on a specific design of RBC was taken out in 1973 (Ames Crosta Mills and Sanderson, 1973) in relation to its ability to remove contaminants from wastewater through adsorption. The use of chitosan as a biosorbent, and some of the methods of cross-linking the polymer to increase its stability, are covered in the next section. However, one potential application of this effective polymer is as a film covering an inert support material; this differs from our previous definition of a biofilm in that the material is not grown in situ but is applied at a later stage. This technique, developed by Onsoyen and Skaugrud (1990), coats the chitosan onto expanded silica, with a large surface area, to produce a filtration medium with a very high binding capacity for metals.

Bio-polymers One of the more commonly studied polymeric biosorbents for metals removal is chitosan and its derivatives. The potential of this material for the removal of impurities in aqueous streams has been known for many years and was the subject of a patent in 1970 (Peniston and Johnson, 1970). Since that time its use as a practical adsorbent has been attempted by a number of workers both in a pure form and only partially derivatised. Coughlin et al. (1990) chopped and sieved (1–10 mm) crab shells which were partially deacetylated by a rapid hot alkali treatment of the material after it had first been decalcified using HCl. The process was found to be much less costly than making pure chitosan from the crab shells for two reasons: • chopping was found to be more economical than grinding of shells to a small particle size, and made the materials more manageable and useful as a practical sorbent • the cost of chemicals for deacetylation was lower. The material, known as PCSFW (partially converted shellfish waste) has only been used in laboratory experiments, although its application to an Ni and Cu containing electroplating wastewater has been demonstrated. Metals could be desorbed from the 135


PCSFW by acid treatment and the material regenerated using NaOH. However, with repeated treatment the sorptive capacity decreased. Elson et al. (1980) undertook packed column experiments using a 60 mesh sieved chitin/chitosan mixture, which gave better column characteristics, to remove arsenic from solution and contaminated well water. Unfortunately, the efficiency of the material was only 10-4 times that of alternative ion exchange resins. One of the problems encountered with chitosan is that it forms viscous, watersoluble salts in acidic solution. In the case of metal binding, this precludes the use of acids to elute the metals in any regeneration step. To overcome this problem, Masri et al. (1978) partially crosslinked the native chitosan, in its solid state, with the polyfunctional agent gluteraldehyde. In this state, the chitosan retained its original native structure and metal binding capability and was suitable for use as column packing. Preparation of gluteraldehyde crosslinked chitosan beads has been further investigated by Rorrer et al. (1993) and tested as an adsorbent for cadmium in batch adsorption experiments. Other crosslinking materials are available for the stabilisation of chitosan as a porous bead structure. Seo and Kinemura (1989) varied the preparation conditions of coagulation of a chitosan dope in alkaline solution to give beads of varying porosity and size; the beads were then crosslinked with diepoxy or diisocyanate derivatives to form a stable structure named Chitopearl. These beads were used in small scale trials for the treatment of mine water in batch experiments. Comparison of reactor designs As indicated in the section on bio-reactor types, different designs are likely to offer differing degrees of performance even when the same biomass is exposed to the same wastewater. Few comparative studies of this type have been reported, with most workers choosing to work with one particular type of system. White and Gadd (1990) studied four different bio-reactor designs in relation to the removal of thorium by Rhizopus arrhizus grown as a pelleted mycellium: upflow and downflow packed beds; a slowly stirred reactor; and an airlift reactor. Neither type of the static beds nor the stirred reactor gave satisfactory thorium removal, and they all demonstrated a poorly defined breakthrough into the effluent. The airlift system showed 90–95% removal over a prolonged period and a rapid breakthrough on saturation of the biomass; the improvement was attributed to better contact between the metal-containing solution and the biomass. The static bed designs were criticised as allowing channelling to occur, thus short-circuiting the biosorbent. It is, of course, possible to model the process of biosorption within any particular system and to derive the most important parameters which are likely to impact on the process performance. Such a model for a batch reactor has been developed by Tsezos and Deutschmann (1992), using immobilised biomass, with the finding that biomass particle size, initial solute concentration, maximum biomass uptake capacity and mass transfer coefficients are the most significant parameters affecting the process. References AMES CROSTA MILLS & COMPANY LTD and SANDERSON, F., 1973, Apparatus for the biological treatment of waste water by the biosorption process, Application: GB 69–49332 691008. GB 1 324 358 730725. 136

Scavenging Trace Concentrations of Metals APEL, M.L., BARNES, J.M. and TORMA, A.E., 1991, Biosorption kinetics of metal removal from uranium mill tailing effluents. In: Mineral Bioprocessing, SMITH, R.W., and MISRA, M. (Eds), The Minerals, Metals and Materials Society, Warrendale, Pennsylvania, pp. 339–353. ASHLEY, N.V. and ROACH, D.J.W., 1990, Review of biotechnology applications to nuclear waste treatment, Journal of Chemical Technology and Biotechnology, 49, 381–394. BANKS, C.J. and PARKINSON, M.E., 1992, The mechanism and application of fungal biosorption to colour removal from raw waters, Journal of Chemical Technology and Biotechnology, 54, 192–196. BENNETT, P.G., FERGUSON, C.R. and JEFFERS, T.H., 1991, Biosorption of metal contaminants from acidic mine waters, In: Process Mineralogy XI—Characterization of Metallurgical and Recyclable Products, HAUSEN, D.M., PETRUK, W., HAGNI, R.D. and VASSILIOU, A. (Eds), The Minerals, Metals and Materials Society, Warrendale, Pennsylvania, 213–222. BRADY, D., STOLL, A. and DUNCAN, J.R., 1994, Biosorption of heavy metal cations by non-viable yeast biomass, Environmental Technology, 15, 429–438. BRIERLEY, C.L. and BRIERLEY, J.A., 1993, Immobilization of biomass for industrial application of biosorption, In: Biohydrometallurgical Technologies. Vol. 2, TORMA, A. E., APEL, M.L. and BRIERLEY, C.L. (Eds), The Minerals, Metals and Materials Society, Warrendale, Pennsylvannia, pp. 35–44. BRIERLEY, J.A., BRIERLEY, C.L. and GOYAK, G.M., 1986a, AMT-Bioclaim: a new wastewater treatment and metal recovery technology, Process Metallurgy, 4, Fundamentals of Applied Biohydrometallurgy, 291–304. BRIERLEY, J.A., GOYAK, G.M. and BRIERLEY, C.L., 1986b, Considerations for commercial use of natural products for metals recovery, In: Immobilisation of Ions by Biosorption, ECCLES, H. and HUNT, S. (Eds), Ellis Horwood, Chichester, pp. 105– 117. BRIERLEY, J.A. and VANCE, D.B., 1988, Recovery of precious metals by microbial biomass , In: Biohydrometallurgy, Proceedings of International Symposium, Warwick, 1987, NORRIS, P.R. and KELLY, D.P. (Eds), Science and Technology Letters, Kew, pp. 477– 485. BROWN, H.G., HENSLEY, C.P., MCKINNEY, G.L. and ROBINSON, J.L., 1973, Effi-ciency of heavy metal removal in municipal sewage treatment works, Enviromental Technology Letters, 5, 103–114. BROWN, M.J. and LESTER, J.N., 1979, Metal removal in activated sludge: the role of extracellular polymers, Water Research, 13, 817–837. BROWN, M.J. and LESTER, J.N., 1982, Role of bacterial extracellular polymers in metal uptake in pure bacterial culture and activated sludge 1—effects of metal concentration, Water Research, 16, 1539–1560. CORDER, S.L. and REEVES, M., 1994, Biosorption of nickel in complex aqueous waste streams by cyanobacteria, Applied Biochemistry and Biotechnology, 45/46, 847–859. COTORAS, D., VIEDMA, P. and PIMENTEL, J., 1993, Biosorption of metal ions by attached bacterial cells in a packed-bed bioreactor, In: Biohydrometallurical Technologies, Vol 2. TORMA, A.E., APEL, M.L. and BRIERLEY, C.L. (Eds), The Minerals, Metals and Materials Society, Warrendale, Pennsylvania, pp. 103–110. COUGHLIN, R.W., DESHAIES, M.R. and DAVIS, E.M., 1990, Chitosan in crab shell wastes purifies electroplating waste water, Environmental Progress, 9, 35–39. DARNALL, D.W., 1991, Removal and recovery of heavy metal ions from waste waters using a new biosorbent; AlgaSORB, Innovative Hazardous Waste Treatment, Technology Series. Vol. 3, Biological Processes, 65–72. DARNALL, D.W., GREENE B., HENZL, M.T., HOSEA, J.M., M C PHERSON, R.A., SNEDDON, J. and ALEXANDER, M.D., 1986, Environmental Science and Technology, 20, 206–208. 137

Biosorbents for Metal lons DUNBABIN, J.S. and BOWMER, K.H., 1992, Potential use of constructed wetlands for the treatment of industrial wastewaters containing metal, Science of the Total Environment, 111, 151–168. DUGAN, P.R., 1970, Removal of mine water ions by microbial polymers, Proceedings of the 3rd Symposium on Coal Mine Drainage Research, Pittsburgh, Pennsylvania. DUGAN, P.R. and PICKRUM, H.M., 1972, Removal of mineral ions from water by microbially produced polymers, Proceedings of the 27th Industrial Waste Conference, Purdue University, Purdue, USA, pp. 1019–1038. EGER, P., 1994, Wetland treatment for trace metal removal from mine drainage: the importance of aerobic and anaerobic processes, Water Science and Technology, 29, 249– 256. ELSON, C.M., DAVIES, D.H. and HAYES, E.R., 1980, Removal of arsenic from contaminated drinking water by a chitosan/chitin mixture, Water Research, 14, 1307–1311. FRIEDMAN, B.A. and DUGAN, P.R., 1968, Concentration and accumulation of metallic ions by the bacterium Zooloea, Developments in Industrial Microbiology, 9, 381– 388. GADD, G.M. and WHITE, C., 1992, Removal of thorium from simulated acid streams by fungal biomass: potential for thorium desorption and re-use of biomass and desorbent, Journal of Chemical Technology and Biotechnology, 55, 39–44. GEE, A.R. and DUDENEY, A.W. L., 1988, Adsorption and crystallisation of gold at biological surfaces, In: Biohydrometallurgy, NORRIS, P.R. and KELLY, D.P. (Eds), Science and Technology Letters, Kew, pp. 437–451. GREENE, B. and DARNALL, D.W., 1990, In: Microbial Mineral Recovery, EHRLICH, H. L. and BRIERLEY, C.L. (Eds), McGraw-Hill, New York, pp. 303–323. HAMMER, D.A., 1989, Constructed wetlands for wastewater treatment, Proceedings of the First International Conference on Constructed Wetlands for Wastewater Treatment, Chattannoga, Tennessee, Lewis Publishers, Chelsea, Michigan. HARRIS, P.O. and RAMELOW, G.J., 1990, Binding of metal ions by particulate biomass derived from Chlorella vulgaris and Scenedesmus quadricauda, Environmental Science and Technology, 24, 220–228. HARTMEIER, W., SCHUMACHER, R., GLOY, W. and LASSAK, R., 1992, Biosorption of heavy metals using immobilised polymers of plant origin, Medical Faculty Landbouww, University of Gent, 57, pp. 1713–1716. HUANG, C.P., MOREHART, A.L., HUANG, C.J., WESTMAN, D.C. and QUIRK, K., 1989, Removal of toxic heavy metals from contaminated ground water by a fungal adsorption process. Technical Report USGS/G–1292, NTIS, Springfield, Virginia. HUANG, C.P., HUANG, C.P. and MOREHART, A.L., 1990, The removal of Cu(II) from dilute aqueous solutions by Saccharomyces cerevisiae, Water Research, 24, 433–439. ILERI, R., MAVITUNA, F., PARKINSON, M. and TURKER, M., 1990, Repeated use of immobilised dead Rhizopus arrhizus for the removal of heavy metal contaminants from wastewater, Proceedings APBioChEC’90, 564–567. JEFFERS, T.H., BENNETT, P.G. and CORWIN, R.R., 1993, Biosorption of metal contaminants using immobilised biomass—Field Studies, Report of Investigations 9461, United States Bureau of Mines, 1–10. JEFFERS, T.H., FERGUSON, C.R. and BENNETT, P.G., 1991, Biosorption of metal contaminants from acidic mine water, In: Mineral Bioprocessing, SMITH, R.W. and MISRA, M. (Eds), The Minerals, Metals and Materials Society, Warrendale, Pennsylvania, pp. 289–298. KIFF, R.J. and LITTLE, D.R., 1986, Biosorption of heavy metals by immobilised fungal biomass, In: Immobilisation of Ions by Bio-sorption, ECCLES, H. and HUNT, S. (Eds), Ellis Horwood, Chichester, pp. 71–80. LESTER, J.N., 1981, Removal of heavy metals in conventional wastewater treatment, International Conference on Heavy Metals in the Environment, Amsterdam, Holland. 138

Scavenging Trace Concentrations of Metals LEWIS, D. and KIFF, R.J., 1988, The removal of heavy metals from aqueous effluents by immobilised fungal biomass, Environmental Technology Letters, 9, 991–998. LIEHR, S.K., 1995, Effect of pH on metals precipitation in denitrifying biofilms, Water Science and Technology, 32, 179–183. MASRI, M.S., RANDALL, V.G. and PITTMAN, A.G., 1978, Removal of metallic ions by partially crosslinked polyamine polymers, Polymer Preparation, American Chemical Society, Division of Polymer Chemistry, 19, 483–488. MURALEEDHARAN, T.R., IYENGAR, L. and VENKOBACHAR, C., 1991, Biosorption: an attractive alternative for metal removal and recovery, Current Science, 61, 379– 385. NAKAJJIMA, A. and SAKAGUCHI, T., 1986, Selectiveaccumulation of heavy metals by microorganisms, Applied Microbiology and Biotechnology, 24, 59–64. NEUFIELD, R.D. and HERMANN, E.R., 1975, Heavy metal removal by acclimated activated sludge, Journal of the Water Pollution Control Federation, 47, 310–329. NORBERG, A.B. and RYDIN, S., 1984, Development of a continuous process for heavy metal removal by Zoogloea ramigera, Biotechnology and Bioengineering, 26, 265–268. OLIVER, B.G. and COSGROVE, E.G., 1974, The efficiency of heavy metal removal by a conventional activated sludge plant, Water Research, 8, 869–874. ONSOYEN, E. and SKAUGRUD, O., 1990, Metal recovery using chitosan, Journal of Chemical Technology and Biotechnology, 49, 395–404. PENISTON, Q.P. and JOHNSON, E.L., 1970, Method for treating an aqueous medium with chitosan and derivatives of chitin to remove an impurity, U.S.P. No. 3533940, 1–14. RORRER, G.L., HSIEN, T.Y. and WAY, J.D., 1993, Synthesis of porous-magnetic chitosan beads for removal of cadmium ions from waste water, Industrial Engineering and Chemical Research, 32, 2170–2178. SAKAGUCHI, T. and NAKAJIMA, A., 1991, Accumulation of heavy metals such as uranium and thorium by microorganisms, In: Mineral Bioprocessing, SMITH, R.W. and MISRA, M. (Eds), The Minerals, Metals and Materials Society, Warrendale, Pennsylvania, pp. 309–322. SEO, H. and KINEMURA, Y., 1989, Preparation and some properties of chitosan porous beads, In: Chitin chitosan: sources, chemical, biochemical, physical properties, Proceedings 4th International Conference, SKJAAK-BRAEK, G., ANTHOSEN, T. and SANDFORD, P.A. (Eds), Elsevier, London, pp. 585–588. STERRITT, R.M. and LESTER, J.N., 1981, The influence of sludge age on heavy metal removal, Water Research, 15, 59–65. STOVELAND, S., ASTRUC, M., PERRY, R. and LESTER, J.N., 1979, The balance of heavy metals through a sewage treatment works II: chromium, nickel and zinc, Science of the Total Environment, 12, 25–34. TOBIN, J.M., L’HOMME, B. and Roux, J.C., 1993, Immobilisation protocols and effects on cadmium uptake by Rhizopus arrhizus biosorbents, Biotechnology Techniques, 1, 739– 744. TSEZOS, M., 1988, The performance of a new biological adsorbent for metal recovery. Modelling and experiment results, In: Biohydrometallurgy, NORRIS, P.R. and KELLY, D. P. (Eds), Science and Technology Letters, Kew, pp. 465–475. TSEZOS, M., 1990, Engineering aspects of metal binding by biomass, In: Microbial Mineral Recovery, EHRLICH, H.L. and BRIERLEY, C.L. (Eds), McGraw-Hill, New York, pp. 325–339. TSEZOS, M. and DEUTSCHMANN, A.A., 1990, An investigation of engineering parameters for the use of immobilised biomass particles in biosorption, Journal of Chemical Technology and Biotechnology, 48, 29–39. TSEZOS, M. and DEUTSCHMANN, A.A., 1992, The use of a mathematical model for the study of the important parameters in immobilized biomass biosorption, Journal of Chemical Technology and Biotechnology, 53, 1–12. 139

Biosorbents for Metal lons WALES, D.S. and SAGAR, B.F., 1990, Recovery of metal ions by microfungal filters, Journal of Chemical Technology and Biotechnology, 49, 345–355. WEIDER, R.K., 1989, A survey of constructed wetlands for acid coal mine drainage treatment in the eastern United States, Wetlands, 9, 299–314. WHITE, C. and GADD, G.M., 1990, Biosorption of radionuclides by fungal biomass, Journal of Chemical Technology and Biotechnology, 49, 331–343. WHITLOCK, J.L. and MUDDER, T.I., 1985, The homestake wastewater treatment process: biological removal of toxic parameters from cyanidation wastewaters and bioassay effluent evaluation, Proceedings of the 6th International Symposium on Biohydrometallurgy, Vancouver, British Columbia, pp. 327–339.

140

7

Low-cost Biosorbents: Batch Processes D.A.J.WASE, C.F.FORSTER and Y.S.HO

Introduction This chapter deals with a range of biosorbents and their potential for adsorbing metal ions while, as far as possible, excluding materials dealt with in other chapters. Biological materials have long been known to adsorb metal ions. Early reports spoke of substantial reductions in metal content when aqueous wastes were run through peat bogs, though the precise extent of reduction and the adsorbent life were not documented and almost certainly not understood. Further reports of bioadsorption of metals have continued to this day. Currently, recycling waste biomaterials is politically sound: so is cleaning up the environment. Less welldeveloped countries are expanding their industries, often with serious accompanying pollution. Even well-developed countries like the UK have their share of leachates from waste tips and acid mine drainage. The number of literature reports in this area has recently expanded sharply; the methods, metals adsorbed and materials used are diverse. This, in turn, makes comparisons complex. Peat is perhaps the most common of all the bio-materials examined and provides the most data for comparisons with other bioadsorbents. Even so, Table 7.1 shows wide variation between the amounts of copper taken up by peats from differing sources, the largest value being some fifteen times that of the smallest. Moreover, in some cases (numbers 15, 17, 18, 19 and 21), the maximum adsorption capacity, Q 0 (see Chapter 2), is quoted; in others the figures are arbitrary adsorption capacities, with many of the larger values from chemically pretreated material, and little indication of whether pH is optimal. Nevertheless, as there is so much more information on peat than on any other biosorbent, this material will be used as the base-line for making comparisons.

Peat The extent of metal adsorption on peats varies with factors such as pH and the adsorbed ion. Table 7.2 compares the Q0 values at the optimum pH for chromium, copper, lead and nickel. This shows that, of the four, nickel has the lowest value. 141

Biosorbents for Metal lons Table 7.1 Copper adsorption by peat: a selection of maximum capacities of various peats and related substances from the literature

* Acid-treated; † for energy ; ‡ for agriculture.

The data also show that peat removes a much higher percentage of copper than of nickel over a wide range of metal concentrations (Table 7.3), irrespective of the peat doses (Table 7.4). There is no obvious reason why the absorptive abilities of a particular peat should differ between nickel and, say, copper, an anion of similar size and weight. Nevertheless, this observation about nickel seems typical for many bioadsorbents. For instance, Gould and Genetelli (1978) have suggested that, with anaerobic sludges, there are some sites which, although suitable for other metals, are not available to nickel. Rudd et al. (1984) have noted with a different bioadsorbent, extracellular polymer (ECP) from Klebsiella aerogenes, that it continued to bind copper, cadmium and cobalt even after the complexation capacity had been reached, whereas with nickel it did not. It would thus seem that for a range of bioadsorbents, including peat, if all else is equal, nickel is adsorbed less Table 7.2 A comparison of the maximum capacities of peat (Langmuir Q 0) for chromium (VI), lead (II), copper (II) and nickel (II) at a peat dose of 4 g/l

142

Low-cost Biosorbents: Batch Processes Table 7.3 Metal removal (R) by peat at initial pH of 5 and various initial metal concentrations (C0)

Data from Ho et al., 1994, 1995.

readily than other heavy metals, although in compensation it is bound more tightly, and is more resistant to desorption (Gosset et al., 1986). Precise reasons for this are not fully understood. Finally, this also suggests that the binding phenomenon involves similar mechanisms for a wide variety of bioadsorbents, an important point to bear in mind. Although, then, some evidence suggests similarities in binding/adsorption mechanisms for a wide variety of adsorbents, it is by no means clear how much the process owes to straight physical adsorption, how much arises from ion exchange, and how much is due to chelation. Some light may be thrown on this by considering the structure of peat. Peat is generally dark brown in colour, of spongy consistency and is partially fossilised plant matter which is formed in wet areas under partially anaerobic conditions, where the rate of accumulation of plant matter is greater than that of decomposition, since decomposition is incomplete. Peat is a complex material which is in many ways typical of many bioadsorbents, in that it is derived from woody material. This latter consists mainly of an intimate mixture of cellulose and hemicellulose, much of which can readily be microbially degraded, with organic residues such as lignin (Figure 7.1) and cork-like tissues which are the main constituents of plants. During conversion to peat, almost all the readily degradable components are decomposed. Other substances present can include resins, waxes, proteins, and dyes (Walmsley, 1977). The physical and chemical properties of the peat are said to depend on a variety of factors, such as the nature of the plants from which it has originated, the properties of the water in which the plants were growing (in particular the oxygen status, the pH and the dissolved bases), and the moisture relations during and following its formation and accumulation (Stanek, 1977).

Table 7.4 A compar ison of the removal of copper and nickel by varying the dose of the adsorbent, peat

* Data from Ho et al., 1994, 1995.

143


Figure 7.1 The structure of lignin

The basic component of peat before biomodification is, therefore, seen to be structurally stable, and to comprise aromatic rings with three-carbon side-chains, arranged so that a material whose overall integrity is difficult to modify results. Peat, in contrast, has lost many of the three-carbon side-chains, and has become much more condensed (Figure 7.2). This change allows the development of charged groups, permitting good chelation, and it is this basic structure which makes peat such a good general adsorbent, although in this connection Bloom and McBride (1979) suggest that while ions seem to be located at carboxylate sites, chelation mechanisms or sites of greatly different acid strength are not necessarily involved. In addition to chelation, cation exchange with the various phenolic hydroxyl, heterocyclic and carboxyl groups has been invoked (Martin, 1991), as has hydrogen bonding and anion-cation bonds (Couillard, 1994). In spite of this confusion, it is fairly clear that chemical adsorption can occur in one way or another through the polar functional groups of lignin, which include alcohols, aldehydes, ketones, acids, phenolic hydroxides and ethers as potential providers of chemical bonding (Adler and Lundquist, 1963). The peat particles are surface-active and have a high cation-exchange capacity (90–150 meq/100 g dry matter) (Stewart, 1977). Indeed, Szalay (1969) has reported that a great number of cationic elements are sorbed, and 144

Low-cost Biosorbents: Batch Processes

Figure 7.2 Chelating structure of peat

that electrochemically equivalent quantities are sorbed from cations of different valencies by given peat preparations. Even then there is confusion, for, as Bencheikh-Lehocine (1989) points out, while the main removal mechanism for zinc at acid to neutral pH values was found to be adsorption, at pH values in the alkaline range other processes considerably enhanced zinc removal. Because of the fairly polar character of peat, the specific adsorption for dissolved solids such as transition metals and polar organic molecules is very high. According to Asplund et al. (1972), peat has a high specific surface area (>200 m2/g) and is extremely porous (95%). These two characteristics of peat, its polar and porous nature, clearly help in adsorption. The porosity of peat can be seen in Figure 7.3 (Ho, 1996). In this connection, Poots and McKay (1979) have reported the specific surface area of a peat moss using dye solutions. The specific surface areas depend both on the chemical nature of the solute and adsorbent and also on the molecular dimensions of the solute. A value of 27 m2/g was obtained for nitrogen, 11.8 m2/g for acid dye, and 100 m 2/g for basic dye (Poots and McKay, 1979). This indicates the presence of negatively charged molecules associated with the polar functional groups of peat as well as exchange adsorption occurring with hydrogen ions. As in many natural materials, the natural capacity of peat to retain cations is related to the pH of solution. At pH values above 8.5, peat itself is not stable. At low pH, below 3.0, most metals will be leached from peat. Between these values, most metals are adsorbed fairly efficiently (Coupal and Lalancette, 1976). 145


Essentially, then, the precise mechanisms governing the sorptive activities of peat have received extensive attention over the years, but are still by no means fully understood.

Other biosorbents While peat is a good model system, and is the most studied system, there are many other adsorbent materials. For heavy metals, one can compare biosorbents, such as tree bark, straw, peanut husks, peat moss and insoluble starch xanthate, with other adsorbents such as activated carbon, crushed coal, inorganics like activated alumina, and specially fabricated ion exchangers. Comparisons show that some of these materials, such as ion-exchange resins and activated carbon, are effective but expensive; other adsorbents such as coal and straw are inexpensive and poor adsorbents; still others fail for other reasons, for instance activated alumina is effective in removing heavy metals but is readily soluble under extreme pH conditions. Other, perhaps less stable, bio-materials have also been tested as adsorbents, and for these there is the possibility that still further mechanisms are involved in the adsorption process. For instance, black liquor, a paper industry waste material, arises when the lignin is solubilised by sodium sulphite: a brown tarry highly condensed material results. It is, therefore, hardly surprising that Srivastava et al. (1994) found that uptake of lead and zinc follows the Freundlich and Langmuir models typical of peat, since it is a very similar material. They found that uptake of lead was always greater than that of zinc, and that the sorption capacity increased with increasing pH, as might be expected. Low et al. (1993) used natural and dye-treated oil-palm fibres from al palm-oil mill to remove copper and related heavy metals from solution. The dye coating was found to have a pronounced effect on metal sorption, increasing it very markedly. When the dyed oil-palm fibre was tested on cations from electroplating wastes, they found that it had considerable potential for treating metalpolluted waters. This work has more recently been extended by Quek et al. (1996), who compared oil-palm parenchyma tissue with sago waste and coir. Again, these materials comprise lignocellulosic residues which have partially oxidised, and the adsorption of various cations, notably lead, was shown to obey Freundlich and Langmuir kinetics. It is interesting to note that sago waste, a better-defined material than peat, and one which contains a much higher proportion of cellulose, was much more effective than moss peat for adsorbing lead. Coir and related coir composts have also received attention from other workers. For instance, Guijarro et al. (1996) noted that coir-based compost has a high capacity for nickel removal, while Sharma and Forster (1994) showed that this material adsorbed hexavalent chromium well, particularly at low pH values; pH 2 giving the best results. However, although it was shown in the latter paper that intra-particle diffusion and chemical binding reactions were the rate-controlling steps, the precise mechanisms of metal ion binding were not established. The range of materials used in adsorption trials is, therefore, very considerable, extending from highly humified and polymerised lignocellulosics like peat, through less lignified materials like coir, to substituted cellulosics, and biomass of one sort or another containing substantial amounts of amino material. Most of these materials have been investigated as adsorbents only recently, compared with peat’s long history. Since the adsorptive 146


processes in peat are not, by any means, understood fully, it is hardly surprising that there is still much to learn about these other biosorbents. Peat certainly occupies a central place in the range of bio-materials used as adsorbents for metals. However, although it is essentially beyond the scope of this book to consider pollutants other than metals, it is worth noting one or two unusual ones because of the material they are adsorbed on. For instance, alkyl benzyl sulphonates have been absorbed by algae and algal residues (Fernandez et al., 1995, 1996a), and their use as metal ion adsorbents is discussed elsewhere (Chapter 2). Linseed cattle cake has also been used recently for dye adsorption (Liversidge et al., 1997), and there seems no reason why it should not adsorb metals also. Other studies reported recently have described the use of seeds of Molinga oleifera, pods of Albizia lebbek or compost to remove calcium, and thus soften water (Muyibi and Evison, 1995; Fernandez et al., 1996b). It seems that all sorts of unusual wastes have been tried; modified chitosan for uranium, for instance, and tea leaves for lead, calcium and zinc (Saucedo et al., 1992; Tee and Khan, 1988).

Novel activated carbons In this connection, it is of interest to note that since about 1980, much effort has been put into converting waste materials into activated carbon—with very limited success. Examples range from those that essentially fall outside the range of materials which one can call ‘biowastes’, such as waste car tyres (Lucchesi and Maschio, 1983; Paprowicz, 1990) to true agro-surplus materials. McKay and Roberts (1982) pyrolysed lignocellulosic materials as diverse as computer paper, walnut shells, cattle manure and newsprint. Fruit stones have been used in Spain, principally by Rodriguez-Reinoso and his team who tested olive stones (Rodriguez-Reinoso et al., 1984), almond shells (Rodriguez-Reinoso et al., 1982) and apricot and peach stones (Rodriguez-Reinoso, 1989), and characterised the resultant high-quality microporous carbon (Rodriguez-Reinoso, 1986). Pollard et al. (1992) have tabulated a range of feedstocks proposed for preparation of activated carbon adsorbents, together with appropriate chemical activants (Table 7.5), but admit that ‘carbon adsorption remains an expensive treatment process’: the same conclusion as reached above. Table 7.5 Examples of chemical activant-feedstock couples (from Pollard et al., 1992)

147


Copper It has already been shown that there are wide variations between the amounts of copper taken up by peats from differing sources (Table 7.1). This is, of course, further confused by comparisons of Q 0 with arbitrary adsorption capacities, and the conditions used for sorption tests. Although this makes absolute comparisons essentially impossible, useful information can still be deduced. Table 7.6 lists

Table 7.6 Maximum capacities for adsorption of copper by various adsorbates

148


chronologically, for comparison, amounts of copper taken up by a range of bioadsorbents other than peat, comparing them with peat and typical non-bioadsorbents. Surprisingly, Q0 figures for clays, iron hydroxide and activated carbon, which are frequently used adsorbents, are reasonably comparable with the biosorbents listed, bearing in mind the provisos above. This suggests that biosorbents derived from waste bio-materials should at any rate be economically viable.

Nickel and lead Tables 7.7 and 7.8 list comparable Q0 values for nickel and lead. Again, there are a few sample values for materials other than bioadsorbents for comparison, and again it can be concluded that the performances of many of the biosorbents listed are reasonably comparable with more conventional materials.

Chromium It is also clear from Tables 7.6, 7.7 and 7.8 that many bioadsorbents are as effective as the inorganic materials in common use, and this point will be laboured no Table 7.7 Maximum capacities for adsorption of nickel

149

Biosorbents for Metal lons Table 7.8 A comparison of maximum adsorption capacities for lead of various materials with those of peats

further. There are, however, still other metals to be considered more briefly. Chromium has received considerable attention because its compounds are known to be detrimental to humans and animals. Hexavalent chromium is the most poisonous form, and exposure to dust or mist has caused illnesses. Although less widely used for decorative plating than formerly, chromium still finds wide application in electroplating, metal finishing, paints, particularly yellow road paint, pigments and dyes, and as an anticorrosion agent in cooling waters. While lime precipitation is a 150


stan-dard treatment, preliminary reduction of hexavalent chromium is often necessary, and the contaminated lime slurry itself has to be disposed of. As with other metals, activated carbon is very effective in adsorbing chromium. Indeed, it is clear from Table 7.9 that the best activated carbon still outstrips performance of agro-industrial biowaste adsorbents, but it is a general rule that cost rules out such material, particularly overseas where the need is generally most pressing. Quek et al. (1996) have made similar comments, pointing out that it is only for gold and precious metals that activated carbon is a viable solution, particularly when the means of recovery is to burn the carbon and recover the precious metal. It is therefore interesting to note that coconut shell was used to try to make a suitable activated carbon (Alaerts et al., 1989) long before it was realised that the coconut husk fibre which surrounds it is a much more effective adsorbent, particularly when cost is considered.

Zinc Zinc is another metal which is highly contaminating and very widespread. It occurs in combination with other metals and tends to be liberated when these metals are refined; it is also used widely in galvanising, where the zinc coating is sacrificed rather than the structural integrity of the component. Surprisingly, in view of its ubiquity, there are fewer quantitative data than for many other cations, though there are many reports of its removal by bioadsorbents; representative ones include reports of a peat-based device for continuously removing cations from drinking water (Coupal and Lalancette, 1976) and comparisons with other cations by Gosset et al.

Table 7.9 A comparison of maximum adsor ption capacities of various biosorbents for hexavalent chromium

151

Biosorbents for Metal lons Table 7.10 Adsorption of metal ions onto defatted, extrusion-stabilised rice bran (from Marshall et al., 1993)

(1986). Sludges are dicussed in Chapter 6, and zinc is reported by, for example, Gould and Genetelli (1978). It would appear that Langmuir and Freundlich type isotherms both describe adequately adsorption of zinc by such materials, suggesting that this is true adsorption of the metal, and that little precipitation occurs, though in many cases this is merely of theoretical value, as recovery of the metal is not considered. It is therefore surprising that there is such widespread interest: for instance, one paper reports on eleven Chinese peats as being suitable for zinc removal, though data are scant, while another confirms a 93–96% removal efficiency (Zhipei et al., 1984; Bencheikh-Lehocine, 1989). Biosorbents such as peat are likely, therefore, to be of value in applications such as the treatment of landfill site leachate, where Brady (1988) discusses zinc adsorption. As far as other adsorbents are concerned, a notable paper is that of Marshall et al. (1993), who used a range of rice products including rice hulls, rice bran and defatted bran for zinc adsorption. The defatted material was superior to the untreated bran: presumably the solvents removed fat from pores, making more adsorption sites available. Comparisons with other metal ions confirmed that, as with other materials, even this novel material adsorbed nickel less readily than the other cations such as zinc (see Tables 7.2 and 7.10). Although there are fewer data for zinc than for many other cations, typical values are given in Table 7.11.

Table 7.11 Biosorbents for zinc

152


Manganese Most of the work on manganese has been done in connection with soils and soil organic matter. Thus, Bloom and McBride (1979) noted that manganese and cobalt were bound to peat during electromagnetic spin resonance studies aimed at establishing binding mechanisms. They concluded that, except for copper, metals were bound as hydrated ions. This explains the high affinity of copper for peats: indeed, Eger et al. (1980) have reported the ability of white cedar bog to adsorb both copper and cobalt.

Cobalt and cadmium The amounts of cobalt and cadmium have been shown to vary with the depth of the peat layer in situ; however, it was not possible to explain this merely on the basis of Langmuir or Freundlich adsorption (Markert and Thornton, 1990). In contrast, Bunzl et al. (1976) suggested that adsorption and desorption of a range of metal ions on peat and humic acids could be modelled, and claimed that cadmium behaved similarly to zinc in respect of adsorption and desorption, though their results were only from the laboratory, and they did not confirm them with on-site measurements. Similar early reports for cadmium are those of Coupal and Lalancette (1976) and Gosset et al. (1986); the latter reported maximum capacities of 180 mmol/kg for eutrophic peat and 200 mmol/kg for oligotrophic peat. It therefore appears that the situation in situ is more complex than had earlier been realised. There are few quantitative data on cadmium for other biosorbents. Tee and Khan (1988) have shown that cadmium can be removed by waste tea leaves, though they provide little comparative information, and is more strongly held to sludges than zinc if one uses a gravimetric rather than a molar basis (Gould and Genetelli, 1978). As far as other metals are concerned, mercury concentrations have been shown to decrease, while cadmium and zinc sulphides were simultaneously removed from polluted waters (McLellan and Rock, 1988). Parkash and Brown (1976) have used peat and coal for recovering zirconium from solution, while iron concentrations in acidic mine waters were shown to be reduced by between 50% and 96% in 25 specially constructed wetlands in the USA by Girts and Kleinmann (1986), and selenium, chromium, arsenic, beryllium and silver concentrations were all seen to be reduced in sanitary landfill leachates when peat beds were used (Brady, 1988). There is clearly a considerable gap between reports of data from the laboratory and in situ semiqualitative information, and our understanding of the adsorption-desorption processes underlying these reports.

Competitive adsorption Multi-ion systems are the norm for such metal wastes as effluents or landfill leachates, which have received far less attention than single-ion systems. One of the 153


more comprehensive studies (Trujullo et al., 1991), investigating competitive adsorption of six metal ions from a single solution, led to a model applicable to their batch and semi-continuous packed beds. This was used to predict equilibrium constants, mass-transfer coefficients and adsorption capacities. Binding capacity of their adsorbent was highest for copper, independent of the other ions, and copper also exerted the largest competing effect. A very similar order of affinity— Pb 2+>Cu 2+>Cd 2+>Mn 2+—was reported by Pakarinen et al. (1980), confirming the more restricted results of Bunzl et al. (1976), who listed Pb 2+ >Cu 2+ >Cd 2 ⬇Zn2+ >Ca2+ in the pH range 3.5–4.5. Masslenilov and Kiselva (1989) mentioned a generally similar order of cation exchange, Cu 2+>Zn2+>Fe 3+>Ca2+ , and Brady et al. (1994) also came to roughly the same conclusions. However, with granulated peat, Chistova et al. (1990) found differences, giving an order Fe 3+>Cu2+>Cr 3+>Zn2+>Ni 2+. Since adsorption is claimed to depend as much on ‘impurities’ in the peat as anything else (Couillard, 1994), such differences are hardly surprising. Low et al. (1993) tested dye-treated and natural oil-palm fibres, and Kumar and Dara (1982) treated bagasse, while Suemitsu et al. (1986) used red and yellow dyestuff-treated rice hulls, finding differences between the two and the untreated hulls. Ho et al. (1996) examined multicomponent systems, finding that the competitive effect affected the three ions in the order nickel>lead>copper. It therefore seems that metals are not necessarily adsorbed by exactly similar mechanisms for all bioadsorbents, and that each needs to be tested to determine its characteristics. Adsorption of lead is usually greater than of copper, though copper is a more aggressive competitor, and adsorption of nickel is usually weaker than that of the others, though once adsorbed it is firmly bound.

Practical aspects of using peat During the past 20 years, most bioadsorbent work on full-scale metal and organic removal from industrial waste streams has been focused on peat. The adsorption capacity of peat varies with the type of material being adsorbed, but because of its relatively low-cost compared to more conventional adsorbents, it can still compete successfully for particular applications. The periodic table (Table 7.12) illustrates the elements that can be sorbed in peat; the peat industry has been well established in many countries and employs, directly or indirectly, specialists in such diverse fields as engineering, agriculture, forestry, horticulture, geobotany, anthropology, chemistry, and medicine (Stewart, 1977). There is therefore a wealth of experience to draw on. The majority of potential processes are along the lines proposed by Ruel et al. (1977), who proposed a division into three basic steps, as follows. (1)

Contacting. The wastewater is contacted with peat after treatment with a precipitating agent such as sodium sulphide.

(2)

Dewatering and drying. If regeneration is not economic, the wet peat needs dewatering to remove part of the mechanically held water, with some further drying, if this is possible, to assist handling and transport to the central incinerator.

(3)

Recovery of the metals by burning the peat. Since the ash content of peat is very small, combustion will result in formation and easy recovery of metals from

154

Low-cost Biosorbents: Batch Processes Table 7.12 Perodic table illustrating the elements sorbed by peat (information from Szalay, 1969 with additional data from Parkash and Brown, 1976; Coupal and Lalancette, 1976; Illarionovich et al., 1976; Pakarinen et al., 1980; Rock et al., 1984; Smith et al., 1984; McLellan and Rock, 1988; Sheppard et al., 1989; Trujullo et al., 1991; James et al., 1992; Green et al., 1994)

their oxides. However, for some metals, notably lead, the oxides disperse in the air and are reprecipitated over a wide area. Lead residues occur particularly around busy city motorway junctions, leading to concerns regarding children, as lead is an accumulative poison. The problem of recovering lead seems currently ill-addressed. An example where regeneration can be applied is that of Moore (1954), who reported on the ability of peat to adsorb uranium from solution: a maximum of 98.0% uranium was removed from aqueous solution by peat. In contrast to such a specialised system, the first full-scale peat system for water pollution control was a sacrificial peat system; constructed in 1978, it has consistently produced high-quality effluent. Coupal and Lalancette (1976) give working drawings for a 20000 gallon/day plant capable of removing a range of metal ions (1 gallon=4.5 1). Flow-through bogs have also been successfully used to remove metals (e.g. Eger et al., 1980; Aho and Tummavuori, 1984), though no one has ever hazarded a guess at the possible life of such a process, and peat on-site wastewater treatment systems are now in operation in Maine, Alaska, Canada, and Ireland, treating wastewater and protecting groundwater on sites not appropriate for conventionally designed treatment systems (Brooks, 1991). All these are examples of systems where the emphasis is on ion removal, and the peat is treated as a utilisable, disposable resource. 155

Biosorbents for Metal lons Table 7.13 Cost comparison of wastewater treatment systems at Asarco’s Globe plant (information from Green et al., 1994)

1 gallon=4.5 l.

Over the past few years, however, the political emphasis has swung strongly towards using only those resources which can readily be renewed. This has had two consequences. In the first place, systems involving peat have become much more engineered and much more specific. One example is that of Harrison Western Environmental Services Inc. of Lakewood, Colorado, which developed a process, called membrane-media extraction, which uses peat moss ‘capsules’, effective at reducing concentrations of arsenic, cadmium, lead, nickel, selenium, and other metals from electroplating rinsewater, pulp and paper mill discharge, municipal wastewater, and acid mine drainage (Green et al., 1994). The performance and cost of the membrane-media extraction system compared to the existing chemical precipitation unit at the ASARCO plant is illustrated in Tables 7.13 and 7.14. Another alternative is to take a pure component of the system and use it in a loadregeneration cycle system. Examples of this type of system can involve chemically

Table 7.14 Performance of chemical precipitation versus the retrofitted membrane-media system at Asarco’s Globe plant (information from Green et al., 1994)

156


modified cellulosics, for example, as the pure adsorbent, but information about them is currently scanty, although it is known that column-based systems such as the above have found application in recovering specific metals from waste industrial components. It has been observed above that sago waste, a cellulosic, is more effective than peat in lead adsorption (Quek et al., 1996). Perhaps chemically modified fabricated particles or packings for such biowastes point the way to future developments. Even with such developments, it is clear that there is still considerable mileage in peat as a metal adsorbent, since it removes such a wide range of cations. While it must be appreciated that the amount of ion adsorbed at any one time is not an absolute figure, and the information in Table 7.1 suggests wide variabilities between amounts of ion adsorbed under similar conditions by differing samples, peat-based systems work well. As far as comparisons between different sorts of peat are concerned, it is probably better merely to list the range of metal ions adsorbed by peat qualitatively, by the modified periodic table mentioned earlier (Table 7.12). The second consequence is the use of non-peat substances which are renewable, and which are already waste materials. It has been shown earlier that although absorption capacities of peat vary, it is likely for some applications to be competitive with more conventional adsorbents, because of its low-cost. This must also be a necessary criterion for the successful application of a biowaste, particularly as its capacity is usually somewhat less than peat. However, few reports of commercial applications of such material in full scale systems exist. Overall, then, the only comprehensively investigated bioadsorbent is probably peat. Peat is still freely available, and has been shown to adsorb a wide range of metal ions. It can be made into capsules, and used in continuous devices. It can be burned to recover many of the sorbed metals, with notable exceptions such as lead. As pointed out above, there is just one snag: supplies are finite, and this is probably not the best use one can make of this resource. To match the properties of peat one needs a lowcost or zero-cost waste, which is relatively clean, will form beds and cohere readily, with the physical and sorptive properties of the peat it is to replace. The search has only recently begun. It is our belief that it will be solved not by a single substance but by a range of such substances. We have no doubt that the search will be successful.

References ADLER, E. and LUNDQUIST, K., 1963, Spectrochemical estimation of phenyl coumaran elements lignin, Acta Chemica Scandinavia, 17, 13–26. AHO, M. and TUMMAVUORI, J., 1984, On the ion-exchange properties of peat—IV. The effect of experimental condition on ion exchange properties of sphagnum peat, Suo, 35, 47–53. AKSU, Z., SAG, Y. and KUTSAL, T., 1992, The biosorption of Cu(II) by C. vulgaris and Z. ramigera, Environmental Technology, 13, 579–586. ALAERTS, G.J., JITATURUNT, V. and KELDERMAN, P., 1989, Use of coconut shell-based activated carbon for chromium(VI) removal, Water Science and Technology, 21, 1701– 1704. ALLEN, S., BROWN, P., M C KAY, G. and FLYNN, O., 1992, An evaluation of single resistance transfer models in the sorption of metal ions by peat, Journal of Chemical Technology and Biotechnology, 54, 271–276. 157

Biosorbents for Metal lons ALVIM-FERRAZ, M.C., 1988, Preparation of activated charcoal for air pollution control, Fuel, 67, 1237–1241. ASPLUND, D., EKMAN, E. and THUN, R., 1972, Counter-current peat filtration of waste water, Proceedings of the 4th International Peat Congress, Poznan, pp. 358–371. BANSAL, R.P., DONNET, J.P. and STOEKLI, F., 1988, Active Carbon, Marcel Dekker, New York. BENCHEIKH-LEHOCINE, M., 1989, Zinc removal using peat adsorption, Environmental Technology Letters, 10, 101–108. BLOOM, P.R. and MCBRIDE, M.B., 1979, Metal ion binding and exchange with hydrogen ions in acid-washed peat, Soil Science Society of America Journal, 43, 687–692. BOWERS, A.R. and HUANG, C.P., 1981, Activated carbon process for the treatment of chromium (VI) containing industrial wastewaters, Water Science and Technology, 23, 629– 650. BRADY, M.J., 1988, A case study of sanitary landfill leachate attenuation by a natural claypeat soil in the Gulf Coast region, American Institution of Chemical Engineering Symposium Series, 84, 192–197. BRADY, D., STOLL, A. and DUNCAN, J.R., 1994, Biosorption of heavy metal cations by non-viable yeast biomass, Environmental Technology, 15, 429–438. BROOKS, J.L., 1991, Peat as a raw material for waste water treatment, Proceedings of seminar at Bord na Mona Peat Research Centre, FUCHSMAN, C.H. (Ed.), Newbridge, Ireland, pp. 13–27. BUNZL, K., SCHMIDT, W. and SANSONI, B., 1976, Kinetics of ion exchange in soil organic matter—IV. Adsorption and desorption of Pb2+, Cu2+, Cd2+, Zn2+ and Ca2+, Journal of Soil Science, 27, 32–41. CHEN, X.H., GOSSET, T. and THEVENOT, D.R., 1990, Batch copper ion binding and exchange properties of peat, Water Research, 24, 1463–1471. CHISTOVA, L., ROGACH, L.M., SOKOLOVA, T.V. and PEKHTEREVA, V.S., 1990, Removal of heavy metal ions from electroplating wastewaters by granulated peat, Torf. Prom-st., 2, 25–28. COUILLARD, D., 1994, The use of peat in waste water treatment, Water Research, 28, 1261– 1274. COUPAL, B. and LALANCETTE, J.M., 1976, The treatment of wastewaters with peat moss, Water Research, 10, 1071–1077. DISSANAYAKE, C.B. and WEERASOORIYA, S.V. R., 1981, Peat as a metal-trapping material in the purification of industrial effluents, International Journal of Environmental Studies, 17, 233–238. EGER, P., LAPAKKO, K. and OTTERSON, P., 1980, Trace metal uptake by peat: introduction of a white cedar bog and mining stockpile, Proceedings of the 6th International Peat Congress, Duluth, Minnesota, pp. 542–547. EHRBURGER, P., ADDOUN, A., ADDOUN, F. and DONNET, J. DE-P., 1986, Carbonisation of coals in the presence of alkaline hydroxides and carbonates: formation of activated carbons, Fuel, 65, 1147–1149. FARRAH, H., HATTON, D. and PICKERING, W.F., 1980, The affinity of metal ions for clay surfaces, Chemical Geology, 28, 55–68. FERNANDEZ, N.A., CHACIN, E., GARCIA, C., ALASTRE, N., LEAL, F. and FORSTER, C.F., 1996b, Pods from Albizia lebbek as a novel water softening biosorbent, Environmental Technology, 17, 541–546. FERNANDEZ, N.A., CHACIN, E., GUTIERRES, E., ALASTRE, N., LLAMOSA, B. and FORSTER, C.F., 1995, Adsorption of lauryl benzene sulphonate on algae, Bioresource Technology, 54, 111–115. FERNANDEZ, N.A., CHACIN, E., GUTIERRES, E., ALASTRE, N., LLAMOSA, B. and FORSTER, C.F., 1996a, An analysis of the absorption of alkyl benzene sulphonate by algal residues, Process Biochemistry, 31, 333–336. 158

Low-cost Biosorbents: Batch Processes FERRO-GARCÍA, M.A., RIVERA-UTRILLA, J., RODRIGUEZ-GORDILLO, J. and BAUTISTA-TOLEDO, I., 1988, Adsorption of zinc, cadmium and copper on activated carbons obtained from agricultural by-products, Carbon, 26, 363–373. GIRTS, M.A. and KLEINMANN, R.L. P., 1986, Constructed wetlands for treatment of acid mine drainage: a preliminary review. Proceedings 1986 National Symposium on Mining, Hydrology, Sedimentology and Reclamation, University of Kentucky, Lexington, pp. 166– 171. GONZALEZ-DAVILA, M. and MILLERO, F.J., 1990, The adsorption of copper to chitin in sea-water, Geochemica et Cosmochimica Acta, 54, 761–768. GOSSET, T., TRANCART, J.L. and THEVENOT, D.R., 1986, Batch metal removal by peat: kinetics and thermodynamics, Water Research, 20, 21–26. GOULD, M.S. and GENETELLI, E.J., 1978, Heavy metal complexation behavior in anaerobically digested sludges, Water Research, 12, 505–512. GREEN, D.H. et al., 1994, Peat moss beads remove metals from wastewater, The Hazardous Waste Consultant, 12, 1.19–1.21. GUIJARRO, L., WASE, D.A.J. and FORSTER, C.F., 1996, Investigation of nickel removal from aqueous solutions using natural adsorbents. Proceedings of the 1996 L Chem. E. Research Event/Second European Conference for Young Researchers, Institution of Chemical Engineers, Rugby, pp. 624–626. HASSLER, J.W., 1967, Activated Carbon, Leonard Hill, London. HITCHCOCK, S.J., MCENANEY, B. and WATLING, S.J., 1983, Fibrous active carbons from coir, Journal of Chemical Technology and Biotechnology, 33A, 157–163. Ho, Y.-S., 1996, Adsorption of Heavy Metals from Waste Streams by Peat, PhD Thesis, The University of Birmingham, Birmingham, UK. Ho, Y.-S., WASE, D.A. J. and FORSTER, C.F., 1994, The adsorption of divalent copper ions from aqueous solution by sphagnum moss peat, Transactions of the Institution of Chemical Engineers Part B, 17, 185–194. Ho, Y.-S., WASE, D.A. J. and FORSTER, C.F., 1995, Batch nickel removal from aqueous solution by sphagnum moss peat, Water Research, 29, 1327–1332. Ho, Y.-S., WASE, D.A. J. and FORSTER, C.F., 1996, Removal of lead ions from aqueous solution using sphagnum moss peat as adsorbent, Water SA, 22, 219–224. HUANG, C.P. and Wu, M.H., 1977, The removal of chromium(VI) from dilute aqueous solution by activated carbon, Water Research, 11, 673–679. HUANG, J.P., HUANG, C.P. and MOREHART, A.L., 1991, Removal of heavy metals by fungal (Aspergillus oryzae) adsorption , In: Trace Metals in the Environment: I; Heavy Metals in the Environment. VERNET, J.P. (Ed.),Elsevier, Amsterdam, The Netherlands, pp. 329–349. ILLARIONOVICH, B.P., ALEXANDROVNA, G.K. and ROMANOVNA, C.L., (1976), Peat and peatland in the natural environment protection. Proceedings of the 5th International Peat Congress, Poznan, pp. 328–348. JAMES, B.R., RABENHORST, M.C. and FRIGON, G.A., 1992, Phosphorus sorption by peat and sand amended with iron oxides or steel wool, Water and Environmental Research, 64, 699–705. KIM, J.I. and ZOLTEK, J., 1977, Chromium removal with activated carbon. Progress in Water Technology, 9, 143–155. KORCZAC, M. and KURBIEL, J., 1989, A new adsorbent for wastewater treatment, In: Chemistry for the Protection of the Environment, PAWLOWSKI, L., MENTASTI, E. and LACEY, W. (Eds), Elsevier Science Publishers, Amsterdam, The Netherlands. KUMAR, P. and DARA, S.S., 1982, Utilization of agricultural wastes for decontaminating industrial/domestic wastewaters from toxic metals, Agricultural Wastes, 4, 213–223. LAINE, J., CALFAT, A. and LABADY, M., 1989, Preparation and characterisation of activated carbons from coconut shell impregnated with phosphoric acid, Carbon, 27, 191–195. 159

Biosorbents for Metal lons LEE, C.K. and Low, K.S., 1989, Removal of copper from solution using moss, Environmental Technology Letters, 10, 395–404. LIVERSIDGE, R.M., LLOYD, G.J., WASE, D.A.J. and FORSTER, C.F., 1997, Removal of Basic Blue 41 from aqueous solution by linseed cake, Process Biochemistry, 32, 257–264. LOPEZ-GONZALEZ,J.DED.,VALENZUELA-CALAHORRO,C.,NAVARETTE-GUIJOSA, A. and GOMEZ-SERRANO, V., 1980, Preparation and characterisation of active carbons from olive stones, Carbon, 18, 413–418. Low, K.S. and LEE, C.K., 1987, The sorption characteristics of lead and copper by the moss Calymperes delessertii Besch, Pertanika, 10, 321–326. Low, K.S., LEE, C.K. and LEE, K.P., 1993, Sorption of copper by dye-treated oil-palm fibers, Bioresource Technology, 44, 109–112. Low, K.S., LEE, C.K. and LEO, A.C., 1995, Removal of metals from electroplating wastes using banana pith, Bioresource Technology, 51, 227–231. LUCCHESI, A. and MASCHIO, G., 1983, Semi-active carbon and aromatics produced from the pyrolysis of scrap tyres, Conservation and Recycling, 6, 85–90. M C KAY, D.M. and ROBERTS, P.V., 1982, The dependence of char and carbon yield on lignocellulosic precursor composition, Carbon, 20, 87–94. MCLELLAN, J.K. and ROCK, C.A., 1988, Pretreating landfill leachate with peat to remove metals, Water, Air and Soil Pollution, 37, 203–215. MARKERT, B. and THORNTON, I., 1990, Multi-element analysis of an English peat bog soil, Water, Air and Soil Pollution, 49, 113–123. MARSHALL, W.E., CHAMPAGNE, E.T. and EVANS, W.J., 1993, Use of rice milling byproducts (hulls and bran) to remove metal ions from aqueous solution, Journal of Environmental Science and Health, A28, 1977–1993. MARTIN, A.M., 1991, Peat as an agent in biological degradation: peat biofilters, In: Biological Degradation of Wastes, MARTIN, A.M. (Ed.), Elsevier, London, pp. 341– 362. MASSLENILOV, B.I. and KISELVA, S.A., 1989, Physico-chemical basis for the use of peat in ion-exchange technology and adsorption processes, Torfianaia Promyshlennost, 5, 23–25. MOORE, G.W., 1954, Extraction of uranium from aqueous solution by coal and some other materials, Economic Geology, 49, 652–658. MUSTAFA, S. and HAQ, I., 1988, Adsorption of Cu(II), Co(II) and Ni(II) on amorphous iron hydroxide from aqueous electrolyte solution, Environmental Technology Letters, 9, 1386– 1397. MUYIBI, S.A. and EVISON, L.M., 1995, Moringa oleifera seeds for softening hard water, Water Research, 29, 1099–1104. NAGENDRA ROA, C.R., IYENGAR, L. and VENKOBACHAR, C., 1993, Sorption of copper(II) from aqueous phase by waste biomass, Journal of Environmental Engineering Division, Proceedings of the American Society of Civil Engineers, 119, 369–377. NIU, H., Xu, X.S. and WANG, J.H., 1993, Removal of lead from aqueous solutions by penicillium biomass, Biotechnology and Bioengineering, 42, 785–787. ONG, H.L. and SWANSON, V.E., 1966, Adsorption of copper by peat, lignite and bituminous coal, Economic Geology, 61, 1214–1231. ÖZER, D., AKSU, Z., KUTSAL, T. and CAGLAR, A., 1994, Adsorption isotherms of lead (II) and chromium (VI) on Cladophora crispata, Environmental Technology, 15, 439–448. PAKARINEN, P., TOLONEN, K. and SOVERI, J., 1980, Distribution of trace metals and sulfur in the surface peat of Finnish raised bogs, Proceedings of the 6th International Peat Congress, Duluth, Minnesota, pp. 645–648. PANDAY, K.K., PRASAD, G. and SINGH, V.N., 1985, Copper(II) removal from aqueous solution by fly ash, Water Research, 19, 869–873. PAPROWICZ, J.T., 1990, Activated carbons for phenols removals from waste waters, Environmental Technology, 11, 71–82. 160

Low-cost Biosorbents: Batch Processes PARKASH, S. and BROWN, R.A. S., 1976, Use of peat and coal for recovering zirconium from solution , Canadian Mining and Metallurgical Bulletin, 69, 59–64. POLLARD, S.J.T., FOWLER, G.D., SOLLARD, C.J. and PERRY, R., 1992, Low cost adsorbents: a review, Science of the Total Environment, 116, 31–52. POOTS, V.J.P. and MCKAY, G., 1979, The specific surfaces of peat and wood, Journal of Applied Polymer Science, 23, 1117–1129. QUEK, S.Y., FORSTER, C.F. and WASE, D.A.J., 1996, Adsorption of lead by natural organic waste materials, Proceedings of the 1996 I. Chem. E. Research Event/Second European Conference for Young Researchers, Institution of Chemical Engineers, Rugby, pp. 570– 572. RENOUPREZ, A. and AVOM, J., 1988, Characterisation of active carbons from palm tree fibres using nitrogen adsorption and small angle X-ray scattering, In: Characterisation of Porous Solids, UNGER, K.K., ROUQUEROL, J. and SING, K.S.W. (Eds), Elsevier Science Publishers, Amsterdam, The Netherlands, pp. 49–54. ROCK, C.A., BROOKS, J.L., BRADEEN, S.A. and STRUCHTMEYER, R.A., 1984, Use of peat for on-site wastewater treatment: I; Laboratory evaluation, Journal of Environmental Quality, 13, 518–523. RODRIGUEZ-REINOSO, F., 1986, Preparation and characterisation of activated carbons, Proceedings of the NATO Advanced Study Institute on Carbon and Coal Gasification, NATO ASI Series E, 105, Martinus Nijhoff Publishers, Dordrecht, The Netherlands, pp. 603–643. RODRIGUEZ-REINOSO, F., 1989, Process for the preparation of activated carbon, European Patent 329 251. RODRIGUEZ-REINOSO, F., LINARES-SOLANO, A., MOLINA-SABIO, M. and LOPEZGONZALEZ, J. DE D., 1984, The two-stage air-CO2 activation in the preparation of activated carbons. I: Characterisation by gas adsorption, Advances in Science and Technology, 1, 211–222. RODRIGUEZ-REINOSO, F., LOPEZ-GONZALEZ, J. DE D. and BERENGUER, C., 1982, Activated carbons from almond shells, I—Preparation and characterisation by nitrogen adsorption, Carbon, 20, 513–518. ROWLEY, A.G., HUSBAND, F.M. and CUNNINGHAM, A.B., 1984, Mechanism of metal adsorption from aqueous solutions by waste tyre rubber, Water Research, 18, 981–984. RUDD, T., STERRITT, R.M. and LESTER, J.M., 1984, Complexation of heavy metals by extracellular polymers in the activated sludge process, Journal Water Pollution Control Federation, 56, 1260–1268. RUEL, M., CHORNET, S., COUPAL, B., AITCIN, P. and COSSETTE, M., 1977,Industrial utilization of peat moss, In: Muskeg and the Northern Environment in Canada, RADFORTH, N.W. and BRAWNER, C.O. (Eds), University of Toronto Press, Toronto and Buffalo, pp. 221–246. SAUCEDO, I., GUIBAL, E., ROULPH, Ch. and LE CLOIREC, P., 1992, Sorption of uranyl ions by a modified chitosan: kinetic and equilibrium studies, Environmental Technology, 13, 1101–1105. SHARMA, D.C., 1994, An examination into the treatment of hexavalent chromium using lowcost adsorbents, PhD Thesis, University of Birmingham, Birmingham. SHARMA, D.C. and FORSTER, C.F., 1993, Removal of hexavalent chromium, using sphagnum moss peat, Water Research, 27, 1201–1208. SHARMA, D.C. and FORSTER, C.F., 1994, A preliminary examination into the adsorption of hexavalent chromium using low-cost adsorbents, Bioresource Technology, 47, 257– 264. SHARMA, Y.C., PRASAD, G. and RUPAINWAR, D.C., 1991, Removal of Ni(II) from aqueous solutions by sorption, The International Journal of Environmental Studies, Section B; Environmental Science and Technology, 37, 183–191. SHEPPARD, M.I., THIBAULT, D.H. and SMITH, P.A., 1989, Iodine dispersion and effects on 161

Biosorbents for Metal lons groundwater chemistry following release to a peat bog, Manitoba, Canada, Applied Geochemistry, 4, 423–432. SMITH, C.M., NAVRATIL, J.D. and MACCARTHY, P., 1984, Removal of actinides from radioactive wastewaters by chemically modified peat, Solvent Extraction and Ion Exchange, 2, 1123–1149. SMITH, E.F., MACCARTHY, P., Yu, T.C. and MARK, H.B., JR, 1977, Sulfuric acid treatment of peat for cation exchange, Journal of the Water Pollution Control Federation, 49, 633– 638. SMITH, E.F., MARK, H.B., JR, and MACCARTHY, P., 1978, Investigation of chemically modified forms of peat as inexpensive means of wastewater treatment, In: Chemistry of Wastewater Technology, RUBIN, A.J. (Ed.), Ann Arbor Science Publishers, Ann Arbor, Michigan, pp. 349–372. SRIVASTAVA, S.K., PANT, N. and PAL, N., 1987, Studies on the efficiency of a local fertilizer waste as a low-cost adsorbent, Water Research, 21, 1389–1394. SRIVASTAVA, S.K., SINGH, A.K. and SHARMA, A., 1994, Studies on the uptake of lead and zinc by lignin obtained from black liquor—a paper industry waste material, Environmental Technology, 15, 353–361. SRIVASTAVA, S.K., TYAGI, R. and PANT, N., 1989, Adsorption of heavy metal ions on carbonaceous material developed from the waste slurry generated in local fertilizer plants, Water Research, 23, 1161–1165. STANEK, W., 1977, Classification of muskeg, In: Muskeg and the Northern Environment in Canada, RADFORTH, N.W. and BRAWNER, C.O. (Eds), University of Toronto Press, Toronto and Buffalo, 31–62. STEWART, J.M., 1977, Canadian muskegs and their agricultural utilisation. In: Muskeg and the Northern Environment in Canada, RADFORTH, N.W. and BRAWNER, C.O. (Eds), University of Toronto Press, Toronto and Buffalo, pp. 208–220. SUEMITSU, R., UENISHI, R., AKASHI, I. and NAKANO, M., 1986, The use of dyestufftreated rice hulls for removal of heavy metals from waste water, Journal of Applied Polymer Science, 31, 75–83. SUZUKI, A., SEKI, H. and MARUYAMA, H., 1994, An equilibrium study of adsorption of divalent metal ions onto a metal oxide adsorbent, Journal of Chemical Engineeering of Japan, 27, 505–511. SZALAY, A., 1969, Accumulation of uranium and other micrometals in coal and organic shales and the role of humic acids in these geochemical enrichments, Arkiv For Mineralogi Och Geologi, 5, 23–55. SZARBO, I., 1958, Adsorption of cations on humus preparations, Communications of the Third Maths-Physics Class of the Hungarian Academy of Sciences, 8, 393–402. (In Hungarian.) TAN, W.T., OOI, S.T. and LEE, C.K., 1993, Removal of chromium(VI) from solution by coconut husk and palm press fibres, Environmental Technology, 14, 277–282. TEE, T.W. and KHAN, A.R. M., 1988, Removal of lead, cadmium and zinc by waste tea leaves, Environmental Technology Letters, 9, 1223–1232. TIEN, C.T. and HUANG, C.P., 1991, Formation of surface complexes between heavy metals and sludge particles, In: Trace Metals in the Environment: 1. Heavy Metals in the Environment, VERNET, J.P. (Ed.), Elsevier, Tokyo, pp. 295–311. TRUJULLO, E.M., JEFFERS, T.H., FERGUSON, C. and STEVENSON, Q., 1991, Mathematically modeling the removal of heavy metals from a wastewater using immobilized biomass, Environmental Science and Technology, 25, 1559–1564. TUMMAVUORI, J. and AHO, M., 1980, On the ion-exchange properties of peat. Part I: on the adsorption of some divalent metal ions (Mn2+, Co 2+, Cu2+, Zn2+, Cd2+ and Pb2+) on the peat, Sou, 31, 45–51. VIRARAGHAVAN, T. and DRONAMRAJU, M.M., 1993, Removal of copper, nickel and zinc from wastewater by adsorption using peat, Journal of Environmental Science and Health, Part A—Environmental Science and Engineering, 28, 1261–1276. 162

Low-cost Biosorbents: Batch Processes WALMSLEY, M.E., 1977, Physical and chemical properties of peat, In: Muskeg and the Northern Environment in Canada, RADFORTH, N.W. and BRAWNER, C.O. (Eds), University of Toronto Press, Toronto and Buffalo, pp. 82–129. YADAVA, K.P., TYAGI, B.S. and SINGH, V.N., 1989, Fly-ash for the treatment of water enriched in lead(II), Journal of Environmental Science and Health, Part A—Environmental Science and Engineering , 24, 783–808. YADAVA, K.P., TYAGI, B.S. and SINGH, V.N., 1991, Effect of temperature on the removal of lead (II) by adsorption on chinal clay and wollastonite, Journal of Chemical Technology and Biotechnology, 51, 47–60. ZHIPEI, Z., JUNLU, Y., ZENGHUI, W. and PIYA, C., 1984, A preliminary study of the removal of Pb2+, Cd2+, Zn 2+, Ni2+ and Cr6+ from wastewater with several Chinese peats, Proceedings of the 7th International Peat Congress, Dublin, Ireland, pp. 147–152.

163

8

Biosorption Using Unusual Biomasses R.G.J.EDYVEAN, C.J.WILLIAMS, M.W.WILSON and D.ADERHOLD

Introduction The term ‘biosorption’ has come to be accepted as referring to ‘passive’ or physicochemical attachment of a sorbate to a biomass, essentially the binding of a chemical species to biopolymers. The definition, thus, specifically excludes metabolic or ‘active’ uptake by living, metabolising, cells. When biosorption occurs in living biomass, the physico-chemical step (biosorption) is often followed by active uptake across the cell membrane and into the cell itself. Such a step usually requires energy from the cell for the transfer process and can be highly selective and often irreversible unless the living system is destroyed. This process is often termed bioaccumulation (Volesky, 1990). Thus, while biosorption and bioaccumulation can occur on both living and dead biomass, only biosorption occurs on dead biomass. Biosorption tends to be very rapid and reversible while bioaccumulation tends to be slower and irreversible. The nature of the binding processes in biosorption are largely unknown. Gadd (1990) comments that ‘a large variety of physical, chemical and biological mechanisms may be involved, including adsorption, precipitation, complexation and transport’. With respect to metals, Hunt (1986) notes that ‘As a generalisation the binding of metal ions involved two mechanisms, the first of these being simple ionexchange and the second through the formation of complexes (co-ordination compounds) which may be chelates. Because of the complexity of most biopolymers it is very likely that both of these processes of binding will take place in a system at the same time.’ Essentially, biosorption can be represented as an adsorption process (transfer from a bulk fluid to a surface), and Gadd (1992) has stated that ‘virtually all biological material has biosorptive properties’. An increasing range of biological, and biologically derived, materials have been investigated for their biosorptive properties. The realisation that biosorption occurs on dead biomass and the development of industrially useful processes have been slow but lead to interesting possibilities for low-cost industrial biosorption systems. Such systems may be based on biomasses derived from waste products from other industries or low-cost material harvested or collected for the purpose of biosorption. This chapter describes some of the more unusual biosorbents and concentrates in more detail on two of them. 165


Types of biomass

Seaweeds and seaweed derivatives There are a considerable number of scientific publications on the metal binding properties of microscopic algae, but fewer on macroscopic algae (seaweeds). The first studies were into the use of algae as ‘bioindicators’ of pollution, and from these developed studies of metal bioaccumulation for pollution control. Eventually, studies started into the biosorption properties of macroalgae (seaweeds) in an effort to remove the expense and problems associated with growing biomass and maintaining living systems. References to macroalgae (seaweeds) are far fewer than for microalgae: only a relatively few species have been examined and even fewer considered in terms of the technology required for commercialisation. Biosorption by seaweeds has recently been reviewed by Volesky and Holan (1995), who state that ‘most of even the recent contributions studying the uptake of toxic metals by live marine and, to a lesser extent, freshwater algae focused on toxicological aspects, metal accumulation and pollution indicators’ and that focus on technological aspects of metal removal by algal biomass has been rare. However, Volesky and Holan do cite some 23 examples of algal biomass/metal adsorption. The interest in seaweeds lies in the combination of their efficient biosorption (Figures 8.1 and 8.2) and their cost. Perhaps surprisingly to those not involved, there is a considerable industry in seaweed collection, such as the highly priced edible seaweeds which are collected or farmed in various parts of the world. However, there is another industry collecting, drying and shipping large tonnages of non-edible

Figure 8.1 Cadmium removal from a 100 mg/l solution by a variety of brown seaweeds: Mp=Macrocystus pyrifera, Lf=Lessonia flavicans, Ln=Lessonia nigrescens, Em=Ecklonia maxima, Dp=Durvillea potatorum, Lh=Laminaria hyperborea, W=dealginated waste seaweed (Wilson, 1993)

166

Biosorption Using Unusual Biomasses

Figure 8.2 Percentage uptake of cadmium from a 10 mg/l solution by different types of biomass (Aderhold, 1995)

seaweeds for use as fertiliser and as raw material for several processes, the chief of which is alginate production. Alginate is used as a stabiliser in paint, food (as an edible thickening agent) and pharmaceutical products (where it is harmless on skin and stabilises foam) (McDowell, 1957). Usually, a broad mixture of brown seaweeds, particularly Macrocystis pyrifera, Lessonia flavicans, Lessonia nigrescens, Ecklonia maxima, Durvillea potatorum, Laminaria hyperborea and Ascophyllum nodosum is used, and these are collected from South America, South Africa, Australia, Ireland and other sites, dried, ground and shipped to the alginate producers. A number of researchers have shown that both seaweeds and alginates can accumulate metal ions (Holan and Volesky, 1994; Carvalho et al., 1994; Volesky, 1990; Jang et al., 1990; Apel and Torma, 1993). However, alginate is relatively expensive, particularly in any suitable (e.g. fibrous) form, and is therefore likely to be unsuitable for all but highly critical biosorption purposes, except perhaps for its use in combination with other biosorbents. Metals whose uptake by seaweeds have been studied include gold (Kuyucak and Volesky, 1988, 1989a), cobalt (Kuyucak and Volesky, 1989b), lead, cadmium, mercury, nickel, copper (Wilson, 1993), zinc, lead, cadmium, chromium, nickel and iron (Ramelow et al., 1992; Holan et al., 1993; Holan and Volesky, 1994; Volesky and Prasetyo, 1994), and in comparative tests seaweed biomass has been shown to outperform ion-exchange resins (Kuyucak and Volesky, 1988). 167


However, while readily available and capable of good biosorption, seaweeds can cause problems in process systems. Dried seaweeds swell on contact with water, and both dried and freshly collected seaweed can release large amounts of extracellular polymeric material. Both these factors can cause an unacceptable pressure drop through column adsorption systems and it is likely that other contact systems, such as stirred tank reactors, would have to be used, perhaps together with pretreatment with an acid or alkali wash to remove the extracellular material. The alginate producing industry extracts alginate by treating the seaweed with sulphuric acid followed by extraction with soda. The process leaves a considerable amount of waste material. This waste product or ‘dealginated seaweed’ is currently disposed of by tipping into the sea. Therefore, it would be very useful if this waste material were found to be a good biosorbent. The processing opens up the structure of the seaweed and a more open-structured material should be conducive to biosorption. However, the seaweed has had most of the alginate removed and it was thought that the alginate was solely responsible for the metal binding and that, on removal of the alginate, the biosorption capacity would be lost. More recent research has shown that alginate, while capable of metal binding, may not be the only, or indeed the main, component in seaweeds responsible for metal sorption (Wilson, 1993; Carvalho et al., 1994) and thus, the waste product may be suitable for biosorption. Tests carried out by Wilson (1993) and Aderhold (1995) show that waste seaweed with no alginate is often as good as or better than the parent seaweed material (Figures 8.1 and 8.2).

Bryophytes Bryophytes, particularly the mosses and, especially, species of Sphagnum, have also been studied (e.g. Coupal and Lalancette, 1976). Sphagnum moss, dried and ground and immobilised in polysulphane beads, has been shown to be an effective biosorbent for a variety of metals including cadmium, manganese and zinc (Trujillo et al., 1991). Sphagnum is also a component of peat, whose biosorption properties are described in Chapters 7 and 9.

Higher plants Plants have been used for many years as bioindicators of metals, due to their bioaccumulation properties. Certain species are known to be associated with, or tolerant of, certain metals and these are used as indicators of metal deposits in remote sensing programmes (Baker, 1989). In several studies a ‘rapid’ adsorption period (in the order of a few hours) is followed by a slower bioaccumulation period (Lee and Hardy, 1987; Jamil et al., 1985). Living plants have thus been suggested and used for the decontamination of both water and soils. Duckweed (Lemna minor) and the water hyacinth (Eichornia crassipes) have been shown to remove a variety of metals from water (Lenka et al., 1990; Jamil et al., 1985) while terrestrial plants, such as maize and a solonaceae (Datura inoxia), have been shown to adsorb calcium. Many higher plant biomass materials—the plant, parts of the plant or derivatives— have been shown to have good biosorbent properties. These include ground maize 168


cobs (Zea mays) (Okieimen et al., 1988), cypress leaves (Salim, 1988), jute (Shukla and Sakhardande, 1991) and bark of trees (Khangan et al., 1992; Shukla and Pandey, 1990), including Techtona grandis (Deshkar and Dara, 1988) and Hardwickia binata (Deshkar et al., 1990). Deshkar and Dara (1988) found that Techtona grandis bark could remove up to 55% of 100 mg/l mercury solution. True waste materials have also been investigated (see Chapter 7). These include sawdusts (Bhargava et al., 1987; Bryant et al., 1992; Chan et al, 1992), straw (Aderhold, 1995), wool, rice-straw, coconut husks and walnut shells (Ferro-Garcia et al., 1988) and apple waste (Maranon and Sastre, 1991). Waste tea (Tee and Khan, 1988), coffee, nut and walnut shells have been investigated by Orhan and Buyukgungor (1993) with a high removal of aluminium, chromium and cadmium when compared to that of activated carbon. Waste tea leaves have also been investigated by Tee and Khan (1988), rice-bran by Verma and Rehal (1994) and Suemitsu et al. (1986) and oil-palm fibres by Low et al. (1993). Adsorption capacities range from 15 to 34 mg/g and near 100% removal at lower contamination levels (5–50 mg/l), but the contact times can range from 10–15 minutes to 1–2 hours, which precludes some of these biomasses from all but the most basic and high volume contact mechanisms.

Chitin/chitosan Chitin occurs widely in lower animals, fungi and Crustacea. Exoskeletons of crabs, prawns and lobsters are particularly rich in chitin. The prawn processing industry in India produces some 60000 tonnes of waste containing about 5% chitin per year. The ‘pen’ of squid provides almost pure chitin (Ramachandran Nair and Madhavan, 1982). Fungal mycelia (e.g. Aspergillus niger) are also a source of chitosan, which is isolated as the insoluble fraction on boiling in 40% sodium hydroxide. Chitin is deacetylated to chitosan for use in foods, pharmaceuticals, flocculation, textiles and printing. Both chitin and chitosan have considerable potential for metal binding (Chapter 9). There is evidence that chitin/chitosan derived from different species of fungi behaves differently in metal uptake. Chitin derived from Streptomyces spp. was found to have good metal ion uptake for chromium, manganese, cobalt, nickel, copper, zinc and lead but not mercury, while that derived from Aspergillus niger was excellent at removing mercury (Muzzarelli and Tanfani, 1982). While ‘activation of chitosan’ by sulphuric acid and differences in pH have been reported to influence metal ion uptake, this has not been found by some authors in chitin (Ramachandran Nair and Madhavan, 1982). However, Muzzarelli et al. (1989) report considerable differences in uptake between pH 1 (zero uptake for many metal ions) and pH 5 (up to 90% uptake) and an equilibrium capacity of 125 mg lead/g. The differences between research findings and between different sources of chitin are probably due to the differing components and compositions of the isolated ‘chitin’. Unless very costly purification steps are taken, the crude isolations may contain a mixture of chitin and chitosan and other substances (e.g. glucans). The binding sites of chitin and chitosan for metal ions are believed to be free amine groups. However, the binding ability of the polymers for various metal ions is not directly proportional to the degree of free amine content. Rather, the binding is related to the conformation of the polymer and chelation depending on the pH, ionic 169


strength and metal ion to chitosan ratio (Park et al., 1986). This has meant that the polymer can be manipulated to improve or make selective the metal binding and various chitin derivitaves and modifications have been studied for their ability to adsorb metal ions. Chitin derivatives have been shown to remove low levels of metal contamination (0.1 mg/l) from brines, with around 100% removal for iron, cobalt, nickel and copper, and to remove over 90% of lead, chromium and cadmium from drinking water supplies (Muzzarelli et al., 1989). Ohga et al. (1987) describe the performance of cross-linked metal-complexed chitosan using copper, cadmium, nickel and iron as complexation ions and (chloromethyl)oxirane to achieve the cross-linking. All metals used in complexation studies resulted in an improved ability to remove copper from solution when compared to chitin in the absence of complexing metals. Cadmium-complexed chitosan was found to be effective in adsorbing mercury. Using a similar cross-linked copper-complexed chitosan, Inoue et al. (1988) showed a pH dependent selectivity, with gallium and indium being adsorbed at much lower pH values (pH 3) than zinc (pH 6). Thus, the selective adsorption of small amounts of gallium and indium in the presence of much larger amounts of zinc could be achieved. Similar results were found for iron and aluminium when compared to zinc. Inoue et al. (1988) cite commercial possibilities for this chitosan adsorbent in the recovery of gallium and indium from zinc refinery by-products or zinc lechate residue, the selective removal of iron and aluminium impurities in zinc plating baths, and highly effective separations of nickel from cobalt to produce high-purity cobalt.

Performance It is very difficult to demonstrate performance of adsorbent systems without going into great detail as to the conditions prevailing in the system used. It is even more difficult to give meaningful comparisons between different reports. Some authors give the following experimental conditions: • mass of sorbent • volume and concentration of sorbate • contact time and temperature. and report the removal percentage of the sorbate. Other authors report specific removal figures (mg sorbate/g sorbent) without adequate specific details about the experimental conditions. For example, the adsorption capacities of 15 to 34 mg/g reported for higher plants are not very high when compared to some seaweeds, which can reach 270 mg/g (Holan and Volesky, 1994). However, not all studies take the sorption to equilibrium; some are looking at percentage removal. Thus, Verma and Rehal (1994) give sorption rates of copper on ricebran as 9.44 mg/g. However, in the experimental system they used, this resulted in 94.3% removal of copper from a 100 ppm solution. Similarly, Volesky and Prasetyo (1994) report an average metal loading of 30 mg/g for cadmium in a packed-bed flow-through system, but the system was removing 99.985% of a 10 mg/l cadmium effluent. The differences in experimental technique and the reporting 170


of results reflect the diverse nature and background of research in this area. It is only now that engineering design is beginning to be applied to these systems in order to test whether they will actually work on an industrial scale. There is not the scope in this review to give experimental detail, and only some broad results can be shown. Examples are Kuyucak and Volesky (1989a), who cite up to 180 mg/g for gold adsorbed onto the seaweed Sargassum natans; Holan et al. (1993), who report over 100 mg/g cadmium for Ascophyllum nodosum; Holan and Volesky (1994), who found up to 270 mg/g lead for Fucus spp.; and Ramelow et al. (1992), who report greater than 80% removal of cadmium from a 5 ml solution by an undefined Sargassum species. Kaplan (1988) showed that taxonomically similar species of seaweed can have very different biosorption capacities. Some typical uptake data for cadmium removal from a range of brown seaweeds and dealginated seaweed waste (W) are shown in Figure 8.1. Using 0.1 g wet weight biomass in 50 ml metal solution, between 95 and 99% removal could be achieved in under 30 min contact time. Some data comparing the components of seaweed (whole, alginate only and waste with alginate removed) are shown in Figure 8.2, along with modified linseed straw as an alternative biomass comparison. While alginate gives almost 100% removal, whole seaweed and seaweed waste are very similar, with the seaweed waste giving a more consistent performance as indicated by the very small standard deviations. Much information is derived from ‘batch equilibrium sorption’ experiments which tend to be carried out in shaken or stirred flasks. There can be some anomalies in the results of such experiments, and the next stage is to determine uptake in a suitable process system. Commonest of these is a packed-bed flow-through column (again not without problems when using waste biomass). Examples of column systems are given by Volesky and Prasetyo (1994), who found nearly 100% removal of cadmium from a 10 mg/l solution with a mean loading of the Ascophyllum nodosum at the breakthrough time of 30 mg/g. Wilson (1993) studied a column of bed volume 0.5 litres of waste seaweed material which reduced a 10 mg/l solution of cadmium supplied at 75 ml/min to 0.05 mg/l for a period of 24 hours. Biosorption need not take place in fixed-bed reactors. Systems from large slowflow culverts to stirred tanks to small cartridge filters, as well as variations on fixedbed systems (fluidised bed, pulsed bed and compartmentalised systems) could all be used.

Factors affecting adsorption

Multi-metal systems Industrial effluents are unlikely to be homogeneous, either for metals or for other pollutants which may compete for adsorption sites. This is an area which is now being addressed, although most of the available literature is still concerned with single metal experimentation. Volesky and Holan (1995) cite some work under way with two-metal systems by Chong which shows marked differences between different pairs of metals. Using an Ascophyllum nodosum biosorbent, Chong shows an enhancement of uptake over equilibrium conditions for cadmium, for copper and 171


cadmium (over 200% the uptake of cadmium alone) and for zinc and cadmium (a maximum of 160% the uptake of cadmium alone). These results show that the total uptake of copper and cadmium was at the expense of the cadmium (40% cadmium, 180% copper, expressed as a percentage of the cadmium equilibrium), while the zinc/cadmium uptake was about 80% cadmium, 80% zinc (Volesky and Holan, 1995). Using the same metals but a different macroalgal biosorbent (dealginated seaweed waste), Aderhold et al. (1996) found no significant difference in the uptake of metal ions from a single metal and two- or three-metal combinations (Figure 8.3). However, in this case the metal concentrations were not high enough for equilibrium conditions for any or all of them. Thus, as may be expected, below equilibrium conditions there is little competitive effect between different metals but at equilibrium conditions for any one metal uptake appears to be enhanced, thus the total cumulative equilibrium metal uptake for two metals can be greater than a single-metal sorption equilibrium with the partition of available adsorption depending on the metal pairs. This has several implications. Firstly, in the examples of metal pairs described by Volesky and Holan (1995), while total metal uptake was enhanced, the cadmium was not above 100% of its equilibrium. This implies that the equilibrium for the other metal is much greater and/or that there are specific adsorption sites for each metal. As it is difficult to imagine metal ion specific sites, it may be that the binding capacity of each site varies and that, for a given number of sites on a biomass, a proportion, x, may have the binding capacity for cadmium, while 10x may have the capacity to bind a ‘more easily fixed’ metal such as copper. Thus, in the case of copper and cadmium, there would be plenty of sites on which the copper can bind but fewer with enough binding energy to hold cadmium, and there would be direct competition for these cadmium capable sites. It would be expected that, given equal amounts of copper and cadmium metal ions, about 50% of cadmium capable sites

Figure 8.3 The uptake by dealginated seaweed of copper (¢), cadmium (£) and nickel ( ) from single- and multi-metal ion solutions (from Aderhold et al. (1996); (Reprinted with kind permission from Elsevier Science Ltd, The Boulevard, Langford Lane, Kidlington, UK)

172


would end up occupied by cadmium. In the case of zinc and cadmium, the binding requirements may be much more similar, resulting in less total capacity and an equilibrium between the two metals. Considerable further work is required on multi-metal systems, particularly at equilibrium concentrations. However, equilibrium concentrations themselves are difficult to determine due to different effects of contacting methods, different effects of biomass/volume ratios and other factors. An interesting finding by Aderhold et al. (1996) is the effect of a semi-continuous batch adsorber system on selection between different metals. In this experiment, a biomass of dealginated seaweed waste was exposed to a fresh solution of 10 mg/l each of copper, cadmium and nickel in each adsorption cycle without the adsorbed metal being removed from the biomass between each cycle (i.e. no forced desorption). This means that a pseudo-equilibrium condition is reached after about the second cycle. The results (Figure 8.4) show that, after 11 cycles, copper is still being removed almost completely, but only 70% of the cadmium and 25% of the nickel is being removed. However, the cumulative effect is much greater than the equilibrium conditions for any of the metals. Such results could be important in industrial scale-up, particularly where the metals of most interest will be in much lower proportion to others.

pH It seems that for the majority of biosorbents the optimum pH is slightly acid to around neutral (4–7). However, some metals which form anionic complexes (e.g. some chromium, selenium, platinum and gold complexes) adsorb most strongly at

Figure 8.4 The selective adsorption of copper (¢), cadmium (F) and nickel () by dealginated seaweed over eleven successive adsorption cycles (from Aderhold et al. (1996); permission from Bioresource Technology sought)

173


very acid pH (1–2) (Kuyucak and Volesky, 1989a). The optimum pH of uptake of gold for the seaweeds Chondrus crispus, Palmaria palmata and Porphyra tenera was found to be 4, and was even lower (pH 2.5) for Codium taylorii and Halimada optunia (Kuyucak and Volesky, 1988), while Ramelow et al. (1992) found maximal uptake of Cu, Zn, Pb, Cd, Cr, Ni and Fe at pH 5–6 for the seaweeds Gracillaria conferta, Eisania bicyclis and Sargassum spp. It is thus important to investigate a range of pH values for any biomass with any metal, as considerable variation is found between biomasses and between metals. Desorption tends to be by treatment with acids at a low pH, but when the metal binding shows little pH dependence desorption can be accomplished by adding a competing ligand (e.g. mecaptoethanol) (Greene et al., 1987).

Temperature Temperature can have a significant effect on biosorption. Greene and Darnall (1988) found that the binding of metals to non-living biomass of the alga Chlorella pyrenoidosa increased by an order of magnitude between 0 and 60ºC. Aksu et al. (1992) reported optimal metal uptake at 25ºC for non-living biomass of both Chlorella vulgaris and the bacterium Zoogloea ramigera. In general, optimal temperatures mostly lie in the range 10–25ºC, with some interactions enhanced by higher temperatures. Any residual heat in industrial wastewaters would therefore not suppress biosorption and, if anything, might enhance it.

Time The speed of uptake is critical to the design and economics of any adsorption system. Leaving aside the bioaccumulation phase of metal entry into living cells, many authors have found times of between ten minutes and two hours for the biosorption process (usually quoted as 90% or 95% of maximal time dependent uptake or to equilibrium) (de Rome and Gadd, 1987). Aksu et al. (1992) report biosorption of copper to Chlorella vulgaris to be complete in 20–30 minutes, and similar or shorter times to equilibrium have been reported for seaweeds (Wilson, 1993), straw (Taylor, 1996), tree bark (Deshkar and Dara, 1988) and treated oil palm fibres (Low et al., 1993). However, others can take several hours (Deshkar et al., 1990; Bhargava et al., 1987). Alginate, produced as a fibre, has been found to remove 95%+ of its capacity in the first two minutes (Williams and Edyvean, 1996). Several authors have concluded that any commercially successful method of biological metal uptake would exploit this passive, rather than any active, uptake as it has been suggested that processes occurring in less than one hour will find much more ready acceptance in the engineering community than those requiring longer periods (Boutwood, 1990). Add to this the ease of using a non-living and, perhaps, processed biomass and the potential is encouraging.

174


Pretreatment and modification Modification of the biomass to improve adsorbent properties or physical performance has already been mentioned in the case of seaweeds, where pretreatment with acids or alkali opens more of the structure of the biomass and removes problem extracellular polymeric materials. In the case of seaweeds, this pretreatment has already been carried out by the alginate industry. Similar pretreatment can also be beneficial for other biomasses. Chemical pretreatments have been shown to be both beneficial and detrimental to the biosorptive properties (Ross and Townsley, 1986; Ramelow et al., 1992). Pretreatment with acid (nitric acid) and formaldehyde or formaldehyde in sulphuric acid has been shown to be beneficial to the biosorption properties of coniferous wood barks for the removal of uranium from seawater (Funjii et al., 1988) and to those of the bark of Techtona grandis (Deshkar and Dara, 1988). Ramelow et al. (1992) report that a prewash with hydrochloric acid increases cadmium binding to the seaweed Sargassum spp. but decreases binding to the seaweed Eisenia bicyclis. Several authors have reported that pretreatment with alkali increases the metal uptake of a range of biomasses (Rao et al., 1993; Brierley, 1990; Verma and Rehal, 1994; Salah Azab and Peterson, 1989; Low et al., 1993). Figure 8.5 shows the effect of a pretreatment with 40% NaOH solution at 100ºC for three hours on a linseed straw derivative. However, also shown in the figure are results for boiling the linseed straw in distilled water without NaOH, and the resulting uptake of cadmium, nickel and copper is even greater. There is an arbitrary dividing line between pretreatment and modification. Possibly a suitable division is one of cost and complexity: pretreating with an alkali is low-cost while the extraction and purification of chitin from fungal or crustacean biomass is relatively high cost. However, between the two are modified cellulose (Shukla and Sakhardande, 1991; Svoboda et al., 1992) and modified starch (Wang et al., 1991; Zhang et al., 1990). A considerable amount of work has been put into the modification of straw for various uses, including biosorption. With the introduction of a ban in the European Community on burning straw in open fields, there has been considerable

Figure 8.5 The removal of copper, nickel and cadmium from 10 ppm initial solution concentrations by untreated, boiled and sodium hydroxide treated linseed straw (Aderhold, 1995)

175


interest in alternative uses for straw or products derived from straw, including for biosorption (Aderhold, 1995). A recent development in this area is the patented process using linseed flax (and other) straws. Linseed flax fibre is extracted from the straw by decortication. This mechanical process follows the critical complex microbiological process of retting, which frees the fibre-containing ‘bast’ from the brittle stem of the straw. The subsequent separation of fibre is carried out using a patented system of open cage breaker rollers and a pinned rotor. This technique enables a high throughput of straw to be processed in a continuous operation. The process has been developed into a commercial scale operation by a consortium of industrial companies and academic institutions funded by the UK Ministry of Fisheries and Food in the ‘fibrelin’ project. The fibre produced is not the same as that extracted by traditional methods, and produces a biomass with different biosorbent properties (Taylor, 1996). The biomass can be processed into various loose or tight non-woven mats and other forms, giving strength and resistance to compaction. While the original straw is substantially modified, the process is relatively low cost, and with straw at around £100 per tonne, such a modification is feasible in comparison with other biosorbents. Such material is capable of removing between 80 and 90% of cadmium, copper and mercury from 100 mg/l solutions (Figures 8.2 and 8.3).

Competition Very little work has been carried out on competition for adsorbtion sites between metals and other components of a waste stream. Kuyucak and Volesky (1989b) showed that, while there is little effect on total gold uptake by Sargassum natans of other cations, some ‘contaminants’ such as PO4 and NO3 reduced uptake by 75% and 50% respectively. Some work has been carried out on coloured compounds, particularly dyes, and results have indicated that dye-treated biomass can have a greater metal adsorption capacity than untreated biomass (see Chapter 7).

Sorption/desorption cycles The critical cost factor for an adsorbent is often not its single exposure capacity but its life-time exposure capacity, i.e. the number of adsorption cycles the material can stand while remaining structurally sound and capable of adsorption. This presupposes that the metal adsorbed can be desorbed. Usually, with the biosorbents described in this chapter, desorption can be achieved with an acid wash, thus concentrating metals from a large volume of effluent into a small volume of acid wash. The process of desorption itself is important in the eventual disposal of the used biomass adsorbent. If the metal ions can be thoroughly removed the spent biomass will be much easier to dispose of than if it remains contaminated. Another factor is the extent to which the desorbed biomass retains adsorption ability. A shift in pH has been demonstrated to be an effective desorption mechanism by several authors (Greene et al., 1987; Cotoras et al., 1992; Wilson, 1993). Acids used include hydrochloric, sulphuric and nitric acid. However, some metals, such as gold, silver, and mercury, which showed less or no pH dependent adsorption, could not be removed by this method (Greene and Darnall, 1990) and other elutants are needed. 176


Kuyucak and Volesky (1990) found that a mixture of thiourea and formic ammon-ium sulphate was required to desorb gold from the seaweed Sargassum natans, and Nakajima et al. (1982) showed that sodium carbonate could be used to desorb uranium from the alga Chlorella regularis. Greene and Darnall (1990) found that anionic metal complexes (as formed by selenium, molybdenum and chromium), which had the best adsorption at low pH values, were best desorbed by adjusting the pH to higher values and that selective desorption could be achieved by using a ‘pH gradient’ and isolating the wash products from different pH ranges. The percentage of the adsorbed metal recovered is usually high once the right desorption system is found. Wilson (1993) reports values of 90–90% for copper and cadmium on the seaweed Ecklonia maxima but little desorption of mercury when using three different acids (sulphuric, nitric and hydrochloric). Some authors have reported that adsorption/desorption cycles are possible. Zimnik and Sneddon (1988), working with the alga Chlorella vulgaris and using nitric acid to desorb aluminium, report adsorption/desorption being possible across twelve cycles, the first seven having little effect on the metal loading capability of the biomass. Bangert (1996) reports adsorption/desorption experiments over eleven cycles with lead remaining consistently high, and cadmium and mercury falling initially and then remaining at a consistent level. It appears that in many cases, while there might be a slight drop in the ability of the biomass to adsorb metal ions after the first two or three cycles, the main limit to adsorption/desorption cycling is likely to be the structural integrity of the biomass, eventually breaking down and becoming unusable.

Mixed biomasses There are differences in uptake capacity and in metal selectivity of different biomasses, which make the use of mixed biomasses interesting. However, there is very little reported work in this area and such studies may have to wait until development of systems for actual industrial effluents.

Industrial scale systems Biosorption may have a potential marketing advantage over other treatment techniques, in that it can be perceived as environmentally friendly and, particularly for the use of waste biomass, a natural and maximal use of gathered material. Clearly, much of the development work is not available in the open literature and some reports of ‘commercial’ systems turn out to have only reached laboratory and cost analysis stages. However, several biosorption systems have been commercialised, with varying degrees of success. Attempts to commercialise bacterial systems include the AMT-Bioclaim (Brierley, 1990), where biomass of Bacillus subtilis was obtained as a by-product of a fermentation process, heat and alkali treated to improve binding capacity for silver and copper, immobilised using a cross-linking agent and then extracted, dried and sieved. The material was recommended for use in either packed beds or fluidised pulsed beds (for larger flows) and was claimed to be capable of removing 99% of the metal in a solution (see Chapter 3). 177


Cost is a major factor of biosorbent systems: to be commercially viable the raw material must be considerably lower in cost than ion-exchange resins, activated carbon and other systems. Bulk production costs of specifically cultured algae and fungi are in the order of £2000–£10000 per tonne, depending on the species, costs of immobilisation, etc. The bulk cost of commercially supplied, dried seaweed is in the order of £200–£300 per tonne; waste biomasses, such as seaweed waste left after alginate removal and agricultural wastes such as linseed straw, even after processing, will be considerable cheaper than this. Biosorbents may, perhaps, find their greatest utility in treating metal solutions of concentration 1–10 mg/l, due to high removal capabilities at low initial metal ion concentrations. Their use as a polishing or secondary stage to other removal systems would be in competition with ion-exchange resins, where material costs will be very much in the biosorbents’ favour. The competitiveness will, therefore, be based on using factors such as consistency, reliability and re-usability in the final cost-benefit analysis. While large scale systems have been proposed and developed, e.g. sphagnum based ‘Bio-fix’ beads in a culvert system in which the biosorbent beads are contained in sacks (Jeffers, Bennett and Corwin, 1993) (see also Chapter 9), the greatest use for biosorption systems may be in modular systems for small companies.

Conclusions The investigation of biosorption, as defined by the passive or physico-chemical uptake of a sorbate by non-living biomass, has been slow to develop. This is partly for historical reasons. The initial investigations were rooted in toxicological and ecological studies of, for example, metal ion uptake and a knowledge of the workings of wastewater treatment works. This led to investigations of the use of living organisms to remove pollutants from wastewaters. While such investigations showed that the removal efficiencies could be as good as, or better than, conventional systems such as ion-exchange resins, living systems are costly to produce and to maintain and suffer from problems associated with shock loading. It was not until the recent investigations into the use of dead biomass and, more specifically, biomass wastes from other industries, that the economics and reliability of such systems made them potentially much more attractive. The use of non-living systems allows the material to be treated like any other adsorbent and thus considerably decreases the cost of running the system. However, this still leaves the cost of producing the biomass. Investigations have shown that biomass wastes from other industries, such as straw, waste seaweed and bagasee pith have cut the cost of the adsorbent ten-fold, at least. A great deal of basic research has been carried out and has recently been published or will be published in the next couple of years and, while there are engineering design problems to be overcome, the scene is now set for industrial trials. A very interesting area of the research is the findings of the effects of pretreatments to enhance the biosorption process and, along with pH control and the variation of other parameters, to engender specificity. This, together with the mixed ability of biosorbants to remove metals and compounds, such as dyes (Low et al., 1993; Doherty, 1996) leads to the possibility of developing highly tailored systems for specific waste streams where, by careful analysis of the waste, coupled with 178


matching the biosorbent (and other treatments), both sequential removal and specific recoveries could take place within a multistep system. The understanding and detailed studies of uptake capacities, desorption and biosorption behaviour have now reached a point where specific treatment plants can and should be developed.

References ADERHOLD, D., 1995, Adsorbers/Biosorbers for the Removal of Pollutant Metals from Industrial Aqueous Waste Streams, Diplomarbeit, Universitat Dortmund, March 1995. ADERHOLD, D., WILLIAMS, C.J. and EDYVEAN, R.G. J., 1996, The removal of heavy metal ions by seaweeds and their derivatives, Bioresource Technology, 58, 1–6. AKSU, Z.,SAG,Y. and KUTSAL, T., 1992, The biosorption of copper(II) by C. vulgar is and Z. ramigera, Environmental Technology, 13, 579–586. APEL, M.L. and TORMA, A.E., 1993, Determination of kinetics and diffusion coefficients of metal sorption on Ca-alginate beads, Canadian Journal of Chemical Engineering, 71, 652–656. BAKER, A.J.M., 1989, Terrestrial higher plants which hyperaccumulate metallic elements. A review of their distribution, ecology and phytochemistry, Biorecovery, 1, 81–126. BANGERT, N., 1996, The Ability of Waste Biomass to Remove Heavy Metal Ions from Dilute Aqueous Effluents, Diplomarbeit, Ruhr-Universität, Bochum, Germany and The University of Sheffield, UK. BHARGAVA, D.S., GUPTA, M.S. and VARSHNEY, B.S., 1987, Use of sawdust for the adsorption of heavy metals, Asian Environment, 9, IV, 29–37. BOUTWOOD, J., 1990, Personal Communication. BRIERLEY, J.A., 1990, Production and application of a Bacillus-based product for use in metals biosorption, In: Biosorption, VOLESKY, B. (Ed.), CRC Press, Boca Raton, Florida, pp 305–311. BRYANT, P.S., PETERSEN, J.N., LEE, J.M. and BROUNS, T.M., 1992, Sorption of heavy metals by untreated red fir dust, Applied Biochemistry and Biotechnology, 34–35, 778– 788. CARVALHO, R.P.DE, CHONG, K.H. and VOLESKY, B., 1994, Effects of leached alginate on metal biosorption. Biotechnology Letters, 16, 875–880. CHAN, W.H., LAM-LEUNG, S.Y., CHENG, H.W. and YIP, Y.C., 1992, Synthesis and characterisation of imminodiacetic-cellulose sorbent and its analytical and environmental applications in metal ion extraction, Analytical Letters, 25, 305–320. COTORAS, D., VIEDMA, P., CIFUENTES, L. and MESTRE, A., 1992, Sorption of metal ions by whole cells of Bacillus and Micrococcus, Environmental Technology, 13, 551–559. COUPAL, B. and LALANCETTE, J.M., 1976, The treatment of waste waters with peat moss, Water Research, 10, 1071–1076. DESHKAR, A.M., BOKADE, S.S. and DARA, S.S., 1990, Modified Hardwickia binata bark for adsorption of mercury (II) from water, Water Research, 24, 1011–1016. DESHKAR, A.M. and DARA, S.D., 1988, Sorption of mercury by Techtona grandis bark, Asian Environment, 3–11. DOHERTY, S.M., 1996, PhD Thesis, The University of Leeds, UK. FERRO-GARCIA, M.A., RIVERO-UTRILLA, J., RODRIGUEZ-GORDILLO, J. and BAUTISTA-TOLEDO, I., 1988, Adsorption of zinc, cadmium, and copper on activated carbons obtained from agricultural by-product, Carbon, 26, 363–373. FUNJII, M., SHIOYA, S.I. and ITO, A., 1988, Chemically modified coniferous wood barks as scavengers of uranium from sea water, Holsforschung, 42, 295–298. GADD, G.M., 1990, Biosorption, Chemistry and Industry, 13, 421–426. 179

Biosorbents for Metal lons GADD, G.M., 1992, Microbial control of heavy metal pollution, In: Microbial Control of Pollution, FRY, J.C., GADD, G.M., JONES, C.W. and WATSON-CRAIK, I. (Eds), Cambridge University Press, Cambridge, pp. 59–88. GREENE, B. and DARNALL, D.W., 1988, Temperature dependence of metal ion sorption by Spirulina, Biorecovery, 1, 27–41. GREENE, B. and DARNALL, D.W., 1990, Microbial oxygenie photoautrotrophs (Cyanobacteria and algae) for metal-ion binding , In: Microbial Mineral Recovery, EHRLICH, H.L. and BRIERLEY, C.L. (Eds), McGraw-Hill, New York. GREENE, B., M CPHEARSON, R. and DARNALL, D., 1987, Algal sorbents for selective metal ion recovery, In: Metals: Speciation, Separation and Recovery, PATTERSON, J.W. and PASSINO, R. (Eds), Lewis Publishers Inc., Chelsea, Michigan. HOLAN, Z.R. and VOLESKY, B., 1994, Biosorption of lead and nickel by biomass of marine algae, Biotechnology and Bioengineering, 43, 1001–1009. HOLAN, Z.R., VOLESKY, B. and PRASETYO, I., 1993, Biosorption of cadmium by biomass of marine algae, Biotechnology and Bioengineering, 41, 819–825. HUNT, S., 1986, Diversity of biopolymer structure and its potential for ion binding applications, In: Immobilization of Ions by Biosorption, ECCLES, H. and HUNT, S. (Eds), Ellis Horwood, Chichester. INOUE, K., BABA, Y., YOSHIZUKA, K., NOGUCHI, H. and YOSHIZAKI, M., 1988, Selective series in the adsorption of metal ions on a resin prepared by crosslinking copper(II)-complexed chitosan, Chemistry Letters (Japan). The Chemical Society of Japan, pp. 1281–1284. JAMIL, K., JAMIL, M.X., RAO, P.V. R. and THYAGARAJAN, G., 1985, The role of water hyacinth in abating aquatic pollution, Pollution Research, 4, 67–75. JANG, L.K., BRAND, W., RESONG, M., MAINIERI, W. and GEESEY, G.G., 1990, Feasibility of using alginate to adsorb dissolved copper from aqueous media, Environmental Progress. 9, 269–274. JEFFERS, T.H., BENNETT, P.C. and CORWIN, R.R., Biosorption of metal contaminants using immobilised biomass-field studies. Report of Investigations—US Department of the Interior, Bureau of Mines, vol. 9461. KAPLAN, D., 1988, Binding of heavy metals by algal polysaccharides, In: Algal Biotechnology, STADLER, T. et al. (Eds), Elsevier Applied Science, London, pp. 122– 136. KHANGAN, V.W., BANKER, D.B. and DARA, S.S., 1992, Effectiveness of Terminalia bellirica bark for scavenging zinc ions, Chem. Environmental Research, 1, 87–94. KUYUCAK, N. and VOLESKY, B., 1988, Biosorbents for recovery of metals from industrial solutions, Biotechnology Letters, 10, 137–142. KUYUCAK, N. and VOLESKY, B., 1989a, Accumulation of gold by algal biosorbent, Biorecovery, 1, 189–204. KUYUCAK, N. and VOLESKY, B., 1989b, Accumulation of cobalt by marine alga, Biotechnology and Bioengineering, 33, 809–814. KUYUCAK, N. and VOLESKY, B., 1990, Biosorption by algal biomass, In: Biosorption, VOLESKY, B. (Ed.), CRC Press, Boca Raton, Florida, pp. 173–198. LEE, T.A. and HARDY, J.K., 1987, Copper uptake by the water hyacinth, Journal of Environmental Science, A22,141–160. LENKA, M., PANDA, K.K. and PANDA, B.B., 1990, Studies on the ability of the water hyacinth (Eichhornia crassipes) to bioconcentrate and biomonitor aquatic mercury, Environmental Pollution, 66, 89–99. Low, K.S., LEE, C.K. and LEE, P.K., 1993, Sorption of copper by dye-treated oil-palm fibres , Bioresource Technology, 44, 109–112. MARANON, E. and SASTRE, H., 1991, Heavy metal removal in packed beds using apple wastes, Bioresource Technology, 38, 39–44. MCDOWELL, R.H., 1957, Properties of alginates, Alginate Industries Ltd, London. 180

Biosorption Using Unusual Biomasses MUZZARELLI, R.A.A. and TANFANI, F., 1982, The chelating ability of chitinous materials from Aspergillus niger, Streptomyces, Mucor rouxii, Phycomyces blakesleeanus, and Choanephora cucurbitarum, Proceedings 2nd International Conference on Chitin and Chitosan, Ancona, Italy, 1982, pp. 183–186. MUZZARELLI, R.A.A., WECKX, M. and FILIPPINI, O., 1989, Removal of metal ions from industrial waters, nuclear effluents and drinking water with the aid of cross-linked Ncarboxymethyl chitosan, Carbohydrate Polymers, 11, 293–306. NAKAJIMA, A., HORIKOSHI, T. and SAKAGUCHI, T., 1982, Recovery of uranium by immobilised microorganisms, European Journal of Microbiology and Biotechnology, 16, 88–91. OHGA, K., KURAUCHI, Y. and YANASE, H., 1987, Adsorption of Cu2+ or Hg2+ ion on resins prepared by crosslinking metal-complexed chitosans , Bulletin of the Chemical Society of Japan, 60, 444–446. OKIEIMEN, F.E., MAYA, A.O. and ORIAKHI, C.O., 1988, Sorption and cadmium, lead and zinc ions on sulphur-containing chemically modified cellosic materials, International Journal of Environmental and Analytical Chemistry, 32, 23–27. ORHAN, Y. and BUYUKGUNGOR, H., 1993, The removal of heavy metals by using agricultural wastes, Journal of Water Science and Technology, 28, 247–255. PARK, J.W., PARK, D.M. and PARK, K.K., 1986, Characterisation and metal ion binding properties of carboxymethylchitosan, Polymer (Korea), 10, 641–645. RAMACHANDRAN NAIR, K. and MADHAVAN, P., 1982, Metal binding property of chitosan from different sources. Proceedings of the 2nd International Conference on Chitin and Chitosan, Ancona, Italy, 1982, p. 187. RAMELOW, G.J., FRALICK, D. and ZHAO, Y., 1992, Factors affecting the uptake of aqueous metal ions by dried seaweed biomass, Microbios, 72, 81–93. RAO, C.R.N., IYENGAR, L. and VENKOBACHAR, C., 1993, Sorption of copper(II) from aqueous phase by waste biomass, Journal of Environmental Engineering, 119, 369–377. ROME, L.DE and GADD, G.M., 1987, Copper adsorption by Rhizopus arrhizus, Cladosporium resinae and Penicillium italicum., Applied Microbial Biotechnology, 26, 84–90. Ross, I.S. and TOWNSLEY, C.C., 1986, The uptake of heavy metals by filamentous fungi, In: Immobilisation of Ions by Bio-sorption, ECCLES, H. and HUNT, S. (Eds), IRL Press, Chichester, pp. 49–58. SALAH AZAB, M. and PETERSON, P.J., 1989, The removal of cadmium from water by the use of biological sorbents, Journal of Water Science and Technology, 21, 1705–1706. SALIM, R., 1988, Removal of nickel from polluted water using decaying cypress leaves, Journal of Environmental Science and Health, A23, 321–334. SHUKLA, N. and PANDEY, G.S., 1990, Charred waste of oxalic acid plant as an adsorbent of toxic ions and dyes, Biological Wastes, 32, 145–148. SHUKLA, S.R. and SAKHARDANDE, V.D., 1991, Novel method of using reactive dyes for effluent treatment, American Dyestuff Report, July, 38–42. SUEMITSU, R., UENISHI, R., AKASHI, I. and NAKANO, M., 1986, The use of dyestufftreated rice hulls for removal of heavy metals from waste water, Journal of Applied Polymer Science, 1, 75–83. SVOBODA, L., UHLIR, J. and UHLIR, Z., 1992, Sorption of lead by the selective ion exchange Ostsorb DETA, Collect. Czech Chem. Commun., 57, 1393–1404. TAYLOR, D.T., 1996, Personal Communication, Department of Textile Industries, The University of Leeds, Leeds. TEE, T.W. and KHAN, R.M., 1988, Removal of lead, cadmium and zinc by waste tea leaves, Environmental Technology Letters, 9, 1223–1232. TRUJILLO, E.M., JEFFERS, T.H., FERGUSON, J. and QUINN-STEVENSON, H., 1991, Mathematically modelling the removal of heavy metals from a wastewater using immobilized biomass, Environmental Science and Technology, 25, 1559–1564. 181

Biosorbents for Metal lons VERMA, N. and REHAL, R., 1994, Bioscavenging of Cu(II) ions from aqueous solution with ricebran, Bioresource Technology, 49, 277–278. VOLESKY, B., 1990, Biosorption and Biosorbents, In: Biosorption, VOLESKY, B. (Ed.), CRC Press, Boca Raton, Florida, pp. 3–5. VOLESKY, B and HOLAN, Z.R., 1995, Biosorption of heavy metals, Biotechnology Progress, 11, 235–250. VOLESKY, B. and PRASETYO, I., 1994, Cadmium removal in a biosorption column, Biotechnology and Bioengineering, 43, 1010–1015. WANG, Y., HAN, Q., HUANG, Z. and TANG, Y., 1991, Removal of heavy metal ions from water by water-insoluble carboxyl-containing grafted starch polymers, Water Treatment, 6, 339–342. WILLIAMS, C.J. and EDYVEAN R.G.J., 1997, Optimization of metal adsorption by seaweeds and seaweed derivatives. Trans. I. Chem. E., 75(B), 19–26. WILSON, M.W., 1993, Biosorption for the Treatment of Industrial Wastewaters, PhD. Thesis, The University of Leeds, Leeds. ZHANG, L., Hou, W., ZHANG, L. and ZHANG, B., 1990, Removal of dissolved heavy metals from industrial wastewater with improved insoluble starch xanthate, Water Treatment, 5, 87–94. ZIMNIK, P.R. and SNEDDON, J., 1988, Binding and removal of aluminium ions in waters by an algal biomass, Analytical Letters, 21, 1383–1396.

182

9

Low-cost Adsorbents in Continuous Processes G.M C KAY and S.J.ALLEN

Introduction

Sorption processes There are several methods available for contacting sorbent materials with solutions containing metal ions. These are batch contacting processes and column contact methods. This chapter deals mainly with the removal of metal ions in sorption columns although, if applicable, the literature survey includes batch contact systems. Specifically, it presents a summary of the relevant research on the applications of a range of sorbent materials, particularly peat and lignite, to the removal of dissolved metal contaminants from wastewaters in which the use of sorption beds would be an effective means of bringing about liquid-solid contact. It should be noted that there are inherent difficulties in comparing the metal ion investigations carried out on peat, and to a lesser extent lignite. The main problems are the heterogeneity of peat, variable chemical and physical properties of different peat and lignite types and the variations in pretreatment and sampling procedures. Similar research work can be carried out in a number of ways, which causes great diversity in the results obtained by different researchers. Thus, the accuracy of comparisons of the results given by several authors working on similar experimental programmes has been difficult and sometimes impossible to assess. There are several mechanisms by which metal ions may be removed from a solution and attached to a sorbent particle surface. This is due largely to the complex chemistry of the available surfaces of many of the less traditional sorbents covered in this book. The mechanism may be due to ion exchange, physical sorption, chemisorption, chemical reaction, lone pair electron sharing or donating plus a number of other mechanistic processes. For the purposes of this chapter, the terms sorbent and sorption are used in a generic manner to cover all possible mechanisms or combinations of mechanisms resulting in the removal of metal ions from aqueous solutions onto sorbent surfaces. Numerous industrial processes produce aqueous effluents containing heavy metal contaminants (Table 9.1). These pollutant concentrations must be reduced to meet ever-increasing legislative standards (see Chapter 1), particularly from European 183

Table 9.1 Heavy metals found in major industries (from Dean et al., 1987)

All values given as mg/l. A=annual average, T=total, M=maximum allowable concentration, D=dissolved, P=95% of samples (b).

Table 9.2 EU water quality standards


Union Directives (Table 9.2), or, where feasible, recovered. Consequently, improved and innovative methods of wastewater treatment are continuously being developed to deal with waste metals not readily compatible with traditional methods, or economically unfeasible to remove by these methods. Activated carbon has traditionally been used for wastewater treatment, but the use of other carbonaceous materials including the incorporation of peat and lignite in wastewater treatment has received increasing attention over the past 20 years, and these materials currently offer a very attractive method of pollution remediation. Besides being plentiful, inexpensive and locally available, peat and lignite possess several characteristics which make them effective media for the removal of dissolved metal pollutants from wastewater.

Current methods of removal The most common methods of removal of metals from industrial effluents include chemical precipitation, solvent extraction, dialysis or electrodialysis, electrolytic extraction, cementation, reverse osmosis, evaporative methods, ion-exchange resins, carbon sorption and dilution. In recent years carbon sorbents have generally become the accepted medium for physiochemical treatment of wastewater, while the most common metal treatment processes have been chemical precipitation and electrolytic methods (McLellan and Rock, 1987). Conventional physical and chemical treatment of low concentration, large volume wastes tends to be costly. Consumptive processes, such as chemical precipitation, entail large capital and operating costs. Attention has thus focused on non-consumptive methods which include ion exchange and other sorption processes. Even then, commercially available resins and activated carbons can be costly. Among the naturally occurring sorbent materials, peat and lignite have been found to compare favourably in capacity to synthetic resins and activated carbons (Brown et al., 1992).

Carbon-based sorbents Carbon is the oldest adsorbent used in water treatment. There are references to its use as far back as 1500 BC, and it has been suggested that the tree which Moses cast into the ‘bitter’ water at Marah to make it ‘sweet’ for the Israelites to drink (Exodus 15:23–25) functioned as an ion exchanger (McKeown, 1987). In the 19th century the highly polluted water from the River Thames was frequently treated through charcoal before being used for drinking purposes, and the same process was applied in many American cities (Arden, 1966). Although carbons can be classified in many different ways, the first and most important differentiation is between bone charcoal, manufactured by the high temperature carbonisation of bones in the absence of air, and all the other types of carbon which are of vegetable origin. For water treatment vegetable carbons are now used almost exclusively. The second most important classification is the age of the carbon: these values are given in Table 9.3. The most widely investigated sorbent is activated carbon. The conversion of cellulosic matter to activated carbon is carried out either by chemical activation, in 186

Low-cost Adsorbents in Continuous Processes Table 9.3 Classification of peat and lignite by age (courtesy of Nl geological survey)

which the structure is dehydrated and broken down using sulphuric or phosphoric acid at 200–650ºC, or by the high-temperature oxidation of a previously charred vegetable material at around 900ºC in an atmosphere of steam or flue gas. Both these processes introduce a porous structure into a starting material of relatively low surface area. Finally, different physical forms of the activated carbon are produced depending on the desired application: granular forms are used in sorption columns and powder forms are used for batch sorption followed by filtration (Arden, 1966). There are drawbacks, however, to activated carbon usage. As Chermisinoff and Ellerbusch (1978) noted, the large surface of activated carbon (500–1400 m2/g) is, for the most part, non-polar. This makes carbon a good sorbent for organics and, although some carbons have been found effective, sorption of inorganic electrolytes is more difficult (Netzer and Hughes, 1984). The idea of using low-cost carbons and agricultural products and by-products for the removal of toxic metals from wastewater has been investigated by a number of authors. Studies to assess the ability of scrap rubber to adsorb dissolved metal ions from wastewater found it to be a moderately effective adsorbent (Henderson et al., 1977). Comparatively, however, it was generally found not to be as effective as peanut husks or composted bark in removing all types of heavy metals from solution (Henderson et al., 1977). Netzer et al. (1974) investigated the possible use of sand filtration for removing trace metals from wastewater. They concluded that the minimal sorption achieved made the process unattractive from economical and technical points of view. Wood fines were found not to be very effective, as only copper and nickel were retained in appreciable amounts (Henderson et al., 1977). Corn-cob fragments, however, even though relatively large in size, removed a sizeable percentage of copper and cadmium from low concentration solutions. From a study of barley straw, Larsen and Schierup (1981) drew the comparison that 1 g of straw was able to sorb amounts of Zn, Cu, Pb, Ni and Cd ranging from 4.3 to 15.2 mg, and 1 g of pine sawdust 1.3 to 5.0 mg, whereas 1 g of activated carbon removed 6.2 to 19.5 mg. Bhattacharya (1983) found both bituminous coal and crushed coconut shells to be effective in removing cadmium from water. Even human hair has been investigated for its metal trapping potential by Krishnan et al. (1987), who concluded that it was effective in removing arsenic, mercury and cadmium.

Structure and properties of peat and lignite Peats and lignites are composed of both organic and inorganic components. The organic part comprises the carbonaceous matrix and the mineral (inorganic) matter is non-uniformly distributed throughout (Lowry, 1945). In proceeding down the series the carbon content increases at the expense of oxygen. 187


Peat structure and properties Peat is enormously variable in composition. It varies from material so slightly decomposed that plants can be readily identified from bits of leaves and roots and stems, to soil so highly decomposed that it seems to be a structureless thick mud (Kadlec and Keoleian, 1986; Kadlec and Rathburn, 1983). Consequently, peat is a rather complex material containing lignin and cellulose as major constituents (Souci, 1938). These constituents, especially lignin, bear polar functional groups such as alcohols, aldehydes, ketones, carboxylic acids, phenolic hydroxides and ethers which can be involved in chemical bonding (Alder and Lindquist, 1963). Because of the very polar character of this material, the specific sorption potential for dissolved solids, such as metals and polar organic molecules, is quite high. These properties have consequently lead to the examination of the potential of peat as an agent for the purification of wastewaters contaminated with dissolved metals (Lalancette, 1974).

Lignite structure and properties Macroscopically, lignite is dark brown or black in colour when moist, turning light brown when dry. The mechanical strength of lignite is generally poor; consolidation increases as coalification proceeds. Its density lies in the range 1.0–1.35 g/cm 3. Lignites are amorphous and fibrous or woody in texture, the ‘as mined’ material possessing numerous water-filled pores and capillaries. This feature means that the lignites have high moisture contents (30–70%). Pore diameters vary and include micropores (Pope, 1984). As a consequence of their highly porous nature lignites possess high surface areas, typically 100–200 m2/g (Gan et al., 1972). As with peats, the organic constituents consist of complex polymeric materials formed during the coalification process. Geologically young coals, such as lignites, often contain plant materials which have undergone little change. For example, substantial proportions of cellulose are present because of its high resistance to fungal attack (Wilson et al., 1987). Lignites are, consequently, less reduced than more mature coals and, therefore, contain more. The polymeric matrix is composed of polycyclic aromatic ring systems joined by straight and branched chain aliphatic groups which retain varying numbers of functional groups, depending on origin and mode of formation. Typically, there are carboxylic, carbonyl, phenolic and hydroxyl groups and ether linkages. These functionalities impart enhanced reactivity to the lignites and make them suitable to bond with polar organic and inorganic materials.

Peat, lignite and chitosan as sorbents for metal ions

Applications The application of peat and lignite to pollution control in water treatment systems has received increasing attention over the past 20 years, and currently offers a very attractive method of wastewater treatment. Besides being plentiful, inexpensive and locally available, both materials possess several characteristics which make them effective media for the removal of dissolved metal pollutants. 188

Low-cost Adsorbents in Continuous Processes

Peat is reported to have excellent ion-exchange properties similar to those of natural zeolites and green sands (Belkevich et al., 1976). The natural capacity of peat for heavy metal retention has been recognised in studies which have investigated metal contaminants in peat bogs. Pakarinen et al. (1980) studied the vertical distribution of trace heavy metals in sphagnum peat of southern Finnish ombrotropic bogs. They found selective sorption of Pb>Cu>Zn, Mn in the surface of the bogs. Eger et al. (1980) studied the drainage of mining stockpile leachate through a white cedar bog. They found that most of the flow occurred across the surface and in the top 30 cm of the bog. This flow pattern resulted in 30 to 100% removal of trace metals from solution, >30% removal of Ni, and >99% removal of Cu. Glooschenko and Capobianco (1982) found low and consistent concentrations of Zn, Pb, Cr, Cu and Hg in several peatland ecosystems located in Ontario, Canada. In addition to studying the natural constituents of peat bogs, considerable attention has been focused on the potential of peat as a commercial adsorbent for the removal of toxic metals in contaminated water. As far back as 1939, Nikol’skii and Paramonova realised the potential of peat as an ion-exchange medium for metals such as copper, zinc, lead and mercury. Using laboratory columns, Schwartz (1968) found that peat was capable of removing Ca and Mg ions from wastewater. He concluded that cation removal was due to a weak acid ion-exchange mechanism. In batch and column studies Lalancette and Coupal (1972) found peat to be an efficient means of removing Hg from water. They concluded, therefore, that peat would also be effective for removing all metals having sulphides of low solubility, including Cd, Zn, Pb, and Cu. Later, Coupal and Lalancette (1976) found that Hg, Cd, Zn, Cu, Fe, Ni, Cr(VI), Cr(III), Ag, Pb and Sb can be treated efficiently by contacting wastewater with peat moss following a settling process. They found sorption to be quite high due to the polar character of peat. They concluded that the main advantages of utilising peat for wastewater treatment are its broad scope in terms of pollutants eliminated and its ability to accept rather wide variations of effluent composition. Ruel et al. (1973) reviewed the ability of peat to treat polluted water. They found that peat can compete with other sorbents to remove metals, provided low flow rates are maintained and the quantity of wastewaters to be treated is small. Leslie (1974) studied the treatment of dyehouse effluent using a Hussong-Couplan wastewater treatment system, which is based on peat moss. He attained heavy metal removals of >99% for Cd, Cr (III and VI), Cu, Fe, Pb, and Ni and 98.9% removal for Zn. In a later study, Zhipei et al. (1984) investigated the removal of Pb, Cd, Zn, Ni, and Cr from wastewater. They concluded that, in treating water containing heavy metal ions with peat, the best results were obtained when the ion concentrations in wastewater were low. Parkash and Brown (1976), in their investigation of the sorption of zirconium and titanium from aqueous solutions by peat, found that peat accommodated almost four times more zirconium than titanium. The data obtained by Smith et al. (1976, 1977) showed that peat treated with sulphuric acid or a sulphuric/ phosphoric acid blend possessed enhanced cation exchange capacities.

Loading Coupal and Lalancette (1976) concluded that peat can adsorb most metals in a very efficient way, up to 4% of the weight of dry peat—that is, for solutions of metals less 189


than 10 ppm, 0.68 kg of peat can purify 0.95 m3 of wastewater. It was also noted that the metal removal efficiency in unbuffered solutions is significant in a very large concentration range from 0.01 to 100 mmol, with maximum extraction ratios being obtained in the 0.1–1.0 mmol range. Ong and Swanson (1966) found that the maximum copper concentration in peat and lignite by sorption ranges between 0.6 and 2.2%; thus copper concentrations exceeding 2.2% include processes other than sorption. In a study of various classifications of peat Gosset et al. (1986) revealed that the maximum binding capacities in 10 mmol/1 metal cation solutions were very similar regardless of metal or type of peat used: all values fall in the range 180–200 mmol/1 dry weight.

Mechanism of metal sorption The nature of metal binding in soil organic matter, including peat and lignite, has been extensively investigated but a common viewpoint on the exact mechanism of sorption has yet to be reached. Additionally, the comparison of results is very difficult because pretreatment methods vary among authors, and the type of peat (and to lesser extent, lignite) investigated will have a significant effect on the ion-exchange capacity (Tummavuori and Aho, 1980); this view contradicts the statement by Gosset et al. (1986). Clymo (1963) proposed that there is a good correlation between the content of unesterified polyuronic acids in the cell wall of sphagnum peat and the cation exchange capacity. De Mumbrum and Jackson (1956) proposed that the sorption of copper and zinc ions occurs by the formation of complexes with the carbonyl and nitrile groups in peat. Kashirtseva (1960) proposed that the presence of humic acids in peat was primarily responsible for its ability to sorb metals. Furthermore, many workers have implicated carboxylic acid (COOH) groups in the reaction of divalent metals with humic acids (Schnitzer, 1978; Schnitzer and Khan, 1972; Van Dijk, 1971; Vinkler et al., 1976; Boyd et al., 1981). They support the general view that the reaction of metal ions, such as Cu and Fe, with humic acids is one of chelate ring formation involving adjacent aromatic carboxylate COOH and phenolic OH groups or, less predominantly, two adjacent COOH groups which participate in ionexchange reactions by binding metal ions with the release of H + ions. Others believe that there is no direct evidence for chelation: NMR studies (Deczky and Langford, 1978) and an ESR study (Alberts et al., 1976) have shown that the Mn 2 + ion does not form an inner sphere and is bound electrostatically. This was supported by Bloom and McBride (1979) who, after extensive investigations with acid washed peat, concluded that peat and humic acids are likely to bind most divalent metal ions, with the exception of Cu2+ , largely as hydrated ions. The binding of copper appears to involve the exchange of one or two aquo ligands by carboxylate oxygens. Thus, neither chelation by adjacent functional groups nor heterogeneity with respect to acidity constants can be postulated to explain the binding of metal ions by peat and humic acids. Ong and Swanson (1966) carried out studies to challenge the theory that humic acids were primarily responsible for metal binding in peat and lignite. Their investigation centred around the sorption of copper by peat and lignite. They found that humic acids extracted from peat and lignite are able to complex copper in solution. IR studies indicated that the complex involved was probably the carboxylic 190


group and could be considered a chelating complex similar to that found in soil organic matter, thereby agreeing that the humic acids may have been, at least in part, responsible for the sorption of copper. However, the sorption of copper by peat and lignite from which the humic acids had been removed did not decrease, as expected, but actually increased. Accordingly, the sorption processes could not be solely explained by the formation of humic acid complexes. The increased copper sorption capacity of peat and lignite from which humic acids had been removed was explained as a consequence of the increased surface area in the organic matter that is exposed for metal sorption. Thus, humic acids in their soluble form are responsible for the fixation of metals, but in the solid form have quite different properties and can play only a very minor part in the sorption process. They concluded that the sorption of metal ions by peat and lignite can be considered as an attraction between the negatively charged surface and a positively charged metal ion, i.e. sorption. It was also noted that sorption capacity decreases as the degree of metamorphism increases (resulting from compaction which reduces surface area). The order of capacity is peat>lignite>coal. This theory is supported by more recent work. Bencheikh-Lehocine (1989) set out to determine whether sorption or ion exchange was responsible for zinc removal from peat at low pH. The theory is that sorption is an irreversible process whereas ion exchange is reversible, and this was the key to differentiating between the two processes. The effect of increasing sodium concentrations on the sorption of zinc in batch experiments was found to be negligible. It was then concluded that at low pH the process is irreversible and must be sorption of a strong type, encountered usually in chemisorption processes. In contrast, at moderate to high pH mechanisms other than that of sorption must be involved, and the mechanism of metal removal from solution will be generally termed sorption throughout this chapter.

Kinetics and rate of metal ion uptake Kinetics. In order to elucidate the transport and retention of pollutants in peat, Loxham (1980) developed a mathematical model. In applying the model to a field situation, it was found that the overall efficiency dropped quickly at high loadings. This was attributed to the slow mobilisation of the sorption capacity in the inactive zone. Loxham (1980) concluded that care must be taken in assuming the availability of sorption sites in determining the total sorption capacity of the peat. BencheikhLehocine (1989) found that the Langmuir model (see Chapter 2) was a better fit to the experimental data than the Freundlich model for peat-metal ion systems. However, Bhattacharya (1983), investigating the efficiency of crushed coal in removing cadmium from solution, concluded that the sorption data at equilibrium can be approximated by the Freundlich isotherm. Rate of sorption. Authors differ in their theories concerning the mechanism of sorption. They do, however, appear to be in agreement on one issue: the rate of sorption is rapid. Ong and Swanson (1966) noted that, with all systems studied, sorption by the organic matter was rapid, with nearly maximum sorption being achieved within 20 min. This was confirmed by Lee and Low (1989). Thus, as it was assumed to be a surface reaction, hydrolytic sorption could be considered as a simple condensation reaction between hydrolytic products of metal ions and sorbent functional groups. For batch processes a one hour residence time is 191


necessary for complete treatment of solutions on columns. This result led Gosset et al. (1986) to conclude that the concepts proposed by Lalancette (1974) would seem very unlikely to achieve complete metal removal by complexation, sorption or ion exchange. In a study of the removal of mercury from soil, Ahmad and Qureshi (1989) found that the time required to reach sorption equilibrium between the solid and liquid phases was two minutes. Also, Bunzl (1974a, 1974b) and Bunzl et al. (1976) studied the kinetics of metal ion binding on peat in a series of batch experiments using either continuous or discrete metal cation addition to preacidified sphagnum peats. They determined the reaction times to be between 10 and 30 seconds—much shorter than those quoted by most authors, who seem to agree on a figure in the range 20–60 min. Such a difference of magnitude in kinetic constants may be related to large differences in experimental conditions: Bunzl shredded and sieved peat samples and allowed time for it to establish swelling equilibrium in water, whereas Gosset et al. (1986), for example, adopted a drying period of 24 h at 70ºC.

Factors affecting metal ion sorption To identify the physiochemical conditions which are most effective for the removal of metal ions from wastewater, it is essential to have information about the effect of system parameters under varying process conditions. This is required for optimum system design criteria because wastewaters from different metallurgical and chemical processes may contain mixtures of metals and other waste compounds. Effect of pH. All authors discussing the subject agree that the natural capacity of peat to retain cations is related to the pH of the solution. In fact pH is critical to the sorption process (Ong and Swanson, 1966). Gosset et al. (1986) concluded that the percentage of metal extraction from solution onto peat varies from 0% to almost 100% within 4–5 pH units. The pH of the system should be lower than 8 for sorption to be significant, but a pH greater than 3.0 is required to prevent the metal ions from being exchanged by hydrogen ions, resulting in leaching or stripping of metals from the peat (Coupal and Lalancette, 1976; Lee and Low, 1989). In the pH range 3.0 to 3.5, the removal of most metal ions from solution ceases. Optimum sorption occurs within the pH range 3.5 to 6.5, and there is little difference in the amount of sorption of copper ions onto peat moss within this pH range (Lee and Low, 1989). BencheikhLehocine (1989) noted that at low pH, no leaching of humic or fulvic substances occurs, as shown by the absence of colour in the solution, and that zinc sorption by peat was moderate but at high pH zinc removal was considerably enhanced. The formation of a humic acid-zinc complex was advanced as a possible explanation. This view is supported by Ong and Swanson (1966). Coupal and Lalancette (1976) noted that at pH above 8.5 peat itself is not stable. Effect of surface area. Sorption increases as the surface area of the organic substance increases (Ong and Swanson, 1966). This explains the increase in sorption capacity after humic acids were extracted from the peat and lignite. Increasing metamorphism in natural carbonaceous deposits, i.e. peat>lignite>coal, results in compaction which produces a decrease in the surface area and a corresponding decrease in sorption capacity. Herman (1952) reported natural moisture contents for lignite ranging from 30 to 67%, indicating that they contain extensive porosity and consequently a large 192


internal surface area which facilitates access, for the cations, to carboxylic and phenolic functional groups. Brown coal from Victoria, Australia, in the dry condition, contains 2.33–3.28 meq/g carboxylic acid and 3.27–5.33 meq/g phenolic hydroxyl and has an internal surface area of 210–305 m2/g—determined by carbon dioxide sorption (Allardice et al., 1977). Peat typically has a surface area of over 200 m2/g, and is highly porous (95%) (Asplund et al., 1976). Effect of metal mixtures. In an effort to elucidate further the sorption process of heavy metals onto peat and lignite, studies have been performed to evaluate the effect of metal ion interaction. Chaney and Hundermenn (1979) found that, due to its polar nature, peat can adsorb large amounts of most metals. They noted, however, that pH and the presence of other ions may affect the sorption of a particular ion, having found, for example, that residual cyanide, chloride or other strong complexing agents might reduce the efficiency of Cd removal from peat columns. Lee and Low (1989) investigated the effect of the presence of equimolar amounts of Pb, Ni, Zn and Cd ions on the sorption of Cu onto moss. They concluded that there was no effect on the uptake of Cu. At low metal concentrations all the metals were adsorbed in almost equal amounts. At high concentrations Pb was sorbed to the same extent as Cu, whereas Cd, Ni and Zn were less competitive than Cu. The order of sorption is given as Pb, Cu>Cd>Ni, Zn. This order of sorption is almost identical to that obtained by Henderson et al. (1977), who studied the sorption of mixtures of metal ions onto ground scrap rubber. The only difference is that mercury was studied instead of lead. The order given is Hg, Cu>Cd>Ni, Zn. Ong and Swanson (1966), in a study of peat and lignite, concluded that several metals in a solution will compete against each other for open sites available on the organic matter. Those having the greatest ionic potential (ratio of charge to ionic radius) will be sorbed first and if the sites are still undersaturated, then those having lower ionic potential will be sorbed in sequence. Modified forms of peat and lignite. Some studies have pursued the development of peat and lignite to their maximum potential as sorbents of metals. This has been achieved by physical and chemical pretreatment of the peat and lignite prior to contact with metal-bearing wastes. A number of authors have observed that peat and lignite exhibit enhanced cation exchange capacities after treatment with phosphoric and/or sulphuric acid (Smith et al., 1976, 1977; Krishnamoorthy and Shanmugam, 1986), while others have described the characterisation of acid-treated forms of peat and lignite from various parts of the world (Diaconescu and Chirac, 1969). The development of activated carbons from peat and lignite would seem a natural progression following the success of coal-based activated carbons. Indeed, lignite-derived activated carbons are commercially available, whereas peat-based carbons are a relatively new idea. Fica (1972) investigated the production of activated carbons from peat. He concluded that, by means of zinc chloride activation, it is possible to obtain from peat or from a peat-lignite mixture activated carbons with good physical, chemical, mechanical and sorption properties and a microporous structure. Wolf et al. (1977) studied the effects of Ca on the sorption of Pb, Cd, Cu and Zn. They found that the selective uptake of metals was influenced by the presence of Ca, both in solution and as a natural constituent of peat, but in reverse correlation. Increasing the Ca concentration in peat enhanced sorption of heavy metals, while increasing the Ca concentration in solution reduced sorption. Dissanayake and 193


Weerasooriya (1981) found that peat can be used as a metal trapping material, but in the presence of Na the sorption of metals was very low. They determined that the reduction of sorption when monovalent ions, such as Na, are present indicates a buffering action of sorption sites. The ion-exchange properties of peat have been recognised for some time. In 1939, Nikol’skii and Paramonova used peat as an ion-exchange medium for metals such as copper, zinc, lead and mercury. More recently, Bunzl (1974a, 1974b) and Bunzl et al. (1976) have carried out extensive studies on the kinetics of the ion-exchange reactions during adsorption of heavy metal ions onto peat and humic acids. The mechanism of interaction of metal cations with peat soils has been reported by Gamayunov and Maslennikov (1992) as being primarily due to ion-exchange reactions along with adsorption via van der Waals interactions. The ion-exchange potential of lignite for the removal of dissolved metals from solution has been studied and the rate of uptake of metal ions onto peat and lignite compared by Baruah and Upreti (1994), who actually extracted the humic acid fraction from lignite for investigation of metal ion uptake. Interactions of the humic acid with seven cations showed preferential uptake of the Fe(III) ion. The authors concluded that, of the metal ions studied, the Fe(III) solution had the lowest pH and as such had the greatest tendency to form complexes with the humic acid functional groupings. The use of Australian lignite for the removal of quaternary ammonium salts present in effluents from minerals processing and hydrometallurgical plants has produced very successful results (Rice and Ikwue, 1990). The high cation-exchange capacity of peats has generated interest in the potential use of constructed and naturally occurring wetland systems as a low-cost, low maintenance method for the treatment of metal polluted waters (Wieder, 1990). Laboratory scale experiments have resulted in the production of a simple model for the quantitative assessment of the adsorption capacity of sphagnum peat and sawdust, when exposed to metal polluted water of particular chemical composition. Western Environmental Services, Colorado, has developed ‘peat moss beads’ specifically for the removal of metals from wastewaters. Developmental tests have demonstrated that the highly porous beads are capable of adsorbing more than 95% cadmium, zinc and manganese ions from wastewaters. In order to determine the maximum potential of peat and lignite as adsorbents, a number of authors have investigated methods for pretreatment of these materials prior to contacting with effluent. Ammoniated and acid-treated forms of peat have been shown to exhibit enhanced cation exchange capacities and coalescing properties by Smith et al. (1977). Kertman et al. (1993) found that increasing the number of sulphonated groups on the peat particle surface, by treatment with sulphuric acid, raises the cation-exchange capacity of peat almost to the level of synthetic organic ion exchangers. In particular, this chemically modified peat showed an increased capacity for the adsorption of lead ions, and after 20 sorption regeneration cycles the capacity of sulphonated peat for copper ions remained unaltered. Sulphonation of lignite has also been shown by Ibarra and Moliner (1984) to markedly increase the ion-exchange capacity of this adsorbent. Kinetic studies on the use of sulphonated lignite for the removal of copper ions from effluent have shown that ion exchange of Cu2+, on the surface acidic groups, occurs in the initial stages of the reaction. Once these acidic groups become saturated, the rate of removal of copper ions from solution is decreased as the Cu2+ diffuses into the pores of the adsorbent. 194


Investigations have also been carried out by Foersterling et al. (1990) into the potential of lignite, in its native state and after modification with dimethylglyoxime (DMG), for the removal of palladium from nuclear fuel wastes. It was found that treating the lignite with DMG greatly increased its capacity for palladium uptake. However, given sufficiently long contact times, the native lignite adsorbed approximately the same amount of palladium as DMG-treated adsorbent. Activated carbon can be produced from a wide variety of carbonaceous raw materials. Peat and lignite have proved to be promising starting materials for the production of activated carbons because of their ready availability at low price.

Chitosan Shellfish wastes from all Crustacea species are primarily mixtures of calcium carbonate, protein and chitin, but there is considerable variation between species as to the relative amounts of these substances and also as to minor components such as lipids, pigments and phosphates. Chitin is a naturally occurring polysaccharide which is found abundantly in both the animal and plant kingdoms. It is a major component of the exo-skeletons of anthropods such as crabs, lobsters and insects. Chitin may be regarded as a derivative of cellulose in which the hydroxyl group on the C-2 position has been replaced by the acetylamido group –NHCO–CH3. It is, therefore, a b -(1 4) linked polymer comprised of 2-acetylamido-2-deoxy- b -Dglucose. A derivative of chitin, namely chitosan, has proved to have considerable potential as a sorbent for the removal of metal ions from effluents. Chitosan can be regarded as a product of the deacetylation of chitin using concentrated alkali at high temperature. Chitosan is not a single substance but rather the name for a group of products which have been deacetylated to various extents. Chitin may be regarded as poly-N-acetyl-glucosamine deacetylated as little as possible, whereas chitosan is poly-N-acetyl-glucosamine deacetylated as much as possible but still not enough to be called polyglucosamine. Industrial wastes may be treated with chitosan which sorbs toxic metals such as copper, mercury or radioactive metals. Muzzarelli and Isolati (1971) showed that methyl mercury acetate can be removed with the aid of chitosan from solution, though organometallic compounds are not collected as well as simple ions. For mercury recovery Muzzarelli and Rocchetti (1974) found that powdered chitosan had a capacity of 114 mg of mercury per gram of chitosan.

®

Sorption column design

Contacting systems This section briefly reviews the methods of contacting sorbent and metal ion solution in column or bed-type systems. Consequently, the interpretation of ‘columns’ has been used fairly widely and the scope of this section includes the following: • moving mat filter • golf-green type filter 195


• fixed bed sorption • pulsed bed sorption • moving bed sorption • two-phase fluidised bed sorption • three-phase fluidised bed sorption.

Moving mat filter The basis of the Hussong-Couplan wastewater treatment plant described by Leslie (1974) and referred to earlier is a moving mat filter of peat. This purification process is based on the scrubbing action of a moving mat of peat as the effluent is passed through. The system successfully removed over 98% of several metals in a process involving both sorption and complexing mechanisms. The process system comprises three stages: • preparation of the peat • contacting the peat with the effluent • removal of the used peat. Figure 9.1 shows the contacting system. Peat is loaded into a preparation hopper and some water is added to minimise dust emissions and facilitate handling operations. This moist peat with additional water is fed at a controlled rate into a mixing tank from which the peat slurry is fed evenly onto a perforated moving belt which permits drainage of any excess water. The conveyor belt takes the peat mat filter along under

Figure 9.1 Hussong-Couplan system for treating textile effluent using peat

196


discharge pipes from which the metal ion-bearing textile effluent is sprayed onto the peat. The treated effluent solution passes through the mat filter onto collection trays. The effluent-peat contacting process is a two-stage operation; initially the raw effluent contacts partially used peat that has already completed purification of partially treated effluent. In the second stage, the partially treated effluent is contacted with fresh peat to maximise the removal of pollutants. After the second stage of the liquid treatment the treated effluent may be recycled or may be discharged. The proposed disposal methods for the used peat are sanitary landfill, with no danger that the metals, now in a highly insoluble condition, will leach and with recovery of the heavy metals providing some return on investment.

Golf-green type filter Although this device has been primarily used for the treatment of domestic wastewater and sewage it has achieved success by the removal of large amounts of phosphorus, as described by Surakka and Kamppi (1971). A peat and peat-sand filter arrangement has been used on a pilot scale, and a full-scale fluid plant as shown in Figure 9.2 has been proposed by Farnham and Brown (1972). The system is relatively simple and comprises a mound type or golf-green type filter system and a rock or gravel irrigation system to take the treated wastewater away from the peat/peat-sand filter. The main disadvantages concern the ease or difficulty of regeneration and assessing whether or not leaching is a problem.

Fixed bed sorption Sorption breakthrough curve. In general, from a practical and design viewpoint, the behaviour of a pollutant-sorbent system in a column is expressed graphically using the breakthrough curve concept. The breakthrough curve represents the sorption section within a fixed bed of sorbent (sometimes called the mass transfer zone or MTZ). The sorbable impurities are removed as the liquid passes through the carbon bed, and the portion of the bed in which sorption is occurring is defined as the sorption zone or MTZ.

Figure 9.2 Golf-green type filter bed

197


In Figure 9.3, the wavefront or MTZ can be expressed in terms of time or volume of effluent treated. The area below the wave (agdea) reflects unused sorbent capacity, and the area ratio (agdea)/(abdea) is the fraction of unused sorbent bed in the MTZ. A vertical straight line drawn through g, the 50% breakthrough point, and the area (afcba)=(agdba) and (fedcf)=(agdea) yields an equivalent stoichiometric front for the system at time t s. The rectangle (hfckh) corresponds to sorbent at its equilibrium loading, Y e; it is defined as the equivalent equilibrium section and is specified in terms of the length of this section, Ze. The area (fedcf) corresponds to adsorbent at its initial loading, Y0 ; it is specified in terms of the length of unused bed, Z 0. The adsorbent bed at breakthrough actually comprises an equilibrium zone and a mass transfer zone. A downflow system is shown in Figure 9.4. The sorption zone is the distance between the exhausted carbon at the top of the zone and the unused sorbent at the bottom of the zone. Sorbent at the top of the sorption zone, which is always contacted by a fresh, full-strength wastewater, becomes exhausted and the solute concentration becomes Cd. After the lower boundary of the sorption zone has progressed through the full depth of the bed and reached the base at Cc, breakthrough begins. As the sorption zone moves out of column, the exhaustion curve rises sharply. After the zone has moved all the way out, the entire bed of sorbent is exhausted, and the impurity concentration Cd of the effluent is the same as that of the feed, C0. Experimentally determined sorption breakthrough curves for mercuric ions at different bed depths are shown in Figure 9.5. Fixed bed systerns can be operated as a single column or as multiple columns in parallel, in series, or in combined parallel and series. A single-column system will

Figure 9.3 Mass-transfer zone and sorption breakthrough curve

198


Figure 9.4 The sorption wave

Figure 9.5 Breakthrough curves for the sorption of mercuric ions on activated carbon with different bed depths

199


operate effectively if (1) the breakthrough curve is steep, (2) the sorbent in the column will last so long at the desired processing rate that the cost of replacing or regenerating it becomes a minor part of operating expenses, or (3) the capital cost of a second or third column cannot be justified because not enough sorbent cost can be saved to pay for additional equipment. Otherwise a multiple-column system, which may be either series or parallel, should be designed. A multiple-column system is usually best if (1) the process cannot be interrupted for unloading, reloading, or regeneration, and a standby column is not available, and (2) the size of or height of a single column to hold the required amount of sorbent will not fit in the existing space. A parallel-column system (in which the columns are placed on-stream at evenly spaced time intervals) will receive the same feed and discharge into a common manifold. A system of the type shown in Figure 9.6 is usually employed if pressure drop is likely to be a problem. With this system, pumps can be smaller, power requirements lower, and pressure specifications for column and piping less stringent, especially if the flow will be high. The sorbent in a single column, or in a parallel system which is moved for regeneration or discarded, is usually not completely spent. That is, the adsorption zone is still within the column when the carbon is discharged. A series-column system, in which the effluent from one column becomes the feed for the next (Figure 9.7), is usually adopted if (1) the breakthrough curve is gradual and the highest possible effluent purity is desired, or (2) the combination of a gradual breakthrough curve and a high carbon requirement per unit of wastewater treated makes it economically necessary to exhaust the sorbent completely. When the sorbent in the bed column of a series system becomes spent, it is removed for dumping or regeneration and a fresh column is put on stream at the end of the series. The second column in the series becomes the lead column.

Figure 9.6 Schematic diagram of a parallel column system

200


Figure 9.7 Schematic diagram of a series column system

The sorbent in the series system is fully used because the leading column is removed after the sorption zone has passed completely through it. The zone continues to pass through the ‘downstream’ columns. Therefore, the operating costs would usually be lower than those for a single column or parallel system. A combined series and parallel system couples the high efficiency of series operation with the longer residence times of a parallel system. A major advantage of downflow operation is filtration. Figure 9.8 shows construction details of a downflow contactor. Suspended solids will be removed by the finer sorbent particles at the top of the backwashed bed. However, pumps must be sized to accommodate the pressure drop through the system, which could be substantial if a high degree of suspended solids is present. To operate an upflow system, the piping is simpler because the direction of flow is the same during sorption and washing cycles. Pressure drops are low because the sorbent bed expands with the flow. The problems of sorbent regeneration, reactivation and metal recovery are dealt with later. Breakthrough curve analysis using LUB theory. If a stable MTZ moves down through an adsorber with a uniform velocity U, it is conceptually possible to superimpose a stoichiometric transfer front on the actual transfer front. Providing the S-shaped breakthrough is reasonably symmetrical, the assumption of being able to superimpose a stoichiometric front onto the actual breakthrough curve is reasonable and is used as the basis for design analysis and sizing in the length of unused bed (LUB) model and the bed depth service time (BDST) model. This approach is based on determining the LUB at the moment the solute breaks through the bed to the design breakthrough concentration. The L value is determined by the following methodology. The total amount of solute removed, w, from the bulk fluid between time t and time t+Dt is expressed as (9.1) 201


Figure 9.8 Downflow contactor

As w kg of solute is adsorbed, the shape of the MTZ is unaltered although its position changes and there is an increase in bed length, DZ; then (9.2) where r is the bulk density of the adsorbent and A is the surface area for sorption. Combining Equations 9.1 and 9.2 and rearranging gives (9.3) or (9.4) During steady-state operation, the values of L, r, DC, and DY are fixed so that at any time, t, the position of the stoichiometric front, Zs, is given by (9.5) and by definition at the breakthrough time, tb (9.6) 202


Therefore (9.7) At the stoichiometric time, ts (9.8) Therefore (9.9) but (9.10) Combining Equations 9.8, 9.9 and 9.10 and eliminating U gives (9.11) Thus, simple breakthrough curves provide the data required for the correlation of rate data in dynamic systems, and the basic equation which may be used for the analysis of breakthrough data is (9.12) Breakthrough curve analysis using BDST theory. In the operation of a sorption column, wastewater with an influent solute metal ion concentration C0 is fed into the column. It is desired to reduce the solute’s concentration in the effluent to a value not exceeding Cb, as determined by water quality requirements. At the beginning of the operation, when the sorbent is fresh, the effluent concentration is actually lower than the allowable concentration C b. As the operation proceeds and sorbent reaches saturation, effluent concentration reaches C b . This condition is called the breakthrough point. Fixed bed data can be used in the BDST analysis. The BDST method presented by Hutchins (1973) is based on work by Bohart and Adams (1920). The service time of a fixed bed of adsorbent, treating a dilute solution of a single adsorbate, can be expressed as a fixed function of operational parameters in the form of a modified Bohart-Adams equation: (9.13) which has the form of an equation for a straight line (9.14) and Cx can be written (9.15) When design data are required for a change in the initial solute concentration fed to the adsorption system, Hutchins (1973) proposes that the new gradient, m´x, can be written as 203


(9.16) and the new intercept as

(9.17)

If design data are required for a change in volumetric flow rate of solute to the adsorption system, Hutchins recommends that the new gradient can be written as (9.18) and the new intercept should be unchanged: (9.19) Table 9.4 gives the equation constants for various pollutants, as follows. (a) Bed height. The critical bed depth Z 0 is the theoretical depth of carbon sufficient to prevent solute concentration from exceeding value Cb at t=0. Z 0 is obtained from an equation which, in its original form as suggested by Bohart and Adams (1920), is (9.20) By letting t=0, Z0 is obtained from Equation 9.20 by solving for Z (which is equal to Z0 in this case). Since the exponential term is usually much larger than unity, the unity term within brackets on the right-hand side of Equation 9.20 is often neglected. The final result is (9.21) (b) Pollutant velocity. In the design of fixed bed adsorption columns contact time is most important and, therefore, bed depth and solute flow rate are the major design parameters. Figure 9.9 shows the BDST plots for Hg 2+. It also shows different percentage breakthroughs. The BDST model describes the Hg 2+ sorption quite accurately. Table 9.4 BDST equation constants for Hg 2+ on carbon Pollutant

204


Figure 9.9 Effect of breakthrough by the BDST model for the sorption of mercuric ions (C 0=25 mg/l, d p=302 µm)

(c) Effect of percentage breakthrough. The primary objective of operating a fixed bed column adsorption system is to achieve a certain degree of reduction in the solute concentration, which may be called a percentage breakthrough value. The BDST model can predict service time versus bed depth according to the desired percentage breakthrough value. Figure 9.9 shows the effect of changing percentage breakthrough in the case of reducing Hg2+ concentration from 25 mg/l to 2%, 5% and 10% breakthrough values. The general observation is that no significant increase in throughput is achieved by varying the percentage breakthrough because the sorption breakthrough curves are rather steep and, consequently, the mass transfer zone is very short. The broken line in Figure 9.10 is the predicted BDST curve for changing the initial concentration from 25 mg/l to 10 mg/l at a fixed breakthrough value of 0.5 mg/l. The BDST model offers a rapid method of designing fixed bed columns. A great deal of the design and operational data can be extracted from log books (i.e. operating records) of wastewater treatment plants so that changes and modifications in operational conditions can be predicted. Optimisation of fixed bed sorber design using EBRT theory. The selection of contacting and regeneration systems for sorbent utilisation in wastewater treatment processes is integrated into a total system design. Although contacting and regeneration systems are essentially independent, they are linked by the sorbent utilisation rate. 205


Figure 9.10 Theoretical BDST line for the sorption of mercuric ions

Many variables may influence the design of fixed bed sorbers, and it is advisable to utilise the most economical conditions. A procedure referred to as the empty bed residence time, or EBRT, provides a simplified method for minimising adsorbent usage. Two types of variables are recognised as important in optimising sorption systems. First, there are primary process variables which are mainly determined by sorbent exhaustion rates. Second, there are those process variables which affect the operating costs as a result of changes in the required level of effluent purity, influent liquid purity, temperature and viscosity, as well as sorbent particle size. For a given system, in which the liquid flow rate, impurity concentrations, and sorbent characteristics are fixed, the capital and operating costs are basically a function of two variables: the sorbent exhaustion rate, which is usually expressed as mass sorbent used per volume of liquid treated; and superficial liquid retention time, empty bed contact time (EBCT), more often referred to as empty bed residence time (EBRT), which is the time that the liquid would take to fill the volume of the sorbent bed and is a direct function of liquid flow rate and sorbent volume. 206


For a given system, in which the liquid flow rate, impurity concentrations and sorbent characteristics are fixed, the costs depend almost entirely on the above two variables. The total capital cost is established primarily by the volume of the sorbent columns, and usually, to a lesser extent, by the size of the reactivation or regeneration system. The operating costs are determined by the sorbent exhaustion rate since the largest variable is usually the cost of the make-up sorbent. In the case of low-cost sorbents, attention must be paid to the regeneration and disposal costs. The relationship between these variables can be established by tests and plotted as shown in Figure 9.11. For a given system to achieve a certain level of performance, there is a single line relating these two variables which is called the operating line. The operating line concept can be used to optimise the basic design to achieve the lowest cost or other objectives. The operating line generally approaches a minimum on both axes. The minimum exhaustion rate for a given sorption duty is that which is achieved when the exhausted sorbent is in equilibrium with the influent liquid, and the minimum retention time represents the minimum volume of sorbent necessary to achieve the desired effluent purity at infinitely high sorbent exhaustion rates. The arrangement of the sorber vessels, the influent liquid purity, temperature, sorbent particle size and other variables affect the total cost. Figure 9.12 shows the effect of sorber flow arrangement. Once the operating line is established, it is possible to select the combination of sorbent exhaustion rate and liquid retention time which gives the optimum or lowest cost design. Since both capital and operating costs are involved, these must be

Figure 9.11 EBRT operating line plot

207


Figure 9.12 Effect of sorber flow arrangements

reduced to a common annual cost which is appropriate to the situation (Allen et al., 1967). The annual capital value varies from 5% to 40% depending on the application. Typical annual costs are shown in Figure 9.13. The EBRT plot for Hg 2+ sorption onto carbon (Figure 9.14) shows the effect of percentage breakthrough. Higher breakthrough means a higher concentration of pollutant in the final effluent and this also indicates that more effluent is being

Figure 9.13 Typical relative annual costs to remove ABS from water

208


Figure 9.14 Fixed bed optimisation for the removal of mercuric ions: effect of percentage removal ( C0=25 mg/l, flow rate=2.78 ml/s, dp=302 µm)

processed. Figure 9.14 also shows that, for a fixed EBRT, the sorbent (in this case carbon) exhaustion rate is higher for lower percentage breakthroughs. However, as the sorbent bed is made larger, an optimum value of EBRT is approached. Therefore, it is possible to seek different design alternatives regarding sorbent bed size and column configurations. This fixed bed operating line concept has been used to determine the optimum sorbent exhaustion rate. However, the effects of flow rates, sorbent particle size, and percentage breakthrough need to be studied and defined. The operating line plots yield valuable information regarding minimum sorbent exhaustion rate and minimum EBRT. The EBRT is found to approach a limiting value with respect to sorbent usage rate. The concept of operating line optimisation has shown that the kind of data obtained from pilot-scale plant can provide valuable information for the design of sorption systems.

Pulsed bed sorbers In pulsed beds, the carbon bed moves in the opposite direction to the wastewater flow. The wastewater flows upwards while spent carbon is removed from the 209


bottom of the column and an equal volume of fresh carbon or regenerated carbon is added to the top. This resembles a number of stacked fixed bed columns operating in series. Pulsed beds (Figure 9.15) are normally operating with the columns filled with carbon, so there is no free space to allow bed expansion during operation or cleaning. It is desirable to maintain ‘plug’ flow of carbon so that a sharp adsorption zone will be maintained.

Moving bed sorbers Steady-state conditions require continuous movement of both fluid and adsorbent through the equipment at a constant rate, with no change in composition at any point in the system with passage of time. If parallel flows of solid and fluid exist, the net result is at best a condition of equilibrium between effluent streams or the equivalent of one theoretical stage. However, counter-current operations of such systems enable separations equivalent to many stages to be developed. A schematic

Figure 9.15 Pulsed bed system

210


Figure 9.16 Schematic diagram of the continuous counter-current process

of a continuous counter-current sorption system is shown in Figure 9.16. A solute balance around the entire tower is given by (9.22) and the mass balance around the upper part of the column is given by (9.23) From the mass-balance equations, the operating line can be constructed on the equilibrium curve (Figure 9.17). The operating line is the straight line of slope V/G joining the terminal conditions (C0, Y0) and (C1, Y1). The concentration of solute C and Y, at any height within the column, will lie on this line. If one assumes that the column operates under approximately isothermal conditions and that film and internal diffusional resistances can be represented by an overall mass-transfer term, k0, then the incremental mass-transfer rate is (9.24) 211


Figure 9.17 Operating diagram for the continuous contact sorption

where dZ is an incremental height in the adsorber, a is the external surface of the adsorbent particles, m is the slope of the equilibrium curve, and Ye is the equilibrium solute concentration in the solid corresponding to composition C. The driving force, DY=Y e , is obtained from the vertical distance between the operating line and the equilibrium curve in Figure 9.17 and the number of masstransfer units, N, may be defined by (9.25) The number of mass-transfer units may be obtained by graphical integration, and the height of the bed, Z, is obtained by combining Equations 9.24 and 9.25 to give (9.26) where the height of the mass-transfer unit, Ht, is defined by: (9.27) 212


Mass-transfer correlations for evaluating k0 have been reviewed by Heister et al. (1977). Large-scale devices for the continuous contacting of granular solids and fluids were developed after certain operational problems had been overcome. These difficulties included obtaining uniform flow of solid particles and fluid without channelling or local irregularities, as well as those of continuously introducing and removing the solid from the vessel to be used. Such devices include the hypersorber (Kehde et al., 1948; Berg, 1946, 1951) and the Higgins contactor (Kidnay and Myers, 1966), the latter being specifically designed for solid-liquid ion-exchange operations. The former systems utilise the principles of fluidisation, and recent work on fluidisation has been applied specifically to colour removal.

Fluidised bed sorbers In wastewater treatment systems, it is advantageous to keep the particle size as small as possible, so that high rates of sorption may be achieved. Columnar operation, in which the solution to be treated flows upwards through an expanded bed of the particulate sorbent, is one method of taking advantage of small particle size. The main problems of the use of small particles in fixed beds are excessive head loss, air binding and fouling with particulate matter. The design of a continuously operating counter-current fluidised bed is relatively simple (Kunii and Levenspiel, 1969), although the solution of the design equations can become quite complex (McKay and Alexander, 1977). Determination of the mass-transfer coefficient, k0, provides the fundamental design parameter for fluidised bed systems. An intraparticle diffusion-controlled adsorption model has been proposed, and also a ‘completely mixed column’ model has been used for the removal of alkylbenzene sulphonate from wastewater using a fluidised bed of carbon (Weber and Morris, 1965) and basic dye has been removed from effluent using a fluidised bed of silica (McKay and Alexander, 1977). Ganho et al. (1975) and McKay and Bino (1987) studied the adsorption of phenol onto activated carbon. These models should also be applicable to the removal of metal ions from solution, providing the same mass transfer resistances are rate controlling. Assuming steady-state conditions, the rate of mass transfer to the sorbent across a film is (9.28) The driving force is the product of the concentration gradient between adsorbent and solution and the total area across which transfer is taking place. The resistance is assumed to be uniform throughout the system, and the reciprocal of this resistance is the transfer coefficient k0. Therefore (9.29) where A is the effective transfer area per unit mass of adsorbent, M is the mass of adsorbent, DC is the concentration difference across the film, and r is the rate of transfer. 213


Since the investigations were performed under non-steady conditions, the rate of transfer may be expressed in terms of the concentration in solution at any time and the concentration gradient across the film at that time. Due to the strong affinity of the solute for the sorbent, a partition coefficient, P, is incorporated to increase the concentration gradient term. Its value is the mass concentration of sorbate in solution, at time t, divided by the mass concentration in the sorbent at time t, (9.30) One of the potential advantages of fluidised bed reactors is the reduction of intraparticle mass transport resistance in comparsion with that in fixed bed reactors, which are simple to construct and operate and are particularly suitable for the sorptive removal of strongly sorbed solutes. The mechanistic processes which are considered in representing single component sorption are: • film diffusion from the fluid phase to the surface of the particle • sorption on the surface • pore diffusion in liquid-filled pores onto the particle surface. Models considering internal mass transfer, as well as external mass transfer and a ‘completely mixed’ column, are available (McKay, 1983). The following assumptions are made in the theoretical analysis of the problem: • the sorption isotherm corresponds to a Langmuir isotherm • the concentration of solute remains constant in a cross-sectional area • perfect mixing of solids occurs in the bed • for short intervals of time a quasi-stationary regime is developed, and C is a function only of adsorber length, l. The differential mass balance. The differential material balance represented in Figure 9.18 may be written as: (9.31) and for the whole fluidised bed: (9.32) Boundary conditions are at t=0, CL=C0, Y=0. The following assumptions are made in this analysis: • C is constant in a cross-sectional area • perfect mixing of solids occurs and so Y does not change with position but is a function of time • for a short interval of time a quasi-stationary regime is developed and C is a function only of l 214


Figure 9.18 Material balance relation for column

• the sorption isotherm corresponds to a Langmuir isotherm written as (9.33) Fluid phase mass transfer coefficient. Using the fluid phase mass transfer coefficient, it is found that (9.34) Equation 9.34 can be rearranged to determine the fluid phase mass-transfer coefficient: (9.35) and Figure 9.19 shows the theoretical diagrams used to define mass-transfer coefficients. Solid phase mass-transfer coefficient. The following equation is written for the solid phase mass-transfer coefficient: (9.36)

(9.37)

215


Figure 9.19 Theoretical diagrams to define mass-transfer coefficients

This method enables the mass-transfer term to be determined, and this parameter can now be used to predict the performance of fluidised bed sorbers.

Regeneration and metal recovery The inverse process of removing metals from peat and lignite has received only scant attention. Many authors have observed the effect of low pH on stripping metals from peat, but have failed to connect the phenomena with possible regeneration. This may be due to the fact that in the past recovery and regeneration were not as economically or politically attractive as they have become in recent years. Coupal and Lalancette (1976), working with peat, and Cullen and Siviour (1981), working with lignite, proposed that spent materials could be disposed of by incineration or landfill. Cullen and Siviour (1982) suggested that the metal residues could be recovered by subsequent treatment of the post-combustion ash. Gosset et al. (1986) observed that, except for nickel, which seems so strongly complexed on peat that at pH 1.2–2.0 half of the maximum sorption capacity is attained, metals may be easily removed from peat by acid treatment. Aho and Tummavuori (1984) studied the effect of the repeated use of peat columns on the structure of the peat, its capacity to remove metals, acid elution profiles and the maximum column flow rate. After several cycles of sorption, elution and washing with distilled water it was found that the ionexchange capacity of the peat remained within narrow limits during repeated Cu 2+ sorptions, thus the cations sorbed can be released with a small volume of acid and the peat sorber repeatedly used. However, the release of bound ions is strongly dependent on the chemical composition of the stripping solution. This desorptive ion exchange is important in the design of metal recovery processes and in controlling the leaching of metals from peat and lignite used in metal clean-up operations and subsequently disposed of in landfills. 216


References AHMAD, S. and QURESHI, I.H., 1989, Fast mercury removal from industrial effluents, Journal of Radioanalytical and Nuclear Chemistry, 130, 347–352. AHO, M. and TUMMAVUORI, J., 1984, On the ion-exchange properties of peat-IV. The effects of experimental conditions on ion-exchange properties of sphagnum peat, SUO, 35, 47–53. ALBERTS, J.J., SCHINDLER, J.E., NUTTER, D.E. and DAVIS, E., 1976, Elemental, infrared spectrophotometric and electron spin resonance investigations of non-chemically isolated humic material, Geochimica Cosmochimica Acta, 40, 369–372. ALDER, E. and LINDQUIST, K., 1963, Spectrochemical estimation of phenylcoumaran elements in lignin, Acta Chemica Scandinavica, 17, 13–26. ALLARDICE, D.J. et al., 1977, Peat for Power Generation. State Electricity Commission of Victoria, Melbourne, Australia, Report 342. ALLEN, J.B., JOYCE, R.S. and KASCH, R.H., 1967, Process design calculations for adsorption from liquids in fixed beds of granular activated carbon, Journal of the Water Pollution Control Federation, 39, 217–223. ARDEN, T.V., 1966, Water Purification by Ion Exchange, Butterworths, London. ASPLUND, D., EKMAN, E. and THUN, R., 1976, Counter-current peat filtration of waste water, Proceedings of the 5th International Peat Congress, Poznan, 1, pp. 358–371. BARUAH, M.K. and UPRETI, M.C., 1994, Preferential uptake of Fe(III) ions by humic acid extracted from lignite, Fuel, 73, 273–275. BELKEVICH, P.I., GAYDUK, K.A. and CHRISTOVA, L.R., 1976, Role of peat in decontamination of environment, Proceedings of the 5th International Peat Congress, Poznan, 1, pp. 328–348. BENCHEIKH-LEHOCINE, M., 1989, Zinc removal using peat adsorption, Environmental Technology Letters, 10, 101–108. BERG, C., 1946, Adsorption using hypersorbers, Transactions of the American Ind. Chem. Engrg, 42, 665–673. BERG, C., 1951, Hypersorption design modern advancements, Chemical Engineering Progress, 47, 585–593. BHATTACHARYA, A.K., 1983, Adsorption of Cadmium and Complexes. Ph.D Thesis, Indian Institute of Technology, Kanpur. BLOOM, P.R. and MCBRIDE, M.B., 1979, Metal ion binding and exchange with hydrogen ions in acid-washed peat, Soil Science Society of America Journal, 43, 678–692. BOHART, G.S. and ADAMS, E.Q., 1920, Adsorption in columns, Journal of the Chemical Society, 42, 523–529. BOYD, S.A., SOMMERS, L.E. and NELSON, D.U., 1981, Copper (II) and iron (III) complexation by the carboxylate group of humic acid, Soil Science Society of America Journal, 45, 1241–1243. BROWN, P.A. et al., 1992, The removal of toxic metals from wastewaters using peat and lignite, Proceedings of the Institution of Chemical Engineering Research Symposium, Queens College Cambridge, pp. 46–56. BUNZL, K., 1974a, Kinetics of ion exchange in soil organic matter. II. Ion exchange during continuous of Pb2+ ions to humic acid and peat, Journal of Soil Science, 25, 343–356. BUNZL, K., 1974b, Kinetics of ion exchange in soil organic matter. III. Differential ion exchange reactions of Pb2+ ions in humic acid and peat, Journal of Soil Science, 25, 517– 532. BUNZL, K., SCHMIDT, W. and SANSONI, B., 1976, Kinetics of ion exchange in soil organic matter. IV. Adsorption and desorption of Pb 2+, Cu2+, Cd2+, Zn2+ and Ca2+, Journal of Soil Science, 27, 32–41. CHANEY, R.L. and HUNDEMENN, P.T., 1979, Use of peat moss columns to remove cadmium from wastewaters, Journal of the Water Pollution Control Federation, 51, 17–21. 217

Biosorbents for Metal lons CHERMISINOFF, P.N. and ELLERBUSCH, F., 1978, Carbon Adsorption Handbook, Ann Arbor Science Publishers Ltd., Ann Arbor, MI, 2–3. CLYMO, R.S., 1963, Migration of metal ions, Annals of Botany, 27, 300–324. COUPAL, B. and LALANCETTE, J.M., 1976, The treatment of waste waters with peat moss, Water Research, 10, 1071–1076. CULLEN, G.V. and SIVIOUR, N.G., 1981, Equilibrium relationships in the extraction of metals from ammoniacal solution with low rank coals. Journal of Mining Engineering, 371–377. CULLEN, G.V. and SIVIOUR, N.G., 1982, Removing metals from waste solutions with lowrank coals and related materials, Water Research, 16, 1357–1366. DEAN, J.G., BOUSQUI F.L. and LANOUETTE K.H., 1987, Removing heavy metals from waste water. Environmental Science and Technology, 6, 518–522. DECZKY, K; and LANGFORD, C.H., 1978, Application of water nuclear magnetic resonance relaxation times to the study of metal complexes of the soluble soil organic fraction fulvic acid, Canadian Journal of Chemistry, 56, 1947–1951. DE MUMBRUM, L.E. and JACKSON, M.L., 1956, Metals ions in peat soils, Proceedings of the Soil Science Society of America, 20, 334–337. DIACONESCU, E. and CHIRAC, F., 1969, Cation exchangers from peat of Dersca, Dorohoi. Bull Inst. Politech Iasi, 15, 41–45. DISSANAYAKE, C.B. and WEERASOORIYA, S.V. R., 1981, Peat as a metal-trapping material in the purification of industrial effluents, International Journal of Environmental Studies, 17, 233–238. EGER, P., LAPAKKO, K. and OTTERSON, P., 1980, Trace metal uptake by peat: interaction of a white cedar bog and mining stockpile leachate, Proceedings of the 6th International Peat Congress, Duluth, Minnesota, pp. 542–547. FARNHAM, R.S. and BROWN, J.L., 1972, Use of peat and peat-sand filtration media, Proceedings of the 4th International Peat Congress, Helsinki, Finland. FOERSTERLING, H.U., BEER, M. and HALLMEIER, K.H., 1990, Investigation of the adsorption of palladium on carbonaceous adsorbents modified with dimethylglyoxime, III—The adsorption of palladium on a lignite in its unmodified form and modified with dimethylglyoxime, Carbon, 28, 503–508. GAMAYUNOV, N.I. and MASLENNIKOV, B.I., 1992, Mechanism of interactions of cations with the adsorptive complex of peat soil, Eurasian Soil Science, 24, 32–37. GAN, H., NARDI, S.P. and WALKER, P.L., 1972, Nature of the porosity in American coals, Fuel, 51, 272–275. GANHO, R., GILLBERT, H. and ANGELINO, H., 1975, Adsorption in fluidized beds, Chemical Engineering Science, 30, 1231–1240. GLOOSCHENKO, W.A. and CAPOBIANCO, J.A., 1982, Trace element content of northern Ontario peat, Environmental Science and Technology, 16, 187–188. GOSSET, T., TRANCART, J.L. and THEVENOT, D.R., 1986, Batch metal removal by peat: kinetics and thermodynamics, Water Research, 20, 21–26. HEISTER, N.K., VERMEULEN, T. and KLEIN, G., 1977, Chemical Engineers Handbook, PERRY, R.H. (Ed.), McGraw-Hill, New York, pp. 16.1–16.48. HENDERSON, R.W., LIGHTSEY, G.R. and POONAWALA, N.A., 1977, Competitive adsorption of metal ions from solutions by low-cost organic materials, Bulletin of Environmental Contamination and Toxicology, 18, 340–343. HERMAN, H., 1952, Brown Coal, State Electricity Commission of Victoria, Melbourne, Australia, pp. 4–7. HUTCHINS, R.A., 1973, New simplified design of activated carbon systems, American Journal of Chemical Engineering, 80, 133–138. IBARRA, J.V. and MOLINER, R., 1984, Copper removal from effluents using a sulphonated lignite, Fuel, 63, 373–377. KADLEC, R.H. and KEOLEIAN, G.A., 1986, In: Peat and Water: Aspects of Water Retention 218

Low-cost Adsorbents in Continuous Processes and Dewatering in Peat, FUCHSMAN, C.H. (Ed.), Elsevier Applied Science Publishers, London, New York. KADLEC, R.H. and RATHBURN, M.A., 1983, Copper sorption on peat, Proceedings of the International Symposium on Peat Utilisation, Bemidji State University, FUCHSMAN, C.H. and SPIGARELLI, S.A. (Eds), Bemidji, Minnesota. KASHIRTSEVA, M.F., 1960, Peat soils and ion exchange, International Geological Reviews, 2, 52–59. KEHDE, H., FAIRFIELD, H.G., FRANK, J.C. and ZAHNSTECKER, L.W., 1948,Design of adsorption columns, Chemical Engineering Progress, 44, 575–585. KERTMAN, S.V., KERTMAN, G.M. and CHIBRIKOVA, Z.S., 1993, Peat as a heavy metal adsorbent, Russian Journal of Applied Chemistry, 66, 382–383. KIDNAY, A.J. and MYERS, A.L., 1966, Simplified method for prediction of multicomponent adsorption equilibria from single gas isotherms, American Institute of Chemical Engineers Journal, 12, 981–989. KRISHNAMOORTHY, H.S. and SHANMUGAM, T., 1986, Effluent treatment with lignite, Indian Journal of Environmental Protection, 6, 169–171. KRISHNAN, S.S., CANCILLA, A. and JERVIS, R.E., 1987, Industrial wastewater treatment for toxic heavy metals using natural materials as adsorbants, Journal Radioanalytical Nuclear Chemistry, 110, 373–378. KUNII, D. and LEVENSPIEL, O., 1969, Fluidisation Engineering, J.Wiley and Co., New York. LALANCETTE, J.M., 1974, Removal and Recovery of Metals from Polluted Waters, U.S. Pat. 3,790, 370. LALANCETTE, J.M. and COUPAL, B., 1972, Recovery of mercury from polluted water through peat (moss) treatment, Proceedings of the 4th International Peat Congress, Otaniem, Finland, pp. 213–218. LARSEN, V.J. and SCHIERUP, H.J., 1981, Phosphorus in soil—advances in agronomy, Environmental Quality, 10, 188–193. LEE, C.K. and Low, K.S., 1989, Removal of copper from solution using moss, Environmental Technology Letters, 10, 395–404. LESLIE, M.E., 1974, Peat. New medium for treating dye house effluent, American Dyestuff Reporter, 63, 15–18. LOWRY, H.H., 1945, Chemistry of Coal Utilisation, John Wiley & Sons, New York. LOXHAM, M., 1980, Proceedings of the 6th International Peat Congress, Duluth, Minnesota, pp. 600–606. MCKAY, G., 1983, The adsorption of dyestuffs from aqueous solutions using activated carbon, Journal of Chemical Techology and Biotechnology, 33A, 196–204. MCKAY, G. and ALEXANDER, F.A., 1977, Kinetics of the removal of basic dye from effluent using silica, The Chemical Engineer, 243–247. MCKEOWN, D., 1987, Metal Removal by Flocculation, BSc. Dissertation, Queen’s University Belfast. MCLELLAN, J.K. and ROCK, C.A., 1987, Pretreating landfill leachate with peat to remove metals, Environmental Monitoring and Assessment, 9, 203–215. MUZZARELLI, R.A.A. and ISOLATI, A., 1971, Metal ion uptake on modified chitosan, Water, Air and Soil Pollution, 1, 65–71. MUZZARELLI, R.A.A. and ROCCHETTI, R., 1974, The deacetylation of chitosans, Journal of Chromatography, 96, 115–121. NETZER, A. and HUGHES, D.E., 1984, Adsorption of copper, lead and cobalt by activated carbon, Water Research, 18, 927–933. NETZER, A., WILKINSON, P. and BENEDICT, S., 1974, Removal of trace metals from waste water by treatment with lime and discarded automotive tires, Water Research, 8, 813–817. NIKOL’SKII, B.P. and PARAMONOVA, V.I., 1939, Peat for treatment of metal ions, Uspekhi Khim, 8, 1535–67. 219

Biosorbents for Metal lons ONG, H.L. and SWANSON, V.E., 1966, Adsorption of copper by peat, lignite, and bituminous coal, Economic Geology, 61, 1214–1231. PAKARINEN, P., TOLONEN, K. and SOVERI, J., 1980, Distribution of trace metals and sulfur in the surface peat of Finnish raised bogs, Proceedings of the 6th International Peat Congress, Duluth, Minnesota, pp. 645–648. PARKASH, A.S. and BROWN, R.A.S., 1976, Use of peat and coal for recovering zirconium from solution, Canadian Mining and Metallurgy Bulletin, 69, 59–64. POPE, G.G., 1984, Pore structure and size for lignites, Fuel, 63, 1681–1685. RICE, N.M. and IKWUE, A.E., 1990, Treatment of solvent extraction raffinates with lignite as an alternative to activated carbon, Transactions of the Institute of Mining and Metallurgy, Section C—Mineral Processing and Extractive Metallurgy, 99, 131–136. SCHNITZER, M., 1978, Soil organic matter. In: Soil Organic Matter, SCHNITZER, M. and KHAN, S.U. (Eds), Elsevier, New York, pp. 47–52. SCHNITZER, M. and KHAN, S.U., 1972, Humic Substances in the Environment, Marcel Dekker, New York. SCHWARTZ, W.A., 1968, Removal of calcium and magnesium from waste water using peat, US Dept. Int. F.W.P.C.A. SMITH, E.F., MCCARTHY, P. and MARK, H.B., JR, 1976, Ammoniated and acid treated peat for metal ion exchange, J. Environ. Sci. Health—Environ. Sci. Eng., A11, 727–734. SMITH, E.F., MCCARTHY, P., Yu, T.C. and MARK, H.B., JR, 1977, Sulfuric acid treatment of peat for cation exchange, Journal of the Water Pollution Control Federation, 49, 633–638. SOUCI, S.W., 1938, The chemical characteristics and analytical investigation of peat, Kolloid K, 82, 87–99. SURAKKA, S. and KAMPPI, A., 1971, Reduction of BOD and other factors using peat, SUO, 22, 51–57. TUMMAVUORI, J. and AHO, M., 1980, On the ion exchange properties of peat—I. The adsorption of some divalent metal ions (Mn2+, Co 2+, Cu2+, Zn2+, Cd2+ and Pb2+) on the peat, SUO, 31, 45–51 and 79–83. VAN DIJK, H., 1971, Reaction of metal ions with humic acids, Geoderma, 5, 53–67. VINKLER, P., LAKATOS, B. and MEISEL, J., 1976, Infrared spectroscopic investigations of humic substances and their metal complexes, Geoderma, 15, 231–242. WEBER, W.J., JR and MORRIS, J.C., 1965, Adsorption of surfactants on carbon, Journal of the Water Pollution Control Federation, 37, 425–437. WIEDER, R.K., 1990, Metal cation binding to sphagnum peat and sawdust: relation to wetland treatment of metal polluted waters, Water, Air and Soil Pollution, 53, 391–400. WILSON, M.A., VERHEYEN, T.V., VASSALLO, A.M., HILL, R.S. and PERRY, G.J., 1987, Selective loss of carbohydrates from plant remains during coalification, Organic Geochemistry, 11, 265–269. WOLF, A. et al., 1977, Resistance of lignites to fungal attack, Chemosphere, 5, 207–213. ZHIPEI, Z., JUNLU, Y., ZENGHUI, W. and PRYA, C., 1984, A preliminary study of the removal of Pb2+, Cd2+, Zn2+, Ni 2+ and Cr6+ from wastewater with several Chinese peats, Proceedings of the 7th International Peat Congress, Dublin, Ireland, 3,147–152.

220

10

Biosorption: The Future C.F.FORSTER and D.A.J.WASE

Introduction Although metals are disseminated into the environment naturally by both geological and biological activity, human activity today produces a greater input. The toxicity of many of these metals is well known (Chapter 1); indeed, one of the earliest reports associating human disease with metals is attributed to Hippocrates in 370 BC. It is, therefore, not surprising that there has been considerable effort to develop technologies for reducing metal emissions. A significant proportion of these emissions are wastewaters arising from industrial concerns which can range from heavy industry, such as metal processing, to the production of fine chemicals. There can also be very wide variations in the concentration of metals which require treatment. Sorption has been one of the processes on which much attention has been focused. In general terms, activated carbon must be thought of as being a most effective sorbent and, as such, its performance in removing metal ions has been examined widely. For example, Wilczak and Keinath (1993) have studied the kinetics of the sorption and desorption of copper and lead while Perez-Candela et al. (1995) have reported the efficiencies of nine different carbons (both granular and powdered) for the removal of chromium. Metal recovery and sorbent regeneration are also of significance, and Reed et al. (1995) have evaluated a range of different regeneration schemes for a granular activated carbon/lead system. Taken overall the performances imply that, all other factors being equal, activated carbon should be considered the most appropriate choice for a sorption-based treatment of metal-polluted wastewater. However, despite the excellence of its performance, activated carbon is expensive to use and, as such, cannot be thought of as a truly viable option in many parts of the world. Because of this, attention has turned to the sorptive capabilities of biological materials and the preceding chapters have illustrated how wide a range of materials has been evaluated. Without a doubt the most successful and the most rigorously scrutinised biocompound is peat, either in its natural state (Viraraghaven and Dronamraju, 1993; Allen et al., 1994) or in a modified form (Davila et al., 1992) and there are valid arguments for suggesting that if biosorption is to be considered commercially then 221


peat sorption is the process which ought to be developed. Indeed, as has been shown in Chapter 9, designs for large-scale peat-sorption processes are available so that the technology could be utilised immediately. However, peat is a finite resource and, if a truly environmental code were to be followed, it would not be considered as a process option. Futhermore, in many countries there is a need for metal removal technologies but there are no peat reserves. Thus, the development of other forms of biosorption is essential. Predicting the future for biosorption, as a commercially used process, really requires the use of the proverbial crystal ball. In the absence of this technology, it is suggested that perhaps three variants will emerge as dominant options. These would be based on: • algae • fungi • wastes. What also needs to be judged is how well the various sorbents, which have been evaluated in laboratories, perform in scaled-up reactors.

Algal biosorption Any consideration of the use of algae as biosorbents must assess both micro and macro forms (Chapters 2, 6 and 8). At the same time there needs to be an evaluation of the process configurations which are required. In particular, there is a need to define how the wastewaters and the sorbents are contacted and how the clean liquor is obtained after the treatment stage. With microalgae there appear to be two options: either to use a stirred tank contactor in conjunction with downstream processing techniques, such as flotation, to separate the biomass from the clean liquors, or to immobilise the algae so that they can be used in a fixed or expanded bed reactor. Both techniques have been used. A non-immobilised biomass, Chlorella vulgaris, in conjunction with froth flotation, was able to remove over 80% of the silver in a simulated photographic wastewater (Cordery et al., 1994). AlgaSORB™, as has been discussed in Chapter 2, is Chlorella in an immobilised form and field tests on mercury-contaminated ground water indicate that it could compete economically with ion-exchange resins (Barkley, 1991). With the macroalgae much of the attention has focused on marine algae and, after initial screening studies, on ways in which the structure of the biomass can be strengthened using, for example, cross-linking with formaldehyde (Holan et al., 1993) or glutaraldehyde (Leusch et al., 1995). Trials with cadmium have shown that a formaldehyde-linked sorbent outperfomed a commercial ion-exchange resin (Holan et al., 1993).

Fungal biosorption Although, as has been shown in Chapters 4 and 6, considerable effort has been expended in the evaluation of the uptake of metals by fungi, the assessment of their potential at the large-scale/industrial level is very restricted. This is despite the fact 222

Biosorption: The Future

that the procedures for the immobilisation of the fungal cells either in/on reticulated foam (Kiff and Little, 1986) or within plastics/resins (Tobin et al., 1993) are established and fungi have efficient metal-binding characteristics. For example, a recent survey has shown that fungal mycelia can bind metals up to 25% of the biomass dry weight (Volesky, 1994).

Bio-wastes Within most countries a significant amount of biological waste will be generated each year. This can be produced by agriculture, for example wood (Bryant et al., 1992) and waste linseed fibre (Taylor et al., 1994); by the fine chemical industry; from seaweed residues after the extraction of agar (Fernandez et al., 1996b); or by nature (the pods produced by Albizia lebbek (Fernandez et al., 1996a) being a good example). All of these have the potential to act as biosorbents, and many of them have been examined in this role (Chapters 7 and 8). However, little has been done about examining their potential beyond laboratory scale. Essentially, the main requirements of an industrial sorption system will be that the sorbent: • can be used as a fixed or expanded bed • will not cause an excessive pressure drop across that bed. This probably means that most of the available bio-wastes will require some degree of pretreatment such as drying, comminution, sizing, pelletising or even chemical modifications similar to those described for algae. Any chemical treatment of a biowaste should be aimed at: • obtaining a structure suitable for use in a bed reactor • enhancing the number of metal-specific binding sites. Fourest et al. (1996), for example, have highlighted the importance of carboxyl groups, hence an oxidative step might be worth considering. Similarly, other studies into surface modification have focused on producing groupings which have established binding/chelating properties. For example, N-carboxymethyl groups constitute the active sites of several commercial ion-exchange resins. Therefore, Brezny et al. (1988) produced derivitives of lignins obtained from cellulose pulping, which contained these groupings. Rusina et al. (1977) have also applied the Mannich reaction, using amino acids as the amino components, together with formaldehyde, to furnish chelate-forming structures in mononuclear phenols. Thus, lignins were modified by • • • •

o-carboxymethoxylation (CM-lignins) the Mannich reaction with glycine (Gly-lignins) the Mannich reaction with iminodiacetic acid (IDA-lignins) carboxymethylation of Gly-lignins (Gly-CM-lignins).

While the starting lignins were almost inactive, the modified lignins exhibited considerable ion-exchange ability, with the highest capacities being given by Gly-CM and IDA lignins (up to 0.6 mmol/g). Rusina’s group also found the o-carboxymethyl groups were centres of non-specific binding and of lower affinities towards Cu2+, Cd2+ and Zn 2+ ions, whereas for the same ions, n-carboxymethyl groups showed 223


preferential adsorption in the order CuCu2+>Ni2+>Zn2+>Cd2+>Mg2+,whereas for 2+ 2+ dithiocarbamate wood it was Hg >Cu > >Zn2+>Cd2+>Ni2+>Mg2+=0. The main focus of the Biocomposites Centre of the University of Wales has been on combining one-step modifications with a complete characterisation of the derivative. For example, using the packed column method, succinated wood has been shown to have a cation exchange capacity of 0.55 meq/g. Although this is an order of magnitude lower than typical synthetic resins, the modified wood fibres are quicker to adsorb metal ions by virtue of their higher specific surface area and

Table 10.1 Metal binding (mmol/g adsorbent) to wood derivatives in relation to a commercial dithizone resin (Morita et al., 1987)

PEI, polyethyleneimine; DTC, dithiocarbamate.

224


Figure 10.1 Copper-sodium exchange by succinated wood fibre (Suzuki, 1991)

fibrous structure (Suzuki, 1991). The packed column method also allows the demonstration of true ion exchange as shown in Figure 10.1. Here, copper sulphate solution containing 1000 mg/l Cu 2+ ion was passed through a known weight of succinated wood fibre. The concentrations of sodium and copper in the effluent as well as the conductivity were monitored. The results clearly show that sodium was displaced by copper until the fibre was saturated, at which point copper was detected in the effluent.

The future In view of these results, the low numbers of industrial plants based on this technology must be of considerable concern. Why are there so few? The sorbents are effective, the underlying technology is sound and there is an environmental awareness within industry (to say nothing of the persuasiveness of legal sanctions). So where does this leave biosorption/wastewater treatment as an embryonic industry? Not all the companies which generate metal-polluted wastewaters will have the capability, or indeed the interest, to do anything other than the basic treatment to comply with the legalities. What can therefore be envisaged is a series of specialist, centralised facilities which could accept used sorbent, possibly in exchange for clean material. Such facilities would be capable of regenerating or processing the metal-loaded sorbents and then converting the recovered metal into a re-usable form (Figure 10.2). Such a development would mean that there would need to be a significant input of metallurgical skills which would then be integrated with the established base of sorption and wastewater treatment. The basis of this concept, which is certainly credible, is not new. Earlier workers have advocated 225


Figure 10.2 Schematic diagram illustrating the ‘Metalsplex’ concept

centralised, multidisciplinary facilities for the reclamation/recycling of wastes without success (Porteus, 1969; Forster, 1974; Forster and Hughes, 1974). However, the political climate for environmental improvement in the 1990s is different from that in the 1970s, a point endorsed by a recent report which outlined a strategy for centralised anaerobic digestion in the UK (Dagnal, 1995). Therefore, it is pertinent to consider what might be needed within a metal recovery centre. The processing of sorbents which had been used to treat wastewaters containing a single metal ought not to pose any significant problem. The use of acids to regenerate sorbents is well established (Philip et al., 1995; Aldor et al., 1995), as is the use of electrochemical techniques to recover metals from solution (Campbell et al., 1994). Alternatively, if the original sorbent had been based on a waste product and could be viewed as ‘disposable’, incineration could be used to produce a metal-rich slag. Handling used sorbents from the treatment of multi-metal wastewaters would pose some more interesting and complex questions. Could sequential sorption be used, with the process being based on more than one material, each of which was specific for a particular metal? Alternatively, would a differential regeneration of a single sorbent, using a sequence of different eluants to remove different metals, be more appropriate?

Conclusions Even when technologies are sound, costs are of considerable importance in determining the viability of a process or concept (Chapter 6). A recent review (Eccles, 1995) has produced a precise appraisal not only of the costs of various 226


biosorption options but also of the overall concepts of biosorption. The results were encouraging. For example, the total costs associated with metal removal, sorbent regeneration and metal recovery for the the AMT-Bioclaim process, which is based on a granular biomass (Bacillus subtilis), were estimated as being only 45% of those of a chemical precipitation process. Thus, the basic conclusion reached by Eccles was that the costings were sufficiently attractive to merit further effort, in particular with a multi-disciplinary approach.

References ALDOR, I., FOUREST, E. and VOLESKY, B., 1995, Desorption of cadmium from algal biosorbent, Canadian Journal of Chemical Engineering, 73, 516–522. ALLEN, S.J., MURRAY, M., BROWN, P. and FLYNN, O., 1994, Peat as an adsorbent for dyestuffs and metals in wastewater, Resources, Conservation and Recycling, 11, 25–39. BARKLEY, N.P., 1991, Extraction of mercury from groundwater using immobilized algae, Journal of the Air and Waste Management Association, 41, 1387–1393. BREZNY, R., PAZNER, L., MICKO, M. and UHRIN, D., 1988, The ion-exchanging lignin derivatives prepared by Mannich reactions with amino acids, Holzforschung, 42, 369–373. BRYANT, P.S., PETERSON, J.N., LEE, J.M. and BROWNS, T.M., 1992, Sorption of heavy metals by untreated red fir sawdust, Applied Biochemistry and Biotechnology, 92, 777– 788. CAMPBELL, D.A., DALRYMPLE, I.M., SUNDERLAND, J.G. and TILSTON, D., 1994, The electrochemical recovery of metals from effluent and process streams, Resource, Conservation and Recycling, 10, 25–33. CORDERY, J., WILLS, A.J., ATKINSON, K. and WILLS, B.A., 1994, Extraction and recovery of silver from low-grade liquors using microalgae, Minerals Engineering, 7, 1003–1015. DAGNAL, S., 1995, UK strategy for centralised anaerobic digestion, Bioresource Technology, 52, 275–280. DAVILA, J.S., MATOS, C.M. and CAVALCANTI, M.R., 1992, Heavy metals removal from wastewater by using activated peat, Science and Technology, 26, 2309–2312. ECCLES, H., 1995, Removal of heavy metals from effluent streams—why select a biological process? International Biodeterioration and Biodegradation, 35, 5–16. FERNANDEZ, N., CHACIN, E., GARCIA, C., ALSTRE, N., LEAL, F. and FORSTER, C. F., 1996a, Pods from Albizia lebbek as a novel water softening biosorbent, Environmental Technology, 17, 541–546. FERNANDEZ, N., CHACIN, E., GUTIERREZ, E., ALSTRE, N., LLAMOZA, B. and FORSTER, C.F., 1996b, The adsorption of lauryl benzene sulphonate on algal residue, Bioresource Technology, 54, 111–115. FORSTER, C.F., 1974, Bioplex, a recycling concept, Water and Waste Treatment, June, 26. FORSTER, C.F. and HUGHES, D.E., 1974, Recovery of organic matter from wastes and its effect on the environment, Paper presented at the International Public Works Conference, Montreal, Canada. FOUREST, E., SERRE, A. and Roux, J.C., 1996, Contribution of carboxyl groups to heavy metal binding sites in fungal wall, Toxicological and Environmental Chemistry, 54, 1– 10. HAYATSU, H., 1992, Cellulose bearing covalently linked copper phthalocyanine trisulphonate as an adsorbent selective for polycyclic compounds and its use in studies of environmental mutagens and carcinogens, Journal of Chromatography, 597, 37–56. HOLAN, Z.R., VOLESKY, B. and PRSETYO, I., 1993, Biosorption of cadmium by marine algae, Biotechnology and Bioengineering, 41, 819–825. KIFF, R.J. and LITTLE, D.R., 1986, Biosorption of heavy metals by immobilised fungal 227

Biosorbents for Metal lons biomass, In: Immobilisation of Ions by Bio-sorption, ECCLES, H. and HUNT, S. (Eds), Ellis Horwood, Chichester, pp. 71–80. LEUSCH, A., HOLAN, Z.R. and VOLESKY, B., 1995, Biosorption of heavy metals (Cu,Cd, Ni, Pb, Zn) by chemically reinforced biomass of marine algae, Journal of Chemical Technology and Biotechnology, 62, 279–288. MEYER, V., CARLSSON, F.H. H. and] OELLERMANN, R.A., 1992, Decolourisation of textile effluents using a low-cost natural adsorbent material, Water Science and Technology, 26, 1205–1211. MORITA, M., HIGUCH, M. and SAKATA, I., 1987, Binding of heavy metal ions by chemically modified woods, Journal of Applied Polymer Science, 34, 1013–1023. PEREZ-CANDELA, M., MARTIN-MARTINEZ, J.M. and TORREGROSA-MACIA, R., 1995, Chromium (VI) removal with activated carbon, Water Research, 29, 2174–2180. PHILIP, L., IYENGAR, L. and VENKOBACHAR, C., 1995, Biosorption of copper(II) by Pseudomonas aeruginosa, International Journal of Environment and Pollution, 5, 92–99. PORTEUS, A., 1969, The recovery of industrial ethanol from paper in waste, Chemistry and Industry, 1763. REED, B.E., ROBERTSON, J. and JAMIL, M., 1995, Regeneration of granular activated carbon (GAC) columns used for removal of lead, Journal of Environmental Engineering, 121, 653–662. RUSINA, M.N., TIMAKOVE, L.M., YAROSHENKO, G.F., BRANZBURG, M.Z. and TEMKINA, V.Y., 1977, Synthesis and complexing properties of 3-[bis(carboxymethyl) amino]methyl-2, 4-dihydroxy benzoic acid, Zh. Obshch. Khim., 47, 2112–2117. SUEMITSU, R., VENISHI, R., AKASHI, I. and NAKANO, M., 1986, The use of dyestufftreated rice hulls for removal of heavy metals from wastewater, Journal of Applied Polymer Science, 31, 75–83. SUZUKI, M., 1991, Application of fiber adsorbents in water treatment, Water Science and Technology, 23, 1649–1658. TAYLOR, D.T., EDYVEAN, R.G.J. and JOHNSON, D.J., 1994, Investigations into the use of linseed flax straw for the removal of metals from wastewater, Proceedings of the 2nd International Symposium on Environmental Biotechnology, Institution of Chemical Engineers, Rugby, 116–118. TOBIN, J.M., L’HOMME, B. and Roux, J.C., 1993, Immobilisation protocols and effects of cadmium uptake by Rhizopus arrhizus biosorbents, Biotechnology Techniques, 7, 739– 744. VIRARAGHAVEN, T. and DRONAMRAJU, M.M., 1993, Removal of copper, nickel and zinc from wastewater by adsorption using peat, Journal of Environmental Science and Health, Part A: Environmental Science and Engineering, 28, 1261–1276. VOLESKY, B., 1994, Advances in biosorption of metals—selection of biomass types, FEMS Microbiology Reviews, 14, 291–302. WILCZAK, A. and KEINATH, T.M., 1993, Kinetics of sorption and desorption of copper (II) and lead (II) on activated carbon, Water Environment Research, 65, 238–244.

228

Index

Absidia orchidis 71 acacia bark 149, 150, 151 actinides 96–7, 109–10 radium as daughter product 106–7 see also radium see also thorium; uranium activated alumina 146 activated carbon as biosorbent 8, 150, 221 for chromium 8, 152 for copper 149, 221 for lead 221 for thorium 107 for uranium 99 formation of 186–7 novel sources 8, 147–8 peat and lignite 193 see also carbon activated sludge see sludge active uptake see bioaccumulation adsorption isotherms BET (Brunauer-Emmett-Teller) 68 Freundlich 12, 13, 51 Langmuir 12–13, 68 and bacterial surfaces 51 and metal binding mechanisms 48, 49–50, 51 Advanced Mineral Technologies 131 age of carbon 186, 187 of fungal cells 71 of sludge and efficiency 117–19 air lift reactors 124, 128 Alcaligenes eutrophus 87, 96 algae 11–37 and anion concentration 22 as biosorbent 222 for lead 151 mechanisms 12–20 and cation concentration 22 cell size 20 cell walls 19–22 costs of 23–6 commercial systems 31–3, 177–8 crosslinked biomass 18 cyanobacteria see cyanobacteria factors affecting 20–3 immobilised algae 18, 23, 26–7

lyophilized biomass 18 macroalgae see seaweeds and macroalgae microalgae see microalgae and pH of external medium 20–2 physiological state 23 processes 27–30 production of 23–6 and temperature 23 in waste stabilisation ponds 120–1 AlgaSORB 26, 32, 131–2, 222 alginate 27, 167, 168, 171 for algae immobilisation 26 time for biosorption 174 alginate beads 23 Alicaligenes sp. 90, 92 almond shells 148 aluminium biosorption by bacteria 57 by living microalgae 14 by lyophilized (algal) biomass 18 by waste materials 169 effect of pH on 20 ionic competition effects with uranium 103–5 americium 96 AMT-Bioclaim process 32, 131, 177, 227 Anabaena cylindrica 16 Anabaena variabilis 16 Anacystis nidulans 16, 18 anions and algal biosorption 22 sorption by bacteria 51–2 Ankiistrodesmus sp. 14 antimony 189 Aphanocapsa sp. 16 apple waste 169 apricot stones 148 aquatic life shellfish see shellfish toxicities of heavy metals for 2, 3, 4 water quality standards for 185 arsenic 52, 76, 154, 187 Arthrobacter sp. 57, 90 Ascophyllwn nodosum for alginate production 167 in commercial systems 32 crosslinked biomass 18

Index desorption 29 metals sorbed 17, 18, 171 in multi-metal systems 17, 19, 22, 171–2 Aspergillus niger and biosorption capacity 70, 99, 107 and cell age 71 and copper 149 and gold 69 and mercury 169 pretreatment of 75 and silver 69 and thorium 107 and uranium 99 Aspergillus oryzae biosorbent capacity 70 and copper 149 and lead 151 and nickel 150 pretreatment of 75 Aspergillus terreus 107 Asterionella formosa 14 Bacillus licheniformis 42, 49, 57 Bacillus sp. 53, 54, 133–4 Bacillus subtilis cell walls of 44, 45, 48, 49, 52 in industrial systems 177, 227 maximum complexation capacities 54, 57 stability constants for 53 bacteria sorptive processes 39–66, 90–6 anions 51–2 cations 46–51 cell walls 40–1, 46–7 modelling 57–8 surfaces 39–46 bagasse 8, 149, 150, 151 competitive adsorption with 154 sugar cane 152 banana pith 8, 149 barium 77 bark 8, 146, 169, 187 acacia bark 149, 150, 151 pretreatment 175 techtona see Techtona grandis time for biosorption 174 batch processes 141–63 BDST (bed depth service time) theory 203–5 beads and granules 78, 88, 129–33 AlgaSORB 26, 32, 131–2, 222 AMT-Bioclaim process 32, 131, 177, 227 Bio-fix 32, 33, 129, 130, 178 Chitopearl 136 polyacrylamide gel beads 132, 133 beryllium 20, 154 BET (Brunauer-Emmett-Teller) isotherm 68 binding constants 52–7 from Scatchard plots 13, 15, 17 see also biosorptive capacity; capacity; maximum complexation capacity; uptake capacity Bio-fix 32, 33, 129, 130, 178 bio-polymers 135–6 Bio-recovery Systems Inc see AlgaSORB bio-wastes 8–9, 223–5

230

bioaccumulation by algae 12 definitions 11, 67–8, 87, 165 mechanisms 96 biofilm systems 126, 133–5 biofilms 43 in biological filters 119 biological filters 119 biosorption definitions 11, 87 future of 221–8 as passive uptake 12, 67, 165 biosorptive capacity 70–1 see also binding constants; capacity; maximum complexation capacity; uptake capacity Biotechna 25 biotechnology 58–9 black liquor 146–7 Black List 4–5 blue-green algae see cyanobacteria breakthrough curves 197–200 for aluminium 104 BDST theory for 203–5 in EBRT plots 208–9 LUB theory for 201–3 for uranium 101, 102 brine see salt water Brunauer-Emmett-Teller isotherm (BET) 68 bryophytes 168 B.V.Sorbex 32 cadmium biosorption 153–4 binding constants 55, 56 and cell wall binding 47, 48, 50 by chitin 170 by crosslinked algal biomass 18 by cyanobacteria 16 effect of pH on 20 by fungi 69, 70, 77 by immobilised algal biomass 18 low-cost biosorbents for 154, 187 by lyophilized algal biomass 18 by macroalgal biomass 17, 19, 22, 171, 175 in marine conditions 22 maximum complexation capacities 54 by microalgae 14, 15 in multi-metal systems 17, 19, 22, 171–3 by peat 193 by seaweeds 166, 167, 171 stability constants for 53 by waste materials 169 regulations controlling 4–5 as secondary pollutant 6 toxicity of 2 caesium 69 calcium binding constants 56 binding to bacteria 42, 49, 50 and fungi 73, 77, 79 low-cost biosorbents for 147 and peat 189, 193 Caloxthrix Crustacea 16 capacity

Index for chromium 142, 152 for copper 142, 149 for lead 142, 151 for nickel 142, 150 of peat, lignite and coal 142, 191 see also binding constants; biosorptive capacity; maximum complexation capacity; uptake capacity capsules 42, 53 carbon dioxide 25 carbon-based sorbents activated carbon see activated carbon as biosorbent for chromium 149, 221 in industrial scale systems 186 in pulsed bed sorbers 209–10 types and history of 186–7 see also charcoal; coal; lignite; low-cost biosorbents; peat carboxyl groups 19, 48–9 cations 22, 46–51 cell envelope, bacterial 39–43 cell size, algal 20 cell walls and surfaces algal 19–22 bacterial 39–43, 46–7 charge on 43–6 metal binding to 46–52 charcoal 186 charge on cell surfaces 43–6 Chemelec cells 8 chitin 149, 169–70, 195 in fungal cells 79, 90 see also chitosan Chitopearl 136 chitosan 135–6, 169–70, 195 see also chitin Chlamydomonas reinhardtii 14, 18 Chlamydomonas sp. 14 Chlorella emersonii 14 Chlorella homosphaera 14 Chlorella pyrenoidosa 18 Chlorella regularis 14, 21, 132, 177 Chlorella saccarophila 18 Chlorella salina desorption solutions for 29 metals adsorbed 14, 18 cobalt 12, 13, 21 Chlorella sp. 14, 22, 24, 27 immobilisation system see AlgaSORB mixture with Scenedesmus 15 Chlorella vulgaris adsorption/desorption cycles 177 desorption solutions for 29 in immobilised state 132 lyophilized 18 metals adsorbed 14, 18 copper 149 gold 90 silver 222 optimum temperature 174 time for biosorption 174 Chondrus crispus 17, 18, 32 chromates 16, 51–2 chromium biosorption

by activated carbon 8, 149–51, 152 by carbons 8, 149, 221 by chitin 170 by fungi 70 by low-cost biosorbents 147, 154 by microalgae 15 by peat 189 by seaweeds 167 by Streptomyces sp. 169 by waste materials 169 capacity of peat for 142 of various biosorbents for 152 effect of pH on 20 toxicity of 2, 51–2, 150 see also chromates Citrobacter sp. 87, 92, 96 Cladophora 12, 23–4 clay 148, 149, 150, 151 coal 146, 148, 187, 193 cobalt binding constants 55, 56 biosorption 153–4 by algae 12, 13 by chitin 170 by Chlorella salina 21 effect of pH on 20, 21 by fungi 70 by immobilised algae 18, 27 by macroalgal biomass 17 by microalgae 14, 15 by seaweeds 167 by Streptomyces sp. 169 maximum complexation capacities 54 stability constants for 53 toxicity of 2 coconuts coir 147, 148, 151 husks 150–1, 152, 169 shells 148, 150, 152, 187 Codium taylori 17 coffee 169 coincidental sorption systems 116–21 coir see coconuts columns for continuous processes 195–216 combustion of algal biomass 30 commercial systems 30–1, 115–40 for algae 23–6, 30–3, 177–8 AlgaSORB 26, 32, 131–2, 222 AMT-Bioclaim process 32, 131, 177, 227 batch processes 141–63 Bio-fix 32, 33, 129, 130, 178 coincidental 116–21 columns for 195–216 continuous processes 183–220 current methods 7, 186 design of 121–2 Filtrasorb 400 8 immobilisation matrices 128–36 see also immobilised biomass processes see processes reactors for see reactors specific 121–3 time in 174 see also engineering applications; industrial scale systems

231

Index competition and algal biomass 22 and cell wall sites 49–50 and fungal biosorption 76–7 with low-cost biosorbents 154, 167 and peat sorption 193 in radium uptake 107–8 in uranium recovery 103–6 see also selectivity complexation capacity see maximum complexation capacity compost 152 conventional stirred tank reactors (CSTR) 123–5 copper binding constants 55, 56 binding sites 50 biosorption 148 by activated carbon 149, 221 by chitin 170 by crosslinked algae 18 effect of pH on 20 by fungi 69, 70, 77 by immobilised algae 18, 27 by lyophilized algal 18 by macroalgal biomass 17, 19, 22, 167 by microalgae 14 in multi-metal systems 17, 19, 22, 171–3, 193 by peat 141–3, 189, 190–1, 193 by Saccharomyces cerevisiae 71 by Streptomyces sp. 169 capacity 149 of fungi for 70 maximum complexation capacities 54, 56–7 of peat for 142 desorption of 121 stability constants for 53 toxicity of 2–3 corn cobs 8, 187 costs of activated carbon 150 of algal biomass 23–6 commercial systems 31–3, 177–8 of biosorption 226–7 in fixed bed sorber design 207–8 of fungal biomass 80 low-cost biosorbents see low-cost biosorbents in sorption/desorption cycles 176–7 counter-current sorption 210–13 Cricosphaere elongata 14 crosslinked biomass 18, 125 see also immobilised biomass CSTR (conventional stirred tank reactors) 123–5 cyanobacteria 16, 19–20, 23–4 cypress leaves 169 Datura inoxia (solonaceae) 168 dead biomass advantages of 23, 73, 122–3 algal 23 bacterial 23 cells 11, 23 fungi 67–8, 73–7 reactors for 126 dealginate 167, 168, 171, 172 desorption

232

cycled with sorption 176–7 from seaweeds 174 systems for algae biosorption 29–30 see also elution; recovery; regeneration Desulfovibrio desulfuricans 96 Desulfovibrio sp. 92, 96 differential mass balance 214–15 diffusion processes 40, 99, 100 domestic wastewater, golf-green type filters for 197 downflow systems 198–201, 202 drinking water 2–3, 185, 186 duckweed see Lemna minor (duckweed) Dunaliella tertiolecta 14 Durvillea potatorum 167 dyes removal with bio-waste 224 treatment with 147, 150 EA (Environment Agency) 5 EBRT (empty bed residence time) theory 205–9 Ecklonia maxima 167, 177 Eichornia crassipes (water hyacinth) 168 elution 176–7 of fungal biomass 77–8 for radium 108 for uranium 99, 101–6 see also desorption; recovery; regeneration engineering applications of biosorption 88–9 of uranium biosorption 101–3 see also commercial systems; industrial scale systems environment, metal damage to 1–5 Environment Agency (EA) 5 Environmental Protection Act and regulations 5 EPS (extracellular polysaccharide) 42, 43 maximum complexation capacities 54 and metal binding 44, 48–9 modification 45 production 51 stability constants for 53 Escherichia coli 40 Escherichia coli K12 41 Euglena sp. 14 European Union 4–5 water quality standards 185 exhaustion curve 198 exopolymers 57 expanded bed reactors 124, 127, 128 extracellular polysaccharide see EPS (extracellular polysaccharide) filters 116 biological 119 golf-green type 197 microfungal 133 moving mat filters 196–7 Filtrasorb 400 8 fish see aquatic life fixed film processes see biofilms; filters fixed-bed sorption 197–209 see also packed bed reactors fluid phase mass transfer coefficient 215

Index fluidised bed systems reactors 29, 124, 127–8, 213–16 for suspended solids 31 fly-ash 149, 151 Fragilaria crotonensis 14 fresh water, water quality standards for 185 Freundlich adsorption isotherm 12, 13 and bacterial surfaces 51 fruit pits 8 fruit stones 148 Fucus sp. 171 Fucus vesiculosus 17, 18 fungi 67–85, 222–3 biosorption by dead cells 73–7 by immobilised fungal mass 78 by living cells 69–73 mechanisms for 78–9 chitin/chitosan from 79, 90, 169 microfungal filters 133 regeneration 77–8 surface charge on 72 uptake modes 67–8 see also Saccharomyces cerevisiae; yeast future of biosorption 221–8 gallium 46, 170 Ganoderma lucidum 71, 149 gold biosorption by cyanobacteria 16 by fungi 70 by immobilised algae 18, 27 by macroalgal biomass 17, 167, 171 by microalgae 14 mechanisms 90 desorption of 121 elutants for 176–7 recovery 131 golf-green type filters 197 Gram-negative walls 19, 40–1 metal binding to 46–7 Gram-positive walls 40–1 metal binding to 46–7 granules see beads and granules Grey List 4

in stirred tank reactors 125 for uranium uptake 99 see also immobilisation matrices indium 170 industrial scale systems 177–8 AMT-Bioclaim process 131, 177 Bio-fix 32, 33, 129, 130, 178 current methods 7, 186, 225 future of 225–6 processes see processes see also commercial systems; engineering applications industry effluent from 184 metal finishing 5, 6 wastewater from 221 see also mines Integrated Pollution Control (IPC) 5 intracellular uptake see bioaccumulation ion-exchange in fungal sorption 79 in peat sorption 190–1, 194 ion-exchange resins 223 compared to seaweeds 167 thorium uptake capacity 107 uranium uptake of 99 ionic competition 103–6 see also competition IPC see Integrated Pollution Control (IPC) iron binding constants 56 biosorption by chitin 170 effect of pH on 20 low-cost biosorbents for 154 by peat 189 by seaweeds 167 in wetlands 120 maximum complexation capacities 54 stability constants for 53 iron hydroxide 148, 149, 150 isoelectric point 21 isotherms see adsorption isotherms jute 169

hair, human 187 Halimeda opuntia 17, 32 Halobacterium sp. 41 Hardwickia binata 169 history of metal usage 1 Hussong-Couplan system 196–7 hydroxide precipitation 7 Hymenomonas elongata 15

kinetics 12, 123–4 of fungal biosorption 69, 71 and non-viable biomass 126 of peat sorption 191 of radium biosorption 109 of thorium biosorption 106 see also mechanisms Klebsiella aerogenes 44, 53, 54, 57 Klebsiella pneumonia 50

imidiazol 19 immobilisation matrices 128–36 see also immobilised biomass immobilised biomass 88–9 algae 18, 23, 26–7 fungal 78 microbial 109, 110 for radium sorption 109, 110

lagoons see ponds Laminaria hyperborea 167 Langmuir adsorption isotherm 12–13, 68 and bacterial surfaces 51 lanthanides 96, 97, 109–10 lanthanum 47, 70, 77, 96 laurel bark 149, 150, 151 LC50 2, 3

233

Index lead binding constants 55, 56 biosorption 148–9 by activated carbon 221 by chitin 170 by crosslinked algae 18 by cyanobacteria 16 effect of pH on 20 by Fucus sp. 171 by fungi 69, 70, 77 by macroalgae 17, 167 by microalgae 14 by peat 189, 193 by sago waste 147 by Streptomyces sp. 169 capacity for 142, 147, 151 in mine wastewater 23–4, 31 recovery 156 toxicity of 3 leaf mould 152 Lemna minor (duckweed) 168 Lessonia flavicans 167 Lessonia nigrescens 167 light in algae cultivation 24 lignin 144, 146–7, 151, 223–4 lignite age of 187 ion-exchange properties 194 modification of 193–5 regeneration and recovery 216 sorption mechanisms 190–1 structure and properties 187, 188 linseed 147, 176, 223 lipopolysaccharide see LPS (lipopolysaccharide) living biomass algal 23 cell walls 23 compared to dead 23 fungi 67–8, 69–73 reactors for 123 living systems 58 see also ponds; wetlands loading of peat 189–90 low-cost biosorbents for batch processes 141–63 in continuous processes 183–220 from waste see waste see also peat LPS (lipopolysaccharide) 41 LUB (length of unused bed) theory 201–3 lyophilized biomass 18 macroalgae see seaweeds and macroalgae Macrocystis pyrifera 167 magnesium binding to bacteria 49, 50 biosorption 73, 77, 189 maize 152, 168, 169 manganese binding to bacteria 49 biosorption 153 by fungi 70 by immobilised algae 18, 27 by lyophilized algae 18 by microalgae 14, 15

234

by peat 189 by Streptomyces sp. 169 stability constants for 53 manganese dioxide 150 marine conditions see salt water mass transfer zone see MTZ (mass transfer zone) mass-transfer coefficients 213–16 maximum complexation capacity 54, 56–7 see also binding constants; biosorptive capacity; capacity; uptake capacity mechanisms of biosorption 183 for activated sludge 116–17 and bioaccumulation 89–96 by algae 12–20 by bacteria 90–6 by fungi 78–9 ionic competition effects 103–6 for metal binding to bacteria anions 51–2 cations 46–51 by organic matter 190–1 see also adsorption isotherms membrane-media system 156, 157 mercury BDST equation constants 204 biosorption by bark 169 breakthrough curves 198, 199 by chitin/chitosan 169, 170, 195 by Chlorella 222 by cyanobacteria 16 by fungi 70 by immobilised algae 27 by low-cost biosorbents 154, 187 by lyophilized (algal) biomass 18 in marine conditions 22 by microalgae 14, 167 by peat 189, 193 time for 192 desorption of 121 regulations controlling 4–5 toxicity of 3 metal finishing effluent 5, 6 Metalsplex concept 225–6 Methanospirillum hungatei 41, 42–3 microalgae 222 costs and production 24–5 desorption solutions for 29 immobilization of 26 in solution 14–15 Micrococcus luteus 44, 53, 54 Micrococcus sp. 53, 54 mines algae for 12, 23–4, 31 biofilms for 135 polymeric bead systems for 129, 130–1 uranium mine leachate 87, 103–6 mixed biomasses 177 modelling 57–8, 68–9 modification 175–6, 193–5 see also pretreatment molybdenum 14, 21 moss 8, 151 peat moss 146 Sphagnum moss peat 149, 150, 168

Index moving bed sorbers 210–13 moving mat filters 196–7 MTZ (mass transfer zone) 197–8 Mucor miehei 70, 76, 80, 129 Mucor mucedo 70 Mucor rouxii 70, 75 multi-metal systems 22, 171–3 and peat sorption 193 two-metals 17, 19 see also competition; selectivity nickel binding constants 55 binding to bacteria 49, 50 biosorption by chitin 170 by cyanobacteria 16 effect of pH on 20 by immobilised algae 18 by low-cost biosorbents 147 by microalgae 14, 15 in multi-metal systems 172, 173, 193 by peat 141–3, 148–9, 189, 193 by peat bed system 120 by seaweeds 167 by Streptomyces sp. 169 capacity of biosorbents for 142, 150 maximum complexation capacities 54 stability constants for 53 toxicity of 3 non-living cells see dead biomass non-viable biomass see dead biomass Nostoc calicola 18 Nostoc sp. 16 nuclear industry 96, 97 see also radionuclides; uranium nuts peanut husks 187 shells 8, 169 oil palm fibres 147, 149, 150, 151, 169 competitive adsorption with 154 time for biosorption 174 olive stones 148 operating lines 207–9 Oscillatoria, Red 16 Oscillatoria sp. 16, 23–4, 53 Oscillatoria woronichinii 16 packed bed reactors 124, 125–7 for algal biomass 26 media for biofilm growth 126 three-column 33, 129–30 see also fixed-bed sorption Padina gymnospora 17 palladium 91–2, 94, 195 palm press fibres 152 palm tree cobs 148 Palmaria palmata 17, 32 paper 148 parallel-column systems 200 partially converted shellfish waste (PCSFW) 135–6 see also chitin; chitosan

particle size 213 passive uptake see biosorption PCSFW (partially converted shellfish waste) 135–6 see also chitin; chitosan peach stones 148 peanut husks 187 Pearson’s classification 106 peat age of 187 as alternative to activated carbon 221–2 for batch processes 141–6 compared to others 146–54 practical aspects of use 155–8 beds for nickel removal 120 columns for 196–7 competitive adsorption with 154, 193 factors affecting 192–5 ion-exchange properties 190–1, 194 kinetics of sorption 191–2 metals sorbed by 188–95 modification 193–5 moss beads 194 peat bogs 189 regeneration and recovery 216 selectivity 189, 193 Sphagnum moss peat 149, 150 structure and properties 187–8 variability of 183, 188 Penicillium chrysogenwn 71, 74, 76 uptake of radionuclides 98, 99, 107–8 Penicillium Ci 99 Penicillium digitalum 74 Penicillium italicum 107 Penicillium sp. binding capacity 77 biosorptive capacity 70 and cell age 71 elutants for 77 pretreatment of 74 sorption of lead 151 sorption of uranium 76 Penicillium spinulosum 71, 74 pH and bacterial cell charge 44–5 and biosorption algal 20–2, 25 by chitin 169, 170 fungal 72, 76 by peat 191, 192 by radium 107, 109 and desorption 176–7 for hydroxide precipitation 7 and peat stability and efficiency 145 for seaweeds 173–4 phosphorus 197 pilot testing 101–6, 109 plants 168–9 see also algae platinum 46 Plectonema boryanum 16 plutonium 96 political climate 226 polyelectrolytes 45 polymers

235

Index beads and granules from see beads and granules bio-polymers 135–6 see also chitin; chitosan polysaccharides 11, 19, 79, 116 extracellular see EPS (extracellular polysaccharide) ponds for algae cultivation 24–5 for living systems 58 waste stabilisation ponds 120–1 for wastewater treatment 23–4 see also wetlands Porphyra tenera 17 pretreatment 175–6 of algal biomass 18 for bio-wastes 223 with dyes 147, 150 for fungal biomass 73, 74–5 see also modification processes activated sludge 116–19 for algal biosorption 27–30 batch 141–63 continuous 183–220 see also commercial systems; industrial scale systems production of algal biomass 23–6 proteins in cells 19 Pseudomonas aeruginosa 54, 57 Pseudomonas fluorescens 50, 57 uptake capacities 98, 99, 107, 108 Pseudomonas putida 53 Pseudomonas sp. 49, 57, 58, 90 maximum complexation capacities 54 Pseudomonas stutzeri 52 pulsed bed sorbers 209–10 radionuclides 11, 76, 79, 88 biosorption by fungi 69 see also actinides; radium; strontium; thorium; uranium radium 77, 87, 106–9, 110 raffination earth 148 rare earths see lanthanides rate of sorption 191–2 reactors 123–8 air lift 124, 128 for algae biosorption 28–9 comparisons 136 conventional stirred tank reactors (CSTR) 123–5 for cultivating algae 25 expanded bed 124, 127, 128 fluidised bed 124, 127–8, 213–16 packed bed reactors 124, 125–7 recovery 121 from algal biomass 29–30, 96 from bacteria 57 using Chemelec cells 8 from peat and lignite 155–6, 216 of uranium 96, 99, 101 ionic competition effects 103–6 pilot testing 101–6

236

see also desorption; elution; regeneration recycling see recovery; regeneration regeneration of activated carbon systems 221 of fungal biomass 77–8, 80 of peat and lignite 216 for radium 109, 110 of Rhizopus arrhizus 101 using three-column systems 129–30 see also desorption; elution; recovery regulations 4–5, 183, 185–6 reloading see regeneration Rhizoclonium 23–4 Rhizopus arrhizus binding sites 50 in biofilm system 134–5 biosorptive capacity 70 immobilisation matrix for 133 pretreatment of 75 and radium uptake 107, 109 and thorium 106 biosorption mechanism 90, 92 uptake capacities 107, 108 and uranium 76, 99, 100 biosorption mechanism 90, 91 ionic competition effects 103 reloading following elution 101 uptake capacity 98, 99 Rhizopus nigricans 70 Rhizopus oryzae 149 Rhizopus sp. 77 rice 149, 150, 151 bran 169 defatted 152, 153 hulls 8, 154 straw 169 rubber 187, 193 tyres 8, 148, 151 S-layers 42, 59 Saccharomyces cerevisiae 71, 75, 99, 134 see also yeast sago waste 147, 151, 157 salt bridging 48–9 salt water biosorption in 22, 175 chitin as biosorbent from brine 170 water quality standards for 185 sand filtration 187 Sargassum fluitans 17, 18 Sargassum natans 17, 29, 32, 171, 177 Sargassum sp. 171, 175 Sargassum vulgaris 17 sawdust 148, 152, 169 pine 187 Scatchard plots 13, 15, 17, 68–9 Scenedesmus obliquus 14, 19 Scenedesmus quadricauda 18, 132 Scenedesmus sp. 14 sea water see salt water seaweeds and macroalgae 17, 25–6, 166–8, 222 alginate see alginate biosorption

Index of cadmium 171 comparison with others 170–1 of gold 171 pH for 173–4 time for 174 immobilization of 26 pretreatment of 175 waste see dealginate selectivity 88 and algae biosorption 22, 29–30 and fungal biosorption 76–7 and lignins 224 in peat sorption 189, 193 in uranium recovery 101, 102 see also competition Selenastrum sp. 14 selenium 58, 154 series column system 200–1 sheaths 20, 42–3 shellfish PCSFW (partially converted shellfish waste) 135–6 as source of chitin 169, 195 see also chitin; chitosan Shewanella algae 51, 57, 96 shock loads 115 siderophores 56 affinity constants for 54 production from algae 11, 27 stability constants for 53 silicates 52 silver biosorption 91, 92 by cyanobacteria 16 by fungi 77 low-cost biosorbents for 154 by lyophilized algae 18 by macroalgal biomass 17 by peat 189 desorption of 121 toxicity of 4 single-column systems 198, 200 sludge activated process using 116–19 thorium uptake capacity 107 uranium uptake of 99 anaerobically digested 149, 150 as by-product of water treatment 7 solids 149, 150, 151 for zinc 153 slurry 148, 151 sodium 49, 56 solid phase mass transfer coefficient 215–16 solonaceae (Datura inoxia) 168 sorption, defined 39 sorption zone 198, 199 speciation 22 specificity, usefulness of 178 Sphagnum 168 Sphagnum moss peat 149, 150 Spirogyra 23–4 Spirulina platensis 16 Spirulina sp. 24

stabilisation ponds 120–1 see also pond; wetlands stability constants 53 see also binding constants Staphylococcus sp. 40 Stichococcus bacillaris 18, 19 stirred tank reactors (CSTR) 123–5 straw 146, 169 barley 187 pretreatment of 175–6 time for biosorption 174 Streptomyces albus 132 Streptomyces niveus 99, 107 Streptomyces sp. 169 Streptomyces viridochromogenes 132 strontium binding to bacteria 49, 50 biosorption bacterial 96 by fungi 69, 70 by macroalgal biomass 17 by yeast 90 sugar beet pulp 152 sugar cane bagasse 152 surface area, effect of peat sorption 192–3 surfaces of bacteria 39–43 charge on 43–6 of fungi, charge on 72 Synechococcus elongata 16 Synechococcus PCC 6301 16 Synechococcus sp. 20 Synechocystis PCC 6803 16 Synechocystis aquatilis 16 systems see commercial systems; industrial scale systems; processes tea 169 leaves of 8, 149, 151, 154 technetium 14, 16, 21 Techtona grandis 169, 175 bark 149, 150, 151 temperature 23, 77, 174 Thalassiosira rotula 14 thorium biosorption 106 by fungi 70, 78, 90, 92 from mine waters 87 uptake capacities 107, 108 radium as daughter product 106–7 three-column systems 33, 129–30 time 174 rate of sorption by peat 191–2 tin 4 titanium 189 titanium (IV) oxide 151 toxicity of metals 2–4, 122 chromium 2, 51–2, 150 treatment for biomass see pretreatment Trichoderma reesii 71 Trichoderma viride 71 two-metal system 17, 19 see also multi-metal systems tyres, waste 8, 148, 151

237

Index UASB anaerobic digestion process 128 upflow systems 201 uptake see biosorption uptake capacity for thorium 107, 108 for uranium 98–9 see also binding constants; biosorptive capacity; capacity; maximum complexation capacity uranium biosorption 23, 97–101 by chitin/chitosan 79, 147 by Chlorella 22, 27 by fungi 69, 70, 71, 77 by macroalgal biomass 17 by Penicillium 76 by cyanobacteria 16 engineering applications 101–6 from mine waters 87 by immobilised fungal mass 78 in marine conditions 22, 175 by microalgae 14 by peat 156 by Pseudomonas sp. 49 by R.arrhizus 76, 90, 91, 99, 100 breakthrough curves for 101, 102 elution 77, 99, 101, 176–7 oxides binding constants 56 effect of pH on 20 maximum complexation capacities 54 stability constants for 53 radium as daughter product 106–7 recovery 96, 99, 101–6 uptake capacities 99

treatment of 5–9 using algae 23 using wetlands 119–20, 194 see also mines water drinking water 2–3, 185, 186 salt water 22, 170, 175, 185 treatment of contaminated water 5–9 waste see wastewaters water quality standards 185 water hyacinth 168 Water Industry Act 5 water quality standards 185 wetlands 116, 119–20, 154 for removing metals from wastewaters 194 see also ponds wollastonite 151 wood 224–5 wood fines 187 wool 169

Vaucheria 17 Visa Tech Ltd see AMT-Bioclaim process

Zea mays (maize) 152, 168, 169 zinc binding constants 55, 56 binding to bacteria 49, 50 biosorption 151, 152, 153 by activated sludge 117–19 by crosslinked algae 18 by cyanobacteria 16 effect of pH on 20 by fungi 69, 70 by immobilised algae 18, 27 by lyophilized algae 18 by macroalgae 17, 19, 22 by microalgae 14, 15 in multi-metal systems 17, 19, 22, 171–3 by peat 189, 193 by seaweeds 167 by Streptomyces sp. 169 desorption of 121 stability constants for 53 toxicity of 4 zinc sulphide 154 zirconium 14, 16, 154, 189 Zoogloea ramigera 54, 57, 149, 174

walnut shells 169 washout 125 waste bio-wastes 8–9, 223–5 for biomass 129, 169 fungal 80 black liquor 146–7 as low-cost biosorbents 147–8, 157, 178 metal finishing effluent 5, 6 PCSFW (partially converted shellfish waste) 135–6 sago waste 147 seaweed (dealginate) 167, 168, 171, 172 tyres 8, 148, 151 water see wastewater waste stabilisation ponds 120–1 see also ponds wastewaters 184, 221 domestic 197 as source of biomass 107

238

Xanthan 53, 54 yeast as biosorbent 67 for copper 149 for lead 151 for nickel 150 for strontium 90 and cell age 71 immobilisation 131 see also fungi; Saccharomyces cerevisiae yttrium 91, 92, 95

Biosorbents for Metal Ions

Metal Ions in Fungi

Metal Ions and Neurodegenerative Disorders

Neurodegenerative Diseases and Metal Ions: Metal Ions in Life Sciences Vol 1

Interplay between Metal Ions and Nucleic Acids

Electro-diffusion of ions

Electro-diffusion of ions

Ions in Polymers

Metal

Bioelectrochemistry.. Ions, Surfaces, Membranes

The.Ethical.EquOI^ions.(1945)

Physics of Atoms and Ions

Physics of Atoms and Ions

Macromolecules Containing Metal and Metal-Like Elements, Metal-Coordination Polymers

Physics of Atoms and Ions

Nickel and Its Surprising Impact in Nature: Metal Ions in Life Sciences (Vol. 2)

Transport in Metal-Oxide-Semiconductor Structures: Mobile Ions Effects on the Oxide Properties

The Ubiquitous Roles of Cytochrome P450 Proteins (Metal Ions in Life Sciences, Volume 3)

Metallothioneins and Related Chelators (Volume 5) (Metal Ions in Life Sciences)

Organometallics in Environment and Toxicology (Metal Ions in Life Sciences, Volume 7)

Metal Fire

Metal Swarm

Metal Reign

Metal Fire

Metal Fire

Der Metal

Metal Murderer

Metal Nanoparticles

Metal Fırtına

Metal Fire

Warlord Metal

Biosorbents for Metal Ions

Metal Ions in Fungi

Metal Ions and Neurodegenerative Disorders

Neurodegenerative Diseases and Metal Ions: Metal Ions in Life Sciences Vol 1

Interplay between Metal Ions and Nucleic Acids

Electro-diffusion of ions

Electro-diffusion of ions

Ions in Polymers

Metal

Bioelectrochemistry.. Ions, Surfaces, Membranes

The.Ethical.EquOI^ions.(1945)

Physics of Atoms and Ions

Physics of Atoms and Ions

Macromolecules Containing Metal and Metal-Like Elements, Metal-Coordination Polymers

Physics of Atoms and Ions

Nickel and Its Surprising Impact in Nature: Metal Ions in Life Sciences (Vol. 2)

Transport in Metal-Oxide-Semiconductor Structures: Mobile Ions Effects on the Oxide Properties

The Ubiquitous Roles of Cytochrome P450 Proteins (Metal Ions in Life Sciences, Volume 3)

Metallothioneins and Related Chelators (Volume 5) (Metal Ions in Life Sciences)

Organometallics in Environment and Toxicology (Metal Ions in Life Sciences, Volume 7)

Metal Fire

Metal Swarm

Metal Reign

Metal Fire

Metal Fire

Der Metal

Metal Murderer

Metal Nanoparticles

Metal Fırtına

Metal Fire

Warlord Metal

Recommend Documents