Use of Local Minerals in the Treatment of Radioactive Waste

O = 0(0H) •=Si(AI)

O = 0(0H) # = AI, Mg, Fe, etc.

TECHNICAL REPORTS SERIES No. 136

INTERNATIONAL ATOMIC ENERGY AGENCY, VIENNA, 1972 USE OF LOCAL MINERALS IN THE TREATMENT OF RADIOACTIVE WASTE The following States are Members of the International Atomic Energy Agency:

AFGHANISTAN GUATEMALA PAKISTAN ALBANIA HAITI PANAMA ALGERIA HOLY SEE PARAGUAY ARGENTINA HUNGARY PERU AUSTRALIA ICELAND PHILIPPINES AUSTRIA INDIA POLAND BELGIUM INDONESIA PORTUGAL BOLIVIA IRAN ROMANIA BRAZIL IRAQ SAUDI ARABIA BULGARIA IRELAND SENEGAL BURMA ISRAEL SIERRA LEONE BYELORUSSIAN SOVIET ITALY SINGAPORE SOCIALIST REPUBLIC IVORY COAST SOUTH AFRICA CAMEROON JAMAICA SPAIN CANADA JAPAN SUDAN CEYLON JORDAN SWEDEN CHILE KENYA SWITZERLAND CHINA KHMER REPUBLIC SYRIAN ARAB REPUBLIC COLOMBIA KOREA, REPUBLIC OF THAILAND COSTA RICA KUWAIT TUNISIA CUBA LEBANON TURKEY CYPRUS LIBERIA UGANDA CZECHOSLOVAK SOCIALIST LIBYAN ARAB REPUBLIC UKRAINIAN SOVIET SOCIALIST REPUBLIC LIECHTENSTEIN REPUBLIC DENMARK LUXEMBOURG UNION OF SOVIET SOCIALIST DOMINICAN REPUBLIC MADAGASCAR REPUBLICS ECUADOR MALAYSIA UNITED KINGDOM OF GREAT EGYPT, ARAB REPUBLIC OF MALI BRITAIN AND NORTHERN EL SALVADOR MEXICO IRELAND ETHIOPIA MONACO UNITED STATES OF AMERICA FINLAND MOROCCO URUGUAY FRANCE NETHERLANDS VENEZUELA GABON NEW ZEALAND VIET-NAM GERMANY, FEDERAL REPUBLIC OF NIGER YUGOSLAVIA GHANA NIGERIA ZAIRE, REPUBLIC OF GREECE NORWAY ZAMBIA

The Agency's Statute was approved on 23 October 1956 by the Conference on the Statute of the IAEA held at United Nations Headquarters, New York; it entered into force on 29 July 1957, The Headquarters of the Agency are situated in Vienna. Its principal objective is "to accelerate and enlarge the contribution of atomic energy to peace, health and prosperity throughout the world".

Permission to reproduce or translate the information contained in this publication may be obtained by writing to the International Atomic Energy Agency. Kärntner Ring 11, P.O. Box 590, A-1011 Vienna, Austria.

Printed by the IAEA in Austria

June 1972 TECHNICAL REPORTS SERIES No. 136

USE OF LOCAL MINERALS IN THE TREATMENT OF RADIOACTIVE WASTE

INTERNATIONAL ATOMIC ENERGY AGENCY VIENNA, 1972 USE OF LOCAL MINERALS IN THE TREATMENT OF RADIOACTIVE WASTE IAEA, VIENNA, 1972 STl/DOC/10/136 FOREWORD

A great deal of information has been made available on the various techniques in use for safely managing radioactive waste. The International Atomic Energy Agency periodically convenes panels or other meetings to bring this information up to date and publishes Safety Series, Technical Reports Series and Guidebooks to help disseminate this information. Technical Reports Series No. 78, 'Operation and Control of Ion- Exchange Processes for Treatment of Radioactive Wastes' , published by the Agency in 1967, deals with the use of high capacity organic and inorganic exchange materials. A need has been expressed by developing countries for more information on the use of locally available ion-exchange and sorbent materials. The present publication, which is the outcome of a panel meeting, on the Use of Local Minerals in the Treatment of Radioactive Waste, held at the Agency's Headquarters in Vienna on 5-9 May 1969, presents, for the first time in a single volume, the large amount of practical information available. A list of the panel participants and the advisers appears at the end of the book. These experts kindly provided the material for the publication, which was organized by Mr. E.W. Wiederhold of the IAEA.

CONTENTS

I. INTRODUCTION 1

II. NATURAL MATERIALS FOR RADIOACTIVE WASTE TREATMENT 3

II. 1. Introduction 3 11.2. Types of reaction mechanism 3 11.2.1. Distribution of the radioactive microcomponent between solid and liquid phases 3 11.2.2. Co-precipitation 3 11.2.3. Coagulation and flocculation of colloids 6 11.2.4. Adsorption from solutions 9 11.2.5. Ion exchange 10 11.2.5.1. Ionic crystals 10 11.2.5.2. Aluminosilicates 12 11.2.6. Mineral replacement reactions 17 11.2.7. Oxidation-reduction mechanisms 19 11.2.7.1. Redox exchangers 19 11.2.7.2. Examples 19 11.2.7.3. Applications 20 11.3. Materials 20 11.3.1. Introduction 20 11.3.2. Mineral classification scheme 21 11.3.3. Minerals of use in waste treatment 21 11.3.4. Oxides and hydroxides 23 11.3.5. Halides 23 11.3.6. Carbonates 24 11.3.7. Phosphates 25 11.3.8. Sulphates 26 11.3.9. Silicates 26 11.3.9.1. Layer silicates 26 11.3.9.2. Zeolites 27 11.3.9.3. Crystal chemistry of the clays and zeolites 27 11.3.9.4. Waste treatment applications — 29 11.3.10. Other naturally occurring substances 30

III. CHARACTERIZATION OF MATERIALS 31 III.l. Sampling 31 III. 1.1. Sampling in the field 31 III.1.2. Laboratory sample preparation 32 III.1.2.1. Microscopic examination 32 111.1.2.2. Crushing 32 111.1.2.3. Sieving 32 111.1.2.4. Separation methods 33 111.2. Identification of the material 33 111.2.1. X-ray diffraction 33 111.2.2. Electron microscopy and electron diffraction 34 111.2.3. Petrographic microscopy 34 111.2.4. Thermal methods 35 111.2.5. Infrared spectrometry 35 111.2.6. Chemical methods of material identification 36 111.3. Physico-chemical characterization 36 111.3.1. General remarks 36 111.3.2. Capacities 36 111.3.2.1. Definitions 36

111.3.2.2. Pure ion exchange capacity (Kr) 37 111.3.2.3. Sorption capacity (K ad) 37

111.3.2.4. Total exchange capacity (Ktot) 38 111.3.2.5. Procedures 38 111.3.2.6. Complications 39 111.3.3. Selectivity 39 111.3.3.1. Ion exchange equilibrium 40 111.3.3.2. Ion exchange isotherm 40 111.3.3.3. Distribution coefficient (K*) 41 41 111.3.3.4. Separation factor (OB ) 111.3.3.5. Selectivity coefficient (NKB) 42 111.3.3.6. Procedures 42 111.3.4. Kinetic properties 42 111.3.4.1. Definitions 43 111.3.4.2. Procedures 43 111.3.5. Swelling properties 44 111.3.5.1. Definitions 44 111.3.5.2. Procedures 45 111.3.6. Reduction-oxidation properties 45 111.3.6.1. Redox capacity 45 Definition 45 Procedure 45 111.3.6.2. Redox potential 46

IV. MATERIALS 47 IV.1. Material preparation 47 IV.1.1. Crushing and grinding 47 IV.1.2. Sieving (screening) 48 IV.1.3. Washing 51 IV.2. Chemical and heat pre-treatment 52 IV.2.1. Chemical pre-treatment 52 IV.2.2. Heat pre-treatment 54 IV.2.3. Pelletizing 56 IV.2.3.1. Heat treatment of bentonites 56 IV.2.3.2. Heat treatment of alumina and clinoptilolite 56 V. PLANT SCALE APPLICATIONS 57 V.l. General 57 V.l.l. Ion exchangers 57 V.l.1.1. Batch process 57 V.l.1.2. Column operation 57 V.l.2. Additives and product conditioners 59

V.2. Operational experience 59 V.2.1. United Kingdom (Harwell) 59 V.2.2. India (Trombay) 61 V.2.3. United States of America 61 V.2.3.1. National Reactor Testing Station (NRTS), Idaho 61 V.2.3.2. O&k Ridge National Laboratory 64 V.2.3.3. Battelle North West Laboratory, Richmond 65 V.2.3.4. Savannah River 65 V.2.4. Federal Republic of Germany 65 V.2.5. Czechoslovak Socialist Republic 66

VI. FINAL PRODUCT CONDITIONING 67 VI.1. Conditioning of exhausted inorganic ion exchangers 67 VI.1.1. Treatment at Idaho Falls 67 VI.1.2. Treatment at Harwell 67 VI. 1.3. Incorporation into bitumen or concrete 67 VI.1.4. Fixation in glasses 68 VI.1.5. Reduction of leachability by heat treatment 68 VI.2. Use of inorganic minerals for improvement of fixed radioactive wastes 68

VII. ECONOMIC ASPECTS 71 VII.1. Capital costs 71 VII.2. Operating costs 72

APPENDIX I: NATURAL MATERIALS FOR USE IN WASTE TREATMENT 73 APPENDIX II: DEFINITION AND DETERMINATION OF CAPACITIES OF NATURAL ION EXCHANGERS .... 81 1. Introduction 81 2. Elementary principles 82 3. Definitions of ion exchange capacity 83 4. General methods of determination 92 References to Appendix II 95

APPENDIX III: COMMUNICATIONS CONCERNING NATIONAL EXPERIENCE IN VARIOUS COUNTRIES 97 1. France 97 2. Germany, Federal Republic of 97 3. Italy 98 4. Korea 99 5. Union of Soviet Socialist Republics 99 Special communication by V.M. Sedov, USSR: Research in the USSR on the use of natural sorbents for radioactive waste treatment 99

REFERENCES 107

LIST OF PARTICIPANTS 113 I. INTRODUCTION

In 1962 the Agency convened an ad hoc panel on Radioactive Waste Disposal into the Ground. This panel was primarily concerned with the rate of movement of ground water and radioactive isotopes through soil. Ion exchange with and sorption by naturally occurring zeolites and clays were discussed. Replacement reactions, in which a slightly soluble mineral in the soil is gradually replaced by a less soluble species by reacting with an appropriate ion in the solution, were also discussed. In 1964 the Agency convened another panel relating to the use of minerals in the treatment of radioactive wastes. This panel was on the Application of Mineral Reactions in Radioactive Waste Treatment. Much useful information was presented and a variety of mineral reactions were discussed. Among the topics discussed were mineral species used as aids in flocculation, as sorbents or exchange media in beds or columns, as sorbents or exchange media which remove the activity from discharges to the ground and those species which tend to concentrate activity in the sediments of streams, lakes and oceans. Much work had been done on the various clays, vermiculite, zincite, pyrolusite and zeolites. It was established that many synthetic inorganic, as well as organic, sorbents could be prepared for the solution of specific waste management problems. Much of the past research and development work on the use of minerals in the treatment of radioactive wastes has been a result of the solving of specific problems that arose in the course of developing nuclear power and other nuclear applications in the more highly developed countries. The process economics in these countries tend to favour those applications involving smaller columns or other equipment and thus the use of sorbents with relatively high capacities and good regeneration properties. Both synthetic, inorganic compounds and chemically and thermally treated natural minerals are finding applications in waste management in these areas. Applications of this type, for the main part, are concerned with the treatment of high-or intermediate-level wastes of small volume. In some instances, the high degree of selectivity of inorganic sorbents and exchangers permits a more economic treatment of certain wastes. The economics, however, are somewhat different for the processing of high- volume, low-level wastes. Here the cost of treatment chemicals may represent a large part of the real processing costs. In this case the utilization of inexpensive minerals may be justified because of the overall savings. In some instances, the use of natural minerals in sorption and exchange reactions is possible after suitable chemicals and/or thermal treatments to improve capacity and stability. Here the exhausted material may be disposed of as a contaminated solid and not regenerated as would be the more expensive materials. Waste management problems in the developing countries can in some cases be quite different and locally available resources can play a larger role in waste processing. A panel was convened on applications of mineral

1 resources in radioactive waste management with the emphasis placed on the use of local minerals in developing countries and areas where, due to the availability of inexpensive raw materials, nearness of burial site or other economic or technical considerations, the use of local minerals can be economically feasible. Topics such as types of minerals, e. g. vermiculite, clays, calcite, peat (although not a mineral in the same sense), and naturally occurring ion exchangers were discussed. A number of possible applications were presented, such as the use of minerals as flocculant aids, sorption columns for mixed fission products, lining or filling material in liquid disposal fields, the stabilization of small amounts of activity during transport and final product conditioning. Since highly processed inorganic materials and synthetic organic and inorganic ion exchangers had been well treated in the Agency's Technical Reports Series No. 78, they were not discussed. This report presents a great deal of practical information which it is hoped will be of value to those studying inorganic ion exchange or evaluating processes for solving their particular waste management problem. The classification and selection procedures are of particular value to those who must begin their evaluation without the benefit of many years of experience in this field.

2 II. NATURAL MATERIALS FOR RADIOACTIVE WASTE TREATMENT

II. 1. INTRODUCTION

Many naturally occurring materials exhibit one or more of the useful sorption or chemical reactions with radioactive or stable trace elements. These natural materials are seldom pure chemical species and thus a number of different reactions may occur in the testing or use of a single material. To evaluate these natural materials for use in the sorption of radioactive nuclides, it is necessary to understand the various useful reaction mechanisms and classification schemes.

II. 2. TYPES OF REACTION MECHANISM

II. 2. 1. Distribution of the radioactive microcomponent between solid and liquid phases

For the distribution of the radioactive microcomponent between the solid and liquid phases, the validity of the simple distribution law, Ci/Ca = K, was proved by Khlopin, where Ci and C2 and the concentrations of the dissolved microcomponent in the respective phases and K is the distribution constant.

II. 2.2. Co-precipitation

The oldest method in radiochemistry for separating the radioactive microcomponent from solution is to add a macroamount of a salt containing an ion chemically similar to that of the microcomponent in solution, and to precipitate both macro- and micro-ions by adding an excess of a salt containing the anion forming an insoluble salt with the former ions, or by raising the pH of the solution to form insoluble hydroxides. The magnitude of the action on the microcomponent in co-precipitation depends on the crystallographic compatibility between the macro- and microcomponent ions and other factors, such as the concentration of macro-ions (hence the precipitation rate), the relative excess of either ion, and temperature. For the case of mixed crystal formation between the micro- and macrocomponent within the crystal, essentially two kinds of distribution of microcomponent within the crystal have been observed, (1) homogeneous or equilibrium distribution, and (2) heterogeneous or logarithmic distribution.

3 The equilibrium (homogeneous) distribution of the microcomponent (x) between the crystal phase and solution conforms to the equation derived by Henderson and Kracek expressing Khlopin1 s law:

(1) y b - y where x and y are the amounts of micro- and macrocomponent respectively in the solid phase, a and b are their contents in the system, and D is the equilibrium co-crystallization coefficient, which is independent of the relative quantities in the solid and liquid phases and of the reaction path. For D > 1, the microcomponent will be concentrated relative to the macrocomponent in the solid phase (enrichment systems), while for D < 1, the microcomponent will be concentrated in solution (depletion systems). Consequently, the more D differs from 1, the more complete is the separation of the elements in the process of co-crystallization. Heterogeneous distribution is obtained with precipitation from a slightly supersaturated solution when there is a limited number of crystallization centres (nuclei) on which the growth of crystals can take place. Such crystals possess a rather perfect structure and a small surface area. The diffusion rate within the crystal is so slow that at a given time an equilibrium is established only between the surface of the growing crystals and the solution. Then the distribution of the microcomponent between the solid phase and the solution conforms to the relation derived by Doerner and Hoskins:

In —-— = X In — (2) a - x b - y

The logarithmic distribution coefficient depends upon the precipitation rate. Generally, at a very slow rate A. -» D;, whereas for very fast rates X -> 1 and co-precipitation is non-selective, producing neither enrichment nor depletion of the microcomponent in the solid phase relative to the solution [131], Some data on values of D and X obtained for various systems by different authors are presented in Table I. Note that for X, it is necessary always to state the precipitation conditions (rate). It must be emphasized that both equations (1) and (2) for homogeneous and heterogeneous co- precipitation respectively are valid only for the co-precipitation of ions of equal valence. Besides truly isomorphic co-crystallization, other forms of this process are known, which take place through the formation of mixed crystals of the "new type" or of "anomalous" mixed crystals. The "new type" mixed crystals are formed by ions the charges of which are different. The system BaS04- KMn04 serves as an example. By X-ray structural and physico-chemical analysis, it has been shown that these compounds form a solid solution and that a state of thermodynamic equilibrium may be attained in this system. However, the co- crystallization mechanism is different here. During the formation of the 2+ true mixed crystals (e, g. Ra with BaS04), any lattice position occupied by Ba2+ can be replaced by Ra2+, and co-crystallization can occur with as small a concentration of microcomponent as may be desired. However,

4 TABLE I. DISTRIBUTION COEFFICIENTS OBTAINED FOR SOME PRECIPITATION/CO-PRECIPITATION SYSTEMS3

Distribution coefficient Temperature Carrier Co-precipitant Ref. (°C) - D X

BaCr04 RaCr04 25 -5.5 173

BaS04 RaS04 - -1.2 174

BaCOj RaC03 90 0.2 175

BaC03 RaCOg 20 0.51 177

SrS04 RaS04 100 30 177

CaS04 RaS04 Hot <0.01 177

AgCl AgBr 30 211.4 177

AgBr AgCl 30 0.0036 177

BaS04 RaS04 20 1.8 178

BaS04 PbS04 25 0.026 179

BaS04 SrS04 83 0.03 176

PbCr04 RaCr04 0 ~0 177

A&Cr04 RaCr04 Room ~0 177

a A very comprehensive table of D values relating to co-precipitation of 226Ra in different solution compositions with diverse substrates and at different temperatures can be found in Ref. [ 133].

the formation of "new type" mixed crystals takes place by detached aggregates (microcrystals) of the microcomponent entering the crystallized phase of the macrocomponent, so that at a certain low concentration of the microcomponent, co-crystallization does not take place at all. For this reason "new type" mixed,crystals are of little use for practical purposes. By measuring the distribution coefficients it is possible to determine whether a given separation is of a quantitative nature or is dependent on compensating errors. Normally, the analyst wants to precipitate 99.8% (or more) of the primary substance and co-precipitate 0. 2% (or less) of the impurity. For this hypothetical system,the logarithmic distribution coefficient is 3.2X10"4. Hence, any separation for which A. ^ the latter value may be considered one which is ordinarily termed a "quantitative separation". The distribution coefficients may also be used to calculate the degree of separation which is attainable by double (or multiple) precipitation, and are also of use for considering fractional precipitation. The distribution coefficients for enrichment systems in which mixed crystals are formed usually decrease with increasing temperature, while for depletion systems the contrary may usually be expected. Consequently, an engineer responsible for waste management, in order to ensure, for example, the highest possible removal of radium from solution, should

5 carry out the precipitation of BaS04 at room temperature or lower (enrichment system), while for optimum removal of radiostrontium, the precipitation should be performed at the highest economically attainable temperature (depletion system). For the practical purpose of decontammating the radioactive solution, it is necessary to remove from solution as much radioisotope as possible with the smallest possible quantity of the least soluble precipitate (or preformed precipitate-sorbent). The degree of solution purification given by the ratio x/(a - x) increases as D increases and as the content of macro- element in the solution decreases [130], Hence, to achieve maximum purification of the solution, it is necessary to carry out the precipitation of the solid phase of the macrocomponent in such a way that its concentration in the dissolved state will be at a minimum during the entire precipitation process. For example, the co-precipitation of radiostrontium with BaS04 is best carried out by slowly adding the soluble salt of the macrocomponent Ba2+ to be precipitated to a solution containing the radioactive microcomponent (Sr2+) and the corresponding precipitating ions (SO|"), the latter being always in excess over the precipitated ions of the added macrocomponent (Ba2+).

II. 2. 3. Coagulation and flocculation of colloids

Naturally occurring waters always contain solids in different states of dispersion, e. g. either as true solutions of ions or molecules, as colloidal dispersions containing particles (10"7-10"5 cm) or as fine suspensions (10"4-10"2 cm). While true ionic or molecular solutions form one homogeneous phase only, the colloidal dispersions and suspensions are heterogeneous in that they contain both solid and liquid phases. The solid- liquid interphase area is the location of very complex solid-liquid interactions. These interactions acquire importance as the extent of the solid surface increases, a:nd are most significant among colloidal dispersions. These are formed by solid colloidal particles (sols) and a dispersing medium (usually water or water solutions). Lyophobic sols form substances practically insoluble in the liquid medium in which they are dispersed (usually water or water solutions). At the interface between the two phases, adsorption and ion exchange phenomena of the utmost importance for the stability of the colloid take place. The sol plus the water solution adhering to it form micelles (e. g. clay micelles). When the micelles are small (<10"4 cm), they do not usually sediment due to Brownian motion, unless conditions are imposed to promote their coagulation. When the particles are large (> 10"4 cm), the sedimentation is sufficiently rapid to be observable. By the application of Stoke's equation, the particle radius can be calculated from the sedimentation rate. This method, when applied to heterodispersed systems, can give a distribution curve of the sizes of particles present in a sample, and is frequently used in soil science. The lyophobic sols move along the lines of force when subjected to an electric field (electrophoresis), which demonstrates that the particles are electrically charged. This primary charge at the surface in contact with a polar solvent (water) may be acquired by ionization and/ or ion adsorption. For example, proteins acquire their charge mainly through

6 the ionization of carboxyl and amino groups to give COO" and NHJ ions. The ionization of these groups and thus the primary charge on the protein molecule depend strongly on the pH of the solution. At low pH the protein molecule will be positively charged and at high pH it will be negatively charged. The pH at which the net charge is zero is called the isoelectric point. If the primary charge is acquired by the ionization of strongly acidic groups (e. g. sulphonated polystyrene sols), then it will be practically constant with pH. Since the sol as a whole is electrically neutral, an equal charge of opposite sign must be present in the intermicellar liquid, and is supplied by ions of the opposite charge (counter-ions). Owing to the attraction between them, these charges of opposite sign remain adjacent, and each particle is thus surrounded by an electric double layer. The ions are the carriers of the double layer charges. The potential difference over the whole double layer in equilibrium is equal to the Nernst potential difference (e) between the two phases and is thus determined by the concentration of the so-called potential-determining 2+ + ions (co-ions), e. g. Ba or SO|" for BaS04, H and OH" for oxides and hydroxides, etc. Since the counter-ions always remain close to the interface, the electrophoretic velocity of a particle is not simply proportional to its charge, but it is retained more or less by the ions of the double layer. The potential difference between the bulk solution and the surface of the adhering part of the double layer is equivalent to the zeta potential, which is consequently smaller than the Nernst potential. Electrophoretic velocity and colloid dispersion stability are proportional to the zeta potential. Colloid particles agglomerate, coagulate and flocculate, if the zeta potential is zero or tends to zero, as also do isoelectric particles. The influence of electrolytes on the structure of the double layer and on the zeta potential is discussed in Ref. [132] (see Fig. 12, p. 79). The loss of stability by coagulation and flocculation brings about an irreversible cohesion of the colloidal particles in the form of loose and irregular clusters, in which the original particles can still be recognized. Coagulation and flocculation is usually obtained by the addition of electrolytes to a hydrophobic sol. The electrolyte concentrations necessary for flocculation are strongly dependent upon the valency of the electrolyte, or more specifically upon the valency of the ions that are oppositely charged to the sol. Moreover, flocculation values are practically independent of the specific character of the ions and only slightly dependent upon the concentration and even the nature of the sol. According to the rule of Schulze and Hardy, flocculation values occur over the following ranges: 0.150 -0.025 M (for monovalent counter-ions); 0.002 -0.0005M (for divalent counter-ions); and 0.0001 -0.00001M (for trivalent counter-ions). Exceptions to the rule of Schulze and Hardy occur when the counter- ions are such to be specifically adsorbed (e. g. large organic ions) or to react chemically with the ions building up the double layer. It should also be noted that some sols, e. g. Fe(OH)3, are stable in both slightly acidic and slightly alkaline media, since they carry a positive charge in the former and a negative charge in the latter case. In between there is a flocculation region. The zeta potential (ZP) characterizes in the given medium the properties of the colloid dispersion, such as the dispersion stability, the electrophoretic velocity, etc. According to Smoluchowski:

7 47r r;v Dx

where r/ = viscosity of the medium (usually aqueous solution); v = electrophoretic velocity (determined by observation under the microscope); x = potential gradient; D = dielectric constant of the medium; ZP = zeta potential.

A zeta potential in excess of 20 mV usually prevents sols from agglomerating and the dispersion is stable. The lowering of the zeta potential to the critical level, e. g. the level at which the coagulation begins, is attained by (1) the admixture of ions bearing a charge opposite to that of the potential- determining ions (see the rule of Schulze and Hardy); (2) changing the concentration of the potential-determining ions; or (3) raising the ionic strength of the solution. The first method above for flocculation by zeta potential control is of practical importance, and is exemplified by the clarification process widely used in water technology for removing colloidal and finely suspended solids. The sequential steps in clarification are: electrolyte admixture -> coagulation flocculation -» separation. Aluminium sulphate (A12(S04)3) is most generally used as it is hydrolyzed by the alkalinity of water, according to:

A12(S04)3 . 12 H20 + 3 Ca(HC03)2 = 2 A1 (OH)3 + 6 C02 + 3 CaS04 + 18 HjO

In an alkaline medium the formation of Al(OH)3 particles may be promoted by the addition of "activated silicic acid" (colloidal dispersion of stabilized hexasilicic acid). The silicic sol bears a positive charge at pH < 5. 5 and a negative charge at pH > 5. 5. By the admixture of A12(S04)3 as coagulant into the alkaline solution, the chemical precipitation immediately forms a large number of highly dispersed sols of Al(OH)3 bearing a positive charge at pH < 7. When two hydrophobic sols of the same charge that do not react chemically with each other are mixed, the resulting system is a stable sol containing two different kinds of particles. If, however, one sol is positively charged and the other negatively, e. g. positively charged sol of Al(OH)3 and negatively charged clay and other impurity sols, they flocculate with each other.

By the admixture of A12(S04)3 with proper pH control, the positive sols created flocculate with clay and other negatively charged sols and settle out. Flocculation of positively charged sols is also promoted by the action of the divalent sulphate ions. A correct dose of coagulant is necessary to ensure quantitative coagulation, and either insufficient or excessive doses should therefore be avoided. In the first stage, coagulation must be promoted by vigorous mixing to ensure uniform distribution of coagulant. Brownian motion promotes the close approach of sols, which is necessary for coagulation (perikinetic coagulation). As long as the agglomeration by coagulation creates particles of sufficient diameter (about 5 • 10"4 cm), the velocity gradient of the liquid and not the Brownian motion ensures

8 particle collisions. In this so-called orthokinetic coagulation, the particles aggregate to floes (flocculation), which are finally allowed to sediment and are separated.

II. 2. 4. Adsorption from solutions

Adsorption lowers the surface tension. Substances which produce a marked reduction of interfacial tension are said to be surface active. For adsorption, in contrast to ion exchange, the whole molecule (polar or non-polar) is deposited usually on a non-ionic solid surface, e. g. carbon black, alumina or silica. This is usually done in exchange for another less firmly held molecule located on the surface. This displaced molecule may be an adsorbed solvent molecule (water). The firmness with which the adsorbed molecules are held on the surface is generally determined by the amount of heat that is released by the adsorption. Physical adsorption involves either polar attraction between substrate and adsorbate, or dispersive forces ("van der Waals forces") which are weak relative to the chemical forces that bind atoms together to form molecules. Consequently, if the heat of adsorption is comparable to the heat of a chemical reaction (10-100 kcal/mole), chemical bonding is operative (e. g. hydrogen bonding) and a type of chemisorption takes place. The adsorption equilibrium that is established is determined by the heat of adsorption (AH) or surface tension change (6a), the solute concentration (c) and adsorbate concentration (x). This equilibrium is usually very temperature-sensitive. The mathematical relation linking these variables is called the adsorption isotherm, Surface tension (cr) and adsorption, expressed by the excess concentration (T) of a solute on the surface, are related by the Gibbs adsorption isotherm as follows:

6a r = - — (1) RT 6c

If AH varies with the surface coverage, the most active surface sites are filled first. Then the equilibrium may often be defined by the purely empirical Freundlich isotherm:

x = k c 1/n (2) where k and n are constants and (l/n) <1. If AH does not vary with the surface coverage, the Langmuir isotherm based on kinetic considerations may hold:

!s£ m— = -1 + k-c (3) where k is a constant and m is the adsorbent mass. The constant k is proportional to AH in both equations (2) and (3). For very low values of c, equation (3) reduces to x/m = kc, while for very high values of c, it becomes x/m = k. Hence for intermediate values of c, an expression of the type x/m = kc1/" (where l/n lies between 0 and 1) may be expected to

9 hold, which is identical to the Freundlich relation (2). If 1/n = 1, the adsorption equation would be equivalent to the distribution law. Examples of adsorption are the uptake of molecules of dyes or humic acids and uncharged (not dissociated) complex molecules on charcoal, carbon black, silica, alumina or clay micelles (e. g. kaolinite), or of weak (not dissociated) electrolytes (e. g. organic acids) on phenolic or carboxylic ion exchange resins (under experimental conditions preventing the dissociation of the respective plenolic or carboxylic groups). The latter case can be explained by the interaction of van der Waals forces between the resin and the carbohydrate part of the respective acids.

II. 2. 5. Ion exchange

The reasons why a solid phase may act as an ion exchanger are different for different types of solids, e.g. slightly soluble ionic crystals, slightly soluble inorganic polymers such as aluminosilicates, organic polymers of a carbohydrate type such as cellulose, or of a polypeptide type such as protein, and synthetic organic polymers such as sulphonated cross- linked polystyrene. A given solid, such as a clay mineral, may also have more than one ion exchange mechanism, and the situation here is very complex. Consequently, it is necessary to mention a few distinct types of ion exchanger that may occur in combination.

II. 2. 5. 1. Ionic crystals

Ionic crystals (precipitates) immersed in a solution containing their constituent ions rapidly exchange the constituent ions in solution with the lattice ions on the surface. Except for these constituent ions, only ions forming isomorphic compounds within the structure, or at least (less rigorously) on the surface, can exchange with the lattice ions, e. g. 2+ 2+ 2+ SO|", Sr or Ra with BaS04, or Ca on the surface of a BaS04 crystal. This is primary exchange sorption and all considerations relative to isomorphic co-crystallization apply here. The equilibrium in the case of primary sorption of an isomorphous microcomponent is given by:

-—— m = D^ = constant (1) where (x/(l -x)) is the distribution coefficient of the microcomponent, p is the adsorbent weight, S is the specific surface of adsorbent, m is the molar concentration of the macrocomponent (i. e. the ion displaced from the surface by the exchange against microcomponent) in the bulk solution, V is the solution volume, and D is the coefficient (similar to the co- crystallization coefficient) the value of which equals unity in the case of isotopic exchange. The primary adsorption is slightly dependent on the concentration of the non-isomorphic ions in the solution, independent of the surface charge, and proportional to the overall surface of the adsorbent (pS). The isomorphic ions held in the surface layer in excess over the constituent lattice ion of opposite charge make the ionic crystal electrically charged. These ions are called potential-determining ions and are

10 balanced by the non-isomorphic ions of the opposite sign located in the outer part of the double layer. For example, AgCl immersed in a KC1 solution bears a negative charge imparted by the primary adsorbed potential-determining chloride ions; this charge is compensated by the potassium ions in the outer part of the electric double layer. Potassium ions can exchange against other ions in bulk solution by what is called secondary exchange. All ions present in the bulk solution can take part in the secondary exchange, irrespective of whether they are isomorphous or not. In the case of microcomponent sorption, we may apply the distribution coefficient (x/(l-x)) and the proportion of microcomponent held by secondary exchange adsorption is then given by:

where a is inversely proportional to the concentration of all ions in the bulk solution and consequently practically independent of the microcomponent concentration, z is the valency of the microcomponent cation, V is the solution volume, S is the specific surface of the sorbent, and K0 is the mass action coefficient of the microcomponent sorption. It may be seen from equation (2) that the uptake of the microcomponent by secondary exchange sorption depends exponentially on the valency of the sorbed microcomponent, while it depends on the concentration m of the potential-determining anions according to the relation:

•l/z = a + $ In m (3) 1 -x

As ionic solids we may consider, e. g. slightly soluble sulphates, carbonates, phosphates and halides. Although theoretically there is no direct relation between solubility and sorbability on ionic solids, it is known by experience that such a relation exists in that those ions are sorbed well which form less soluble compounds with the opposite ion in the lattice (the Paneth-Fajans rule). There are, however, notable exceptions to this rule. The ion exchange on hydrous oxides and hydroxides can, as a first approximation, be explained in similar terms, considering hydroxyl ions as the potential-determining ions. Amphoteric oxides such as hydrous alumina may sorb either cations or anions depending upon the pH of the solution, and this has been ascribed to the following equilibria:

+ Al(OH)£ + OH" ^ Al(OH)3 =52 A10(0H£ + H

Anion exchanger Cation exchanger

Alkali metal cation exchange studies on hydrous MnO(OH)2 have shown the existence of a well-marked affinity series Cs+ > NH|>K+>Na+>Li+ similar to that found for other inorganic exchangers.

11 The hydrous aluminium and iron(III) oxides are a frequent seat of phosphate fixation in soils (Swenson et al. , 1949) according to the equilibrium:

A1(H20)3(0H)3 + HgPOj === Al(H20y0H)2H2P04 + OH"

The resulting dihydroxy-dihydrogen phosphate of iron(III) or aluminium gradually decomposes as the pH increases above 6. 5. In the range pH3 -7, however, the polymerized phosphate of aluminium forms cation exchanging groups, such as: I 0 O - 1 11 ! Al-O-P-O" ' Na+ + 1/2 Sr2+ I I i ' 0 OH 1 1 which have a remarkable affinity for radiostrontium. At higher pH the phosphate group gradually splits under the formation of tertiary phosphates and aluminates. In the presence of Ca2+, there ensues the precipitation of basic calcium phosphates in the alkaline region, connected with strong radiostrontium capture by co-precipitation. These combinations of ion exchange and co-precipitation are probably operative when using the mineral crandallite (CaAl3(P04)2 (OH)5 HzO) as sorbent in alkaline solution (Irving et al. , 1963).

II. 2.5.2. Aluminosilicates

Cation exchange in clay minerals can be simply stated:

Na-clay + H+ ===• H-clay + Na+

However, ion exchange phenomena are not simple; they vary with the type of clay mineral, nature of the replacing ion, pH of solution, concentration of the replacing ion in the solution, the associated ions in the solution, and the cations already in the exchange positions of the clay minerals. Table II gives the cation exchange capacity and relevant structural information for some clay minerals. The structural causes for cation exchange in the clay minerals may be summarized as follows: (1) unsatisfied valences produced by "broken bonds" at surfaces and edges of particles; (2) unbalanced charges caused by isomorphous substitution of cations, e. g. Al3+ substituted for Si4+, giving one net negative charge; (3) dissociation of structural OH" radicals, the H+ of which may be replaced by metallic cations; and (4) accessibility of structural cations other than H+ which become exchangeable under certain conditions, e. g. at low pH values Al3+ ions move from the octahedral units to the exchange positions. The principal cause of cation exchange in montmorillonites, "illites", and vermiculites is isomorphous replacement. "Broken bonds" are the most important cause of cation exchange in kaolinite, halloysite, and in fine particles of other minerals such as quartz. In montmorillonite, exchange takes place at three sites: on the flat surfaces, on the edges, and between the silica and alumina layers where ions are loosely held to neutralize deficiencies in these layers caused by isomorphic replacements.

12 TABLE II. CATION EXCHANGE CAPACITY OF CLAY MINERALS [22]

Exchange capacity Mineral Structural control (meq/100 g at pH7)

Kaolinite Unsatisfied valences on edges of 3-15 structural units

Halloysite (2H20) As kaolinite 5-10

Halloysite (4H20) Unsatisfied valences on edges of 40-50 structural units and on internal surface between the layers

Montmorillonite Substitutions in the octahedral 7-100 and tetrahedral units giving excess negative charge, unsatisfied valences on edges of units

"lllites" As montmorillonite, plus deficiency 10-40 (hydrous micas) of K+ between the layers

Vermlculite Replacement of interlayer cations, 100 - 150 substitution within the units, and unsatisfied valences on edges of units

Chlorite No data; possible deficiency of 10 - 40 ( ?) charge due to substitution in the brucite layer

Glauconite As "illites" 11-20

Palygorskite Substitution of Al3+ for Si4+ in 20-30 structural units, unsatisfied exchange sites within channels in the structure

Allophane Porous amorphous structure with ~70 unsatisfied valences

The excess negative charge on the mineral is neutralized by an equivalent number of positive ions. In natural situations the most common exchangeable cations are Ca2+, Mg2+, H+, Na+ and K+. Calcium is the dominant cation in soil clay minerals. Ion exchange takes place in a water film that surrounds a micelle of clay or a mineral grain. This water film is considered to be a diffuse double layer. It contains water and an ion swarm that is dependent on the surface-charge density of the mineral surface (the surface of the clay micelle), the kinds of exchangeable cations, the concentration of electrolytes in the solution, and, to a lesser degree, on the temperature. The exchangeable cations are held to the mineral surface by coulomb forces. The sorption affinity of ions for a clay surface increases with the valence, and the less hydrated the cation, the more tightly it can be held. The anions known to be adsorbed or exchanged on clay minerals are CI , NO3, SO4 , PO4" and ASO4". They function as counter-ions and are exchangeable with other anions in the same way as cations are exchanged. However, the exchange capacity for anions compared with cations is small in montmorillonite (few broken bonds compared to the charge induced by isomorphous replacement). In soil there is a higher anion-exchange

13 capacity than in clay minerals, because of the presence of hydrous ferric and aluminium oxides. Anion exchange of fluoride and phosphate by dis- placement of OH" from clay minerals and hydrous oxides may occur. Certain cations can be so firmly held on clay minerals that they are virtually fixed. This phenomenon is known to occur with K+, NH+, Rb+ and Cs+, i. e. with ions having a large polarizability, in hydrous micas, "illites" and vermiculites. Lattice contraction of these clay minerals following their saturation with such cations causes the latter to be trapped. Cs is generally very strongly held by major clay minerals which is of advantage for the decontamination of water containing radioactive caesium. Phosphate is the principal anion fixed by clay minerals; this fixation is mainly due to the formation of insoluble salts of Fe, Al, and alkaline earths. Besides clay minerals, zeolites and felspathoids are typical cation exchangers of the aluminosilicate type in nature. Zeolites may be defined as crystalline aluminosilicates with a tetrahedral framework structure enclosing cavities occupied by cations and water molecules, both of which have enough freedom of movement to permit cation exchange and reversible dehydration [111]. Felspathoids are crystalline aluminosilicates in which the cavities contain occluded salt molecules as well as water molecules and cations. Zeolites comprise essentially crystalline, cross-linked, polymeric macromolecules. The smallest units of these macromolecules are

Si(0/2)4 and Al(0/2)4 tetrahedra, where 0/2 represents the bridging oxygen atoms. The isomorphous replacement of Al3+for Si4+ gives rise to a negative charge on the lattice. Such a replacement is common to all aluminosilicates, clays included, and the net negative charge must be compensated by cations to preserve the electroneutrality. These cations are freely exchangeable against others of the same charge. In the aluminosilicate lattice (see Fig. 1), all four oxygen atoms are bridging atoms except where the macromolecule terminates at the crystal faces, in which case a proton co-ordinates to the non-bridging oxygen atom. The concentration of these protons (forming a hydroxyl group with the oxygen ion) grows as the dimensions of the aluminosilicate crystal diminish. Con- sequently Si-OH groups are rather important in colloidal clay minerals.

HO-Si -O-Al-O-Si-O-Al-OH I I I i 0 0 0 0 HO-Al — 0 - Si - 0-Al-0 - Si -OH I I I I 0 0 8 Me+ I I I HO-Si -0 - Al- 0 -Si -0 - Al-OH i i I i 0 0 0 0 i I I I HO-Al—0-Si— O-Al-O-Si -OH i I I I. 0 0 0 0 H H H H

FIG. 1. Two-dimensional idealized representation of an aluminosilicate lattice. Eight net negative charges on the lattice must be compensated by eight positive charges supplied by exchangeable cations.

14 There are enough metal cations, such as Na+, K+, Ca2+, Mg2+, or Srs+, present in the interstices of the aluminosilicate lattice to make the crystal electrically neutral. These cations are usually mobile in zeolites and are responsible for their ion exchange properties, while in feldspars they are not mobile due to the closed structure of these minerals and are therefore not available for exchange. The crystallographic character of zeolites is determined by the anionic lattice, which is manifold and complex. In this crystallographic arrangement, there exist empty spaces or channels in the aluminosilicate lattice which are occupied by water molecules and exchangeable ions. The zeolites are usually not very acid-resistant; this prevents their application in acid solutions (generally under pH5). However, the acid- resistance rises with increasing Si/Al ratio in the structure. Consequently, erionite (Si/Al = 3) is more acid-resistant than chabazite (Si/Al = 2), while mordenite or clinoptilolite (Si/Al = 5) possess notable acid-resistance and can be applied in aqueous solutions of pH substantially under 5. For application as radioactivity scavengers, it is of paramount importance that zeolites are generally resistant towards ionizing radiations, and, with certain exceptions, are also resistant to neutral water solutions at temperatures up to 100°C. The cation exchange capacities of zeolites range from about 230 meq/100 g (mordenite) to about 620 meq/100 g (thomsonite). The pore dimensions in zeolites may limit the entrance of exchanging cations for steric reasons. This ion sieve effect of zeolites may give rise to unusual selectivities. The effective pore size in various zeolites decreases in the following series: faujasit > sieve A > chabazite > gmelinite > mordenite > levynite > harmotome > analcite. In faujasite the extent of exchange among alkaline earths decreases in the order: Ba > Sr > Ca > Mg, suggesting that the hydrated ions are involved, although in the more compact structures such as ultramarine the affinity towards the alkali metal cations indicates that in this case the unhydrated ions are concerned. The full'ion sieve effect is observed when the pore diameters are such as to exclude certain cations in solution from entering the zeolite. This permits the separation of, e. g. Rb+ (1. 48 A) and Cs+ (1. 63 A) using analcite, since only Rb+ can enter this zeolite. A partial ionic sieve effect occurs with all ions which diffuse into the zeolitic pores with different rates due to their different ionic sizes. When extreme effects due to the ion sieve action are absent, the zeolites may still display marked selectivities towards certain ions due to differing thermodynamic affinities. The affinity varies rather un- predictably and in many cases it is strongly dependent on the cationic composition of the exchanger, since frequently the solid phase does not behave ideally. For example, in the exchanges

mordenite-Na + Cs+ mordenite-Cs + Na+ or clinoptilolite-Na + Cs+ clinoptilolite-Cs + Na+ the mass action quotient (K^a) and consequently the selectivity of the caesium uptake is high when the caesium coverage of the zeolite is low. The selectivity of some silica-rich zeolites such as clinoptilolite for

15 TABLE III. DEPENDENCE OF KJ FOR CATION EXCHANGE ON SOLUTION COMPOSITION AND PHYSICAL STATE OF THE RADIOACTIVE ISOTOPE [31]

K(j change Concentration Change of No. Solution composition change (V/m) Cation exchange Colloid adsorption

Non-complexing salt Constant Decreases Increases or remains constant

Non-complexing salt Constant Varies Constant Varies proportionally

Non-complexing salt plus Salt constant, Constant Decreases Increases or remains non-complexing acid acid Increases approximately constant

Non-complexing salt plus Constant Varies Constant Varies non-complexing acid proportionally

Non-complexing acid plus Constant Constant Decreases Decreases complexing agent

Non-complexing acid plus Constant Varies Constant Constant complexing agent

XV a Constant where the law of mass action is obeyed: Kj = -—- . — , where x is the portion of radioactive tracer in the sorbent, (1 - x) that in the solution, V is the solution volume, and g is the adsorbent mass. caesium is of great significance in the case of wastes containing higher concentrations of stable salts, such as the high-level wastes normally stored in underground tanks. Chemically, the radioactive fission products represent trace constituents in systems containing, in many cases, gross concentrations of non-radioactive salts. For example, a solution containing 0. 1 Ci 137Cs per litre is actually about 7. 10"6 M Cs+. In unselective ion exchange processes the capacity of the adsorption bed would be rapidly exhausted by the stable cations of the macrocomponent. Sorbents having a selectivity for certain fission products will display a larger effective decontamination capacity for these materials than would otherwise be the case. This is also very favourable for the performance of deep beds filled with such zeolites, when both large decontamination and large concentration factors are needed. In the case of sorbing the microcomponent (A), the distribution coefficient of'this microcomponent (Kj) is inversely proportional to the total solution concentration:

K*d » K*B (4) c0 where Kg is the mass action coefficient, q0 is the total saturation capacity of the exchanger (meq/g), and c0 is the total concentration of the solution (meq/ml). Generally for such a tracer component (A) of valency (z), Kj is inversely proportional to the macrocomponent concentration (co) raised to z the z-power, i.e. proportional to (c0" ). Consequently, if the law of mass action is obeyed, the plot of log Kj versus log c will be a straight line of slope -z. The K

The dependence of Kd on solution composition, sorbent concentration and solution concentration may yield valuable qualitative information on the reaction mechanism of sorption and on the physical state of the macrocomponent as may be seen from Table III.

II. 2. 6. Mineral replacement reactions.[68]

Mineral replacement reactions are those between a mineral and a constituent of the solution, resulting in the formation of a new crystal lattice and the inclusion of radioisotopes in the final product. Such reactions are really a special form of precipitation reaction. The mineral solid is contacted with a solution containing an anion or cation that can react with the appropriate ion from the mineral to form a compound which is less soluble in that system. Solution of the original solid takes place on a microscale at a reaction interface, permitting lattice rearrangement of the original crystal to that of the replacing mineral. Replacement

17 reactions include (1) anion replacement reactions, (2) cation replacement reactions, and (3) heterogeneous replacement reactions. Examples of anion replacement reactions include calcite-phosphate, calcite-fluoride, apatite-sulphate and gypsum-fluoride reactions. The calcite-phosphate replacement reaction can be formulated:

2 3 PC^" + 5 CaCOtj + OH" Ca5(P04)3(0H) + 5 CO " Crystalline Crystalline

Using this reaction the following isotopes can be removed from solution: Sr, Pu, Ca, Mg, Zn, rare earths, (Ru), (Zr). Reactions involving the above calcium mineral will generally favour the removal of bone-seeking radioisotopes, such as 90Sr, from solution. This is particularly true of the calcite-phosphate reaction, because the apatite product is chemically similar to bone mineral. Cation replacement reactions are of less interest for waste decontamination. Such a replacement takes place when a bed of gypsum (CaS04. 2 H2O) is contacted by a solution of barium chloride. Upon reacting, the barite

(BaS04) that is formed will incorporate other anions that fit into the new mineral lattice, such as the phosphate ion. Heterogeneous mineral replacement reactions result in the formation of a product that contains neither cations nor anions from the replaced mineral. That is, the gradual dissolution of the initial mineral creates conditions which promote the precipitation of a new compound. For example, when a dilute (15-150 p. p.m. ) solution of ferrous sulphate is passed through a bed of limestone chips, a film of ferric hydroxide forms on the mineral surface. This continuously regenerated film will remove certain forms of ruthenium from solution. The reason that metasomatic replacement reactions proceed at all is that the original material being replaced (e. g. calcite) is more soluble in the system than the final product (e. g. apatite). Radioactive ions that fit into the new mineral lattice are concurrently removed from solution by being fixed in the solid. Consequently, this process has all the features in common with mixed crystal formation by the multiple co-crystallization process effected under conditions of constantly renewed solution. If both components (e. g. Ca2+ and Sr2+) could be kept constant, a form of the homogeneous distribution law should be applicable. In anion replacement, the anion forms the least soluble compound with the cation or cations of the available solid. If the initial solid is the least soluble in a given system, there will be no replacement. The solubility difference in a given system between the original solid and a possible alteration product determines whether or not a given replacement reaction will occur and greatly affects the rate at which the reaction occurs. Given several possible alteration products less soluble than the original solid, and assuming a completed reaction, only the least soluble product is stable in the system. Several variables affect the difference in solubility between original and final product, including active anion concentration, extraneous ion concentrations, temperature and pH. With systems containing hydroxyl and carbonate compounds, we may assume that solution pH would have a significant effect on this solubility difference. For the extent of 85Sr removal from solutions of widely differing properties, see Ref. [32].

18 Besides the possible application of replacement reactions for the in situ treatment of wastes, they may prove to be an extraordinary potential means for immobilizing bone-seeking isotopes in certain soils after the disposal of waste on land by accident or design.

II. 2.7. Oxidation-reduction mechanisms

Oxidation-reduction reactions of naturally occurring ion exchange minerals are mentioned in the literature very rarely. Although they are by no means so important as ion exchange reactions, e. g. in the soil, several clay minerals are known to show such reduction or oxidation reactions. Humic acids and their derivates, e. g. in peat, show reducing properties due to the presence of hydroquinone/quinone groups in the organic molecule. From soil science it is well known that humus soil is able to reduce iron (III) to iron (II). Peat is sometimes also used in the treatment of radioactive wastes. These natural ion exchange materials showing reduction and oxidation properties may be able to change the valency of ions in waste solutions and therefore the ion exchange behaviour drastically. Furthermore, it has been shown recently that in some cases the ions of some noble metals in these materials can be precipitated as metal [134, 135],

II. 2.7.1. Redox exchangers

This unusual phenomenon can be discussed in terms of redox exchangers [134, 136], which are here defined as solid and insoluble substances in both oxidation states, and which because of their porous structure can react stoichiometrically with solutions of proper redox potentials. They are to some extent regenerable [134], Two types are known: redoxites, which are insoluble both in the reduced and oxidized forms and are strictly reversible, and redox ion exchangers, which are the usual ion exchangers (rich in redox ions) and are therefore not stable in all solutions, especially at high concentrations of electrolyte. Redox exchangers have been developed in analogy to ion exchange resins [134, 136], As a pure inorganic material, so far as is known, only nontronite has been investigated thoroughly as a redox exchanger [134],

II. 2.7.2. Examples

Nontronite is a clay mineral having the crystal structure of montmorillonite, in which, however, the structural Al3+ ions are replaced in 3+ part by Fe . The latter connect two tetrahedral Si04 layers, forming a non-swelling octahedral double layer of silicate. The Fe3+ ions are therefore not fixed by ion exchange, and nontronite belongs to the redoxite type of redox exchanger. Due to its Fe3+ content, nontronite has a yellow colour, which after treatment with sulphite or dithionite changes to deep green. This is due to an effect of the mixed Fe3+/Fe2+form, and not to Fe2+ alone. Only about one third of the Fe3+ can be reduced to Fe2+. The reduced form of nontronite can be re-oxidized to the Fe3+ form by excess of Fe3+, dissolved oxygen or hydrogen peroxide in water, or by the ions of some noble metals. This mineral, therefore, has been used for removal of oxygen and hydrogen peroxide from water and for selective

19 separation from solution of some noble metals, such as gold, silver, mercury and radioruthenium, by precipitation in the mineral as metals [134, 135,137, 138], Unfortunately, pure nontronite, like most clay minerals, is not very stable in acid or alkaline media; it can be used only in the relatively small pH range of about 4 to 7. In more acid solutions, iron is leached out and the silicate matrix is damaged. Because of its very small particle size, nontronite alone cannot be used in column operation. There exists, however, a form of nontronite, very rich in quartz ("chloropal"), which is very hard and can be used in the form of granules with 0. 3 -0.5 mm diameter in column operations [134, 137], For characterization of redox exchangers, redox capacity, analogous to ion exchange capacity, and redox potential (Eg^J for the insoluble material have been defined [134, 136], These properties are discussed in more detail in section III. 3. 6. , and in the following, there are given only some numerical values [134, 137], The chloropal from Ficht/Opf investigated contained more than 90% quartz, and had a total redox capacity (measured with Fe3+) of about 0.16 meq/g; the value for a pure nontronite from St. Andreasberg was about

4 meq/g. The redox potential E50

According to the pH dependence of the redox potential E5ffy0 of about 20 mV/pH, an electron transfer mechanism without overlapping of hydrogen ion transfer occurs.

II. 2.7.3. Applications

The mineral chloropal has been used for removal of oxygen and hydrogen peroxide from water and for selective separation of some noble metals, e. g. gold, silver and mercury, from solution by precipitation in the mineral as metal [134, 135, 137], For example, 37 g chloropal of particle size 0.3 - 0.5 mm in a column of 55 X 0.8 cm reduce the concentration of dissolved oxygen in 1.5 litres of water by a factor of more than 200. Other inorganic redox exchangers, e. g. zirconium oxide hydrate-dithionite, have been used recently for the removal of radioruthenium from fission product solutions [138].

II. 3. MATERIALS

II. 3. 1. Introduction

A mineral can be defined as a naturally occurring chemical element or compound formed by inorganic processes. Most of the natural materials of use in radioactive waste treatment are minerals, and the bulk of this section will therefore be devoted to a discussion of the various mineral classes with particular emphasis placed upon those minerals of use in waste treatment. A brief discussion of certain naturally occurring organic materials used in waste treatment is also included.

20 II. 3.2. Mineral classification scheme

Minerals can be classified in several different categories; genetical, geochemical, crystallochemical or even on the basis of the type of interaction shown with radionuclides. The modern crystallochemical classification is the most logical and useful in that it uses the relationships existing between chemical composition and structure of a mineral. This is the classification scheme presented here [99], (1) Elements. Only about 20 elements in the uncombined or native state are found as minerals, an example of which is gold. (2) Sulphides. This class consists mainly of combinations of various metals with sulphur, selenium or tellurium. The majority of the common ore minerals fall in this class, e. g. sphalerite (ZnS). (3) Sulphosalts. Minerals composed of lead, copper or silver in combination with sulphur and arsenic, antimony or bismuth are the only ones in this class. An example is the ore mineral enargite (CU3ASS4). (4) Oxides. The minerals in this class are those in which a metal is combined with oxygen. These substances are usually classified as A, simple or multiple oxides or B, hydroxides. Hematite (Fe203) belongs to group A, while gibbsite (Al(OH)3) belongs to group B. (5) Halides. This class includes the salts of a metal ion and the halides such as fluoride, chloride, bromide and iodide. Common minerals in this class are halite (NaCl) and fluorite (CaF2 ). (6) Carbonates. The minerals in this class are those whose formula contains the carbonate radical, COs. Examples of minerals in this class are calcite (CaC03) and dolomite (CaMg(C03)2). (7) Nitrates. This class includes those minerals containing the nitrate radical, N03. A mineral of this class is niter (KNOj). (8) Borates. These minerals contain the borate radical, BOs. An example of a common borate mineral is borax (NagB^. 10H20). (9) Phosphates. Minerals whose formula includes the phosphate radical are included in this group. An important mineral in this class is apatite (Ca5(P04)3(F, CI, OH)). (10) Sulphates. Minerals whose formula includes the sulphate radical

(S04) are included in this group. Barite (BaS04) is a common mineral in this class. (11) Tungstates. The relatively few minerals in this class have the tungstate radical (W04) in common. The mineral scheelite belongs to this class. (12) Silicates. The silicates form the largest class of minerals and contain various elements, most commonly calcium, sodium, potassium, magnesium, aluminium and iron, in combination with silicon and oxygen. The silicates are sub-divided on the basis of their structure into six sub- classes as shown in Table IV [99], For further information pertaining to the structure and mineralogy of the substances listed, Refs [99-108] should be consulted.

II. 3. 3. Minerals of use in waste treatment

Of the preceding 12 classes of minerals only half contain minerals useful for the treatment of radioactive wastes, e. g. silicates, oxides, halides, carbonates, phosphates and sulphates. With few exceptions,

21 TABLE IV. SILICATE MINERAL CLASSIFICATION [99]

Arrangement Ratio Class of Si04 Mineral example Si: 0 tetrahedra

Nesosilicates Isolated 1:4 Olivine, Mj^eSi04

Sorosilicates Double 2:7 Hemimorphite,

Zn4(Si207) (OH) -H20

Cyclosilicates Rings 1:3 Beryl, BejAljfSijO^

Inosilicates Chains

Single 1:3 Enstatite, Mg2(Si20&) Double 4: 11 Tremolite,

Ca2Mgs(Si8022)(OH)2

Phyllosilicates Sheets 2:5 Talc, MgjS^OnjtOHjj

Tectosilicates Framework 1:2 Quartz, Si02

such as the use of borate minerals for the production of glasses in pot calcination, the remaining classes of minerals are of little use in waste treatment and will therefore not be further discussed here. It is interesting to note why the silicates, etc. , are generally useful in waste treatment, while the sulphides, etc. , are not. The reason for this lies in the nature of the waste treated. For the most part, these wastes are aqueous and contain the principal radionuclide contaminants as ions. The goal of waste treatment is to remove these ions from solution and retain them in a concentrated and immobile form. For this, substances which will enter into ionic reactions are needed, and thus only those minerals which are dominantly ionic compounds are useful. The covalent sulphides and sulphosalts are therefore of little use. Only a few of the minerals in the above six classes of dominantly ionic compounds are useful in waste treatment; again the reason is chemical. To be useful, the mineral must combine with the various radionuclides in such a way that the resultant compound is insoluble in the waste stream. Thus, many compounds, such as most of the halides, nitrates, etc. , are excluded on the basis of solubility. Finally, it should be noted that frequently the major nuclides of interest in waste treatment are the alkalis or alkaline earths, e. g. 137Cs or 90Sr. Although an over-simplification, a useful general guide to the selection of minerals for radioactive waste treatment is that the minerals should be ionic solids capable of forming insoluble compounds with alkali or alkaline earth cations. Thus, it is no accident that barite or fluorite and the natural ion exchangers such as the clay minerals and zeolites are useful in waste treatment. With the preceding principle as a guide, the different minerals useful in waste treatment will now be discussed. Except where pertinent to the discussion of waste treatment, the structure and mode of occurrence of the different minerals will not be discussed. The previously cited references should be consulted for this information.

22 TABLE V. OXIDE AND HYDROXIDE MINERAL CLASSIFICATION [99]

Oxides

Hematite _groug_ Rutile group Goethite_grou£

Corundum A1203 Rutile TiOj Diaspore HAIO,

Hematite Fe^ Pyrolusite Mn02 Goethite HFeO'2;

Ilmenite FeTiOj Cassiterite Sn02

Uraninite U02

Hydroxides

Brucite Mg(OH)2 Limonite Fe0(0H)-nH20

Manganite MnO(OH) Bauxite Al hydrates

+2 4 Psilomelane BaMn MnJ 016(0H)4

II. 3. 4. Oxides and hydroxides

The minerals within this group may be classified as shown in Table V [99], which includes only those oxide and hydroxide minerals found useful in waste treatment. Tamura [1] has shown that 90Sr is sorbed on the sesquioxides, particularly AI2O3. The removal of 90Sr by some soils containing appreci-

able quantities of A1203 is thought to be due to a reaction involving the alumina. The oxides and hydrated oxides of iron and aluminium are particularly common minerals and are, for example, the principal mineral phases of the laterite soils, which may be quite useful in waste treatment [97], The oxides and hydroxides of manganese are also useful sorbents of radionuclides. Pyrolusite has been shown to be an effective sorbent of strontium by investigators in the USSR [3], Other oxide compounds of manganese should also be effective. Manganese nodules and manganese coatings on stream pebbles are noted for the variety and concentration of trace elements they contain. These different elements appear to be incorporated during the formation of the manganese dioxide layer. It is possible to make use of this same process in waste treatment. For example, the removal of cobalt and chromium from waste streams canbeeffectedbyco-precipitationonmanganese

dioxide [5], Similarly, co-precipitation reactions involving Fe(OH)3 and Al(OH)3 have been studied for the removal of a number of different radionuclides [2],

II. 3. 5. Halides

The major halide minerals are as follows [99]: halite (NaCl), cryolite

(NagAIFg), carnallite (KMgCl3. 6Hp), sylvite (KC1), fluorite (CaF2), and cerargyrite (AgCl).

23 Excluding halite mines as sites for the burial of radioactive wastes, only fluorite and possibly cryolite of the common halides are useful in waste treatment. The other halides are either too rare or too soluble to be of use. Although fluorite will remove strontium from a waste stream [98], it is much more likely to be the product of waste treatment rather than the reactant. Ames [98] has shown that strontium can be removed by passing a fluoride-containing waste stream through a column containing calcite. The calcite is converted to the more insoluble fluorite, with strontium isomorphously replacing the calcium in the newly-formed fluorite. Cryolite may be useful in the sorption of strontium but it does not appear to have been studied. Although this mineral is considerably more soluble in water than is fluorite, it may still be sufficiently insoluble for some waste-treatment purposes.

II. 3. 6. Carbonates

This class includes several minerals which have proved to be useful in waste treatment. The common carbonates can be divided into two groups as shown in Table VI [99], The calcite structure is assumed by the smaller cations, while the aragonite structure is formed by cations larger than about 1 A in diameter.

TABLE VI. CARBONATE MINERAL CLASSIFICATION [99]

Calcite _grouj)_ Aragonite_graup_

Calcite CaC03 Aragonite CaC03

Dolomite CaMg(C03),, Stronianite SrCOs

Magnesite MgCOj Witherite BaC03

Siderite FeC03 Cerussite PbC03

Rhodochrosite MnC03

Smithsonite ZnCOs

Solid solution of the smaller ions in the calcite group is quite extensive. Thus, Fe2+ substitutes readily for Mg2+, but substitution of Ca for Mg is not as extensive and leads ultimately to the ordered dolomite structure. In this structure, the carbonate ion layers are separated by a layer of calcium ions, then a layer of magnesium ions, and so on. This ordered structure is also formed by the smallest and largest ions of the aragonite group. Here, in the mineral barytocalcite, calcium and barium form an ordered structure similar to the calcium-magnesium structure of dolomite. The carbonates are useful in waste treatment because of their reactivity and abundance. Further, strontium occurs in the carbonates and it is therefore reasonable to assume that the carbonates could be useful in the removal of radiostrontium. The several examples below serve

24 to illustrate the manner in which the carbonates have been used in waste treatment. It has been shown that strontium can be removed from neutral and basic waste streams by sorption on calcite [1, 9,10, 35], The process is one of ion exchange in which the strontium substitutes for calcium in the calcite structure. A more important reaction is one in which, because of the chemical reactivity of the calcite, the mineral is converted to the very insoluble apatite by reaction with alkaline phosphate solutions. Here, the strontium isomorphously replaces the calcium in the apatite structure [7, 8,11-13], Recently, investigators in the USSR have shown dolomite to be an excellent sorbent for a number of radionuclides [3], In this application, the dolomite is first calcined at 7 50°C to selectively convert the MgCOa layers to MgO, leaving the CaCOg layers intact. This is then a highly reactive structure for utilization in waste treatment. Other carbonates have not been as thoroughly investigated for use in waste treatment. Of these, siderite, which is relatively abundant, may be of use for the sorption of Co or Ru.

II. 3.7. Phosphates

This class of minerals includes a large number of oxy-salts with the anionic group (XO4)"11, where X can be phosphorus, arsenic or vanadium and n equals three [108], There is extensive substitution between P and As and between As and V. Some of the several groups of minerals in this class are listed in Table VII [108],

TABLE VII. PHOSPHATE MINERAL CLASSIFICATION [ 108]

Anh^draa^iwrjMy)hos£hates_ Unrated normal £hosj>hates_

Xenotime YP04 Variscite A1P04-2H20

Monazite CeP04 Strengite FeP04-2H20

A^^drojis_jA^^ates_with hydroxyl_or halogen^

Apatite series Ca5(P04)3 (F, CI, OH, C03)

Phosphorite, fine-grained fluor- or hydroxy-apatite, is the substance forming the vast rock phosphate deposits of the world. Various members of the apatite group have been investigated as waste treatment media [7, 8,11-13], As in the case of fluorite and calcite, apatite can be used as a primary removal medium for such radionuclides as 244Pu and 90Sr. As previously discussed, apatite is also formed as the result of contacting calcite with alkaline phosphate solutions. Because of the insolubility of apatite and its ability to incorporate strontium, this mineral is a useful waste treatment medium. The aluminium phosphate, variscite,has also been studied for use in waste treatment. Like apatite, it is a good sorbent of strontium [11], Although they could be of use in waste treatment, the rare-earth phosphates, such as monazite, do not appear to have been studied.

25 II. 3. 8. Sulphates

The sulphate class can be divided into two groups consisting of anhydrous and hydrous salts, some of which are listed in Table VIII [99], By virtue of its abundance and insolubility, barite is the most important mineral of the sulphate group for waste treatment. With the exception of the abundant gypsum, which is quite soluble, the other sulphates are too rare to be of practical use [ 6, 9], The removal of radionuclides on barite has been studied extensively by Berak and co-workers [14-16], In this work, a strontium-selective sorbent was synthesized from barite and anhydrite at temperatures greater than 1000°C, where calcium is incorporated in the high-temperature modification of barite. If the solid solution so obtained is quenched in cold water and the calcium leached out, an activated barite residue, highly selective for strontium, is produced. The compound barium sulphate is also used in several synthetic substances prepared for waste treatment [89],

TABLE VIII. SULPHATE MINERAL CLASSIFICATION [99]

A ^^drous_sul£hates Hydrousjulghates

Barite group Gypsum CaS04-2H20

Barite BaS04

Celestite SrS04

Anglesite PbS04

Anhydrite CaS04

II. 3. 9. Silicates

Of all the minerals, the silicates have been investigated most extensively. There are two reasons for this: firstly, the silicates are the largest class of minerals and include the common mineral constituents of soils and rocks, and secondly, because of the practice of discharging radioactive wastes in the ground as a means of disposal, the interactions of the soil minerals and the radionuclides are of considerable interest. Minerals representative of all the silicate sub-classes (see Table IV) have been studied for use in waste treatment. Of these, only the clay minerals (phyllosilicates) and zeolites (tectosilicates) have any utility in waste treatment. For this reason, classification schemes will be presented for these two groups only.

II. 3. 9. 1. Layer silicates

Table IX [109] gives one of several possible classification schemes for the layer silicates. For alternative formulations, Ref. [104] should be consulted.

26 II. 3.9.2. Zeolites

The zeolites are not as yet adequately classified. The following general classification is, however, useful [102], For additional information regarding the structure and properties of the zeolites, see Refs [110-112], (1) Natrolite group, which includes natrolite, mesolite, scolecite, thomsonite, gonnardite and edingtonite. A chain-like unit is fundamental to the structures within this group, of which all members have a fibrous morphology. (2) Harmotome group, which includes harmotome, phillipsite, gismondine and garronite. The silica tetrahedra of this group form chains similar to those in the feldspars, except that the linkage of the chains is different. The network contains many four-fold and eight-fold rings, the latter constituting the channel openings. (3) Chabazite group, which includes the zeolites chabazite, gmelinite, levyne and erionite. Single and double six-fold rings of tetrahedra perpendicular to a triad or hexad axis are found in this group. Eight- and twelve-fold rings also occur, forming wide channel systems. (4) Faujasite group, which includes faujasite and some synthetic zeolites, e. g. Type A. Here, the tetrahedra are linked to form cubo- octahedral cage-like units which in turn are joined so as to give the structure a cubic symmetry. (5) Mordenite group, which includes the minerals mordenite and dachiardite. The characteristic feature of these structures is a chain containing five-fold rings of tetrahedra. These chains can be linked in a variety of ways. Wide channels are formed by twelve-fold rings in mordenite and by ten-fold rings in dachiardite. (6) Heulandite group, which includes heulandite, clinoptilolite, stilbite, epistilbite, ferrierite and brewsterite. The structure of the minerals in this group is at present under study [113],

II. 3. 9. 3. Crystal chemistry of the clays and zeolites

Unlike the other mineral classes in which co-precipitation and isomorphous replacement are the major reactions involving radionuclides, the silicates are ion exchangers. The phenomenon of ion exchange is exhibited to a marked degree only among the clays and zeolites. These ion-exchange properties can be explained by the crystal chemistry of these minerals. Because of its small size, Si4+ occurs in tetrahedral co-ordination with respect to oxygen, i. e. each Si4+ ion is surrounded by four oxygen anions to form a tetrahedral-shaped co-ordination polyhedron. These tetrahedra are the building blocks of the silicate structures, and the different ways in which they are joined give rise to the different silicate sub-classes. The aluminium ion (Al3+) is only slightly larger than the Si4+ ion and can isomorphously replace Si4+ in a number of silicate structures. This substitution is one of the main reasons why the silicates possess ion- exchange properties. Consider, for example, the mineral montmorillonite. As shown in Table IX, this mineral has a 2:1 layer structure, i. e. its structure consists of an octahedral brucite sheet sandwiched between

27 TABLE IX. PRINCIPAL GROUPS OF LAYER SILICATES [109]

2-layer = 1:1 = 7 A 3-layer = 2:1=10 A 4-layer = 14 A r Groups { KAOLIN-SEPTECHLO RITE PYROPHYLLITE- EXPANDED 2: 1 MINERALS MICA CHLORITE TALC

dioct. trioct. dioct. trioct. dioct. tiioct. dioct. trioct. trioct.

Kaolinite Chrysotile Pyrophyllite Talc Montmorillonite Hectorite Celadonite Polylithionite (SERPENTINES)

Mineral Dickite Stevensite or mineral Nacrite families3 Halloysite

Endellite

BEIDELLITES SAPONITES PHENGITES LEP1DOLITES -o

.3 a SEPTECHLORITES dioct. VERMICULITES Muscovite Phlogopite CHLORITES SO vermiculites

rt P BIOTITES Q> O C Margarite 1 Common mixed layers mixed layers heteropoly types

a Family names are in capital letters; names of minerals that have close to idealized end-member compositions are written with initial capitals. two silica tetrahedral sheets. Magnesium substitutes for aluminium and aluminium substitutes for silicon giving rise to the formula (Al3.x Mg^) (Si4-y Aly) Oio (OH)2, where x and y indicate the extent of subsitution of Mg2+ for Al3+ in the octahedral sheet and of Al3+ for Si4+ in the tetrahedral sheet respectively. As a result of this substitution, there is a net negative charge on the lattice of about 0.3. This charge is satisfied by large cations such as Naf, K+ or Ca2+, which occupy positions between the 2:1 layer units. Because these positions are external to the aluminium-silicon- oxygen framework, these cations are exchangeable. Although the zeolites have a different structure from the clay minerals, similar crystallochemical principles apply. Thus, through a combination of charge imbalance and open structure, minerals such as the clays and zeolites exhibit ion-exchange properties. In addition to charges arising by isomorphous replacement, surface charges due to broken bonds exist. In minerals such as kaolinite, which does not show aluminium-for-silicon substitution, these surface charges are responsible for the small cation exchange capacity shown by this mineral. The clays and zeolites are dominantly cation exchangers as would be expected by the nature of the charge imbalance. However, these minerals exhibit anion exchange behaviour as well. The anion exchange capacity involves the structural hydroxyl ions of the clays and the hydroxyl ions from the silanol (Si-OH) groups on the zeolite surface. Several other properties of the clays and zeolites are of interest. Certain of the clays, notably montmorillonite, possess the ability to adsorb large amounts of water, swelling tremendously in the jjrocess, e. g. expanding from a normal interlayer distance of about 12 A to a distance of more than 100 A. This swelling property is useful in some applications such as the use of bentonites to make water-impervious barriers around buried wastes. This same property makes it difficult to use bentonites in ion-exchange columns because of the great difficulty in passing fluids through the expanded clay. It is partly for this reason that vermiculite and illite are used in preference to bentonite for waste treatment. The montmorillonite minerals have the ability to absorb polar organic liquids as well as water. This property is useful in the treatment of tributyl phosphate wastes [60, 93], An interesting property of the zeolites is that these minerals contain water as an essential part of their structure. This water can be driven off at temperatures in excess of 100°C and it is this property from which the name zeolite, boiling stone, is derived. The dehydrated zeolites are capable of rehydrating or of sorbing a variety of gases. These minerals serve as gas sorbents in a number of industrial and analytical applications. The use of tuffs as sorbents has been discussed [58], Tuff or vitric tuff is the name given to rocks composed of compacted shards of volcanic glass, which are very reactive and readily convert to hydrous minerals such as clays and zeolites. Usually when tuffs are used for waste treatment, the underlying reason for the tuffs to function in this manner is that they have already been altered to various zeolites.

II. 3. 9. 4. Waste treatment applications

A great deal of work has been done using the clays and zeolites as waste treatment media. Only two current applications of these minerals

29 for waste treatment in the USA are included as examples here. Two minerals which find much use in the USA are the clay mineral illite and the zeolite clinoptilolite, both of which are caesium selective. Currently, illite is used as an additive to fix 137Cs in the lime-soda treatment process used for low-level wastes at Oak Ridge National Laboratory [85], This application is discussed in detail in the reference cited. Note that grundite, a commercial variety of illite, is used here. An example of the use of clinoptilolite in the treatment of low-level wastes is that at the Idaho Chemical Processing Plant [44, 51], Here, a partially zeolitized tuff is packed in disposable ion-exchange columns and used for the sorption of Cs and Sr. Once loaded to the strontium breakthrough, the columns and their contents are buried as solid waste.

II. 3. 10. Other naturally occurring substances

Numerous naturally occurring inorganic and organic substances other than minerals in the strict sense have been studied for use in waste treatment. The various rocks and substances such as perlite and diatomite belong to this group. Perlite is a hydrated volcanic glass which when heated expands to many times its original volume. It is useful as a sorbent because of its large surface area. In some applications, such as the incorporation of radioactive wastes into cement or ceramic blocks, perlite or diatomite (siliceous diatom tests and sponge spicules) are used as additives [3], Ceramic bodies in which various radionuclides are incorporated have been made from igneous rocks, particularly basalts [114], These bodies retain the radionuclides 131Cs, 90Sr and 106Ru and have good chemical and physical stability. Natural organic substances such as peat, humus, lignite, bitumen and sawdust have been thoroughly studied for use in waste treatment [5, 80, 81, 86, 87, 89, 91, 95, 96], Generally, the chemical stability of the organic ion exchangers is poor and they must be treated to improve their stability. Sulphonation is a common process and has been applied especially to coals. Recently, Belgian scientists [86] have sulphonated bitumen to produce an excellent polyfunctional ion exchanger. A listing of the various naturally occurring materials which have been studied for use in waste treatment is given in Appendix I. This table lists the type of substance, mineral name, radionuclide studied, type of reaction, and references to various investigators who have studied the particular substance. The table is reasonably complete from the standpoint of the minerals listed, but the listing of radionuclides and references is incomplete. A similar table for various inorganic and organic ion exchangers is given in Ref. [107],

30 III. CHARACTERIZATION OF MATERIALS

III. 1. SAMPLING

One of the factors which makes the use of minerals difficult in waste treatment is the unpredictable nature of the properties of the mineral. Thus, one batch of mineral will frequently differ greatly in behaviour from the preceding batch. Much of this inconsistent behaviour and consequent difficulty in use stems from the heterogeneity of natural materials. In part, the solution to this problem rests in thorough and proper sampling of the material to be used. The problem of obtaining a representative sample of natural material for use in waste treatment is difficult. For example, beds of zeolitic tuff may be metres thick and extend over many square kilometres. The extent of zeolitization of the tuff differs from one locality to the next; one outcrop of tuff may be completely zeolitized while another consists of unaltered glass. When one considers that the sample of the material evaluated in the laboratory is of the order of a few hundred grams and that the disposal facility may use several hundred tons of the material, the problem and importance of proper sampling becomes clear. A knowledge of the statistics of sampling and various sampling strate- gies as outlined in Chapter 7 of Ref. [115] is essential to a realistic ap- praisal of the potential of various materials for use in waste treatment. The statistical aspects of geological sampling cannot be overemphasized. Concerning the techniques of sampling, the discussion presented in Ref. [116] is very useful. The following brief discussion is taken mainly from this source.

III. 1.1. Sampling in the field

Sampling strategy should be carefully planned [115] . The actual collection of samples in the field should be undertaken by a geologist who is knowledgeable regarding the potential use of the mineral and the physico- chemical tests to be performed. Ideally, this field man should be accompanied by the laboratory specialist, since the samples taken routinely may be inadequate for the special tests envisaged. During the initial sampling, it is best to collect a quantity of material sufficient for bulk analysis as well as for separate analysis of the different constituents. It is preferable if all the required material is collected at one time. It is also desirable to keep a representative sample of the material after testing for future reference. The method of sampling differs with different types of materials. For minerals and rocks, collecting methods such as blasting with explosives, hammering out samples with a sledge hammer, channel sampling or core drilling may be used. Unconsolidated material may often be collected with a shovel or trowel. Samples of plastic or fibrous materials such as peat or clay must sometimes be cut out with a knife or saw.

31 It is not possible here to provide detailed directions regarding the amount of sample, the number of samples to be taken and the method to be used, which will be adequate for all materials and all purposes. A knowledge of how the material is to be used and the amount required for this use, coupled with the expertise of the field and laboratory scientists, should suffice to guide the sampling of natural materials in the field.

III. 1.2. Laboratory sample preparation

The samples gathered in the field must be prepared for the various laboratory tests envisaged. The material analyzed should be as representative of the material in the field as possible. This means that contamination of the sample and selective loss of mineral constituents must be held to a minimum. Excellent discussions of the preparation of natural materials for analysis are provided in Refs [116, 117].

III. 1.2.1. Microscopic examination

The first step in sample preparation should be a microscopic examination of the material. For rocks and consolidated materials, petrographic thin sections should be prepared and studied. Unconsolidated materials should be examined under the stereo-microscope. Microscopic examination provides information on which minerals are present, the mode of occurrence of the different minerals, as well as the type and extent of undesired phases. The particular sample treatment followed will depend upon information gathered from the microscopic examination.

III. 1.2. 2. Crushing

Generally, laboratory tests are performed upon fine particles of the material; as a rule, the material should not be crushed any finer than is required by the tests to be performed. Solid materials such as tuffs may be fragmented by flaking off chips with a hammer, breaking the specimen against a steel anvil or by breaking the sample between the jaws of a rock splitter (see Fig. 1, Ref. [116]). The small pieces of rock obtained by the above processes may be crushed further either in a percussion mortar, a motor-driven roller or a vibrating ball mill. If necessary, the material may be reduced further in size by grinding in an agate mortar. Clays and unconsolidated materials are usually treated more gently since the aim here is to separate the clay from the other minerals of the sample matrix and not to crush the primary particles. Crushing with a hardwood rolling pin and further reduction with a rubber-capped pestle in a porcelain mortar is usually employed for clay-like aggregates.

III. 1.2.3. Sieving

The crushed materials are usually sieved to ensure that all the sample is less than a particular particle size and to aid in the crushing process. By removing the fine material during crushing, the coarser particles can be crushed more efficiently and over-crushing is avoided.

32 Depending upon the tests to be performed, stainless steel, brass or nylon screens can be used. If chemical analysis of the material is to be carried out, contamination of the sample by metal from the screens may be a problem. In this case, nylon screens are preferable. A further use of sieving is to isolate or enrich a portion of the sample with the desired mineral constituent. For example, zeolites might occur in a tuff as small grains, while the quartz, feldspar and mica may be much larger in grain size. Sieving may remove much of the coarse impurities, leaving the fine-grained portion enriched in zeolites.

III. 1.2.4. Separation methods

It is sometimes useful to obtain a concentrate of a given mineral. In addition to sieving, mineral separation can be achieved by a number of methods. Those which utilize the difference in density between minerals include separation by suspension in heavy liquids and the use of such devices as panners and shaking tables. Minerals possessing varying degrees of permanent magnetism can be separated by such devices as a Frantz Isodynamic Mineral Separator (see Fig.8, Ref. [117]). With instruments such as this, even weakly magnetic minerals such as tourmaline and monazite can be separated from practically non-magnetic minerals such as quartz or zircon. In addition, methods such as froth flotation, electrochemical techniques, elutriation and even hand-picking of grains under a microscope can be used. A description and discussion of various mineral separation techniques is given in Refs [116, 117] .

III. 2. IDENTIFICATION OF THE MATERIAL

In characterizing the natural materials for use in waste treatment, it is extremely important that the identity of the materials be known. It is not enough to know, for example, only that clay minerals are present in a given sample, but it is mandatory to know what specific clay minerals, e.g. kaolinite or illite, are present. For this problem, structure-sensitive tools are required. The most powerful and useful of these is X-ray diffraction, supplemented by petrographic microscopy and infrared spectrometry. The techniques discussed below should be used by those thoroughly familiar with the problems of their utilization in mineral identification. The main purpose of presenting a discussion of the different methods of mineral identification here is to draw the attention of the laboratory scientist to those methods which the structural chemist and mineralogist have found most useful. Unless the scientist engaged in waste treatment research is experienced in using the various techniques, it is suggested that the materials to be identified or studied be referred to experts in the various techniques outlined.

III.2.1. X-ray diffraction

Because X-ray diffraction is the most useful mineral identification tool, a brief discussion of its use for this purpose is presented here. The various techniques of X-ray diffraction are described in a number of texts, one of

33 the more complete being Ref. [118]. X-ray methods of mineral identification are discussed in Ref. [120], while those methods devoted to clay minerals are described in Refs [109, 119] . In practice, collimated X-rays of a fixed wavelength are permitted to strike a powdered sample. Only certain atomic planes of the crystalline substances in the sample will be in the proper orientation relative to the incident X-ray beam such that the Bragg condition of nX = 2 d sin 0 is satisfied. In this equation, n is the spectral order of the diffracted radiation, X is the wavelength, d is the distance between like planes of atoms in the substance, and 0 is the angle of incidence of the X-ray beam to the "reflecting" atomic planes. Because the wavelength of the X-rays is fixed, the variable is d, the interatomic spacings, which are unique to a given mineral structure. When different lattice planes are rotated within the X-ray beam, certain angular positions 9 are obtained at which diffraction of the X-ray beam occurs. These diffracted beams are recorded photographically (powder camera) or by means of a scintillation detector (X-ray diffractometer) as a function of the angular rotation 2 0. For the X-ray diffractometer, the output of the scintillation detector is recorded on a strip-chart recorder, a series of peaks of different heights (intensities) being recorded at various distances along the chart. These distances measured in degrees 2 6 are converted to "d spacings" in angstrOm units using the Bragg equation. The relative intensities of the diffracted peaks can be electronically scaled while the peaks are being recorded or visually estimated from the chart. A table of "d spacings" and relative intensities is prepared. To identify the crystalline materials in the sample, the values in the table must be compared with other tables listing similar parameters for known crystalline substances. In addition to the X-ray data listed in the ASTM Powder Diffraction File, X-ray data for the clay minerals are given in Refs [109, 119, 121] . Similar Hata for the zeolites are also given in Ref. [102] . The problem of mineral identification is difficult where several crystalline phases are present in a sample. Recently, a computer program for the identification of crystalline substances has been developed which greatly assists in this problem [122] .

III. 2. 2. Electron microscopy and electron diffraction

These techniques may also be used in certain cases for the characterization of natural substances, but they are not as generally useful as X-ray diffraction methods, nor is the necessary equipment as readily available in most laboratories. These methods will therefore not be discussed here. A useful discussion of these topics is given in Ref. [123] .

III. 2. 3. Petrographic microscopy

The problem of mineral identification is easier the more information there is available about the unknown phases. The microscopic study of minerals in thin sections or as grain mounts provides information which greatly assists in the identification of minerals by X-ray diffraction. Fre- quently the sedimentary zeolites and clay minerals are so fine-grained that they cannot be identified by optical microscopy; accompanying minerals such as quartz and feldspar can, however, be easily identified optically.

34 A knowledge of the presence of these minerals permits the subtraction of their X-ray diffraction patterns from the complex unknown pattern, thus greatly simplifying its analysis. Optical techniques are capable of identifying minute amounts of a mineral phase present in a matrix of other minerals. X-ray diffraction requires at least 5% of a phase to be present before its X-ray pattern is evident. Optical microscopy serves both as a primary identification tool as well as a supplement to X-ray methods. Like X-ray diffraction, optical mineralogy is a specialized field of study and requires some experience in its utilization. The techniques of optical mineralogy are discussed in Refs [124, 125], and the optical properties of the rock-forming minerals are listed in Ref. [109] .

III. 2. 4. Thermal methods

Although most minerals can be identified by the preceding methods, additional instrumental techniques are often useful in certain cases. For problems involving clay minerals, methods such as thermal gravimetric analysis (TGA) and differential thermal analysis (DTA) are especially useful. These techniques also provide thermochemical data on such phenomena as phase transitions and dehydration reactions. Various thermal techniques are described in Ref. [126] .

III. 2. 5. Infrared spectrometry

Infrared absorption spectrometry is an extremely valuable technique for materials characterization. An excellent discussion of this technique is given in Ref. [127], from which this brief summary is taken. The infrared spectrum of a substance arises from the absorption of infrared radiation by vibrating atoms and molecules. The particular absorption spectrum produced is a complex function of interatomic distances, bond strengths and bond angles, as well as the mass of the constituent atoms. Infrared absorption spectrometry is very sensitive to short-range atomic ordering, in contrast to X-ray diffraction which is dependent upon long-range ordering and a periodic repetition of atoms. Thus, unlike X-ray diffraction, infrared spectrometry can be applied to the study of non-crystalline substances, such as glasses, as well as to organic compounds and crystalline materials. In this regard, it should be noted that infrared absorption spectrometry is by far the most useful of the instrumental techniques for the characterization of organic compounds. Like the X-ray diffraction pattern, the infrared spectrum of a substance is unique and serves as a "finger print" of the particular substance. In addition to such qualitative data, infrared absorption spectrometry can yield semi-quantitative data on the amounts of phases. A description of the various techniques used in infrared absorption spectrometry as well as a listing of the infrared absorption spectra for selected mineral groups is given in Ref. [12 7] . As in all instrumental techniques, careful attention to detail is required if good results are to be obtained with infrared absorption spectrometry. With care, a great deal of chemical and structural information is obtainable which powerfully complements those data obtained by X-ray and optical methods.

35 III. 2. 6. Chemical methods of material identification

The previously discussed methods have been mainly structurally oriented, that is, they are dependent upon how the atoms are arranged and less so on which atoms are present. In contrast, chemical methods are dependent upon the particular atoms present. Prior to the advent of X-ray diffraction, etc., standard analytical chemical schemes were used for the identification of minerals. Information on chemical composition is now used mainly to supplement the data obtained by the structure-sensitive methods. Data on the chemical composition of materials are obtainable by a variety of methods, including wet chemical analysis, atomic absorption analysis, emission spectroscopy, X-ray fluorescence and neutron activation analysis. Of these, emission spectroscopy and atomic absorption analysis are most frequently applied to mineral analysis. A description of the standard chemical tests for the identification of minerals can be found in most mineralogy texts, e.g. Ref. [99] . Good summaries of the application of emission spectroscopy and atomic absorption spectrometry to mineralogy are given in Refs [128, 129] .

III. 3. PHYSICO-CHEMICAL CHARACTERIZATION

III.3.1. General remarks

The physico-chemical characterization of natural ion exchange materials with respect to applications in waste disposal of radioactive solutions is relatively easy for pure ion exchange components. The definitions given below can be applied even to mixtures of ion exchangers, if a phenomenolo- gical description of the overall process is sufficient. This becomes difficult, however, for complex mixtures of constituents reacting according to mechanisms other than ion exchange (section II. 2.). This, unfortunately, is often true with soils or natural minerals including those used in local waste disposal.

III. 3. 2. Capacities

III. 3. 2.1. Definitions

The ion exchange capacity gives the number of ions in a definite amount of material, under specified experimental conditions, which is available for the ion exchange process. Because of different possibilities in specifying the amount of material,and different mechanisms and kinds of operation, up to 19 definitions of the capacity are currently used. They are listed and discussed in more detail in Appendix II. For waste disposal applications, the following are relevant: pure ion exchange capacity, sorption capacity, total exchange capacity, apparent or useful capacity, maximum capacity, pure volume ion exchange capacity, sorption volume capacity, total volume capacity, useful volume capacity, breakthrough capacity, and S-, T- or V-Values. In the following discussion, only the three most important, namely, pure ion exchange capacity, sorption capacity, and total exchange capacity are considered.

36 III. 3. 2. 2. Pure ion exchange capacity (Kr)

From a theoretical point of view, this is one of the most clear and satisfactory definitions, though difficult to measure in more complex cases.

The pure ion exchange capacity (Kr) gives the milliequivalents of exchangeable ions in a specified amount of ion exchanger. The latter is by convention 1 g material after removal of absorbed water by drying. For a cation exchanger, it must be the H+ - form, and for an anion exchanger, the CI -form. The absorbed water is removed under specified conditions, e.g. drying at 110°C to constant weight. This need not imply, however, that all the water is desorbed; because of possible irreversible changes during heating of, e.g. natural organic exchange material, the dry weight has always to be determined for separate samples. The H+-form of an ion exchanger can only be used if it is sufficiently stable. To overcome this difficulty, another cation may be used, e.g. sodium. The capacity then has to be corrected for the weight of the H+-form by calculation. Since, for example, the dry weight of the Na+-form exceeds that of the H+-form, too small a capacity will result. It can, however, be correlated to the H+-form of the definition according to:

K Me Me -3 (1) K (EW-1.008) • 10 where Me denotes a metal ion, for which the capacity was determined experi- Me mentally, and EW its equivalent weight in grams. Kr should be given in milliequivalents per gram dry exchanger in the Me-form in order to obtain K^ in the same dimensions for the H+-form. An analogous equation holds for anion exchangers. Here, the OH -form, comparable with the H -form of cation exchangers, is used only very rarely, because of its relative instability for a number of ion exchange materials with respect to time, temperature and C02-uptake from the air. In most cases, therefore, the CI -form is used experimentally.

Ill. 3. 2. 3. Sorption capacity (Kad)

This quantity gives the amount of electrolyte or non-electrolyte fixed by other than ion exchange mechanisms to the material. It is defined as the number of milliequivalents electrolyte or millimoles non-electrolyte which can be adsorbed by 1 g of the dry ion exchanger in addition to exchangeable ions. All values again refer to the dry H+- or CI -forms (see equation (1)). Depending to some extent on the method used for determination of the sorption capacity, it may not only include ions or molecules absorbed by a Donnan- type equilibrium, but also by other mechanisms, e.g. precipitation or mineralization. It is sometimes difficult to separate the sorption capacity from the pure ion exchange capacity. In this case, it may be derived as the difference between the total and pure ion exchange capacities.

37 III. 3. 2. 4. Total exchange capacity (Ktot)

The total exchange capacity is easily obtained by simple equilibrium experiments, because it represents the total amount of ions fixed on the exchange material without reference to the mechanisms involved. Although of limited theoretical interest, this quantity has practical importance for applications in waste disposal.

Ktot represents the total amount of ions absorbed by different mechanisms per gram of the dry material in the H+- or CI -form. These mechanisms can be ion exchange, absorption, precipitation, mineralization, electrolyte uptake or combinations of these effects. Again the capacity for the Me-form can be corrected according to equation (1) for the H+- or CI -form. The total exchange capacity can also be defined as the sum of the pure ion exchange and sorption capacities according to:

III. 3. 2. 5. Procedures

Numerous special procedures have been developed for the determination of capacities for minerals or soils. By no means all of them consider all the different definitions mentioned in Appendix II; no distinction is therefore made in these cases between the different effects mentioned above. Different methods may thus give different results. Nevertheless, such procedures are frequently useful for a comparison of the properties of different materials. In the following, only one chemical and one radiochemical procedure for the determination of the cation exchange capacity are given as examples. For other methods, reference should be made to Appendix II.

(1) Ammonium acetate method. The sample is equilibrated with a IN ammonium acetate solution, containing about a 20-fold excess of the ammonium acetate. The difference in the concentration of the NH^-ions before and after exchange permits the calculation of the total capacity. If the number of the NH^-ions remaining in the clay is to be determined, excess ammonium acetate absorbed in the sample has to be removed first by washing with water. If alcohol is used as a working liquid, too high a capacity will result. As shown by Weiss [139], this is not due to hydrolysis, but rather to adsorption of free ammonia.

(2) 110Ag and 90Sr/90Y method [140] . In contrast to conventional methods, no elution of the exchanged ions is necessary here. The activity of the exchanger loaded with radioactive ions is measured directly to obtain the capacity. Conversion to the desired ionic forms is performed with IN silver nitrate labelled with 110Ag, IN strontium chloride labelled with 90Sr/90Y, or IN yttrium chloride labelled with 9cSr/90Y. In the case of an exchange with Ag+-ions, bright daylight has to be avoided. After centrifuging, the samples are washed with water, filtered, dried and counted. Ion exchange capacities thus obtained agree well with values obtained with conventional methods. For further tracer ions, 22Na, MNa, 42K, 86Rb, 134Cs, and 137Cs are recommended [140],

38 III. 3. 2. 6. Complications

Only carefully pretreated ion exchangers in well-defined ionic forms can be expected to give reproducible results. In order to determine the dry weight of the ion exchanger, the application of temperatures, which may destroy the material to some extent, is sometimes necessary. Numerous factors can affect the determination of capacities, as follows: (1) the rate of the ion exchange process can be so slow that attainment of the equilibrium is difficult to ascertain; (2) capacity can depend on the nature of the counter-ion, e.g. for steric reasons; (3) complex formation of counter- and co-ions takes place in solution as well as in the exchanger; (4) in the case of incomplete dissociation of ionic groups, the capacity will be a function of the solution pH; (5) some methods give rise to the so-called "salt-f-ree water film" (see Appendix II); (6) during the washing procedure, absorbed counter-ions can be washed out by the hydrogen ions of the water (hydrolysis); (7) in addition to equivalent exchange, equimolar exchange can take place; (8) some ions are absorbed so strongly that they can hardly be removed by leaching; (9) conversion of the ion exchanger to the H+-form destroys most mineral exchangers; (10) ion exchange capacity may depend on the particle size; (11) the samples may be soluble to some extent; (12) the nature of the co-ion may have some influence on the capacity; and (13) oxidation or organic poisoning can cause considerable capacity losses. Since so many different factors can influence the capacity of clay minerals, one cannot expect the different methods to give identical values for the same material. It should be emphasized, therefore, that ion-exchange capacities are meaningful quantities only in those cases where all experimental details of the procedure used for the determination are reported completely. Capa- cities of different materials should in any case be compared only if they have been determined by the same method under identical experimental conditions.

III. 3. 3. Selectivity

All the quantities described in this section characterize the preferential uptake of an ion by the ion exchanger with respect to other ions in solution. The property of the ion exchanger to discriminate between different ions is also called "selectivity", and depends on the nature of the sample as well as on the experimental conditions. For a graphical representation of the selectivity, the ion exchange isotherm is most widely used, since it shows immediately the mole fractions of the ions of interest in the solution phase and in the ion exchanger. Distri- bution coefficient, separation factor and selectivity coefficient on the other hand are mathematical expressions of the experimentally observed mole fractions and characterize the extent of selectivity by the magnitude of their numerical value. In the field of radioactive waste disposal, the distribution coefficient is used in most cases, since its experimental determination is relatively simple. The above quantities for the characterization of ion exchange selectivity can be used not only with uniform materials, but can also be applied to mixtures of ion exchangers and to materials which absorb ions according to mechanisms other than ion exchange.

39 III. 3. 3.1. Ion exchange equilibrium

If an ion exchanger is placed in a solution containing ions different from those of the exchanger, a reversible ion exchange reaction occurs in the case of a cation exchanger according to:

A++B+^B + + A+ (3) and in the case of an anion exchanger according to:

A + B" S B + A" (4) where the bars refer to the ions in the solid exchanger. After a given time, an equilibrium distribution of the ions A and B in the solution as well as in the ion exchanger will be established. Convenient quantities to characterize this distribution are the equivalent ionic fractions

•yA and yB of the ions A and B respectively in the solution, and the corresponding values yA and yB in the solid ion exchanger. They are given for the solution by:

7 = ; T = A IAT^+TBT B TAT^[B] ' [A] + [B]-C (5) and for the ion exchanger phase by:

^ = HT+TBT' ^b • [A/+][B] ' [A] + [B] = C (6) where [A] and [B] and [A] and [B] denote the concentrations (meq/ml) of the ions A and B in solution and in the ion exchanger phase respectively. The formulation "per ml solid exchanger" (not 1 g ion exchanger material) is used, because the dimension of a concentration is mol/litre. C and C denote the total concentration of both ions in the solution and in the ion exchanger respectively.

III. 3. 3.2. Ion exchange isotherm

The ion exchange isotherm is obtained by plotting yA against yA, as shown in Fig. 2. Depending on the ions used and the experimental conditions, several types of curve will be observed. The curves 1 to 5 in Fig. 2 are obtained under the following conditions: 1; no preference for ion A with respect to ion B; 2: ion exchanger prefers ion A with respect to ion B; 3: ion exchanger prefers ion B with respect to ion A; 4: ion exchanger prefers ion A at low equivalent ionic fractions of ion A, and ion B at high equivalent fractions of ion A; and 5: ion exchanger prefers ion B at low equivalent ionic fractions of ion A, and ion A at high equivalent fractions of ion A. Except for curve 1, which occurs only in the case of isotopic ion exchange, all the above curves will be observed with mineral ion exchangers.

40 FIG. 2, Ion exchange isotherms.

Ill. 3. 3. 3. Distribution coefficient (K*[)

The distribution coefficient is defined for the ion A as;

A .. [A] _ rAc K (?) d [A] yAC

Since radioactive ions are present in solution and in the ion exchanger in general only as trace components, [A] « C and [A] « C, so that from equation (7) we can write:

A A C - [A] A C *B " Kd C - 15] " Kd C (8)

Equation (8) permits the calculation of the separation factor (a£) from experimentally obtained values of K^. The distribution coefficient is not constant, but depends for a given ion exchanger and a given pair of ions on the concentration C of the solution, the equivalent fractions of the ions in the exchanger and on the temperature.

A III. 3. 3. 4. Separation factor (o^)

The separation factor is another quantitative measure for the selectivity of an ion exchanger and is defined as:

[A] [ 1 Tic A [A] { C - [A]} (9) 7B 7A [A] [1 7JC [A]{C - [A]}

41 The separation factor expresses the preference of the ion exchanger for ion A with respect to ion B. If is greater than unity, then ion A is preferred and if its magnitude is less than unity, ion B is preferred by the ion exchanger. Similarly to the distribution coefficient, the separation factor is not a constant, and depends on the concentration [C] of the ions A and B in the solution, the equivalent fractions of the ions in the exchanger, and on the temperature.

III. 3. 3. 5. Selectivity coefficient (nk£)

This quantity is defined as:

7 %A _ A Tfe nn. B (10) v B lzA l 7A B

where ZA and ZB are the valencies of the ions A and B respectively. As for the separation factor, the selectivity coefficient is a measure of the preference for the ion A with respect to ion B. It is used mainly in theoretical considerations.

III. 3. 3. 6. Procedures

The experimental determination of the concentrations [A] and [A], which

are necessary to calculate yA or 7A, can be achieved either by a batch method or by column operation. In the batch method; the ion exchanger is mixed successively with solutions of different compositions; after attainment of equilibrium, the ion exchanger is separated from the solution. The ions in the solution and also the ions from the ion exchanger after elution are determined quantitatively. From these concentrations, K^ or can be calculated. In column operation, the ion exchanger is converted first to the pure A-form, and then eluted by a solution containing an ion B, which is less tightly bound to the exchanger than A. From the concentration profile of the ion A measured in the effluent, the whole ion exchange isotherm can be calculated [141-144], This method requires, however, that local equilibrium is attained in the exchanger during elution of the column. This can be achieved only if the rate of ion exchange is sufficiently high, the flow rate low enough and the particle size very small. Since these conditions cannot always be satisfied with mineral ion exchangers, serious errors may arise. In several cases of natural local minerals or soils, the small grain size of, e.g. clay particles, prevents column operations at all. In all such cases, recourse has to be made to the more time-consuming batch method.

III.3.4. Kinetic properties

For the application of a natural ion exchange material it is necessary to know not only the ion exchange equilibrium, and how many ions can be bound and which ions are preferred by the ion exchanger, but also the time needed

42 for attainment of equilibrium. If this is too long, column operation becomes difficult and batch operation more convenient, particularly when working with very finely dispersed material as is the case with clays or soils.

III. 3. 4.1. Definitions

The process of ion exchange can be considered as proceeding according to the following steps: (1) migration of the ions in the outer solution; (2) diffusion of the ions through the solvent film adhering to the exchanger grain; (3) chemical exchange reaction of the ions with the fixed ionic groups; and (4) diffusion of the ions within the structure of the ion exchanger. In order to obtain the kinetics of the overall ion exchange process, we have to consider the rates of each of these steps separately. Migration of the ions within the solution is usually achieved by a con- vection process, i.e. by forcing a solution flow through an exchanger bed or by stirring the solution in a batch operation. In this way, concentration gradients of the ions in the solution can be avoided, and process (1) will not become the rate-determining step. Even at a high rate of stirring, an adherent film of solvent persists on the surface of the ion exchanger particle, whereby the movement of the ions is diffusion-controlled only and independent of the rate of stirring. The thickness of this diffusion layer is of the order of 10"3 cm and depends on the rate of stirring. When exchanging only two counter-ions A and B with a common co-ion C, the diffusion process can be described by two diffusion coefficients (Dj), which are functions of the four diffusion coefficients (Djk) of the corresponding ternary mixture of AC and BC in the solvent. The rate of ion exchange for the film diffusion process will also depend on the concentration (C) in solution, the capacity Kr, the surface of the sample, the separation factor, and the thickness (d) of the film. Since two of these quantities, namely D; and d, are functions of the temperature, film diffusion will also depend strongly on this variable. In general, the rate of ion exchange increases with increasing temperature. The rate of the chemical exchange reaction within the ion exchanger is usually fast compared to the diffusion processes. If, however, special chemical reactions are associated with the ion exchange, e.g. chelation with organic groups and metal ions, which are rate-determining, they have to be characterized by the order and the rate constants of the corresponding process. The pure chemical parts of reactions other than ion exchange, e.g. precipitation, mineralization, can of course be rate-determining and have in these cases to be investigated thoroughly. Diffusion within the ion exchanger particle can be described by the so-called "interdiffusion coefficient". This quantity is a function of the self-diffusion coefficients of the corresponding ions and, as shown recently, also of their affinity for the exchanger lattice. Particle diffusion depends also on particle size and on the temperature. It does, however, not depend on the concentration of the ions in the outer solution or on the rate of agitation.

III. 3. 4. 2. Procedures

The simplest way to characterize the rate of ion exchange for a given material and given concentrations of ions is to determine the half-time of

43 ion exchange. The ion exchanger of interest is stirred in a solution containing the radioactive ions and the resulting decrease of the solution activity is monitored continuously as a function of time. The observed time interval, necessary to remove half of the activity which can be absorbed by the ion exchanger after equilibrium is attained, is called the half-time of exchange. It is an experimental quantity which yields diffusion coefficients or rate constants only when the mechanism of the ion exchange or precipitation is known. To investigate experimentally whether film or particle diffusion is the rate-determining step, a method first proposed by Levi and co-workers may be used [145, 146] . After crushing, the grains of the mineral ion exchanger are attached to a glass rod by means of a thin film of glue. The rod is then dipped in a solution containing the radioactive ions to be investigated and is rotated with a motor. For the determination of the fractional attainment of the equilibrium, the activity of the rod is counted in a well-type scintillation counter after different time intervals. If the fractional attainment of the equilibrium is independent of the speed of rotation of the rod as well as of the solution concentration, particle diffusion is the rate-determining step. If an increase in the solution concentration or in the speed of rotation increases the rate of ion exchange, film diffusion is the rate-determining step. Note, however, that by this method, no diffusion coefficients can be determined, since part of the ion exchanger particles will be covered with glue.

III. 3.5. Swelling properties

III. 3. 5.1. Definitions

Mineral ion exchangers can absorb water depending on (1) the nature of the mineral; (2) the nature of the exchangeable cation; (3) the concentration of the solution; (4) the temperature; and (5) the pre-treatment of the sample (e.g. heating). Depending on the experimental conditions, water molecules can be bound to the hydrogen or oxygen atoms of the lattice by hydrogen bonds, and to the exchangeable cations by ion-dipole interaction. In either case, the structure of the absorbed water will be different from that of ordinary water. If layer clay minerals are used, water absorption will also cause a volume increase of the sample. In these cases, changing the ionic form of the mineral in a column experiment by elution will also result in a volume change of the exchanger bed. In order to characterize the ability of a clay mineral to absorb water, the water sorption isotherms of the sample are determined. In general, they are plots of the amount of water sorbed (in mg water/g clay, or molecules water per cell) as a function of the relative vapour pressure of water at a certain temperature. Depending on the material used, water sorption isotherms are in many cases higher if they are determined by desorption of water vapour rather than by an absorption process. These hysteresis effects are especially large for montmorillonite minerals and indicate that a certain activation energy is necessary to change from one stable interlayer distance to another. Another method used to characterize the swelling properties of clay minerals is to plot differential thermal analysis curves. Depending on the energy necessary to desorb the water molecules, these curves will exhibit

44 endothermic peaks. If the water is desorbed from different sites within the samples, multiple peaks will arise. For crystalline minerals, the progress of water absorption can also be demonstrated in many cases by a corresponding increase in the interlayer distance, as shown by X-ray diffraction methods.

III. 3. 5. 2. Procedures

The water sorption isotherms can be determined most conveniently by isopiestic methods. The sample is equilibrated with water vapour of different partial pressures and the corresponding amount of absorbed water determined by weighing. Constant partial vapour pressures are achieved with the help of saturated aqueous solutions of various salts. For differential thermal analysis and X-ray diffraction methods, commercial apparatus is available.

III. 3. 6. Reduction-oxidation properties

Most solid, insoluble and reversible redox systems can be considered as redox exchangers [134, 136], For their characterization, redox capacities and redox potentials can be used. Both quantities, however, can be determined in a meaningful way only if the redox reactions involved are more or less reversible and not accompanied by a leaching of redox components from the material. In most cases, it is impossible to get true thermodynamic potentials. Nevertheless, the capacities and potentials measured under constant conditions can be used for comparison of different materials.

III.3.6.1. Redox capacity

Definition

Analogous to the ion exchange capacity, the reduction-oxidation capacity is defined as the number of electrons lost or taken up by a unit quantity of the material under investigation. The number of electrons can be measured in milliequivalents and the unit of material in grams or millilitres. Similar to the situation with ion exchangers, we have to distinguish between theoretical, analytical, total, reversible, irreversible, maximum, usable, technical and break-through redox capacities. For details, the reader is referred to Refs [134, 136] . In the following, only the total redox capacity is considered.

Procedure

The material under investigation is reduced or oxidized as completely as possible in column operation, washed without exposure to the air in oxygen-free water, and then oxidized or reduced by an excess of a suitable reagent with an appropriate redox potential. The stoichiometric portion of the reagent reduced or oxidized is determined in the effluent by standard analytical procedures. The redox equivalents thus found are related to the dry sample weight. The mineral nontronite, for example, can be reduced from the yellow iron (III) form to the dark-green iron (II) form with dithionite, sulphite, or

45 titanium (III). The reduced form can then be oxidized with hydroquinone, excess of iron(III), or dissolved oxygen in water. In the second case, iron (II) found in the effluent can be determined with permanganate or cerium (IV) (134],

III. 3. 6. 2. Redox potential

The redox potential of a redox system in solution is proportional to the negative logarithm of the equilibrium constant and to the change in free energy of the redox reaction. It can, therefore, be used as a measure of the strength of the reduction or oxidation power of the system under con- sideration. Furthermore, predictions of possible or impossible redox reactions with other systems can be made. The redox potential of a solid redox system cannot be measured directly, since the latter is by definition insoluble. If, however, the solid system is in contact with the solution of a redox system of suitable potential, its redox potential becomes equal to that of the solution after equilibration. Since the potential of the solution can be measured, the potential of the solid phase is also determined. In the following, we will consider an insoluble redox exchanger (A A +9) with a standard redox potential E„ . In equilibrium with the redox system in solution, one obtains [134, 136]:

En0 = E - In (11) n- F [A] where E = potential in solution under equilibrium conditions, R = the gas constant, T = temperature, n = number of electrons transferred per mole, F = the Faraday constant, and [ ] represents activities. In order to obtain the Eso^-potential of the redox exchanger in a particular medium, E is determined by a potential measurement. The ratio [A+]/[A] becomes unity, and therefore the second term on the right-hand side of equation (11) becomes zero if the redox exchanger, at the time of determination of E inthe solution, is 50% in the reduced form and 50% in the oxidized form; semiquinone formation must be excluded. This is achieved either by a redox titration or by mixing 50% of the reduced form with 50% of the oxidized form; both methods have been used [134] . With this method, the thermodynamic standard redox potentials of the solid phase are not generally obtained, but only the apparent Ejq^- potentials in the given medium. This is because not all the activities involved are known and not all the reactions are always completely reversible, and because interfering diffusion potentials may arise. Nevertheless, these

E50

46 IV. MATERIALS

IV. 1. MATERIAL PREPARATION

IV. 1.1. Crushing and grinding

The substances of interest such as silicate rocks are usually obtained in large blocks or pieces. For the present purposes, it is necessary that they possess a large contact surface at a sufficiently uniform particle size. The large pieces are first reduced in size by crushing and the resulting lumps are ground. After grinding, the product must be sieved, usually in a wet suspension, to obtain the proper size fraction in a dust-free state. To minimize the dust fraction it is necessary that the residence time in the crusher be as short as possible. This is attained by the prompt removal of grains of the proper size from the crusher. The insufficiently crushed material either remains in the crusher or is removed, and after screening returned to the crusher (closed cycle) or transported to further crushing machines (multistage process). The multistage process minimizes the losses of material via the fine fraction, and generally leads to better utilization of the energy supplied. Closed-cycle crushing (in conjunction with sieving) is used in small-scale processes or in large-scale processes when a relatively crude final fraction is required. The most regular grains are obtained by the multistage process using sieving between the stages. The greater the size reduction carried out in a single stage, the larger will be the proportion of finer materials. According to Kasatkin, the degree of size reduction in the case of crude and very hard pieces varies between 2 and 6, for medium-size pieces between 5 and 10, for small pieces between 10 and 50, and for fine particles it is greater than 50. According to Rittinger, the work connected with size reduction is proportional to the increase in the surface area. We can compare the work exerted for different degrees of size reduction by using the relation:

Wi : W2 = (Sl - 1) : (S2 - 1) (1) where s is the degree of size reduction (s = D/d, where D and d are the edge of a hypothetical cube, respectively before and after crushing). Measurements with a 1 m3 granite block showed that for crushing to pieces of 0.1 m3, 2 kWh are necessary, while for crushing to pieces of 10 mm3 and 1 mm3, 6 and 20 kWh respectively were consumed, i.e. very much less than indicated by equation (1). The validity of the above and of other existing size reduction rules is dependent on the nature of the crushed material (hardness and cleavage). According to Bond and Wang, the energy (W) necessary for the reduction in size of various materials is related to the degree of size reduction (s),

47 defined as the relation of the dimension of the sieve mesh through which 80% of the incoming material passes, to the dimension of the sieve mesh through which 80% of the crushed material passes; it is also related to the dimension of the particles after crushing (d, inches), according to:

W = k • s^/d* (horse-power • hour/ton) (2)

The value of the constant k is 0. 25 for soft materials, 0. 50 for medium- hard materials and 1„ 0 for hard materials. There are available a broad range of crushing machines for specific purposes. Jaw-crushers are generally most suitable for crushing very crude pieces into about centimetre pieces (and above). The hammer (impact) mill crusher is usually applicable both for crude crushing (size reduction ~ 10- 15) and also for fine crushing (size reduction ~ 30 - 40). A cog-cylinder crusher has applications similar to those of a hammer mill but has advantages in crushing medium-hard and cleavable material (e.g. limestone). A cone crusher (with grooved eccentric truncated cones) can be used both for crude and fine (except very fine) crushing and it can be properly adjusted to either operation. Operation of the cone crusher is smooth and economical, but it is not suitable for soft materials such as gypsum since it gets clogged easily. Rotating wheel mills are applicable both for medium-hard and soft materials, for crude and fine crushing (grinding included), under both dry or wet conditions; they supply fine particles penetrated with dust and under proper circumstances a homogeneous, finely-grained mixture of a few components can be produced. By grinding in mills, material in the form of powders is obtained. The ball mill is very frequently used for grinding medium-hard and soft materials. The optimum operation of a ball mill requires that the rotation rate of the drum containing the balls and ground material is carefully adjusted. For drums of diameter (D) under 0.8 m, the optimum number of rotations per minute (n) is given by the relation n = 37/s/b, and for larger drums by n = 31 /%/b. The diameter of the balls (!)[,, mm) increases with that of the

material to be ground (dm, mm). The most suitable values according to Razumov are given in Table X. The optimum mass of the balls (Q,, kg) can be roughly calculated by the relation of Perov and Brand: Gj, = 37.7 (O • D2* L), where O is the volume of the balls as a percentage of the drum volume, D is the inside diameter of the drum (m), and L is the length of the longitu- dinal axis of the drum (m). The assumed apparent weight of the balls is 4800 kg/m3. The weight of the crushed material amounts to ~ 8 - 10% of the whole weight of the drum contents, and can therefore be neglected in the above rough calculation. The optimum ratio of drum length to diameter is 1.56 - 1.64. In the ball mill the process can be made most economical by the continuous removal of the proper sieve fraction. This is done either by dry screening in the drum or by washing out the fine fraction with flowing water. Other types of crushing and grinding equipment are available and are described in the literature.

IV. 1.2. Sieving (screening)

The separation of particulate solids according to grain size can best be done by sieving (screening). Separation of grains of sizes smaller than

48 about 0.05 mm is, however, best accomplished in water suspensions utilizing the different sedimentation rates (see section IV. 1. 3.). Sieves are characterized by mesh size, mesh number per unit length, and wire size. There are two principal systems in use, one in which the mesh size increases arithmetically (i. e. 0.1, 0.2, ... 0.9, 1.0 mm) and one in which the mesh size increases geometrically. Sieves for which mesh sizes increase as an arithmetic series have a serious drawback when handling extremely fine material. The sieve system based upon mesh size increasing in a geometric series has a relatively even gradation in aperture size. The Tyler scale of sieves is based on this system; 200-mesh for example indicates that there are 200 mesh (spaces) in the unit length of 1 in. The basic modulus (coefficient of the geometric series) of the Tyler system was originally \/~2 (and eventually became 2\ for a finer gradation). The Tyler sieve scale (modulus = 2|) is given for the metric system in Table XI. The current Tyler sieve series uses a ratio of screen opening sizes (linear dimensions) varying from one screen to the next by a factor of 2j, corresponding to an aperture area ratio of 2|. Sieve analysis is accomplished by screening the crushed or ground material through a series of sieves (Tyler system, DIN, GOST) so that the grains are divided into separate sieve fractions. This analysis allows the assessment of the mass distribution of different grain sizes in the given material, crushed or ground by a certain size reduction procedure. A logarithmic plot of the weight portions against the mean particle size within each sieve fraction yields roughly a linear dependence, if the finely ground material has the same crystal texture. We can therefore approximately extrapolate the portion of small grains in the mixture, if a sieve system with constant modulus (such as the Tyler system) was used. The main factors governing the screening effectiveness are;

(1) The shape and dimensions of the sieve spaces (mesh) and the shape of the material grains; for round particles, round sieve spaces can be used, while for elongated particles, screens with oblong spaces are preferred. (2) The thickness of the crushed material layer on the screen, the kind of screening and the residence time on the screen. The thinner the material layer, the more efficient is the screening; however, the output drops unless residence time is decreased. Decreased residence time causes decreased screening efficiency. In general, therefore, the charge on the screen must be as uniform as possible, and the material on the screen must be agitated to enhance contact of the grains with the screen surface; consequently the screen is shaken or vibrated. (3) The moisture of the screened material; moisture promotes grain agglomeration which increases the apparent grain diameter and promotes clogging of the sieve mesh. (4) The effect of electrostatic forces disturbs the sieving process by promoting adhesion of particles to the pad. (5) The mesh size must be slightly greater than the desired grain size. The large scale screening of finely grained material is usually done on vibrating screens or drum screens. For detailed information, the reader is referred to the specialized literature.

49 TABLE X. DIAMETERS OF MATERIAL TO BE GROUND IN BALL MILL AND OF BALLS TO BE USED

dm (mm) 38-53 27-38 13-19 6.7-9,5 4.7-6.7 2.4-3.3 1.2-1.7 0.6-0.8

Db (mm) 100 89 70 57 49 40 31 25

TABLE XI. TYLER SIEVE SCALE (MODULUS = 2i)

Mesh No. 3 4 6 8 10 14 20 28 35 48 65 100 150 200 270 400

Mesh size (mm) 6.68 4.699 3.327 2.362 1.651 1.168 0.833 0.589 0.417 0.295 0,208 0.147 0.104 0,074 0.053 0,038 IV. 1.3. Washing

The larger grains obtained by fine crushing or grinding contain a certain portion of very fine particles (dust), unless they are carefully screened. However, dry screened particles still have some dust sticking to them, which is best removed by washing in a stream of water. This process can be accomplished in many devices, the application of which depends upon specific circumstances, especially upon the sedimentation rate of the single particles. This in turn depends upon the particle size and density, so that we may separate particles differing either in size or density. If the particles differ in both, it may happen that their sedimentation rates will be identical, and consequently they cannot be separated by washing procedures. The possibility of separating particles (a, b) differing both in size and density can be qualitatively assessed by means of the sedimentation diagram, an example of which is given in Fig. 3. For the construction of this diagram, it is necessary to experimentally determine the sedimentation rates of species (a) and (b), differing in density, as a function of the particle diameters. Let us consider a system containing the substances (a) and (b) of diameters between M and N in Fig. 3. The smallest particles of (a) sediment faster than the largest and most rapid particles of substance (b); in this case these substances can be separated. For cases given by points R and S, the conditions are far more complicated and one cannot separate all particle sizes of (a) from all particle sizes of (b). However, when washing out the crushed or ground substance from its powdered fraction, the case is very simple compared with that in Fig. 3, since we are working with particles varying only in size. The question then becomes one of washing the given batch of material with the smallest possible volume of water, i. e. most economically. The quantity of ground material to be washed out and its mean particle size determines the type of equipment to be used. Smaller lots of materials with grains larger than about 0.2 mm are usually washed during the operation

PARTICLE DIAMETER

FIG. 3. Sedimentation diagram: sedimentation rate vs particle diameter.

51 of column filling; the water suspension of the degassed sorbent is inserted in small lots into the column with water flowing upwards at rates sufficient to remove the fines. This operation is successful and economical if the content of fines is small (screened material). If the material contains a substantial portion of fines (unscreened crushed or ground material), the process is not sufficiently effective and results in bed channelling. If channelling occurs, the stepwise insertion of degassed suspension is supplemented with a subsequent wash of the whole bed until the flowing water is clear of fines. If the quantity of material is large and/or if the column design does not permit use of the previous operation, it may be advisable to remove the fines in hydrocyclones, which utilize centrifugal force. The inherent advantage of hydrocyclones is their large output, and consequently the superior economy of the process. The operation of a hydrocyclone (see Fig. 4) is as follows: the suspension enters tangentially to the periphery of the upper cylindrical part of the cyclone (1), and thereafter it follows a spiral descending path along the cone-shaped walls. In the lower part (2), the water stream advances to the centre, rises again and leaves the cyclone in the centre orifice of the upper cover (3). The solid phase accelerated by the centrifugal force is driven to the sloping inner walls where it slips down by gravity into the bottom narrow part of the cone (2), and thus is removed from the cyclone.

FIG. 4. Hydrocyclone.

IV. 2. CHEMICAL AND HEAT PRE-TREATMENT

IV. 2.1. Chemical pre-treatment

By treating various inorganic and organic substances with strong mineral acids, such as sulphonic and phosphoric acid, alkaline hydroxides or salt solutions, under simultaneous or subsequent heat treatment, a product

52 of better sorption properties is usually obtained as compared with the starting material. Since the reaction mechanism involved varies with the chemical nature of the substrate and reacting solution, such treatment is usually known as "activation". If the acid reacts chemically with the substrate, the salt-like product of the anion of the oxy-acid with the cation or hydrolysed cation of the solid body is generally obtained. For example, if aluminium or iron oxides are contacted with phosphoric acid, a certain type of aluminium or ferric phosphate is obtained. The hydrogen, hydroxy-phosphate of aluminium is a potential sorbent of cations in neutral and slightly acidic media, while aluminium oxide or iron oxide are not. In other cases, especially if the substance does not react with the anion of the acid, only the dissolution of the acid-soluble component may be expected. For example, if bentonite is treated with concentrated hydro- chloric acid, the aluminosilicates decompose with aluminium chloride formation, and essentially silica remains. Insofar as the sorption ability is connected with the existence of the aluminosilicate structure, as is the case with strontium sorption for example, then the sorption capability breaks down. If dilute acids are applied, the aluminosilicate may not break down completely, but cations of bases are substituted for hydrogen, probably together with the partial decomposition of the aluminosilicate framework. For example, activated bleaching earth is obtained if bentonite or similar clays are treated with dilute HC1 or H2SO4 at elevated temperatures. Such treatment also removes the undesirable components and loosens the texture of the clay. The resulting state has much in common with the "active" state in that the specific surface area and surface imperfections are increased, with a corresponding increase in the reactivity. The resulting "clay acid" reacts by ion exchange with electrolytes, while salt-like compounds may arise by reaction with molecularly-dispersed substances. In reaction with colloidal substances, coagulation can take place, essentially by ion exchange [ 54]. Such "clay acids" may be viewed as a product of decomposition and hydrolysis, i.e. as having both H+ and Al3+ (or (A10H)2+) in the outer layer, containing exchangeable bases. The aluminium ions (or their hydrolysed forms) are generally not exchangeable. The natural clays and zeolitic materials, for example clinoptilolite, frequently contain insoluble calcium carbonates which may interfere with the base exchange in certain cases. The treatment of such materials with dilute acid (10% HNO3 or HC1) together with subsequent washing removes the interfering component prior to use as a base exchanger. Humus-rich coals, especially lignite, contain acid hydroxyl and carboxylic groups. They are relatively stable in acidic media, but easily peptize in alkaline media. The ability to peptize is usually associated with their solubility in the alkaline medium. Various organic materials, such as wood, peat, brown coal, black coals and anthracites, can be artificially humified by means of water-extracting and partially oxidizing substances. The product possesses many properties of activated carbons, and moreover has ion exchange properties. Treatment with sulphuric acid creates the substantially dissociated sulphonic acid groups by sulphonation, in addition to the carboxylic and hydroxylic groups created by the humification process. Then the base exchange capacity in an acidic or a neutral medium is essentially determined by the sulphur content. Sulphonation is carried out using

53 sulphuric acid or its derivatives, such as oleum, pyrosulphuric acid, potassium disulphite, chlor-sulphonic acid or amidosulphonic acid. When the coal-containing substance is contacted with one of the above reagents, the reaction starts suddenly and heat evolution maintains the temperature between 150 and 250°C. Activation is always accompanied by a volume increase and a deterioration of mechanical strength of the material treated, which results in a fine-grained or even pulverized product. Prior application of dilute reagents often leads to a more uniform and solid product. Acti- vation can also be effected by other strong mineral acids, such as nitric acid and especially phosphoric acid. In the latter case, there is neither sulphonation nor oxidation and consequently high temperatures can be applied (500 - 800°C), and the product, beyond having base exchange ability, also possesses the properties of activated carbon.

IV.2. 2. Heat pre-treatment

The reactivity of solids is determined primarily by the magnitude of the specific surface area that is in contact with the other reacting phase, but it is also affected by the character of this surface, especially the degree of structural imperfection. The surface area and structural imperfections are in turn largely determined by the preparation, i. e. the history of the sample. For example, lime obtained from calcining calcite at 900°C reacts violently with water. However, lime formed at 1600°C hydrates very slowly (dead burnt). Similarly, magnesium carbonate yields an active oxide if calcined at 400 - 600°C, and an unreactive oxide from calcination at 1000°C (dead burnt). Active substances may thus be prepared by the decomposition of other solid substances by heat treatment. A substance with a large and imperfect surface area can also be prepared by very fine grinding, or by condensation from the gaseous phase, e.g. carbonyl of iron produces a finely dispersed reactive oxide of trivalent iron. Precipitation from the liquid phase also produces substances with large specific surface areas, if the conditions are such that the forming precipitate has a high nucleation rate and a relatively low crystal growth rate. By drying such a precipitate, the substance in an active state can be prepared. The activity of such substances, however, diminishes with time. By heat treatment to a certain temperature, an active state can be created, and by further heat treatment to higher temperatures, this active state is destroyed (e.g. active and dead burnt lime). Active lime is composed of minute crystallites of CaO 0. 3 nm) forming a pseudomorph of calcite crystal. Between 1150 and 1200°C, a partial sintering of the CaO crystallites takes place and larger crystals are formed. At 1400°C, the apparent density of the product is near to that of CaO and the activity is at a minimum. With MgC03, the processes are qualitatively the same, but are shifted to lower temperatures. There are numerous reports of heat-treated substances being used as sorbents; the following are only a few examples. According to research carried out in the USSR [55, 59, 90], dolomite calcined at 720-750°C into the "magnesium mass" formula (MgO • CaC03) can be used for the decontamination of radioactive liquid waste containing 9(Sr, 95Zr, 95Nb, 106Ru, 144Ce,and 137Cs. The mechanism involved is assumed to be as follows: strontium removal is by chemisorption on the grain surface and co-precipitation with recrystallization of the magnesium mass; cerium

54 and phosphorus are adsorbed on the surface of the material due to the formation of weakly soluble compounds (cerium hydroxide and phosphates of calcium and magnesium), and caesium removal is by means of fixation on the silicate components of the dolomite. As a result of laboratory studies carried out with dolomite [ 147], pre- liminary information for designing a pilot plant is available. These data indicate that: (l)itis more advantageous to use the sorbent in dynamic conditions; (2) the size of the sorbent grains must be 0. 5 - 1 mm; (3) the pH of the solution does not have a significant effect on the decontamination factor; (4) the liquid flow rate through the column should be 1. 5 m/h; (5) the height of the sorbent bed should be 1.5 m; (6) the overall decontamination factor to be expected for betas is about 25; and (7) almost complete removal of 144Ce and 106Ru is achieved with the passage of 490 and 700 column volumes respectively. In the USA, efficient sorbents have been obtained by the heat treatment of hydrous oxide minerals. A selective sorbent for strontium was obtained at ORNL by heating gibbsite (Al(OH)3) above its decomposition temperature at 150°C. The procedure resulted in formation of aluminium oxides with high specific surface areas (greater than 200 m2/g), in contrast to the low 2 90 specific surface area (about 0.3 m /g) for gibbsite. The KDfor Sr ranged from 4000 to 40 000 when the solid-to-solution ratio was increased from 0.001 to 0.05 g/50 ml. A raw aluminium ore (bauxite from Arkansas) containing 40% gibbsite also exhibited good strontium sorbing properties when the gibbsite component was decomposed by heating. The product was applicable for neutral or slightly alkaline solutions. In distilled water the capacity of the product was approximately 4 meq/100 g for caesium and 12 meq/100 g for strontium. The amount of caesium sorbed remained constant over the pH range 6-10, and caesium could be readily leached with an elutriant containing 0. 01N NaN03. The passage of sodium ions equivalent to 1. 7, 3. 3 and 23 times the concentration of caesium sorbed on the column resulted in the removal of 57, 83 and 96% respectively of caesium. Removal of strontium was more difficult: the passage of

0. IN NaN03 equivalent to over 100 times the concentration of strontium on the sorbent resulted in only 50% of the strontium being desorbed; the passage of sodium ions (as IN NaN03 solution) of over 730 times the strontium concentration desorbed 77%, and the passage of 5N NaN03 solution equivalent to about 7000 times the concentration of strontium on the sorbent was necessary for the removal of over 95% of the strontium. The difficulty with which strontium is desorbed is evidence of a strong bond between the sorbent and this element. However, strontium was desorbed quite readily when the pH of the sodium nitrate solution was dropped below neutral [ 148, 150, 153 ]. Another selective sorbent of strontium was obtained by the heat treatment of limonite giving a product with a specific surface area of 93. 2 m2/g compared to 20.8 m2/g for the original mineral. The heat-treated limonite removes strontium best from neutral and weakly alkaline solutions.

Strontium uptake was about 99.6% even from 0.1M NaN03 solutions containing IX 10"5 MSr2+ as carrier. Heat-treated limonite is more selective for strontium than heat-treated gibbsite, and is probably the most efficient of all the heat-treated hydrous oxides. The efficiency of natural limonite as an adsorber is proTaably due to its relatively large surface area [153]. In the Czechoslovak Socialist Republic, a selective sorbent for strontium was obtained by the heat treatment of barium sulphate with calcium sulphate

55 above 1000°C, with subsequent rapid cooling. The product is a metastable

substance in which the CaS04 component adjusted structurally to the BaS04 upon rapid cooling, giving rise to sorption-selective sites for cations, the sulphates of which are isomorphous with barium sulphate (radium, strontium, etc.). The product is efficient for the uptake of radiostrontium even from solutions containing substantial amounts of soluble calcium salts; uptake is enhanced by the presence in solution of sulphate ions. This product is best used in a finely ground state (particles under about 0.06 mm diameter) in the form of slurries with the radioactive water to be decontaminated. A s Kd of up to 10 can be obtained in the case of one-stage addition of the sorbent, and may be increased to 106 if similar amounts of sorbent are added in a few increments without removal of the existing admixture from the reaction vessel. The reaction mechanism of this activation process involves

the polymorphy of BaS04 and CaS04 and the different miscibilities of these polymorphic components with SrS04.

IV. 2. 3. Pelletizing

The non-uniform size distribution and shape of mineral fragments makes it difficult to standardize the operating characteristics of mineral ion- exchange columns. This problem of poor hydraulic properties can be circum- vented by pelletizing the minerals. A technique for the conversion of clinoptilolite to uniform pellets has been described [18]. Attapulgite clay in a sodium hydroxide solution is used to plasticize and bind the finely ground clinoptilolite particles. The resultant plastic mass is extruded and dried to form pellets, which have good chemical and physical stability. The use of calcium chloride and sodium silicate with attapulgite rather than sodium hydroxide may be preferable since the latter tends to dissolve the clinoptilolite.

IV. 2.3.1. Heat treatment of bentonites

The caesium selectivity of bentonite can be improved by potassium saturation and/or heating to 500 - 700°C [ 20, 21 ]. It is thought that this improved selectivity arises from the collapse of the bentonite structure to a 10A spacing as a result of the above treatment. The significance of the 10A spacing on caesium sorption has, however, been contested [27].

IV. 2.3.2. Heat treatment of alumina and clinoptilolite

Heating of various hydrous alumina compounds to 400°C has been shown to greatly improve the strontium selectivity of these substances [ 1]. Heat treatment of clinoptilolite also increases the strontium selectivity of this mineral [ 1] . The mechanism whereby heat treatment improves strontium selectivity in these materials is not clearly understood, but compound formation is considered to be a possibility in the case of the alumina compounds.

56 V. PLANT SCALE APPLICATIONS

V.l. GENERAL

The principal techniques that have been considered at various installations for use of clay minerals on a plant scale can be broadly categorized under two headings: (1) use of clay minerals as ion exchangers in both the batch and column type of contacting devices; and (2) use of clay minerals as additives and product conditioners as well as barriers in disposal pits and trenches.

V.1.1. Ion exchangers

V.l.1.1. Batch process

The use of the mixer-settler type of unit has been considered for batch contacting of waste solutions with clay materials. This technique has not, however, been adopted for regular operation on a large scale at any installation. It should be noted that the extent of ion exchange in this method is limited by the selectivity of the mineral under equilibrium conditions, and therefore, unless the selectivity for the radioactive ion is very favourable, the efficiency of removal will be poor.

V.l. 1.2. Column operation

(1) Fixed-bed column: Operation of a single column or a series of columns as fixed-bed ion exchangers appears to be the most commonly followed method on a plant scale. Although column operation is essentially a large number of batch operations, it is much less dependent upon the selectivity than the batch contactors, and theoretically a column using a clay mineral can be considered effective regardless of the selectivity coefficient, since large quantities of the clay minerals can be obtained for use at low cost. Accounts of the operational experiences of some installations using the fixed-bed type of clay mineral exchangers are given in section V. 2.

(2) Centrifuge: A modified fixed-bed column method, using a centrifuge, has been developed at Harwell, UK, and Mol, Belgium.

(3) Moving-bed contactors: The use of moving-bed, continuous counter-current exchanger, or pulsed-bed contactors for treatment of wastes on a plant scale using mineral sorbents is not known.

57 TABLE XII. ION EXCHANGE PLANT DATA

Capacity of „ . , Bed dimensions Country Mineral , Particle size . , Flow rate Principal isotopes Decontamination mineral , Diameter x height s „ 2 , pv H (Plant) used (mm) /(m3/v /h, - m) removed factor (meq/g) (m)

United Kingdom Vermiculite 0.60 0.5-1.0 1.52 x 0.45 0.63 Cs, Sr 11.5 25 (Harwell)

India Vermiculite 0.6 0.4-0.8 1.8x0.9 1.29 "Cs, 90Sr 10.5 10-15 (Trombay)

United States of America Clinoptilolite 0.35 0.25-2.0 0.56 x 1.2 0.24 "Cs, Sr 8.2 (National Reactor Testing Station (NRTS), Idaho)

Federal Republic "Filtrolit" 0.22 0.3-1.0 0.3 x 1.8 7.1 "Cs 500 (max) of Germany (Pelagonite tuff) 22 Czechoslovak Baiyte 1.0-3.0 3.4 x 2.0 4.0 %a 3.3 Socialist Republic V. 1. 2. Additives and product conditioners

Clay minerals have been utilized in certain installations as additives to adsorb and immobilize radioelements that are fixed in cement or bitumen. Available details on this aspect of clay mineral utilization are also presented below.

V. 2. OPERATIONAL EXPERIENCE

In this section accounts are given of the operational experience of installations in various countries which use clay minerals on a plant scale. Relevant data on these ion exchange plants are given in Table XII.

V. 2. 1. United Kingdom (Harwell)

For several years, a treatment system using a sludge blanket precipitator and a vermiculite column was in operation at Harwell, UK. After a calcium-iron phosphate treatment at pHll, the supernatant was passed at a rate of 2. 3 m3/h through two parallel columns, each 1. 52 m in diameter with beds 0. 45 m deep. The apparent contact time was 40 min, i.e. a flow rate of 1. 5 bed volumes/h. Grade 1 vermiculite was used. Since the ion exchange efficiency was impaired with treatment at high pH, the vermiculite was converted to the Na-form. The cation exchange capacity of this product was 0. 6 meq/g. Operations over a period of more than nine years showed that the columns could treat 800 bed volumes before a break-through occurred. The decontamination factor achieved by the vermiculite was 25, giving an overall factor of 200 for the two steps. After 1400 bed volumes these figures dropped to 3 and 33 respectively. The particle sizes of the grade 1 and grade 0 vermiculite used at Harwell are given in Table XIII. Disadvantages experienced in the plant were as follows: (1) Due to poor flow characteristics and the low rate of exchange, large columns were needed in order to give the required flow rate, while still

TABLE XIII. PARTICLE SIZES OF GRADE 1 AND GRADE 0 VERMICULITE USED AT HARWELL, UK

Percentage by weight Mesh range Grade 1 Grade 0

16 0.8 0

16-30 89.6 2.9

30-72 8.1 54.8

72 -100 0.9 28.8

100 -200 0.3 10.8

200 - 240 0.2 1.3

240 - 300 0.1 1.4

59 FIG. 5. Flow diagram of centrifuge ion exchange plant, Harwell, UK. maintaining the long contact time necessary for efficient ion exchange. (2) After a short time, the columns were sufficiently contaminated to be a radiation hazard, requiring the use of shielding. (3) Emptying the columns after exhaustion of the vermiculite was a very difficult operation. Digging out was impracticable because of the radiation; thus the only acceptable method was fluidization. This, however, was not very efficient and increased the volume discharged by an amount equivalent to the water necessary to displace the bed. In order to overcome these difficulties, a basket-type centrifuge was developed for the ion exchange treatment [82, 149]. The centrifuge is a vertical discharge machine with a perforated basket, 1. 22 m in diameter, 0. 36 m in height and 0. 15 m in depth. It is connected to two feed systems, one for supplying the effluent to be treated and the other for charging vermiculite to the basket. Discharge is made by a hydraulically operated traversing plough which pushes the vermiculite down the discharge tube. Figure 5 shows the flow diagram of the centrifuge plant [107]. Before use, the vermiculite is washed free of fines in a tank and converted to the Na-form in the centrifuge. The plant used at Harwell has a capacity of 2.3 m3/h. The effluents are pre-treated by a calcium-copper ferrocyanide precipitation in a sludge blanket precipitator. Since a decrease in particle size leads to an increased rate of exchange, the finer grade 0 vermiculite is used in the centrifuge to compensate for the shorter contact times. The bed is about 3. 8 cm thick (50 litres). The results obtained with the basket centrifuge have so far proved promising. It should be mentioned that the very high concentrations of sodium ions tend to shorten the life-time of the vermiculite. Recent work has shown that the varying composition of the waste and the chemical treatment used

60 are the main causes of the occasional problems experienced. In general, however, the procedure has proved satisfactory and is being continued at Harwell.

V. 2. 2. India (Trombay)

The ion exchange unit at Trombay, India, has been in operation since June 1966 [19,91]. With a flow-through capacity of 225 m3 of waste per day, the unit consists of eight cylindrical mild steel columns of 1. 8 m diameter, four of which operate in parallel while the other four constitute a stand-by. Each column holds about 2. 5 t of natural vermiculite of particle size 20 - 40 mesh. The flow diagram of the ion exchange unit is shown in Fig. 6. When the vermiculite bed is exhausted, it is loosened from below using compressed air and water under pressure. The vermiculite is then transferred as a slurry by gravity feed to an underground sump where de-watering of the slurry by settling is carried out. The vermiculite containing 40-50% moisture is collected in 45 gal steel drums with a 2 in. concrete lining. Since the level of radiation at the surface of these containers does not exceed 10-15 mR/h, there is no need for any additional shielding. It has been found that, although vermiculite is very effective for the preferential uptake of radiocaesium, the overall performance of the ion exchange unit is dependent upon the type of chemical treatment adopted in the first stage. Operational data averaged over a period of six months are summarized in Table XIV. It was observed that if a ferrocyanide complexing was combined with the phosphate treatment step, the pH had to be maintained around 9. Under these conditions, Sr removal was not very satisfactory and also the Cs decontamination in the succeeding ion exchange step was not complete. Consequently, the ferrocyanide treatment was omitted and phosphate flocculation was carried out at pHlO. 5. Both Sr removal in this step and Cs decontamination in the vermiculite columns thereby improved considerably. Another modification that was found necessary was to reduce the flow rate in the ion exchange columns; instead of the initially used rate of 2. 58 m3/h • m2, the plant is now operating at a flow rate of 1. 22 m3/h • m2. Under the present conditions, a decontamination factor of 10-15 for the ion exchange unit is obtained, and each column can conveniently handle about 850 - 1000 bed volumes of waste before exhaustion. In general no problems of radiation hazard have occurred in the ion exchange unit operating with a maximum feed concentration of 1 X 10"2/jCi/ml. The maximum radiation levels on the surface of the columns were between 250 and 400 mR/h.

V. 2. 3. United States of America

V. 2. 3. 1. National Reactor Testing Station (NRTS), Idaho

The design of the unit containing the mineral clinoptilolite was based on the results of laboratory work at the Idaho Chemical Processing Plant (ICPP) [44] and other locations [151,152]. This unit, called the Isotope Removal System (IRS), consists of four columns in parallel, each column

61 WASTE LIQUID FROM CLARIFLOCCULATOR

VERMICULITE FEED TANK

LIQUID TO MONITORING AND DISCHARGE

DE-WATERING

WET VERMICULITE (50'/. MOISTURE) FOR MIXING WITH CEMENT AND DISPOSAL FIG. 6. Flow diagram of ion exchange plant, Trombay, India. TABLE XIV. OPERATIONAL DATA OF ION EXCHANGE PLANT, TROMBAY, INDIA

Ion exchange decontamination factor Overall decontamination factor Type of chemical treatment Cs Gross Cs P Gross

75 ppm

P04" = 130 ppm 29.8 8.1 1.8 8.6 66.0 169.9 19.8 40.4

Fe" = 25 ppm Phosphate CaJ+ = 150 ppm

3 P04 " = 300 ppm 10.0 1.54 1.1 3.8 45.6 264.9 11.6 56.5

Fe" = 50 ppm

Ca2^ 150 ppm

PO°- = 300 ppm

4 Fe(CN)6 - = 60 ppm 10.6 2.6 2.0 7.6 37.0 164.5 34.8 35,6

Fe 50 ppm

Phosphate Cu" = 40 ppm +

Ferrocyanide Ca* = 75 ppm

PQ.- = 130 ppm

4 Fe(CN)6 - = 30 ppm 3.7 2.3 2.7 2.8 55.3 40.6 12.5 14.8

Fe" = 25 ppm

Cu" = 20 ppm

a Overall decontamination factor includes that obtained in the chemical treatment unit of the plant. SUPPLY HEADER

FIG. 7. Clinoptilolite column consisting of two drums connected in series.

comprising two 200 litre mild steel drums connected in series. Each drum contains about 60 cm of clinoptilolite making an effective total column length of 1. 2 m. The bottom drum, when exhausted, is drained and used to replace the top drum. When removed from the system, the top drum is capped, the drain on the bottom is sealed and the drum with its contents is buried with other solid radioactive wastes. A single unit consisting of two drums [51] in series is shown in Fig. 7. The total life of each drum is 800 column volumes, and the average decontamination factor achieved is approximately 200. The clinoptilolite has a high selectivity for 137CS (and 90Sr), and the equilibrium distribution coefficient for 137Cs over the pH range 1.0- 10. 0 was found to exceed 1000. For the columns at Idaho Falls, the clinoptilolite is crushed to the size range 0. 25 - 2. 00 mm diameter; initial use of a larger size (1-6 mm diameter) resulted in unsatisfactory removal of 9°Sr. Each month, approximately 375 000 litres of the waste water produced at ICPP is decontaminated by clinoptilolite. The average composition of such waste is summarized in Table XV [44].

V. 2. 3. 2. Oak Ridge National Laboratory

At Oak Ridge, 'Grundite1, a commercial variety of illite, is being used as a caesium sorbent on a plant scale in the soda-lime softening process for liquid wastes [ 25,26 ]. Illite has also beenused as an additive in the process of solidifying wastes in a cement block, and in the same process attapulgite clay has been applied as a rheological conditioner [154]. Either attapulgite or bentonite is added as a conditioner to TBP wastes at Oak Ridge for inclusion either in asphalt or polyethylene [155], In this application, the purpose of the clay is to sorb the TBP so that it may be

64 TABLE XV. COMPOSITION OF WASTE TREATED BY CLINOPTILOLITE AT THE IDAHO CHEMICAL PROCESSING PLANT, USA

Chloride 0.5 mg/1

Nitrate 200 mg/1

Calcium 40 mg/1

Magnesium 12 mg/1

Sodium 95 mg/1

137Cs (0.8-1.3) x 10"3 jiCi/ml

90Sr (3.5-10) x 10"4 (jCi/ml

PH 8.2

included in the asphalt. The general proportions of the waste mixtures studied [156] are: 30% asphalt, 10-30% TBP and 40-50% bentonite or attapulgite. In the incorporation of wastes in asphalt, it has been found that leaching of caesium from the product is considerably reduced by the addition of 2% by weight each of Grundite and sodium metasilicate [156],

V.2.3.3. BattelleNorth West Laboratory, Richmond

Mordenite has been used at Battelle North West as a caesium sorbent in the cement mortar method of waste immobilization [77], In this work, the leaching of caesium from a cement mortar consisting of 540 g Portland cement, 11 g bentonite, 5 g synthetic mordenite and 300 ml water was investigated. It was observed that the addition of mordenite reduced the initial leach rate of caesium from 0. 47 g cm2 day"1 for untreated mortar to 0. 0086 g cm2 day"1 for treated mortar.

V. 2.3.4. Savannah River

Bentonite mineral has been used to provide a water-proof barrier around buried solid wastes to reduce the leaching of radionuclides. This technique has been investigated at the Savannah River facility [157], but does not appear to be routinely used at this or other installations in the USA.

V. 2. 4. Federal Republic of Germany

Prior to its use for removal of radiocaesium, the 'Filtrolit' (pelagonite • tuff), which is heterogeneous and contains several silicate minerals, is treated with sodium chloride which significantly increases its capacity. Pilot plant studies [ 107, 158] have shown that this material has increasing selectivity, not only for Cs but also for Sr. Ba and Ce, with decreasing relative concentration of these ions in the solutions. The pilot plant comprises two columns, 18 and 30 cm in diameter and 180 cm in height, filled with

65 45 and 100 litres respectively of the tuff, and has been used for the decontamination of various laboratory effluents. Using 0. 3 - 1.0 mm tuff, it has been found that the decontamination is very good for normal effluents from chemical laboratories (DF up to 500), but is rather poor in the case of laundry effluents, with a DF value of 15 - 20. For laundry effluents and strongly acidic solutions, the tuff becomes irreversibly damaged.

V. 2. 5. Czechoslovak Socialist Republic

Pilot plant studies using a number of minerals have been carried out in the Czechoslovak Socialist Republic, but the industrial plant scale use of natural minerals for liquid waste treatment has been limited to barytes [159], which are used in two ways [160]. Natural crushed baryte is used for the removal of radium from industrial effluents, which after settling in tailing ponds (catchpits), are pumped onto fixed beds of crushed baryte of 1 - 3 mm grain size. These beds are located in open, high-rate trickling filters in twin rectangular boxes (each box has two parallel beds) made of concrete. Two pairs are connected in series side by side so that the bottoms of the first pair are above the bottom level of the second pair. Each bed has a load area of 9 m2, a volume of 18 m3 and is 2 m deep. Wire netting is used to separate the barytes from the sand base, preventing mixing during the backwash cycle. The drainage system for the purified water is located underneath. Water throughput for each bed is 10 1/sec, while during the backwash, which lasts at least 15 min, it is 5 l/sec. The purified water is stored in a reservoir of 55 m3 capacity (the contents of which are also used for the backwash), from where it overflows into a second reservoir of 19 m3 capacity. From this second reservoir, the purified water is released into a river. The spent backwash is collected in a spherical basin of 32 m3 capacity, from where it is pumped back into the tailing pond. The average decontamination factor for 226Ra thus obtained is 3.3. Radium-contaminated pit water is purified at the rate of 10 m3/min by cation exchangers, and spent regenerating liquors are purified by finely crushed barytes (or a special particulate barium sulphate) in upward-flow columns or by co-precipitation in mixer-settlers. These operations have proved to be relatively successful and are being modified further.

66 VI. FINAL PRODUCT CONDITIONING

VI. 1. CONDITIONING OF EXHAUSTED INORGANIC ION EXCHANGERS

After exhaustion of their exchange capacity, inorganic ion exchangers must be prepared for storage or final disposal. There is no generally applicable treatment for these materials; the preferred method of conditioning depends on the history of the exchanger in question and particularly on which method of disposal has been selected.

VI. 1. 1. Treatment at Idaho Falls

At the Idaho Chemical Processing Plant, USA, as mentioned earlier, clinoptilolite is used for the treatment of water from storage basins for irradiated fuel elements. It is contained in 200-litre mild steel drums. After exhaustion of the exchange capacity of the mineral, the drum is capped and sealed, and buried in the ground along with other solid radioactive wastes. Since the disposal ground lies in an arid area with a low water table and the radionuclides are firmly fixed, further treatment of the spent ion exchangers is not necessary. Complicated handling of the minerals is thus completely avoided [51] .

VI.1.2. Treatment at Harwell

At Harwell, UK, spent vermiculite is discharged directly into shielded disposal containers. Removal of the mineral from the columns, however, was a difficult operation. Since digging was impracticable due to the high level of radiation, the only feasible method was fluidization. The efficiency of this process was, however, low, since the volume discharged was increased by an amount equivalent to the water necessary for displacing the bed. Subsequently, the vermiculite was placed into a basket-type centrifuge, the discharge from which is made by a hydraulically operated traversing plough, which pushes the spent mineral down the discharge chute [82] .

VI. 1. 3. Incorporation into bitumen or concrete

Since the radioactive ions are strongly fixed on the ion exchangers, such materials areina safe form for disposal. However, unplanned regeneration or leaching can take place if contact occurs with, for example, saline solutions. Physical dispersion may also take place. One method of preventing these processes is the incorporation of spent ion exchange materials into bitumen or concrete. Experiments carried out at Harwell indicate that up to 30% of crude vermiculite can be incorporated into bitumen [161] . Investigations at Mol, Belgium, showed that grade 0 or grade 1 vermiculite can be incorporated into bitumen, whereas larger particles of the mineral do not intermix and

67 separate completely from the bitumen. Incorporation of spent inorganic ion exchangers into bitumen has, however, not been practised up to now on a routine basis anywhere; neither has incorporation into concrete.

VI. 1.4. Fixation in glasses

At Trombay, India, a process for the incorporation of low- and intermediate-level radioactive residues in glass has been developed. This process, which is described in more detail in section VI. 2., permits the permanent fixation of vermiculite in a glass matrix. The volume reduction achieved is in the range 3-5. The cost of incorporating vermiculite in glass is 30-90% higher than direct disposal without preceding treatment [162] .

VI. 1.5. Reduction of leachability by heat treatment

Early work carried out at Brookhaven National Laboratory, USA, showed that radionuclides sorbed on montmorillonite and other clay minerals became very firmly fixed after firing the loaded material to 1000°C and above [163]. These investigations to develop a process for the treatment of high-level radioactive liquid wastes have, however, been abandoned in favour of direct incorporation of the fission products in glasses. At Casaccia, Italy, a fine powder of yellow Neopolitan tuff, a natural inorganic ion exchanger of the zeolite type with high exchange capacities (Cs+, 2.1 meq/g; Sr2+, 0.7 meq/g; Ce4+, 0.5 meq/g), is used on a routine basis as an additive in the chemical treatment of low-level radioactive effluents (10"3-10"4 yCi/ml). For use in columns, larger grains of the product (50-70 mesh) have to be selected. Recent investigations have shown that the exhausted ion exchange material can be rendered virtually un- leachable by heating for 6 h to 1000°C or for 2 h to 1200°C. At temperatures up to 1000°C, the tuff retains its granular structure, whereas at 1200°C vitrification takesr place. For safe disposal, the mineral can also be converted into a leach-resistant mortar by reaction with Ca(OH)2 in the presence of water. The former process may be used for the treatment of higher level radioactive effluents, whereas the latter process is intended for low-level wastes [164] .

VI. 2. USE OF INORGANIC MINERALS FOR IMPROVEMENT OF FIXED RADIOACTIVE WASTES

In addition to the use of minerals as ion exchangers in the treatment of low- and intermediate-levelradioactive effluents, these materials can also be applied as conditioners in the processing of radioactive residues. It has been found [156] that the leach rates of caesium from wastes incorporated into asphalt can be lowered by a factor of 2 - 10 when 2% by weight of Grundite (commercial name for a type of illite) and the same amount of sodium metasilicate are added. The purpose of the latter additive is to keep the mineral particles suspended in the aqueous waste. Reduction of the caesium leach rate from a cement mortar by the addition of mordenite, a mineral of the zeolite type, has been investigated at Battelle Northwest Laboratories, USA. This additive lowers the initial leach rate of caesium from 0.47 g • cm-2 • d"1 for untreated mortar to 0. 0086 g • cm-2 • d-l for

68 treated mortar. Although this method is promising and shows the marked effect of mordenite in retaining caesium, it should be noted that the leach rate of caesium from the mordenite mortar was still about ten times that from an asphalt mixture [77], At Oak Ridge National Laboratory, USA, extensive investigations have been carried out on the improvement of cement mixes with a view to the disposal of radioactive wastes by hydraulic fracturing [165]. The composition of the mix (kg/litre liquid waste) which best met the specifications at the lowest cost was as follows: Portland cement, 0.36; fly ash, 0.24; attapulgite, 0.09; illite, 0.05; and retarder, 0. 4004 (g/litre liquid waste). The cement enhances the hardening and the strength of the mix and also combines chemically with radiostrontium in the waste. Fly ash, a highly siliceous pozzolanic material, substitutes a part of the cement and improves strontium retention. Attapulgite clay is added as a rheological conditioner to prevent any phase separation of the slurry, resulting from the low cement content. Retention of caesium is achieved by the addition of Grundite (illite clay). At Harwell, UK, small volumes of highly active liquids or organic solvents are sometimes absorbed on expanded vermiculite as a method of treatment prior to disposal. At Oak Ridge National Laboratory, USA, either attapulgite or bentonite is added as a conditioner to tributylphosphate (TBP) waste to be incorporated into bitumen. In this application, the purpose of the clay is to absorb the TBP. The general composition of the waste mixture studied is 30% (by weight) bitumen, 10-30% TBP and 40-60% bentonite or attapulgite. Mixtures containing 10% TBP were too hard, and those containing 30% TBP were too fluid [156] . The use of cement for conditioning radioactive residues, such as chemical sludges, evaporator concentrates, spent ion exchangers, etc. , has been widely studied [166-171], and will not be treated in detail here. Sometimes an improvement of the product can be obtained by addition of vermiculite. Recent investigations at Karlsruhe, Federal Republic of Germany, have shown that replacement of 50% or more of the cement by oil shale ash yields a product which binds about the same amount of water as pure cement. Since the water percolates easily through the product, mechanical mixing can be avoided if only small amounts are to be treated. Oil shale ash is a cheap waste product, and its use as a partial replacement- for cement leads to considerable savings. The costs of the product stem mainly from crushing and transport. A mixture of oil shale ash with cement is being used on a routine basis at Karlsruhe for the fixation of evaporator concentrates (100 kg oil shale ash and 100 kg cement per 100 litre evaporator concentrate) until the large bituminization plant is in operation [171] . At the Bhabha Atomic Research Centre, India, bench scale and pilot plant experiments have been carried out in order to develop a process for the fixation of low- and intermediate-level radioactive wastes in low-melting glasses [162] . Glasses have been produced containing spent vermiculite (50-75 wt%), sludge and sludge cakes arising from chemical treatment of radioactive effluents, or incinerator ash. The incorporation of varying weight fractions, and the addition of combinations of fluxing agents such as

Na^O, CaO, B2Os and Pb304, and additives such as Si02 and A1203, have been investigated. With the exception of incinerator ashes, glass pouring temperatures of 900 - 950°C have been obtained by proper choice of fluxing agents. The chemical stability of the product was found to be good; leach rates in water comparable to those of pyrex glass were obtained. Volume reduction

69 factors for sludge, sludge cake, vermiculite and incinerator ash are in the ranges 10-15, 8-10, 3-5 and 2-3, respectively. Based on actual expenditure incurred with a pilot plant having a capacity of 90 litres of glass per day, a unit cost of $0.83/litre of product was achieved. This process, plus subsequent disposal, is more economical than the disposal of secondary wastes without preceding treatment, in cases where a volume reduction of 8-10 or more can be achieved, i.e. in the case of sludge and sludge cake. Further- more, the incorporation of secondary wastes into glasses ensures safe disposal and makes intensive monitoring of the environment unnecessary.

70 VII. ECONOMIC ASPECTS

Natural ion exchange materials are known to be cheaper than synthetic ones. It should be borne in mind, however, that economic comparisons of the use of these two types of material can be made correctly only on the basis of the total operating and capital costs. Economic aspects are reported in detail in Ref. [172] . If the natural materials are used, several factors most be thoroughly investigated and taken into account in an economic evaluation of their use, as follows: efficiency and selectivity for radioisotope removal, required decontamination level, exchange kinetics, column dimensions required, etc.

VII. 1. CAPITAL COSTS

Capital costs include those of the main and auxiliary equipment and also of the buildings and structures. The principal advantages of the use of natural ion exchange materials compared with synthetic materials are as follows: (1) possible use of cheaper structural materials for equipment (carbon steel can be used), since in nearly all cases natural sorbents are not regenerated; (2) no need for equipment for ion exchange material regeneration; (3) elimination of equipment for the treatment of radioactive regenerating solutions; and (4) use of cheaper pumps in liquid pumping systems. Disadvantages in the use of natural materials are: (1) ion exchange column dimensions need to be considerably larger; and (2) in certain cases, several ion exchange stages may be necessary, using a number of columns each with a different natural selective sorbent. Such requirements can result in an increase in total capital costs. Thus, when comparing capital costs and evaluating any economic advantages of natural materials, it is necessary to determine the sorption capacity of these materials, the degree of treatment needed, the radiochemical, physical and chemical compositions of the liquid wastes to be treated, and other factors. The choice of ion exchange materials is governed not only by the ion exchange capacity, but also by their selectivity. In addition, it is necessary to assess the total quantity of the materials required for the columns, in order to estimate costs for material handling and storage, shielding, labour for sorbent loading and unloading, and disposal of spent sorbents. At present, a limited number of pilot and industrial radioactive waste treatment plants in various countries use natural sorbents. Two examples of total capital costs are as follows. (1) At Harwell, UK, a plant for radioactive waste water treatment with vermiculite using a vertical centrifuge is in operation; the cost of the plant, equipment and building is $22 000.

71 (2) At the National Reactor Testing Station (NRTS), Idaho, USA, fuel element cooling water is treated with a mineral ion exchange material, clinoptilolite. The plant capacity is 3800 m3/yr, and the capital cost of the equipment is about $12 500.

VII. 2. OPERATING COSTS

The following items of expenditure must be taken into account in the estimation of operating costs:

(1) Labour, i.e. wages, retirement, social benefits, etc.; (2) Material, i.e. cost of ion exchange materials and of sorbent preparation; (3) Power, i.e. charges for water, steam, electricity, etc; (4) Maintenance, i.e. charges for maintaining equipment and buildings under normal conditions; (5) Depreciation, i.e. annual costs or charges according to the depreciation rate appropriate for the country concerned. For comparison with foreign plants, it would be sound practice to accept depreciation charges based upon 7. 5-year equipment life and 20-year life of the plant buildings and structures; (6) Direct and indirect overhead charges, i.e. expenses for administrative and service staff, postage, factory transport, etc. Operating costs may be expected to be lower when natural ion exchange materials are used in preference to synthetic materials. This is due to a considerable relative decrease in the cost of the materials themselves, and to elimination of charges for ion exchange material regeneration. These costs amount to about 10% of the total with synthetic materials, a percentage which may be reduced by a factor of 5-10 if natural materials are used. Operational experience of some plants shows that the cost of radioactive liquid waste treatment with natural sorbents is not very high, as can be seen from the following three examples. (1) The total cost of water treatment with clinoptilolite at NRTS, Idaho, USA, amounts to $0.28/m3, with an annual plant capacity of ~ 3800 m3; clinoptilolite preparation costs total $309/m3. (2) At the Nuclear Research Institute, Czechoslovak Socialist Republic, radioactive wastes are treated with calcium chloride and domestic kieselgur containing natural aluminium silicate of the illite type; treatment of 1 m3 of wastes costs about 400 Czechoslovak crowns (~40 roubles) with low plant capacity (~ 1000 m3/yr) and expenses for materials amount to <1% of this sum. (3) The cost of liquid waste treatment with vermiculite at Harwell, UK, is $7. 93/m3 of wastes treated, with the equipment in use for 15% of the time.

72 APPENDIX I

NATURAL MATERIALS FOR USE IN WASTE TREATMENT

Material Radionuclides Reaction Selected Name Formula type studied mechanism references

Oxide or Corundum A1Z03 Sr Compound formation? 1 hydroxide 90 Diaspore AI2O3 .H2O Sr Compound formation? 1

90 Boehmite AI2O3. H2O Sr Compound formation? 1

90 Gibbsite A1(0H)3 Sr Compound formation? 1

90 Goethite Fe203 .H20 Sr Compound formation 1,2,91 and co-precipitation

90 131 s 106 Limonite Fe203,HzO Sr. I, S. Ru Compound formation 1,2 (Goethite) and co-precipitation

90 51 60 Pyrolusite Mn02 Sr, Cr, Co 3,4,5

Halide Fluorite CaF2 Sr, U, Pm, Pu Isomorphous replacement 7,8, 98

85 133 Carbonate Calcite CaC03 Sr. Ba Isomorphous replacement 1,9,10,35 SrC0 85 133 Stronianite 3 Sr, Ba Isomorphous replacement 6, 9 BaCOj 133„ Witherite Ba Isomorphous replacement 6, 9 PbC0 Cerrussite 3 Pb Isomorphous replacement 6,9 ZnC0 Smitsonite 3 Zn Isomorphous replacement 6,9 CaMg(C0 ) 144 55 95 Dolomite 3 2 Ce, Zr, Nb, Compound formation 3 "Ru. "P, Sr, "S and adsorption

3Z Phosphate Apatite Ca5(P04)3(F,Cl,0H) Sr, P, U, Pu, Pm Isomorphous replacement 7,8,11

Phosphorite Ca5(P04)3(F,C1.0H) °Sr, Co, Cu, Fe Isomorphous replacement 12,13 (Apatite) and sorption APPENDIX I (cont.)

Material Radionuclides Reaction Selected Name Formula type studied mechanism references

85 Phosphate Variscite A1P04.2H20 Sr Compound formation? 11 (cont.)

133 85 Sulphate Gypsum CaS04.2H20 Ba, Sr Isomorphous replacement 6,9

90 Anhydrite CaS04 Sr Compound formation 14,15

85 Celestite SrS04 Sr Isomorphous replacement 9

85 50 226 35 Barite BaS04 Sr, Sr, Ra, S Isomorphous replacement 6,9,14,15,16

Silicate 137 90 (Neso-) Olivine (MgFe)2Si04 Cs, Sr Ion exchange 17

137 90 Humite Mg(0H,F)3Mg2Si04 Cs, Sr Ion exchange 17

137 90 Chrondrodite Mg(0H.F)2.2Mg2Si04 Cs, Sr Ion exchange 17

137 90 Leocophenacite Mn,(Si04)3(0H)2 Cs, Sr Ion exchange 17

137 90 Taumasite Ca3H2Si04(C03)(S04).13H20 Cs, Sr Ion exchange 11

137 90 Zircon ZrSi04 C"s, Sr Ion exchange 17

137 90 Topaz Al2(Si04)(0H,F)2 Cs, Sr Ion exchange 17

137 90 Kyanite Al2Si205 Cs, Sr Ion exchange 17

137 90 Sillimanite Al2Si205 Cs, Sr Ion exchange 17

137 90 Stuarolite (Fe,Mg)2(A1, Fe )906 (Si04)4(O.OH)2 Cs, Sr Ion exchange 17

137 90 Almanidne Fe 3Al2Si3012 Cs, Sr Ion exchange 17

Grossular Cci j A12 Si^ O^ 137 Cs, 90Sr Ion exchange 17

137 90 Andradite Ca3(Fe,Ti)2Si3012 Cs. Sr Ion exchange 17

137 90 Uvarovite Ca3Cr2Si3012 Cs, Sr Ion exchange 17

1I7 Plazolite Ca3Al2Si20s(Si04)(0H) Cs, ">Sr Ion exchange 17 137 90 Vesuvian Ca10(MgFe)2Al4(Si2O7)2(SiO4)5(OH,F)4 Cs, Sr Ion exchange

13, 90 Sphene CaTiOa Cs, Sr Ion exchange

137 90 Rinkite Na(Ca, Ce)2 Ti(Si04) 2F Cs, Sr Ion exchange

137 90 Axinite (Ca, Mn, Fe)3 Al 2B03 (Si4 012 )OH Cs, Sr Ion exchange

137 90 (Soro-) Hemimorphite Zn4Si20, (OH)2 .H20 Cs, Sr Ion exchange

137 90 Me li lite Ca2 (Al, Mg)(Si, Al)2 O 7 Cs, Sr Ion exchange

137 90 (Cyclo-) Beryl Be3 Alz Si6018 Cs, Sr Ion exchange

137 90 Dioptase Cu6 Si6018,6HzO Cs, Sr Ion exchange

137 9 Tourmaline (Na, Ca)(Mg, Fe, Li)3 Al6 B3 Si6 0 2J (OH)„ Cs, °Sr Ion exchange

Wollastonite CaSiO, 111 Cs. 90Sr Ion exchange

Rhodonite MnSiOj 137 Cs, 90Sr Ion exchange

137 90 (Ino-) Hypersthene (Mg,Fe)2Si206 Cs. Sr Ion exchange

137 90 Aegerine NaFeSiz06 Cs, Sr Ion exchange

137 90 Augite Ca(Mg, Fe, Al)(Si, AI)2 q Cs, Sr Ion exchange

137 90 Diopside CaMgSi206 Cs, Sr Ion exchange

137 90 Jeffersonite CaZnSi206 Cs, Sr Ion exchange

137 90 Tremolite Ca2Mg5Si,022(OH)2 Cs, Sr Ion exchange

137 90 Amphibole NaCa2 (Mg, Fe, Al)3(Si, A1)8022 (OH)2 Cs, Sr Ion exchange

137 90 (Phyllo- ?) Chrysotile Mg6Si4°io(OH>s Cs, Sr Ion exchange ]37 90 Sepiolite M H Si OH 6H O Cs, Sr Ion exchange g8 6 l2°30( )l0- 2 137 90 Attapulgite M H Si O OH 6H O Cs, Sr Ion exchange ,17,18,19,37 g8 6 l2 30( )l0- 2

H Si H 2 i37 90 Palygorskite Mg3 2 8°22( 20)6- H20 Cs, Sr Ion exchange

137 90 Zoisite Ca, Al3Si3012(0H) Cs, Sr Ion exchange -J CJ1 APPENDIX I (cont.)

Material Radionuclides Reaction Selected Formula type studied mechanism references

Silicate (cont.) 90 (Phyllo-) Talc Mg3Si4O10(OH)2 Cs, Sr Ion exchange 17

137 90 Pyrophyllite Al2Si4O10(OH)2 Cs, Sr Ion exchange 17

137 90, Biotite K(MgFe)3(AlSi3)O10(OH,F)2 Cs, Sr Ion exchange 3,17,20,21,58,69

131 Muscovite KA12(AlSi3)O10 (OH, F)2 Cs, °Sr Ion exchange 17,49

137 Pennine (MgAl)s (AlSij)O10 (OH)2 Mg3 (OH), Cs, °Sr Ion exchange 17

131 Dellesite (FeMg)3 Si4O10(OH)2 (MgFe)3(0, OH)6 Cs, °Sr Ion exchange 17

137 Vermiculite Mg3(AlSi3)O10(OH)2.nH2O Cs, "Sr. Ion exchange 1,11,17,19,20,23, 131 Xe 5 Kr, 33,40,41,46.48,49, 60,62,67,69,70,82, 83,84,94

137 90 13l Chlorite MgjSi4O10(OH)2Mg3(OH)6 Cs, Sr, I Ion exchange 17,62,66

137 85 90 Hydrobiotite K,H2O(FeMgAl)2(AlSi)Si3O10 (OH)2 Cs, Sr, Sr, Ion exchange 17,36,43,47,69,70 86Rb, 242Cm

Fluorophlogopite K(Mg)3(AlSi3)O10F2 Cs, Sr Ion exchange 20,21

Hydromuscovite K,H2OAl2AlSi3O10(OH)2 Cs, Sr Ion exchange 17

Sericite KAl2AlSi3O10(OH)2 Cs, Sr Ion exchange 17,38

137 90 85 86 Illite KH2O(Al2)(AlSi)Si3O10(OH)2 Cs, Sr, Sr, Rb, Ion exchange 17,20,30,41,46,47, 65 Zn, 60Co, KMn, aNa 50,61,62,65,85,88, 92

Glauconite KH2O(FeMgAl)2(AlSi)Si3O10(OH)2 Cs, Sr. Rb Ion exchange 3,17,20,21,61

Celadonite KH2O(FeMg)2Si4O10(OH)2 Cs, Sr Ion exchange 17

Serpentine Mg3Si2Os(OH)4 Cs Ion exchange 17 137 90 85 Kaolinite Al2Si2Os(OH)4 Cs, Sr, Sr, Ion exchange 1,17,20,23,28,34, 86Rb, 60Co, 42K, B 39,40,42,45,46,47, 48,49,50,61,62

137 90 Halloysite Al2Si205(0H)4.2H20 CS, Sr Ion exchange 17,23

I37 Allophane Al2Oj.Si02.nH20 Cs Ion exchange 17

l37 Hisingerite Fe203 .Si02.nH20 Cs Ion exchange 17

137 134 133 Montmorillonite (MgAl) 3(AlSi)4 O10 (OH)a Cs, Cs, Ba, Ion exchange 1,17,19,20,21,23, or Bentonite 90Sr. 85Sr, 56Mn, 58Co, 24,28,29,33,34,40, 4JK, 24Na, MNa 47,49,50,56,61,63, 67,69,71

137 90 Nontronite Fe2 Si4O10 (OH)2 .nH20 Cs, Sr Ion exchange 1,17

I3, 90 Apophyllite KCa4Si4OI0F,8H2O CS, Sr Ion exchange 17

137 90 (Tecto-) Orthociase KAlSi3Os Cs, Sr Ion exchange 17

137 90 Sanidine KAlSi308 Cs, Sr Ion exchange 17

137 90 Albite NaAlSisOe (= Ab) Cs, Sr Ion exchange 17,28

I 137 90 Oligoclase NaAlSi3Os +10-307oCaAl2Si208( An) Cs, Sr Ion exchange 17

Andesine Ab + 30-50foAn 137Cs, »Sr Ion exchange 17

Labradorite Ab+ 50-70% An 137Cs, 90Sr Ion exchange 17

42 85 Celsian BaAlSi308 Ca, Sr Ion exchange 76

137 90 Scapolite (NaCa)4 Al(AlSi)Si208 (C1,C03) Cs, Sr Ion exchange 17

137 90 Leucite KAlSi206 Cs, Sr Ion exchange 17

137 90 Analcite Na2Al2Si4012.6H20 Cs, Sr, Na Ion exchange 17,57,58

137 90 42 Wairakite CaAl2 Si4012. 2H20 Cs, Sr, Na, Ca Ion exchange 57,76

137 90 Pollucite CsNaAlSijOg (H20) Cs, Sr Ion exchange 17

137 90 Nepheline NaAlSi04 Cs, Sr Ion exchange 17

137 90 Sodalite Na8(AlSi04)6Cl2 Cs, Sr Ion exchange 17 APPENDIX I (cont.)

Material Radionuclides Reaction Selected Name Formula type studied mechanism references

Silicate I37 90 (Tecto-) Cancrinite (NaKCa)3.4(AlSi)6Ol2 (S04C03C5).nH20 Cs. Sr Ion exchange 17 (cont.) Mordenite and CaAl SijoO^ .7H 0 1I7Cs, 90Sr, 42Ca Ion exchange 2 2 17.53.76.77 Ptilolite

137/- Stilbite and CaAl2 Sij 018.6H20 Ion exchange 17,75 Desmine

Heulandite and CaAl2Si7018 .6H20 Ion exchange 5,17,24,33,40,44.51, Clinoptilolite 52,53,56,61,64,72, Na 73.74.76.78

137 90 Fauja site Na2Ca(A)jSi4012)2.16H20 Cs, Sr Ion exchange 17,78

137 Harmotome BaAl2Si6016.6H20 Cs Ion exchange 17

I37 Brewsterite (SrBa)Al2Si60ls.5H20 Cs Ion exchange 17

13T 90 Phillipsite (K2Ca)Al2Si4012.4H20 Cs, Sr Ion exchange 17,74,78

137 Chabazite CaAl2Si4012.6H20 Cs Ion exchange 17,53

137 Natrolite Na2Al2Si3O10.2H2O Cs Ion exchange 17, 58

137 Scolecite CaAl2Si3O10.3HzO Cs Ion exchange 17

137 Thomsonite NaCa2Al5Si5O!0.1H2O Cs Ion exchange 17

Far8elite Ca Thomsonite 137Cs Ion exchange 17

137 24 Erionite (NaKCa)3Al3Si8024.8H20 Cs. Na Ion exchange 33,53,56,74,78

85 4Z Ion exchange Ferrierite (Na2K2Ca)Alz Si8O20. 7H20 Sr. Ca 76

Yugawaralite 85 Sr, 42Ca Ion exchange 76

137 90 Ion exchange Quartz Si02 Cs. Sr 17

Agate SiOj 137 Cs Ion exchange 17

137 Chalcedoney SiO, Cs Ion exchange 17 Opal SiOj.nHjO 137Cs Ion exchange

Geysrite SiOj.nHjO 137 Cs Ion exchange

137 Silica Gel Si02.nH20 Cs Ion exchange , 89

137 Diatomite Si02.nH20 Cs Ion exchange 17

137 Kacholong Si02.nH20 Cs Ion exchange

Igneous rocks Rhyolites 137 Cs, 90 Sr Ion exchange

Volcanic glass 137 Cs, 90Sr Ion exchange

Perlite 137Cs, 90 Sr Ion exchange

Rhyodacite 137 Cs, 90Sr Ion exchange

Dacite 137 Cs, 90 Sr Ion exchange

Quartz porphyry '"Cs, 90Sr Ion exchange

Trachytes 131 Cs, 90 Sr Ion exchange

Andesites 137Cs, 90Sr Ion exchange

Spillites 137Cs, 90Sr Ion exchange

Diabases 137 Cs, 90Sr Ion exchange

Melaphyres 137Cs, 9°Sr Ion exchange

Basalt 137Cs, 90Sr Ion exchange ,79,114

Teschenite picrites 137Cs, 90Sr Ion exchange

Teschenites with olivine 137Cs, 90 Sr Ion exchange

Picrites 137 Cs, 9QSr Ion exchange

Peridotites 137Cs, 9°Sr Ion exchange

Pyroxenites 137Cs, 9»Sr Ion exchange

Phonolites 137Cs, 90 Sr Ion exchange

Tephrites 137Cs, 90Sr Ion exchange APPENDIX I (cont.)

Material Radionuclides Reaction Selected Name Formula type studied mechanism references

Igneous rocks Basanites I37Cs, 90Sr Ion exchange 17 (cont.) Nephelinites 137Cs, 90Sr Ion exchange 17

Leucites 137Cs, 90Sr Ion exchange 17

Melilitites 137Cs. 90Sr Ion exchange 17

Olivine basalt 137Cs, 90Sr Ion exchange 17

Organic Peat moss MNa, 60Co, 6SZn, Ion exchange 80,89 85Sr. 137Cs. I54Eu, J03Hg, 106 Ru, 204 TI

Lignite 90Sr, 137Cs, 60Co, Ion exchange 5,81,95,96 51Cr, MMn

Bitumen 137Cs, 90 Sr Ion exchange 86

Sawdust U, 137 Cs, 106RU Ion exchange 87,91 APPENDIX II

DEFINITION AND DETERMINATION OF CAPACITIES OF NATURAL ION EXCHANGERS*

1. INTRODUCTION

Natural ion exchangers often show more complicated reactions than synthetic organic ion exchangers. Therefore, a quantitative treatment or even qualitative prediction of their behaviour requires a thorough knowledge of their properties. Quantities of interest are the ion exchange capacity, distribution coefficients, degree of dissociation of the ionic groups, diffusion coefficients, rate constants, swelling properties, resistance to chemical and mechanical attack, pore size, and grain size. The ion exchange capacity is one of the most important and the most often determined quantity. It is defined as the number of ions, which, in a definite amount of material and under specified experimental conditions, is available for the ion exchange process. This definition appears to be rather clear and simple. A closer examination, however, reveals that about 18 different definitions of the capacity are at present used concurrently by scientists from different fields. Of these at least 14 are essential. The reason for this variety arises from numerous complications which can arise in the determination of the capacity. The main factors are: (1) Only carefully pretreated ion exchangers in well-defined ionic forms can be expected to give reproducible results; (2) In order to determine the dry weight of the ion exchanger, temperatures are sometimes necessary, which may destroy the material to some extent; (3) The rate of the exchange process can be so slow that attainment of the equilibrium is difficult to ascertain; (4) The capacity can depend on the nature of the counter-ion, e. g. for steric reasons; (5) Complex formation of counter- and co-ions in solution as well as in the exchanger; (6) In addition to the counter-ions, additional ions can be absorbed by a Donnan-type electrolyte sorption; (7) In the case of incomplete dissociation of the ionic groups, the capacity will be a function of the solution pH; (8) Some methods give rise to errors by the so-called "salt-free water film"; (9) During the washing procedure, absorbed counter-ions can be washed out by the H+-ions of the water (hydrolysis);

* This Appendix was specially prepared for the May 1969 Panel by K. Bunzl and B. Sansoni, Gesellschaft filr Strahlenforschung mbH Munchen, Institut fur Strahlenschutz, Radiochemisch-Analytische Abteilung, Neuherberg bei Munchen, Federal Republic of Germany.

81 (10) Apart from equivalent exchange, equimolar exchange can also take place; (11) Some ions are absorbed so strongly that their removal by leaching is difficult; (12) Conversion of the ion exchanger to the if'-form destroys many mineral exchangers; (13) Ion exchange capacity may depend on the particle size; (14) Partial solubility of the sample; (15) The nature of the co-ion may have some influence on the capacity; and (16) Oxidation or organic poisoning of the exchangers can cause considerable capacity losses. The great number of different definitions of the ion exchange capacity is further favoured by the fact that the phenomena of ion exchange are studied in so many different fields of science, such as analytical, inorganic, physical, and colloid chemistry, chemical engineering, mineralogy, agriculture, molecular biology, nuclear science, waste disposal and hydrology. Scientists in these various fields have developed their own concepts of ion exchange capacity, and have, particularly in the past, not always defined them accurately.

2. ELEMENTARY PRINCIPLES

In contact with an electrolyte solution, the solid ion exchanger can absorb ions from the liquid phase and exchange them for an equivalent amount of its own ions of the same sign. In order to achieve a reasonably fast exchange reaction, the material also has to be able to absorb sufficient solvent to facilitate the diffusion of the ions. If, for steric reasons, the ions or the solvent cannot penetrate in depth, the ion exchange process will occur mainly on the surface of the sample. In these cases, crushing of the material will increase the capacity and the rate of exchange. The exchange of monovalent cations may be written as:

A+ X"+ B+ [c]" ^ A+ [c]" + B+ X" (1) or A+ + [B]+^ [A]++B+ where C denotes the exchanger, A and B the counter-ions and X the co- ions in the solution and in the solid phase C; dl means solid phase. For polyvalent cations or anions, similar equations can be formulated. The equilibrium constant of the above reaction indicates whether the ion A or B is absorbed preferentially by the ion exchanger. If the material has to be saturated completely with only one kind of counter-ion, the equilibrium in equation (1), has to be shifted to one side by applying an excess amount of the corresponding electrolyte solution. In this connection it is important to note that the co-ions X can, in equilibrium with the external electrolyte, penetrate the exchanger to some extent. However, due to the simultaneously arising Donnan-potential, their concentration will be much lower than in the electrolyte solution. Since the electro-neutrality of the sample has to be conserved, an additional amount of counter-ions A, equivalent to the absorbed co-ions, will also migrate in the exchanger. This electrolyte penetration depends on the

82 concentration of the external electrolyte, and will become very small if this concentration is sufficiently small (Donnan-exclusion). If no further complicating factors arise (e. g. hydrolysis), this absorbed electrolyte can be washed out completely with water. For amore detailed introduction to ion exchange, the reader is referred1 to the monographs by Amphlett [1], Blasius [2], Griessbach [3], Helfferich[4], Inczedy [5], Kitchener [6], Nachod and Schubert [7], and Samuelson [8],

3. DEFINITIONS OF ION EXCHANGE CAPACITY

In order to characterize the efficiency of an ion exchanger for given applications, it is important to know how many counter-ions can be exchanged by a certain amount of material. As mentioned above, this ion exchange capacity is strongly dependent upon the experimental conditions. Its different possible definitions and units are listed in Table A. In the following, they are discussed in more detail.

3.1. Pure ion exchange capacity Kr

This quantity gives the number (in meq) of exchangeable ions of a specified amount of ion exchanger. According to definition, this amount is, after removal of the absorbed water, 1 g in the H+-form for a cation exchanger and 1 g in the CI"-form for an anion exchanger. The condition of weighing the ion exchanger in the H+-form cannot be met in cases where this form is not stable. In this case the capacity has to be determined with another ion, e. g. potassium (K+). Since the dry weight of the K+-form.' exceeds that of the H+-form, too small a capacity will result. The capacity of the H+-form, however, can be correlated to the experimentally determined capacity of the Me-form, according to:

KMe (meq/g) = l-K^(EwJ- 1.008). 10"s (2)

and for an anion exchanger, which is weighed in the A" -form instead of the Cl"-form, according to:

Ka (meq/g) = l-K^Ew/- 35.453). 10-3 (3)

Here EW denotes the equivalent weight of the Me+ or A" ion. and K^ should be given in meq/ g dry exchanger in the Me+- and A"-form respectively, in order to obtain K^1 and Kj'1 in meq/g dry exchanger in the H+- and CI"-form, respectively. For a comparison of ion exchange capacities determined with different ions, reduction of these values to a standard state, i. e. the H+-form, is always necessary.

1 The References to this Appendix are listed separately at the end of the Appendix.

83 TABLE A. DEFINITIONS AND UNITS OF ION EXCHANGE CAPACITY

Name Unit

Pure ion exchange capacity meq/g (Reine Ionenaustauschkapazitat) Kr

Backbone capacity meq/g (Geriist-Austauschkapazitat) K0

Theoretical ion exchange capacity meq/g (Theoretische Ionenaustauschkapazitat) Kt[j

Analytical ion exchange capacity meq/g (Analytische Ionenaustauschkapazitat) K

Sorption capacity meq/g (Adsoiptionskapazitat) Kac]

Total exchange capacity meq/g (Gesamt-AustauschkapazitSt) Ktot

Apparent or useful capacity meq/g (Scheinbare oder nutzbare Austauschkapazitat) Ks

Maximum capacity meq/g (Maximale Austauschkapazitat) Kmax

Pure volume ion exchange capacity meq/ml bed (Reine Volumen-IonenaustauschkapazitSt) Ky

Sorption volume capacity meq/ml (Adsorptions-VolumenaustauschkapazitSt) Kv a(j

Total volume capacity meq/ml bed

(Gesamt-Volumenaustauschkapazitat) Ky tot kg CaO/ft3 bed, Useful volume capacity kg CaO/m3 (Nutzbare Volumen-Austauschkapazitat NVK) Kv nutz

Pure grain ion exchange capacity neq/grain (Reine Korn-IonenaustauschkapazitSt) K^

Microscopic pure volume ion exchange capacity meq/ml exchanger, (Mikroskopische reine Volumen-Ionenaustausch- ^eq/mm3

kapazitat) Kmi Vj r

Breakthrough capacity meq/column or (DurchbruchskapazitSt) K,j meq/g

Concentration of fixed ionic groups m, X meq/g adsorbed water (Konzentration der Festionen)

S-, T-, V-value meq/100 g dry weight

Surface capacity meq/m2 (Oberflachen-Austauschkapazitat) Kg

84 Since drying at high temperatures sometimes destroys the exchanger considerably, it is not always possible to remove all the absorbed water. In these cases the exact drying conditions should be stated.

3.2. Backbone capacity Kp

Another possibility to compare materials, whose ion exchange capacity has been determined with different ions, is to calculate their backbone capacity. This quantity gives the number of exchangeable ions in meq/ g dry backbone or matrix, which is the amount of dry exchanger minus the weight of the countei>ions. It can be obtained from the pure ion exchange capacity, determined for the Me+-form, according to:

KMe (4) K0 (Wg^.j^EW^-lO-a

where EWMe is again the equivalent weight of the Me-ion. For an anion exchanger the index A should be substituted for Me. If no complications arise during the exchange, this quantity should be independent of the particular ion used. If for different ions deviations from a constant value occur, one can conclude that the exchange capacity is different for different ions (e. g. for steric reasons). These possibilities for standardization of ion exchange capacities according to 3.1. and 3. 2. are very useful, but even so they have hardly been used until now.

3. 3. Theoretical ion exchange capacity Kth

If the chemical composition of an ion exchanger is already known, the exchange capacity can be calculated from the molecular weight of the unit which carries one exchangeable ion. This method, which gives good results with resin ion exchangers, can be used to check different structure models of the material. For inorganic ion exchangers, the method has been demonstrated by Baetsle and Pelsmaekers [9],

3.4. Analytical ion exchange capacity Kan

In some cases an analytical determination of the heteroatom of the ionic group in the exchanger is possible. The analytical ion exchange capacity thus obtained may, however, not agree with the pure ion exchange capacity, since not all ionic groups may be available for exchange processes.

3. 5. Sorption capacity Kad

This quantity is the amount of electrolyte (meq) or non-electrolyte (mmol) which can be absorbed by 1 g of dry ion exchanger apart from the exchangeable counter-ions. For comparative measurements, all values should again refer to the dry exchanger in the H+- or CI"-form. If the sorption capacity was determined in the Me-form, its value for the corresponding H+-form will be:

85 KMe (5 (meq/g) = d f3 > ^ l-K^f(EWM - 1.008)- 10'

is again the equivalent weight of the Me-ion. For an anion exchanger, whose dry weight was determined in the A- form, the sorption capacity of the material when weighed in the CI"-form will be: KA (me

where EWA denotes the equivalent weight of the ion A. Here again, it is advisable to eliminate the arbitrary state of the H+- form and the CT-form completely, and refer the amount of adsorbed electrolyte only to the dry backbone of the exchanger (i. e. the weight of the exchanger minus the weight of the counter-ions). This quantity, Ko,ad< can be calculated from the experimentally determined sorption capacity in the Me-form, according to: pjMe 3 (7) Ko.ad (meq/g) = ^ . EWMe• 10"

In the case of an anion exchanger, the index A should be substituted for Me, and EW is then the equivalent weight of the ion A. Depending on the method used for determination of the sorption capacity, it may also include ions adsorbed by other mechanisms, e. g. precipitation.

3.6. Total exchange capacity KIot

If one is interested in the total amount of ions absorbed (counter-ions, absorbed electrolyte, precipitated ions, etc. ), a total capacity of the exchanger can be defined, as follows:

Ktot = Kr + Kad (8)

For comparative measurements K,.ot should again be referred to the standard state of an exchanger in the H+- or CI"-form. If the dry weight was determined in the Me-form, we obtain from equations (2) and (5) for a cation exchanger: KMe + KMe

(9) K"ot (meq/g) = x ,KMe (EWMe - l^OOB)-10"8 and for an anion exchanger: a kA -f K (10) Kg\ (meq/g) =!_KA . 35.453)-10"3

If the dry backbone of the exchanger is taken as the standard state, we obtain: pj-Me pWvle 1 n K0, tot (meq/g) = x . KMe • ewJ" . 1q-3 ( )

86 where.the symbols have the same meaning as above. For an anion exchanger, A should be substituted for Me.

3.7. Apparent capacity Ks

Exchangers with weak acid or weak base groups are not always completely ionized. In such cases, depending on the experimental conditions, not every ionic group contributes to the ion exchange process, and only an apparent capacity, Ks , will therefore be measured. Obviously, this quantity will depend strongly on the pH of the external electrolyte solution. If the exchanger has weak acid groups, the apparent capacity will be lower in solutions with low or medium pH values than in those with higher pH values. The reverse is true for anion exchangers with weak base groups. Furthermore, if the external electrolyte is itself a weak electrolyte, Ks will depend on the solution pK. K. is also called in many cases the "useful capacity".

3.8. Maximum capacity Kmax

If experimental conditions are selected where all weak acid (or weak base) groups in an exchanger are ionized, a limiting value of the capacity v/ill be obtained, which is called the maximum capacity, Kmax. It may be considered, therefore, as a special case of the apparent capacity.

3. 9. Pure volume ion exchange capacity Kv,r

Instead of referring the number of exchangeable ions to the dry weight of the exchanger, it can also be referred to the unit volume of packed bed in the H+-or CI"-form in equilibrium with pure water. This definition is especially convenient in column operations. The volume ^ of the packed bed is the sum of the volume of the swollen exchanger and the interstitial volume Vz:

3 vs = VA + Vz (cm ) (12) and the density of the swollen exchanger is given by:

3 dA = ^p(g/cm ) (13) where M is the dry weight and w the water content of the sample. From the definition of the pure ion exchange capacity Kr:

Kr (meq/g) (14)

(n = number of exchangeable ions of the sample) and the pure volume capacity Kv r:

3 Kv,r (meq/cm ) (15)

87 we obtain from equations (12) to (15) the correlation between Kr and K^ r, as follows:

Kv r = dA (me cm3 - 1 + w/M q/ ) (16)

Here r (= Va/Vs) is the packing fraction of the bed and w/M is the amount of absorbed water per g dry exchanger. If we denote the density of the bed as:

3 ds = Vg (g/ cm ) (17)

and since ds = dA" r, we obtain:

The bed density ds can be measured easily by weighing the swollen exchanger and determining the bed volume Vs in a graduated column. It is obvious that the pure volume capacity depends strongly on the swelling properties of the exchanger, its ionic form and the temperature.

3. 10. Sorption volume capacity Kv.ad

The sorption volume capacity is analogous to the sorption capacity

Kad (see section 3. 5. ), except that the number of absorbed ions is referred to the unit volume of the packed exchanger bed. The correlation between

Kv, ad and Kad is given by:

Kv'ad = U^j (meq/cm3) (19) where the symbols have the same meaning as in section 3. 9.

3.11. Total volume capacity Kv.tot

This quantity is analogous to the total exchange capacity Kt0t (see section 3. 6. ), except that the total number of absorbed ions is referred to the unit volume of the packed ion exchanger bed. This means that:

Kv,tot = Kv,t + Kv,ad (20) and dc • K.., Kv,tot = rVw/M" (21)

3.12. Useful volume capacity Ky.nutz

In technical column operations, the capacities are generally expressed as kg CaO/m3 exchanger bed. Instead of the equations (16) and (18), we now obtain:

88 or

28 4 d S KS Knutz = :° ' , M (kgCaO/m3). (23) 1 + w/M where the symbols are the same as in section 3.9.

In equations (22) and (23), the apparent capacity Ks has been used instead of Kr, since due to the kinetics of the exchange or the pH of the solution, it is not possible to use the pure ion exchange capacity. In order 3 to obtain Knutz in kg CaO/m , Ks has to be inserted in equations (22) and (23) as meq/g dry exchanger. Besides the unit kg CaO/m3, other units such

as lb CaC03/cu. ft, lb CaO/cu.ft, g CaO/lOOO ml or eq/cu. ft are also in use. In the case of anion exchangers, CaO equivalents are employed.

3. 13. Pure grain ion exchange capacity Kk. r

By a combination of, e. g. radioactive tracer or microcoulometric methods with microscopic methods, the ion exchange capacity of single exchanger grains can be determined. The obtained values (fieq/mg dry exchanger or neq/ grain) should again be converted to some standard state, i. e. the H+-form, the Cl'-form, or to the backbone of the exchanger. Capacity measurements of single exchanger particles yield information about the homogeneity of the material.

3. 14. Microscopic pure volume ion exchange capacity Km. v.r

In some cases the volume of a single ion exchanger particle can be determined by microscopic methods. It is then possible to refer the number of exchangeable ions to the dry volume \ or the wet volume Vq of the exchanger particle (peq/mra3 or meq/ml). The correlation between these two capacities is:

r (meq/ml dry particle) = ' 1 r (meq/ml swollen particle)(24) vt '

The correlation between the volume system and the weight system is:

(meq/g dry particle) = j ' K^ r (meq/ml dry particle) (25)

where dt denotes the density of the dry exchanger particle.

3. 15. Breakthrough capacity Kd

This quantity is used only in column operations. When a column, filled with an exchanger in the A-form, is leached with a solution containing counter-ions B, the exchange A for B will occur. At the beginning the effluent will contain only the ions A, but after some time the ion exchanger is converted to a large extent to the B-form, and B-ions will appear in

the effluent with a concentration cB, which is small compared to the

89 FIG. A. Breakthrough curve of the ions B. Originally only ions A were in the exchanger. original concentration of B, cB 0 (breakthrough). The amount of B sorbed by the ion exchanger up to this point is called the "breakthrough capacity" of the column. This quantity is always smaller than the pure ion exchange capacity, where the material is converted to the B-form under equilibrium conditions. The behaviour described above is illustrated graphically in Fig. A. Theoretically, the point of exhaustion of the column is reached after an infinite time or an infinite effluent volume. In practice, however, attainment of the equilibrium value between 1 and 5% will be sufficient. The same is true for the point of breakthrough (dashed lines in Fig. A).

The volume breakthrough capacity Kd v corresponds in Fig. A to the area F-j , if we refer to the volume of the packed bed. We thus have:

Kd,v = :f f (l--^)dV (26) VSvJq cBi0

The overall capacity of the column, corresponding to the area FJ+F2 is:

3 Kv.tot = ff (I-" ") dV (27) v B sv=o -°

The degree of utilization (efficiency, E) of the column is given by:

Kdy Fj

Kv.tot Fi + Ej

In practice it is convenient not to determine the whole breakthrough curve but only Vs, VD and VG. We then obtain approximately:

K (29) vs

(30)

90 It is clear that in connection with the value of the breakthrough capacity all operating conditions have to be specified in detail.

3. 16. Concentration of fixed ionic groups m,X

In many theoretical problems, the molality m of the fixed ionic groups in the resin is a very convenient quantity. Its unit is meq/g solvent in the exchanger phase. For the same reason, the volume concentration X of the fixed ionic groups can be defined, which is given as meq/ml swollen ion exchanger without interstitial volume. Both quantities depend strongly on the swelling properties of the sample and are not, therefore, characteristic constants of an exchanger. They vary with the nature of the absorbed ion, the concentration of the external solution and the temperature. If the water content of the exchanger, w' =w/M, is given as g solvent/g dry material, m is derived from the pure ion exchange capacity Kr as:

m =—p (meq/g solvent in exchanger) (32) w

If the amount of solvent w in the exchanger is given as g solvent/ g swollen ion exchanger, we have:

W m = Kr (meq/g solvent in exchanger) (33)

It should be noted that the value of Kr has to be determined in the same ionic form of the exchanger in which the water content has been determined. The volume concentration X of the fixed ionic groups can be determined from the density dA of the exchanger (g swollen exchanger/volume swollen exchanger) by:

X = ^ (meq/cm3 swollen exchanger) (34) or

3 X = Kr • dA (1 - w) (meq/cm swollen exchanger) (35)

The quantities m and X are correlated by:

- 3 X = dA w • m (meq/cm swollen exchanger) (36)

3.17. S-, T-, and V-values [10, 11]

When determining the ion exchange capacity of a clay mineral, it may be sometimes advantageous to denote the amount of exchangeable base forming cations, e. g. Ca2+, Mg2*, Na+, or K+/100 g dry weight as the S- value. The total cation exchange capacity of a clay, including ions such as H+ and Al3+, is then called the T-value. The fraction S/T is sometimes also called the V-value and is given as a percentage. For the difference between the T- and S-values, the term "rest capacity" is used.

91 3.18. Surface capacity Kfi

For comparative measurements of the ion exchange capacity of plant roots, it is advantageous to refer the capacity to the surface area of the roots rather than to their weight or volume. A reasonable unit in this connection is meq/m2 [12],

4. GENERAL METHODS OF DETERMINATION

4.1. Principles

The determination of ion exchange capacity includes: (1) quantitative determination of the milli-equivalents of the ion A in a given amount of ion exchanger AC, according to equation (1) (section 2), either by a non- destructive method, e. g. radiometry, after ashing or fusion of the loaded ion exchanger in solution, or after elution of A and its determination in solution by standard analytical methods; (2) determination of the dry weight of the exchanger in the A-form; (3) calculation from (1) and (2) of the quantity meq A/g dry AC; and (4) discussion of the errors. This procedure appears to be simple, but a number of conditions have to be fulfilled as follows: (1) before determination, the ion exchanger should be completely in the A-form; (2) the A-form should be H+, CI", or another defined standard ion. If other ions have to be used, the capacities obtained must be reduced to the standard forms according to equations (2) - (4) (sections 3.1., 3. 2.); (3) the dry weight has to be determined for the A-form of the exchanger; (4) if no precautions are taken, all ions A in the solid phase AC will be determined as a sum, regardless of whether they are bound by ion exchange, adsorption or precipitation; (5) a definite grain size has to be used if the exchanger shows an inhomogeneous distribution of ionic groups within the grain; (6) the sample has to be a single, pure, ion exchanger or at least a reproducible mixture; and (7) sampling of the exchanger material has to be representative for the whole material to be judged — a trivial, but very often unfulfilled requirement in the determination of natural ion exchangers. In the following, some of these points are briefly discussed in the order in which they occur during the determination.

4.2. Sampling

Errors due to sampling may exceed errors in the determination of the ion exchange capacities by an order of magnitude. It is, therefore, of fundamental importance that the sample to be analysed has exactly the same composition as the total material under investigation. The rules for sampling in chemical analysis should be observed; a geologist or soil chemist should be consulted if necessary.

4. 3. Sample preparation

4.3.1. Sieving

From inhomogeneous inorganic material, either only one distinct grain size fraction is taken or, if the whole material has to be analysed, e. g.

92 in the case of soil, a sieving curve is measured and samples from different grain size fractions are taken according to their proportion.

4.3.2. Purification

Natural inorganic ion exchange material can be contaminated by e. g. humic acids, calcite and other carbonates, quartz, feldspar, mica or graphite. Since these components also show ion exchange capacities (capacities for humic acids exceed those for clay minerals by about two orders of magnitude), a purification or at least exact characterization of the different solid phases is necessary. The latter may be achieved by methods of phase analysis, e. g. chemical methods, X-ray fluorescence, electron microscopy, light microscopy, spectroscopy or thermoanalytical methods. Humic acids in clay minerals may be destroyed by treatment with hydrogen peroxide [13, 14], Inorganic contaminants maybe separated by sedimentation or flotation.

4. 3. 3. Loading

According to equation (1) (section 2), a random ionic form of the ion exchanger can be converted in most cases to the A+ -form by treatment with a concentrated solution of the electrolyte A+ X". In order to get a complete shift of the equilibrium (1) to the right-hand side, repeated treatment is desirable. This can be done in column or batch operation. The latter, however, is commonly preferred in the case of samples with too small a grain size for column operation or long exchange durations. Criteria for a complete shift of the equilibrium and thus for preparation of the pure A+-form are: (1) concentration A+ of the loading solution before and after treatment of the ion exchanger remains constant; (2) after treatment, ions B+ are no longer in solution; and (3) no changes of the concentration of A+ and B+ according to (1) and (2) with time. The time necessary for complete conversion to the A+-form can vary from several hours to several months in the case of natural inorganic ion exchangers. In order to remove absorbed quantities of the electrolyte A+ X" after loading, the exchanger has to be washed thoroughly, e. g. with distilled water, until no further ions A+ can be detected in the effluent. If concentrated solutions of A+ X" were used, very long washing times may be necessary. In the case of weak acid ion exchangers, e. g. some clay minerals, washing with pure water can cause hydrolysis according to Me+ (clay mineral)" + + + + HzO ==5 H (clay mineral)" + Me + OH", and thus elution of ions A .

4.4. Quantitative determination

The amount of ions A+ per unit weight or volume ion exchanger A+ C" can be determined by direct and non-destructive, destructive or elution methods.

4. 4.1, Direct methods

The direct determination of the ions A+ in the ion exchanger is most rapidly and simply made by one of the modern non-destructive instrumental analytical methods. Of these, only the radiometric method has found

93 general acceptance until now, while X-ray fluorescence and some others have been used only occasionally.

4.4. 1. 1. Radiometric methods

Before loading the ion exchanger, a radioactive isotope of the same element in the same ionic form is added to the solution of the ion A+ in a sufficient amount to obtain precise measurements. For non-destructive radiometric measurements, gamma-ray emitters are preferred. They can be measured in a Nal-detector without mechanical or chemical preparation of the sample. Beta-ray emitters may also be used, but fine grinding to small particle size and spreading to a thin layer, as well as preparation of standard samples, are sometimes necessary. Alpha nuclides do not appear to be suitable. The radioactive ions used with the inactive carrier A+ have to fulfill several conditions: (1) monovalent ions in preference to polyvalent ions; (2) preferably long half-lives; (3) if possible, no radioactive daughter nuclides; (4) if radioactive daughter nuclides are unavoidable, their valency should be below or at least equal to that of the mother nuclide; and (5) gamma and especially beta energies not too low. Although radiometric determinations are rapid and convenient, problems of correct loading to the A+-form of the ion exchanger cannot be avoided. In the case of beta nuclides, preparation of standard samples is necessary.

4.4.2. Destructive methods

A standard method of analysis of A+ in A C" is the wet chemical determination, after destroying the matrix C" by incineration or fusion. Con- venient procedures for destroying organic ion exchangers are fuming with nitric acid in combination with sulphuric or perchloric acid, or treatment with hydrogen peroxide alone or with iron(II) (-OH radicals). Inorganic silicate ion exchangers may be fused with soda and/or hydro- fluoric acid.

4. 4. 3. Elution methods

With the methods described in 4. 4. 1. and 4. 4. 2. , it is not possible to determine whether the ions A+ are bound by ion exchange, precipitation, or by adsorption. In the elution (solution) methods, according to:

A+ + B+ [c]" =^A+ jc]" + B+

either the eluted ions B+ in the effluent or the replaced ions A+ in the loading solution can be determined as a difference in both column or batch operations. For the latter, titration curves of the ion exchanger may also be used. The direct determination of the eluted ions B+ in the effluent allows the determination of pure ion exchange capacities by the choice of a suitable elution agent. They are the true reversible capacities if they remain constant with repeated loading and elution. As in most other cases, complete and thorough loading of the exchanger with ions A+ is necessary. Since B+ is directly determined in solution, the method is also well suited for the determination of small ion exchange capacities (for B+).

94 The determination of the amount of ions A+, fixed during loading of the ion exchanger with a highly concentrated solution of A+, is based on the determination of the small concentration of ions A+ left in solution after loading. No differentiation of the binding state of A+ can be made. It is, however, the only method which also allows incomplete loading of the exchanger with ions A+ before starting. Column operations have the advantage of automatic and therefore simple operation. In the case of long reaction times and very small grain sizes, however, batch operation is preferred.

REFERENCES TO APPENDIX II

[1] AMPHLETT, C.B., Inorganic Ion Exchangers, Elsevier, New York (1964). [2] BLASIUS, E., Chromatographische Methoden in der analytischen und prSparativen anorganischen Chemie, Enke Verlag, Stuttgart (1958). [3] GRIESSBACH, R., Austauschadsorption in Theorie und Praxis, Akademie Verlag, Berlin (1957). [4] HELFFERICH, F., Ion Exchange, McGraw Hill, New York (1962). [5] INCZEDY, J., Analytische Anwendungen von lonenaustauschern, Verlag der ungarischen Akad. d. Wiss., Budapest (1964). [6] KITCHENER, J. A., Ion Exchange Resins, Wiley, London (1957), [7] NACHOD. C., SCHUBERT J., Ion Exchange Technology, Academic, New York (1956). [8] SAMUELSON, O. , Ion Exchange Separations in Analytical Chemistry, Wiley, New York (1963). [9] BAETSLE, L., PELSMAEKERS, J., J. inorg. nucl. Chem. 21 (1961) 124. [10] STACH, H., Angew. Chem. 63 (1951) 263. [11] HOFMANN, U. , GIESE, K., Kolloidzeitschrift 87 (1939) 21. [12] SMITH, R. L., WALLACE, A., Soil Sci. 81(1956) 97. [13] WEISS, A., Z. anorg. allg. Chem. 297 (1958) 232. [14] BOEHM, H.P., LIESER, K.H., Z. anorg. allg. Chem. 304 (1960) 207.

APPENDIX III

COMMUNICATIONS CONCERNING NATIONAL EXPERIENCE IN VARIOUS COUNTRIES

1. FRANCE

Studies are being made on the solidification of evaporator concentrates into a concrete, improved by the addition of minerals. The various minerals investigated include sepiolite, activated aluminium oxide and crude vermiculite. The best results, with respect to the hardness of the final product, the overall volume reduction and the homogeneity, were obtained with sepiolite. A typical composition is as follows: evaporator concentrates, 50%; cement, 25%; and sepiolite, 25%. The average chemical composition (in g/litre) of the evaporator concentrates (containing 800 g dry matter/litre) is: NaN03, 357; Ca(OH)2, 90; Fe, 11.5; S04, 108; CI, 22.5; P, 22.5; and F, 22. 5. Evaporator sludges are also being solidified in a vermiculite mortar. The facility includes a storage bin for the cement, introduced by blowing, and a storage bin for the vermiculite, introduced by gravity. The sludge and the vermiculite-cement mixture are both introduced into 400 1 or 750 1 drums by means of a special loading lid; the contents of the drums are mixed thoroughly with an electric mixer. The drums are then sealed in a prefabricated concrete container. A typical dosage (litres) for a 400 I drum is as follows: concentrate, 250; cement, 125; and vermiculite, 80.

2. GERMANY, FEDERAL REPUBLIC OF

At the Hahn-Meitner-Institut fur Kernforschung (HMI), Berlin, a pelagonite tuff which is commercially available under the name "Filtrolit" has been used for some time for selective removal of Cs and some other radionuclides from radioactive effluents. The tuff is found in the Eifel area near Bonn. Filtrolit has been used for about 30 years for water softening and filtration. The material has an excellent column performance. The pilot plant used for both test runs and laboratory effluent decontamination consists of two columns, 18 and 30 cm in diameter and 180 cm in length, filled with 45 1 and 100 1 respectively of the tuff. Usually a specific flow rate of 50 1/h per litre was chosen. The grain size of the tuff particles was 0.3-1 mm and the tuff was pre-treated with NaCl solution. Test runs showed that about 1500 bed volumes of tap water containing 10-5 N CsCl + 137Cs could be passed through the tuff until breakthrough occurred. The mean decontamination factor achieved was of the order of 500. Further, it was observed that the tuff has increasing selectivity not

97 only for Cs, but also for Sr, Ba and Ce, with decreasing relative concentration of these ions in the solution. This behaviour renders the tuff suitable for the decontamination of laboratory effluents, although the Sr-selectivity is rather poor. The pilot plant has been used for the decontamination of various laboratory effluents, and it has been demonstrated that the decontamination is very good for normal effluents from chemical laboratories but it is rather poor for laundry effluents (DF < 50). In the case of laundry effluents and for strongly acid solutions, the tuff becomes irreversibly damaged.

3. ITALY*

A natural mineral of the zeolite type has been used on a large scale at the Radioactive Waste Treatment Station of the Casaccia Nuclear Study Centre. The grain sizes of this zeolite mineral are such that it can be used for different techniques depending upon the particle size selected: (1) the fraction above 230 mesh size is suitable for contact methods; (2) the 100 - 230 mesh fraction can be used for fluidized bed techniques; and (3) the 20 - 100 mesh fraction can be used for fixed bed techniques. In some pilot scale studies, zeolitic tuff has been investigated in suspension alone and also as an additive during ferrocalcium phosphate flocculation treatment. These investigations have shown that this zeolitic tuff has practical advantages, particularly in view of its low cost, and that there exists a definite possibility of decontaminating solutions by a single- stage treatment, instead of the two treatments in series now commonly used. Investigations on the use of the zeolitic tuff in a fixed bed operation have revealed that the mineral has similar permeability characteristics to those of resins. In large-scale fixed bed operations, the effect of three parameters, i.e. the nature of the exchangeable ion, the depth of the bed and the flow rate, was studied. The zeolitic tuff has also been studied in a mixed installation consisting of a flocculator and a column arranged in series. The operating conditions were: throughput, 10 litres/h; retention in clariflocculator, 2 h; height of column, 50 cm; and pH of solution, 9.5. Flocculation with ferrocalcium phosphate alone gave an activity removal of 72%. The addition of the zeolitic tuff (230 mesh) raised this figure to 92%, and the further addition of a column resulted in a removal of more than 99. 9%. The breakthrough of Sr and Ru was reached after passage of 440 bed volumes, while for caesium, breakthrough was reached after 300 bed volumes had passed. This zeolitic tuff has been in use for more than a year on a plant scale as an additive in the flocculation process used to decontaminate low activity solutions (10"3 - 10"4 juCi/ml) containing decayed fission products. The treatment is carried out in a static flocculator with a capacity of 3.5 m3 . Tuff (1000 ppm) in powder form (230 mesh) is used with the ferrocalcium phosphate treatment. The activity removal obtained under these conditions is of the order of 90%.

Material provided by G. Branca andG. Gresson, CNEN, CSN Casaccia, S.M. Galeria, Rome.

98 4. KOREA

Readily available and extensively studied clay minerals at the Atomic Energy Research Institute, Seoul, Korea, are kaolinite, montmorillonite and vermiculite. A three-stage waste treatment facility, in which low-level liquid waste (10-4 -10'6 /uCi/ml) is absorbed by montmorillonite clay, was constructed in the waste treatment plant in 1964 and has since been in operation. The facility consists of three lead-lined 3 mm stainless steel reactor tanks and three settler tanks. It was found that optimum operating conditions are given by a waste feed flow rate of 2 litres/min and a montmorillonite clay dose of 2% by weight. Additions of higher percentages of clay were found not to be useful.

5. UNION OF SOVIET SOCIALIST REPUBLICS

The two natural sorbents investigated for large-scale treatment of radioactive effluents in the USSR are dolomite and pyrolusite. The studies carried out with dolomite have established that: (1) a satisfactory degree of decontamination is obtained when the height of the dolomite bed is greater than 1 m and the grain size is 0.5-1.0 mm; on increasing the bed height to greater than 1.5m, there is no increase in the degree of decontamination; (2) i44Ce is effectively and almost completely removed with the passage of 490 column volumes; 106Ru is removed even more effectively, and with the passage of 700 volumes, no breakthrough of 106Ru was observed; (3) the mineral is significantly more useful for removing radiostrontium than radiocaesium; and (4) the pH of the feed does not have a significant effect on the decontamination factor. The manganese ore (pyrolusite) is another mineral which has been found to decontaminate strontium quite effectively. The results in dynamic conditions establish that decontamination factors as high as 100 000 are obtainable using a grain size of 0. 25 - 1 mm. Almost complete desorption

of strontium is achieved with two volumes of 5% HN03 .

Special Communication, by V. M. Sedov, USSR

RESEARCH IN THE USSR ON THE USE OF NATURAL SORBENTS FOR RADIOACTIVE WASTE TREATMENT

INTRODUCTION

In the USSR, within the last decade, synthetic ion exchangers have come into wide use for the treatment of radioactive waste water of low salt content. They possess the advantages of a high ion exchange capacity, good exchange kinetics, and poor packaging (compaction) ability. Soviet scientists have also carried out extensive investigations on the use of various natural local sorbents for liquid waste treatment. Although these materials have to date found no wide application on an industrial scale, a number of scientists consider that these ion exchangers will be of

99 TABLE B. DATA FOR PILOT AND INDUSTRIAL RADIOACTIVE WASTE TREATMENT PLANTS

Sorbent Solution flow Sorbent layer Expected Sorbent use Solution Ion exchange Sorbent grain size through column thickness decontamination under dynamic conditions pH capacity (mm) (m/h) (m) factor

Dolomite In use 0.5-1.0 No effect on 1.5 1.5 For total isotope a decontamination content: factor 35S, 9«Sr, «Zr, »5Nb, i06Ru, 144Ce, -25

90 Pyrolusite To be used 0.25-1.0 - - - For Sr, 100 000 Some meq/g

Iron May be used - 4-7 1.5 - For 337 Cs, 2600 7 mg Cs/g ferrocyanide precipitate

a Nearly complete removal of 144Ce by passage of 490 column volumes; 106 Ru is fully removed by passage of 700 column volumes. major importance in liquid radioactive waste treatment, and investigations of the problems of large-scale, use of natural sorbents are therefore being made. Compared with synthetic resins, natural ion exchangers offer the following advantages: (1) comparatively low cost, for example, in the USSR the cost of mineral sorbents is between 20 and 100 roubles/ton, whereas that of synthetic resins is between 800 and 8000 roubles/ton; (2) stronger fixation of radioactive isotopes, a property which can be considerably improved by thermal treatment; (3) higher selective sorption capacity in the case of certain natural sorbents, e.g. that of clinoptilolite for Cs; and (4) selective sorption of some radioactive isotopes at high salt concentrations. The shortage of plants producing synthetic resins is the principal reason for recourse to local natural minerals for radioactive liquids treatment. This is particularly the case in a number of developing countries, since the import of resins from other countries involves a large financial outlay. In the choice of natural sorbents of local origin, the following factors have to be considered: (1) physical, chemical and radiochemical composition of the wastes to be treated; (2) degree of decontamination required; (3) ion exchange capacity of the sorbents; (4) stability of radioisotope fixation; (5) mechanical strength of the sorbent; (6) conditions of used active sorbents disposal; (7) cost of extraction and partial treatment of sorbents; and (8) method of sorbent application, which may depend on the conditions of waste treatments, i. e. static or dynamic, filtration or pre coat.

USE OF NATURAL SORBENTS FOR RADIOACTIVE WASTE TREATMENT IN THE USSR

In the USSR, research has been carried out, and is still underway, on radioactive waste treatment with the aid of the following natural sorbents: dolomite, glauconite, vermiculite, biotite, minerals containing phosphorus, various zeolites, pyrolusite, perlite, diatomite, peat, and other artificial inorganic ion exchangers. Some data for pilot and industrial radioactive waste treatment plants in the USSR are presented in Table B; characteristics of a number of natural sorbents are discussed in more detail below.

Dolomite

Most frequently, dolomite is used in the calcined form (calcination at 720 - 750° C). The mechanism involved is assumed to be as follows: strontium removal is by chemisorption on the grain surface and co-precipitation with recrystallization of the magnesium mass; cerium and phosphorus are adsorbed on the surface, due to the formation of compounds of low solubility (cerium hydroxides and calcium and magnesium phosphates); and caesium is fixed on the silicate components of the dolomite. As a result of research carried out with dolomite in water containing 144 Ce, 95Zr, 95Nb, 106Ru, 32P, 85Sr and 35S, Soviet scientists have found

101 that: (1) a satisfactory degree of decontamination for all the isotopes is obtained when the thickness of the dolomite layer is greater than 1 m and the grain size is 0.5-1.0 mm; (2) a dolomite layer thickness of more than 1.5m does not increase the degree of decontamination; (3) the pH of the solution does not affect the decontamination factor; (4) it is more advantageous to use the magnesium mass in dynamic conditions; (5) 144Ce is effectively and almost completely removed by the passage of 490 column volumes. 106Ru is removed even more effectively; after the passage of 700 column volumes no breakthrough of radioruthenium could be observed; and (6) dolomite has poor selectivity for 9°Sr and is ineffective in the case of 137Cs; it has no effect whatsoever in the extraction of 131I. In these studies, the fraction of semicalcined dolomite was 0.5-1 mm and 5-7 mm; column diameter 13 mm; column height 0. 5 - 4 m. The flotation of dolomite gives the best result, the volume of waste pulp being 0. 3% instead of 0. 6 - 1% in the case of coagulation. The activity of a solution containing 90Sr, 95Zr, 95Nb, 10(3Ru, 144Ce and 137Cs was reduced by a factor of 25.

Glauconite

Glauconite is a hydrous ferroaluminosilicate of potassium, sodium, calcium, magnesium and some other metals, and belongs to the group of hydromicas. The cation exchange is an irreversible process. The mineral is a satisfactory material for removal of strontium and caesium in dynamic conditions. The Soviet scientist Gornak has proposed a simple and effective method for increasing the ion exchange capacity of glauconite by heating in a reducing atmosphere, which makes it possible to increase the glauconite exchange capacity by a factor of six, from 0.17 to 1 meq/g.

"Vermiculite

Vermiculite is a natural magnesium-potassium or magnesium-alumina silicate. It is used mainly in the magnesium form, Mg2+, but is, however, preferably used in the Na+ form after treatment with a solution of NaCl, since at high pH values magnesium hydroxide is formed, which precipitates on the sorbent granules thereby reducing diffusion. Another way of eliminating this undesirable phenomenon is treatment at pH7. Vermiculite has a fairly high exchange capacity (up to 1. 5 meq/g; in dynamic conditions, up to 0.6 meq/g), and is more efficient in the removal of 137Cs and 90Sr in comparison with synthetic ion exchangers. The decontamination factors for 137Cs and 90Sr in dynamic conditions are in the range 102 - 103. Soviet scientists have shown that the heating of vermiculite to 150° C improves its filtration properties but reduces its ion exchange capacity. It would appear that, upon heating, a compaction of the inter-packet water layers takes place in the vermiculite and access of the ions to the Mg2+ exchange ions becomes difficult.

Biotite

Biotite is a mineral of the hydromica group, and has the approximate formula: (OH)4K2(Si6Al2) (Mg, Fe)602o. The structure of biotite is

102 similar to that of montmorillonite except that K+ is a compensating cation which is distributed in the spaces between the packets. Soviet scientists have studied the sorption of 137Cs and 90Sr on biotite in the presence of macroquantities of alkaline metal and magnesium ions. Biotite of 80 - 140 mesh was used, the finer fractions being separated beforehand. The initial waste contained a mixture of 90Sr, 90Y and 137Cs, and had a specific activity of 10"2 Ci/l, and pH6. 2-6.4. The exchange capacity was found to be 1X10"5 mol Me2+/g of sorbent. Biotite was also used for the treatment of another solution which had a specific activity of 1 X10-5 Ci/l, and contained Ca2+ in an amount equivalent to its content in rivers. The degree of 90Sr removal was slightly dependent on the composition of the solutions, amounting to 70 - 90%.

Minerals containing phosphorus

Spitsyn and associates have studied the sorption capacity of various minerals containing phosphorus. In the treatment of solutions with phosphorite, it was found that the degree of 90Sr removal varied between 80 and 96%. Phosphorite adsorbs 90Sr fairly effectively at Ca2+ ion concentrations of approximately 100 mg/1. The initial activity of the effluents was 1X10-5 Ci/l. Kibardin has investigated hydroxyapatite for use in removing 90Sr and some other stable metal isotopes. It was found to be efficient only where the pH was not less than 8. For a stable Sr content of 5X10'5 M, the degree of decontamination was 99%; if the stable strontium content was increased by a factor of 100, the degree of decontamination became 50%. Hydroxyapatite can be used most effectively in dynamic conditions, as was confirmed by a number of experiments in which this mineral was used for extracting Co, Cu and Fe ions from solutions of their sulphates at pH6. 8. These metals were not found in the filtrate. The disadvantage of using hydroxyapatite in dynamic conditions is its high degree of dispersion.

Manganese ores — Pyrolusite and ferrocyanides

A method has been developed and experiments carried out using mineral sorbents for removing radiostrontium and radiocaesium from solutions after alkaline precipitation. Because of the high salt content, the use of organic sorbents is not very effective. For 90Sr removal, domestic manganese ores with a grain size of 0. 25 - 1 mm were used. The sorption capacity was a few meq/g of sorbent, and the decontamination factor was as high as 100 000 in dynamic conditions. Almost complete desorption of strontium was achieved with two volumes of 5% HNO3. The removal of caesium was carried out using iron ferrocyanide on a granular carrier of activated charcoal. The iron ferrocyanide was reduced on the surface of the charcoal to a compound which can be considered as ferrocyanide. Quantities in the range 100 - 200 g of precipitate were applied per litre of the charcoal (weight 250 g). For a caesium content of 5 mg/1, approximately 7 mg Cs/g of sorbent were sorbed. Sorption took place at pH4-7, and the decontamination factor was 2600. The sorbent was

103 regenerated with 5% NaOH. The original waste containing Sr and Cs with a specific activity of 10"2 Ci/1 was brought to the maximum permissible concentration (MPC) by two-stage ionization.

FIG.B. Incorporation of radioactive sludge into a ceramic material based on aluminium phosphate binding.

SOLIDIFICATION OF RADIOACTIVE CONCENTRATES

A method has been developed in the USSR for incorporating active sludge into a ceramic material based on an aluminium phosphate binding (see Fig. B). The ceramic material is obtained as a result of binding the sludge particles with the aluminium phosphate. The latter is formed by introducing into the sludge two components of refractory clay and phosphoric acid (refractory up to 1900°C). The ceramic blocks are resistant to hydration and retain about 65 - 80% of wastes. The volume of the wastes is reduced by a factor of 40, and the elution rate is 10"5 - 10-4 g/cm2/day.

BIBLIOGRAPHY TO SPECIAL COMMUNICATION BY V. M. SEDOV

KUZNETSOV, Yu.V. etal., Principles of Water Deactivation, Atomizdat (1968) (in Russian).

SPITSYN, V.I., BALUKOVA, V.D., NAUMOVA, A . F., GROMOV, V.V., SPIRIDONOV, F.M., VETROV, E.M., GRAFOV, G.I., "A study of the migration of radioelements in soils", Int. Conf. peaceful Uses atom. Energy (Proc. Conf. Geneva, 1958) 1J3, UN, Geneva (1958) 439.

104 BAGRETSOV, V.F., PUSHKAREV, V.V., Radiokhimiya 2 (1960) 446.

PUSHKAREV, V.V., Atomn. Energ. 20 (1966) 53.

ZAITSEV, B.A., Rep. Conf. Members COMECON, Brno (1964) (in Russian).

PUSHKAREV, V.V. et al., Radiokhimiya 4 (1962) 49.

TYUTRINA, A.P., Atomn. Energ. 18 (1965) 56.

BALUKOVA, V.D., KULICHENKO, V.V., NAZAROV, A.I., SIBIREV, A.V., RAUZEN, F.V., Practices in the Treatment of Low-and Intermediate-level Radioactive Wastes (Proc. Symp. Vienna, 1965), IAEA, Vienna (1966) 841.

KIBARDIN, S.A., Prikl. himija 36 (1964) 2757.

BALUKOVA, V.D. et al., The incorporation of radioactive waste, Atomn. Energ. 8 (1967) (in Russian).

GROMOV, V.V., Atomn. Energ. 17 (1964) 13.

BAGRETSOV, V.F. et al., Radiokhimiya 6 (1965) 137.

VOZNESENSKY, S.A. et al., Radiokhimiya 3 (1961) 510.

GORNAK, V.M., Ion Exchange and Sorption from Solutions, Minsk. Publ. Acad. Sci. Byelorussian SSR (1963) 149 (in Russian).

BAGRETSOV, V.F., Prikl. himija 34 (1961) 11.

KHONIKEVICH, A.A., Dezaktivacija Sbrosnych Vod, Atomizdat, Moscow (1964).

105

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112 LIST OF PARTICIPANTS

PANEL ON THE USE OF LOCAL MINERALS IN THE TREATMENT OF RADIOACTIVE WASTE, IAEA, VIENNA, MAY 1969

Chairman

C. GAILLEDREAU France

Panel Members

L. BERÁK Czechoslovak Socialist Republic D. HAWKINS United States of America H. KRAUSE Federal Republic of Germany Sang Hoon LEE Korea V. SEDOV Union of Soviet Socialist Republics K.T. THOMAS India N. Vàn de VOORDE Belgium

Advisers

M. PIRS Yugoslavia B. SANSONI Federal Republic of Germany

Scientific Secretary

E.W. WIEDERHOLD Division of Health, Safety and Waste Management (now Division of Nuclear Safety and Environmental Protection), IAEA, Vienna

113

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ARGENTINA Comisión Nacional de Energia Atómica, Avenida del Libertador 8250, Buenos Aires AUSTRALIA Hunter Publications, 58 A Gipps Street, Collingwood, Victoria 3066 BELGIUM Office International de Librairie, 30, avenue Marnix, Brussels 5 CANADA Information Canada, Ottawa C.S.S.R. S.N.T.L., Spálená 51, Prague 1 Alfa, Publishers, Hurbanovo námestie 6, Bratislava FRANCE Office International de Documentation et Librairie, 48, rue Gay-Lussac, F-75 Paris 5e HUNGARY Kultura, Hungarian Trading Company for Books and Newspapers, P.O.Box 149, Budapest 62 INDIA Oxford Book and Stationery Comp., 17, Park Street, Calcutta 16 Prakash Publishers, Film Colony, Chaura Rasta, Jaipur-3 (Raj.) ISRAEL Heiliger and Co., 3, Nathan Strauss Str., Jerusalem ITALY Agenzia Editoriale Commissionaria, A.E.I.O.U., Via Meravigli 16, 1-20123 Milan JAPAN Maruzen Company, Ltd., P.O.Box 5050, 100-31 Tokyo International NETHERLANDS Martinus Nijhoff N.V., Lange Voorhout 9—11, P.O.Box 269, The Hague PAKISTAN Mirza Book Agency, 65, The Mall, P.O.Box 729, Lahore-3 POLAND Ars Polona, Centrala Handlu Zagranicznego, Krakowskie Przedmiescie 7, Warsaw ROMANIA Cartimex, 3-5 13 Decembrie Street, P.O.Box 134-135, Bucarest SOUTH AFRICA Van Schaik's Bookstore, P.O.Box 724, Pretoria Universitas Books (Pty) Ltd., P.O.Box 1557, Pretoria SWEDEN C.E.Fritzes Kungl. Hovbokhandel, Fredsgatan 2, Stockholm 16 U.S.S.R. Mezhdunarodnaya Kniga, Smolenskaya-Sennaya 32-34, Moscow G-200 YUGOSLAVIA Jugoslovenska Knjiga, Terazije 27, Belgrade

Orders from countries where sales agents have not yet been appointed and requests for information should be addressed directly to: /j£L\ Polishing Section, nternat VryP ß ' '°nal Atomic Energy Agency, ^^ Kärntner Ring 11, P.O.Box 590, A-1011 Vienna, Austria INTERNATIONAL ATOMIC ENERGY AGENCY VIENNA, 1972

PRICE: US$4.00 SUBJECT GROUP: II Austrian Schillings 93,- Health, Safety and Waste Management/ (£1.60; F.Fr. 20,-; DM 1270) Waste Management