the thermal conversion of contaminated soil into
TRANSCRIPT
THE THERMAL CONVERSION OF
CONTAMINATED SOIL INTO
CARBONACEOUS ADSORBENTS
A thesis submitted to the University of London in partial fulfilment of the
requirements for the degree of Doctor of Philosophy in the Faculty of Engineering and
for the Diploma of Imperial College
by
Geoffrey David Fowler B.Sc. (Hons) G.R.S.C.
December 1995
Centre for Environmental Control and Waste Management
Department of Civil Engineering
Imperial College of Science, Technology and Medicine
University of London
ABSTRACT
Contaminated soil is a major problem in many industrialised countries. The contamination is often
detrimental to the health of flora, fauna and people living on or near the land and can be reactive
towards any buildings and services. Most treatments for these sites rely upon contamination
excavation and disposal into a landfill, or its covering with clean soil. In both of these scenarios,
the solution is not sustainable because the pollution still has potential to cause harm. Permanent
treatments for contaminated soils are required. The research presented within this thesis has
demonstrated one possible solution, for an especially problematic contaminated soil resulting from
the operation of gaswork facilities. The nature of the gas manufacturing process and the poor
waste disposal practices operated were such that the typical pollutants in gaswork soils include:
coal tars, cyanide, sulphur and heavy metals. Cost-effective treatment techniques for this type of
mixed pollution are not available because current practical methodology will treat only one of the
components.
The increased stringency of environmental legislation has required industry to clean their effluents
to much higher standards. Historically, one of the most versatile materials to perform this task
is activated carbon, consequently a large market for activated carbon exists. Activated carbon
is manufactured mainly from non-renewable feedstock, using established industrial processes, but
with concerns over the depletion of natural resources, the need for new feedstock has resulted in
waste materials being evaluated as possible precursors. Contaminated gaswork soil has been
demonstrated to be a novel activated carbon feedstock.
Gaswork soil was sampled and characterised for carbon, free sulphur, total sulphate and total
cyanide content. Thermogravimetric analysis was conducted to establish the thermal characteristics
of the soil upon exposure to various activated carbon manufacturing procedures. The ZnCl;
process was shown to be most suitable to carbonise and activate the heavily polluted soil and
produce porous carbonaceous materials. The temperature range used was between 175 and
600 °C, which is considerably lower than other thermal processing technologies for contaminated
soil. The pollution levels of the soils were reduced significantly by this treatment. The cyanide
content was 99% lower and up to 60% sulphate and sulphur reduction was realised. Utilising X-
Ray Fluorescence, the retention of metal species by the char was also observed. Testing of the
soils and carbons for leachable TOC showed a >99 % reduction, implying that the process was
totally effective in eliminating organic contamination from the soil.
The adsorption characteristics developed by the soil were shown to be highly variable, which was
attributed to the heterogenous nature of the samples. Typically, the samples exhibited nitrogen
adsorption surface areas of 100 - 600m^/g. Aqueous adsorption studies demonstrated that effective
adsorption of phenol and p-nitrophenol (up to 95%) was achieved. These adsorbents were
especially adept in the removal of selected toxic metals, particularly Cd and Hg. In this respect,
the soil derived carbons were far superior to commercial activated carbons. The subsequent
application of the carbon product to the treatment of landfill leachate and metal plating waste
water was demonstrated and shown to be feasible.
ACKNOWLEDGEMENTS
Acknowledgements are never easy to write, and the task of writing this one has been made
infinitely harder by the sudden death of Professor Roger Perry, just before the completion of this
thesis. The expression of my thanks to him for providing me with the opportunity to conduct this
research, and advice which he Aeely provided (which at the time I did not perhaps realise was
useful) seems wholly inadequate when considering the tremendous loss which his death means
to everyone who knew or worked with him.
I wish to express my deepest gratitude to Dr Chris Sollars for his infinite patience and the many
hours spent in meetings examining data and suggesting possible alterations or indicating possible
flaws with the research being performed.
The technical assistance of the following people greatly contributed to the research by keeping
the laboratory equipment operating. Especial thanks to Chris Pitches and his colleagues who
repaired and modified the carbolite furnace on several occasions, as well as numerous other
things; John Burgess for TGA assistance; Mark Cowan of Coulter Electronics Ltd. for Omnisorp
assistance; Ken Wheatley; Mark Scrimshaw and Richard Kermack.
My interest in the most obscure references, and the patience with which it was always sated by
Janice Lewis in Lyon Playfair Library or Kay Crooks and Susanah Parry in Civil Engineering
Library was greatly appreciated.
Finally, I must thank my colleagues at Imperial and friends who provided support and friendship
when it was needed, particularly Ed Butcher and Ben Purcell. Thanks also to; Judith Barritt,
Jenny Bubb, Sarah Buss, Ivan Gee, Nikki Meakins, Alastair McCullough and Paul Johnston.
Too many other people have been in contact with the research presented within this Thesis to be
mentioned personally, so I thank them now for the time and patience which they may or may not
have afforded me.
" If one buys land on which there is a slag heap 120 ft. high and it costs £100,000 to remove that
slag, that is not land speculation in the sense that we condemn it. It is land reclamation "
Lord Harold Wilson, Parliamentary speech, Hansard 4th April 1974, col. 1441,
" It is exciting to have a real crisis on your hands, when you have to spend half your political life
dealing with humdrum issues like the environment "
Lady Margaret Thatcher, on the Falklands campaign, 1982.
A speech to the Scottish Conservative Party conference.
This thesis dedicated to my family
TABLE OF CONTENTS
ABSTRACT 2
ACKNOWLEDGEMENTS 3
TABLE OF CONTENTS 5
LIST OF TABLES 11
LIST OF FIGURES 13
LIST OF PLATES 17
GLOSSARY OF ABBREVIATIONS AND UNITS OF MEASURMENT 18
CHAPTER I
INTRODUCTION AND LITERATURE REVIEW 20
1.1 Introduction 20
1.2 Waste definitions, arisings and waste legislation 21 1.2.1 Defining waste 21 1.2.2 Waste arisings 23 1.2.3 A European perspective to environmental legislation 24
1.2.3.1 United Kingdom environmental legislation 25 1.2.3.2 European Community (EC) environmental legislation . . . 26 1.2.3.3 EC environmental directives and action programmes . . . 27
1.2.4 The nett effect of waste legislation 29 1.3 Contaminated soil 30
1.3.1 Introduction 30 1.3.2 Defining contaminated soil or land 30 1.3.3 The causes and hazards of contaminated land 31 1.3.4 Gaswork sites - their history and the nature of the contamination
resulting from their operation 35 1.3.5 Policy and practice towards contaminated land 38
1.3.5.1 The EC 38 1.3.5.2 The United Kingdom 39 1.3.5.3 Other industrialised countries 43
1.3.5.3.1 The Netherlands 43 1.3.5.3.2 The United States of America 44
1.3.6 Guidance for assessing the degree of contamination to land 45 1.3.6.1 Canada 45 1.3.6.2 The Netherlands 46 1.3.6.3 United Kingdom 46
1.3.7 Redevelopment of contaminated land 46 1.3.8 Remedial treatment methodology for contaminated soil 47
1.3.8.1 Thermal treatment technologies for contaminated soil . . . 49 1.3.8.1.1 Incineration technologies 49 1.3.8.1.2 Incinerators specifically designed for
contaminated soil 52 1.3.8.1.3 Thermal evaporation systems 53
1.3.8.1.4 Vitrification 55 1.3.8.2 Other soil treatment methods 57
1.3.8.2.1 Bioremediation 57 1.3.8.2.2 Chemical 57 1.3.8.2.3 Encapsulation 58 1.3.8.2.4 Non-thermal extraction 58
1.3.9 The economics of soil rehabilitation technologies 59
1.4 Adsorbents 61 1.4.1 Introduction 61 1.4.2 The characteristics of adsorption 61
1.4.2.1 Mathematical treatment of adsorption 64 1.4.2.1.1 The Freundlich equation 64 1.4.2.1.2 The Langmuir equation 66 1.4.2.1.3 The Brunauer, Emmett and Teller (BET)
equation 67 1.4.2.1.4 The t-plot 68
1.4.3 Porosity 69 1.4.3.1 Determination of porosity 70 1.4.3.2 Type IV isotherms, capillary condensation and the Kelvin
equation 70 1.4.3.3 The Horv^th-Kawazoe equation 72
1.5 Carbonaceous adsorbents 73 1.5.1 The history of activated carbons 74 1.5.2 The structure of activated carbons 74 1.5.3 The manufacture of activated carbon 75
1.5.3.1 Carbonisation and activation 76 1.5.3.1.1 The chemistry of carbonisation 76 1.5.3.1.2 The chemistry of activation 78
1.5.4 The surface characteristics of activated carbons 79 1.5.5 Applications of activated carbons 80
1.6 Problematic aqueous waste streams 81 1.6.1 Metal species in the envirormient 81 1.6.2 Sources of metal effluents 82
(i) Landfill sites 82 (ii) Metal plating effluents 84
1.7 Summary 85
CHAPTER TWO EXPERIMENTAL RATIONALE 88
CHAPTER THREE SOIL CHARACTERISATION AND SELECTION OF AN ACTIVATION PROCEDURE
89
3.1 Introduction 89 3.1.1 Nomenclature for sample identification 89
3.2 Experimental protocols 93 3.2.1 Introduction 93 3.2.2 Soil sampling, post-treatment and fractionation 93
3.2.3 Chemical characterisation of the soil 94 3.2.3.1 Introduction 94 3.2.3.2 CHN Analysis 94 3.2.3.3 Total cyanide, total sulphate and free sulphur analysis . . 94 3.2.3.4 Metal analysis 94
3.2.3.4.1 Acid digestion by method 3050 95 3.2.3.4.2 Acid Digestion for Hg 95
3.2.4 Screening of soils as potential activated carbon feedstock 95 3.2.4.1 Introduction 95 3.2.4.2 Thermal analysis 96
3.2.5 Activation agent selection 98 3.2.5.1 Physical activating agents 98 3.2.5.2 Chemical activating agents 98
3.2.6 Laboratory scale production of carbon adsorbents from STA05 99 3.2.6.1 The carbolite HTR 11/150 rotary furnace 99 3.2.6.2 Production of carbon from contaminated soil 99 3.2.6.3 Post-activation treatment of the product 100
3.2.7 Characterisation of soil derived activated carbons 101 3.2.7.1 CHN analysis 101 3.2.7.2 Total cyanide and sulphate and free sulphur analysis . . . 101 3.2.7.3 Evaluation of aqueous adsorption ability of carbons . . . . 101
3.2.7.3.1 Introduction 101 3.2.7.3.2 Determination of adsorption ability of soil
carbons 102 3.2.7.4 Surface area and porosity analysis by gas adsorption . . . 102
3.2.7.4.1 Introduction 102 3.2.7.4.2 Sample preparation for Omnisorp® analysis . . . . 103
3.3 Results and Discussion 104 3.3.1 Characteristics of the air-dried soil 104 3.3.2 Total cyanide, total sulphate and free sulphur analysis 106 3.3.3 Metal analysis of soil samples 108 3.3.4 Thermal analysis: Survey of activation procedures I l l
3.3.4.1 Carbonisation alone and physical activation I l l 3.3.4.2 Chemical activation 113
3.3.5 Characterisation of bulk carbon samples prepared from contaminated soil 116 3.3.5.1 CHN results 116 3.3.5.2 Aqueous adsorption results 116
3.3.6 Adsorption isotherm study of soil-carbons 117 3.3.6.1 Interpretation of the phenol and 4-nitrophenol adsorption
data 118 3.3.6.2 Langmuir adsorption isotherms 119 3.3.6.3 Freundlich adsorption isotherms 120 3.3.6.4 Gas adsorption data 122 3.3.6.5 Comparison of the Langmuir and BET surface areas . . . 124
3.4 Summary 125
CHAPTER FOUR ZINC CHLORIDE ACTIVATION OPTIMISATION 127
4.1 Introduction 127
4.2 Experimental protocols 127 4.2.1 Preparation of the soil-ZnCl2 samples 128 4.2.2 Thermal analysis of STAOl - STA04 and G1 128 4.2.3. Bulk carbon production 128
4.2.3.1 Optimisation of the activation agent dosage 128 4.2.3.2 Optimisation of carbonisation and activation residence
times 128 4.2.4 Analysis and characterisation of the soil-carbons 128
4.2.4.1 CHN analysis 128 4.2.4.2 Total cyanide, total sulphate and free sulphur analysis . . 129 4.2.4.3 Aqueous adsorption studies 129 4.2.4.4 Gaseous adsorption analysis 129
4.2.5 Effect of residence time 129 4.2.6 Optimised sample production 129 4.2.7 X-Ray analytical techniques 129
4.2.7.1 X-Ray Fluorescence analysis PCRF) 129 4.2.7.2 X-Ray Diffraction analysis (XRD) 130
4.3 RESULTS AND DISCUSSION I: ZnCl^ dosage optimisation 130 4.3.1 Thermal analysis of STAOl - STA04 and G1 130
4.3.1.1 Mechanistic aspects of activation and carbonisation by ZnClz 137
4.3.2 Sample yields and CHN results 138 4.3.3 Total cyanide, total sulphate and free sulphur analysis 140 4.3.4 Aqueous adsorption results 143
4.3.4.1 Influence of carbon on solution pH 143 4.3.4.2 Single point adsorptions 144 4.3.4.3 Mechanistic aspects of phenolic adsorption on activated
carbon 151 4.3.4.4 Adsorption isotherms 152
4.3.5 Gaseous adsorption 156 4.3.5.1 Surface area 156 4.3.5.2 Pore-size distribution analysis - soil carbons 161
4.3.6 Summary of the effect of varying ZnCl; loadings 169
4.4 RESULTS AND DISCUSSION II; Optimisation of carbonisation and activation dwell time 170
4.4.1 Introduction 170 4.4.2 C.H.N, analyses 170 4.4.3 Total cyanide, total sulphate and free sulphur analysis 171 4.4.4 Aqueous adsorption results 172 4.4.5 Gaseous adsorption 172 4.4.6 Adsorption results summary 174 4.4.7 Analysis of the bulk sample of optimised soil carbon 174
4.4.7.1 Chemical composition 174 4.4.7.2 Analysis of the wash waters 175 4.4.7.3 Aqueous adsorption 176 4.4.7.4 Gas adsorption 177 4.4.7.5 Summary 178
4.5 X-Ray Analysis 178 4.5.1 X-Ray Fluorescence 178 4.5.2 XRD Results 183
4.6 Summary 188
CHAPTER FIVE
METAL ADSORPTION AND APPLICATION EVALUATION 190
5.1 Introduction 190
5.2 Selection of the metals for investigation 190 5.2.1 Sources of the metal-containing effluents 191
5.3 Experimental protocol 192 5.3.1 Introduction 192 5.3.2 Preparation of reaction solutions 192
5.3.2.1 Buffer solution 192 5.3.2.2 Metal solutions 193
5.3.3 Adsorption studies 193 5.3.3.1 Single point adsorptions 193 5.3.3.2 Adsorption Kinetics 193 5.3.3.3 Isotherm analysis 193
5.3.4 Waste liquor analysis 194 5.3.4.1 Landfill leachate 194 5.3.4.2 Metal plating effluent 194 5.3.4.3 Adsorption testing 194
5.3.5 Leach testing 195
5.4 Results and Discussion 195 5.4.1 Effect of buffer upon solution metal concentrations 195 5.4.2 Single point adsorptions 196 5.4.3 Adsorption kinetics 197 5.4.4 Metal adsorption isotherms 199
5.4.4.1 Control of pH 199 5.4.4.2 Cadmium adsorption isotherms 201 5.4.4.3 Chromium adsorption isotherms 207 5.4.4.4 Copper adsorption isotherms 213 5.4.4.5 Mercury adsorption isotherms 219 5.4.4.6 Adsorption isotherm for a mixed metal solution 225
5.4.5 Analysis of the landfill leachate 231 5.4.6 Adsorption testing with the aqueous wastes 233
5.4.6.1 Landfill leachate 233 5.4.6.2 Electroplating effluent 236 5.4.6.3 Summary 238
5.4.7 Leach testing of the soils and carbons 238
5.5 Summary 242
CHAPTER SIX CONCLUDING DISCUSSION 245
6.1 Introduction 245
6.2 Assessment and treatment of the contaminated soils 248 6.2.1 Chemical analysis and activation agent selection for soil treatment 248 6.2.2 Production of the carbon with optimum adsorption properties and
minimum contaminant potential 250
6.2.3 Demonstrating potential applications for the soil-carbons 254
6.3 Summary 258
CHAPTER SEVEN
CONCLUSIONS AND SUGGESTIONS FOR FURTHER WORK 260
7.1 Conclusions 260
7.2 Suggestions for further work 263
CHAPTER EIGHT REFERENCES 266 APPENDIX I
INTERNATIONAL GUIDANCE VALUES FOR CONTAMINATION IN SOIL . . . 285
APPENDIX II THE LANGMUIR EQUATION . . . , 293
APPENDIX III THE BET EQUATION 296
APPENDIX IV FREUNDLICH AND LANGMUIR PHENOL ISOTHERM PLOTS FOR THE SOIL AND COMMERCIAL CARBONS 301
APPENDIX V FREUNDLICH AND LANGMUIR 4-NITROPHENOL ISOTHERM PLOTS FOR THE SOIL AND COMMERCIAL CARBONS 306
APPENDIX VI PUBLICATIONS DERIVED FROM THIS RESEARCH 311
10
LIST OF TABLES
Table 1.1
Table 1.2
Table 1.3 Table 1.4 Table 1.5
Table 1.6 Table 1.7 Table 1.8
Table L9 Table 1.10
Table 1.11 Table 1.12 Table 1.13 Table 1.14
Table 3.1 Table 3.2 Table 3.3 Table 3.4
Table 3.5 Table 3.6 Table 3.7 Table 3.8 Table 3.9
Table 3.10
Table 3.11
Table 3.12 Table 4.1 Table 4.2 Table 4.3 Table 4.4 Table 4.5 Table 4.6 Table 4.7 Table 4.8 Table 4.9 Table 4.10 Table 4.11 Table 4.12
Table 4.13 Table 4.14 Table 4.15
Disposal methods for solid waste in the EC and the USA expressed as a percentage 24 Important acts of Parliament controlling pollution of the environment in the UK 26 List I and list II substances, as defined by EC directive (76/464/EC) . . . 28 EC action programmes on the environment 29 A list of sites which will be potentially contaminated and the type of contamination which may be encountered 32 The proportion of contaminated sites in the UK 35 The five stages recommended when conducting a site restoration 47 The five sub-categories for soil treatment technologies and their applicability to soil treatment 48 Relative costs of different clean-up techniques 60 Interpretation of isotherm shape in terms of the physical characteristics of the solid under investigation 63 Feedstock suitable for activated carbon manufacture 75 Activated carbon usage (1972) 81 Summary of literature data on landfill leachate quality 84 USEPA effluent guidelines for electroplating plants discharging t 37.85 m' (10 000 gallons) per day into a municipal sewerage system . . . 85 Description of the soil samples 89 Parameters used for thermal analysis of soil samples 97 Soil sample characteristics 104 Total cyanide, total sulphate and free sulphur in soils STA01-STA05 and G1 107 Metal concentrations detected in the soils STAOl - STA05 and G1 108 Carbonisation and activation temperatures established for STA05 115 CHN values for the raw soil and soil derived activated carbons 116 Aqueous adsorption studies results 117 The adsorption capacity of each activated carbon for phenol and 4-nitrophenol estimated from data in figures 3.18 and 3.19 118 Langmuir and Freundlich constants derived from the phenol adsorption isotherms 121 Langmuir and Freundlich constants derived from the 4-nitrophenol adsorption isotherms 122 Compiled gas adsorption data 123 Manufacturers technical details for the commercial carbons 127 Temperatures chosen for carbonisation and activation 137 Sample identity, yields and CHN values 139 TotaJ sulphate, total cyanide and free sulphur analysis 141. Aqueous adsorption results 145 Chemical and adsorption data for FeSO^ activated sample 151 Langmuir and Freundlich constants for phenol adsorption 153 Langmuir and Freundlich constants for 4-nitrophenol adsorption 154 Langmuir surface areas for each sample 155 Parameters determined from the gas adsorption data 158 Pore volume parameters calculated from the gas adsorption data 169 Carbonisation and activation times investigated during process optimisation 170 Sample C. H. N. values 171 Total cyanide, total sulphate and free sulphur analysis 171 Single point adsorptions 172
11
Table 4.16 Surface area parameters determined from the gas adsorption data 173 Table 4.17 Comparison of runs 041 and 045 175 Table 4.18 Metid analysis of HCl wash liquors for optimised sample 175 Table 4.19 Langmuir and Freundlich constants for phenol adsorption 176 Table 4.20 Langmuir and Freundlich constants for 4-nitrophenol adsorption 176 Table 4.21 Langmuir surface areas for runs 041 and 045 177 Table 4.22 Metals detected by XRF in original soils and corresponding carbons . . . 179 Table 4.23 Crystalline phases detected by XRD 184 Table 5.1 Maximum permissable concentrations in water for selected metals
according to EU Drinking Water Directive (EEC, 1980a) 191 Table 5.2 Effect of buffer upon metal concentration in solution 195 Table 5.3 Adsorption by run 045 of Cd, Cr, Cu, Hg, Ni and Pb from 10 ppm
solutions 197 Table 5.4 Variation of solution pH in the buffered Cd isotherm solutions 199 Table 5.5 Adsorption isotherm parameters for Cd(II) 204 Table 5.6 Adsorption isotherm parameters for Cr 209 Table 5.7 Adsorption isotherm parameters for Cu(II) 216 Table 5.8 Adsorption isotherm parameters for Hg(II) 223 Table 5.9 Isotherm parameters for Cd(II) and Hg(II) adsorption from the mixed
metal solution 229 Table 5.11 pH and TOC value for the landfill leachate samples 231 Table 5.12 Landfill leachate metal analysis 232 Table 5.13 The treatment of an electroplating effluent with activated carbon 236 Table 5.14 Analysis of the metal plating sludge 237 Table 5.15 pH and TOC results of the 3 hour leach test 239 Table 5.16 pH and TOC results of the DIN leach test 240 Table AI.l Canadian assessment and remediation criteria. (All values, except pH, in
ppm) 285 Table AI.2 Dutch A, B and C values for selected soil pollutants 286 Table AI.3 The revise Dutch reference and intervention values for selected soil
pollutants 287 Table AI.4 Guidelines for contaminated soils, compiled by the GLC 288 Table AI.5 Tentative 'trigger concentrations' for selected inorganic contaminants . . . 289 Table AI.6 Tentative 'trigger concentrations' for contaminants associated with former
coal carbonisation sites 291
12
LIST OF FIGURES
Figure 1.1
Figure 1.2 Figure 1.3 Figure 1.4
Figure 1.5 Figure 1.6 Figure 1.7 Figure 1.8
Figure 1.9 Figure 1.10
Figure 1.11
Figure 1.12 Figure 1.13 Figure 1.14 Figure 1.15 Figure 1.16 Figure 1.17 Figure 1.18 Figure 3.1 Figure 3.2 Figure 3.3 Figure 3.4
Figure 3.5
Figure 3.6 Figure 3.7
Figure 3.8
Figure 3.9
Figure 3.10
Figure 3.11 Figure 3.12
Figure 3.13 Figure 3.14 Figure 3.15 Figure 3.16 Figure 3.17 Figure 3.18
Figure 3.19
Exposure pathways for animals, plants, humans and buildings to contaminants present in soils 34 The unit processes involved in the manufacture of gas 36 Selected PAHs which can all be found in coal tar 37 The proposed application of available soil treatment methods to contaminated sites in the USA during the period 1987-1989 49 A typical fluidised bed combustion chamber 50 A typical fixed hearth combustion chamber 50 A typical rotary kiln and after-burner combustion chamber 51 The combustion processes occurring within the kiln of a rotary incinerator 51 The operating sequence for in-situ vitrification of contaminated soil . . . . 56 A typical flow-sheet for the physical extraction of contamination from soils 59 Typical van der Waals forces between a molecule and a planar surface as a function of separation in molecular diameters 62 The BDDT classification for adsorption isotherms, types I to VI 63 The hysteresis loop classification proposed by lUPAC 64 The four general pore shapes for adsorbent systems 65 Idealised t-plot scenarios 69 The structure of activated carbon 76 Acidic oxide structures found on the surface of activated carbon 80 The five stages in the chemical life of a landfill 83 A schematic of the ST A 1500 simultaneous thermal analyzer 97 The Carbolite HTR 11/150 rotary furnace 100 The Coulter® Omnisorp® 100 103 The effect of different drying temperatures upon the soil moisture content 105 The effect of different drying temperatures upon the soil carbon content 105 The particle size distribution of sample G1 105 The distribution of contamination between the soil particle size fractions 107 The effectiveness of the digestion procedure using a standard reference material, BCR 143 109 The spike recoveries exhibited by the digestion of STAOl, STA04 and STA05 110 The effect of the soil sulphur content upon the spike recoveries for Co, CuandNi 110 Thermal analysis of STA05 in Nj I l l Evolution of the resistance of carbons during heat-treatment. An insulator-conductor transition occurs in the high temperature treatment range (H.T.T.) = 600 - 700 ^C 112 Thermal analysis of STA05 with CO; 113 Thermal analysis of STA05 in Nj then CO; 113 Thermal analysis of STA05 with H^SO^ 114 Thermal analysis of STA05 with HNO3 115 Thermal analysis of STA05 with ZnClj 115 Isotherms for phenol adsorption by activated soil carbons prepared using different activation procedures 118 Isotherms for p-nitrophenol adsorption by activated soil carbons prepared using different activation procedures 119
13
Figure 3.20 Figure 3.21 Figure 3.22 Figure 3.23 Figure 3.24 Figure 3.25 Figure 3.26 Figure 3.27
Figure 4.1 Figure 4.2 Figure 4.3 Figure 4.4 Figure 4.5 Figure 4.6 Figure 4.7 Figure 4.8 Figure 4.9 Figure 4.10 Figure 4.11 Figure 4.12
Figure 4.13
Figure 4.14 Figure 4.15 Figure 4.16 Figure 4.17 Figure 4.18 Figure 4.19
Figure 4.20 Figure 4.21 Figure 4.22 Figure 4.23 Figure 4.24
Figure 4.25
Figure 4.26
Figure 4.27
Figure 4.28
Figure 4.29
Figure 4.30 Figure 4.31 Figure 4.32 Figure 4.33 Figure 4.34 Figure 4.35 Figure 4.36
Langmuir isotherm for phenol adsorption 119 Langmuir isotherms for 4-nitrophenol adsorption (€0% and H^SOJ . . . . 120 Langmuir isotherms for 4-nitrophenol adsorption (ZnCy 120 Freundlich isotherm for Phenol adsorption 121 Freundlich isotherms for 4-nitrophenol adsorption 121 Nj isotherms for the carbonised and activated soil samples 123 Pore-size distributions for the activated soil carbons 124 Comparison of the surface areas calculated for the activated soil carbons from aqeuous and gaseous adsorption data 125 The thermal analysis traces for STAOl 131 The thermal analysis traces for STAOl plus ZnClj 131 The thermal analysis traces for STA02 132 The thermal analysis traces for STA02 plus ZnCI; 133 The thermal analysis traces for STA03 133 The thermal analysis traces for STA03 plus ZnCl; 134 The thermal analysis traces for STA04 134 The thermal analysis traces for STA04 plus ZnClj 134 Thermal analysis of sulphur flowers 135 The thermal analysis traces for G1 136 The thermal analysis traces for G1 plus ZnCl^ 136 The variation of solution pH for adsorption of phenol by the soil-carbons 143 The variation of solution pH for adsorption of p-nitrophenol by the soil-carbons 144 The effect of different ZnCl^ additions upon STAOl derived carbons . . . 146 The effect of different ZnClz additions upon STA02 derived carbons . . . 147 The effect of different ZnCl; additions upon STA03 derived carbons . . . 147 The effect of different ZnCl; additions upon STA04 derived carbons . . . 147 The effect of different ZnCl; additions upon STA05 derived carbons . . . 148 Effect of soil sulphur species content upon development of phenol relative activities 149 Thermal analysis of STA05 and FeSC^ 150 Thermal analysis of FeS04.7H20 150 The effect of ZnClj upon the phenol Langmuir monolayer capacity . . . . 154 N; adsorption isotherms for runs 001, 002, 004, 005 and 006 156 N; adsorption isotherms for the commercial carbons Norit SA4 and Type C 157 The effect of ZnCl; upon the gas adsorption data for carbons manufactured from soil STAOl 159 The effect of ZnCI; upon the gas adsorption data for carbons manufactured from soil STA02 160 The effect of ZnCl; upon the gas adsorption data for carbons manufactured from soil STA03 160 The effect of ZnCl; upon the gas adsorption data for carbons manufactured from soil STA04 161 The effect of ZnClj upon the gas adsorption data for carbons manufactured from soil STA05 161 Mesopore size distribution for STAOl carbons 162 Mesopore size distribution for STA02 carbons 162 Mesopore size distribution for STA03 carbons 163 Mesopore size distribution for STA04 carbons 163 Mesopore size distribution for STA05 carbons 164 Mesopore size distribution for commercial carbons 164 Horv&th and Kawazoe micropore size distributions for STAOl carbons . . 165
14
Figure 4.37 Horv^th and Kawazoe micropore size distributions for STA02 carbons . . 166 Figure 4.38 Horv&th and Kawazoe micropore size distributions for STA03 carbons . . 166 Figure 4.39 Horv^th and Kawazoe micropore size distributions for STA04 carbons . . 167 Figure 4.40 Horv^th and Kawazoe micropore size distributions for STA05 carbons . . 167 Figure 4.41 Horvkth and Kawazoe micropore size distributions for the commercial
carbons 168 Figure 4.42 N; adsorption-desorption isotherm for run 041 173 Figure 4.43 Mesopore size distribution for run 038 to run 044 173 Figure 4.44 Horvith and Kawazoe micropore size distributions for runs 038 to 044 . . 174 Figure 4.45 Mesopore size distribution for run 041 and 045 177 Figure 4.46 Horv^th and Kawazoe micropore size distributions for runs 041 and
045 178 Figure 4.47 XRF spectrum for run 001 180 Figure 4.48 XRF spectrum for run 002 181 Figure 4.49 XRF spectrum for run 004 181 Figure 4.50 XRF spectrum for run 005 182 Figure 4.51 XRF spectrum for run 006 182 Figure 4.52 XRF spectrum for run 045 183 Figure 4.53 X-Ray Diffraction pattern for soils STA01-STA05 and G1 185 Figure 4.54 X-Ray Diffraction pattern for carbons run 007 - run 011 and run 045 . . 186 Figure 4.55 X-Ray diffraction patterns for the commercial carbons Norit SA4 and
Chemviron Type C 187 Figure 5.1 Contact time study of Cd adsorption by carbon 197 Figure 5.2 Contact time study of Cr adsorption by carbon 198 Figure 5.3 Contact time study of Cu adsorption by carbon 198 Figure 5.4 Contact time study of Hg adsorption by carbon 198 Figure 5.5 Effect of pH upon solubility of Cd, Cr, Cu and Hg 200 Figure 5.6 Effect of pH upon solubility of Cd, Cr, Cu and Hg in a mixed solution . 201 Figure 5.7 Variation of solution pH for Cd(II) adsorption isotherm using solutions
without pH adjustment 202 Figure 5.8 Variation of solution pH for Cd(II) adsorption isotherm using pH adjusted
solutions 202 Figure 5.9 Cd(II) adsorption isotherm determined using pH adjusted solutions . . . . 203 Figure 5.10 Langmuir isotherms for Cd(II) adsorption from pH adjusted solutions . . . 203 Figure 5.11 Freundlich isotherms for Cd(II) adsorption from pH adjusted solutions . . 204 Figure 5.12 Variation of solution pH for Cr adsorption isotherms using solutions
without pH adjustment 207 Figure 5.13 Variation of solution pH for Cr adsorption using pH adjusted solutions
208 Figure 5.14 Langmuir isotherm for Cr adsorption by Run 045 from pH adjusted
solutions 208 Figure 5.15 Langmuir isotherm for Cr adsorption by Type C and Norit SA4 from pH
adjusted solutions 209 Figure 5.16 Freundlich isotherms for Cr adsorption from pH adjusted solutions . . . . 209 Figure 5.17 Variation of solution pH for Cu(II) adsorption isotherms using solutions
without pH adjustment 213 Figure 5.18 Ce vs x/m for Cu(II) adsorption from solutions without pH adjustment.
214 Figure 5.19 Langmuir isotherm for Cu(II) adsorption from solutions without pH
adjustment 214 Figure 5.20 Freundlich isotherm for Cu(II) adsorption from solutions without pH
adjustment 214 Figure 5.21 Variation of solution pH for Cu(II) adsorption isotherms using pH
adjusted solutions 215
15
Figure 5.22 Langmuir isotherms for Cu(II) adsorption from pH adjusted solutions . . . 215 Figure 5.23 Freundlich isotherms for Cu(II) adsorption from pH adjusted solutions . . 215 Figure 5.24 Variation of solution pH for Hg(II) adsorption isotherms using solutions
without pH adjustment 219 Figure 5.25 Hg(II) adsorption isotherms determined using solutions without pH
adjustment 220 Figure 5.26 Langmuir isotherm for Hg(II) adsorption from solutions without pH
adjustment 220 Figure 5.27 Freundlich isotherm for Hg(II) adsorption from solutions without pH
adjustment 220 Figure 5.28 pH of Hg(II) solutions after adsorption, initial pH adjusted to 5.5 221 Figure 5.29 Hg(II) adsorption isotherms determined using solutions with pH
adjustment 221 Figure 5.30 Langmuir isotherm for Hg(II) adsorption from pH adjusted solutions . . . 222 Figure 5.31 Freundlich isotherms for Hg(II) adsorption from pH adjusted solutions . . 222 Figure 5.32 Variation of solution pH for the mixed metal solutions after adsorption . . 226 Figure 5.33 Initial concentrations in mixed metal solution 226 Figure 5.34 Cd(II) and Hg(II) adsorption isotherms for the mixed metal solutions. . . 227 Figure 5.35 Langmuir isotherms for Cd(II) and Hg(II) adsorption by Type C 228 Figure 5.36 Langmuir isotherms for Cd(II) and Hg(II) adsorption by Norit SA4 . . . . 228 Figure 5.37 Langmuir isotherms for Cd(II) and Hg(II) adsorption by run 045 228 Figure 5.38 Freundlich isotherms for simultaneous Cd(II) and Hg(II) adsorption . . . . 229 Figure 5.39 Spike recoveries for Lagoon samples 233 Figure 5.40 Spike recoveries from Well samples 233 Figure 5.41 Concentrations of metals and TOC present in the leachate sample before
and after contact with the carbons 234 Figure 5.42 Percentage reductions in the concentrations of metals and TOC in the
leachate sample after adsorption by the carbons 234 Figure 5.43 Metals leached ft-om Type C and Norit after the 3 hour leach test 239 Figure 5.44 Metals leached from the raw soil samples and the soil-carbons after the
3 hour leach test 240 Figure 5.45 Metals leached fi-om Type C and Norit after the DIN leach test 241 Figure 5.46 Metals leached from the raw soil samples and the soil-carbons after the
DIN leach test 241 Figure AIV.l Freundlich isotherms for phenol adsorption by STAOl carbons 301 Figure AIV.2 Langmuir isotherms for phenol adsorption by STAOl carbons 301 Figure AIV.3 Freundlich isotherms for phenol adsorption by STA02 carbons 302 Figure AIV.4 Langmuir isotherms for phenol adsorption by STA02 carbons 302 Figure AIV.5 Freundlich isotherms for phenol adsorption by STA03 carbons 302 Figure AIV.6 Langmuir isotherms for phenol adsorption by STA03 carbons 303 Figure AIV.7 Freundlich isotherms for phenol adsorption by STA04 carbons 303 Figure AIV.8 Langmuir isotherms for phenol adsorption by STA04 carbons 303 Figure AIV.9 Freundlich isotherms for phenol adsorption by STA05 carbons 304 Figure AIV. 10 Langmuir isotherms for phenol adsorption by STA05 carbons 304 Figure AIV. 11 Freundlich isotherms for phenol adsorption by G1 carbons 304 Figure AIV. 12 Langmuir isotherms for phenol adsorption by G1 carbons 305 Figure AIV.13 Freundlich isotherms for phenol adsorption by the commercial carbons . . 305 Figure AIV. 14 Langmuir isotherms for phenol adsorption by the commercial carbons . . 305 Figure AV.l Freundlich isotherms for 4-nitrophenol adsorption by STAOl carbons . . . 306 Figure AV.2 Langmuir isotherms for 4-nitrophenol adsorption by STAOl carbons Figure AV.3 Freundlich isotherms for 4-nitrophenol adsorption by STA02 carbons Figure AV.4 Langmuir isotherms for 4-nitrophenol adsorption by STA02 carbons Figure AV.5 Freundlich isotherms for 4-nitrophenol adsorption by STA03 carbons Figure AV.6 Langmuir isotherms for 4-nitrophenol adsorption by STA03 carbons
16
306 307 307 307 308
Figure AV.7 Freundlich isotherms for 4-nitrophenol adsorption by STA04 carbons . . . 308 Figure AV.8 Langmuir isotherms for 4-nitrophenol adsorption by STA04 carbons . . . 308 Figure AV.9 Freundlich isotherms for 4-nitrophenol adsorption by STA05 carbons . . . 309 Figure AV. 10 Langmuir isotherms for 4-nitrophenol adsorption by STA05 carbons . . . 309 Figure AV. 11 Freundlich isotherms for 4-nitrophenol adsorption by G1 carbons 309 Figure AV. 12 Langmuir isotherms for 4-nitrophenol adsorption by G1 carbons 310 Figure AV. 13 Freundlich isotherms for 4-nitrophenol adsorption by the commercial
carbons 310 Figure AV. 14 Langmuir isotherms for 4-nitrophenol adsorption by the commercial
carbons 310
LIST OF PLATES
Plate 3.1 SoilSTAOl 90 Plate 3.2 Soil STA02 90 Plate 3.3 Soil STA03 91 Plate 3.4 Soil STA04 91 Plate 3.5 Soil STA05 92 Plate 3.6 Soil G1 92
17
GLOSSARY OF ABBREVIATIONS AND UNITS OF MEASURMENT
AAS ADR/RID
APHA BATNEEC BCR BET BGAM BPEO BPM BSI BRE CCC CEC CERCLA
CFR CCME CCMS CEGB COPA DoE DTA DTG DTI DRE EEC EPA ESR FRO GLC GC HMIP HoCEC HSWA ICP-AES ICRCL
IDL lEHO IPC LDR L r NATO NCP NPL NRA NSCA OTS PAHs PCBs
Atomic Absorption Spectroscopy Accord europ6en relatif au transport international des marchandises Dangereuses per Route; R^gelement International de marchandises Dangereuse par chemin-de-fer American Public Health Association Best Available Techniques Not Entailing Excessive Cost Community Bureau of Reference Brunauer, Emmett and Teller Equation British Gas Analytical Methods Best Practicable Environmental Option Best Practicable Means British Standards Institution Building Research Establishment, Cheshire County Council Cation Exchange Capacity Comprehensive Environmental Response Compensation and Liability Act ('Superfund') 1980 Code of Federal Regulations Canadian Council of Ministers of the Environment Committee on the Challenges of Modern Society Central Electricity Generating Board Control Of Pollution Act 1974 Department of the Environment Differential Thermal Analysis Derivative ThermoGravimetry Department of Trade and Industry Destruction Removal Efficiency European Economic Community Unit^ Kingdom Environmental Protection Act 1990 Electron Spin Resonance Federal Republic of Germany Greater London Council Gas Chromatography Her Majesties Inspectorate of Pollution House of Commons Environment Committee Hazardous and Solid Waste Amendments Act 1984 Inductively Coupled Plasma - Atomic Emission Spectrometry Interdepartmental Committee on the Redevelopment of Contaminated Land Instrument Detection Limits Institute of Environmental Health Officers Integrated Pollution Control Land Disposal Regulations Low Temperature Thermal Treatment North Atlantic Treaty Organisation National Contingency Plan National Priorities List National Rivers Authority National Society for Clean Air and Environmental Protection Overseas Trade Services PolyAromatic Hydrocarbons PolyChlorinated Biphenyls
18
PCDDs PCDFs PCP pHf PH; p H ^ POHC RCRA RCRAA RCEP RTP SARA SCI SITE TCLP TEQ TG TOC TSCA UK VOC(s) UNEP USA USEPA WHO WRAP XRD XRF
PolyChlorinated Dibenzo-para-Dioxins PolyChlorinated DibenzoFurans Penta-ChloroPhenol Solution pH after adsorption Solution pH before adsorption pH at which zero point charge occurs Principal Organic Hazardous Constituents Resource Conservation and Recovery Act 1976 Resource Conservation and Recovery Act Amendments 1984 Royal Commission on Environment^ Pollution Room Temperature and Pressure Superfund Amendments and Reauthorisation Act 1986 Society of Chemical Industry Superfond Innovative Technology Evaluation Toxicity Characteristic Leaching Procedure Toxic EQuivalent Dynamic ThermoGravimetry Total Organic Carbon Toxic Substances Control Act 1976 United Kingdom Volatile HydroCarbon(s) United Nations Environment Protection Agency United States of America United States Environmental Protection Agency World Health Organisation Waste Reduction Always Pays X-Ray Diffraction X Ray Fluorescence
UNITS OF MEASUREMENT
°C k
MS mg g
kg t m nm A pm ml 1 mol mM meq Nm^
ppb ppm keV
degrees Centigrade Kelvin microgramme (l*10'®g) milligramme (l*10"^g) gramme kilogramme (l*10^g) tonnes (l*10^kg) metre nanometre (1*10 'm) Angstrom (P lO ' m) picometre (1*10 '%]) milli-litre(l*10-'l) litre Mole = 6.09 xlCP atoms, milli Mole (l*10'^mol) milli equivalent A cubic metre of gas measured under standard (normal) conditions. These conditons are: 273 k (0 °C);101.3 kPa (760 mmHg); 11 % O2 or 9 % CO2; dry gas. parts per billion (jig/kg) parts per million (mg/kg) kiloelectron volt
19
CHAPTER 1
INTRODUCTION AND LITERATURE REVIEW
1.1 Introduction
Man has had a profound effect upon his surrounding environment. Pollution of the air, land or
water has invariably followed our progress from the stone age up to the modern day. Pollution
of the air or water has not, historically, been regarded as a problem, due to the 'dilute and
disperse' characteristics exhibited by these media. Contamination of land, however, is a far more
serious problem, being the most durable and requiring geological time scales to remove or
ameliorate the contamination by natural processes. Early industrial activities, for example by the
mining of copper and its smelting during the bronze-age (4500 - 5000 years ago), left a legacy
of arsenic contamination in the areas surrounding the early mines and smelters, this has however
proven to be of great value to historians and archaeologists. More recent incidents of pollution,
however, have not been treated with the same historical interest. There have been several major
pollution incidents which have resulted directly in action to curtail the cause of the problem, to
investigate and correct any other situations which potentially may cause similar incidents and to
ensure that future industrial practice does not cause comparable pollution again. Some of the most
notorious examples include: Lekkerkerk (USEPA,1992) Love Canal (RCEP, 1985), Jinzu River
and Minimata Bay (Beecher and Rappaport, 1990).
The development of the industrial society which the developed world lives by has meant that
discarded wastes have become a common occurrence. The harm done to the environment by these
wastes has grown with the complexity of manufacturing and industrial activities and the volume
of the inevitable unwanted by products of these efforts. The routine release of wastes into the
ecosystem was practised until it became obvious that the environment could not cope with this
additional burden. Unfortunately, the actions required to prevent this situation worsening, in the
form of national legislation, were not taken for 200 years after the start of the industrial
revolution, and in the intervening period the environment affected deteriorated and decayed. The
most telling damage can be attributed to the rise of the synthetic organic chemical industry during
the past 40 years, which produced complex, recalcitrant products and wastes. For example,
production of alkenes and aromatics between 1950 and the early 1970s, which constitute the six
primary petrochemicals, increased by 3500 % to almost 70 million tonnes/year (Bentley, 1987),
with a similar increase in the associated wastes.
With the realisation that waste production, disposal and treatment must be controlled, for the
20
benefit of society, research into waste management, minimisation and treatment has become a
subject of great importance. Not only is it important to clean-up the refuse from past industrial
and domestic activities, but in addition, current processes require modification to reduce their
waste products or eliminate them completely from unavoidable waste streams.
One of the most important tools for waste treatment is activated carbon. The origins of the
substance date back to the Egyptian civilisation (1500 BC), when charred barrels were recognised
for their ability to keep water and it was realised that charcoal possessed antiseptic properties
(Hassler, 1963). The unique adsorptive properties of this type of carbon gained the material its
name activated carbon. The extensive adsorptive ability and complex porous structure of this form
of carbon has meant that it has found many modern applications, for example in medicinal
products, gas masks and ore separation processes.
The increased stringency of environmental legislation has forced industry to purify its effluent
streams (aqueous and gaseous) in order to minimise the threat which they pose to the public and
the environment. Activated carbon has fulfilled this purpose. However, supplies of activated
carbon are served mostly from non-renewable feedstock such as coal and peat, whose supplies
are finite. The concerns and longer-term problems regarding the use of these materials has caused
the carbon manufacturers to investigate alternative feedstock such as vegetable by-products and
carbonaceous waste materials.
The work presented within this thesis addresses the problem of contaminated land, by showing
that it is possible to treat this material, by utilising activated carbon technology, to convert the
previously toxic soil into a benign product which can be re-utilised as an activated carbon for the
treatment of aqueous industrial effluents.
1.2 Waste definitions, arisings and waste legislation
1.2.1 Defining waste
The threat presented to human health by many wastes has been the major force which has caused
the development of definitions of waste and pollution. This is an area which has been subject to
much discussion due to the legal necessity for accurate definitions. Pollution is defined, literally,
as 'destruction of the purity of or the sanctity of, to make foul or filthy' (Oxford, 1982).
However, may substances such as heavy metals, dioxins and PCBs are very harmful to health at
concentrations which are too low to be detected by sensory organs.
21
In the UK, the definitions used in the Control of Pollution Act 1974 (COPA) (and retained in the
Environmental Protection Act 1990 (EPA)) splits waste into categories:
Controlled waste: 'Household, industrial and commercial waste or any such waste'
(COPA, 1974).
Waste: 'Any substance which constitutes a scrap material or an effluent or other
unwanted surplus substance arising from the application of any process; and any
substance or article which requires to be disposed of as being broken, worn out,
contaminated or otherwise spoiled, but does not include a substance which is an
explosive within the meaning of the Explosives Act 1875' (COPA, 1974).
Special Wastes: 'Any controlled wastes which:
(a) consist of or contain any of the substances listed in Part 1 of Schedule 1 to SI 1980 No
1709 (which include acids, alkalis. As compounds, asbestos, Cd and compounds,
inorganic cyanides etc.) and by reason of the presence of such substance,
i is dangerous to life
(Waste is to be regarded as dangerous to life for the purposes of the regulations
if:
a a single dose of not more than 5 cm^ would be likely to cause death or
serious tissue damage if ingested by a child if 20 kilograms body weight,
or b exposure to it for 15 minutes or less would be likely to cause serious
damage to human tissue by inhalation, skin contact, or eye contact.)
ii has a flash point of 21 °C or less as determined by the methods and with the apparatus
laid down by the British Standards Institution in BS3900: Part A8: 1976 (EN53),
or
(b) is a medical product, as defined in section 130 of the Medicines Act 1968, which is
available only in accordance with a prescription given by an appropriate practitioner as
defined in section 58(1) of that Act' (DOE, 1981).
In the USA, hazardous wastes are controlled by the Resource Conservation and Recovery Act,
(RCRA) and are defined as:
'Solid waste or combinations of solid waste, which because of its quantity, concentration
or physical, chemical or infectious characteristics, may: Cause, or significantly contribute
to an increase in mortality or an increase in serious irreversible, or incapacitating
reversible, illness or pose a substantial present or potential hazard to human health or the
22
environment when improperly treated, stored, transported, or disposed of, or otherwise
managed' (Lee et al., 1990).
This definition is further split by the United States Environmental Protection Agency (USEPA)
into "waste which is hazardous due to its characteristics, and waste that is specifically listed as
hazardous'. Solid wastes are defined by RCRA to include 'garbage, refuse, sludge and other
discarded material, including solid, semisolid, liquid, or contained gaseous material'
(Davenport, 1992). Whether a waste is RCRA hazardous is dictated by the reactive, corrosive,
ignitable, or toxic characteristics of the substance. The first three parameters are simple to test
for: a violent reaction with water implies reactivity; a pH of less than 2 or greater than 12.5
indicates a corrosive substance and a flash point below 60 °C implies ignitability. Toxicity tests
require use of the Toxicity Characteristic Leaching Procedure (TCLP) (Davenport, 1992).
International organisations use their own definitions of wastes. The UNEPAVHO use a joint
definition: 'Waste is something which the owner no longer wants at a given place and time and
which has no current or perceived market value' (Huismans, 1990). Huismans goes on to describe
the Basle Convention on the Control of Transboundary Movements of Hazardous Wastes and their
Disposal, which defined wastes as: 'Substances or objects which are disposed of, or are intended
to be disposed of, or are requested to be disposed of by the provisions of national law'. The
ADR/RID (Accord Europ^en Relatif au Transport International des Marchandises Dangereuses
per Route; Rfegelement International de Marchandises Dangereuse par Chemin-de-fer) agreements
on road and rail transport in Europe adopted the following waste definition: 'Wastes are
substances, solutions, mixtures, or articles for which no direct use is envisaged but which are
transported for reprocessing, dumping, elimination by incineration or other methods of disposal'
(Huismans, 1990).
The EC 'framework' Waste Directive (EC, 1975) defines waste as '... any substance or object
which the holder disposes of or is required to dispose of pursuant to the provisions of national
law in force'.
In conclusion, material are considered wastes where they present a threat to health. For the
purpose of this work the UK definitions and meanings shall be used throughout.
1.2.2 Waste arisings
In the USA, Municipal Solid Waste (MSW) is produced at the rate of 180 million tonnes per year
23
(1990 figures), and this is forecast to increase 20% by 2000 (Kreith, 1992). Estimates of EC solid
waste arisings (domestic and industrial) are of the order of 2.2 billion tonnes. The disposal routes
used for this waste are detailed in table 1.1.
Table 1.1 Disposal methods for solid waste in the EC and the USA expressed as a percentage.
Country Landfilled Incinerated Recycled Composted
Austria 64 20 - 16
Denmark 31 50 18 1.0
France 47.9 41.9 0.6 8.7
FRG 74 24 - 2
Italy 83.2 13.9 0.6 2.3
Netherlands 51 34 15 -
Sweden 35 60 5 -
Switzerland 20 80 - -
United Kingdom 88 11 1.0 -
EEC Average 54.9 37.2 4.5 3.3
United States 72.7 14.2 13.1 (HOCEC, 1991) and (Kreith, 1992)
Annual special waste arisings in Europe (EC) are estimated to be about 31 million tonnes
(O'Sullivan, 1989). In the USA, RCRA hazardous waste arisings are estimated to be 295 million
tonnes, of which only 14 million tons are solid wastes (Hanson 1989). The apparent large
discrepancy between special/RCRA wastes produced by the Europeans and USA can, in part, be
attributed to the different regulations and waste definitions operated in each country.
In the UK, total waste arisings (1991 figures) are estimated to be 516 million tonnes/year. This
comprises 140 million tonnes of controlled wastes and 2.5 million tonnes special wastes, the
remainder is composed of agricultural wastes (250 million tonnes) and mining/quarrying wastes
(108 million tonnes) (NSCA, 1993).
1.2.3 A European perspective to environmental legislation
The major industrialised countries of the world have, during the last 25 years, developed a series
of laws and regulations aimed at controlling the pollution which has resulted from their industrial
activity. Most of the laws are specific, that is, they deal with gaseous, aqueous or solid
24
pollutants/wastes. The implementation of this type of legislation has often been as a direct result
of a particular incident which was linked to waste disposal, particularly by manufacturing
industries.
1.2.3.1 United Kingdom environmental legislation
The UK has a long history of legislation pertaining to control of emissions. The early laws did
not tended to be aimed at the protection of the environment, but more towards protecting the
population, for example; in 1273 the use of coal was banned in London as it was prejudicial to
health. The UK has treated air and water as separate entities for control, recognising the ability
of each medium to cope with pollution. Land has tended to be considered with water, and at
present there is no UK legislation pertaining only to the prevention of pollution to land or to
enforce the clean-up of contaminated land (McKenna & Co, 1989), this is in contrast to both the
USA and the Netherlands which will be briefly considered. However, the Environment Act 1995
will address some of these omissions. Table 1.2 gives details of relevant UK legislation since
1500. The scientific understanding of pollutant and waste behaviour in the environment, which
has shown that exchange of wastes between air, land and water occurs, has affected the way in
which Parliament formulates legislation, with the older legislation becoming consolidated into
more recent single acts.
The Environmental Protection Act (EPA, 1990) is the current backbone of UK legislation for
environmental management. It has introduced several concepts to the regulation of industrial
processes: Integrated Pollution Control (IPC), Best Practicable Environmental Option (BPEO) and
Best Available Techniques Not Entailing Excessive Cost (BATNEEC). (NB:The British
BATNEEC (Technique) is different to the European BATNEEC (Technology) (NSCA, 1993).)
IPC regulates all major emissions of solids, liquids or gases to land, air or water by part A
processes. Part A processes are distinguished from part B processes because they were regulated
under the Health and Safety at Work Act 1972, give rise to 'significant quantities of special
waste' and give rise to emissions (to sewers or controlled waters) of substances with noxious
effects. This is implemented and regulated by Her Majesties Inspectorate of Pollution (HMIP) in
England and Wales. For Scotland this responsibility is split between HM Industrial Pollution
Inspectorate and River Purification Authorities. Part B processes are controlled by Local
Authorities and are concerned with air pollution control only.
BPEO was defined as: "... for a given set of objectives, the option that provides the most benefit
or least damage to the environment as a whole, at acceptable cost, in the long term as well as the
25
short term' (Harrison, 1990). It is applied as a consequence of compliance with IPC, so that the
pollution control equipment to give the best protection to the environment (through all pathways)
is always applied to IPC processes. This is, however, offset by the implementation of BATNEEC.
BPEO was proposed by the Royal Commission on Environmental Pollution (RCEP) in 1976, as
an extension of the 'Best Practicable Means' concept (BPM) used since the Alkali Act (1874).
The principle has been thoroughly discussed in the RCEP 12th report on BPEO (RCEP, 1988).
Table 1.2 Important acts of Parliament controlling pollution of the environment in the UK
Water Protection Laws General Air Protection Waste Laws Laws
Environment Act 1995
Water Resources Act 1991 (Eng. & Wales)
Water Act 1989
Environmental Protection Act 1990 and Controlled Waste Regulations
1992
Control Of Pollution Act 1974 and Control Of Pollution (Special Waste) Regulations 1980
Water Resources Act 1963 Deposit of Poisonous
Wastes Act 1972
Health and Safety at Work etc Act 1974
Rivers (Prevention of Pollution) Act 1951 and 1961
Clean Rivers (Estuaries and Tidal Waters) Act 1960
Rivers Pollution Prevention Act 1876
Public Health Acts 1848, 1875, 1936, 1961
The Causes Acts 1847 -A series of acts providing a framework for
industrial and domestic water pollution
Bill of Sewers 1531
(Compiled with the aid of information from the NSCA
Alkali etc Works Order 1966
Clean Air Act 1956, 1968
Public Health (Smoke Abatement)
Act 1926
Alkali etc Works Regulation Act 1906
Alkali etc Works Act 1863, 1874
Pollution Handbook (NSCA, 1993).
Much of the most recent UK legislation for the environment has its basis in EC directives. As the
European Community moves towards the ultimate aim of a unitary legislative and administrative
'state', this trend will only increase.
1.2.3.2 European Community (EC) environmental legislation
The formation of the EC by the signing of the Treaty of Rome provided a basis for the
26
formulation and implementation of common environmental laws between member states through
use of the following articles; 100 - applicable when a directive was required to harmonise national
laws which would otherwise perturbed the functioning of the common market and 235 - for
'measures of a more general nature'. As the EC has matured the breadth of these measures have
reflected the changing needs. EC legislation is implemented in two manners: by regulations or
directives. The environmental legislation has been traditionally adopted as a directive. The Single
European Act (1987) adapted the Treaty of Rome to accommodate modern environmental
concerns by introducing articles which were specific to environmental concerns: Articles 130R,
130S and 130T, (Clifford Chance, 1992).
Article 130R - The EC environment policy is set out in this article, with the objectives of
'preserving, protecting and improving the quality of the environment, contributing towards
protecting human health and ensuring rational utilisation of natural resources', under the
principles that 'preventative action should be taken, that environmental damage should as a
priority be rectified at source and that the polluter should pay' (Clifford Chance, 1992).
Articles 130S and 130T - 130S allows the EC council to determine environmental policy, whilst
130T maintains the rights of member states to introduce their own, more stringent, treaty
compatible legislation (Clifford Chance, 1992).
1.2.3.3 EC environmental directives and action programmes
The waste directive (EC, 1975) set the scene for environmental regulation, describing a series of
guidelines which member states should utilise for waste control. It was substantially amended in
1991 making it applicable to a wide range of objects/substances described in an annex, which is
termed the waste list (Clifford Chance, 1992). The waste directive has been followed by many
other directives: dangerous substances in the aquatic environment (EC, 1976), toxic and
dangerous wastes (EC, 1978), groundwater (EC, 1979), drinking water (EC, 1980a), cadmium
discharges (EC, 1983) and hazardous waste (EC, 1991) (which replaced the toxic and dangerous
waste directive on 31/12/1993). The dangerous substances in the environment directive (EC,
1976) details the list I (Black List) and list II (Grey List) substances, which were " . . . selected
mainly on the basis of their toxicity, persistence and bioaccumulation,... and ... substances which
have a deleterious effect on the aquatic environment', respectively. It is important to note that the
ultimate aim of the EC is to issue directives controlling the discharge of each and every list I
substance, as has already been done for Cadmium (EC, 1983) and Mercury (EC, 1982; EC,
1984). Additionally, once toxicity parameters for each of the list II substances have been
27
established, they will be promoted to list I (Haigh, 1990). List I and II substances are detailed in
table 1.3.
Table 1.3 List I and list II substances, as defined by EC directive (76/464/EC)
List I substances List II substances
organohalogen compounds and precursors The following elements and compounds: Zn, Cu, Ni, Cr, Pb, Se, As, Sb, Mo, Ti, Sn,
organophosphorus compounds Be, B, U, V, Co, TI. Te, Ag
organotin compounds
substances carcinogenic in or via water Biocides not on List I
Hg and compounds Substances affecting taste and/or smell on _ . , aquatic derived products for human
an compoun s consumption and their precursors
Persistent mineral oils and petroleum hydrocarbons
Synthetic substances which may interfere Silicon organic compounds of a toxic or with water use, by floating, sinking or persistent nature remaining in suspension
(EC, 1976)
The annexes to the hazardous waste directive describe the manner in which wastes can be
assigned as hazardous and the relevant properties which are exhibited by them (EC, 1991).
Of particular significance to this work is the groundwater directive (EC, 1980a). Article 1 para. 1
of the directive states that: 'The purpose of this directive is to prevent the pollution of
groundwater by substances belonging to the families and groups of substances in lists I or II in
the Annex, hereinafter referred to as "substances in lists I or 11", and as far as possible to check
or eliminate the consequences of pollution which has already occurred' (EC, 1980a). This concern
regarding groundwater pollution originates from the fact that 70 % of mainland Europes' drinking
water is groundwater derived. The list I and II substances discussed in the groundwater directive
are slightly different to those shown in table 1.3 from the dangerous substances in the aquatic
environment directive for waste (EC, 1976). This first paragraph from the groundwater directive
can be used, directly, as a reason for enforcing the restoration of a contaminated site. There is
evidence that this power is being used by the NRA (Anon, 1991a).
The EC has also adopted a series of "action programmes on the environment" each of which
constituted a set of goals for protecting and improving the environment. These programmes (the
fifth was adopted in 1992) are shown in table 1.4 (Clifford Chance, 1992).
28
Table 1.4 EC action programmes on the environment.
Number Starting Purpose
One
Two
Three
Four
Five
22-11-1973
17-06-1977
07-02-1983
19-10-1987
Improvement of the environment by reducing pollution and nuisance and the structuring of the objectives and principles of
the community environment policy
The extension of the first programme by protecting and managing: land, the natural environment and resources and
improving the environment
Development of an overall strategy. Describe measures to be taken to prevent and reduce pollution and nuisance in the different environments, as well as protection and rational
management of land, the environment and natural resources.
To identify pollutants; to determine best focus for control measures; to set objectives to combat acid deposition and forest die-back; to reduce ambient pollutant concentrations; to develop
appropriate management techniques.
(1992) The implementation of sustainable growth. Encouragement of energy efficiency. Addressing the challenges of climate change, air pollution, depletion of natural resources, deterioration of the
urban environment and waste management. Ensuring that further generations will inherit a environment capable of
sustaining them. (Adapted from: NSCA, 1993 and EC, 1992)
1.2.4 The nett effect of waste legislation
The aim of regulatory devices is to prevent pollution, restrict or halt the discharge of wastes and
repair, as far as possible, the damage which has been inflicted upon the environment to date. In
the USA, the HSWA stated that"... wherever feasible, the generation of hazardous waste is to
be reduced or eliminated ... Waste that is nevertheless generated should be treated, stored, or
disposed of so as to minimise the present and future threat to human health and the environment.'
(Hirshhorn & Oldenburg, 1989).
The USEPA's preferred waste management strategy is: 1) waste reduction, 2) waste separation
and concentration, 3) waste exchange, 4) energy / material recovery, 5) Incineration / treatment
and 6) Secure land disposal (Freeman, 1988). The UK RCEP 17th Report (1993) produced a
similar priority for waste management options:'... 1st, wherever possible avoid creating wastes,
2nd, where wastes are unavoidable, recycle them if possible, 3rd, where wastes cannot be
recycled in the form of materials, recover energy from them and 4th, when the foregoing options
have been exhausted, utilise the best practicable environmental option to disposed of wastes'.
The nett result of the environmental legislation over the last 25 years has been to make industry
29
and the public realise that waste costs money. Many industrial processes are based upon
production routes which were developed 40 (or more) years ago. This fact, coupled with the
highly competitive world of commerce in which most industry operates has seen renewed efforts
in the areas of process development (Laing, 1992). Waste minimisation is an important aspect of
modem industrial management, which was successfully implemented by Dow Chemicals, who
lead the way in this field with their WRAP programme; Waste Reduction Always Pays (Gillis,
1994).
1.3 Contaminated soil
1.3.1 Introduction
Soil consists of 50-60 % by weight mineral matter, derived from the underlying rock and of
varying inorganic compounds, 25-35 % water, 15-25 % 7%, soil gases and 5-10 %
organic matter. It is also a complex ecosystem and the shape and structure of soil is extremely
variable (Skeffington, 1987). Soils are very heterogeneous, which means that added chemicals will
interact differently with each soil type making the contamination very difficult to handle and treat
successfully. Thus, land is the most vulnerable medium to pollution and soil contamination has
become a major issue of the industrialised world. The approach that is taken towards soil
contamination by different countries is reflected partly by the different geological conditions in
each country and by their varied experience with soil pollution incidents.
1.3.2 Defining contaminated soil or land
In the UK, the Department of the Environment (DoE) states that contaminated land cannot be
defined unambiguously, and has separate definitions for contaminated land: "... land which
represents an actual or potential hazard to health or the environment as a result of current or
previous use', and derelict land: "... land so damaged by industrial or other development that
it is incapable of beneficial use without treatment'. The House of Commons Environment
Committee (HoCEC) has accused the DoE of treating contaminated land 'as a sub-category of
derelict land' (HoCEC, 1990). The Environmental Protection Act 1990 (EPA), section 143, gave
a definition which was for 'land subject to contamination' meaning 'land which is being or has
been put to a contaminative use', where contaminative use means 'any use of land which may
cause it to be contaminated with noxious substances' (EPA, 1990). However, in the Environment
Act (1995) (Part IIA section 78A), there are definitions of contaminated land included:
30
'"Contaminated Land' is any land which appears to the local authority in whose area it is situated
to be in such a condition, by reason of substances in, on or under the land, that
(a) harm, or
(b) Pollution of controlled water, is being or is likely to be, caused.
"'Harm' means harm to the health of living organisms or other interference with the ecological
system of which they form part and in the case of man, includes harm to his property"
(Environment Act, 1995).
Other definitions which have been proposed include: 'Contaminated land can be defined as any
land which is shown to contain sufficient quantities or concentrations of a substance or substances
such as to pose a direct hazard to a specified target or targets' (lEHO, 1989). 'Land that contains
any substance that when present in sufficient concentrations or amounts, presents a hazard. The
hazard may: a) be associated with the present status of the land, b) limit the future use of the
land, c) require the land to be specifically treated before particular use' (BSI, 1988). 'Land that
contains substances that, when present in sufficient quantities or concentrations, are likely to cause
harm, directly or indirectly, to man, to the environment, or on occasions to other targets' (Smith,
1986).
The USEPA does not have any statutory definitions for contaminated soil, but it classifies soil
according to the components causing the pollution to the soil, which are governed by RCRA (as
previously discussed in section 1.2.1). Thus contaminated soil means 'soil which contains RCRA
hazardous waste(s)...' (listed in the Code of Federal Regulations (CFR) Title 40 CFR Part 261,
Subpart D), ' . . . or soil which otherwise exhibits one or more characteristics of a hazardous waste
as defined in 40 CFR Part 261 Subpart C (Davis and Chou, 1992). The characteristics of
materials which make them wastes under US law were also discussed in section 1.2.1.
1.3.3 The causes and hazards of contaminated land
Table 1.5 details the types of sites which can be expected to contain foreign matter as a result of
the activities of the industry associated with them. The threat that contaminated land poses can
be split into two broad categories: physical and toxic hazards. The toxic hazards tend to be
chemical in nature. Chemical pollution to soil is normally encountered as two distinct forms:
organic and inorganic. The greatest concern is expressed about organic materials, most of which
fulfil the criteria describing special waste or RCRA wastes (section 1.2.1). This does not imply
that inorganic wastes are less dangerous, but they do tend to be less frequently encountered, and
31
when they do occur, they are invariably associated with organic compounds.
Table 1.5 A list of sites which will be potentially contaminated and the type of contamination which may be encountered.
SITE TYPE MET ORG. ASB. ACID ALK. FIRE BIO. OTHER
ABATTOIRS • • ANTHRAX
AIRPORTS / • FUEL SPILLS
CANALS • • METHANE GAS
CHEMICAL -BULK PHARMACEUTICAL PETROCHEMICAL
• •
• •
/ •
/
/
• / •
COAL MINING • •
DENTAL FACTORIES • MERCURY
ELECTRONICS / •
GAS WORKS • • • • • • CYANIDE
GRAVEL PFTS •
LANDFILL SITES • • • / • • • LEACHATE
MILITARY LAND
MUNITIONS FACTORIES
/ •
• •
• •
EXPLOSION HAZARD
NUCLEAR ESTABLISHMENTS
/ • • RADIOACTTVTTY
OIL REFINERIES • • •
ORE EXTRACTION TIN, COPPER, LEAD •
LEACHATE
PAINT FACTORIES • •
PAPER/PRINTING WORKS
/ • • •
PLATING WORKS / • • •
PORTS • • • /
POWER STATIONS • • • •
RAILWAYS • • •
ROADS • •
SCRAP & BREAKERS YARDS
• • • • • • RADIOACTIVrrY
SEWAGE WORKS • • •
SHIP YARDS • • • •
SMELTERS AND STEEL WORKS
• • •
TANNERIES / / •
TEXTILES • • / /
For explanations of abbreviations, please refer to the accompanying notes on This table was compiled from (BSI, 1988), (Corbett, 1982), (ICRCL,1987, (Crowhurst and Beever, 1987) and (Parker, 1984),
the following page. 1986, 1983, 1990),
32
Notes
Org.= Organic material, carbon containing material i.e. tars, oils, solvents etc.
Met= Metal contamination; refers mainly to metals and metalloids toxic to man, animals and
plants such as: Pb, Hg, Cu, Cd, Cr, As, B, Ni, Sn, Zn.
Asb.= Asbestos pollution. These materials was a general insulation material, and would be
expected to be present on any site where heat insulation was required.
Alk.= Alkaline material present. For example, ammonia, hydroxides, group 1 and 2 oxides.
Bio.= Biological Hazards; mainly pathogenic organisms and diseases.
Fire = Fire hazards can be from gas (methane), fuel or solvent spills, coal heaps and
underground fires.
Acid= Acidic material present such as: mineral or organic acids, metallic chlorides.
Table 1.5 is not comprehensive, it is merely an illustration of the types of sites and contamination
associated with them. Only the main contaminants associated with each particular industry have
been indicated. It is very important to be aware that any form of contamination can be present
on any site, for example, asbestos is often found on sites where heat or power was required for
the process.
The physical hazards presented by a contaminated site include: explosion and fire, subsidence,
corrosion of structures, effects on mechanical properties of soil. Toxic hazards include:
inhalation of dust and toxic or asphyxiant gases: ingestion of contaminants from fingers
(especially by children) or food contamination: direct ingestion through uptake of contaminants
by edible plants or contamination of water: direct contact with absorption through skin,
retardation or death of plants etc (Smith, 1986). Figure 1.1 summarises the various pathways in
which contamination present in soil can move through the environment and ultimately affect
plants, animals, humans and buildings.
Complicating features of contaminated land are the inherent heterogeneous nature of the
contamination, and the uncertainty about the effect that contamination has upon man and the
surrounding environment. In table 1.5 the types of contamination have been split into general
groups (metals, acids, asbestos etc.). Very often the combinations of wastes make the soil
dangerous (for example: sulphur and acid produces H^S; waste heaps from coal mines are not a
threat because of their contents, but because they can ignite and their inherent instability).
The governments of most industrialised countries have tried to establish guideline values for
background and contaminating concentrations of components in soil. The obstacle to doing this
33
has always been how you can define a 'safe' concentration. For example; As in soil, which is
recommended to not exceed 40 ppm for gardens in the UK, has been found in Cornwall (which
has a history of Sn mining and smelting) at concentrations of up to 900 ppm (Pearce, 1993).
Some of this excess As can be attributed to the Sn industry, but some will be a result of the
natural geology of the area. The cause of the contamination to a particular site is an important
aspect of the assessment of the site. By understanding how the site was polluted, the correct action
can be taken to address this problem, and the prevention of similar incidents may be the result.
Figure 1.1 Exposure pathways for animals, plants, humans and buildings to contaminants present in soils
SOURCE (of contaminant)
Means of release
TO AIR 1
Evaporation 1 Dust
Fumes |
— • TO WATER 1
Leaching | Run off Spillage |
TO SOIL
Leaching Leakage Spillage
inhalation
WATER
ingestion
ingestion m
uptake PLANTS
ingestion skin contact
V w w
ingestion skin contact
ANIMALS ingestion
inhalation ingestion
K ^ HUMANS ^ BUILDINGS
(Figure 1, Laidler et al., 1995)
Contamination of sites may have been due to: disposal of process wastes on site, underground
storage facilities which leaked, accidental spillage during plant operation, reclaiming land using
industrial wastes, inadequate plant shut down prior to demolition, careless demolition or licensed
emissions to air or water.
The exact extent of contaminated land in the UK is not known, and most estimates are based upon
a pilot survey of contaminated land in Cheshire in the mid 1980s, conducted by Cheshire County
Council (CCC) and the Department of the Environment (DoE). They set out to identify those
sites which, irrespective of their current status and use, could contain substances likely to give
34
rise to hazards which might affect the future use of the land' (CCC, 1986). From the findings of
this survey, the number of contaminated sites in the UK was estimated to be 50 000 to 100 000
(HoCEC, 1990). The cost of cleaning up this contamination has been estimated to be between
£10 - 30 billion (Anon, 1991a). The most commonly encountered forms of contaminated sites in
the UK have been estimated by the DoE, by utilising data on the number and type of requests for
assistance made to the ICRCL over a ten year period, and are shown in table 1.6 (Anon, 1989).
This data shows that gasworks are the joint most common forms of contaminated land
encountered in the UK, and the significance of gaswork sites is now considered in detail.
Table 1.6 The proportion of contaminated sites in the UK
Contaminative uses
Gasworks 25.8
Waste tips 25.8
Metal industries 9.0
Sewage works/sludge tips 7.8
Chemical works 7.4
Docks and wharves 4.1
Tar, oil, petrol depots 3.3
Scrap yards 3.1
Munitions 1.2
Tanneries 1.0
Miscellaneous 11.5 (Anon, 1989)
1.3.4 Gaswork sites - their history and the nature of the contamination resulting from
their operation
Abraham Darby is credited with first producing gas from coal, in 1709, with the first commercial
plant opening in 1802 at Soho, Birmingham (Johnston et al., 1993). The sole use of gas produced
by the early plants was to provide light, with Pall Mall in London being the first street to be gas-
lit. The momentum for finding other uses for gas was with the advent of the electric light, but
the success of gas was assured (ERL, 1987).
The manufacture of gas is a very complex process. Figure 1.2 shows the layout of a typical gas
manufacturing process.
35
Figure 1.2 The unit processes involved in the manufacture of gas
^ y _ | L _
—dt'M (0*1
CONDiNStRS mo*T
by HYDRAULIC MAIN
by rtlorit] IXHMJSTU
(Drawt oH *#*!
AMMONIA SCBUBKEK
CONMNSW F«nh#r (oolmg
PUAIFIE&S Cat MiMd *v#, bog u6a or« Ktmi-ftd«r ol
fd
:B8(S%iL0
CM HOlOU BENZOLE HCOVMY CAS DKYER
NATION wm
le prsitu
Coal was fed into retorts (vertical or horizontal) and heated in an oxygen starved atmosphere. The
coal was carbonised, releasing gas and a complex mixture of by-products, which included: coal
tar (containing PAHs), phenolic aromatics, ammonia compounds, sulphur and compounds,
hydrogen cyanide and coke. The principle of the process did not change in the 150 years of the
gas-manufacturing industry's existence, although various modifications to the process, which
principally affected the volume and quality of the gas and by-products did occur. The crude gas
was purified in several stages: The raw gas entered the hydraulic main, which contained an
aqueous solution, where up to 85 % of the tar, ammonia and phenolics were removed. The gas
would then enter a condenser followed by an electrostatic precipitator to remove final tar traces.
A further wash would remove residual ammonia. Finally, the H^S in the gas was removed using
bog ore, a hydrated form of FezO; mixed with a peat-like organic material. This step also
removed the cyanides in the gas, by reaction of HCN with the iron oxide and sulphide present
in the purifier beds. The important reactions which occurred in the purifier beds were as follows:
1. HjS removal.
Gas purification: FegO .H^O + -^FezSs.HoO 4-
Regeneration: 2Fe2S3.H20 + 30; -> ZFe^Os.H^O 4- 6S
36
2. HCN removal.
FeO + FezO] +6HCN ^3Fe(CN)2 +
9Fe(CN)2 + 30 -^Fe^Oj + Fe,( FeCCN).),
3FeS + IFezSg + 18HCN - * Fe / FeCCN) ), + 9HjS
The coal tar produced was a complex mixture of aromatic hydrocarbons, which included
heteroatom-containing molecules such as quinoline, thiophene and so forth. Dragun details over
500 organic compounds which have been identified in coal tar (Dragun, 1988), although there are
an estimated 10 000 different compounds in coal tar (Zander and Collin, 1993). The structures
of selected PAHs, which are priority pollutants according to USEPA guidelines are shown in
figure 1.3.
Figure 1.3 Selected PAHs which can all be found in coal tar-
Flourene Naphthalene Acenapthlenc Accnapthyiene
O
Benzo[k]nounnthene Benzlajanthracene
Benzorklnouranthene
Benzo[b]fiouranthene
O
Beniolghi]peryiene In(leno[lA3^]pyrene Coronene
Virtually every town in the UK had its own gaswork, which was originally constructed on the
edge of the town, usually near a local water supply (surface or ground). There are an estimated
3000 +/- 1000 derelict gaswork sites in the UK (ERL, 1987), but, because of the major
urbanisation of the UK over the last 150 years, these sites were built around and are now
invariably located in central positions within towns and cities. An indication of the widespread
problem of gaswork sites was given by Thomas and Lester (1993a), who produced a review in
which they identified and subsequently visited 35 abandoned gaswork sites in Greater London.
37
They categorised them in respect of the security of the site, pollution by the site due to odours,
noise, soil discolouration, water contamination, visible waste and old process buildings, and
current uses of the land. Thirteen of these site were considered to exhibit evidence of pollution,
despite the fact that they had all closed at least 20 years previously.
During the operation of the site, waste disposal would occur within the site, much of the pipe-
work and the tar tanks would be arranged underground, and they invariably leaked polluting the
soil. The bog ore would be laid out on the ground to regenerate it, allowing leaching of cyanide,
sulphur and metals into the soil. Often, sites would be expanded, with any fill material coming
from the process wastes. When natural gas replaced coal gas in the early 1970s, all the gaswork
sites were closed and demolished. Rarely were tanks or pipes emptied, and uncontrolled
demolition resulted in wide-spread contamination of the site surface. The underground structures
deteriorated and caused wide-spread contamination. Rain water percolation over the site leached
much contamination into the lower soil levels, and water supplies were frequently subject to
contamination by phenols, PAHs and metallic species, especially Fe, Mn, Cr, Pb and Cu, which
were all present in bog ore. Asbestos contamination can also be a hazard, because of its use to
lag hot pipes and the wide-spread application of asbestos containing board in industrial building
cladding.
During their operation, gasworks were considered by neighbours to be a nuisance due to the
odours and wastes produced by the plant. This problem still exists, and abandoned gaswork sites
still are a major problem, their contamination is complex and harmful, and the occurrence of sites
is very widespread. Due to these factors, gaswork soil was chosen as the subject for study for this
work.
1.3.5 Policy and practice towards contaminated land
The approach of most countries to the problem of contaminated land generally reflects their
experiences of the effects of land pollution. The following discussion compares the differing
attitude of the EC, UK and other industrialised countries where contaminated land is concerned.
1.3.5.1 The EC
The EC does not actually have a contaminated land policy, however, there are indications of the
possible future actions of the EC in this area contained in the 5th action programme and the EC
green paper on 'Repair of damage caused to the environment' (COM(93)47). This discussion
38
paper clearly stated that contaminated soil was an issue to be addressed. The main discussion of
the paper was towards the establishment of a legal framework for allocating liability for
environmental damage, based upon the proposed directive on civil liability for damage caused by
waste, however, the problems which the EC has had in constructing an acceptable format for this
document and then applying it to contaminated land has been similar to the problems which the
UK and other industrialised countries have encountered with this issue (Butlin, 1994).
In his discussion of the EC stance on contaminated land, Butlin (1994) summarised that the EC
policy was not implicit in any one document, but by careful examination of several directives,
both current and proposed, as well as the EC green paper, policy documents such as the
Maastricht Treaty and the fifth environment programme a policy could be implied. Some of the
main points of this policy were: "(i) prevention of pollution at source must always be a priority
action, (ii) the polluter must pay for pollution, (iii) the use of EC natural resources must be on
a sustainable basis ... (v) contaminated land is included among the sources of factors causing
'damage to the environment' ... (xi) certain of the water quality objectives also represent part
of the EC's implicit policy on contaminated land by regulating or prohibiting the discharge of
certain dangerous substances from contaminated land to surface or groundwater" (Butlin, 1994).
1.3.5.2 The United Kingdom
The UK approach to contaminated land has not been dictated by the influence of public opinion,
mainly because the UK has been spared a major pollution incident of the kind suffered by other
countries such as the USA (Love Canal) and the Netherlands (Lekkerkerk).
Contaminated land control has always been a sub-set of more general laws. The Nuisance
Removal Act, 1848 (Bentley,1987), the Public Health Act, 1936 and the Public Health (Recurring
Nuisances) Act, 1969 empower local authorities to stop statutory nuisances, which can be applied
to land containing contamination detrimental to health. The Town and Country Planning Act,
1971 allows the regulation of development of sites which are known to be contaminated, even to
the point of refusal of planning permission. The Occupiers Liability Act, 1957 and the Health and
Safety at Work etc.. Act, 1974 can be applied to prevent contamination to land by ensuring
employers conduct their business so as to minimise the danger to their staff, visitors and the
public (RCEP, 1985; Bentley, 1987; lEHO, 1989). The Deposit of Poisonous Waste Act, 1972
(the fore-runner to the COPA Special Waste Regulations, 1980) was introduced as a result of
public concern over indiscriminate waste disposal. The act made it an offence to dispose of
material in a way 'liable to cause an environmental hazard' (Hillman, 1984). More recent
39
legislation includes the Derelict Land Act 1982, the Building Act 1984, the Environmental
Protection Act 1990 and the Water Resources Act 1991. Of especial interest are the
Environmental Protection Act 1990 (specifically section 143, the registers of contaminated land
uses) and the Water Resources Act 1991. Both acts contained several sections relating to the issue
of contaminated land, which are discussed in more detail below.
Environmental Protection Act 1990.
Section 27: The powers of the HMIP inspectors to remedy harm caused by operation of
scheduled processes without IPC authorisation.
Section 33/59 Concerned with waste management licences and the removal or remediation of
unlicensed controlled waste deposits.
Section 34 Duty of care, not known to have been used as yet against persons who have
caused land contamination by failing to follow DoC.
Section 73 Compensation for damage caused by a deposit of controlled waste.
Section 79 Local authorities have a duty to inspect their areas and investigate complaints
from residents with respect to statutory nuisance (inc. deposits prejudicial to
health or a nuisance).[Where 'nuisance' means "deleterious affection of land or
its use and enjoyment which arises outside the land and then proceeds to invade
it"].
Section 80 Serving of abatement notice once they are satisfied a statutory nuisance exists.
Section 81 Allows claiming of expenses arising from implementing sections 79/80.
Section 82 Action by individuals for statutory nuisance.
(sections 79-82 replaced section 92 of the Public health Act 1936).
Section 143 Introduced the idea of 'Public registers of land which may be contaminated',
however, this section soon became known as the 'Contaminated Land Register'. The principle
of the register was for local authorities to compile details of all sites which were considered to
be contaminated due to previous or current use, for which the government had originally intended
16 categories of contaminative usage, but which were then reduced to 8. The uses included the
manufacture of: gas, coke or bituminous material; asbestos or asbestos products; chemicals; or
use as a scrap metal store. Although the stimulus of the register was based upon advice given by
successive Parliamentary Committees (RCEP, 1985 and HoCEC, 1990) this piece of legislation was
never implemented, due to the lobby pressure applied to the Government by many interested
parties (developers, surveyors, land owners etc.). One of their major criticisms of the register was
that once a site was on the register, it would never be taken off, irrespective of any restoration
work that had been performed, consequently the fear of blight to assets was used, to great effect.
40
as an argument against the registers. The only positive aspect of this attempt to regulate
contaminated land was that the public and industry became fully aware of the threat posed by
contaminated land. Perhaps another major factor in the resistance to the section 143 register was
the current legal situation in the UK with respect to liability for contaminated sites. The axiom
caveat emptor applies in UK common law (McKenna & Co, 1989). Thus there were many
landowners who had purchased, unknowingly, land which was contaminated but who were now
responsible for the contamination to these sites, despite the fact that the pollution did not occur
during their ownership.
The cost of site restoration is already escalating (current rates are £100 000 to £300 000 per
hectare), and where government organisations such as the National Rivers Authority (NRA)
become involved, perhaps due to groundwater pollution from a particular site, costs can rise
exponentially. The UK government, unlike the USA or Dutch counterparts, does not directly fund
site decontamination/restoration nor become directly involved in remediating land, although
derelict land grants have been available since 1975 (RCEP, 1985). Hence, some land owners and
developers would have owned an 'asset' that was of negative value.
Water Resources Act 1991
Sections 85/86 These are the principle water offenses, it is a criminal offence for a person to;
"cause or knowingly permit any poisonous, noxious or polluting matter or any
solid waste matter to enter into controlled water". Section 86 are the legal powers
to prevent discharges occurring.
Section 161 Gave wide powers with respect to prevention/amelioration to reduce actual harm
or likely pollution of controlled waters.
As the preceding discussion indicates, legislation specifically controlling the contamination of land
in the UK is non-existent. Most UK powers focus on preventing harm and do not seek to
remediate existing contamination unless it is causing harm. When the government was forced to
reconsider the issue of Section 143 registers they then issued a consultation document "Paying for
our past" (DoE, 1994a), the results of which were incorporated into the 1995 Environment Act.
The main points which arose from the discussion document were published in "Framework for
contaminated land" (DoE, 1994b) and were as follows:
Proposed formation of the Environmental Protection Agency (EPAgency) to enhance
existing regulations to prevent land pollution (related to sustainable development)
Remediation is suitable for use (not multifunctionality)
41
Use of the existing process of development and redevelopment of land as the best means
for solving the contaminated land problem
Replace existing statutory nuisance powers with ones specifically for contaminated land
The EPAgency will become the centre for technical expertise for contaminated land and
development of policies and priorities for tackling contaminated land by consulting
widely.
The borough and district councils require to identify and act upon land contamination -
(including current statutory nuisance investigation powers under part III of EPA 1990,
but extended to new contaminated land nuisance.
The councils are to seek specific/additional advice from EP Agency where water pollution
may occur.
Local authorities will be able to require a person responsible for contaminated land to
undertake work to remedy the contamination and prevent, or where practicable, restore
any consequential damage.
(i) They are required, however, to notify appropriate person of work required on
site, non-compliance with the notice will be an offence. Alternatively, the local
authority can remedy the pollution its-self if the responsible party does not take
any action, and then claim the costs back from the other party, providing that it
does not cause hardship.
(ii) Special arrangements for monitoring any closed landfill sites which fall under the
definition of contaminated land, replacing the never implemented section 61 of
EPA 1990.
There shall not be protection for all the risks of contaminated land to lenders, nor the
borrowers who pollute land, but the "deep pocket" Financial institutions should not be
expected to be held responsible for remedial work costs, irrespective of responsibility.
Section 143 in EPA 1990 to be repealed, with the new regime a register was considered
to be unnecessary because under section 78B of the 1995 Environment Act, local
authorities shall have the requirement to periodically inspect their area to identify
contaminated land, closed landfill sites and closed landfill sites which appear to the
authority to be suitable for designation as special sites.
Section 78C Requires the local authorities to issue remediation statements with respect to
special sites, detailing the requirements for remediation.
Section 78D States that any contaminated land or special sites shall have a remediation notice
served upon the appropriate person detailing what shall be done to remediate the
site and within what time scale, (but only what is responsible - with respect to
42
cost and the seriousness of the harm caused by the site).
Remediation of contaminated land in the UK, historically, has been directed towards a solution
which fits the end use of the development. The government has not interfered in this area,
preferring to leave 'market forces' to guide land reclamation. The new proposals do not change
this regime. Ultimately, THE POLLUTER PAYS, which is also implicit in EC policy, will be
the cornerstone of the contaminated land regime, but caveat emptor is retained, so the land
owners are still responsible and owner occupiers can still be forced to remediate the
contamination, irrespective of blame.
1.3.5.3 Other industrialised countries
Most countries with a manufacturing history are beginning to tackle the contaminated land which
resulted from these activities. Discussion of every country is beyond the scope of this thesis,
however, the review by Visser (1993) which considers most countries is recommended. Two
countries with significant contaminated land problems and comprehensive policies which require
further consideration are the USA and the Netherlands.
1.3.5.3.1 The Netherlands
The Netherlands has a high population density (Hoogendoorn, 1984) and relies upon groundwater
sources for drinking water the threat to these supplies from soil contamination are extremely
severe. The Lekkerkerk incident in 1978 was the impetus for the Netherlands soil protection
policy. As a direct result of Lekkerkerk, the "Ginjaar Inventory" was compiled, a list of soil
contamination instances which detailed over 4000 cases. The Interim Act on Soil Decontamination
was introduced in 1983 as a 'stop-gap' measure until the Soil Protection Act replaced it in 1987,
which was replaced in 1994.
The 1987 act changed the emphasis from soil clean-up to prevention of soil contamination. The
policy of 'care obligation' was proposed which required operators of sites whose activities were,
or were suspected of polluting the soil to stop the pollution or limit the damage caused or
neutralise it. The Dutch Government adopted the policy of multifunctionality of soil and linked
this idea to the A, B and C values of soil quality for selected components, which are considered
in section 1.3.6. The government also adopted the policy of 'sustainable development' to their
environmental strategy in 1989 (before the EC 5th action programme) and incorporated it into the
Implementation Programme for Soil Protection 1990 - 1994. This programme had three aims: i)
43
prevention of soil contamination, ii) integrated approach to specific sources and iii) clean-up of
soil contamination (Keuzenkamp et al., 1990). The Soil Protection Act was revised in 1994, with
changes to the soil assessment values and dilution of the principle of multifunctionality, due to
the escalating cost of this policy. The policy of strict liability was introduced in the 1987 act and
further strengthened by the 1994 act by placing an onus upon the site owner to prove their own
innocence (Soil Protection Act, 1994).
1.3.5.3.2 The United States of America
In the USA, the Love Canal incident of 1974 provided the momentum to the legislative processes
which resulted in the three major pieces of USA legislation concerned with contaminated land;
RCRA, CERCLA and SARA.
RCRA (1976), demanded that wastes were treated with the best demonstrated technologies
(Skinner, 1988) and had the aim of preventing sites currently in use becoming problem sites in
the future. The HSWA (1984) required the USEPA to use RCRA to prevent disposal of untreated
wastes to land (Anon, 1991b). These Land Disposal Regulations (LDR) required that all
hazardous wastes were treated to an extent that protects human health and the environment prior
to land disposal (Anon, 1990a). For contaminated soil this meant that all the hazardous
constituents must be treated by one or more technique so that the residual concentrations meet the
USEPA defined limits for land disposal. Further discussion of the LDR are covered
comprehensively in reports in 'The Hazardous Waste Consultant' (Anon, 1990a and Anon,
1991b,c).
CERCLA (1980) or the 'Superfund' act was designed to identify contaminated sites, required
immediate of planned removal of contaminated material where serious or immediate risks were
identified and also required the production of a National Priorities List (NPL) of sites requiring
less urgent remedial action but which presented a risk to human health, water supplies and other
sensitive targets (by 1991 there were over 1200 sites on the NPL (Bisio, 1991)). The National
Contingency Plan (NCP), incorporated into CERCLA, dictated levels of clean-up required at
Superfund sites and the procedure to follow for discovery, response and restoration of hazardous
waste sites. Funds to implement CERCLA were raised through taxation of the petroleum industry
and the import or manufacture of certain chemicals.
SARA (1986) (re-authorising CERCLA) demanded standards of clean-up, stipulated rules for
remedial action selection and insisted upon compliance with NCP standards (Simms, 1990). The
44
USEPA was charged with implementing remedial actions involving treatment that; "...
permanently and significantly reduced the volume, toxicity, or mobility of hazardous substances,
pollutants, and contaminants...', and that were 'protective of human health and the environment,
... cost-effective ... utilised permanent solutions and alternative treatment technologies to the
maximum extent practicable' (Simms, 1990). The revised NCP emphasised that the choice of
restoration technique was biased towards treatment by innovative technologies, rather than non-
treatment or traditional technologies (Kovalick, 1990).
In the USA, liability for contamination rests with the polluter or the current owner of the site.
The federal government can recoup all costs from the responsible party, but private individuals
cannot, and must resort to common law and prove negligence or trespass on the party responsible
for the contamination (McKenna & Co, 1989). The effectiveness of the USA "Superfund" policy
has been subject to much investigation, as most of the money spent has been on fighting legal
cases for liability. For an in-depth discussion of the Superfund policy reference should be made
to Church and Nakamura, 1993.
1.3.6 Guidance for assessing the degree of contamination to land
'i4// things are poisonous yet nothing is poisonous. Dosage alone determines poisoning
This was stated by Paracelsus, a 16th century Swiss physician (Lawrence, 1986). When
considering contaminated land, the assignment of values to 'uncontaminated' and "contaminated'
soil is difficult because, as stated earlier, soil is extremely heterogenous and variable. The major
parameter used to evaluate soil contamination is the threat the soil may pose to human health.
These threats to man and the environment were specified in section 1.3.3. Many governments and
individuals have produced guidance values assessing contamination to land, these are considered
in depth in the review by Visser (1993), however, some of the more important assessment values
are discussed below.
1.3.6.1 Canada
The Canadian approach to soil quality is based upon two criteria: assessment and remediation.
The assessment values are considered to represent uncontaminated soil, whilst the remediation
values are split into three categories of land use: agricultural, residential/parkland and
commercial/industrial. The remediation values are not absolute, but 'are considered generally
protective of human and environmental health for specified uses of soil ... , based on experience
45
and professional judgement',• (CCME, 1991). Table AI.l, in Appendix I, details selected
contamination criteria, as defined by the CCME.
1.3.6.2 The Netherlands
Dutch assessment criteria were originally based upon the A, B and C values. Table AI.2 in
Appendix I shows some selected values. If a soil sample were to fall within the 'A' range it was
considered to have maintained its multifunctionality ('uncontaminated'). These values were used
to govern the restoration of the soil quality of contaminated sites and ensure that the aim of
purification of the soil and groundwater was achieved. However, these values were considered
to be too rigid with respect to taking the soil type into consideration, and its effect upon the
behaviour of the contaminants, hence the amended values for the A (reference) and C
(intervention) levels (the B level being deleted) were introduced with the 1994 Soil Protection Act.
A means of modifying these values by accounting for the soil type and the organic material in the
soil was also derived, which is explained in Appendix I and example A and C values are provided
in table AI.3.
1.3.6.3 United Kingdom
The Greater London Council (GLC) performed many site restorations from the 1960s up to its
abolition in the early 1980s. The contaminated land unit of the GLC compiled values from the
remediation work performed, this information was published in the SCI Contaminated Land
conference in 1979 (Kelly, 1980) (table AI.4 in Appendix I). These values became accepted as
'definitive' values for assessing contamination, although that is not what they were intended for
and have received much criticism. They are still used for some site assessment purposes, largely
in the absence of any other values. The DoE ICRCL guidance notes have provided the major
reference for site redevelopment in the UK. These figures are not meant to be used as absolute
values, but were provided solely as a means of advice (see Appendix I, tables AI.5 and AI.6).
In summary, site assessment values under development are frequently being based upon risk
assessment criteria. However the continual uncertainty regarding acceptable dosage or exposure
is still a major area of discussion.
1.3.7 Redevelopment of contaminated land
A systematic approach to site redevelopment is normally undertaken which is broken down into
46
five stages: identification; investigation; assessment; remedial action and monitoring, as shown
in table 1.7
Table 1.7 The five stages recommended when conducting a site restoration
Redevelopment Actions required Reasoning for actions.
Identili cation
(i)
Historical use of site Establish if site has been subject to a contaminative use (See table 1.5)
Identili cation
(i)
Visual inspection of site Lack or poor growth of vegetation. Odours from site, discarded drums, surface discolouration
implies contamination.
Identili cation
(i)
Mapping of site Establish positions of hard cover, sumps, old foundations and lagoons for use in investigation.
Investigation (ii)
Design of sampling procedure
The investigation must be performed to ensure that the whole site is covered thoroughly, and that
any results are representative of site conditions.
Investigation (ii)
Sampling of soil Establish levels of chemical contamination to site.
Assessment (ill)
Evaluation of data from investigation.
Confirm whether the site is contaminated Assessment (ill)
Does that site pose a risk to the public?
Is the envisaged end-use of the site compatible
with the contamination found?
If the site is heavily contaminated (see Appendix I), the building of homes on the land is probably
an unrealistic proposition.
Remedial action (Discussed in detail
in section 1.3.8 (iv)
Change in the proposed development
Usually cost is the major influence. To retain the original development proposal may require soil treatment that costs more than the value of the
completed site.
Remedial action (Discussed in detail
in section 1.3.8 (iv)
Treatment to reduce the concentrations of
contamination found on the site
Excavation of the pollution, or mixing affected soil with clean fill can often affect the required
contamination reduction
Monitoring
(V)
Continual checking of the site after remedial
action and re-development
Ensure that the decision made in (iii) or (iv) above was effective.
Adapted from: (BSI, 1988,Griffiths, 1992 and ICRCL, 1987)
Further detailed guidance on site assessment is beyond the scope of this work, but more detail
is found in Leach and Goodger (1991), Bridges (1987), lEHO, (1989), Caimey (1986 & 1993),
Harris and Herbert (1994) and ERL (1987).
1.3.8 Remedial treatment methodology for contaminated soil
47
There are five general categories which classify the available soil treatment technologies. They
are detailed in table 1.8. These technologies are applicable to eidier in-situ or ex-situ and can be
further classified as mobile or non-mobile (Griffiths, 1992). The trend in mainland Europe is to
use soil treatment centres, where the soil is taken for treatment and then returned to the site. In
the USA, the mobility of a treatment is a factor, due to the physical size of the country and the
fact that most of the sites requiring remediation tend to be isolated.
Table 1.8 The Ave sub-categories for soil treatment technologies and their applicability to soil treatment
Treatment category Examples of technology Applicability (in-situ or ex-situ)
Thermal Incineration ex-situ
Thermal desorption ex-situ, in-situ
Vitrification in-situ, ex-situ
Biological Bioremediation in-situ, ex-situ
Chemical Oxidation, dechlorination etc. ex-situ
Extraction Soil washing ex-situ
Mineral processing ex-situ
Vacuum extraction in-situ
Encapsulation Solidification ex-situ, in-situ
Barriers (curtain walls etc.) in-situ
Capping in-situ
The rapid development of laws designed to protect the environment, discussed in the preceding
sections, has had the effect of spurring research efforts into treating the problem contaminated
land presents. This has been evidenced by the multitude of novel methods for soil treatment which
have become available on a commercial basis over the last decade. The USEPA SITE programme,
introduced with SARA (section 1.3.5.3.2) has probably been the most prolific source of new soil
treatment systems. During the nine years since its inception, the SITE programme has encouraged
the development and testing of new methods for contaminated soil treatment. Invariably, most of
the treatment techniques originating from this programme can be regarded as variations on a
theme, but they do demonstrate that removal of pollution from soil, followed by its neutralisation,
can be effected without transfer of the pollution to a second medium. Of the main treatment
options available for contaminated soil, thermal processing in the most often suggested as is
illustrated by figure 1.4, with 36% of all sites using it.
The rationale for the appeal of thermal over non-thermal methods is because the major cause of
48
soil pollution is from the use and disposal of organic chemicals and related compounds, which
are most effectively and permanently treated using thermal methods, and the permanent treatment
of contaminated soil is preferred under US legislation. The emphasis of the following discussion
will be directed towards thermal remediation methods for soil,with only a limited review of the
other soil treatment strategies previously detailed in table 1.8.
Figure 1.4 The proposed application of available soil treatment methods to contaminated
sites in the USA during the period 1987-1989
In—situ vitrification (1.0%)
S/S (25.0%)
Soil washing (3.3%)
Chemical extraction (2.8%)
Bio—remediation (10.5%)
Off-s i te incineration (13.3%) Chemical destruction (1.0%)
On-si te incineration (17.1%)
Thermal desorption (4.3%;
Other (3.3%)
Vacuum/vapour extraction (13.8%)
In—situ soil flushing (4.8%)
S/S — Solidification/Stabilisation
1.3.8.1 Thermal treatment technologies for contaminated soil
Thermal treatments for contaminated soil can be classified as either low temperature (<550 °C)
(Ayen and Swanstrom, 1991) or high temperature (>650 °C) (Cudahy and Eicher, 1989). There
are three general areas of thermal treatment: (i) incineration, (ii) evaporation/pyrolysis and (iii)
vitrification.
1.3.8.1.1 Incineration technologies
Incineration is the largest sub-group and probably the most understood methodology since
incinerators have been used for over 100 years to destroy wastes (Oppelt, 1987), although modern
incinerators, with their complex treatment for the combustion gases, are only very distant relatives
of their fore-bears. The three major classes of incinerator are: fluidised bed, fixed hearth and
rotating kiln, typical examples of which shown in figures 1.5, 1.6 and 1.7. To destroy the solid
or liquid wastes, the feedstock are heated to over 850 °C in an oxidising atmosphere (usually air).
49
The solid residue is expelled and the hot gases are then combusted for at least 2 seconds before
they are cooled, washed and filtered and then released to the atmosphere. The combustion
processes occurring within the kiln of a rotary incinerator are shown in figure 1.8 (Santoleri,
1990).
Figure 1.5 A typical fluidised bed combustion chamber
1.0-9.0 Cetnbu*llen
GM RMktene*
1400f . tMO-F
Auiwtmr Fw
FHitewno
Figure 1.6 A typical fixed hearth combustion chamber
100-200% ExcMS Atr
Diftcharg* 10 Ouanch or
HmI Recovery
0.25-2.5 Seconds Ween Re»lder>ca Time
Secondary Chamber
1400*F - 2000-F
Primary Chamber
1200.F - 1800*F
Auxiliary Fuel
Transfer DItcherfle
Ram
Steam
Auill lary Fuel or Liquid Waate
50 - 60% Stoichiometric atr
Refractory
Aah Dtacharge
50
Figure 1.7 A typical rotary kiln and after-burner combustion chamber
Wast* Liquids
Auxiliary Fual
Watta Sofidt.
Oitcharoa to Quamch or Haat Raoovary
Auxiliary Fual
Liquid Waatf
Rofary M M
RaaManea HUN
Shroud
Rotary Klin Afttrbumer
(figures 1.5, 1.6 and 1.7 are from Dempsey and Oppelt, 1993)
Figure 1.8 The combustion processes occurring within the kiln of a rotary incinerator
KILN ROTATION 0.5 - 2.0 RPM
SOLIDS FEED EXI-IAUST
GAS HC
REFRACTORY LINING
RADIATION FROM REFRACTORY
radiation FROM FLAME
STEAM- CARBON BURNOUT
J VOLATILES
. '.~::SLAG/ASH
INCINERATION RAMSFCRWATION - — COMBUSTION DRYING
WASTE , LIQUIDS/
FUELS
The efficiency of an incinerator in destroying the waste input is measured using the term
Destruction Removal Efficiency (DRE). The DRE is calculated thus:
W -W 2XRE=[-JL__f*j*100
Where: W, =Mass feed rate of waste fed into incinerator
=Mass emission rate of the waste in the gases prior to atmospheric
release.
51
In the USA, a minimum DRE of 99.99% is required for RCRA principal organic hazardous
constituents (POHC) in the wastes. Wastes containing PCBs have a minimum DRE of 99.9999%,
as stipulated by TSCA (USEPA, 1989). Similar PCB DRE values are required for UK and
German incinerators (Butcher, 1990). The emission of acid gases (HCl & HF), particulates,
metals (Cd, Th, Hg gfc), NO^ and SO, are also stringently controlled. Current UK HMIP,
German and proposed EC emission standards for chemical waste incinerators are detailed in the
recent RCEP 17th Report on Incineration of Waste (RCEP, 1993). More detailed discussions on
incineration are given by Oppelt (1987), and Dempsey and Oppelt (1993).
1.3.8.1.2 Incinerators specifically designed for contaminated soil
Infrared incineration uses suspended heating elements to irradiate a conveyor carrying the soil.
The contamination undergoes pyrolysis at temperatures normally around 850 °C. The gaseous
products of the pyrolysis enter a secondary chamber where they are burnt at 1250 °C (Johnson
and Cosmos, 1989). The process was applied in the USEPA SITE program using soil from an
oil re-refining site which was contaminated by PCBs and Pb. The solid residue passed the USEPA
TCLP test for the leaching of Pb. A DRE of 99.99972 % to 99.99880 % was determined for the
PCBs (USEPA, 1989).
One of the most intriguing 'incineration' processes developed is "The ChemChar Process" (Kinner
et ah, 1993). The contaminated soil is mixed with a low-grade granular activated carbon and
packed into a column. The column has a moist oxygen stream passed through it from one end,
whilst the char/soil mix is ignited at the other. The flame front moves towards the oxygen source
(reverse burning), igniting fresh parts of the column, but depriving the previous flame region of
oxygen, thus extinguishing the flame tail. When the flame reaches the oxygen input point, the
burn assumes normal characteristics and burns in the direction of the oxygen flow (forward burn).
Typical temperatures achieved by the reverse burn are between 500 and 1500 °C (Cady, 1990).
An artificial contaminated soil, containing mineral oil and PCBs was treated using the reverse
burn and forward burn technique of the chemchar process. A 99.9999 % DRE of PCBs from the
soil was achieved, leaving an inert carbonaceous solid residue (Kinner et al., 1993).
Schneider and Beckstrom (1990) described a pyrolysing kiln used to decontaminated 35 000 t of
soil from a coke plant which was contaminated, typically, with cyanides and PAHs. The soil was
heated indirectly in the kiln to a maximum temperature of 750 °C for approximately 1 hour. The
overall removal efficiency of 20 PAHs from the soil was 99.8 %. The soil was replaced in the
site, while the off-gases from the pyrolyser were incinerated in an after-burner, residence time
52
3 seconds, at temperatures up to 1300 °C. The cited advantage of this process was that the soil
retained its fertility and could be re-used (Mackenbrock and Horch, 1990).
Mobile incinerators have been considered as a viable option for large contaminated sites where
transport costs would be excessive. A rotary kiln system with a secondary combustion chamber
was used to incinerate soil contaminated with PCBs to levels of 14700 ppm. The kiln operated
at a temperature of 800 °C and the secondary combustion chamber at 1200 °C, with a retention
time of 2 seconds in the secondary chamber. A DRE of 99.9999 % for PCBs was reported
(Leuser et al., 1990). The same incinerator was noted to give concern to the operators because
it was suffering from oxygen deprivation by emitting too much CO during the destruction of a
PCB contaminated oily lagoon sludge which exhibited a high energy value (average 2900 Btu/lb)
and a consequently high organic content. To correct this, oxygen lances were fitted to the
incinerator (Velazquez et al:, 1991).
Although mobile incinerators are not regarded as an acceptable treatment option in many
European countries, particularly in the Netherlands (Koopmans and Reintjes, 1988), some mobile
systems have been developed. Glaser (1988) discussed a rotary kiln incinerator where the soil was
dried first at 200 °C using the combustion air from the kiln, following which the soil was
incinerated at 1200 °C in the rotary kiln for up to 1 hour. The kiln gases were secondary
combusted at 1200 °C before cleaning and discharge. Removal efficiencies of >99.9 % for
PAHs, PCBs and cyanides were reported.
1.3.8.1.3 Thermal evaporation systems
This process can be performed either in- or ex- situ and is utilised by many treatment methods
to remove the volatile compounds from the soil. The components affected are normally organic,
but some volatile metals, inorganic compounds and complexes (for example Hg and cyanides) are
also removed. The advantage of desorption is that the soil structure is not severely damaged and
the treated soil can be returned to the site as 'clean' fill. Pollution by volatile hydrocarbons
(VOCs) (aviation fuel, fuel oils etc) is a problem which is commonly caused by leakage from
underground fuel storage tanks, of which there are an estimated two million in the USA, 25%
of which are thought to leak (Nielson and Cosmos, 1989). In the UK there are 19 000 petrol
stations alone, all of which may be leaking (Henton and Young, 1993). Thermal desorption is a
preferred treatment for soil in the Netherlands as it allows the soil to be re-used and is a cost
effective method of treatment. Ex-situ methods shall be reviewed first.
53
The Low Temperature Thermal Treatment (LT^) uses hollow archimedean screws which are filled
with an oil which is heated to 340 °C. The soil is mixed and heated by the screws to 200 °C and
the moisture and entrained VOCs evaporated from the soil. Initial versions of this process used
an after-burner at 1000 °C to destroy the VOCs (Anon, 1986), but in later systems the VOCs pass
into a gas treatment system, where the gas is condensed. Removal efficiencies of 100 % were
reported for dichloroethene, trichloroethene and tetrachloroethene (Anon, 1986) and >99.9 %
for VOCs (Nielson and Cosmos, 1989).
Asphalt production equipment has also been used to treat VOC soil contamination. The aggregate
used to make asphalt must be dried in a directly heated rotating dryer before it is mixed with the
liquid tar. Utilising the dryer to remove the VOCs from contaminated soil has been shown to be
effective. The VOCs and any water are evaporated at 180 °C before being exposed to the dryer
flame at 1100 °C. The VOCs are subsequently burnt, with removal efficiencies of >98 %
(Anon, 1985 and Gunkel, 1990). A variation on this process uses the soil as an aggregate
replacement material. Aggregate and contaminated soil is mixed in a 95:5 ratio respectively
in the aggregate dryer and then made into asphalt. Removal efficiencies of 99.25 % have been
reported for organic contamination (Czarnecki, 1988).
Kreft (1989) discussed the use of an established industrial process, the carrier gas process (which
has been used extensively in the mineral industry to convert compounds such as CaCOj into CaO
and metal oxides into the metal), for decontaminating soils. The soils are crushed below 3 mm
then pumped into a cyclone where they are mixed with hot air, dried and then separated. The dry
soil then enters a second cyclone, where hot gases from a combustion zone heat the soil and drive
the contamination off the soil into the gas stream. The cleaned soil and contaminated carrier gas
are separated by the second cyclone, with the soil exiting the system to a third cyclone where air
for the drying stage is heated by the cooling of the decontaminated soil. The contaminated carrier
gas enters the combustion zone where all the contaminants are oxidised at 1200 °C. The process
is claimed to be effective for soils contaminated with PAHs, PCBs, cyanide and volatile metals.
Other desorption technologies include the X*TRAX® process. The soil, contaminated with
organics, is placed into an indirectly fired rotating kiln and heated to a maximum temperature of
approximately 425 °C. The water and organic components volatilise into an inert atmosphere
(nitrogen), and are drawn off for washing to remove entrained solids, then cooled to 4 °C in two
stages to condense the vapours. The organics and water are separated and the organics removed
for re-use or disposal. Both the carrier gases and water are reused in the system. The X*TRAX®
process has demonstrated a removal efficiency for chlorobenzene, tetrachloroethene and xylene
54
from soil >99.9 %, and >98 % for mercury (Ayen and Swanstrom, 1991; Eisenhower et al.,
1990).
George et al. (1992) applied microwave energy to decontaminated soil. Microwaving the soil
alone did not prove effective in removing the contamination, with a maximum achievable
temperature of only 200 °C. However, addition of carbon chips enhanced the heating
characteristics of the soil through ionic heating, so that 1000 °C was reached. DREs of 60-99 %
were achieved, with differences in the DREs being attributed to co-evaporation effects (azeotropic
behaviour) between co-contaminants of the test soils.
All the preceding techniques required the soil to be excavated before it was treated. There are,
however, in-situ thermal desorption techniques available. Radio frequency (RF) heating uses the
ability of electromagnetic radiation to excite molecular species causing vibration and rotation.
When applied to soils these motions heat the soil and cause the volatilisation of moisture and
entrained VOC contamination (Downey and Elliott, 1990). Application of the process to soil
contaminated with aviation fuel proved successful. Electrodes were inserted into the ground at
25 cm intervals, to a depth of 2 metres, and a cover was placed over the treatment area to collect
all off-gases. The soil was heated to a temperature of 150 °C over a 9 day period, and maintained
there for a further 4 days. The semi-volatile aliphatic and aromatic were eliminated to the extent
of 94 and 99 % respectively. A VOC removal in excess of 98 % was realised (Dev and Downey,
1988).
Steam stripping of soil is suitable for volatile organic compounds, as was demonstrated at a fuel
transfer station. Two augers capable of boring to a depth of 8 metres, pumped steam (at 200 °C)
and compressed air (at 135 °C) into the soil. The off-gases were collected by a cover system
which was maintained at a reduced pressure. The extracted organics were condensed out of the
gas stream. A removal effectiveness of 85 % of VOCs and 55 % semi-VOCs were achieved. The
soil structure and physical character were not impaired by the process (de Percin, 1991). It was
noted by Hilberts et al. (1986) that steam stripping was most effective in sandy soils.
1.3.8.1.4 Vitrification
This treatment can be either in-situ or ex-situ. Heating soil to 2000 °C, well above its normal
fusion temperature (1100-1400 °C), converts it into an obsidian-like material. The in-situ process
was developed by the US Department of Defence to treat radioactively contaminated soil in-situ,
because excavation presented too high a risk. In the process, graphite coated molybdenum
55
electrodes are inserted into the soil and a starter path of glass frit and graphite is laid across the
soil surface between electrodes (soil does not normally conduct electrical currents). Once the
current is applied the frit melts and this soon spreads down through the soil area bounded by the
electrodes, converting between 3-5 tonnes of soil per hour. A covering hood, at a slight negative
pressure, draws all the off-gases into a treatment system. The organics are pyrolysed by the
melting of the soil, and non-volatile metals are trapped within the matrix. The final product is
also reduced in volume by up to 40 %, depending upon the initial soil composition (Shelley,
1990; Fitzpatrick et al, 1988; Jacobs et al., 1988). Timmons (1990) evaluated vitrification as a
means of treating soils containing Hg, As, pesticides and PCBs. He found that the Hg was
volatilised, 67 % of the As was encapsulated in the residue and 24 % lost by volatilisation. The
pesticides (Aldrin and Dieldrin) were reduced by 100 % and 98.5 % respectively, and the PCBs
were destroyed with 99.9 % efficiency. All the residues passed the USEPA TCLP test (Timmons,
1990). Figure 1.9 shows how in-situ vitrification is applied to contaminated soil.
Figure 1.9 The operating sequence for in-situ vitrification of contaminated soil
GRAPHITE AND FRIT STARTER
OFF GAS HOOD
BACKFILL POROUS GLASS
SUBSIDENCE
ROCK
ELECTRODE MELTING
ZONE VITRIFIED SOIL/WASTE
(After Jacobs et al., 1988)
Ex-situ vitrification is available in several forms. The plasma centrifugal furnace uses the intense
heat ft-om a plasma torch (> 10,000 °C) to melt soil and pyrolyse organic contamination which
has been vaporised. Tests using hexachlorobenzene and ZnO in soil demonstrated 99.99 %
removal of the QClg and the residue passed the USEPA TCLP test (Staley, 1992). The cyclone
furnace is an adaptation of existing boiler technology. The waste soil is injected into a vortex of
air and fuel gas, rapidly reaching 1650 °C, at which point the soil melts and clings to the furnace
sides, pyrolysing the organics. Tests using Pb, Cr and Cd contaminated soil showed that the
residue passed the TCLP test (Czuczwa et al., 1991). The electric pyrolyser melts soil at
56
1650 °C, vaporising liquids and cracking organic compounds. The atmosphere in the pyrolyser
is essentially free of oxygen, so organics are converted to CO, Hj and carbon. Using a soil spiked
with 1,4-dichlorobenzene, the destruction efficiency of the system was assessed to be >99.9 %
in terms of the Total Organic Carbon (TOC) content of the input soil (Cepko, 1987). Finally, a
conventional glass furnace was used to treat soil from a dockyard containing blue asbestos, coal
tars, Pb, Zn, Cu and Fe. The residue passed the USEPA TCLP test, although emissions of
particulates from the furnace were above acceptable limits (Anon, 1990b,c).
1.3.8.2 Other soil treatment methods
The other four major soil treatment methods are bioremediation, chemical, extraction and
encapsulation, each shall be briefly considered.
1.3.8.2.1 Bioremediation
Bioremediation utilises natural organisms (fauna and flora) to take up problematic components of
the soil (organic molecules and metal species), and then convert them into benign forms through
metabolic processes, or just remove them from the soil. The process can be either in- or ex- situ.
The applicability of bioremediation to gasworks sites, and in particular PAHs, has been discussed
in two reviews by Thomas and Lester (1993a,b), and by Wilson and Jones (1993). The major
problem with bioremediation is that it is not readily applied to heavy contamination nor sites of
mixed contaminants (for example gasworks or chemical plants). This is because most organisms
can only metabolise a small number of organic species, but any sites which possess a multitude
of chemicals probably will contain toxins as well as nutrients.
1.3.8.2.2 Chemical
Chemical treatment of soil relies on the reactions between contaminative substances in the soil and
the added process chemicals. Wet oxidation is one potential process which uses oxygen and heat
to oxidise a slurry of water and soil which may possess organic and metallic contamination. The
metals are converted to their highest oxidation states and the organics broken down into CO;,
HjO, acetic acid and other simple organics (Sarensen et al, 1990). Most recalcitrant organic
compounds are chlorinated, therefore chemical degradation to remove the chlorine has been
proposed. The KPEG process, designed for PCB decontamination, operates by mixing
polyethylene glycol with KOH in a suitable solvent. The contaminated soil is mixed into the
reaction vessel, where the chlorine atoms of the PCBs are replaced, then recovered as KCl. DREs
57
of 96 % have been reported (Carpenter, 1986).
1.3.8.2.3 Encapsulation
Encapsulation of soil can be performed in-situ or ex-situ. By mixing the soil with cement, water
and other additives, a monolithic mass can be formed which prevents the migration of the
contaminants associated with the soil into the wider environment. Where large areas of soil area
regarded as a threat, a cementitious wall can be constructed around the soil by injecting liquid
cement grouts into the ground on the edge of the contamination to form 'curtain walls' or
barriers. Detailed descriptions of encapsulation technologies have been given by Smith (1986),
de Percin (1989) and Stinson (1990). These processes are limited in applicability, because some
contaminants interfere with the hydration processes which cause the setting of the cement,
particularly organic contamination. Additionally, many contaminants will be aggressive to the
alkaline encapsulation environment, including sulphates and acids.
1.3.8.2.4 Non-thermal extraction
Non-thermal in-situ extraction of soil relies on the use of low pressures to suck contamination
from soil, or on the ability of air injected into the soil to disrupt the soil gas-contamination
equilibrium and allow the contaminants to be flushed out of the soil. Examples of application of
these processes have been given by Hoag et al. (1989) and Hutzler et al. (1991). This technique
is limited in application to VOC contamination.
Ex-situ non-thermal extractive techniques are basically soil washing technologies. Figure 1.10
shows a typical flow-sheet for the physical extraction of contamination from soils. The expertise
of the mineral-processing industry has been adopted to clean soil and debris typically associated
with contaminated sites. By utilising particle separation techniques, surfactants and complexation
agents, the organic and metallic contamination of soil can be effectively separated. The finest
portion of the soil (clay and silt fraction) is usually separated from the soil bulk as it is the portion
of the soil which is the most adsorptive and tends to concentrate the contamination. Ultimately,
the soil washing processes leave an aqueous and solid residue in which the contamination is
concentrated and requires further treatment (NeeBe and Grohs, 1990; Stinson et al., 1990;
Raghavan et al., 1991).
58
Figure 1.10 A typical flow-sheet for the physical extraction of contamination from soils
Coarse 10mm Screening
2mm Trommel Screen
Scrubber
I Debris
Washing
Feeder
f '1 OZD
Debris
Flocculation Sedimentation
Disposal or Treatment
Spiral Classifier
Slime Attrit oner
Cyclone
Flotation Cells
Conditioner
Organlcs and Fine Metals
Coarse Meta s
Spiral Concentrator
Filtration and Detoxification
Decontaminated Soil Sedimentation and Reuse
(After Armishaw et al., 1992)
1.3.9 The economics of soil rehabilitation technologies
A guide to the relative costs of the soil treatments discussed previously are shown in table 1.9.
Generally, the more resistant contamination is to natural degradative processes, the more
expensive the treatment options. For example, PCBs and other chlorinated organic molecules are
considered to present the greatest threat to the environment and are thus the most expensive
contaminants to treat (see Incineration and the KPEG process in table 1.9).
59
Table 1.9 Relative costs of different clean-up techniques
Treatment classification
Technology Cost (£/tonne)
Reference
Excavation and landfill
Dig and dispose 2 5 - 5 0 Smith and Parry, 1993
Encapsulation Solidification 65 Shelley, 1990
Encapsulation Slurry or curtain walls 3 5 - 7 5 Smith and Parry,
1993
Physical methods Soil washing, vacuum extraction
2 5 - 7 5 Smith and Parry, 1993
Biological Bioremediation 1 6 - 3 5 Smith and Parry, 1993
Thermal Ex-situ Vitrification 50 Anon, 1990c
In-situ vitrification 180 Shelley, 1990
Incineration 160 - 260 Shelley, 1990
X*TRAX 100 - 160 Ayen and Swanstrom, 1991
Mobile rotary incinerator 100 Glaser, 1988
In-situ radio-frequency heating
85 (per m )
Dev and Downey, 1988
In-situ steam/hot-air stripping
110-285 (per m')
USEPA, 1991a
Pyrolysing kiln 5 0 - 6 5 Schneider and Beckstrom, 1990
Chemical KPEG process 140 - 250
(per m ) Carpenter, 1986
Stabilisation of wastes 35-75 Smith and Parry, 1993
Most countries encourage the use of technologies which remove or destroy the contamination to
soil, except for the UK where market forces are allowed to decide the viable treatment. The
redevelopment of Chatham dockyard unearthed asbestos, tars and extensive metal contamination.
The vitrification of the soil, which would have resulted in total destruction of the contamination
was rejected in favour of landfill for the 300 000 tonnes of affected material (Anon, 1990b,c,d).
When a site is to be decontaminated, the cost of the treatment is an important factor. In the UK
cost is the primary factor, with often the cheapest option being used. The historically low cost
of landfill in the UK has meant that treatment of soil has not been a viable option. In the
60
Netherlands, the government's aim has been to ensure that the contamination never poses a threat
to future generations. Since this cost can be excessive, a ceiling of Dfl 250 (£90) per tonne has
been established as the maximum expenditure allowable for soil treatment (DTI/OTS,1992). In
the USA the clean—up criteria set by the USEPA must be met, thus cost is not a major factor.
However, the soil treatment market is vast and extremely aggressive, thus competition between
rival treatment processes acts as a price control mechanism.
1.4 Adsorbents
1.4.1 Introduction
Adsorbents are materials that possess surfaces which exhibit affinities for other substances, but
they should not be confused with absorbents. The term adsorption 'refers strictly to the
existence of a higher concentration of any particular component at the surface of a liquid or solid
phase than is present in its bulk' (Glasstone, 1948), whereas an absorbent exhibits a uniform
uptake throughout the solid (such as by a sponge). When adsorption occurs, for example, by a
gas on a solid surface, the adsorbed gas is referred to as the adsorbate and the solid the
adsorbent (Sing et ah, 1985).
1.4.2 The characteristics of adsorption
Adsorption can be categorised into two types, (i) physisorption and (ii) chemisorption but only
type (i) is of concern to this work. Physical adsorption is the attachment of a molecule to a
surface, using electrostatic forces such as van der Waals forces or London disperson forces.
However, these forces are only very short range and can repel as well as attract molecules. The
fluctuation in van der Waals forces between two molecules as they approach one-another is
illustrated by figure 1.11.
The influence of van der Waals forces is such that they are the cause of non-ideal behaviour of
gases and the condensation of gases as liquids. However, van der Waals forces will only act over
approximately two molecular diameters, so the first adsorbate layer experiences the full van der
Waals attraction but the second layer is much less attracted. Thus, for vapour phase adsorption
the layers above the first experience an adsorption potential (which is, for practical purposes,
equivalent to the latent heat of vaporisation/condensation (AH^p/u,) which is an assumption of the
BET equation, discussed in section 1.4.2.1.3 and Appendix III. Hence, a monolayer will cover
the surface of a solid when AH^p is smaller than this adsorption potential. Utilisation of this
61
phenomenon has resulted in the information from adsorption studies being employed to calculate
surface areas of solid materials (Neely and Isacoff, 1982).
Figure 1.11 Typical van der Waals forces between a molecule and a planar surface as a function of separation in molecular diameters.
' Z o < REPULSIVE
n UJ CO O w a. t
li UJ TU
< ATTRACTIVE
ut K
SEPARATION BETWEEN
MOLECULAR CENTER AND S U R F A C E
IN MOLECULAR DIAMETERS
(Figure 3.1, pp 82, Neely and Isacoff, 1982)
The most common method employed to study physical adsorption is to use a solid adsorbent and
a gaseous adsorbate. By exposing a known weight of the solid to the gas, at a known pressure and
constant temperature (usually below the critical temperature of the gas) and volume, the gas
adsorbed by the solid will cause a drop in the pressure inside the sample holder. From this
pressure change the volume (or quantity, n) of gas adsorbed can be calculated. Measuring a series
of points at increasing gas pressures will allow the construction of an adsorption isotherm (Gregg
and Sing, 1982).
The isotherm can be expressed as: n = f(P/P°)j_
where: n = amount adsorbed , P = pressure in sample holder, P°= pressure in reference holder
It has been recognised that there are six general types of adsorption isotherm which can result
from the occurrence of physical adsorption on to solids. The first five were characterised by
Brunauer, Deming, Deming and Teller (BDDT) (Brunauer et al., 1940), and a sixth type, which
is relatively rare has also been added (Gregg and Sing, 1982). Figure 1.12 shows the six
recognised isotherm types. Through the classification of isotherms by BDDT, suppositions about
the physical characteristics of the solids under study can be made from the shapes of the
adsorption isotherms obtained which are summarised in table 1.10.
The isotherm type will also indicate the nature of the adsorption mechanism which is occurring
within the sample. A type I shape implies that micropore filling is occurring, a type II isotherm
62
implies monolayer-multilayer adsorption and a type IV isotherm shape implies that adsorption-
capillary condensation is occurring (Sing, 1989). The hysteresis exhibited by type IV isotherms
have been classified into 4 different classes (HI to H4) by lUPAC (Sing et al., 1985). From the
shape of the hysteresis loops it is considered possible to make assumptions about the pore
structure in the solid. Figure 1.13 shows these classifications. HI can be associated with
compacted powders or uniform particle sizes and shapes, H2 associated with some silica gels with
poorly defined pore sizes and H3 and H4 associated with slit-shaped pores (Gregg and Sing,
1982).
Figure 1.12 The BDDT classification for adsorption isotherms, types I to VI
I n j y
r m
s
^ Rtlotivt prtswrt, p/p^
(Figure 1.1, pp 4, Gregg and Sing, 1982)
Table 1.10 Interpretation of isotherm shape in terms of the physical characteristics of the solid under investigation
Isotherm Classification of solid
Type I Microporous solids
Type II Nonporous solids
Type III Nonporous or macroporous solids which exhibit weak gas-solid interaction
Type IV Mesoporous solids
Type V Microporous or mesoporous solids which exhibit weak gas-solid interaction
Type VI Surface is energetically uniform
63
Figure 1.13 The hysteresis loop classiHcation proposed by lUPAC
Relative pressure, plp°
(Gregg and Sing, 1982)
The shapes of pores are considered to be of four general types, as shown in figure 1.14. The
cylinder closed at one end, cone and wedge shape pores fill and empty with adsorbate at the same
pressures so do not exhibit hysteresis. The cylinder open at both ends, the ink-bottle pores and
the parallel sided slits fill and empty at different pressures hence exhibit hysteresis.
1.4.2.1 Mathematical treatment of adsorption
The three most commonly applied isotherm equations are: (i) Freundlich, (ii) Langmuir and (iii)
BET equations. Each of these equations require consideration in detail. There are also three
further treatments of adsorption processes to assist in the interpretation of gas adsorption data,
namely the application of the t-plot (Lippens et al., 1964; Lippens and deBoer, 1965), the Kelvin
equation (Barrett et al., 1951) and the Horvdth and Kawazoe equation (Horvdth and Kawazoe,
1983).
1.4.2.1.1 The Freundlich equation
Proposed by H. Freundlich (1907), this equation is essentially empirical in nature (Levine, 1978;
Knettig et al., 1986; McCabe et at., 1985) although it can be derived from the Langmuir equation
if the surface is assumed to be heterogeneous in nature (Levine, 1978; Adamson, 1990a). For
adsorption by a gas onto a solid, the Freundlich equation takes the form:
v=kpy"
Where: V = Volume of gas adsorbed per unit mass of adsorbent
k = constant (equivalent to the adsorbent capacity)
n = constant (equivalent to the intensity of adsorption)
P = pressure at which V is adsorbed
64
Figure 1.14 The four general pore shapes for adsorbent systems
(J ) (6)
ui,
r —
io)
B I
(<7 )
n
\Ji (A)
7 ^
y
(Gregg and Sing, 1982)
65
For adsorption from solution, it is written as:
—=kC]"' m
Where: x/m = Amount of adsorbate (x) taken up by unit mass (m) of adsorbent
C, = Equilibrium concentration of adsorbate (Shaw, 1992a)
The linear form of the Freundlich equation is derived simply by taking logs, resulting in:
log(-)=logii:+(i)logC, m n
The Freundlich equation is only applicable to dilute solutions because it does not show a limiting
value at high concentrations (unlike the Langmuir equation). This indicates that n is always
greater than 1 (Adamson, 1990a). Hassler (1963) suggested that the magnitude of the 1/n value
is considered to indicate the ease with which molecules are adsorbed by the adsorbent, a value
within the range of 0.1-0.5 is considered to indicate that adsorption is favourable.
1.4.2.1.2 The Langmuir equation
This equation was derived by I. Langmuir (1916), to account for the adsorption of gases, which
he considered to be limited to single molecular layers. It is based upon a kinetic treatment of the
process of adsorption, where the surface is comprised of a series of adsorption sites. Three
assumptions were made by Langmuir in his treatment of adsorption: (i) the surface of the
adsorbent is uniform (all adsorption sites are equal), (ii) there is no interaction between
neighbouring adsorbed molecules and (iii) only a monolayer of adsorbate molecules form on the
surface (Levine, 1978).
The equation which was derived is commonly written as:
n Bp
Where: n = Amount of gas adsorbed
p = pressure at equilibrium
B = Empirical constant
For adsorption from solution, the equation is expressed as:
X _
m 1 +BCg
66
Where; x/m = Amount of adsorbate (x) taken up by unit mass (m) of adsorbent
C, = Equilibrium concentration of adsorbate
Qnax = Monolayer capacity of adsorbent
The linear form of the equation is expressed as:
1 1 . . . 1
( 2 ) ^max m
From the monolayer capacity, the surface area of the material is readily calculated, assuming that
the surface area occupied by one molecule of the adsorbate is known. A full derivation of this
equation is given in Appendix II.
1.4.2.1.3 The Brunauer, Emmett and Teller (BET) equation
The basis of the BET equation is the Langmuir equation, except that the BET equation considers
the multimolecular adsorption which will result where physical adsorption onto a heterogeneous
surface occurs. The BET equation, like the Langmuir equation, is also based upon three
assumptions (Gregg and Sing, 1982): "... (i) in all layers except the first, the heat of adsorption
is equal to the molar heat of condensation, (ii) in all layers except the first the evaporation -
condensation conditions are identical and (iii) when P = P°, the adsorptive condenses to a bulk
liquid on the surface of the solid (hence, the number of layers becomes infinite)".
It is usually expressed as:
v^C-^ v=- ^
(1-Z)[1+(C_1)Z]
For data interpretation purposes, the linear form of the equation is preferred:
^ - 1 . ( C - D ^ P v ( l - A v„C po
Where: v = Volume of gas adsorbed
j/„ = Volume of gas adsorbed forming a monolayer on the sample
P / P = Relative pressure of adsorbing gas
C = BET constant
67
The surface area of the sample can be calculated from the monolayer capacity, in a similar
manner to that used for the Langmulr Equation. A full derivation of the BET equation is given
in Appendix III.
1.4.2.1.4 The t-plot
Derived by Lippens et al. (1964) and Lippens and deBoer (1965) the t-plot was proposed as a
means of examining the adsorption processes which occur during the measurement of an
adsorption isotherm and subsequently interpreting them in terms of the sample pore structure
relative to a non-porous reference material. They determined the N; adsorption isotherm for a
non-porous solid (type II isotherm). Application of the BET equation to the data resulted in a
value for (and consequently the BET surface area, SBET)- By assuming that the density of the
adsorbed Nj was equal that of liquid N;, an average layer thickness (a) of 3.54 A was used to
convert the volume of Nj adsorbed into a thickness value (t):
Where: t = average thickness of adsorbate on surface
a = thickness of single molecular layer
= Volume of adsorbate on surface
j'n, = Volume of monolayer of adsorbate
The values for t were tabulated against P/P° providing a standard reference table which enabled
the conversion of P/P° values into corresponding t values. Samples which are subsequently
examined can be tested against the standard isotherm, by plotting against t (rather than P/P°).
There are essentially four scenarios which can result and these are shown in idealised form in
figure 1.15.
A non-porous sample will be identical to the reference sample and so give a straight line plot
(case A). A microporous sample (type I isotherm) will provide more adsorption than expected for
a given t (P/P°) value, hence has a reduced gradient in the -1 plot (case B). The mesoporous
samples (type IV isotherm, with or without micropores - cases C or D respectively) exhibit a
stepped plot, where the first linear region represents micropore filling and the second region the
capillary condensation occurring in the mesopores. Extrapolation of the second linear region back
to the y-axis will provide a value for the micropore volume of the sample (y-axis intercept),
whilst the slope of this region is equivalent to the surface area of the meso and macropores (the
linear regions result from the unhindered formation of multilayers upon the solid surface).
68
Lippens and deBoer (1965) express the surface area as:
15.47 V , S=- ads
t
(the factor 15.47 converts the surface area into m^/g).
Figure 1.15 Idealised t-plot scenarios
VCJLUME
MICRDPORDUS NON-POROUS
FILM THICKNESS, t FILM THICKNESS, t
VOLUME ADSORBED
MICRO & MESOPOROUS MESOPOROUS
FILM THICKNESS, t n U 4 THICKNESS, t
(Figure 29, pp 111; Coulter Corporation, 1991)
Sing (1975) has stated that the t method is limited in its applicability by the reliance of the method
upon the BET equation for supplying the monolayer volume and that application must be limited
to solids which produce a well defined "point B". For this reason, the t-plot has been useful for
comparative purposes only in this work.
1.4.3 Porosity
Porosity is one of the most important features of adsorbent materials and a consideration of
porosity and the means available to evaluate this parameter is necessary. The word "pore"
originates from the Greek word for passage "iropoa" (McEnaney and Mays, 1988). The pores
which are found in adsorbent solids are not, however, simple passages and this fact has made the
analysis and evaluation of pore structures complex, and often resulted in unreliable information.
Porosity is used to describe the channels and cavities contained within the adsorbent. The pores
are classified into three size groups according to their average widths:
69
Macropores > 500 A (50 nm)
Mesopores between 20 A and 500 A (between 2 nm and 50 nm)
Micropores < 20 A ( < 2 n m )
(Gregg and Sing, 1982).
In this classification, "width' means the diameter of cylindrical pores, or distance between two
faces in a slit-shaped pore (Gregg and Sing, 1982, Adamson, 1990b). The pores are important
because they allow the transport of molecules into and out of the body of the adsorbent
(Broekhoff and Linsen, 1970).
1.4.3.1 Determination of porosity
The porosity of a solid material can be measured by two techniques: gas adsorption and intrusion
methods usually using mercury. Both techniques have limitations. Gas adsorption can only be
applied to determine pore size widths between 20 A (2 nm) and 250 A (25 nm) whilst intrusion
methods are limited to the macroporous/upper mesopore range > 500A (50 nm) because the high
pressures [up to 2000 bar ( « 1974 atmospheres)] used to force the mercury into the pore structure
are considered to cause distortion of the pores (Gregg and Sing, 1982, McEnaney and Mays,
1988). In light of the above observations, the remaining discussion is limited to consideration of
gas adsorption data for the evaluation of porosity.
The different types of isotherms resulting from adsorption studies were considered earlier in
section 1.4.2. The study of porosity is limited to materials producing type I and IV isotherms,
although the former type are complex and their interpretation is subject to much uncertainty. Thus
the following discussion is only concerned with the interpretation of type IV isotherms.
1.4.3.2 Type IV isotherms, capillary condensation and the Kelvin equation
Considering the type IV isotherm shown in figure 1.12, once the monolayer coverage of the
surface of an adsorbent has been completed, the processes of pore filling would be assumed to
occur by the successive accumulation of additional layers of adsorbate upon the walls of the
pores. However, instead of assuming a type II shape, the isotherm deviates upwards indicating
a much higher uptake of adsorbate than expected. Upon desorption of the adsorbate, the isotherm
undergoes hysteresis and so does not return along the same path. Zsigmondy (Adamson, 1990c)
has been accredited with providing the explanation of the deviation from type II behaviour by
proposing that capillary condensation occurred within the pores as the pressure of the adsorbing
70
gas was increased (Gregg and Sing, 1982). Zsigmondy had based his proposals upon the work
of Lord Kelvin and the equation which he had proposed to account for the fact that the vapour
pressure of a gas above a curved surface was different to that of a flat surface (Bromberg, 1980).
The Kelvin equation has its basis in the independent work of Young (Lord Kelvin) and Laplace,
who showed that the pressures on opposite sides of a liquid-vapour interface were related to the
radii of curvature r, and r; and the surface tension of the liquid by the Young-Laplace equation
(Gregg and Sing, 1982):
AP=P«-PP=Y(—+—)
Where: P" = Pressure of liquid
= Pressure of vapour
7 = Surface tension or surface free energy of liquid
When capillary condensation occurs within a pore, the liquid will be in equilibrium with its
vapour. By considering both the mechanical equilibrium (given by the Young-Laplace equation)
and the chemical equilibrium (from Gibbs chemical potential theory) it can be shown that:
In-f - ' pO RT
which is recognised as the Kelvin equation. Where: = Molar volume of the liquid
R = Gas constant
T = temperature in K
r„ = mean radius of curvature.
The equilibrium vapour pressure, P, over a concave meniscus of liquid must be less than the
saturation vapour pressure, P°, which implies that a vapour will be able condense to a liquid in
pores of a solid, even when its relative pressure is less then 1. Thus, capillary condensation will
occur at a pressure P, which will be determined by r„ (Gregg and Sing, 1982).
The Kelvin equation has subsequently been applied by numerous researchers to calculate the "pore
size distribution" of porous solids. The historical application of this equation to determine pore-
size distribution has been comprehensively reviewed by both Gregg and Sing (1982) and
Broekhoff and Linsen (1970). Of particular interest to this work is the method employed by
Barrett et al. (1951) to determine pore-size distributions. They utilised the desorption branch of
the isotherm to provide the data for the calculation.
71
The method assumes that: (i) the pores are cylindrical (open ended) and (ii) the amount of
adsorptive in equilibrium with the gas phase is retained by the adsorbent by two mechanisms; (a)
physical adsorption on the walls and (b) capillary condensation in the inner capillary volume. It
is assumed that when P/P° is almost unity, most of the pores are filled with adsorbate (for
example Nj). The in the pore will comprise a "core" of liquid adsorbate, of radius equal to
the Kelvin radius r and an amount adsorbed onto the pore walls, of thickness "t". Lowering P/P°
will cause some of the liquid to evaporate and a slight reduction in t of the adsorbed layer from
pores which are already empty but still retain some physically adsorbed Nj. Thus the volume of
gas desorbed will comprise a volume due to pore emptying and a volume due to thinning of
physisorbed gas on the walls of empty pores. The amount of Nj desorbed from the walls at each
successive pressure reduction is expressed as an area (which can be converted to a volume) which
is summed at each stage to allow the evaluation of the volume of gas desorbed which is solely
attributable to the core of the pore range under investigation. Hence the Kelvin radius is
calculated and to this value is added a value for the thickness of the adsorbed layer remaining on
the pore walls. The actual calculations can be found in the original work of Barrett et al. (1951).
The final "pore-size distribution" plot is of Av'/ArP vs. r, which represents the volume of gas
desorbed (Av"") per unit of radius pore change (Ar"") vs. the calculated pore radius (r"" + t, in
angstroms).
The above method for pore-size distribution calculation assumes cylindrical pores. As discussed
previously, the shape of pores within adsorbents are considered to be of several forms, and as it
cannot be known with certainty which pore shapes are present within a system (although the
isotherm shape may imply the presence of certain pores), all pore-size distribution data must be
considered with caution and can only be of use as a comparative tool.
1.4.3.3 The Horvath-Kawazoe equation
The limitations of the Kelvin equation in determining pores with average radii of < 20 A was
addressed by Horvdth and Kawazoe (1983), who developed a method of evaluating the "effective
pore diameters" in microporous and molecular sieving carbon solids. The term "effective pore
size" is used because in very narrow pores the diameter of the pore is of molecular dimensions,
hence the effect of variation in the effective radius of the carbon atoms which comprise the walls
of the pore will have disproportionate effects upon the pore size.
Horvdth and Kawazoe assumed that the micropores of molecular sieve materials are usually
considered to be slits between two graphitised carbon layers. By considering the interaction
72
between the plates in terms of Lennard-Jones potentials (by assuming that each plate behaves as
an infinite graphite plate) the expression arrived at can be written as:
j g J l n C — + - 2 ! L ] a\l-d) 3il-dJ2f 9(l-dJ2f 3(dl2f 9(^/2)'
Where:
Aa = Constant in Lennard-Jones potential
AA = Constant in Leimard-Jones potential
d = 4 + dA
da = diameter of adsorbent molecule
d* = diameter of adsorbate molecule
K = Avogadro's number
1 = Distance between nuclei of two layers
Ng = Number of atoms per unit area of adsorbent
Na = Number of molecules per unit area of adsorbate
P = Pressure of N;
P° = Saturation pressure of Nj
R = Gas constant
a = Distance between a gas atom and the nuclei of the surface at zero interaction energy.
This equation accounts for the variation in potential which occurs as the distance between the
plates changes. The effective pore size distribution is a function of (1 - d ) and is expressed as:
WAV. = f(l - d)
Where: W = Mass of N; adsorbed into pores smaller than (1 - d j
Woo = Maximum amount of N; adsorbed into the pores at P/P° = 0.9
The constants required for the above equation {a, N^ and Na) were extracted from the literature
by Horvdth and Kawazoe. They proposed that the equation was useful for determining pore
distribution between the range of 3.5 A and 14 A (0.35 - 1.4 nm). By choosing suitable values
for (1 - d j between 3.5 and 14 A a value for P can be calculated (and hence W). Plotting WAV„
vs. (1 - d j provides the pore size distribution.
1.5 Carbonaceous adsorbents
There are many classes of adsorbent materials which include materials such as zeolites and clays
but the type which are of most importance to this research are amorphous carbons. The class of
amorphous carbon of significance to this work is activated carbon. This is a family of solid
carbons, which cannot be readily assigned a structural formula nor identified by chemical
73
analysis, but typically contain varying amounts of heteroatoms (O, S, N) and some mineral
content. They are highly porous and have large surface areas, typically of several hundred m7g
(McDougall, 1991). These materials are of great importance to industry as they perform many
tasks as adsorbents, catalysts and reaction support media.
1.5.1 The history of activated carbons
The historical origins of activated carbon date back to the Egyptians and Hindus (1500 BC), who
recognised the adsorptive properties of wood char for water purification and medicinal usage
(McDougall, 1991). The first scientific report of the adsorbent properties of activated carbon was
made by Scheele in 1773, who noted the ability of wood carbon to adsorb gas, followed by
Lowitz in 1785 who reported its decolorizing ability (Houghton and Wildman, 1971). The ability
of char carbon to adsorb materials is how it became called activated, because other forms of
carbon did not show adsorbent properties unless specially prepared.
The modern activated carbon industry was established at the turn of the 19th century when
carbons superior to bone chars were prepared to purify sugar. The First World War accelerated
activated carbon research by the need for gas masks to protect the troops from the effects of
chemical weapons such as phosgene gas (McDougall, 1991). After the war, activated carbon
demand fell, so new markets had to be exploited. The success of this operation is verified by the
world production of activated carbon which was 365 000 tonnes in 1985 (Bansal et al., 1988).
The world usage of activated carbon is increasing rapidly, because the elevated demands of
environmental legislation have forced industry to re-appraise their releases into the environment.
For example, activated carbon production in the period 1978 to 1985 increased by 90 000 tonnes
(Holden, 1982 and Bansal et al., 1988). Activated carbon is one of the most versatile materials
where the control of pollution in aqueous or gaseous effluent is required. However, current
feedstock are predominantly non-renewable (coal, peat and lignite constitute over 50 % of current
supplies) and with the higher activated carbon needs, the requirement to develop more feedstock
has become an immediate research demand. Table 1.11 lists some of the materials which have
been used to make activated carbons.
1.5.2 The structure of activated carbons
The structure of activated carbon is considered to represent a very disordered graphite-like
structure. Bisco and Warrens (1942) were the first people to characterise the amorphous carbons
using carbon black as the subject. They stated that carbon black was "... a random layer
74
structure, in which single graphite layers are stacked roughly parallel and equidistant, but with
each layer having a completely random orientation with regard to rotation about the layer
normal". They summarised that carbon black was not a wholly amorphous material because it
contained 2D order, and that "... the term turbostratic (disordered layers) is suggested as a
name for this particular class of mesomorphic solids" (Bisco and Warrens, 1942). The currently
accepted structure of activated carbon is shown in figure 1.16, with sheets of aromatic rings
folding over one-another, with pores formed between the sheets. The disorder is readily
discemable (Bansal et al., 1988).
Table 1.11 Feedstock suitable for activated carbon manufacture
Bagasse Blood Bones
Cereals Coal Coal tar
Coconut coir Coffee beans Corn cobs
Cottonseed hulls Distillery waste Fish
Fruit pits Jute stick Kelp and seaweed
Lampblack Leather waste Lignin
Lignite Molasses Municipal waste
Newspapers Nut shells Oil shale
Palm tree cobs Peat Petroleum acid sludge
Petroleum coke Potassium ferrocyanide waste
Pulp mill waste
Refinery wastes Rice hulls Rubber waste
Sewage sludge Spent fullers earth Sugar-beet sludge
Sunflower seeds Wheat straw Wood (Adapted from Pollard et al., 1992)
1.5.3 The manufacture of activated carbon
Early activated carbons were made by heating wood in an air starved vessel or pit. The carbons
so produced were of a variable quality due to the poor control of temperature and environment
(Bancroft, 1920). Ostrejko is credited with pioneering industrial activated carbon manufacture
through his patents of 1900 which detail the production of activated carbons by carbonising
vegetable materials previously mixed with metal chlorides, or pre-charring the raw vegetable
material and exposing it to steam or CO; (SmlSek and Cerny, 1970). The principles he described
are applied today for commercial activated carbon manufacture. There are two stages to activated
carbon manufacture: carbonisation and activation.
75
Figure 1.16 The structure of activated carbon
1
(After Bansal et al., 1988)
1.5.3.1 Carbonisation and activation
Carbonisation is the first stage in the manufacture of activated carbon, and the most important,
since it involves converting the carbonaceous portion of the feedstock into a solid char ready for
activation. The quality and form of the carbonised product will dictate the final properties of the
activated carbon. Carbonisation is effected by heating the material, usually in an inert atmosphere,
to consolidate the carbon atoms and remove hydrogen and heteroatoms such as oxygen, sulphur
and nitrogen from the final product. The retention of carbon during carbonisation is desirable,
so the optimal production of H^O, HCl and H2S is required, with the minimal production of CO,
aliphatic, and sulphur-carbon compounds (Fitzer et al., 1971).
The chemical reactions always proceed in a manner which will produce the most energetically
stable molecule. For aliphatic molecules, as the chain length increases, the molecule is less stable,
but, for aromatic structures, an increase in the number of rings increases the stability of the
molecule (Fitzer et al., 1971). This implies that carbonisation will involve an increase in the
aromatic bulk of the molecule.
1.5.3.1.1 The chemistry of carbonisation
The process of carbonisation is essentially a form of pyrolysis - the chemical degradation reaction
induced by thermal energy, in an inert environment (Ericsson and Lattimer, 1989). The two major
groups of hydrocarbons: aliphatics and aromatics react in slightly different manners, although
76
the products are essentially the same. The pyrolysis reaction is subject to control by temperature,
heating rate and residence time at the pyrolysis temperature, with reactions below 700 °C being
the most important in terms of carbon yield and properties (Fitzer et al., 1971). These authors
simplified the carbonisation process into 3 stages: (i) the cracking of aliphatic molecules, (ii)
aromatisation of hydrocarbon chains and (iii) aromatic condensation into large PAH systems.
(a) Aliphatic compounds
These molecules react to produce aromatic systems. The large chains break up producing lighter
olefins, for example:
C7H16 -»(Pyrolysis 700 °C)-^ H2C=CH2 + HjC^CH-CHj + butenes 4- 1,3-butadiene
These molecules then react to produce cyclic alkenes by reactions such as the Diels Alder reaction
and finally aromatise via de-hydrogenation reactions (Sakai, 1983).
(b) Aromatic compounds
The initial reactions require the fission of C-H and C-C bonds giving free radicals which induce
molecular rearrangement and polymerisation reactions. Hydrogen is eliminated by these reactions,
often through internal transfer. Unstable non-aromatic rings are eliminated by molecular
rearrangements, which are very efficient and frequently involve no loss of carbon. The resulting
molecular species coalesce forming larger polymeric aromatic structures (Lewis, 1982), which
have been described as a collection of molecular solids, containing carbon structural units of 10-
12 condensed aromatic rings (Marchand, 1986). For efficient carbonisation, the aromatic fraction
must retain some mobility, which at atmospheric pressure generally occurs in the temperature
range 300-500 °C (Fitzer et <2/., 1971). The large aromatic structures of coals and chars are
frequently characterised by the high number of stable radicals which can be detected in the
structures (Mrozowski, 1988; Lewis, 1987). This would appear to confirm that most carbonisation
reactions proceed via the free-radical pathway.
The carbonised product is normally very disordered, with numerous structural units arranged
randomly forming channels and cavities, most of which are blocked by the products of
carbonisation, which include tars (Hucknall, 1985; SmRek and Cern^, 1970). The required surface
area and porosity desirable in an activated carbon must be produced by activation by removing
the carbonisation products from the cavities and the channels of the char.
77
1.5.3.1.2 The chemistry of activation
Activation of carbon requires the use of an aggressive agent capable of burning out the porosity
of the carbon and producing a large surface area. To do this two means are used: physical and
chemical activation. Each method is applied in a slightly different manner, and the choice of
activating agent will depend upon the nature of the carbon feedstock.
(a) Physical activation
By heating the carbonisation product in the stream of a reactive gas, at temperatures usually in
excess of 800 °C, physical activation is achieved. The activation agent reacts with the sample
which is eroded, with removal of tar from the pores and the erosion of the edge groups of the
sample. The porosity of the sample and an extensive surface area is developed. The major
physical activating agents are: COj, air and steam. This type of activation is often referred to as
gasification (Wigmans, 1989). The major recognised reactions occurring during physical
activation are:
C + HjO CO2 + H2 AH = +117 kJ/mol
C + CO2 ^ 2C0 AH = + 159 kJ/mol
The products from the use of these activating agents do vary, because CO2 is limited in terms of
its diffusion rate into the carbon, accessibility to the microporous region of the carbon and
reaction rate, which is much lower than steam. Hence, steam carbons are more microporous
(Wigmans, 1989).
(b) Chemical activation
By the application of chemicals to the carbon feedstock before or after carbonisation, the porosity,
surface area and adsorption capability of the activated carbon will be developed. The mechanism
of action for chemical activation is not as well understood as that for physical activation. The
number of possible reactions are large and the chemicals may not remain in their initial form
throughout the reaction. There are far more potential chemical activants than physical activants.
Activating chemicals include: ZnCI^, H3PO4 and H2SO4. The majority of activating agents are
characterised by their ability to dehydrate the carbon feedstock, enhance the aromatising reactions
of carbonisation and retard tar formation so preventing pore blocking (Bansal et al., 1988).
Chemical activation occurs at much lower temperatures than physical activation (500-600 °C), and
enables carbonisation and activation to occur simultaneously. The porous structure of chemically
activated carbons is dependent upon the extent of chemical impregnation of the feedstock. The
78
overall control of porosity during chemical activation is not as effective as in physical activation.
1.5.4 The surface characteristics of activated carbons
The previous discussions have established that activated carbons are porous systems which possess
large internal surface areas, but these features, although important, do not wholly account for the
adsorption properties of activated carbons. The chemical structure of the surface of activated
carbon also plays an important r61e in determining its adsorption characteristics (Cookson, 1978;
Bansal et al., 1988).
It has been shown that activated carbon is comprised of collections of carbon atoms in graphite-
like structures (microcrystallites) which are known to give weak X-ray reflections (section 1.5.2).
These structures invariably gives rise to a large number of edge atoms, where the basal planes
of the graphitic structures are fractured or the small regions of graphite type order ceases
(Mattson et al., 1969). These edge positions also tend to be populated by heteroatoms (Bansal et
al., 1988; Cookson, 1978; Donnet, 1968). The nett effect of these features is that on edges there
are un-paired electrons which will affect the electron cloud arrangement on the carbon surfaces.
Hence, there will be a tendency for oxygen to bond to the edges forming oxide structures which
affect the pH behaviour of the carbon in solution.
Steenberg (1944) proposed the "H" and "L" classification for activated carbons with either alkali
surface groups or acid surface groups respectively. The temperature of preparation and the means
of activation both affect the nature of the surface groups. The actual composition of the feedstock
was considered a minor contributor to the surface oxide structure / pH nature of the final product.
The "H" carbons are formed by firstly heating the carbon in an inert atmosphere to remove any
surface oxygen, then cooling to ambient temperatures and exposing to the atmosphere or by
preparing carbons at temperatures in excess of approximately 700 °C (Bansal et al., 1988). "L"
carbons are formed by exposing carbon surfaces to air at a temperature between 200 °C and
600 °C or by using an oxidising agent under moist conditions.
The elucidation of the nature of the structure of the surface groups has received much attention.
Most research has been focused upon the acid groups on activated carbons which have proven to
be the most readily detected and identified. Cookson (1978) detailed the seven groups most
frequently considered to be present upon the carbon surface, which are shown in figure 1.17.
The acid groups (carboxylic) are considered to impart a hydrophilic nature to the carbon and thus
increase the affinity of the carbon for polar substances whilst reducing the affinity of the carbon
79
for non-polar (for example; aromatic) compounds, where as the carbonyl groups can enhance the
aromatic adsorption by promoting formation of donor-acceptor complexes (Cookson, 1978).
Overall, the adsorption process exhibited by activated carbons will be a combination of physical
adsorption processes (van der Waals forces and hydrogen bonding), donor-acceptor complexes
and direct chemical interactions. Further discussion of this aspect of carbon science is given in
the reviews by Boehm (1966), Donnet (1968) and Bansal et al. (1988).
Figure 1.17 Acidic oxide structures found on the surface of activated carbon
OH
Carboxylic group Phenolic hydroxy group
Normal lactone
Quinone-type carbonyl group Flourescein-type lactone
group
\ / r 1
Normal lactone group
cyclic peroxide group
Carboxylic acid anhydride group
1.5.5 Applications of activated carbons
All the applications of activated carbon rely upon its adsorption characteristics. Although activated
carbon has a history of usage dating back 3500 years, the usage which it is most well known for
is as a water purification agent, for both organic and inorganic pollutants, despite the fact that its
earliest industrial application was in sugar refining. Table 1.12 shows the usage of activated
carbon based upon 1972 figures.
Other uses for activated carbon include;
Medicinal applications - treatment of kidney disease (Cameron et al., 1977)
- treatment for paraquat poisoning (Kitakouji et al., 1989)
- remedy for gastro-enteritis.
Refrigeration - use as the adsorbent in adsorption cycle refrigeration systems
80
(Critoph, 1989)
Further uses of activated carbon, including recovery of gold from ores, flue-gas desulphurisation
and radioactivity emission control are described in detail in the reviews by Juntgen (1977),
Holden (1982) and Bansal et al. (1988).
Table 1.12 Activated carbon usage (1972)
Use % of production Carbon form
Sugar refining 30 powdered
Water treatment 23 powdered / granular
Industrial gases treatment (including solvent recovery)
12 granular
Petrol vapour adsorption 7 granular
Pharmaceuticals 5 powdered
Oil and fat refining/food applications
4 powdered
Miscellaneous 19 powdered / granular (After Table 2, Juntgen, 1977)
1.6 Problematic aqueous waste streams
Most of the applications of activated carbons are well established and understood, particularly
where organic species are involved in the adsorption processes. However, for the adsorption of
metals by activated carbon, which are commonly encountered as charged species, much less
research has been performed, whilst the understanding of the toxic effects of many metals has
much increased.
1.6.1 Metal species in the environment
The introduction of metals into the environment has become an issue of increasing concern. The
behaviour of metallic species is extremely complex, with factors such as surrounding pH,
neighbouring anions and cations (ionic strength) and redox conditions affecting their speciation
and solubility. Many metals can complex irreversibly with bio-molecules with fatal consequences.
The relative toxicity of metals to mammals has been identified as follows (in decreasing toxicity);
Ag, Hg, Tl, Cd > Cu, Pb, Co, Sn, Be > In, Ba > Mn, Zn, Ni, Fe, Cr > Y, La, > Sr, Sc
> Cs, Li, Al (Evans, 1989). Sources of metals include metal plating and processing operations
and leaching from landfills. The latter source has been recognised and subsequently addressed in
the various EC directives on water quality considered previously (section 1.2.3.2.1) and the draft
81
EC Landfill Directive.
When soils and metal species interact, three effects can occur: Cation Exchange Capacity (CEC,
outer-sphere complexes - weak hydrogen bonds, physical adsorption), inner sphere complexes
(strong covalent bonds, chemisorption) and precipitation of new phases onto the mineral surface,
especially oxides, hydroxides, oxyhydroxides and carbonates (Evans, 1989). The extent of soil-
uptake is affected by the concentration of the metal species in the soil solution. pH affects the
surface charges of oxide and (oxy)hydroxide minerals and certain functional groups of some
minerals. Inner sphere complexes are readily formed by transition and rare earth metal ions and
Hg(II) and Pb(II) which all form readily hydrolysable ions. The oxide, hydroxide and
oxyhydroxides of Al, Fe and Mn in particular are effective scavengers of metal ions. Soil organic
matter contains many donor atoms (such as O, N and S) which are retained as basic groupings.
Sulphur is particularly effective at retaining soft lewis acids such as Hg(II) and Cd(II). Mineral
phases which may form due to precipitation reactions include hydroxides of Fe(III), A1(III),
Cu(II), Zn(II) and Cd(II) and carbonates of Ca(II), Zn(II), Cd(II) and Pb(II) (Evans, 1989). Soils
often contain clay minerals which are known to develop negative surface charges, which enables
them to adsorb cations from the surrounding systems. Similarly, the humic acids present in soil
contain acid groups which dissociate in the soil/pore waters producing negative charges
(Skeffington, 1987).
Similarly, carbons have surface charges which are dependent upon the pH of the surrounding
environment. The formation of inner and outer sphere complexes will depend upon the carbon
type (H or L) and the energy involved. L-type carbons, once hydrated, are the most suitable for
adsorption of metal species because they possess a negative surface charge (Huang, 1978).
Corapcioglu and Huang (1987), in their investigation of the uptake of Cu, Ni, Pb and Zn by
selected activated carbons, stated that hydrogen bonds rather than covalent bonds are more likely
to form due to the energy difference between the two processes (2-4 kcal/mol vs 50-100 kcal/mol
respectively). They also stated that it was unlikely that simple electrostatic interaction played a
major part in the adsorption of these metals due to the level of the free energy of bonding
determined from their experiments, which was indicative of hydrogen bonding.
1.6.2 Sources of metal effluents
(i) Landnil sites
Landfill sites are recognised to undergo five distinct chemical phases which are illustrated in
82
figure 1.18. Stage I occurs during waste placement. Stage II is the onset of anaerobic conditions
with the formation of carboxylic fatty acids and acidic leachates. Stage III is a consolidation of
anaerobic conditions, which is accompanied by an increase in the organic components of the
leachate and the production of shorter chain acids, esters and a fall in pH, there may be an
increased occurrence of metal species in the leachate as the acid conditions dissolve or complex
metal ions. Stage IV is more commonly called methanogenesis, as carboxylic acids are consumed
and CH4 production occurs with a corresponding rise in pH, the metals will tend to re-precipitate.
At stage V, the landfill returns to aerobic conditions, CH4 is no longer produced and the major
organic components are similar to humic acid species (Phillips et al., 1994).
Figure 1.18 The five stages in the chemical life of a landfill
P h a s e 100
Settlement
landfill-gas production
10
— 6
X a
— 2
L- 0 Time-
A review of the available landfill leachate data was undertaken with the objective of compiling
a data set representative of landfill leachate quality within all classes of landfill. Landfill leachate
composition is affected by: the nature of the waste, climatic conditions, age of the site and
properties of the soil surrounding the site (Chan et al., 1978). Results of this review are
summarised in table 1.13. These data represent information from both hazardous waste landfills
and domestic landfills operated as mono-disposal or co-disposal sites and a variety of site ages.
The severity of the pollution which can result from poor management or construction of these
sites is clearly indicated by the maximum values for each parameter determined for these
leachates.
83
Table 1.13 Summary of literature data on landfill leachate quality
Parameter Average value Minimum value Maximum value
pH 6.76 5.4 8.60
TOC (mg/1) 1788.60 8.00 8000
BOD (mg/1) 4154.86 48.00 24500.00
COD (mg/1) 11868.94 110.00 56373.00
Temperature (°C) 17.20 13.80 22.00
Cd (ppm) 5.48 0.00 107.00
Cr (total ppm) 1.50 0.00 5.52
Cu (ppm) 14.13 0.01 268.00
Hg (ppm) 73.14 0.01 1264.00
Ni (ppm) 8.25 0.00 79.00
Pb (ppm) 0.26 0.00 1.14
Compiled from information presented by; Copa and Meidl (1986); Johansen and Carlson (1976); Keenan et al., (1983 & 1984); Lavine and Rear (1989); Lema et al. (1988); Liu et al. (1991); McDougall et al. (1980); McLellan and Rock (1988); Robinson and Luo (1991): Robinson and Maris (1985) and Ross (1990).
(ii) Metal plating effluents
The effluents produced by metal plating facilities are heterogeneous, containing both metals and
organic species. Plating operations have very high water requirements hence they produce large
volumes of effluent. There are many methods available for treating these effluents, including;
adsorption, biological operations, cementation, coagulation/flocculation, complexation,
electrochemical operations, evaporation, filtration, flotation, membrane processes and solvent
extraction, the most commonly applied procedure is chemical precipitation using either sulphide,
hydroxide or a combined treatment (Ku and Peters, 1987). However, although precipitation is
very effective in reducing effluent metal-concentrations, very low concentrations of metals tend
to remain in solution, hence further purification of the effluent may be required before
discharging the liquors. Typical concentration limits for metal emissions to municipal drains is
illustrated by table 1.14. Further reductions in the metal concentrations to meet these or similar
discharge standards is costly. One of the most effective methods for achieving reductions in heavy
metal concentrations is to use activated carbons, but the effectiveness of this adsorption can be
affected by the pH of the solution, physical and chemical characteristics of the carbon and the
carbon type (Ku and Peters, 1987).
84
Table 1.14 USEPA effluent guidelines for electroplating plants discharging k 37.85 (10 000 gallons) per day into a municipal sewerage system
Pollutant Pretreatment standard (mg/1) Pollutant
Maximum for any one day Four day average
CN" Total) 1.9 1.0
Cu 4.5 2.7
Ni 4.1 2.6
Cr 7.0 4.0
Zn 4.2 2.6
Pb 0.6 0.4
Cd 1.2 0.7
Total metals 10.5 8.8 (Table 1; Amy and Petersen, 1981) .
1.7 Summary
The disposal of waste materials to land has made contaminated soil an unavoidable feature of
industrialisation. The preceding review of the problems of contaminated land, its regulation,
treatment and reuse has shown that a satisfactory solution is not a simple one. The consequence
of the decline in traditional heavy engineering and the coal-based chemical feedstock industry
resulted in large areas of land becoming disused. However, the historical trend of not regulating
the disposal of waste materials to land unlike air and water, and the recent public health incidents
involving contaminated sites have demonstrated that this policy was flawed. To confound the
situation, a lack of governmental guidance regarding the responsibility for contaminated land
remediation and clean-up standards resulted in the uncertainty which exists regarding the value
of contaminated land and ultimately the liability for the contamination.
The consequence of this uncertainty has had two main effects, both of which are of concern. The
first is a reluctance to re-develop contaminated sites, with greenfield sites being preferred and the
second is the approach which is taken when remediating contaminated sites. Techniques which
destroy or nullify the contamination are very rarely utilised. Low-technology treatments are
preferred such as excavation of the contamination with disposal into a landfill, but this means that
the contaminated material is just moved else-where, and will need to be dealt with by future
generations. Consequently, contaminated land is still being created and existing problem sites are
not being addressed in an sustainable manner.
85
Although there has been research performed in the areas of assessing and treating contaminated
land, much more research is required, particul^ly with respect to means of enabling reuse of the
soils after treatment. The heterogeneous nature of soil has complicated the assessment of the
behaviour of particular pollutants, which are themselves very often in a complex mixture, so that
development of effective soil treatment processes have been often thwarted by the soil nature.
Consequently, the most widely used contaminated soil treatment processes have been thermal
treatment based techniques. These treatments unfortunately tend to destroy the beneficial
components of the soil. The resulting sterile product cannot be considered to be soil, and
consequently has no use except as inert fill or may be disposed of to a controlled waste landfill.
There have been no reported applications of thermal processing of contaminated soil, which
resulted in a product which had potential for resale, not as a soil but as a different product.
The demand for activated carbon is continuously increasing, with the development of much more
strict emission limits for aqueous and gaseous emissions into the environment. Similarly, the use
of traditional feedstock for carbon manufacture, especially non-renewable sources such as peat
and coal cannot be sustained, particularly as peat bogs are becoming a scarce resource with a
value as a natural habitat for flora and fauna. Hence, there is an immediate requirement to
research and develop novel feedstock, which can replace the non-renewable materials.
Application of activated carbon technology to contaminated soil as a means of converting organic
contaminants into a product with adsorbent properties has several advantages which have not been
considered previously. Existing means of treating soil pollution such as incineration, chemical
treatments or solidification must receive approval by appropriate regulatory bodies, and perhaps
most importantly, overcome the increasing public concern for untried or new processes. On both
these accounts, activated carbon manufacturing has distinct advantages, being an established
process and in receipt of approval from regulatory bodies (HMIP, 1992).
Hence, contaminated land shall be investigated as a potential feedstock for activated carbon
manufacture. The gas manufacturing industry has contributed more contaminated sites to the UK
inventory than any other industry. Soils from these sites are a complex mixtures of different
chemical wastes, which are reported to be impossible to treat in a single process. Consequently,
the objectives of the work described in this thesis are to demonstrate that contaminated soil from
derelict gaswork sites can be converted into carbonaceous adsorbents through the application of
activated carbon manufacturing techniques, with all the contamination (organic and inorganic)
being ameliorated simultaneously. The resulting product shall be chemically and physically
characterised and subsequently investigated for its suitability for adsorbing organic and metal
86
species from aqueous systems. Finally, the potential of the carbons for treating two reportedly
problematic waste streams: landfill leachates and the aqueous waste stream from metal plating
operations shall be investigated.
87
CHAPTER TWO
EXPERIMENTAL RATIONALE
Examination of the literature has shown that the potential for production of carbonaceous
adsorbents from contaminated soil has not previously been investigated.
The investigation was thus approached in the following manner:
# Characterisation of the chemical and thermal properties of the soil samples available for
this study, to establish levels of contamination and the suitability of the soil for activated
carbon manufacture.
# A survey of readily available and accepted activated carbon manufacturing methods, to
establish which was the most effective activation method for thermally converting the soil
into carbonaceous adsorbents.
# For the most effective activation method, the process parameters which affect the
adsorbent properties of the product were optimised.
# Complete investigation of the adsorption properties of the optimised carbons through:
• Organic adsorption
• Metal adsorption
• Nitrogen gas adsorption
# Investigation of potential applications of the soil-adsorbent in terms of:
• Wastewater effluent treatment
• Control of leachate within landfills
The experimental work was performed in stages, in such a manner that each set of experiments
produced results and conclusions which provided the information necessary for the development
of the next experimental programme. This rationale is reflected in the layout of the subsequent
chapters in which experimental results and related discussion on specific themes are brought
together to set the context for the following chapter.
88
CHAPTER THREE
SOIL CHARACTERISATION AND SELECTION OF AN ACTIVATION PROCEDURE
3.1 Introduction
This study has been limited to considering gaswork contaminated soil only (see section 1.3.3).
In total, six samples from two gaswork sites were utilised. The initial carbonisation study,
activation agent selection and organic adsorption experiments were performed upon five samples
called STA01-STA05 taken from site I. The optimisation of the activation procedure, application
development and demonstration discussed in chapters 4 and 5 were performed using the sample
G1 which came from site II. Although the soils were sampled at different times (12 months
apart), their characterisation will be discussed simultaneously, to facilitate the comparison of
results.
3.1.1 Nomenclature for sample identification
There were six soil samples studied during this research, their origin and description are given
below and in table 3.1 and photographs of the as sampled soils are shown in plates 3.1 to 3.6.
Site I samples: STAOl - STA05 originated from a gaswork site in Hertfordshire which was
undergoing extensive restoration work.
Site II sample: G1 was taken from a derelict gaswork site in Central London.
Table 3.1 Description of the soil samples.
Identifier Appearance and Odour
STAOl Deep blue sample. A pungent 'coal tar' odour.
STA02 Black sample. Virtually no odour.
STA03 A very stony sample. Little odour and very blue.
STA04 Brown sample. A creosote smell.
STA05 Brown, oily sample. Very pungent creosote odour.
G1 Black sample. Coal tar odour.
89
Plate 3.1 Soil STAOl
STAOl
Plate 3.2 Soil STA02
STAOl
90
Plate 3.3 Soil STA03
STA03 L-.l'
Plate 3.4 Soil STA04
STA04
91
Plate 3.5 Soil STA05
STA05
%
Plate 3.6 Soil G1
G1 (As SainplcUk
92
3.2 Experimental protocols
3.2.1 Introduction
Throughout the experimental work of this research, standard experimental protocols were adopted
with respect to apparatus cleaning, reagent preparation and usage. All apparatus was soaked
overnight in a 2.5 % v/v Decon® solution then thoroughly rinsed using tap water and finally
distilled water. When performing metal analysis, the apparatus was subsequently soaked overnight
in a 10 % HNO3 solution (v/v) and then rinsed thoroughly with distilled water. All chemicals
used were supplied by either BDH Ltd. or Fisons pic. The grade of chemical used depended upon
the analytical accuracy required for the experiment.
3.2.2 Soil sampling, post-treatment and fractionation
All samples were kept refrigerated at 4 °C until used. Prior to analysis, each sample was air-dried
at RTF to constant weight. The dried fraction tended to agglomerate, resulting in the large
particles being coated in finer materials. It has been recognised that when a soil is contaminated,
most of the pollution is associated with the adsorbent clay and silt fractions (BSI, 1990) which
comprise the smallest fraction of soil (the material < 60 nm (White, 1987)). The method adopted
to fractionate the soil, by successive crushing and sieving to loosen the small particles adhering
to the oversize particles, reflected these findings. The samples were broken-up in a pestle and
mortar and sieved to pass a BSI Standard 2 mm mesh (BSI, 1990; BGAM, 1984). The over-size
material was returned to the mortar and broken further, without excessive force, and re-sieved.
This procedure was repeated until only the rocks and rubble, which would require large blows
to split, remained. Representative portions of the < 2 mm fractions were taken and further ground
to pass a 150/im mesh sieve. This reduced particle size fraction was used for the physical and
chemical characterisation of the soil.
Drying the soil at higher temperatures than ambient could result in loss of organic contamination
which would be undesirable for the process under investigation. However, air-drying of the soil
to constant weight was a long process, typically taking 1 month for a stable weight to be
achieved. To investigate whether the drying process could be accelerated by using higher
temperatures, small portions of soils STAOl - STA05 were also dried, in duplicate, at 50, 80 and
110 °C. These samples were then analyzed for their carbon content. Sample G1 was also
subjected to a complete particle-size analysis. A 4 kg sample of G1 was air dried and then sieved
by shaking (not crushing) using 12 sieve sizes, covering the particle size 16 mm to 75 ^m.
93
3.2.3 Chemical characterisation of the soil
3.2.3.1 Introduction
The major pollutants associated with gaswork soils were considered earlier (section 1.3.4). For
the proposed process to be effective and accepted for treating contaminated soil from gaswork and
similar sites, the samples needed to contain sufficient carbon components (PAHs, tars, phenols,
oils, coal and coke residues) for the formation of the required carbon structure. Carbon,
Hydrogen and Nitrogen (CHN) analysis were performed on all the soil samples. The major
pollutants of concern: total cyanide, total sulphate and free sulphur contents of each sample were
also determined, as were the metal contents of the soils. All the analyses were performed upon
the < 150 nm sample discussed in section 3.2.2.
3.2.3.2 CHN Analysis
CHN analysis of the raw soils was performed using a Lehmann 440 CHN analyzer (Butterworths
Laboratories) which was calibrated using acetanilide (CH^CONHC^H;).
3.2.3.3 Total cyanide, total sulphate and free sulphur analysis
The total cyanide, total sulphate and free sulphur content of each soil sample was performed by
Environmental Analysis Ltd. (St Leonard-on-Sea). The cyanide was determined by the
Chloroamine-T-hydrate/pyridine-barbituric acid colourimetric method (APHA, 1992). Total
sulphate was measured by acid extraction, then precipitating as BaSO^ and weighing the dried
precipitate (BSI, 1990). Free sulphur was extracted by acetone and exchanged by water, with the
turbidity of the solution providing the sulphur concentration.
3.2.3.4 Metal analysis
The metals in the soils were determined by Inductively Coupled Plasma-Atomic Emission
Spectrophotometery (ICP-AES) using an Applied Research Laboratories 34000 ICP. All the
metals of interest, except Hg, were determined by the in-house ICP-AES method "GENS" (Dept.
of Geology, Imperial College) which checked for 25 elements (Li, Na, K, Rb, Be, Mg, Ca, Sr,
Ba, Al, La, Ti, V, Cr, Mo, Mn, Fe, Co, Ni, Cu, Ag, Zn, Cd, Pb and P). Mercury was analyzed
separately, also by ICP-AES. Only 11 of the GEN5 elements (and Hg) were considered to be of
interest to this work. These elements have been recognised as deleterious to human health by the
94
various regulation bodies which have produced the contamination classification data detailed in
Appendix I. Samples were always digested in duplicate. A standard reference material, BCR143,
was included among the samples to check the efficiency of the digestion. Blank samples and
spikes were routinely included in all digestion procedures. The solutions used to spike selected
samples were those used to check the instrument calibration and consisted of a high and low
concentration solution. The digestion procedures used for solubilising the metals were a USEPA
method and an in-house method, which are discussed below.
3.2.3.4.1 Acid digestion by method 3050
Approximately 1 g of soil (dry, < 150 /im) was accurately weighed into a 150 ml acid-washed
beaker. 10 ml of 1:1 HNO3 (PrimaR®) was added and the mix slurried then covered by a watch-
glass. The sample was heated to 95 °C and refluxed, without boiling, for 10 minutes. To the
cooled sample, 5 ml of conc. HNO3 (PrimaR®) was added and then further refluxed for
30 minutes. After cooling, 2 ml of distilled water and 3 ml of 30 % (PrimaR®) were added,
followed by warming. Once effervescence had subsided, further 1 ml aliquots of were added
until no obvious change in the sample occurred. This often resulted in the maximum
reconunended volume of 10 ml of being added. The sample was again cooled, then 5 ml
of conc. HCl (PrimaR®) and 10 ml of distilled H^O were added and the sample heated for
15 minutes. The cold sample was transferred to a 100 ml acid-washed volumetric flask and diluted
to the 100 ml mark. For early digestions, particulates remaining in the solution were removed by
filtration through Whatman® N° 42 filter paper, but this was found to be an unsatisfactory
procedure. Later samples were centrifuged at 2000 rpm for 5 minutes to remove the particulates.
The final sample was in a 5 % HCl and 5 % HNQ, matrix (USEPA, 1986,1991b).
3.2.3.4.2 Add Digestion for Hg
To approximately 1 g accurately weighed of dried soil (< 150 /xm) in a 150 ml acid-washed
beaker, 2 ml of conc. HNO3 (PrimaR®) was added. The slurry was heated at 80 °C for 1 hour.
After cooling, the samples were transferred to a 100 ml volumetric flask and diluted up to the
100 ml line (Imperial College, 1991).
3.2.4 Screening of soils as potential activated carbon feedstock
3.2.4.1 Introduction
The activated carbon preparation requires the carbonisation and activation of the feedstock under
95
controlled thermal conditions (section 1.5.3). The only means for studying the thermal behaviour
of a material is the well established technique of thermal analysis. Various factors can affect the
carbon product, including: a) the carbon content of the feedstock; b) the carbonisation
temperature; c) the activation temperature; d) the heating rate; e) the residence time at the
carbonisation and activation temperatures; f) the form of activation used and g) the impregnation
ratio (Kirubakaran et al., 1991). It is important to study each factor as an isolated parameter. The
information on points b - d are provided by thermal analysis.
3.2.4.2 Thermal analysis
Thermal analysis is a generic name for several methodologies which have arisen Arom the basic
technique of measuring the weight change of a solid as the temperature increases. There were
three distinct methods of interest to this work: (i) Dynamic Thermogravimetry (TG); (ii)
Differential Thermal Analysis (DTA) and (iii) Derivative Thermogravimetry (DTG). Two
thermobalance instruments were utilised during this work. Both were Stanton Redcroft
Simultaneous Thermal Analyzers, which determined TG/DTG/DTA from the same sample. The
only major difference between the two instruments was the manner of collection for the DTG.
The earlier model, an STA780, (used for the work involving soils STAOl - STA05) measured the
DTG by direct determination whilst the later model, an STA1500, (used for soil Gl) employed
the instrument software to calculate the first derivative of the TG. Figure 3.1 shows the layout
of the STA instrument.
The thermal analysis of the soils was split into several parts:
1) Characterisation of the thermal response of the contaminated soils to heating in an inert
atmosphere - Nitrogen.
2) Studying the effect of selected activation procedures (chemical and physical) upon the
thermal response of the soils.
3) Evaluating the time for completion of carbonisation and activation.
The < 150 ixm particle-size soil samples were used for the TG/DTG/DTA analysis. The soil was
contained in a fired quartz crucible. A second crucible, containing a similar weight of fired
alumina (AljOj) was used as the reference for the DTA measurement. The experimental
parameters were kept constant for all the measurements, the values used for each analysis are
shown in table 3.2.
96
Figure 3.1 A schematic of the STA 1500 simultaneous thermal analyzer
ELECTRONIC MICROBALANCE DTA
TG
VACUUM CONNECTION (RECORDER
ONLY) 4 SIDE BRANCH FOR BLANKET GAS
CONTROL THERMOCOUPLE WATER COOLED
FURNACE
FURNACE WINDING
DTA HEAD IN . ^ MICRO-ENVIRONMENT CUP
- O WATER IN
•0 WATER OUT
,CONTROL ^ VALVE
GAS GAS
FLOW METER
DC AMPLIFIER
DATA ACQUISITION
BALANCE CONTROL UNIT
TEMPERATURE PROGRAMMING
FACILITY
GAS IN
GAS OUT
(Figure MM 3.2, Stanton Redcroft, 1987)
Table 3.2 Parameters used for thermal analysis of soil samples
Parameter Setting
Gas flow rate 50 ml/min
Sample weight 20 mg +/- 2 mg
Heating rate 5 °C/min
DTA range 100 /iV
DTG range 4000 /ig/min
Previous work on spent bleaching earth (a clay loaded with edible oil residues which could be
likened to coal tar contaminated soil) showed that the lower the heating rate, the better the final
carbon product (Pollard et al., 1991a,b). They found that 5 °C/min afforded the best results,
hence this heating rate was used throughout this study for the preparation of the samples. A slow
97
heating rate gives better control over the reaction and because of the variable nature of the
carbon content of the contaminated soil (15-50 % carbon, Table 3.3), carbon loss due to
volatilisation must be minimised.
3.2.5 Activation agent selection
There have been numerous activation agents applied to many different carbonaceous feedstock to
produce activated carbon. When considering the activating agents for study in this work, there
was a requirement to use those with a history of commercial usage and were in ready supply. The
agents chosen were: a) Physical activating agent: CO2
b) Chemical activating agents: H3PO4; ZnCl;; H2SO4 and HNO3
For this part of the experimental protocol, only STA05 was studied, as it was the soil sample
which consisted of the greatest bulk (approx 8 kg). The other samples, STAOl - STA04, were
considerably smaller samples of only 2-3 kg each and required to be conserved for later
investigation. It must be noted that at this stage, G1 had not been procured.
3.2.5.1 Physical activating agents
Of the physical activation agents available only CO2 was studied. The other possible activating
agents, air and steam, were not used because: i) both agents are particularly reactive towards
carbon, oxidising it very rapidly. In view of the low carbon content of the soil (see table 3.3),
carbon retention was paramount and ii) suitable facilities for steam manufacture were unavailable.
For the TG/DTG/DTA study of CO; activation, the instrument was set up as described in section
3.2.4.2. and the sample firstly carbonised in Nj at 175 °C, a temperature which was established
from the earlier thermal study of the soil alone (see figure 3.11). The gas was subsequently
changed to CO; and the heating continued up to 1400 °C.
3.2.5.2 Chemical activating agents
Chemical activation, unlike physical activation, requires the mixing of feedstock and activant. A
fixed impregnation ratio of 1:0.5 was used for all the activation trials. Pollard etal. (1991b),
established that for their clay and oil waste, the ratio of waste to chemical activating agent which
produced the optimum sample was 1:0.5 (33 wt % impregnation). Kirubakaran et al. (1991)
also reported similar findings for coconut shell activation, when an impregnation ratio of 25 -
50 wt % produced similar adsorption characteristics in the products. In the present study, using
the same impregnation ratio enabled the direct comparison between the products of different
98
activation routes.
The H2SO4 (AnalaR®) and H3PO4 (GPR) were initially added as a concentrated solution (Bansal
et al., 1988), but, the effervescent reaction which ensued (HjS production), caused the mixture
to over-flow the mixing bowl. These activation agents were subsequently diluted and mixed with
the soil as a 50 % (^/w) solution, which reduced the vigour of reaction. Zinc chloride (GPR), as
a 25 % saturated solution (ZnClj water solubility = 4.32 g/ml, (Weast, 1987)), was added to a
known soil weight to give a soihZnClj ratio of 1:0.5 (SmKek and Cern;^, 1970).
Concentrated HNO3 (AnalaR*) was directly mixed into the soil. Excess water was evaporated
from the resulting soil/chemical activant slurry in a ventilated oven at 35 °C until the mix was
sufficiently solid to hold its shape after stirring. The drying was necessary to facilitate ease of
handling of the feedstock. The dried, impregnated samples were thermally analyzed, using
nitrogen as the atmosphere.
3.2.6 Laboratory scale production of carbon adsorbents from STAGS
From the information gained in the thermal analysis of STA05, large samples of each carbon were
produced to allow the evaluation of the different activation agents and the choice of one agent
which produced the best carbon as dictated by the adsorption and surface area characteristic of
the product. All of the carbons produced in this study were done so using a rotary furnace, a
Carbolite HTR 11/150.
3.2.6.1 The carbolite HTR 11/150 rotary furnace
A schematic of the furnace is shown in figure 3.2. This equipment consist of a large "Christmas
cracker" shaped reaction-vessel. The sample is loaded into the vessel, which is suspended by air-
tight rotary fixtures in a box-furnace. The atmosphere in the vessel can be varied, as can the
heating and rate of rotation of the vessel. For all experiments, the rate of vessel rotation was fixed
at 10 rpm, which was judged to be suitable to mix the sample and ensure even "cooking" of the
sample, without promoting excessive volatilisation or attrition of the sample.
3.2.6.2 Production of carbon from contaminated soil
A standard protocol for processing the samples was applied. Once loaded into the furnace, the
sample was mixed for 1 hour at 25 °C, with the selected gas entering the system at 500 ml/min.
The furnace was then heated at 5 °C/min to the carbonisation temperature established by thermal
99
analysis. It was held there for 2 hours, which from examination of the TG/DTA/DTG trace was
judged to be sufficient time for completion of carbonisation. Heating again at 5 °C/min, the
activation temperature was attained and held for 1 hour, which, from inspection of the
TG/DTA/DTG information, was Judged to be sufficient time for activation to be completed. The
sample was cooled overnight, under the reaction atmosphere.
Figure 3.2 The Carbolite HTR 11/150 rotary furnace
Furnace
i
Gas supply Gas IS
Gas flow control
Sample
Silica vessel
Rotation
# • Tb off-gan freatment and (umcxupboard extract.
3.2.6.3 Post-activation treatment of the product
A standard procedure was adopted throughout the manufacture of all the soil carbons, using a
2 M HCl wash at the end of the run. The acid wash is a routine procedure for carbons
manufacture from ZnClj because it allows for the recovery of the un-reacted ZnCI^ which could
then be recycled into the process (SmfSek and Cern^, 1970). It is also regularly used by
experimentalists to improve the adsorption characteristics of physically-activated commercial
carbons by removing ash and other loose inorganics which can block pores and obscure active
sites (for example; Nacco and Aquarone, 1978; Srivastava et al., 1987).
The product was discharged from the reactor into a 2 litre beaker and approximately 2M HCl was
added at a ratio of 1:10 soil:acid. The product and acid were heated to 60-70 °C, with constant
stirring and maintained at that temperature for 30-40 minutes. The acid was then filtered off the
product under vacuum and the product washed repeatedly with hot distilled water and then cold
distilled water to remove all traces of the acid. It was finally transferred to an evaporating dish
in which it was dried at 105 °C for up to 24 hours. The product tended to form "balls" in the
rotary reactor of diameters up to 20 mm which did cause problems when removing the sample
from the reaction vessel. McKay et al. (1985) considered that the particle size of carbon plays
100
a significant part in the adsorption exhibited by the material. Generally, the small particles exhibit
a greater adsorption than larger particles which they attributed to the inability of the adsorbate
molecules to penetrate the internal pores of the particles. Consequently, after drying, the complete
product was ground by pestle and mortar until it passed through a 150 fim BSI standard sieve.
The sample was then ready for characterisation.
3.2.7 Characterisation of soil derived activated carbons
3.2.7.1 CHN analysis
The powdered carbons were analyzed for their CHN content (Department of Chemistry, Imperial
College) using a Perkin Elmer model 2400 CHN analyzer, which was calibrated using
CHgCONHCgH).
3.2.7.2 Total cyanide and sulphate and free sulphur analysis
These were performed by Environmental Analysis Ltd., using the methods described in section
3.2.3.3.
3.2.7.3 Evaluation of aqueous adsorption ability of carbons
3.2.7.3.1 Introduction
The carbons manufactured in this work were all in the powdered form which means that they
would be used for liquid phase adsorption applications. There are several ways of evaluating the
ability of carbons to remove substances from aqueous solution, such as by iodine (Hassler, 1963),
methylene blue dye (Barton, 1987: Potgieter, 1991) or model organic compound adsorption from
solution. By far the most frequently used substances are the last group, of which the phenolic
family of compounds have proven to be extremely suitable. Phenolic compounds are one of the
most common pollutants in industrial waste water streams. They are of particular concern because
of their susceptibility to chlorination during routine disinfection processes. Industries which are
known to produce phenolic waste waters include: petrochemical, petroleum refining, coking
plants, pulp and paper mills and wood preservation plants (Knettig et al., 1986). The high
concentrations of phenolics in waste waters are usually treated by distillation, but there is still a
small level of phenols which remain in solution. Concentrations of up to 300 mg/1 (3 mM) are
not uncommon (Paprowicz, 1990).
101
Phenol (CfiHsOH) (AnalaR®) and 4-nitrophenol (GPR) (OzNC^H^OH) are both UltraViolet active
(UV). They were used to evaluate the aqueous adsorption capability of the manufactured carbons
by measuring the residual phenolic solution concentration after contact with the carbons
manufactured. The peak at 270 nm for phenol and 317 nm for 4-nitrophenol were used for the
measurement of each species. It was found to be necessary to purify the 4-nitrophenol before use
by steam distillation due to isomeric contamination.
3.2.7.3.2 Determination of adsorption ability of soil carbons
Utilising the Beer-Lambert law, calibration graphs for the UV absorption by phenol and 4-
nitrophenol were prepared using a Perkin Elmer Lambda 3 UV spectrophotometer. Phenol obeyed
Beer-Lambert at solution concentrations up to 1.28 mM and 4-nitrophenol at solution
concentrations up to 0.5 mM. Paired quartz cells were used as sample containers and a minimum
of 4 points were determined per calibration graph.
Stock solutions (10 mM concentration) of phenol and 4-nitrophenol were accurately prepared in
distilled water. To 1.00 g of soil carbon in glass 125 ml screw-top bottle was pipetted 100 +/-
0.05 ml of the adsorbate solution. The carbon/adsorbate solutions were end-over-end rotated at
RTP for 3 hours, a time which had been established as sufficient for adsorption equilibrium to
be attained from some initial adsorption studies and the literature (Pollard et al., 1991a: McKay
et al., 1986). The solution and carbon were separated by filtering through fluted Whatman® N°1
paper. The pH of the filtrate was measured (pH meter, RS Components Ltd). All samples were
measured in triplicate, as were the required control solutions. Adsorption isotherms were also
measured for each carbon, over a suitable concentration range, with the results being modelled
to both the Freundlich and Langmuir isotherm equations.
3.2.7.4 Surface area and porosity analysis by gas adsorption
3.2.7.4.1 Introduction
The importance of the surface area and porosity character of activated carbon has already been
considered (section 1.4.2 -1.4.3). All the samples were analyzed for these properties by nitrogen
adsorption, using an Coulter® Omnisorp® 100. The schematic of the system is shown in
figure 3.3. This instrument operates by a continuous volumetric technique which means that the
sample for study is exposed to a steady flow of nitrogen gas, which is regulated by a mass-flow-
controller. The rate at which nitrogen is introduced into the sample holder is so slow that it can
102
be assumed that the nitrogen adsorbed onto the sample surface and that in the sample tube
atmosphere are always in equilibrium. This quasi-constant equilibrium allows for extremely small
changes in pressure to be detected and the production of full adsorption-desorption isotherms
which can contain up to 2000 data points, compared to 200 which are measured by the traditional
"static" method.
Figure 3.3 The Coulter® Omnisorp* 100
FURNACE
HELIUM NITROGEN
V7|| ll V4
MASS FLOW CONTROLLER
V5
PI
PS (2^=1 s , MANIFOLD
VAC. GAUGE
DEWAR
^Po
RELIEF VALVE
A. SAMPLE TUBE B. REFERENCE TUBE C> SAMPLE STOPCOCK
PI, 10 TORR PRESSURE TRANSDUCER P2, 1000 TORR PR. TRANSDUCER Poi REFERENCE PR. TRANSDUCER
HIGH-VAC. PUMP (CPTIDNAL)
UMP NOTARY VAC. PUMP
(Figure 1, p 6, Coulter®, 1991)
3.2.7.4.2 Sample preparation for Omnisorp® analysis
Approximately 0.3000 g of powdered soil-carbon was accurately weighed into a sample tube and
a sample-tube stopcock was fitted. The sample and holder were placed in the furnace in position
"A" (figure 3.3) and degassed for 8 hours at 250 °C to an ultimate vacuum of 10" Torr. The
sample tube, with the stopcock closed, was transferred to the dewar, which was filled with liquid
nitrogen and placed in position "B" (figure 3.3). The instrument is computer controlled, hence,
once the sample parameters had been entered into the software and the instrument started, all data
measurements were fully automated. Data interpretation were by means of the instrument
software, which facilitated direct calculation of surface area by the BET equation (Brunauer et
103
al., 1938) and t-plot analysis (Lippens and de Boer, 1965), estimation of the mesopore size
distribution by application of the Kelvin equation based upon the method of Barrett et. al., (1951)
and limited study of the micropore distribution by the method of Horvath and Kawazoe (1983),
all of which have been discussed previously (section 1.4.2.1).
3.3 Results and Discussion
3.3.1 Characteristics of the air-dried soil
The percentage moisture, carbon, hydrogen, and nitrogen content for each soil, and the
percentage of the air dried soil which passed a 2 mm sieve are shown in table 3.3. The % carbon
value provided a measure of the amount of carbonaceous contamination present in each soil
sample but not the chemical form of that carbon. The effects of carbonates in the soil were
examined by washing the soil with dilute acid before determining the carbon content. The values
obtained agreed within 1 % of the untreated samples, hence, a contribution from carbonate to the
carbon value could be discounted for these samples. STA02 contained the most carbon, but the
behaviour of the samples under physical treatment (grinding) did indicate differences in the form
of the organic contamination. STA04 was problematic to grind below 150 /xm due to the tarry
nature of the sample but the other samples did not suffer from the same problem to any great
extent.
Table 3.3 Soil sample characteristics
Identifier % Moisture % < 2 mm % Carbon % Hydrogen % Nitrogen
STAOl 35.1 70.3 14.6 2.2 1.4
STA02 32.0 91.4 49.7 1.5 1.6
STA03 22.8 38.6 34.3 2.4 1.6
STA04 46.0 77.5 32.9 3.8 14.3
STA05 5.7 44.0 18.8 1.7 1.0
G1 18.1 73.1 12.8 1.3 0.9
In figure 3.4 and 3.5, the effect of increasing the temperature of the drying environment for the
soils upon their carbon and moisture contents are shown. It can be seen that higher temperatures
did not cause any major changes in their carbon content. Slight increases in carbon content when
drying at 110 °C compared to the lower temperatures were observed for STA02, STA03 and
STA05 (figure 3.5). This observation could be partially explained by the concurrent increase in
the % moisture content of the samples.
104
Figure 3.4 The effect of different drying temperatures upon the soil moisture content
60 r
•o 50 a
40 E a 30
1 1 20
10
0 -STADl STAD2 STAD3
Soil identity STAD4 STAD5
S50SC inDSO c raiio^
Figure 3.5
60 p
50
1 40
1 30
20
10
0 -
The effect of different drying temperatures upon the soil carbon content
STAOl STA02 STA03 Soil identity
STA04 STA05
B502C mnsosc raiiosc
The particle size distribution of sample Gl. Figure 3.6
i f r
Size fraction (mm)
of soil • Cumulative data
105
The heterogenous nature of soil will also cause differences between sub-samples of the bulk
sample. The above results indicated that increasing the temperature of the drying process would
not cause excessive loss of carbon volatiles. However, for the duration of this work, it was
decided to continue drying the soil at RTP to minimise laboratory odours.
Figure 3.6 shows both the cumulative and fractional distribution of particle-sizes for sample Gl.
The amount of soil retained on each sieve size is expressed as the % of the whole sample (by
weight). The result of just shaking the soil whilst sieving gave a very different proportion of the
soil which separated into < 2 mm and > 2 mm. Inspection of table 3.3 above, shows that
73.1 % of the soil Gl passed a 2 mm sieve, when subjected to mild crushing. However, this
value was almost reversed when just shaking was used. The particle size analysis showed that
65.3 % of the soil Gl was retained by sieves ^ 2 mm. This result verified that successive sieving
and crushing is necessary to ensure that effective separation of the smallest contaminant containing
particles are achieved for contaminated soils.
3.3.2 Total cyanide, total sulphate and free sulphur analysis
The total cyanide, total sulphate and free sulphur content of the soils are shown in table 3.4.
These results emphasised the variability of contaminated soil samples. With reference to the
contamination assessment values detailed in Appendix I, STA02, in terms of sulphur
contamination, would be classed as uncontaminated in all the countries where values are supplied.
Using the ICRCL guidelines for sulphur in former coal carbonisation sites (table A 1.5), only
STA04 breached the 'action' concentration value. Consideration of the cyanide values, however,
indicates levels of very heavy contamination (except for STA04) when using the GLC guidance
values shown in table AI.3 (Kelly, 1980). If the ICRCL parameters for cyanides associated with
coal carbonisation sites (table AI.5) are compared with the values in table 3.4 the threshold value
for both complex and free cyanides are breached by all the contaminants. Only STA04 falls below
the complex cyanide action value, (assuming that the site is to be landscaped or built upon). The
sulphate values fall astride the 50 000 ppm value. STAOl and Gl exceed the 50 000 ppm sulphate
content and would be regarded as unusually heavily contaminated (Kelly, 1980), whilst STA02 -
STA05 respectively are heavily contaminated. The ICRCL guidelines would require action upon
the sulphate present in all the soil samples before usage in domestic applications and a similiar
situation would exist for STAOl and Gl where buildings are to be erected upon the sites.
However, since each sample contains more than one parameter which is in great excess of the
'uncontaminated' or action values where appropriate, treatment or excavation of the soil prior
to redevelopment of the sites would be a pre-requisite.
106
Table 3.4 Total cyanide, total sulphate and free sulphur in soils STA01-STA05 and Gl.
Identifier total cyanide (ppm)
total sulphate (ppm)
free sulphur (ppm)
STAOl 15400 84600 2900
STA02 16840 17600 100
STA03 109800 31200 1560
STA04 2010 37200 77800
STA05 13480 14300 1040
Gl 10500 122000 3960
The analysis of the size fractions of Gl for contamination are shown in figure 3.7. Overall, the
contamination appears to be evenly distributed between the different particle fractions, at levels
representative of the values identified for the <2mm fraction, shown above in table 3.4. There
is a tendency for the contamination to reduce as the particle size increases (left to right), but this
would be expected. The fact that the amount of contamination associated with each size fraction
does not substantially decrease with increased particle size, when only shaking and sieving the
soil is used, is of interest, especially since there has been a tendency for development of soil
treatment facilities which attempt to separate the contamination from the soil by assuming that
most contamination is associated with the smallest particles. With these samples this assumption
would still leave much contamination behind. This has been recognised where particle separation
has been applied in soil decontamination, dry sieving is rarely used alone (Armishaw et al.,
1992). Addition of crushers, water and chemicals are commonly practised in the separation of
contamination from soil.
Figure 3.7 The distribution of contamination between the soil particle size fractions
I
100000
10000
1000
100 -f- -f ^ -4^
1 1-
Size fraction (mm)
EZDTotal su lphate l i F T e e s u l p h u r E^Total cyanide
107
3.3.3 Metal analysis of soil samples
The metal analysis data for soils STAOl - STA05 and G1 are given in table 3.5. Consideration
of the concentrations of the elements found in samples STAOl - STA05 and G1 show values
which are fairly typical for gasworks sites (ERL, 1987). With reference to Appendix I, only some
of the metals pose a particular contamination problem. The obvious examples are the Cd value
for STA05 (contaminated according to table AI.4), the Hg values for all the samples, especially
STA04 (unusually heavy contamination, table AI.4), Mn in STA05 (heavy contamination, table
AI.4), Pb in STA05 (contaminated, table AI.4) and Zn in STA05 (heavy contamination, table
AI.4). The presence of these individual excesses of metals means that the whole soil sample is
classed as contaminated.
Table 3.5 Metal concentrations detected in the soils STAOl - STA05 and G1
Metal Concentration detected in soil (ppm)
STAOl STA02 STA03 STA04 STA05 G1
Ba 217.7 381.1 554.3 87.1 123.9 235
Ca 11374 959 1449 2365 18562 42600
Cd nd nd nd nd 5.8 1.1
Co 1.8 nd nd nd 8.9 3.9
Cr 56.8 19.8 19.3 34.5 19.4 33.7
Cu 21.8 32.5 3.2 nd 127.9 39.3
Hg 27.9 3.1 2.3 864 15.7 4.7
Mn 615.6 193.9 355.1 152.4 3006.9 297
Mo 29.0 10.0 6.9 2.3 26.5 7.0
Ni 4.6 2.3 nd nd 14.6 14.7
Pb 986.6 458.6 299.7 33.8 1189.8 449
V 18.1 25.4 11.1 107.8 25.4 30.9
Zn 151.7 59.8 155.7 28.5 3172.6 102 nd = below instrument detection limits
The efficiency of the digestion procedure was assessed by two means: Digestion of a standard
reference material (BCR143) and the addition of known concentrations of metal mixtures (spikes).
The results from these analyses are presented in figures 3.8 and 3.9. Considering the behaviour
of BCR143, the recovery of the certified metals from this sewage-sludge amended soil are
expressed as the % difference, which was calculated thus:
108
Digestion value - Certified value % difference = x 100
Certified value
The concentration of the metals determined by ICP (digestion value) and the values supplied by
the BCR for these metals (the certified values) are shown in figure 3.8. The % difference value
allows for the efficiency and applicability of the analysis to be determined. The metals (Cd, Cu,
Pb, and Zn) were within 15 % (as indicated by the shaded region in figure 3.8) and Ni within
20 % of the certified values, but most were lower, which was judged to be acceptable. The only
metal of concern was Hg, for which the % difference was in excess of 90 %. The most probable
explanation for this discrepancy was due to the concentration of the Hg in the analyte solution
which, after taking the dilution factor (xlOO) into account, was much lower than the ICP
instrument detection limit (IDL) for mercury (1 ppm). Thus, the "background noise" of the
instrument and the matrix of the analyte solution would cause disproportionate interference with
the observed Hg signal. It is recognised that as the IDL of the instrument is approached, the
actual precision of the results obtained deteriorates severely and that the ICP signal, which is
sensitive to very small variations in the power from the R.F. supply, is easily perturbed
(Thompson and Walsh, 1989).
Figure 3.8 The effectiveness of the digestion procedure using a standard reference material, BCR 143.
C o
§ 6
I
H g N i
Metal species
•BCR143 - • - R e f e r e n c e v a l u e difference
Spike recoveries were not measured for all the elements of interest, but those which were obtained
are shown in figure 3.9. Most of the elements exhibited acceptable recoveries, again confirming
the suitability of the digestion procedures used for the samples. However, STA04 and STAOl
exhibited curious behaviour with regard to their Co, Ni & Cu spike recoveries. Both of these
samples showed that over half of the spikes were not detected. For STA04, the spike was almost
completely removed for Co, Ni, & Cu. The cause of this loss was not clear, but an indication of
109
the possible reason comes from the relative amounts of sulphur species (free sulphur and total
sulphate) in each sample. There was a definite trend in the magnitude of the spike recovery and
the amount of sulphur species in each sample, as illustrated in figure 3.10 by the decrease in spike
recovery as the sulphur and sulphate content increases. All of the above metals will react with
sulphur after gentle heating and the metal sulphide formed is almost insoluble in water, but will
dissolve to varying extent under acidic conditions. The sulphate salts of these metals are also
soluble (Weast, 1987). However, in a review of the properties of metal sulphides by Kolthoff and
Moltzau (1935), several instances of co-precipitation of metal sulphides were reported,
particularly in the presence of CuS or HgS, even under acidic conditions. Hence, it is possible
that the elemental spikes were consumed by co-precipitation with sulphides and were not
subsequently measured in the digestion method employed.
Figure 3.9 The spike recoveries exhibited by the digestion of STAOl, STA04 and ST AOS
J c«
250
200
Si 150
100
50
0
-50
m ST
Ba Cd Co Cr Cu H g M o N i
Metal species
Pb Z n
isiAOl nnn 8X^04 rasiAos
Figure 3.10 The effect of the soil sulphur content upon the spike recoveries for Co, Cu and Ni.
i I
100000
•s-10000 5
SIA05 SlAOl Soil sample
S T A 0 4
3Co ttUDCu m N i ESS
f 1000 eg
100
110
3.3.4 Thermal analysis: Survey of activation procedures
As discussed in section 3.2.5, only the the thermal traces recorded for STA05 with and without
the influence of the selected activation agents are discussed below.
3.3.4.1 Carbonisation alone and physical activation
The thermal trace of STA05 alone in is shown by figure 3.11. It produced only a single weight
change ( « 8 %), which peaked at about 200 °C and was slightly exothermic in nature, implying
that the reactions occurring were forming more stable structures. These reactions were attributed
to carbonisation reactions occurring between the organic contaminants within the soil. There was
also a gradual weight loss over the whole temperature range which was probably due to continued
reactions caused by the increasing temperature. At approximately 700 °C, there was a slight
increase in the rate of weight loss. One possible cause of this phenomenon has been considered
by Snoeyink and Weber (1967) who report that as the temperature of preparation for carbon
increases above 700 °C, the mechanical strength and electrical conductivity of the material
increases. This same phenomenon was related to the graphitisation reactions which are known to
occur in carbons which exhibit mesophase properties. Marchand (1986) represented the change
in electrical resistivity (conductivity) to increasing temperature (carbonisation/graphitisation) as
shown in figure 3.12. Hence, the observed weight loss could have been due to the elimination of
hydrogen and light hydrocarbons as the graphitisation reactions proceeded, consolidating the
carbon skeleton.
Figure 3.11 Thermal analysis of STAGS in
T G - j
D T G / DTA
i i
TEMPERATUREICI m
The effect of heating STA05 in a CO; atmosphere is shown in figure 3.13. This produced a
steady weight loss on the TG curve at temperatures up to 700 °C as carbon was indiscriminately
111
removed from the soil initially in an exothermic manner, probably through volatilisation and
carbonisation mechanisms. After this point the TG line begins to pleateau whilst the DTA, briefly,
becomes less exothermic. At approximately 700 °C and above the weight loss occurring becomes
very exothermic corresponding to the expected temperature for the start of the reaction between
CO2 and carbon. This reaction, the activation by CO;, has been previously considered and can
be expected to progress according to the reactions outlined in section 1.5.3.1.2 (Bansal et al.,
1988). Once the temperature reached 900 °C, the weight was declining more slowly, but the
reaction was still exothermic. This drop in weight may have been due to the removal of the most reactive
carbon portions. At elevated tmperatures some reaction of the CO2 with the non-carbon
components of the soil matrix could be expected.
Figure 3.12 Evolution of the resistance of carbons during heat-treatment. An insulator-conductor transition occurs in the high temperature treatment range (H.T.T.) = 600 - 700 °C
Q. Carbonization
K . K .
( Figure 15, ppl l4, Marchand (1986))
The investigation of the effect of pre-carbonisation followed by CO2 activation is shown in figure
3.14. In this case the sample was exposed to the CO2 after 1 hour at 175 °C in Ng. A comparison
of figures 3.13 and 3.14 indicate clearly that although the weight loss exhibited by each of the
samples (at 1100 °C) is similar (viz: 25 % and 29 % for figure 3.14 and 3.13 respectively), figure
3.14 shows two distinct reaction/weight loss zones on both TG and DTG which are exothermic.
The DTA traces for each figure also show two distinct exothermic reactions which are
surprisingly similar. The first exothermic reaction is completed at 700 °C on both figures, and
have similar peak values ( » 8/iV). The second exotherm, which is increasing on both figures,
does show slight differences between the two samples, the DTA value at 1100 °C being 8/xV on
112
figure 3.14 and i2^V on figure 3.13. This lower value for the pre-carbonised sample (figure
3.14) suggests that the materials causing the second exotherm are not reacting as easily as they
did when only exposed to CO^. One explanation of this observation may be that the CO2, at
elevated temperatures, reacts with the mineral components of the soil. However, pre-carbonisation
has resulted in the carbon components of the soil coating some of the the mineral components,
reducing the accessibility of this material to the CO2.
Figure 3.13 Thermal analysis of STA05 with CO,
TDPQWTUREICI 14BB
Figure 3.14 Thermal analysis of STA05 in then COj
TEMPERATURE ID
3.3.4.2 Chemical activation
The chemical activation studies produced results which were markedly different to those for
physical activation. The absence of the trace for H3PO4 should be noted. This arose because
H3PO4 and the soil reacted particularly vigorously under heating and each attempt at measuring
113
the thermal trace resulted in the sample exploding out of the crucible and covering the balance
hangdown in soil/HaPO^, rendering further measurement impossible. Consequently the further
investigation of H3PO4 was discounted.
The H2SO4 impregnated sample (figure 3.15) displayed three very sharp and distinct peaks on the
DTG which were endothermic. The first peak at 100 °C was assigned to loss of water associated
with the acid, but the second peak at 200 °C most probably arose from the carbonisation reactions
which were being influenced by the activation agent. The final peak at 530 °C was probably due
to activation reactions which appeared to be very slightly exothermic.
The HNO3 sample, shown in figure 3.16, did not exhibit any major differences in terms of
reactions which could be attributed to the influence of the HNO3 (compared with figure 3.11).
Due to this apparent lack of reactions, HNO3 was not studied further.
Figure 3.15 Thermal analysis of STAGS with H^SO^
ZnClz (figure 3.17) clearly shows the influence of an activating agent. The two peaks evident on
the DTG at approximately 100 °C and 530 °C have been attributed to water evaporation
(endothermic reaction on the DTA) and activation processes (exothermic reaction on DTA),
respectively. The carbonisation peak evident on figure 3.11 appears to be totally absent in
figure 3.17, but, careful examination of the TG line at 150-175 °C does indicate the presence of
a step in the line (magnitude of weight change is about 2.5 % compared to 8 % for the soil alone
in N2) which is exothermic. This is the carbonisation reaction, which has been
controlled/suppressed by the ZnCl;.
114
Figure 3.16 Thermal analysis of STAGS with HNO3
J
TEMPEJWTUREtC)
Figure 3.17 Thermal analysis of STAGS with ZnCl^
TEHPERfiTUREIC)
Table 3.6 Carbonisation and activation temperatures established for STAGS
Figure NO
Activation agent
Number of reaction peaks
Temperature (°C)
Carbonisation Activation
Weight loss (%)
3.11 none one 175 n/a 24
3.13 CO, one n/a n/a 27
3.14 N2/CO2 two 175 675 30
3.15 H2SO4 three 150 475 65
3.16 HNO3 one 175 none apparent 24
3.17 ZnClj two 150 450 37
115
The temperatures derived from these plots and the associated weight changes are summarised in
table 3.6. The three activation methods which showed most effect upon the samples were CO2,
H2SO4 and ZnClz; as a result of which only CO;, H2SO4 and ZnCl; were utilised to prepare
samples for evaluation.
3.3.5 Characterisation of bulk carbon samples prepared from contaminated soil
3.3.5.1 CHN results
It can be seen from table 3.7 that the carbonisation and activation procedures used caused changes
in the CHN ratio for each soil sample. CO2 and ZnCl; are shown to be the most aggressive
activation agents, with reductions in the measured C component of 5 % and 4.9 % respectively
(down firom 18.8 % to 13.8 % amd 13.9 % respectively), which equates to 26 % of the carbon
content of the soils. H2SO4 does not affect the carbon present in the soil to any great extent.
Reductions in the H and N content of all the samples were also observed, the reductions were
expected as elimination of H and heteroatoms such as N are a feature of carbonisation and
activation, as previously discussed in section 1.5.3.1.
Table 3.7 CHN values for the raw soil and soil derived activated carbons
Percentage Activating
agent ^ H N % C change
Raw soil 18.8 1.7 1.0 0.0
CO2 13.8 0.2 0.3 -26.6
H2SO4 17.3 0.4 0.8 -8.0
ZnCl2 13.9 0.4 0.3 -26.1
3.3.5.2 Aqueous adsorption results
The results in table 3.8 show the % removal and subsequent Relative Activity exhibited by 1 g
of carbon in 100 ml of 10 mM phenol or 4-nitrophenol. Each of the soil carbons prepared
contained different amounts of "active" carbon, as shown in table 3.7, the carbon prepared from
different activation methods will not exhibit the same properties. Thus, each soil carbon could be
expected to adsorb the phenolic molecules to different extents. This means that the % adsorption
data alone would not provide readily comparable information regarding the adsorption ability of
the carbon retained on the samples. Hence, to enable direct comparisons between the effect of the
116
different activation agents upon the adsorption ability of the products the parameter "Relative
Activity" has been used, which is calculated thus:
% adsorption
Relative Activity =
% carbon in sample
Table 3.8 Aqueous adsorption studies results
Sample ID % Adsorption Relative Activity
phenol 4-nitrophenol phenol 4-nitrophenol
CO2 5.1 11.3 0.4 0.8
H2SO4 5.6 20.8 0.3 1.2
ZnClz 12.7 27.0 0.9 1.9
These results clearly show the effectiveness of the different activation procedures with respect to
development of adsorption capability by the treated contaminated soil. The ZnClj exhibits a
% phenol adsorption ability which is twice that of the CO; and H2SO4 activated samples. More
4-nitrophenol than phenol was adsorbed by the samples but this is simply explained by the
differences in the relative solubilities and the hydrophobic/hydrophylic nature of the two
molecules.
Comparison of the CHN results with the aqueous adsorption study of the three carbons is
achieved by use of the Relative Activity parameter, shown in table 3.8. If it is assumed that each
soil sample taken from the STA05 batch contains a similar distribution of contaminants and that
each sample has been exposed to identical carbonisation and activation conditions, then the major
factor governing the development of adsorption characteristics will be the activation agent and
the Relative Activity can be used as a measure of its effectiveness. Consequently, these values
indicate that ZnCU has been the most effective activation agent, producing values for phenol and
4-nitrophenol which were at least 50 % greater than the equivalent values exhibited by the CO2
and H2SO4 activated samples. The superior effectiveness of ZnClj as an activation agent has also
been reported for other waste-derived activated carbons, including shells from the oil palm nut
(Chan et al., 1980), rice hulls (Tanin and Giirgey, 1987), sarkanda (Chughtai et al., 1987),
almond shells (Ruis-Beva et al., 1984) and spent bleaching earth (Pollard et al., 1991a).
3.3.6 Adsorption isotherm study of soil-carbons
As discussed previously in section (1.4.2), apart from direct analysis of the raw isotherm data.
117
there are two isotherm models which are widely used; Langmuir and Freundlich. Both models
have been applied to the adsorption isotherm data for each adsorbent/adsorbate pair and from the
isotherms the respective constants have been calculated.
3.3.6.1 Interpretation of the phenol and 4-nitrophenol adsorption data
Figures 3.18 and 3.19 represent, respectively, the raw adsorption data for the uptake of phenol
and 4-nitrophenol by each of the carbons. The best-fit line for each data set has been established
by application of a log regression. From the best-fit line, the adsorption capacity of the carbon
(x/m), when the residual concentration after adsorption (Ce) is 10 mM, has been approximated
from the graphs. These values are given in table 3.9. This is the method that was applied by Giles
and Nakhwa (1962), to estimate the surface capacity and subsequently calculate the surface area
of various solids and is considered later.
Table 3.9 The adsorption capacity of each activated carbon for phenol and 4-nitrophenol estimated from data in figures 3.18 and 3.19
Sample ID Adsorption Capacity, x/m. Sample ID
Phenol (mM/1 / g) 4-nitrophenol (mM/1 / g)
CO. 0.07 0.09
H2SO4 0.05 0.09
ZnClj 0.15 0.26
Figure 3.18 Isotherms for phenol adsorption by activated soil carbons prepared using different activation procedures
0.16
0.14
0.12
i 0.1 0
i 0.08
J 0.06
0.04
0.02
0 4 6
Ce(inMol/l)
-C02 •HZSCM •*-Zaa2
118
Figure 3.19 Isotherms for p-nitrophenol adsorption by activated soil carbons prepared using different activation procedures
4 6 Ce (inMoi/1)
-C02 • H 2 S 0 4 -*.ZnC12
3.3.6.2 Langmuir adsorption isotherms
Figures 3.20, 3.21 and 3.22 show the Langmuir adsorption isotherms for phenol and 4-
nitrophenol when exposed to each of the three carbons. The values obtained from the
interpretation of these graphs are shown in tables 3.10 and 3.11 for the adsorption of phenol and
4-nitrophenol respectively.
Figure 3.20 Langmuir isotherm for phenol adsorption
250
i
2 3 1/Ce
|-»-C02 -«-H2S04 -*-ZnC12
119
Figure 3.21 Langmuir isotherms for 4-nitrophenoI adsorption (CO, and H2SO4)
250
200
i 150
100
1/Ce
•C02 -»H2S04
Figure 3.22 Langmuir isotherms for 4-nitrophenol adsorption (ZnCl,)
1/Ce
•ZnCa2
3.3.6.3 Freundlich adsorption isotherms
The results of the application of the Freundlich isotherm model are shown in figures 3.23 and
3.24. The equation constants, calculated from these graphs are shown in tables 3.10 and 3.11 with
the corresponding Langmuir constants.
The Freundlich parameter Kp, the intercept of the graph, provides an indication of the adsorption
capacity of the carbon, whilst 1/n, the slope of the plot, is indicative of the adsorption intensity,
(Knettig et al., 1986 and Adamson 1990). Similarly, the Langmuir parameter is the
monolayer capacity, but B is not normally quantified (see appendix IV, equation 12). The
Freundlich equation, although empirical in nature, has found wide applicability in the activated
carbon industry and is routinely utilised to establish the required carbon application to reduce the
120
initial concentrations of polluting chemicals to predetermined residual values (Fochtman and
Dobbs, 1980).
Figure 3.23 Freundlich isotherm for Phenol adsorption
0
LogCe
C02 * H 2 S 0 4 -*-ZnC12
Figure 3.24 Freundlich isotherms for 4-nitrophenoI adsorption
0
Log Ce
C02 -»H2S04
Table 3.10 Langmuir and Freundlich constants derived from the phenol adsorption Isotherms
Sample ID Langmuir parameters Freundlich Parameters
Qmax [ (x /m)^ (mMol/g)
B r? Kp (mMol/g)
1/n rz
CO; 0.06 4.08 0.70 0.04 0.28 0.93
HzSO, 0.08 0.26 0.91 0.02 0.47 0.73
ZnClz 0.12 1.09 0.91 0.05 0.52 0.97
121
Table 3.11 Langmuir and Freundlich constants derived from the 4-nitrophenoI adsorption isotherms
Sample ID Langmuir parameters Freundlich Parameters
Qmax B r Kp 1/n r (mMol/g)
(mMol/g)
C02 0.16 0.16 0.85 0.04 0.41 0.78
H2S04 0.07 3.01 0.70 0.05 0.31 0.72
ZnClz 0.23 2.37 0.93 0.15 0.31 0.97
Each of the adsorption equations show differing levels of applicability to the available data. Using
the r values as a guide, it can be seen that both equations show reasonable fit for the adsorption
of phenol and 4-nitrophenol by the ZnClz activated sample, whilst the adsorption by the H2SO4
activated sample is best represented by the Langmuir model for phenol adsorption, and for the
CO2 activated sample by the Freundlich model for phenol adsorption. The and Kp values
also clearly indicate that the carbon with the largest adsorption capacity was derived from the use
of ZnCl; as the activation agent. This carbon possessed capacities for phenol and 4-nitrophenol
which, in all cases except for adsorption of phenol by the H2SO4 activated sample, were at least
twice that exhibited by the other activation routes.
If the Langmuir and the Freundlich Kp values are compared with the x/m value obtained
from the raw isotherm data (estimated from figures 3.18 and 3.19 and shown in table 3.9), it will
be observed that the numerical value obtained is closer to the Langmuir than the Freundlich value,
as would be expected, since the nature of the Freundlich equation is such that as the concentration
of adsorbate increases the equation does not show a saturating or limiting value (Adamson, 1990).
3.3.6.4 Gas adsorption data
The samples prepared using CO2, H2SO4 and ZnCl; activation were fully analyzed using the gas
adsorption technique discussed earlier (section 3.2.7.4). The gas adsorption data is summarised y
in table 3.12. The gas adsorption isotherms for each sample are shown in figure 3.25. Each
isotherm is a type IV, with an H3/H4 hysterisis loop (figures 1.12 and 1.13), which were
discussed in section 1.4.2, and are indicative of a mesoporous solid.
The BET data confirm^ the findings of the aqueous adsorption analysis of the samples, the ZnCI; ^
sample possessed the greatest micropore volume and BET surface area (greater by an order of
magnitude compared to the COj and H2SO4 samples). The pore size distribution data, although
122
not qualitative, is useful for comparative purposes. Comparing the three distributions shown in
figure 3.26, it can clearly be seen that ZnClj had a distribution of pores in the 20-80 A region,
which tended towards a microporous character. This was confirmed by the large micropore
surface area shown in table 3.12, which accounts for 77 % of the total surface area. The COj
sample showed a similar trend in terms of pore-size distribution, but, the micropores accounted
for only 46 % of the measured surface area. Finally, the H2SO4 exhibited a reduced pore-size
distribution which was mesoporous in nature and centred at 40 A . The meso and macropores
accounted for 60% of the BET surface area.
Table 3.12 Compiled gas adsorption data
Surface area (m^/g) Pore volume (ml/g)
Sample Identity BET
range (P/P°)
Rz BET Meso &
macro
Micro total vol. ads.
Micro
CO2 0.01-0.21
0.9999 40.73 22.01 18.72 9.3637 0.0085
H2SO4 0.01-0.21
0.9999 22.56 13.55 9.01 5.1863 0.0040
ZnClz .008-.11
1.0 131.24 30.49 100.75 30.171 0.0425
Figure 3.25 isotherms for the carbonised and activated soil samples
P/PO
These general trends in pore volumes and surface areas can be used very readily to explain the
aqueous adsorption results. The microporous surface of a carbon is considered to contribute most
123
to the aqueous adsorption potential of the carbon. The smaller this is, the less effective an
adsorbent the material is. Consequently, the ZnClj activation method is the most effective in
developing a porous sample with good surface area and porosity characteristics.
Figure 3.26 Pore-size distributions for the activated soil carbons
HP (A)
3.3.6.5 Comparison of the Langmuir and BET surface areas
The surface areas of each for the carbons have been calculated from the aqueous adsorption data
by using either the x/m^, value obtained from the isotherm data or the Langmuir monolayer
capacity (see appendix IV, equation 17), as well as from the gaseous results by the BET method.
The requirements of an adsorbate for determining surface area by adsorption from solution were
described by Giles and Nakhwa (1962): (i) Highly polar
(ii) Partially hydrophobic
(iii) Of small molecular size, preferably planar
(iv) Soluble in polar and non-polar solvents
(v) Coloured (uv/vis active).
They concluded that 4-nitrophenol was most suitable for this task, because it fulfilled all of the
above parameters. A surface coverage value for 4-nitrophenol of 52.5 (52.2 x 10"^ m ) was
quoted, assuming that the molecule was adsorbed flat-on by a charcoal (activated carbon) surface.
The authors used the x / m ^ value from the raw isotherm data to calculate their surface areas, so
for comparison purposes the surface areas of each carbon in this study were calculated using
x/m^ax (table 3.9) obtained in this manner and by use of the Qmax value calculated from the
Langmuir equation (table 3.11). These values are shown in figure 3.27, along with the surface
areas calculated by the BET technique, for ease of comparison. The % difference between each
of the surface area values are also shown. The % difference value was calculated thus:
124
Where:
Figure 3.27
200
%differenc€* 100 B
"A" = Langmuir surface area or Giles and Nakhwah surface area
"B" = BET surface area or Langmuir surface area
Comparison of the surface areas calculated for the activated soil carbons from aqeuous and gaseous adsorption data
200
a 100
1 5 0
100
I C 0 2 H 2 S 0 4
C a r b o n s a m p l e
Z n C I 2
L a n g m u i r S u r f a c e A r e a B E T S u r f a c e area G & N S u r f a c e area
dlff. ( L a n g & B E T ) d i g . ( G & N & B E T ) diff ( L a n g & G & N )
Inspection of figure 3.27 shows that the aqueous adsorption data tends to produce surface area
values which are greater than those calculated from gaseous adsorption data. In particular, the
method of Giles and Nakhwa tended to produce surface areas larger even than those from the
Langmuir method. However, the Langmuir surface area values are of a similar order of
magnitude when compared to the BET values and with the exception of the H2SO4 sample,
agreement of the Langmuir and BET values is within 13 %. The surface areas calculated from
the aqueous adsorption data also confim that ZnClj was the most successful activation agent for
producing large surface area, carbonaceous adsorbents from contaminated soil.
3.4 Summary
The preceeding investigation has shown that contaminated soil, even from the same site, exhibits
extreme heterogeneity with respect to the contamination contained within. For example, the
samples examined contained between 12 % and 50 % carbon (CHN ratio). The nature of this
carbonaceous component was variable, sample STA04 was very tarry whilst STA03 was quite dry
despite the fact that both samples contained similar levels of carbon (32.9 % and 34.3 %
respectively). Although air drying of contaminated soils at RTF is recommended to minimise
contamination volatilisation (BSI, 1988), few changes in the CHN ratio of the contaminated soils
was noted when drying at higher temperatures. Other contaminants present in the soils included;
125
sulphate (present at levels of between 14300-122000 ppm); cyanides (present at levels of between
2010-109800 ppm) and sulphur (present at levels of between 100-77800 ppm). Heavy metals of
especial concern were: Hg (846 ppm in STA04); Pb (986.6 ppm in STAOl and 1189.8 ppm in
STA05); Mn and Zn (3006.9 and 3172.6 ppm respectively in STA05). In general, the
contamination was evenly distributed between the various size fractions of the soils, indicating
that utilisation of particle separation to concentrate the contamination would not be very effective.
Examination of different physical and chemical means of activating the contaminated soil showed
that ZnClz requires the lowest temperatures to effectively carbonise and activate contaminated soil
and that the carbons produced by the ZnC^ activation method exhibited relative activities for
phenol and 4-nitrophenol which were 50% greater than those presented by the soil carbons when
activated with CO^ or H^SO . The ZnCI; activated soil carbon possessed a BET surface area of
131.24 rnVg, which was at least four times as great as the other samples and a micropore surface
area of 100.75 rnVg, which was five times as great. This well developed porosity with extensive
microporous characteristics, was further confirmed by the high phenol and 4-nitrophenol
adsorption capacity of this sample which produced adsorption isotherms for the uptake of phenol
and 4-nitrophenol that were amenable to interpretation by both the Langmuir and Freundlich
models. The adsorption capacity calculated from the Langmuir model was 0.12 mMol/g for
phenol and 0.23 mMol/g for 4-nitrophenol.
The results of this initial activation study have shown conclusively that ZnCI^ is the most suitable
activating agent for contaminated soil conversion into activated carbon. This observation has been
confirmed by other researchers who reached similar conclusions with other waste streams (Ruis-
Beva et al., 1984 and Pollard et al., 1991a). The next stage of this work was to optimise the
activation of contaminated soil with ZnClj and is described in Chapter 4 following.
126
CHAPTER FOUR
ZINC CHLORIDE ACTIVATION OPTIMISATION
4.1 Introduction
The preceding chapter introduced the laboratory techniques and initial study of the feedstock
utilised in this work. This resulted in ZnClj being chosen as the most suitable activation route for
use in the preparation of activated carbon from gaswork contaminated soil. Optimisation of the
ZnClj process is now discussed, describing the production of carbons with prime adsorbent
characteristics with respect to the dosage of ZnCl; and the residence time at the carbonisation and
activation temperatures established by TG/DTA/DTG. The effect of the process upon the
contamination in the soil and the mechanistic aspects of the ZnClj with respect to its interaction
with the soil and the contamination are assessed and discussed.
To enable a realistic evaluation of the properties of the soil-carbons, comparisons of the soil-
carbon characteristics with commercial activated carbons were considered necessary. The
commercial carbons chosen were Norit SA4 and Chemviron Type C. Their technical
specifications are given in table 4.1.
Table 4.1 Manufacturers technical details for the commercial carbons
Carbon Feedstock Activation
Method Temperature
BET area (mVg)
pH Cost (£/tonne)
Norit SA4 Peat Steam 975 °C 650 > 7 1200
Chemviron Type C
Coal Steam n/a n/a < 7 1550
Both carbons are powdered, physically activated and find wide use in water purification
applications.
4.2 Experimental protocols
These are split into two phases: the first considers the effect of varying ZnCl; dosage on the
decontamination of the soils and development of adsorption properties by the soil carbon products.
The second examines the influence of different carbonisation and activation residence times upon
the same factors when activated by the optimum ZnCl; dosage. Consequently, the results and
discussion are separated into two parts, the first for the activation agent study and the second for
the residence time study.
127
4.2.1 Preparation of the soil-ZnClj samples
To 50 g of the < 2 mm soil fraction of each soil sample, the desired amount of GPR ZnClz was
added as a 25 % saturated solution and thoroughly mixed. The sample was subsequently dried
as described in 3.2.5.2.
4.2.2 Thermal analysis of STAOl - STA04 and G1
The thermal analyses of the soil and the ZnClj impregnated soil [1:0.5 ratio (33.33 %)] were
performed as discussed in 3.2.4.2 and 3.2.5.2, using the same experimental conditions. Sample
01 was analyzed in a similar manner to the STA series of soils using a PL-STA1500
TG/DTG/DTA which benefited from improved software control.
4.2.3. Bulk carbon production
4.2.3.1 Optimisation of the activation agent dosage
For this work, samples STAOl - STA05 were utilised. Soil : ZnCl; impregnation ratios of 1:0.5
(33.33 %), 1:1 (50 %) and 1:2 (66.66 %)^/w were investigated for samples STAOl-STA04, with
the additional ratios of 1:0.1 (9.09 %), 1:0.25 (20 %) and 1:0.4 (28 %) being investigated for
STA05. The ZnCl; loaded soils were processed in the carbolite HTR 11/150 rotary furnace as
described in section 3.2.6.2 (figure 3.2). The standard post-preparation treatments utilised in
section 3.2.6.3 were re-applied.
4.2.3.2 Optimisation of carbonisation and activation residence times
Only sample G1 was applied to this work as supplies of STAOl-STA05 had been exhausted. The
optimum ZnClj dose was used throughout the work, with carbonisation times up to 3 hours and
activation times of up to 2 hours being studied. Samples were prepared and treated as described
in 4.2.3.1.
4.2.4 Analysis and characterisation of the soil-carbons
4.2.4.1 CHN analysis
These were determined by the Department of Chemistry, Imperial College, (section 3.2.7.1).
128
4.2.4.2 Total cyanide, total sulphate and free sulphur analysis
These were performed by Environmental Analysis Ltd. as described in 3.2.3.3.
4.2.4.3 Aqueous adsorption studies
The aqueous adsorption ability was assessed using the methodology described in 3.2.7.3.2. Full
adsorption isotherms were prepared for selected samples and modelled to both the Langmuir and
Freundlich adsorption isotherms.
4.2.4.4 Gaseous adsorption analysis
Nitrogen adsorption analyses of the carbons were performed, as discussed in 3.2.7.4.2.
4.2.5 Effect of residence time
The samples prepared with variable carbonisation and activation times were subjected to the
chemical and adsorption analyses, as described above.
4.2.6 Optimised sample production
A bulk sample of the carbon which exhibited the most effective adsorption properties was
prepared using the optimum ZnCl^ dosage and carbonisation and activation times. It was subjected
to the full range of chemical and adsorption analyses which have been used previously. A separate
sample was tested additionally for the metals which were leached from the sample during the acid
wash, using 2 M HCl at 60-70 °C, then rinsing with hot (80-100 °C) water. The filtrates from
successive washes were analysed by ICP-AES for the same metals as previous (table 3.5).
4.2.7 X-Ray analytical techniques
X-rays were utilised in two ways in this work: to provide information on the crystal structure and
on the elemental composition of sample under study.
4.2.7.1 X-Ray Fluorescence analysis (XRF)
By bombarding a solid with high energy electrons, a Bremsstrahlung continuum of radiation is
129
formed. Every element emits X-rays with characteristic energies, which mean that the elemental
composition of a solid can be rapidly determined. In section 3.2.3.4, the total metal content of
the soil was measured by an acid digestion procedure, however, the products of this process
(carbon) were not amenable to analysis by simple digestion, hence, qualitative analysis of the
products for metals were undertaken by XRF. All samples were measured on a Link Instruments
energy dispersive XRF, with radiation supplied by a Cu Ka lamp operating at 10 keV. The
powdered samples were mounted in "eye-glass" shaped sample holders which possessed a
polyethylene window. Spectral assignments were performed using the instrument software.
4.2.7.2 X-Ray Diffraction analysis (XRD)
The use of X-Ray analysis in determining the structure of graphite and carbons was discussed in
section 1.5.2. All the samples were determined as < 150 fim powders on a Phillips PW 1820
based X-Ray diffractometer, with Cua Radiation operating at 14 kV/15mA. Crystal pattern
assignments were performed utilising the instrument software which comprised an on-line version
of the powder diffraction file. The information from powder samples are only of qualitative use
due to their inherent disorder. However, the extensive data-set of X-ray spectra which are
available in the powder diffraction file allows compositional analysis of the samples, which was
the context in which the technique has been applied in the current work.
4.3 RESULTS AND DISCUSSION I: ZnClj dosage optimisation
4.3.1 Thermal analysis of STAOl - STA04 and G1
Figures 4.1-4.11 are the thermal analysis traces for soil samples STAOl - STA04 and G1 with
and without ZnClj. The traces are paired to allow ease of comparison between the soil-only trace
and the ZnClj impregnated soil. Initial inspection of the samples indicates similar characteristics
to those discussed for STA05 with ZnCl; addition (section 3.3.4.2). However, although the soils
exhibit reaction peaks in similar regions of the thermal traces, there are disparities between the
samples in terms of the magnitude and exact positions of these peaks. These thermal analysis
curves strongly support the earlier assertions that contaminated soil is extremely heterogenous.
Considering each soil sample in turn:
The TG/DTG/DTA data for STAOl is shown in figure 4.1. There were only two weight losses
of significant magnitude recorded by the instrument. The major feature of the sample had a DTG
on-set temperature of 175 °C, and a peak weight loss at 240 °C, corresponding to a DTG signal
130
of -150 /ig/min. This reaction was endothermic and has been attributed to the volatilisation and
decomposition of the organic components of the soil (both natural and contamination). The minor
endothermic weight loss had a peak DTG value centred at 120 °C which has been considered to
be traces of moisture in the sample evaporating.
The TG/DTG/DTA data for STAOl with ZnCij are shown in figure 4.2. The effect of the ZnCI;
on the thermal characteristics of the soil are readily identified from comparison of figures 4.1 &
4.2.
Figure 4.1 The thermal analysis traces for STAOl
7
TEMPOWTUREIC)
Figure 4.2 The thermal analysis traces for STAOl plus ZnClj
TDfERATURE(CI 9BB
With reference to figure 4.2, addition of ZnCI^ to the soil caused the water loss peak to
significantly increase with a peak DTG value of 110 °C corresponding to a signal of -75 /^g/min,
which was attributed to the hygroscopic nature of ZnC^. The DTG peak due to organic reactions
131
at 240 °C has a much reduced magnitude (-50 fig/min), indicating that a reduction in volatilisation
was occurring and thus an enhanced carbon retention. There is also a third peak in the DTG trace
at 510 °C, which has been assigned to the presence of ZnCI^ on the basis of studies by other
workers (Ibarra et al, 1991; Jolly et al, 1988; Kandiyoti et al., 1984; Neuburg et al., 1987 and
O Brien et al., 1987). Caturla et al. (1991) reported similar results for the carbonisation and
activation of peach stones using ZnClj activation. They considered that the major carbonisation
reactions occurred in the region 200-500 °C, after which the basic carbon structure was formed
and only minor elimination reactions were occurring.
STA02, figure 4.3, showed no major reaction peaks. There was a small weight loss in the region
of 100 °C attributed to moisture, and a broad weight loss between 200 °C and 500 °C resulting
in a very minor peak on the DTG at 300 °C, which has been assigned to organics volatilisation
from the soil. The sharp peak in the region of 375 °C was due to extraneous causes. This soil
sample possessed the highest carbon content (table 3.3) but exhibited the lowest weight loss under
heating of all the samples. This has been related to the physical nature of the sample, which was
a black solid that was similar in consistency to coke. It is possible that this sample contained
residues from the coke ovens, which due to the nature of the gas manufacturing process, were
virtually devoid of volatile or evaporable carbon compounds (pre-carbonised).
Figure 4.3 The thermal analysis traces for STA02
V V V D T A ^
D T G ^
s i
TEMPQWTUREIC) 900
Addition of ZnCl; (figure 4.4) resulted in changes similar to the those shown by STAOl. The
broad peak was almost totally suppressed and peaks for water and the ZnCl; occurred in similar
positions to those seen in STAOl.
Figure 4.5 shows the thermal trace for the soil STA03. Only a single weight loss could be seen,
with a peak DTG value at 247 °C, which was slightly endothermic. Addition of ZnCl^ (figure
132
4.6) resulted in similar effects as those seen for STAOl and STA02.
Figure 4.4 The thermal analysis traces for STA02 plus ZnCl^
DTA—2
\J . D T G ^ ^ ^ T G
s i
TEMPQWTURECC) 900
Figure 4.5 The thermal analysis traces for ST AOS
jt L
D T A - ^
g§
TOfERATUREtC) 900
STA04 produced a very different series of plots for the thermal analysis data. For the soil only,
figure 4.7, there were three peaks on the DTG, with maximum values at 190 °C, 300 °C and
393 °C of which the peak at 190 °C was slightly exothermic and that at 300 °C was the most
endothermic. The effects of ZnCI; addition are shown in figure 4.8. The endothermic DTG peak,
centred at 90 °C which had been seen in all previous samples, was due to the loss of water. A
reduction in magnitude for the peaks at 190 °C and 295 °C and the disappearance of the peak at
393 °C was evident. This sample was also surprising in that the broad peak in the 500 °C region,
(present in all the other samples) was absent, and a much reduced peak, centred at 590 °C had
appeared.
133
Figure 4.6 The thermal analysis traces for STA03 plus ZnClj
D T A - ^
4
TDrERRTUREtC)
Figure 4.7 The thermal analysis traces for STA04
7
TEMrOWTURECCl see
Figure 4.8 The thermal analysis traces for STA04 plus ZnCl^
D T A ^ ^
^ J ^ D T G s i
TEMPERATURE(CI 9n
134
Tentative assignments for these weight losses from STA04 were made on the basis of the previous
thermal results for soil and on the results for the chemical analysis of the soils. The peaks at
190 °C and 393 °C were considered to be thermal reactions of organic components in the soil.
This sample was extensively contaminated with free sulphur (7.7 % see table 3.4), hence,
the DTG peak at 300 °C was considered to arise from loss of sulphur. To test this hypothesis,
a sample of sulphur flowers were thermally analyzed, using the same conditions as applied to the
soils. Figure 4.9 shows the TG/DTG/DTA results for this analysis. The sulphur was volatilised
completely, with a peak DTG temperature of 303 °C, almost an exact correlation with the peak
seen in figure 4.8.
Figure 4.9 Thermal analysis of sulphur flowers
01 0) 2
-i 1 r ISO 200 250 300
T e m p e r a t u r e ( " C )
Finally, the thermal analyses of soil GI are show in figures 4.10 and 4.11. Figure 4.10 exhibits
similar peaks to those seen previously for the "STA" series of soils. The DTG has a strongly
endothermic peak at 120 °C which was attributed to moisture loss and a second peak, at 220 °C,
of an exothermic nature which was assigned to reactions by organic contaminants. However, this
sample also exhibited two further exothermic reaction peaks at 480 °C and 780 °C. Their source
was unknown but considering that this sample contained a substantial amount of sulphate species
(122000 ppm sulphate; table 3.4) this result suggested that these weight changes may have been
due to the decomposition of sulphate species in Gl. The most likely sulphate source would be
either an iron or calcium derivative as these were the major metals present in all the soil samples
examined (as determined by ICR). To further investigate this, several available sulphate salts were
examined by thermal techniques [Fe(II)S04.7H20, Fe2(in)(S0j3.xH20 and CaSOJ but the results
135
were inconclusive. This suggested that the heterogeneous nature of the samples had caused
perturbation of the thermal responses of some of the contaminating species, thus introducing an
element of uncertainty with regard to the origins of these weight losses.
Figure 4.10 The thermal analysis traces for G1
100-
300 400 500 600 T e m p e r a t u r e C O
600 900
Figure 4.11 The thermal analysis traces for G1 plus ZnClj
a fl) z
-1.0 B 1 j 1
300 400 500 600 T e m p e r a t u r e ( * C )
BOO 900
Addition of ZnCU (figure 4.11) resulted in a broad peak being superimposed upon the first peak,
which may be indicative of the fact that the weight loss seen in figure 4.10 at 120 °C was water
involved in crystallisation and not "free water". The organic peak at 220 °C almost disappeared
136
completely, confirming the above assignment and the broad peak for the ZnClj, centred at 500 °C
became apparent. That previously seen at 800 °C had disappeared, implying that the cause of the
peak had reacted with the ZnClj.
From the interpretation of the thermal data, carbonisation and activation temperatures for each
sample were chosen using the same rationale as that for STA05 (section 3.3.4). These are shown
in table 4.2. The preliminary residence times used for the carbonisation and activation were
established from the TG traces and previous work (Pollard et al., 1991a; 1991b), and were judged
to be sufficient for the carbonisation and activation reactions to have reached completion.
Table 4.2 Temperatures chosen for carbonisation and activation
Sample ID Carbonisation temperature
r c )
Carbonisation time
Activation temperature
r c )
Activation time
STAOl 200 450
STA02 210 120 minutes for 475 60 minutes STA03 175 all samples 475 for all
samples STA04 275 525
samples
STA05 150 450
G1 180 various 375 various
4.3.1.1 Mechanistic aspects of activation and carbonisation by ZnClj
There are an extensive number of available reports studying the effect of Lewis acid type
compounds (such as ZnClj) on the pyrolysis of coals, from which the contamination of these
samples is derived. The proposed mechanism of activation of carbonaceous materials by ZnClj
and by chemical activants consists of chemical dehydration of the starting material, followed by
the attack of the edges of the carbon planes and the removal of tarry material from pores, thus
increasing the pore volume of the carbon lattice (Bansal et. al., 1988; Pollard et al., 1992).
However, it has been shown that addition of these compounds to coals increases char formation
at the expense of gaseous and tar products under slow heating, whilst the reverse is true for flash
pyrolysis. Hydrogen formation readily occurs below 400 °C in the presence of ZnClj and related
compounds, but this is not the case when ZnCL is absent (Ibarra et al., 1991; Jolly et al., 1988;
Kandiyoti et al., 1984; Neuburg et al., 1987 and O'Brien et al., 1987). Work by Jolly et al.
(1988), using XRD, showed that the ZnCl^ added to coal was converted to ZnO and then
subsequently to ZnS after slow heating to 550 °C. They were especially surprised at the detectable
137
presence of ZnS because of the low sulphur content of the precursor coal (0.79 %^/w).
Ibarra et al. (1991) proposed that ZnS and HCl formed when the ZnO reacted with H^S, which
is one of the pyrolysis products of coals. They also noted that the ZnCl^ separated, with the CI
finely dispersed in the coal matrix and the Zn only being detected where there was an abundance
of sulphur. The association of Zn with sulphur has also been noted by other workers
(Kandiyoti et al., 1984 and Ibarra et al., 1991). Zinc has a high affinity for sulphur, as evidenced
by the vigorous manner in which elemental sulphur and zinc powder react when mixed together
and warmed (Bretherick, 1986). The effect ZnCl^ has on coal pyrolysis has thus been suggested
to be a form of catalytic cracking and that this effect is influenced by the presence of associated
compounds (Jolly et al., 1988 and Ibarra et al., 1991). It is very probable that a similar
mechanism is taking place during the carbonisation of these soil samples.
4.3.2 Sample yields and CHN results
The CHN results and sample yield for each carbon produced from soils STA01-STA05 are shown
in table 4.3 with the commercial carbons Norit SA4 and Chemviron Type C included for
comparative purposes. All the soil carbons contain less carbon than the commercial ones, but this
does not necessarily mean that they will possess inferior adsorption properties.
It is immediately evident that the effect of ZnClj was quite pronounced upon the sample yields.
For example, considering soil STAOl; run 016 had zero ZnCij added, whilst run 002 was
33.33 % (^/w) loaded with a nett effect that run 002 had a 21.8 % greater sample yield than
run 016. Secondly, as the dosage of ZnCl^ was increased the sample yield decreased, due to
increased activation of the samples implying greater burning out and/or erosion of the samples.
As a general rule, this effect was repeated with all the samples, although STA04 did deviate from
this trend and in fact this sample exhibited some very curious properties which are discussed later
(section 4.3.4.2).
The CHN results did not exhibit any major trends. Generally, the product carbon content
increased slightly after thermal treatment when ZnClj was added to the soil, compared to the
samples prepared without ZnCl; addition. However, STA04 and STA03 showed a slight drop in
carbon content with addition of ZnCl;, except when the dosage of ZnClj reached 66% by weight
of soil and then it showed an increased carbon retention. When compared to the CHN values for
the raw soil (table 3.4), all the samples (except those made from STA05) showed an increase in
the carbon content. This contrast with ST AOS was thought to occur because the chemical nature
of the organic contamination in soil STA05 was different to that in the other soils. This sample
138
had a pungent smell, reminiscent of creosote (PCP), and an oily appearance. The smell indicated
that the pollution was very volatile, implying that the contamination would tend to volatilise rather
than undergo carbonisation reactions.
Table 4.3 Sample identity, yields and CHN values
Soil identity Run Number Percentage
ZnClj added r /w)
yield C H N
STAOl 016 0.00 44.7 15.90 0.61 1.47
002 33.33 66.5 21.31 0.76 2.25
Oil 50.00 60.0 20.67 0.55 2.05
013 66.67 60.8 21.10 0.73 1.99
STA02 019 0.00 70.3 56.85 0.97 1.42
005 3133 75.9 57.54 0.85 1.37
010 50.00 73.7 56.18 0.88 1.23
014 66.67 68.7 58.72 0.71 1.16
STA03 017 0.00 68.8 41.84 0.68 1.39
004 33.33 74.8 40.09 0.76 1.21
008 50.00 71.2 37.89 0.77 1.22
012 66.67 72.0 42.42 0.66 1.38
STA04 018 0.00 24.5 63.52 2.08 6.93
001 33.33 50.2 60.28 2.08 8.65
007 50.00 60.0 52.66 1.84 6.72
015 66.67 42.8 59.30 1.65 7.12
STA05 020 0.00 68.7 12.95 0.40 0.42
026 9.09 73.0 14.05 0.38 0.34
025 20.00 73.3 15.22 0.51 0.39
027 28.57 71.3 14.37 0.34 0.38
006 33.33 68.2 17.96 0.52 0.36
009 50.00 71.6 18.17 0.56 0.34
034 66.67 70.3 15.60 0.35 0.51
Nor it SA4 - - - 79.12 1.94 -
Type C - - - 81.49 1.24 trace
The CHN value for the samples were determined by combustion and measurement of the CO;
produced. Thus, the apparent increase in carbon content was not due to formation of carbon, but
139
elimination of volatiles. The major factors to influence the CHN would be: (i) the amount of
volatile substances in the soil (for example, STA04 contained a large amount of sulphur) and (ii)
the amount of acid soluble material in the samples (removed during the HCl wash). The hydrogen
content decreased in all the carbonised samples compared to the raw soil, which was to be
expected, given the nature of the carbonisation reactions which were occurring within the samples
(previously discussed in sections 1.5.3.1.land 4.3.1.1). Finally, the nitrogen content of the
samples showed variable behaviour, which can be attributed to the form in which the nitrogen was
contained within the feedstock soil. During carbonisation, heteroatoms are either eliminated or
incorporated into the carbon lattices. Thus if the nitrogen was already part of an organic molecule
(such as aniline, pyridine or quinoline) it would have been more likely to be retained during
carbonisation than if present as ammonia or similar derivatives. Once nitrogen is incorporated into
a char or carbon removal is very difficult and only achieved by heating to > 800 °C in air,
resulting in Nj evolution (Baraniecki et al., 1957).
These observations on sample yield and CHN changes were indicative of the heterogeneous nature
of contaminated soil and also give support to the previous discussion of the considered mechanism
for the mode of action by ZnCl; upon carbonaceous materials.
4.3.3 Total cyanide, total sulphate and free sulphur analysis
In section 3.3.2, the analyses of the raw soils were presented and discussed. Table 4.4 shows how
addition of ZnClj followed by thermal processing affected these contaminants. The greatest effect
which the thermal treatment had upon the soil contamination was the destruction of the cyanide
component. The iron complex (ferric ferrocyanide, Fe^[Fe(CN)j3), which is the major source of
cyanide contamination in gaswork soils, readily decomposes upon heating to temperatures between
450 °C and 500 °C forming HCN, Fe Oy and (CN); (de Leer et al., 1985; Weast, 1987;
Koopmans and Reintjes, 1988; de Leer 1988). The addition of ZnCl; did not make a major
additional effect upon the extent of cyanide reduction, except for carbons prepared from STA04.
Run 018 retained 20 % of the cyanide contamination after processing without ZnClj addition, but
runs 001, 007 and 015 all exhibited cyanide removals in excess of 94 %. The X-ray diffraction
results (discussed later in section 4.5.2) indicated a possible explanation for this observation, since
the only crystalline phase which was detected by X-ray diffraction in the soil STA04 was iron
nickel cyanide hydrate [Ni2Fe(CN)6.xHzO]. This complex appears to have a greater thermal
stability than Fe4[Fe(CN)6]3. Consequently, for its complete decomposition either higher
temperatures than were used for the preparation of the samples or the addition of a catalyst to
assist the decomposition of this complex would be required. For the cyanide analyses CuSO^ was
140
added to catalyse the decomposition of the Fe4[Fe(CN)j3). Thus, the results imply that ZnCl^ is
behaving as a catalyst. There may also have been a contribution towards the removal of the
cyanide contamination by the HCl wash, which is known to decompose any remaining free
cyanides into HCN.
Apart firom carbons prepared firom STA04, the residual cyanide concentrations were all >99 %
lower than for the untreated soil. With reference to Appendix I and the guidance values for
cyanide concentrations, carbon samples prepared from STAOl - STA03 and STA05 (with ZnCl;
addition) were all below the ICRCL threshold value for both free and complex cyanides (table
AI.6). With respect to the GLC values for total cyanides (table AI.4) these same samples could
also be classed as uncontaminated or only very slightly contaminated. The Dutch assessment
values indicated that the samples would meet or only just exceed the 'A' reference value (table
AI,2) and for the Canadian remediation criteria (table AI.l) the carbons would be suitable for
application to agricultural land.
Table 4.4 Total sulphate, total cyanide and free sulphur analysis
Run ZnClz Soil ID Total cyanide Total sulphate Free sulphur number added (ppm) (ppm) (ppm)
r w (% reduction shown in brackets)
016 0.00 STAOl 40 (99.7) 54900 (35.1) 340 (88.3)
002 33.33 STAOl 6.3 (99.6) 34600 (59.1) 3120 (-7.6)
019 0.00 STA02 625 (96.3) 5500 (68.7) 100 (0.0)
005 33.33 STA02 < 1 (99.9) 4200 (76.2) 160 (-1100)
017 0.00 STA03 450 (99.6) 17400 (44.2) 400 (74.4)
004 33.33 STA03 < 1 (100) 6700 (78.5) 740 (52.6)
018 0.00 STA04 400 (80.1) 21800 (41.4) 520 (99.3)
001 33.33 STA04 50 (97.5) 5000 (86.6) 360 (99.5)
007 50.00 STA04 < 10 (99.5) 37000 (0.5) 100 (99.9)
015 66.67 STA04 105 (94.8) 22400 (39.8) 80 (99.9)
020 0.00 STA05 50 (99.6) 14100 (88.4) 2160 (-107.7)
006 33.33 STA05 2.5 (100) 5400 (95.6) 1200 (-15.4)
009 50.00 STA05 17.5 (99.9) 6100 (95.0) 640 (38.5)
034 66.67 STA05 < 10 (99.9) 6500 (94.7) 520 (50.0)
The decomposition of cyanide by thermal means using varying atmospheres (air and N ) and
temperatures was studied by de Leer et al. (1985). They concluded that, at temperatures above
141
400 °C, almost total cyanide decomposition occurred and increasing the temperature fiirther
produced no advantages. The influence of the atmosphere was negligible, since oxidation would
only occur at temperatures > 800 °C, but an inert atmosphere was preferred to minimise the risk
of explosion. They also noted that when sampling the off-gases for analysis by GC-MS, removal
of sulphur vapour was required to prevent clogging of the equipment.
The sulphate and sulphur species within the soil did not exhibit the same removal capacity as the
cyanide and there was an extreme variability between the samples in terms of sulphate and
sulphur reduction. Consideration must be given to the mechanisms of elimination of these species
from the system. The main mechanism for sulphur removal was by volatilisation, as recorded by
the thermogravimetric analysis traces discussed previously and reported by de Leer et al. (1985).
However, a counter-measure for this mechanism may have been the behaviour of Zn in the
presence of elemental sulphur which has been discussed (section 4.3.1.1). With respect to the
carbonisation reactions, some sulphur will also be incorporated into the carbon lattice and as
surface complexes. The presence of iron in the carbon feedstock has been also considered to
promote the retention of sulphur as pyrites (Baraniecki et al., 1957).
Reduction of the sulphate in the soil was not as readily assigned, since the associated cations will
affect the ultimate behaviour of the sulphate. For example: CuSO^ decomposes to CuO at 650 °C;
CaSO^ (natural anhydride) does not decompose but undergoes a crystal phase change from
rhombic to monoclinic at 1163 °C and ZnSO^ (natural zinkosite) decomposes at 600 °C (Weast,
1987). Other possible effects may have been due to reduction of sulphate into elemental sulphur
or HjS by the Hj produced during carbonisation and activation (thus accounting for the apparently
anomalous increases in sulphur concentrations by runs 002, 005, 006 and 020). It is probable
that the HCl wash was quite effective in removing sulphur and sulphate from the samples. The
production of HjS was very noticeable during the washing stage. HCl is also the extracting agent
in the method for determination of total sulphate in soils (BSI, 1990), but the effectiveness of this
method will depend upon the accessibility of the sulphate to the HCl.
With respect to the guidance values in Appendix I, table AI.6, the sulphate values for the ZnCI^
activated samples all exceeded the threshold values and in the case of runs 002, 007 and 015,
exceeded the action value for gardens, allotments and landscaped areas. The sulphur values for
all the ZnClz activated samples were all within the threshold value and could consequently be
classed as uncontaminated with respect to sulphur.
142
4.3.4 Aqueous adsorption results
4.3.4.1 Influence of carbon on solution pH
As previously discussed in section 1.5.4 the surface structures of the carbon influence the pH of
the solution into which it is placed and consequently affect the molecules in the solution which
are to be adsorbed. If the solution pH exceeds the pKa value for either phenol or 4-nitrophenol
their adsorption by the carbon surface will be reduced due to molecular ionisation (Suzuki and
Takeuchi, 1986). The first concerns of this work were to establish the pH characteristics of the
carbons under adsorption conditions. These results are shown in figures 4.12 and 4.13.
The pH measurements indicated that the carbons made were of the 'L' type as the solution had
a pronounced acidic character (section 1.5.4). The average pH for the phenol adsorption
experiments was pH 3.47 and for 4-nitrophenol adsorption pH 3.52. The average nett pH change
upon addition of carbon to the phenol solutions was -2.62 pH units and -1.32 pH units for the 4-
nitrophenol solutions.
Figure 4.12 The variation of solution pH for adsorption of phenol by the soil-carbons
6
%
1 2 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 25 26 27 34 Run number
Phenol pHf B p H change (Blank corrected)
These results indicated that the dominant form of the adsorbates in solution were molecular and
not ionised. The surfaces of activated carbons are considered to be negatively charged (Bean et
al., 1964), hence repulsion of anions will occur which will be of significance when high pH
solutions (when pH > pKa of adsorbate molecules) are being adsorbed (Snoeyink and Weber,
1967). It has been noted that activated carbons are most effective at adsorbing acid-based organic
molecules from acidic solutions (Ward and Getzen, 1970; Semmens et al., 1986). Phenol,
however, exhibits an adsorption maximum at neutral pH and exhibits a decrease in adsorption as
143
the pH becomes more acidic, due to competition from ions for some of the adsorption sites
on the carbon surface. This effect is more pronounced with phenol than 4-nitrophenol due to the
relative affinities of the two organic molecules for the carbon surface, and consequently the
stength of the bonds formed between the carbon surface and the adsorbate (Snoeyink and Weber,
1967). These observations can be related to the adsorption mechanisms which have been proposed
for activated carbons and are discussed in section 4.3.4.3.
Figure 4.13 The variation of solution pH for adsorption of p-nitrophenol by the soil-carbons.
5
4
3
2
^ 1
0
- 1
- 2
-3 1 2 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 25 26 27 34
Run number
• 4-nitrophenol pHf S B l a n k corrected pH change
4.3.4.2 Single point adsorptions
Table 4.5 contains the data for the single point adsorption experiments with both the % adsorption
and relative activity for phenol and nitrophenol shown. The effect of ZnClj was quite pronounced
when comparing the adsorption of phenol or nitrophenol for samples prepared with ZnClj (runs
001-015, 025-027 and 034) and without ZnCl; (runs 016-020). Considering the % adsorption
results in table 4.5; the samples prepared without ZnCl; (runs 016 - 020) exhibited much reduced
adsorption compared to the ZnClj activated samples. The exception to this trend was run 018
(produced from STA04 without ZnCIJ which exhibited a percentage adsorption which was greater
even than carbons prepared from the other soil samples with ZnCl; addition. This sample
appeared to be self activating, and is discussed later in this section.
144
Table 4.5 Aqueous adsorption results
Run ZnClj Soil ID Number added
Phenol 4-Nitrophenol
(% ,^/w) % adsorption
Relative activity
% adsorption
Relative activity
016 0.00 STAOl 0.00 0.00 13.07 0.82
002 33.33 16.63 0.77 29.42 1.36
Oil 50.00 39.22 2.14 57.15 2.76
013 66.67 23.56 1.12 46.07 2.18
019 0.00 STA02 2.45 0.04 14.62 0.26
005 33.33 25.23 0.44 39.45 0.69
010 50.00 44.26 0.79 44.14 0.70
014 66.67 21.13 0.36 37.46 0.64
017 0.00 STA03 3.76 0.09 14.62 0.37
004 33.33 31.55 0.79 49.36 1.23
008 50.00 30.63 0.81 51.34 1.35
012 66.67 30.15 0.71 55.14 1.30
018 0.00 STA04 45.70 0.72 63.98 1.01
001 33.33 69.12 1.21 93.22 1.63
007 50.00 72.56 1.38 95.71 1.82
015 66.67 67.05 1.13 94.36 1.59
020 0.00 STA05 5.49 0.42 6.49 0.50
026 9.09 6.46 0.46 9.04 0.64
025 20.00 15.36 1.01 21.99 1.44
027 28.57 15.60 1.09 24.95 1.74
006 33.33 19.77 1.03 30.91 1.72
009 50.00 19.40 1.07 29.16 1.60
034 66.67 13.28 0.85 31.14 1.99
Norit SA4 76.45 0.89 93.66 1.09
Chemviron Type C 95.41 1.30 99.27 1.39
All the 4-nitrophenol samples exhibited a nominal adsorption even for non-activated samples, a
phenomenon which was explained by Weber et al. (1983); "...in any real system containing a
heterogeneous mixture of solids, hydrophobic pollutants will tend to accumulate in those materials
with the highest organic content". Thus there would almost certainly have been a mutual attraction
between the hydrophobic carbon surface and the hydrophobic 4-nitrophenol.
145
Addition of ZnCI; caused the adsorption character of carbons produced from runs 001 - 015, 025-
027 and 034 to greatly increase for both phenol and 4-nitrophenol. The highest affinities did not
occur with the same addition of ZnCI; for each sample. This was to be expected with samples
which exhibit such variability in their composition (refer to tables 4.3 and 4.4). To enable easy
comparison between the effects of the changing ZnClj addition upon the adsorption, % carbon
and relative activities of each soil sample, this data has been graphically displayed in figures 4.14
to 4.18. These graphs clearly showed which ZnCl; additions produced the best activated carbons.
In discussion of these graphs, only the relative activities have been considered, as the relative!
activity reflects the affinity of the carbon in the product for the adsorbates, irrespective of the
amount of carbon present. The results obtained for the processed soil samples were spread over
a wide range of values, and all the samples demonstrated that the relative activity increased with
the addition of ZnClj up to a maximum of 50 % ZnCl; loading, before declining with further
additions. The results also indicated that the heterogeneous nature of the samples affected the
degree of activation developed by the samples, with their initial composition being a major
influence upon the final product's performance as an adsorbent.
For STAOl and STA04 carbons (figures 4.14 and 4.17), peak relative activities occurred with
50 % ZnCl; addition as exhibited by runs Oil and 007 respectively. The STA02 carbon with
maximum phenol relative activity was run 010 (50 % ZnCl^ addition, see figure 4.15) whilst the
maximum for 4-nitrophenol occurred for run 005 (33 % ZnCI; addition). The STA03 carbon
which showed greatest phenol activity was run 004 (33 % ZnCl^ addition) but run 008 (50 %
ZnCl; addition) exhibited a slightly higher 4-nitrophenol maximum (figure 4.16). Finally, the
STA05 carbon to show maximum phenol relative activity was run 027 (28 % ZnCl; addition)
whilst run 034 gave the highest 4-nitrophenol activity (66.67 % ZnC^ addition, figure 4.18).
Figure 4.14 The effect of different ZnClj additions upon STAOl derived carbons
16(0) 2(33.33) 11 (50) Sample ID (% ZnC12 added)
13 (66.67)
ESJ% Carbon phenol ads. E3% p-nitrophenol ads.
R. A. phenol -*-R. A. p-nitrophenol
146
Figure 4.15 The effect of different ZnCI% additions upon STA02 derived carbons
S 40
19(0) 5 (3333) 10 (50) Sample Id. (% ZnCl2 added)
14 (66.67)
% Carbon ^ % phenol ads. ^ % p-nitrophenol ads.
R- A. phenol A. p-nitrophenol
Figure 4.16 The effect of different ZnCI^ additions upon STA03 derived carbons
& 40
1 0.4 73
17 (0) 4 (3333) 8 (50) Sample Id. (% ZnC12 added)
12(66.67)
% Carbon ^ % phenol ads. ^ % p-nitrophenol ads.
R. A. phenol R. A. p-nitrophenol
Figure 4.17 The effect of different ZnClj additions upon STA04 derived carbons
100
I 18(0) 1(3333) 7(50)
Sanyle Id. (% ZhC12 added) 15 (66.67)
® % Carbon ^ % phenol ads. p-nitrophenol ads.
* R . A. phenol -*-R. A p-nitrophenol
147
Figure 4.18 The effect of different ZnCI^ additions upon ST AOS derived carbons
2.5
£-I-
i i>
20(0) 26(9.09) 25(20) 27(28.57) 6(3333) 9(50) 34(66.67) Sample Id. (% ZnC12 added)
E3 % Carbon ^ % phenol ads. ^ % p-nitrophenol ads.
R. A. phenol -*-R. A. p-nitrophenol
The sample which exhibited the highest overall relative activity was run Oil . However, it is
interesting to note that the feed soil, STAOl, contained the least carbon of all the samples whilst
samples prepared from STA02, which contained the most carbon (50 % produced activated
carbon with the lowest relative activity. This is again indicative of the heterogeneous nature of
the samples and differences in the types of carbon material in the feed soils.
It was noted earlier that carbons produced from soil STA04 exhibited curious activation
behaviour. Run 018, prepared without ZnClj, demonstrated levels of adsorption for phenol and
4-nitrophenol which were greater than the best activated sample prepared from any of the other
soil samples, and it was concluded that STA04 was self activating. The data for the soil samples
treated without ZnClj indicated a relationship between the initial sulphur species content and the
relative adsorption activity developed by the products towards phenol and 4-nitrophenol. This
relationship is clearly shown in figure 4.19, which shows that the relative activity increases as the
sulphur species in each sample increase, except for phenol adsorption by run 020 and run 016.
The latter result was considered to be due to experimental inaccuracy, but due to the limited
availability of run 016, the experiment could not be repeated.
148
Figure 4.19 Effect of soil sulphur species content upon development of phenol relative activities.
0.6 CQ
53 0.4
run 018 run 016 run 017 Run number
run 019 run 020
120 £
100 1 -
80
60 1 40 i
Q 20
1
0 &
-•-Phenol R. A. 4-Nitrophenol R. A. ^Total sulphur ^ecies (ppm)
The deviation of run 020 may be have been influenced by the high Zn content of STA05
compared to the other soils (table 3.5). Thus, although STA05 possessed the lowest level of
sulphur species, as shown in table 3.4, the Zn may have been behaving as an activating agent and
assisting in development of adsorption ability, in a similar manner to the ZnC^. To test the
hypothesis that sulphur species did have an influence upon the activation of the carbonised soil
sample, a sample of STA05 (40 g) was combined with a paste of FeSO^.VH^O (20 g) giving a
1:0.5 (33.33 %) soil:activant ratio. FeSO^ was chosen as the sulphur source because Fe was
already present in great excess in all the soil samples, so that addition of extra Fe should not have
unduly influenced the carbonisation and activation.
Thermal analysis of the STA05 / FeSO^ sample and of FeSO^ alone was performed as described
previously. Figure 4.20 shows the thermal data obtained for the STA05 / FeSO^ sample, which
gave a characteristic moisture loss in the 100-120 °C region. A continuous weight loss then
occurred up to 460 °C where a sharp weight loss was observed, with a peak value at 490 °C.
Comparing the data in figure 4.20 with those for the soil alone (figure 3.11) and FeSO^.VH^O
alone (figure 4.21) the peaks associated with the decomposition ofFeSO^.VH^O are absent, except
for the O2 losses observed at 580 °C and 660 °C, which appear to have been combined at 490 °C.
The weight loss at 200 °C seen on the DTG of soil STA05 (figure 3.11) was absent from figure
4.20, indicating that the FeSO^.VHzO was influencing the reactions occurring in the soil upon
heating.
149
Figure 4.20 Thermal analysis of ST AOS and FeSO^
7
TEMPERATUREICI 9ee
Figure 4.21 Thermal analysis of FeSO^.TH^O
n r 300 400 500 600
T e m p e r a t u r e ( " C )
There was an observable effect with FeSO^ addition to the soil, thus a large sample was
carbonised at 175 °C for 120 minutes (data from figure 3.11) and activated at 470 °C for only
40 minutes (data from figure 4.20) due to the rapid nature of the reaction. The chemical and
adsorption characteristics for the resulting sample are summarised in table 4.6.
The FeSC^ activated carbon possessed less carbon than the CO2, H2SO4 and ZnClj activated
carbons discussed in chapter 3 (table 3.7), although the H and N values were comparable. The
% phenol and 4-nitrophenol adsorption results were similar to the H2SO4 and CO2 samples, but
the lower carbon content of this sample meant that it possessed a superior relative activity (see
table 3.8). The BET surface area and total pore volume were similar to those produced by CO;
150
activation, although the micropore volume was approximately half that of the CO; carbon. This
sample was predominantly mesoporous. These results indicated that FeSO^ was capable of
activating these contaminated soils, a fact which supports the hypothesis that sulphur species
present in the contaminated soils are augmenting the activation process, enhancing the
development of adsorption characteristics.
Table 4.6 Chemical and adsorption data for FeSO^ activated sample
% Carbon 11.23
% Hydrogen 0.25
% Nitrogen 0.28
% phenol adsorption 5.90
% 4-nitrophenol adsorption 8.10
Phenol relative activity 0.53
4-Nitrophenol relative activity 0.72
Freundlich 4-nitrophenol Kp (mMol/l/g) 0.03
Langmuir 4-nitrophenol (x/m)^^ (mMol/l/g) 0.08
Langmuir 4-nitrophenol surface area (m^/g) 26.30
BET surface area (mVg) 37.98
Meso / macropore surface area (mVg) 29.17
BET "c" value 170.9
Total pore volume (ml/g) 8.7303
Micropore volume (ml/g) 0.0038
4.3.4.3 Mechanistic aspects of phenolic adsorption on activated carbon
Factors which affect the adsorption ability of carbons originate from three sources: (i) adsorbate
effects (molecular weight, branching, solubility, polarity and functionality of the adsorbates), (ii)
carbon effects (carbon surface chemistry, surface area and porosity) and (iii) solution effects (pH
and secondary ions) (Caturla et al., 1988: Giusti et al., 1974: Snoeyink and Weber, 1969)
The accepted mode of adsorption for phenol was first proposed by Coughlin and Ezra (1968) as
the interaction of the tt electron ring system of the phenol ring with the active carbon surface.
They also observed that increasing the number of oxygen groups on this surface caused a decrease
in the phenol uptake due to withdrawal of electrons from the graphite basal planes, weakening
the 7r electron attraction. The adsorption of phenols is strongly influenced by oxygen-containing
functional groups. This effect was demonstrated by Urano et al. (1981) and Biniak et al. (1989).
151
They added O2 to carbon surfaces by either heating in flowing air or HNO3 or removed surface
groups by heating in a or He atmosphere respectively. Oxidation resulted in a substantial
increase in acidic functional groups on the carbon surface and as the 0% content of the surface
increased, phenol adsorption decreased (and vice-versa). However, it was Mattson et al. (1969)
who interpreted this effect by studying the change in the infra-red spectra of carbon surfaces with
and without adsorbed organic molecules, specifically: phenol, 3- and 4-nitrophenol and
nitrobenzene. They noted that the nitro-substituted aromatics adsorbed more strongly than phenol
implying a specific interaction between the aromatic substituents and the carbon surface. The IR
signals for the phenolic C-0(H) and -NO2 groups were unchanged when adsorbed but the phenolic
-(0)-H signal vanished indicating the existence of hydrogen bonds between the adsorbed phenolic
and the carbon surface functional groups. Thus, they concluded that the major influence in the
adsorption processes was the aromatic ring of the adsorbates forming a donor-acceptor complex
with the carbon surface, especially surface carbonyl groups. The strong adsorption seen for the
nitro-substituted aromatics was attributed to the -NO; group withdrawing electrons more strongly
from the aromatic ring than the -OH group, enhancing the acceptor nature of nitrophenol and
giving a stronger complex than phenol.
4.3.4.4 Adsorption Isotherms
Adsorption isotherms for phenol and 4-nitrophenol uptake by the soil carbons were constructed
and the data modelled to the Langmuir and Freundlich equations. The Langmuir and Freundlich
parameters for phenol and 4-nitrophenol adsorption are presented in tables 4.7 and 4.8
respectively. The isotherm plots, for completeness, are presented in Appendices IV and V. The
commercial carbons have also been included for comparative purposes. The soil carbons
possessed inferior and Kp values compared to the commercial carbons, as expected. The
adsorption data for both adsorbates investigated was amenable to interpretation by both the
adsorption models, as illustrated by the r* values for each isotherm, the majority of which were
0.90 or greater. The 4-nitrophenol values were greater than the equivalent phenol data,
as discussed previously in sections 3.3.5.2 and 4.3.4.3.
The values for 1/n given in tables 4.7 and 4.8 indicated that the adsorptions which were occurring
were favourable (Hassler, 1963), as discussed in section 1.4.2.1.1. The values for B are related
to the heat of adsorption (Weber, 1985), although, as stated in section 3.3.6.2, B is not normally
quantified. There were no obvious trends with respect to the variation of B with monolayer
capacity indicating that there were several factors involved in the derivation of B, which will all
play a part in affecting the calculated value (as given by equation 12 in Appendix II).
152
Table 4.7 Langmuir and Freundlich constants for phenol adsorption
Run number
Langmuir parameters Freundlich parameters
[(x/m)^J (mMol/l/g)
B rz Kp (mMoI/l/g)
1/n r2
002 0.16 1.08 0.96 0.06 0.43 0.98
Oil 0.14 1.18 0.97 0.06 0.48 0.99
013 0.25 0.96 0.98 0.08 0.49 0.99
005 0.22 1.45 0.97 0.09 0.44 0.98
010 0.16 1.56 0.95 0.07 0.39 0.92
014 0.18 1.45 0.95 0.07 0.44 0.98
004 0.13 3.78 0.89 0.07 0.35 0.90
008 0.33 0.57 0.95 0.09 0.44 0.99
012 0.25 2.24 0.96 0.11 • 0.43 0.99
001 0.82 6.82 1.00 0.37 0.37 0.94
007 0.73 1.76 0.99 0.28 0.44 0.97
015 0.42 8.57 0.97 0.32 0.46 0.99
025 0.13 1.37 0.94 0.05 0.42 0.96
027 0.15 1.22 0.96 0.06 0.41 0.93
006 0.16 1.10 0.95 0.06 0.45 0.98
009 0.13 0.83 0.99 0.00 0.37 0.86
034 0.13 1.21 0.96 0.05 0.43 0.98
SA4 0.88 40.36 0.90 0.88 0.31 0.94
Type C 0.97 42.92 0.98 0.87 0.35 0.93
The optimum Langmuir monolayer capacity for phenol adsorption was produced by
addition of 33.33 % ZnClj to the contaminated soil. Carbons made from soils STAOl and STA03
deviated from this trend, but the addition of extra ZnCl^ would be an extra process cost which
would not be warranted by the adsorption gains, especially considering the unpredictable effect,
previously noted, which the contamination of the soil can have upon the carbonisation and
activation procedure. The effect of increasing ZnClj addition upon the phenol Langmuir
monolayer capacity is illustrated by figure 4.22.
153
Table 4.8 Langmuir and Freundlich constants for 4-nitrophenol adsorption
Run Langmuir parameters Freundlich parameters
Qmax [(x/m)^J
(mMol/I/g)
B rz K (mMol/l/g)
1/n r^
002 0.23 16.17 0.94 0.17 0.35 0.95
Oil 0.26 15.24 0.98 0.17 0.32 0.93
013 0.35 13.57 0.98 0.23 0.36 0.93
005 0.28 26.35 0.97 0.23 0.32 0.96
010 0.26 22.57 0.98 0.17 0.28 0.84
014 0.70 1.92 0.99 0.15 0.40 0.86
008 0.30 37.84 0.97 0.25 0.32 0.97
012 0.39 35.11 0.99 0.30 0.32 0.95
001 0.95 62.87 0.98 0.72 0.33 0.91
007 1.06 20.43 1.00 0.66 0.31 0.88
015 0.72 71.13 0.98 0.87 0.38 0.93
025 0.17 17.65 0.93 0.12 0.31 0.87
027 0.22 5.03 0.99 0.11 0.40 0.84
006 0.27 3.31 1.00 0.11 0.43 0.89
009 0.19 17.20 0.92 0.14 0.32 0.93
034 0.19 29.03 0.85 0.15 0.29 0.96
SA4 0.93 192.86 0.91 1.08 0.34 0.94
Type C 1.32 2054.05 0.85 1.86 0.19 0.91
Figure 4.22 The effect of ZnCl^ upon the phenol Langmuir monolayer capacity
1
0.8
0.6
0.4
0.2 —•
40 50 ZnC12 addition (% wAv)
! &
•STADl •»STAD2 +STA03 -^STAD4 -e-STAD5
154
From the Langmuir values for both adsorbates, the Langmuir surface areas of the carbons
have been calculated. These values are given in table 4.9 and compared with the BET surface
areas (discussed later in section 4.3.5.1). The surface areas calculated from the Langmuir
values are at least 50% lower than the BET values. The probable explanations are that; (i) not
all the surface of the sample was amenable to phenol adsorption (such as siliceous material) and
(ii) the N; adsorbate used in the BET study penetrated much narrower pores (an N2 molecule is
half the diameter of a phenol molecule). The surface areas calculated from the 4-nitrophenol
values were approximately 10% greater than the phenol values, as expected.
Table 4.9 Langmuir surface areas for each sample
Run ZnCl; added phenol 4-nitrophenol phenol 4-nitrophenol number r/w)
Langmuir surface area (m^/g) % difference (Langmuir vs. BET)
002 33.33 49.81 72.51 -54.85 -34.26
Oil 50.00 43.30 82.98 -63.83 -30.68
013 66.67 77.33 110.93 -52.73 -32.20
005 33.33 68.04 87.58 -65.36 -55.41
010 50.00 49.64 81.91 -73.93 -56.98
014 66.67 57.93 221.09 -70.46 12.73
004 33.33 42.24 n/d -80.72 n/d
008 50.00 104.15 94.94 -53.11 -57.26
012 66.67 79.18 123.35 -70.99 -54.81
001 33.33 257.45 301.10 -54.80 -47.14
007 50.00 229.79 336.34 -61.24 -43.27
015 66.67 130.46 229.10 -82.60 -69.44
025 20.00 41.47 55.27 -63.70 -51.61
027 28.57 46.64 69.95 -65.41 -48.12
006 33.33 49.35 86.05 -62.40 -34.43
009 50.00 41.86 59.676 -69.29 -56.21
034 66.67 41.42 58.87 -65.40 -50.82
SA4 278.19 292.74 -64.81 -62.97
Type C 305.20 416.00 -70.97 -60.43 n/d = not determined due to insufficient sample
155
4.3.5 Gaseous adsorption
4.3.5.1 Surface area
All of the soil carbons and the commercial carbons exhibited a characteristic type IV isotherm
shape with a hysteresis loop of mixed H3/H4 characteristics, as shown by figures 4.23 and 4.24
respectively. This indicated that the soil carbons were mesoporous solids and that capillary
condensation was occurring within the mesopores at elevated P/P® values. The mixed H3/H4
characteristics of the hysteresis loops suggested that the samples contained si it-shapeded pores,
although the presence of H4 character also implied that the slit-shapes were associated with the
micropores (Sing et al., 1985). The BET and t-plot surface areas for all the soil carbons and the
two commercial carbons are presented in table 4.10. The difference in the relative surface areas
and pore volumes of the commercial carbons compared to the soil carbons are noticable from
inspection of the volumes adsorbed at P/P° = 0.95, shown in figures 4.23 and 4.24. With the
exception of run 001, which has been recognised previously as being an exceptional sample, the
other soil carbons adsorbed between 3 to 6 times less than the commercial carbons.
Figure 4.23 Nj adsorption isotherms for runs 001 j 002j 004, 005 and 006
Run 01
Run 04
^ u n 05
Run 06
Run 02
P / P o
The low surface areas given by runs 016, 017, 019 and 020 which were prepared without
activation agent emphasised the necessity of the ZnClz. Run 018 exhibited an extraordinarily high
surface area, for an un-activated sample, which was considered to be due to the effect of the
sulphur species (as discussed previously in section 4.3.4.2). It can be seen that although this
sample had a high BET surface area, the relative contributions to this value by the [macropores
+ mesopores] and micropores were different when compared to the other carbons from this soil.
Run 018 had its surface area shared almost equally between these pore classifications, but for runs
156
001, 007 and 015 the micropores contributed over 75 % of the surface area.
Figure 4.24 adsorption isotherms for the commercial carbons Norit SA4 and Type C
P/Po
TyPEC
NORIT SA4
The significant effect of the ZnClj upon the carbon's micropore characteristics are expressed by
all the activated samples. In particular, runs 004 and 005 exhibited micropore surface areas which
had increased in excess of 100 times compared to the non-activated samples (runs 019 and 017
respectively). It is a well established fact that the micropores make the largest contribution to the
surface area and the adsorption capability of activated carbons {eg\ Wildman and Derbyshire,
1991), and that superior micropore structure is produced when chemical activation is used in
carbon preparation. This was noted by Kadlec et al. (1970) and Rodriguez-Reinoso and Molina-
Sabio (1992) for ZnCl; activation compared to CO^ or air activation. Their findings have been
confirmed by this work. The major effect of the ZnCI^ in activation has been in the enhanced
formation of micropores.
157
Table 4.10 Parameters determined from the gas adsorption data
Run Surface area (m^/g) BET "c" r^ for BET range number
BET meso + macropores
micropores value BET plot (P/P°)
016 21.62 17.51 4.11 80.34 0.99 0.04 - 0.25
002 110.31 28.63 81.68 628.16 1 0.01 - 0 . 1
Oil 119.70 4.16 115.54 1169.60 0.99 0.001 - 0 . 1
013 163.61 31.06 132.55 615.30 0.99 0.01 -0 .14
019 19.62 18.45 1.17 59.94 1 0.05 - 0.25
005 196.41 23.17 173.24 377.31 1 0.007 -0.115
010 190.39 71.35 119.04 324.88 0.99 0.003 - 0.2
014 196.12 49.99 146.13 251.69 0.99 0.01 - 0.2
017 24.21 23.21 1.01 57.44 1 0.05 - 0.25
004 219.13 42.50 176.63 629.48 1 0.005 - 0.1
008 222.12 12.36 209.76 383.02 1 0.01 - 0 . 1
012 272.98 59.83 213.15 239.15 0.99 0.01 - 0 . 1
018 433.01 222.07 210.94 503.66 0.99 0.01 - 0.18
001 569.60 125.79 443.81 1775.20 1 0.001 - 0 . 1
007 592.90 121.30 471.60 1276.70 0.99 0.005 -0 .12
015 749.61 166.26 583.35 487.67 0.99 0.01 -0.135
020 19.91 9.37 10.55 33.56 0.99 0.05 - 0.24
026 51.84 12.35 39.50 437.00 0.99 0.01 -0 .12
025 114.23 32.73 81.50 610.64 0.99 0.01 - 0.13
027 134.83 34.18 100.65 737.78 0.99 0.005 - 0 . 1
006 131.24 30.49 100.75 358.95 1 0.01 -0 .11
009 136.29 53.11 83.18 362.48 0.99 0 .01-0 .17
034 119.71 50.97 68.75 193.37 1 0 .015-0 .17
SA4 717.68 125.18 592.25 1645.7 1.0 0.001 - 0.098
Type C 1051.30 71.01 980.92 427.74 0.99 0.01 -0.145
Addition of ZnClj and increased dosage gave the expected increased in BET surface area. The
samples which exhibited the largest surface areas were made from soil STA04, with the largest
BET surface area (749.61 m^/g) and micropore surface area (583.35 m^/g) attributable to run 015.
There was an inconsistent effect upon the surface area of carbons from different soils as the ZnCl^
dosage increased, which was also a feature of the aqueous adsorption studies. This was attributed
to the heterogeneous composition of the soils and their disparate behaviour during carbonisation
158
and activation. The commercial carbons exhibited much greater surface areas than the soil
carbons, as would be expected considering that the feedstocks for them are carbon rich. The only
exception was run 015 prepared from STA04, which had a greater surface area than Norit SA4.
The microporous nature of the ZnCl2 activated carbons and the commercial carbons were
confirmed by the high BET "c" values shown in table 4.10. The BET equation does have
limitations to its applicability and the "c" value is considered to be a good indication of the
accuracy of the BET result. Hence, "c" values between 2 and 150 are reported to be acceptable
(Gregg and Sing, 1982; Sing et al., 1985) but, as "c" increases, this indicates that there are
micropores within the sample and the validity of the BET approach to calculation of the surface
area is in doubt. The "c" values for the ZnCI^ activated samples all exceed this guide value and
as the micropore surface area increased or decreased the "c" value mirrored this change.
However, the surface area data is very usefiil for comparative purposes, and the BET surface
areas correlate well with the aqueous adsorption results. Figures 4.25 - 4.29 summarise the effect
of different ZnClj dosages upon the different surface area components calculated from the gas
adsorption data.
Figure 4.25 The effect of ZnCl^ upon the gas adsorption data for carbons manufactured from soil STAOl
200
!
in
1400
1000 150
100
n
16(0) 2(33.33) 11 (50)
Run number (% ZnC12 added)
13 (66.67)
*c* value QMeso-and macropores E3Micropores
These graphs show the variable behaviour of the "c" value, BET surface areas and the varying
contributions of each pore range to the surface area value of the carbons made from each soil
upon activation by different loadings of ZnClg. No single loading of ZnClz consistently produced
the highest surface areas in each of the soil samples. Carbons prepared from STAOl, STA03 and
STA04 gave the maximum surface area value with addition of 66.67 % (^/w) ZnClj, whilst
STA02 only required 33.33 % ZnClj, and for STA05 the difference in the surface areas between
using 28.57 %, 33.33 % or 50.00 % ZnClj was only 5 m-/g. Clearly, the heterogenous nature
159
of the soil feedstocks was influencing the effect of the ZnCU. These values indicated that a
minimum addition of 33.33 % C/w) was required to maximise surface area production. The
microporous nature of these carbons is clearly shown by the following figures, and the pore-size
distribution analyses, which are discussed next, also confirms this.
Figure 4.26 The effect of ZnClj upon the gas adsorption data for carbons manufactured from soil STA02
250 400
S 100
19(0) 5 (33.33) 10 (50)
Run number (% ZnC12 added)
14 (66.67)
-••"c" value DMeso-and macropores ^Micropores
Figure 4.27 The effect of ZnCIj upon the gas adsorption data for carbons manufactured from soil STA03
g I"
17(0) 4 (33.33) 8 (50)
Run number (% ZnC12 added)
12 (66.67)
"•"c* value EEMeso-and macropores ^Micropores
160
Figure 4.28
! 2 rt
S
800
600
400
200
The effect of ZnClj upon the gas adsorption data for carbons manufactured from soil STA04
2000
1000 g
18(0) 1 (33.33) 7 (50)
Run number (% ZnC12 added)
15 (66.67)
*c" value DMeso-and macropores ^Micropores
Figure 4.29 The effect of ZnClj upon the gas adsorption data for carbons manufactured from soil ST AOS
800
600
400
200
r
20 (0) 26 (9.09) 25 (20) 27 (28.57) 6 (33.33)
Run number (% ZnC12 added)
9 (50) 34 (66.67)
*c* value BIMeso-and macropores ^Micropores
4.3.5.2 Pore-size distribution analysis - soil carbons
Figures 4.30 to 4.34 show the mesopore size distributions for the carbons made from each soil
sample, and figure 4.35 the pore-distribution for the commercial carbons. Table 4.11 details the
pore volumes.
161
Figure 4.30 Mesopore size distribution for STAOl carbons
Run 16
Run 13
Run 02
Run 11
RP U
Refer to table 4.5 for the ZnCl2 loadings used for activation
Figure 4.31 Mesopore size distribution for STA02 carbons
Run OS
Run 14
Run 10
Run 19
RP (A
Refer to table 4.5 for the ZhQz loadings used for activation
162
Figure 4.32 Mesopore size distribution for STA03 carbons
Run 08
Run 12
Run 04
Run 17
BP ( « tSs ' sfeo
Refer to table 4.5 for the ZnCl2 loadings used for activation
Figure 4.33 Mesopore size distribution for STA04 carbons
Run IS
Run 01
Run 07
Run 15
BP U
Refer to table 4.5 for the ZnCl2 loadings used for activation
163
Figure 4.34 Mesopore size distribution for STA05 carbons
/Run 09
•Run 27
Run 26
Run 06
Run 25
Run 20
HP (A)
Run 34
Refer to table 4.5 for the loadings used for activation
Figure 4.35 Mesopore size distribution for commercial carbons
Tao HP (A)
Tiie mesopore size distributions are only useful in a comparative way (as discussed in section
1.4.3.2). All the samples (including the commercial carbons) exhibit mesopore size distributions
which tend towards a microporous nature. Increasing the ZnCIj dosage to the soils appeared to
cause an increase in the total pore volume in the majority of the samples (as shown in table 4.11).
Carbons from STA02 and STA05 showed a small deviation from this trend but the heterogeneous
164
nature of the soil samples will have influenced these results.
With reference to figure 4.30 - 4.34, increasing ZnClj dosage caused a progressive reduction in
the number of pores in the 20 - 40 A region for almost all the samples whilst the number of pores
greater than 40 A increases. This effect was particularly clearly shown by the carbons made from
soils STA04 and STA05. This trend towards increased mesoporous nature as the ZnCl; dosage
was increased was also reported by Rodrfguez-Reinoso and Molina-Sabio (1992) for ZnClj
activated lignocellulosic wastes. The commercial carbons exhibited superior micropore volumes,
indicative of their higher adsorption capacity and surface areas.
Examination of the micropore-size distribution was possible through the application of the
Horvith and Kawazoe equation (section 1.4.3.3). Horv^th and Kawazoe plots for the series of
carbons prepared from each of the soils are shown in figures 4.36 to 4.40 with the commercial
carbons in figure 4.41. The Horvdth and Kawazoe effective pore diameters and maximum dW/dR
values are also summarised in table 4.11.
Figure 4.36 Horvhth and Kawazoe micropore size distributions for STAOl carbons
Run 13 Run 11
Run 02
•n ifg i'.4 i'.6 i t r E f f i c t l v a p o m d I S M t a r , (MM)
Refer to table 4.11 for the ZnCl2 loadings used for activation
165
Figure 4.37 Horv&th and Kawazoe micropore size distributions for STA02 carbons
/Run 05
Run 10- un 14
Run 19 S TS Ti S its it4 ite ItS fe
E f f e c t i v e p o r e d i a m e t e r . (NM)
Refer to table 4.11 for the ZnCl2 loadings used for activation
Figure 4.38 Horvath and Kawazoe micropore size distributions for STA03 carbons
Run 04
Run 08
Run 12
Run 17
E f f e c t i v e p o r e d i a m e t e r . (NHj
Refer to table 4.11 for the ZnCl2 loadings used for activation
166
Figure 4.39 Horvath and Kawazoe micropore size distributions for STA04 carbons
Run 07
Run 01
Run 15
Run 18
Is i l ' . 2 i'.A I ' .G 1 . 8 2
E f f e c t i v e p o r e d l e n e t e r . (KM)
Refer to table 4.11 for the ZnCL loadings used for activation
Figure 4.40 Horvath and Kawazoe micropore size distributions for STA05 carbons
Run 21 • /Run 25
Run 06
Run 09
Run 26
Run 34 Run 20
TS "74 JS t iTS ni TB" E f f e c t i v e p o r e d l e a e t e r , (NM)
Refer to table 4.11 for the ZJ1CI2 loadings used for activation
167
Figure 4.41 Horvath and Kawazoe micropore size distributions for the commercial carbons
Type C
Norit SA4
-rr .8
Effective pore diameter, (NM)
The increased ZnCI; dosage had the effect of reducing the intensity of the maximum dW/dR value
by greater erosion and enlargement of the pores during activation. The effective pore diameters
which corresponded to the maximum dW/dR values are shown in table 4.11. There was an
observed trend (for most of the soil samples), of the maximum effective pore diameter decreasing
as the degree of activation increased, but there was not an optimum effective pore diameter which
corresponded to the maximum aqueous or gaseous adsorption. The commercial carbons, which
were very effective adsorbents, showed intense dW/dR peaks in the 0.6 nm (6.0 A ) region.
Similar dW/dR values are displayed by runs 004, 001 and 007, but not all of these samples
possessed adsorption capabilities comparable to the commercial carbons. This observation is
consistent with previous discussions which indicated that the adsorption ability of an activated
carbon is affected by several parameters including the composition of the feedstock, the particle
size distribution, the total surface area, the chemical composition of the adsorbent and the surface
structure of the adsorbent (Wildman and Derbyshire, 1991, Evans and Marsh, 1979).
168
Table 4.11 Pore volume parameters calculated from the gas adsorption data
Run Number
Soil ID ZnCl, added
(%, " /w)
Total pore
volume (ml/g)
Micropore volume (ml/g)
HorvAh and Kawazoe peak effective pore diameter (A)
Maximum Horv^th and
Kawazoe dW/dR value
016 STAOl 0.00 4.97 0.0015 10.6 0.5
002 33.33 25.36 0.0338 5.8 2.3
Oil 50.00 27.52 0.0471 5.6 3.3
013 66.67 37.61 0.0572 6.0 3.1
019 STA02 0.00 4.51 0.0002 8.8 0.25
005 33.33 45.15 0.0798 6.0 - 6.2 3 . 0 - 3 . 2
010 50.00 43.77 0.0567 5.6 1.5
014 66.67 45.09 0.0775 5.6 1.5
017 STA03 0.00 5.57 0.0000 9.0 0.2
004 33.33 50.38 0.0723 6.0 4.2
008 50.00 51.06 0.0878 5.4 2.6
012 66.67 62.75 0.1078 5.9 2.0
018 STA04 0.00 99.54 0.0913 6.0 1.8
001 33.33 130.94 0.1829 6.4 5.8
007 50.00 136.30 0.1979 6.4 7.0
015 66.67 172.32 0.2501 6.2 3.8
020 STA05 0.00 4.58 0.0054 11.4 0.4
026 9.09 11.92 0.0161 7.0 1.2
025 20.00 26.26 0.0343 6.5 2.6
027 28.57 31.00 0.0211 6.0 2.8
006 33.33 30.17 0.0425 6.0 1.6
009 50.00 31.33 0.0364 6.0 1.4
034 66.67 27.52 0.0341 5.8 1.0
SA4 Type C
164.98
241.69
0.2483
0.4386
6.1
6.0
4.6
5.5
4.3.6 Summary of the effect of varying ZnCIj loadings
The results up to this stage in the work have indicated that increasing the ZnClj loading beyond
33.33 % by weight of soil (Soil : ZnCI; ratio = 0.5 ^/w) produced marginal gains in terms of
169
contaminant destruction (which are more affected by other stages in the process) and development
of adsorption capability. The addition of 33.33 % ZnCU has been established as an optimum
loading of ZnCljfor development of activated carbon characteristics by the soil-carbons. The next
stage of the work required an examination of the results from the study of dwell time variation
for carbonisation and activation for the development of carbons with prime adsorbent
characteristics.
4.4 RESULTS AND DISCUSSION II: Optimisation of carbonisation and activation
dwell time
4.4.1 Introduction
The optimum ZnCl; dosage of 33.33 % /w was used for each of the carbonisation and activation
times described in table 4.12 below.
Table 4.12 Carbonisation and activation times investigated during process optimisation
Run number Carbonisation time (minutes)
Activation time (minutes)
038 120 60
039 60 60
040 90 60
041 60 120
042 120 120
043 180 60
044 180 120
4.4.2 C.H.N, analyses
Very variable C.H.N, yields were obtained by the different dwell times used, as illustrated by
table 4.13. The explanation for this variability in all the values has already been considered
(section 4.3.2). From the earlier discussions of the mechanism of carbonisation and the mode of
action of ZnClj during activation (section 4.3.1.1), it is possible from inspection of these data to
predict which samples underwent the most activation and may, therefore, exhibit the highest level
of adsorption. Runs 040 and 041 had the lowest carbon and nitrogen contents (implying higher
levels of activation). Run 041 also possessed the lowest hydrogen content, indicating that the
carbon formed by this sample possessed more unsaturation and thus a greater aromatic content
than the other carbons.
170
Table 4.13 Sample C. H. N. values
Run number % Carbon % Hydrogen % Nitrogen
soil G1 13.96 1.34 1.13
038 16.86 0.32 0.79
039 16.41 0.36 0.88
040 15.80 0.38 0.76
041 15.71 0.15 0.76
042 16.22 0.27 0.80
043 17.97 0.28 0.92
044 18.10 0.28 1.11
4.4.3 Total cyanide, total sulphate and free sulphur analysis
The effects of varied carbonisation and activation times upon cyanide, sulphate and sulphur level
in soil G1 are shown in table 4.14. As previously observed (section 4.3.3), high cyanide
reductions (>99%) were recorded for all samples, with the residual concentration at levels below
the ICRCL trigger values shown in table AI.6, Appendix I. The sulphate and sulphur reductions
were all of the order of 50 %. With reference to table AI.6, the sulphur levels were below the
action values but the sulphate levels were still too high for any of the applications detailed. These
results confirmed that the greatest destruction of cyanide was achieved through application of
thermal energy, whilst thermal and chemical effects influenced the sulphate and sulphur contents
(section 4.3.3).
Table 4.14 Total cyanide, total sulphate and free sulphur analysis
Sample ID total cyanide (ppm)
total sulphate (ppm)
free sulphur (ppm)
(% reduction shown in brackets)
038 25 (99.7) 46300 (62.1) 2400 (39.4)
039 10 (99.9) 63200 (48.2) 1960 (50.5)
040 15 (99.8) 68600 (43.8) 2180 (44.9)
041 10 (99.9) 66700 (45.3) 2360 (40.4)
042 5 (99.9) 65500 (46.3) 2160 (45.5)
043 10 (99.9) 66400 (45.6) 2280 (42.4)
044 5 (99.9) 62700 (48.6) 1940 (51.0)
171
4.4.4 Aqueous adsorption results
The % adsorptions exhibited by these carbons, given in table 4.15, were lower than those
observed for the carbons made from the STA series of soils, but soil G1 possessed the lowest
carbon content of all the soils examined, thus explaining the observed results. The varied
residence times did not have a marked effect upon the adsorption capabilities displayed by these
carbons, and only one sample was noticeably different: run 41, which exhibited the greatest
adsorption and relative activity for phenol and the second highest adsorption and relative activity
for 4-nitrophenol.
Table 4.15 Single point adsorptions
Run number Phenol 4-nitrophenol Run number
% adsorption relative activity % adsorption relative activity
038 11.16 0.66 19.92 1.18
039 12.91 0.79 15.56 0.95
040 11.19 0.71 13.65 0.86
041 13.32 0.85 17.91 1.14
042 10.10 0.62 14.25 0.88
043 11.89 0.66 14.43 0.80
044 11.38 0.63 16.08 0.89
4.4.5 Gaseous adsorption
The gas adsorption data is shown in table 4.16. The adsorption isotherms for these samples were
all type IV in character with H3/H4 hysteresis loops and figure 4.42 shows a typical isotherm for
run 041. The gas adsorption results were corroborated by the aqueous adsorption results with run
041 exhibiting the largest BET and micropore surface areas.
The results of the pore-size distribution analysis of these samples were not wholly conclusive with
respect to the optimum carbonisation and activation time. The mesopore size distributions for the
carbons were all very similar and required plotting on an expanded scale to facilitate their
separation, as shown by figure 4.43. A similar situation occurred for the Horv&th and Kawazoe
micropore size distributions in figure 4.44. The pore-size distribution plots did not provide any
information which explained the reason for the higher adsorption ability of run 041 compared to
the other carbons.
172
Table 4.16 Surface area parameters determined from the gas adsorption data
Run Surface area (m^/g) BET "c" r for BET BET range number
BET meso + micropore macropore
value plot p/po
038 61.94 30.93 31.01 282.81 1.0 0.01 -0 .15
039 74.13 34.81 39.32 334.12 1.0 0.01 -0 .15
040 67.40 33.78 33.62 307.46 1.0 0.01 -0 .16
041 81.96 36.31 45.65 345.49 1.0 0.01 -0 .14
042 68.97 31.34 37.63 323.68 1.0 0.01 -0.155
043 77.29 32.86 44.43 375.54 1.0 0.01 -0 .15
044 78.08 32.67 45.41 363.06 1.0 0.01 -0.154
Figure 4.42 adsorption-desorption isotherm for run 041
Run 41
Run 45
Figure 4.43 Mesopore size distribution for run 038 to run 044
Run 41 /
Run 39
Run 43
Tl Run 40
Run 42
Run 38 Run 44
RP A
—go
173
Figure 4.44 Horv&th and Kawazoe micropore size distributions for runs 038 to 044
Run 41
Run 44
Run 43
Run 39
Run 42
Run 40
Run 38
.5 . 6 7 7 Effective pore diameter. (NM)
. 8
4.4.6 Adsorption results summary
The maximum adsorption ability and surface area was exhibited by run 041, prepared with a
carbonisation time of 60 minutes and an activation time of 120 minutes, but the effect of the
varied times upon the pore-size distributions was minimal. With the available information, run
041 was considered to be the most promising sample, so a bulk sample of this carbon was
prepared (run 045) and subjected to aqueous and gaseous adsorption analyses to ensure that the
properties of the carbon were not adversely affected by the increased sample bulk treated.
4.4.7 Analysis of the bulk sample of optimised soil carbon
4.4.7.1 Chemical composition
Comparing the equivalent samples 041 and 045 indicated whether scaling-up the production would
adversely affect the decontamination and conversion of the soil. The chemical components of the
soil (C.H.N., cyanide, sulphate and sulphur) are given in table 4.17. There was some variability
between the samples. Particularly, the H results between 041 and 045 indicated that less H was
eliminated from run 045 than run 041, an indication that run 045 may have undergone less
activation than run 041. Additionally, run 045 displayed a much higher sulphur removal than
run 041. The remaining parameters were not markedly different. The extent to which these
differences affect the adsorption performance of the carbon are considered in section 4.4.7.3.
174
Table 4.17 Comparison of runs 041 and 045
Chemical component Run 045 Run 041 % difference between 041 and
045*
Carbon (%) 15.05 15.71 -4.20
Hydrogen (%) 0.54 0.15 260
Nitrogen (%) 0.67 0.76 - 11.84
Total Cyanide (ppm)* 11.2 (99.89) 10 (99.90) 12.00
Total Sulphate (ppm as SO;)* 75600 (37.00) 66700 (44.42) 13.34
Free Sulphur (ppm)* 420 (89.39) 2360 (40.40) - 82.20
# % difference calculated using 041 data as reference value
4.4.7.2 Analysis of the wash waters
The effect of multiple washing upon the metals leaching from the carbonised soil was studied with
a sample prepared in the same way as run 045.
Table 4.18 Metal analysis of HCl wash liquors for optimised sample
Element WASH 1 WASH 2 WASH 3 WASH 4
Mg % Mg % Mg % Mg %
Ba 0.04 0.02 0.05 0.02 13.15 5.59 1.30 0.55
Cd 0.05 4.22 0.003 0.23 0.00 0.00 0.00 0.00
Co 0.10 2.48 0.006 0.15 0.01 0.20 0.00 0.37
Cr 1.68 4.97 0.09 0.26 0.13 0.39 0.12 6.47
Cu 0.01 0.02 0.19 0.50 0.83 - 2.12 2.55 0.00
Hg 0.04 0.88 0.005 0.11 0.00 0.00 0.00 0.00
Mn 4.43 1.49 0.76 0.25 0.03 0.01 0.00 0.44
Mo 0.00 0.00 0.003 0.05 0.01 0.13 0.03 2.41
Ni 1.13 7.65 0.10 0.68 0.15 1.02 0.35 0.04
Pb 3.93 0.88 1.24 0.28 0.67 0.15 0.19 0.12
V 0.17 0.55 0.07 0.23 0.06 0.18 0.04 4.69
Zn 3154.7 3092.8 50.83 49.83 17.80 17.46 14.99 1.46
of soil processed and the % of the total metal in the soil which this value represents (The total values were determined by digestion and are shown in table 3.5).
The above results show how successive washing of the same sample of soil results in
175
progressively less metals being removed, as would be expected. The amounts of metal removed
by the HCl washes were insignificant, except for Zn. However, the first wash removed most of
the Zn, with further washes containing less Zn. Removal of the other elements was not excessive.
These results indicated that retention of the metal by the carbonaceous char during processing
does occur, a factor which is confirmed by the XRF results (section 4.5.1).
4.4.7.3 Aqueous adsorption
Table 4.19 Langmuir and Freundlich constants for phenol adsorption
Run number
Langmuir parameters Freundlich parameters
Q [ (x /m)^ (mMol/g)
B Kp (mMol/l/g)
1/n
041
045
0.09
0.09
1.01
0.86
0.98
0.94
0.03
0.03
0.47
0.46
0.98
0.96
The Langmuir and Freundlich phenol adsorption isotherm data for the optimised soil carbon
(33.33 % /w ZnCl;, 1 hour carbonisation and 2 hour activation) shown in table 4.19 indicate that
although the energy of the adsorption (B) and the intensity of adsorption (1/n) is lower for the
bulk sample (run 045) than the test sample (run 041), the monolayer capacities for the samples
are exactly the same, and both samples display a good degree of fit to the adsorption equations,
as shown by their r values.
Table 4.20 Langmuir and Freundlich constants for 4-nitrophenol adsorption
Run number
Langmuir parameters Freundlich parameters
Q m a x
[(x/m)mJ (mMol/g)
B Kp (mMol/l/g)
1/n
041
045
0.17
0.16
9.20
6.68
0.94
0.97
0.10
0.10
0.33
0.31
0.96
0.96
The Langmuir and Freundlich 4-nitrophenol isotherm parameters in table 4.20 exhibit the same
trends as the phenol data. The 4-nitrophenol and B values are greater than for phenol, as
expected, although the value for run 045 is 6% lower than the 041 value, but this was not
considered to be significant. The adsorption results indicated that with respect to activity and
adsorption development, increasing the amount of soil treated by a factor of three did not
adversely affect the development of adsorption properties by the products.
176
Table 4.21 Langmuir surface areas for runs 041 and 045
Sample ID phenol 4-nitrophenol adsorption adsorption
phenol 4-nitrophenol adsorption adsorption
Langmuir surface area % difference (Langmuir vs. BET)
041 27.67 52.39 -66.24 -36.08
045 27.92 51.13 -60.93 -28.46
Conversion of the Langmuir value into surface area gave the results in table 4.21. The
phenol surface area was lower than those calculated for 4-nitrophenoI adsorption and from the gas
adsorption data with the BET equation. The factors influencing these features were discussed
previously (section 4.3.4.3).
4.4.7.4 Gas adsorption
The gas adsorption data for run 045 provided a surface area of 71.46 mVg (33.95 mVg in the
meso- & macro-pores and 37.52m^/g in the micropores). The equation constsants were are
follows: "c" value = 298.50, r = 1.0, BET range = 0.01-0.16 P/P°. The BET surface area was
lower in run 045 than run 041, with the greatest reduction occurring in the micropore
contribution. The reduction in the micropore surface area was reflected by the reduced "c" value.
Although run 045 developed a lower surface area, the aqueous adsorption isotherm results (tables
4.19 and 4.20) indicated that the surface activity of the carbon was not affected.
Figure 4.45 Mesopore size distribution for run 041 and 045
Run 41
Run 45
""35" 3o RP (A)
"35" "So
177
Figure 4.46 Horvath and Kawazoe micropore size distributions for runs 041 and 045
Run 41
Run 45
75 Ts TT Effective pore diameter. (NK)
. 8
The mesopore size distribution analysis of runs 041 and 045 (figure 4.45) shows that run 041
possessed a slightly higher distribution of pores greater than 30 A than run 045, an observation
which is supported by the surface area analysis. Similarly, the micropore analysis in figure 4.46
shows that run 041 has a more intense dW/dR peak than run 045, confirming the surface area
analysis which gave run 041 a higher micropore area than run 045.
4.4.7.5 Summary
The bulk carbon, run 045, was prepared with the optimum ZnClj dosage and at the carbonisation
and activation times which were shown to produce the best adsorption characteristics. Increasing
the weight of soil treated did not adversely affect the adsorption properties of the carbon product.
Some changes in the behaviour of the contamination was observed. There was a lower reduction
in the hydrogen content of the carbon in run 045 than run 041 but an increased sulphur removal.
Overall, run 045 performed as well as run 041 as an adsorbent.
4.5 X-Ray Analysis
4.5.1 X-Ray Fluorescence
Table 4.22 summarises the results of the X-ray fluorescence analysis of the powdered carbons
178
compared to the powdered soil samples.
Table 4.22 Metals detected by XRF in original soils and corresponding carbons
Sample ID Elements detected
Soil STAOl
Carbon 002
S, K, Ca, Sc, Ti, Mn, Fe, Cu, Zn
S, K, Ca, Ti, Fe, Ni, Cu, Zn
Soil STA02
Carbon 005
S, K, Ca, Sc, Ti, V, Mn, Fe, Cu,
S, K, Ca, Ti, V, Fe, Cu, Zn
Soil STA03
Carbon 004
S, K, Ca, Sc, Ti, Mn, Fe, Ni, Cu, Zn
Ca, Ti, V, Fe, Cu, Zn
Soil STA04
Carbon 001
S, K, Ca, Sc, La, V, Cr, Mn, Fe
S, Ca Ti, V, Cr, Fe, Ni, Zn
Soil STA05
Carbon 006
S, K, Ca, Sc, V, Cr, Mn, Fe, Cu, Zn,
S, K, Ti, Fe, Zn
Soil G1
Carbon 045
S, Ca, Sc, Ti, Fe
Ca, Ti, Sc, Fe, Zn
Norit SA4
Chemviron Type C
Ca, Sc, Fe
Ca, Ti, Fe
Reference back to the analysis of the soils by acid digestion and ICP finish (table 3.5) indicates
the differences in detection ability of the two techniques employed, where elements such as Pb
and Hg, although present in elevated concentrations (up to 1189 ppm for Pb and 864 ppm for Hg)
were not detected by the XRF instrument. One of the reasons for this disparity is that it was more
difficult to detect elements of elevated atomic masses (such as Pb or Hg), because they gave
multiple XRF line spectra which were often obscured by other elements. This is a common
problem when examining the XRF traces for the 1st transition series of heavy metals like Cr, Mn,
Co, Ni and Cu. These elements produce peaks on the XRF detector at 5.494 keV, 5.898 keV,
6.93 keV, 7.477 keV and 8.047 keV respectively, but sandwiched between these relatively weak
signals are the strong twin peaks for a and jS Fe (located at 6.403 keV and 7.057 keV
respectively) and a and jS Zn (located at 8.638 keV and 9.571 keV respectively). Hence, the Fe
and Zn peaks frequently obscured the other elements. This effect can be seen in the XRF spectra
of the carbons runs 001, 002, 004, 005 006 and 045, which are shown in figures 4.47 - 4.53. The
particularly intense CI signal in figure 4.53 was due to the use of a PVC plastic in the sample
holder as the cell window. The Ar peak was an artefact of the discharge tube used in the
instrument.
179
The XRF results indicated that the metal species present in the contaminated soil prior to its
conversion into activated carbon were not lost by the procedure. It is considered that the metal
contamination of the soils become incorporated into the carbon lattice as it forms during the
carbonisation process. The Zn signal is much increased in magnitude due to the activation process
resulting in the inclusion of Zn into the activated carbon. The only detected metal which did not
appear to be retained by the treatment was Mn. The reason for the disappearance of the Mn is
unknown, but part of the explanation may be due to obscuration of a weakened Mn signal by the
Fe a signal thus preventing its detection. There is also evidence that Mn, in particular, is
susceptible to enhanced leaching after thermal treatment. A study performed by Helsel and Groen
(1989) for the US Gas Research Institute (USGRI) examined the effect of heating gaswork soils
in N2 or air at temperatures up to 600 °C. Subsequent leach testing of the soils showed that
generally, the thermally treated soil did not leach metals any more readily than if it were
untreated, except for Mn. However, the analysis of the wash liquors (section 4.4.7.2) did not
indicate that Mn was more susceptible to enhanced leaching, with only 1.49 % of the Mn in the
soil entering the HCl liquors. Helsel and Groen also noted that Hg and Pb were lost from the
soils by volatilisation mechanisms, although the Pb loss was unexpected. Mercury boils at 356 °C
(Weast, 1987) and would be emitted in the off-gases. It was not possible to confirm whether Hg
and Pb were in the carbonisation off-gases because means of safely trapping the gases were not
available.
Figure 4.47 XRF spectrum for run 001
I
11III n I r p i 11 {1111111II11
0 1 2 3 4 5
^ 1111 | i 11111111111 m '
6 7 8 9 10
180
Figure 4.48 XRF spectrum for run 002
Figure 4.49 XRF spectrum for run 004
' |T i i I | n Ti 'p 1111111 n 11 n 11111| 1111 | i 11 ^
0 1 2 3 4 5 6 7 8 9 10
181
Figure 4.50 XRF spectrum for run 005
I I I I I I ' l l I I I I 11 I I I I 11 I I I I I I I I I I I I I
0 1 2 3 4 5 6 7
Figure 4.51 XRF spectrum for run 006
F e a j
Z n a
1 A r T i FeP ^
I K 1
182
Figure 4.52 XRF spectrum for run 045
1111 III I
F e a
111 II I I II I
Z n a
Zn
0 9 10
4.5.2 XRD Results
The compiled XRD spectra for the soils and resultant carbons are shown in figures 4.53 and 4.54
respectively. The key to the legends used on these figures is given in table 4.23 below. The
spectra for each sample have been arranged to enable easy comparison, and only plotted between
20 and 55 26 since this region exhibited the major peaks of interest. Many compounds were
common to each sample. However, when assigning the peaks, the only ones marked were those
which could be assigned unambiguously.
Considering the soil samples first, the major crystalline component in all samples except STA04
was, as expected, silica. In general, the XRD patterns confirmed the chemical analyses: STAOl
and G1 contained 84600 and 122000 ppm sulphate respectively with the XRD confirming the
presence of large amounts of hydrated calcium sulphate. However, G1 also exhibited a sulphur
peak which was not expected. STA03 possessed 109800 ppm cyanide which the XRD implied was
a mixture o f F ^ O , and Fe^(Fe(CN)g)3. STA02 also possessed peaks in that region of the spectrum
which have been attributed to F^D; and Fe4(Fe(CN)6)3. STA05 only exhibited peaks for silica.
STA04 produced a spectrum which implied that it was almost pure iron nickel cyanide hydrate,
as there was an absence of any other crystalline phases. This was surprising, since this sample
possessed the lowest cyanide concentration according to chemical analysis. This result implied that
183
the chemical analysis was not capable of detecting this form of complex cyanide. Additionally,
this soil was known to contain a high level of sulphur (77800 ppm) but no sulphur signals were
detected by the XRD, suggesting that the sulphur was amorphous. Finally, there was a slight
displacement of peak positions noted for several of the samples, an effect quite clearly seen by
inspection of the silica peak at 26.6 26, which was due to the amorphous nature of the samples
(Gill, 1994).
Table 4.23 Crystalline phases detected by XRD
Compound Formula Powder index
number
Abbreviation
Hydrated Calcium Sulphate (Gypsum)
CaSO,. 2H2O 06 - 0046 HCS
Hydrated Calcium Sulphate (Gypsum)
CaSO,. 2H2O 36 - 0432 HCS2
Anhydrous Calcium Sulphate (Gypsum)
CaSO* 06 - 0226 ACS
Silicon Oxide (silica) SiOz 05 - 0490 SO
Iron Oxide (Hematite) FejOj 33 - 0664 10
Iron Cyanide (Ferric Ferrocyanide)
Fe,(Fe(CN)j3 01 - 0239 IC
Iron Nickel Cyanide Hydrate
Ni2Fe(CN)«.xH20 14 - 0291 INC
Sulphur S 24 - 0733 S
Zinc Sulphide (Wurtzite - 2H)
ZnS 05 - 0492 ZS
Carbon (not Graphite)
C 26 - 1077 C
Similar XRD results were reported by Helsel and Groen (1989) for as-received gasworks soil.
They identified silica as the major component of all their samples, with one of the samples
(heavily contaminated with sulphate) also showing traces of sulphur and hydrated gypsum.
The XRD traces for the carbon samples are shown in figure 4.54. The effect of the thermal
treatment and subsequent wash were quite profound. The only unchanged crystalline phase was
the silica. The hydrated gypsum peaks seen in STAOl and G1 disappeared from their respective
carbons, run Oil and run 045, although run 045 exhibited a weak anhydrous gypsum signal. The
iron oxide and cyanide signals were no longer detectable in runs 008 and 010. Finally, runs 007
and Oil developed new signals for zinc sulphide.
184
Figure 4.53 X-Ray Diffraction pattern for soils STA01-STA05 and G1
19600
O O LA
[counts] .
14400-
HCS2 SO
1 0 0 0 0 -
6400-
HCS2
HCS2 S
jA_L1 HCS2
SO HCS2 SO
61
INC
SO 3600-
l O IC
J
- s o
1 6 0 0 -
s o s o
- A —
INC
400-
INC
l o I C ^ s 8 ^ V V/v
IC l o l o
•• \J\^—/ - ...— ..
s o
STA05 k .
INC ST AO 4
SO s o .—..A.,.,,.
SO s o
l O s o l o
i STAOaj 'J
s o IC lOk STA02
STAOl
0 . 0 HCS SO SO HCS SO HCS HCS
jo ' do [•29]
Figure 4.54 X-Ray Diffraction pattern for carbons run 007 - run Oil and run 045
19600-,
[counts] -
14400-
1 0 0 0 0 -
6400.
3600-
1 6 0 0 -
400-
0 . 0
ACS SO
li Run 045
_ACS SO Run 009
SO so
Run 007
so so
Run 008
SO s o so so Run 010
p^«V»HyywTYv
SO so |l zs Run Oil
Figure 4.55 X-Ray diffraction patterns for the commercial carbons Norit SA4 and Chemviron Type C
6400
23
[counts] -
4900-
3600-
2500-
1 6 0 0 -
900-
400-
1 0 0 -
0 . 0
Norit SA
Type C
I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I I
I'o ' do ' Wo Jo so ['591
This result is particularly interesting, because it confirms the findings of Ibarra et al. (1991) and
the other authors discussed in section 4.3.1.1 with respect to the behaviour of ZnCl; in the
presence of sulphur. It also helps to confirm the earlier supposition that sulphur was formed from
the sulphate species in the soils during carbonisation and activation which could then react as
discussed previously (section 4.3.1.1).
The XRD traces for Norit and Type C are shown in figure 4.55. Both these samples are highly
amorphous and exhibit the typical traces for non-graphitising carbons which consists of two broad
peaks corresponding to the strongest peaks of graphite (Ergun, 1968; Masters and McEnaney,
1979). None of the soil carbons exhibited similar powder patterns.
4.6 Summary
Five different contaminated soil samples from the same gasworks site were converted into
activated carbon using different ZnCl^ loadings and by heating in an Ng atmosphere at
temperatures up to a maximum of 525 °C. The products were subsequently evaluated with respect
to the reduction in the contamination levels relative to the feedstock and their performance as
adsorbents. The components contributing to the contamination which were examined in detail
were: total cyanide, total sulphate and free sulphur. Adsorption ability was assessed using phenol
and 4-nitrophenol as ideal aqueous adsorbents. Nitrogen was used as the gaseous adsorbent to
determine surface area and pore-size distributions. After establishing the ZnCl; dosage which gave
the best adsorbents, the residence times for carbonisation and activation were varied to study their
effects upon the final product. The effects of these changes upon the contamination and adsorption
properties of the carbons were studied using the same parameters as above. Optimum residence
times were established and a bulk sample of the best carbon prepared and characterised for use
in the final stage of the work which comprised investigation of applications for the soil-derived
activated carbons. The feedstock and products were also assessed for crystalline phases by XRD,
and the presence of metals by XRF. Comparisons with commercial carbons Norit SA4 and
Chemviron Type C were made.
The following conclusions were made with respect to the observed results:
Thermal analysis studies showed that soil heated without ZnCl^ only exhibited two major thermal
events, one in the 100 °C region due to water loss, and the second between 150 and 275 °C due
to organic pyrolysis/carbonisation reactions. ZnCU addition caused a reduction in the magnitude
of the carbonisation peak and the creation of a third event in the 450 - 525 °C region. Usage of
188
ZnCI; increased the sample yield and the carbon content of the product whilst reducing its
hydrogen content. The XRD analyses demonstrated that the soils were amorphous with little
crystallinity. The major crystal phases were quartz, calcium sulphate and iron cyanides. Only
quartz was unaffected by the thermal processing. A reduction in the cyanide concentration of the
soil (> 90 %) was due to thermal cracking effects. The sulphate content of the soil was not as
readily removed by the processing, but was considered to be reduced to sulphur by the
carbonisation reactions and decomposition to HjS by the acid wash. Sulphur removal mechanisms
were through volatilisation, reaction with H; from carbonisation reactions and H^S production by
the HCI wash. However, the Zn from the ZnCIj acted as a retention mechanism for sulphur, by
combining with it to give ZnS. Minimal leaching of metal species from the carbon during the HCI
wash were noted and through XRF analysis metal retention within the carbons was seen to have
occurred. There was a large amount of Zn recovered firom the samples during the wash, as
expected. The presence of excessive amounts of sulphur species (sulphur and sulphate) caused the
development of adsorption properties in some samples without addition of activation agents. This
phenomenon was confirmed when addition of iron sulphate to soil was shown to be able act as
an activation agent.
The dosage of ZnCl^ which exhibited the greatest adsorption characteristics for phenol and 4-
nitrophenol was not the same for all the carbons, loadings of 33.33 % or 50 % (^/w) ZnClj
produced the highest adsorptions, but 33.33 % was chosen as the optimum dose. All the carbons
produced were of an "L" type. The adsorption isotherms for aqueous adsorption were amenable
to interpretation by the Langmuir and Freundlich isotherm equations. Gas adsorption data showed
that the carbons were essentially microporous in nature, with only one of the surface areas
exceeding those of the commercial activated carbons. The micropore surface areas increased with
greater ZnCl; additions. Only one of the carbons produced from the soil samples exhibited
adsorption properties comparable to the commercial activated carbons.
Changing the carbonisation and activation dwell times produced minimal effects upon the sample
adsorption ability and contamination reduction. Sixty minutes carbonisation and 120 minutes
activation produced the sample which exhibited the highest N; surface area and aqueous
adsorption ability. Increasing the mass of soil processed by a factor of three did not unduly affect
the aqueous adsorption ability of the product, although a slight reduction in surface area was
noted.
The conditions which resulted in the carbon which was the most effective adsorbent were 60
minutes carbonisation, 120 minutes activation with the addition of 33.33 % by weight of ZnClz.
189
CHAPTER FIVE
M E T A L ADSORPTION AND APPLICATION EVALUATION
5.1 Introduction
The carbons discussed in Chapter Four exhibited extremely variable adsorption properties which
were attributed to the heterogeneous nature of the contaminated soil feedstock. Only two of the
samples possessed greater than 50% carbon after processing. The major crystalline phase of the
samples was siliceous, as shown by the XRD analysis, which was essentially unchanged by the
thermal treatment. The adsorption ability of the carbons towards organic molecules was variable.
Due to the nature of the feedstock, the acceptance of these carbons in drinking water treatment
was considered to be unlikely, thus the application of this material to the treatment of industrial
waste streams, particularly those which contained toxic heavy metals, was considered to be a
suitable alternative use to investigate. Metals were chosen as the target waste steam because (i)
there has been very limited study of the use of activated carbon in metal adsorption compared to
organics adsorption; (ii) the carbons made by the ZnCl, activation procedure were L-type which
are the most suitable for adsorption of cationic species and (iii) the efficient removal of trace
concentrations of metal species is very costly. Hence, a systematic study of metal adsorption
capability by the soil carbon run 045 and the two commercial carbons were performed. The
carbons were subsequently investigated for their metal adsorption capability as a final polishing
step for metal plating effluent. Together with their established organic adsorption capability, the
potential of these carbons as additives to systems to control the composition of landfill leachates
(both for metals and organics) was also evaluated.
5.2 Selection of the metals for investigation
The EC directives controlling water quality were considered previously in section 1.2.3.2.1.
Selected metallic elements have been classified by the EC according to their toxicity and allocated
to either list I or list II. Maximum permissible concentrations in water were assigned to eleven
of these elements by the Drinking Water Directive (EC, 1980b), and these are detailed in table
5.L
Six metals were chosen from table 5.1 for initial study of their removal from aqueous solution
by activated carbon: Cd, Cr, Cu, Hg, Ni and Pb. From the previous discussion of landfill
leachate quality in section 1.6.2 (i), inspection of table 1.13 indicated that organic compounds
generally contributed the greatest part of leachate contaminant loads and the quantity of metals
190
present in leachates is highly variable. The pH of landfill leachate is largely dependent upon the
age of the landfill, previously discussed in section 1.6.2 (i), and since the aim of this work is to
demonstrate that the carbons can be used to control metal transport through a landfill system, it
was necessary to control the pH of the test solutions to reproduce the conditions within a typical
landfill. The interpretation of the literature data indicated that a pH in the range 5.4 to 8.6 would
be necessary.
Table 5.1 Maximum permissable concentrations in water for selected metals according to EU Drinking Water Directive (EEC,1980a)
Element List classification
Maximum concentration
(Mg/1)
As II 50
Cd & compounds I 5
Cr (VI) II 50
Cu II 100
Hg & compounds I 1
Ni II 50
Pb II 50
Sb II 10
Se II 10
Zn II 100
5.2.1 Sources of the metal-containing effluents
The landfill leachate samples were obtained irom a landfill in Essex. This landfill site is a co-
disposal operation, where domestic and industrial wastes are placed into individual waste cells,
rather than being intimately mixed. However, the combined leachate from the site is collected by
a perimeter drainage system which subsequently drains into a central lagoon, after which the
liquors are treated by chemical and biological means to adjust the COD, BOD, ammonia and pH
to levels acceptable for discharge into a near-by stream. Two samples were obtained from the
landfill, one from the lagoon and the other from a sampling well built into the main body of the
landfill.
The metal plating effluent originated from a zinc plating operation based in Surrey. The liquor
was derived from the vacuum filters used to de-water the ferric sludges produced by the
precipitation of the waste liquors from the pickling baths and the plating operations. This effluent
191
is normally discharged directly to the sewerage system under existing discharge agreements with
the area water company.
5.3 Experimental protocol
5.3.1 Introduction
The investigation of the soil carbon and the two commercial carbons for metal adsorption was
conducted in stages. As pH control was an important parameter, this was the concern of the first
stage of the work. Buffering of the solutions was investigated and a study of the effect of buffers
upon the metal concentrations in solution was undertaken. Single point adsorption experiments
were then conducted to investigate if all the metals exhibited an affinity for the carbon surfaces.
The most promising metals were then subjected to a study of the time taken for the adsorption
to reach equilibrium, followed by isotherm analysis as individual metals and finally as a mixed
metal solution. Finally, application of the carbons to real landfill leachate and liquors from a
metal plating plant operation were tested for their treatability by aqueous adsorption analysis.
5.3.2 Preparation of reaction solutions
5.3.2.1 Buffer solution
The buffer solution chosen was based upon the KH^PO^/NaOH system and was designed for
buffering solutions over the pH range 5.8 - 8.0. By varying the amount of NaOH added, the
desired pH could be achieved (Vogel, 1954). To 50 ml of 0.2 M KH^PO , 17.74 ml of 0.2 M
NaOH was added and the mixture then diluted to 200 ml total volume; the resulting pH was 6.6.
There are however concerns with the use of phosphate buffers. They are recognised as metal
complexation agents and can produce insoluble precipitates with metals. Additionally, the ICP-
AES system to be used for subsequent analysis of the metals required an HCl matrix and failure
to match the matrix will cause errors in the analyses. To this end, it was feared that the excess
of phosphate and sodium could cause matrix effects. Hence an initial study of the effect of adding
Cd(II), Cr(IH), Cu(II), Ni(II) or Pb(II) to the proposed buffer solution was performed to examine
whether: (i) the buffer would cause significant precipitation of the metals from solution and (ii)
the phosphate would cause significant matrix effects. Three metal concentrations were examined:
1 ppm, 10 ppm and 50 ppm. Two matrices were prepared: buffer + metal and buffer + metal
+ HCl.
192
5.3.2.2 Metal solutions
All the metals were utilised as 1000 ppm standard solutions (AAS grade, Fisons), with HNO3 as
the solution medium. The required volume of metal stock solution was diluted to 11 with the
buffer solution. This procedure was used for all the metal/buffer solutions required, with the
metal addition varied as required. For high concentration solutions (> 20 ppm), as used in the
isotherm analysis, the buffer solution was found to be incapable of maintaining the solution pH,
due to the excessive acidity introduced by the stock solutions. Hence, after examining different
solution systems, the most suitable isotherm solutions were found to be those where metal stock
solutions and distilled water were mixed to give the required metal concentration, after which the
solution pH values were adjusted to the desired value using 10 M NaOH.
5.3.3 Adsorption studies
5.3.3.1 Single point adsorptions
For these adsorption studies three carbons were investigated: Norit SA4, Chemviron Type C and
rRun 045) To a 125 ml acid washed glass bottle 50 ml of the required metal solution at 10 ppm
concentration and 0.5 g of activated carbon were added. After end-over-end rotation at 20 °C +!-
1 °C for an empirically chosen period of 5 hours, following which the solution was filtered off
the carbon through N° 1 filter paper (Whatman Ltd.) into an acid washed flask. The solution pH
was measured using a calibrated pH meter (RS Components Ltd.) with a glass electrode. Nine ml
of the filtrate were pipetted to a sample tube and 1 ml of conc. HCl was added to this sample,
except for Hg analysis where the acid was conc. HNO3. The solutions were refrigerated at 4 °C
until analyzed by ICP-AES. Each experiment was performed in triplicate.
5.3.3.2 Adsorption Kinetics
Using the same solution compositions as 5.3.3.1, solutions were analyzed after 15, 30, 45, 90,
120, 150, 180 and 240 minutes contact time with Run 045. The solution pH and residual metal
concentrations were determined.
5.3.3.3 Isotherm analysis
Isotherms were constructed using the method described in 5.3.3.1 above, using varied initial
metal solution concentrations of up to 300 ppm. Due to the problems with the buffer system
193
discussed above, it was decided to not use the buffer but to use two different approaches; ie a
series of metal solutions were adsorbed which had not been pH adjusted, whilst a second set were
adsorbed which had been pH adjusted. The solutions were modelled to the Langmuir and
Freundlich isotherms where possible.
5.3.4 Waste liquor analysis
5.3.4.1 Landfill leachate
The source of the leachate has been discussed previously (section 1.6.2(i)). The leachate was
digested using a USEPA method (USEPA, 1991b). To 100 ml of homogenous sample was added
2 ml of 1:1 HNO3, and 10 ml of 1:1 HCl. Without boiling, the sample was maintained at about
95 °C until the volume had reduced to 25-50 ml. The sample was then cooled and transferred to
a 100 ml volumetric flask and the volume made up to the mark with distilled water. If the
solutions did not settle, a portion was centrifuged prior to analyses by ICP-AES. Both filtered
(0.45 ^m WCN filter, Whatman Ltd.) and unfiltered leachate samples were analyzed and
analytical control spike solutions containing Ni, Cu, Cd, Pb, Cr, Mn, Zn, and Co were used to
check the digestion procedure effectiveness. Total Organic Carbon (TOC) analysis (Dorhmann
DC-80, Santa Clara, California) analysis and pH measurement were performed upon the as-
received unfiltered leachate from the lagoon and the deep well of the landfill. Mercury analyses
were performed upon the unfiltered leach at es only, using the digestion procedure described in
section 3.2.3.4.2.
5.3.4.2 Metal plating effluent
The pH of the plating liquor was determined upon receipt. The as-received liquor was not pre-
treated or digested prior to ICP-AES analysis, except that the matrix of the solution was modified
by adding 1 ml of conc. HCl (or HNO3 for Hg analysis) to 9 ml of liquor to matrix match the
ICP-AES calibration solutions.
5.3.4.3 Adsorption testing
When testing the landfill leachate and the metal plating effluent, 0.5 g of carbon and 50 ml of
liquor was used for each experiment which was performed in duplicate. Analyses were performed
as discussed in 5.3.3.1.
194
5.3.5 Leach testing
The leaching behaviour of the untreated contaminated soil, the carbons made from the soil and
the commercial carbons were studied. This was considered to be important because if the carbon
samples exhibited the potential to leach hazardous materials, then their application would be
limited. Secondly, the effectiveness of the process to convert the contaminated soil into activated
carbon could be evaluated by comparison of the leaching result.
Two leaching procedures were used: (1) The first test was a repeat of the single point adsorption
experiments described in 5.3.3.1: 0.5 g of carbon or soil was mixed with 50 ml of distilled water
(w/s ratio = 100:1) for 3 hours, and then filtered off. The pH of the leachate was measured and
the liquors analyzed for TOC and metals content. (2) The second test was a modified form of
DIN 38 414 (Deutsche Norm, 1984). This test is a German standard test for solid wastes, and
utilises a w/s ratio of 20:1. To 1.5 g of solid, in a 30 ml glass bottle, 15 ml of distilled water was
added. The mixture was shaken, end-over-end, at constant speed for 24 hours after which time
it was filtered to remove the solid. The filtrate was analyzed for pH, metals content and TOC.
5.4 Results and Discussion
5.4.1 Effect of buffer upon solution metal concentrations
Table 5.2 Effect of buffer upon metal concentration in solution
Metal concentration in ppm
stock 1 10 50 1 10 50 solution
Buffer only matrix Buffer and HCl matrix
Cr 108 111 110 107 102 104 100
% difference 2.78 1.85 -0.93 -5.56 -3.70 -7.41
Ni 113 111 111 110 105 108 106
% difference -1.77 -1.77 -2.66 -7.08 -4.43 -6.20
Cu 107 85.8 31.1 104 99.2 101 99.1
% difference -19.81 -70.93 -2.80 -7.29 -5.61 -7.38
Cd 95 90.3 108 138 101 107 128
% difference -4.95 13.68 45.26 6.32 12.63 34.74
Pb 113 90 96.1 107 104 104 103
% difference -20.35 -14.96 -5.31 -7.97 -7.97 -8.85
195
The results shown in table 5.2 were dilution corrected to enable direct comparison of the stock
solutions with the 1, 10 and 50 ppm solutions. The % differences indicate that the use of the
phosphate buffer did not interfere excessively with the solubilities of Cr(III) and Ni(II) although
Cd(Il), Cu(II) and Pb(II) were susceptible to the buffer. However, matrix matching the solutions
by the addition of HCl did help to enhance the solution concentrations of Cu(II) and Cd(II),
although the Cd(II) 50 ppm solution was still apparently affected by the buffer (34.74 % greater
concentration than the stock solution). The improvement in the accuracy of the detection was
probably due to two factors: (i) the increased acidity of the solution enhancing the solubility of
the metal ions; (ii) the matrix of the solutions matched the ICP standard solution matrix thus
improving the accuracy of the measurement. The observed deviations from the stock solution
concentrations were less than 9% for all the metals except Cd(II) (up to 35 % deviation at
50 ppm), but for the 10 ppm Cd(II) solution the deviation was only 13 %, which was considered
to be acceptable when considering that this included errors arising from the dilution stages and
instrument variability.
5.4.2 Single point adsorptions
The results of the adsorption of each metal from the buffered 10 ppm solution are shown in table
5.3. Inspection of these results indicated that Cu(II), Hg(II) and Pb(II) were susceptible to
precipitation effects, as shown by their depressed initial solution concentrations (5.2, 6.8, and
1.3 ppm respectively). Due to this effect further study of Pb(II) was discounted. Additionally,
further study of Ni(II) was considered unnecessary due to the poor affinity shown by the carbon
for this metal, since only 40.2 % was adsorbed compared to >98 % for the remaining metals.
Thus only Cd(II), Cr(III), Cu(II) and Hg(II) were studied further. The drop in the solution pH
was also noted, which was considered to be caused by the buffering capacity of the carbon (run
045), which had been previously identified as an L-type carbon. This pH drop also indicated that
the buffer was not capable of moderating the solution pH in the presence of the activated carbon,
and it was also evident that the pH had dropped from pH 6.6 to between pH 6.03 and 6.32 upon
addition of the metal ions to the buffer solution (see "initial pH" column in table 5.4). These
results meant that the application of the phosphate buffer was not going to provide the required
pH control in the adsorption isotherm experiments, and direct adjustment of the pH with a NaOH
would be required.
196
Table 5.3 Adsorption by run 045 of Cd, Cr, Cu, Hg, Ni and Pb from 10 ppm solutions
Concentration (ppm) pH
Metal initial final % adsorption initial final change
Cd 10.1 0.0 99.9 6.12 5.17 -0.95
Cr 10.0 0.1 98.9 6.03 5.22 -0.81
Cu 5.2 0.0 99.8 6.13 5.19 -0.94
Hg 6.8 0.0 98.8 6.32 5.64 -0.68
Ni 9.7 5.8 40.2 6.09 5.14 -0.95
Pb 1.3 0.0 98.4 6.14 5.22 -0.92 NB: Initial pH of buffer prior to metal and carbon addition was 6.6.
5.4.3 Adsorption kinetics
Figures 5.1 to 5.4 show how adsorption was very rapidly completed within the first 15 minutes,
with removals of > 99 % for all the metals under study. The final solution pH after adsorption
(pHf) was similarly equilibrated within this time, levelling off within a pH range of 5.2 - 5.7.
However, there was a slight deviation in pHf after 240 minutes on the Cd(II) and Hg(II) graphs
(figures 5.1 and 5.4 respectively), but the pH returned to its pre-240 minute level at 300 minutes.
The reason for this is unknown, since a similar effect was not seen with the Cr(III) or Cu(II)
studies, so it is possible that an experimental inconsistency was the cause of this effect. From
these results an adsorption time of 3 hours was chosen as the contact time, which was the same
contact time as used for the organic adsorption experiments.
Contact time study of Cd adsorption by carbon Figure 5.1
45 90 120 180
Contact time (minutes)
— Cd + p H f
210 240
g r w
197
Figure 5.2 Contact time study of Cr adsorption by carbon
45 90 120 180 Contact time (minutes)
210
-Cr - k p H f
Contact time study of Cu adsorption by carbon
45 90 120 180 Contact time (minutes)
210 240
-Cu -t-pHf
Contact time study of Hg adsorption by carbon
45 90 120 150 Contact time (minutes)
Hg 4-pH f
180 240
e/5 i -
%
240
Figure 5.3
300
Figure 5.4
300
f %
f %
198
Several researchers have reported the rapid adsorption of metal species from aqueous solution by
activated carbon, including Netzer and Hughes (1984) who noted that after 30 minutes the
equilibrium concentration of Cu(II) had been achieved. Abdel-Shafy et al. (1989) reported that
for Cd(II), adsorption was almost complete with 20 minutes of carbon contact with the solution,
but complete equilibrium was not considered to have been reached until shaking had occurred for
2 hours.
5.4.4 Metal adsorption isotherms
Adsorption isotherms for each of the metals were performed using solutions which covered a
concentration range of 10 to 300 ppm, using the following steps: 10, 20, 30, 50, 75, 100, 200
and 300 ppm. Comparison with the two commercial carbons utilised in the previous part of the
work were also undertaken.
5.4.4.1 Control of pH
As stated in section 5.4.2, the use of a phosphate buffer to control the solution pH during the
single solution concentration tests was not very successful, since by increasing the metal
concentration to 50 ppm, the buffering capacity of the system was rapidly consumed. This
observation was confirmed with a range of solutions at concentrations of 1 ,5 , 15, 20, 30, 50,
75 and 100 ppm which it was intended to use to construct an adsorption isotherm for Cd(II).
Using the 1000 ppm Cd(II) stock solution, 500 ml of the required solution concentrations were
prepared in the KHgPO^/NaOH buffer system. Table 5.4 shows how the solution pH varied as the
Cd(II) concentration was increased. Problems were immediately noted with this solution matrix;
the very low concentrations (1 and 5 ppm) formed a precipitate, whilst with the higher solution
concentrations (greater than 15 ppm), the pH rapidly decreased outside the desired pH range of
5.4 - 8.6 (table 1.13).
Table 5.4 Variation of solution pH in the buffered Cd isotherm solutions
Cd concentration (ppm)
1 5 15 20 30 50 75 100
pH 6.53 6.37 5.60 3.44 2.67 2.18 1.83 1.58
Thus, a revision to the solution matrix was required, and direct modification of the solution pH
with alkali was chosen. Secondly, to ensure that the starting solution pH was not precipitating
199
metals due to hydroxide formation, the effect of the solution pH upon the solubility of the metals
was studied. Thus, 50 ppm metal solutions were prepared from 1000 ppm standard solutions, and
the pH values were then adjusted using NaOH, to give solutions covering the pH range 1-10 in
unit steps. The solutions were filtered through N° 1 paper and the residual solution concentrations
determined by ICP-AES.
Figure 5.5 shows how the solution concentration of each metal in separated solutions varied with
pH. The onset of precipitation is clearly shown by each graph, from which it can be seen that
Hg(II) was the least susceptible to the effect of pH changes, with the residual concentration after
filtering remaining constant within the concentration range 48 - 53 ppm over the pH range
studied. Cadmium was the next least affected metal, with the solution concentration remaining
at 50 ppm until pH 7 was reached where a drop to 21.9 ppm was recorded. Finally, Cu(II) and
Cr(III) proved very prone to precipitation between pH 5 and 6. The Cu(II) concentration dropped
from 48 ppm at pH 5 to 34 ppm at pH 6 and similarly the Cr(III) concentration fell from 50 ppm
to 37 ppm. These results were similar to those reported by Koshima and Onishi (1986), for a
selection of metals which included Cd(II), Cr(III) and Cu(II),
Figure 5.5 Effect of pH upon solubility of Cd, Cr, Cu and Hg
Solution pH
-Cd -»Cr -*-Cu -Hg
From the above data, the only pH which did not result in large-scale removal of the metals from
solution by a precipitative mechanism and, most importantly, was within the literature range for
landfill leachate pH was pH 5.5, hence this was chosen as a suitable initial pH for the solution
adsorption studies. Two different solution matrices were tested: the first consisted of the solutions
without any pH adjustment and the second of solutions which were pH adjusted with NaOH, prior
to contact with the activated carbon. As an extension of this study, because all these metals could
be present in solution together in landfill leachate or metal plating effluent, a mixture of all the
200
metals was prepared with each metal present at a concentration of 50 ppm. The solutions were
adjusted to the pH values in the range 3-10 in the same way as for the single solutions, and
filtered prior to analysis by ICP-AES, the result of which is shown in figure 5.6.
Figure 5.6 Effect of pH upon solubility of Cd, Cr, Cu and Hg in a mixed solution
I S
1 1 "a
I 4 6 8
Solution pH
I Cd Cr Cu Hg"
Comparing figure 5.5 with figure 5.6, there were no major differences between the metal
behaviour as lone species and in mixed metal solutions. There was a fluctuation in the Hg(II)
concentration, an effect which may have been partially due to matrix effects in the ICP-AES. The
Cu(II), Cr(III) and Cd(II) exhibited the same trends with respect to their precipitation from
solution as single species and there did not appear to be any co-precipitation effects.
5.4.4.2 Cadmium adsorption isotherms
Figure 5.7 shows how the solution pH measured after contact with the activated carbons (which
were not adjusted with NaOH) varied across the concentration range used for the Cd(II)
adsorption isotherms with each of the carbons studied.
The data produced from these analyses were not amenable to interpretation by the Langmuir or
Freundlich equations. The isotherm plot (Ce vs x/m), which is not shown, exhibited a similar
shape to that shown for the pH in figure 5.7, with x/m decreasing as Ce increased (and the pH
increased). This showed how dependent the adsorption was to the solution pH and also implied
that physical adsorption mechanisms were the major cohesive forces operating for Cd(II)
adsorption by these activated carbons rather than chemisorption.
201
Figure 5.1 Variation of solution pH for Cd(II) adsorption isotherm using solutions without pH adjustment
4
%
1
100 150 200 250 Initial nominal solution concentration (mg/l)
350
•Blank - l y p c C Norit -*-Run 045
Figure 5.8 Variation of solution pH for Cd(II) adsorption isotherm using pH adjusted solutions
100 150 200 250 Initial nominal solution concentration (mg/l)
300 350
-Blank -«-TypeC Norit -#-Run 045
Adjustment of the solution pH prior to addition of the carbon (figure 5.8) had a profound affect
upon the pH of the solutions after adsorption and the extent of adsorption observed for each of
the carbons. The alkaline nature of the H-type Norit carbon can be clearly seen, with the solution
pH starting at pH > 9 and only slowly decreasing with increasing initial Cd(II) concentration.
The Type C and run 045 pH values were almost constant over the concentration range examined,
with only a very slight decrease in the solution pH being caused by run 045 carbon whilst Type
C caused a slight increase in the solution pH. These pH changes are a characteristic feature of
activated carbons which have already been considered. The important factor with the pH change
was whether the metal has been precipitated or adsorbed. It was very likely that the Type C and
Norit carbons caused some precipitation (if not total precipitation in the case of Norit) of the
metal species.Comparison of the pH plots in figure 5.8 with the pH/metal solubility plot for Cd
in figure 5.5 illustrates this point; at pH 7, only 25 % of the Cd from a 100 ppm solution remain
202
in the system. This problem does not exist with run 045 as the pH is below the point where
precipitation becomes an issue.
The isotherms for Cd(II) uptake are shown in figure 5.9, and all exhibited the characteristic type
I Langmuirian shape. The elevated x/m values exhibited by the commercial carbons Type C and
Norit could not, however, be wholly assigned to adsorption effects, as discussed above, where-as
the x/m value for run 045 was probably wholly attributable to adsorption effects.
Figure 5.9 Cd(II) adsorption isotherm determined using pH adjusted solutions
LOO 150 Ce(mg/1)
200 250 300
-Type C -•-Norit SA4 Run 045
Figure 5.10 Langmuir Isotherms for Cd(II) adsorption from pH adjusted solutions
Type C -•Norit SA4 -*-Run 045
203
Figure 5.11 Freundlich isotherms for Cd(II) adsorption from pH adjusted solutions
2
1.5
S 1
s 0.5
0
-0.5
Log Ce
TVpeC Norit SA4 Run 045
The Langmuir and Freundlich plots for the pH adjusted Cd(II) solutions are shown in figures 5.10
and 5.11 and the adsorption equation parameters are given in table 5.5. The run 045 data fits the
Langmuir model very well. The poor fit of the Langmuir adsorption equation to the data from
Type C and Norit (r = 0.84 and 0.88 respectively) when compared to the run 045 data (i =
0.99) is indicative of other influences upon the removal mechanism for the Cd(II). It is also
noticeable that the Freundlich Kp values were consistently lower than the Langmuir values.
The monolayer capacities were very variable between the carbons; Norit exhibited the highest
Qmax (37.04 mg/g) whilst Type C had the lowest (5.18 mg/g).
Table 5.5 Adsorption isotherm parameters for Cd(II)
Carbon Langmuir parameters Freundlich parameters Comments
r2 (mg/g)
B r2 Kp (mg/g)
1/n
run 045 0.99 9.90 0.99 0.94 3.37 0.39 pH adjusted to 5.5
Norit 0.88 37.04 1.50 0.79 10.21 0.27 pH adjusted to 5.5
Type C 0.84 5.18 1.17 0.96 2.36 0.22 pH adjusted to 5.5
The above results have been explained through consideration of the available literature. Huang
and Ostovic (1978) studied die uptake of Cd(II) by various commercially available activated
carbons, and they were particularly interested in the effect of the surface groups upon the
adsorption processes. They noted that pH influenced the carbon surface charge and that surface
charge characteristics varied among commercial carbons. A study of the kinetics of adsorption
showed that equilibrium was achieved within 30-60 minutes of carbon addition, which agreed with
the findings of this work (section 5.4.3). They stated that increasing the solution pH caused an
increase in the extent of Cd(II) adsorption by the carbons, irrespective of carbon type, but, they
204
also noted that several factors were considered to affect the adsorption process: pH, concentration
ratio and the physical-chemical characteristics of the adsorbate and adsorbent and the presence
of other solutes. By determining the pH at which zero electrophoretic mobility occurred (pH^)
for the carbons under study, they were able to interpret their results with respect to the changes
in pH which subsequently affected the mode of Cd(II)-carbon interaction. They proposed that the
main mechanisms of adsorption involved not only electrostatic attraction-repulsion but the
formation of surface complexes, because some adsorption was noted even at the pH^. Below the
point of pH^ the surface possesses a nett positive charge, whilst above it the surface has a nett
negative charge (Reed and Matsumoto, 1993).
Huang and Wirth (1982) studied the Cd(II) removal from an artificial plating water by activated
carbon. They noted that precipitation (of Cd(0H)2) only became a major consideration at pH >
9. These values are much higher than the findings of this work, but, the solutions they were
utilising did contain other salts which may have influenced the solubility of the Cd(II). They
considered both powdered and granular activated carbons, but concluded that powdered carbons
were much better for heavy metal removal with 3-4 times the adsorption capacity. Their
explanation of this was that powdered carbons had a more negatively charged surface than
granular carbon, hence attracted Cd(II) much more strongly. A similar result was reported by
Reed et al. (1994) who performed a comparative study using granular activated carbon (GAC)
colunms to remove Cd(II) or Pb(II) from solution. They found that at the pH of the experiment,
pH 5.4, Pb(II) was removed much more effectively than Cd(II). They attributed this to the
formation of hydroxides by the Pb(II), stating that Cd(II) would only precipitate out at pH 8.5-9.
In the review of inorganic adsorption by activated carbons by Huang (1978), he stated that L-type
carbons are much more effective adsorbents for metal adsorption than H-type carbons. This can
be related to the work on fungal adsorbents for Cd(II) (Huang et al., 1988) where they reported
that as pH^ increased, the extent of adsorption/removal of Cd(II) decreased. The lowest
adsorption for all the adsorbents studied occurred for Filtrasorb 400, a commercial granular
carbon which is an H-type carbon and also possessed the highest pH^ of the adsorbents studied.
Cadmium (II) adsorption by two commercial powdered activated carbons was studied by Reed and
Matsumoto (1993). They reported that the surface area of the carbon is not an important
parameter for determining the extent of adsorption of metals compared to organics, since the
greatest influence comes from the pH of the carbon surface and bulk solution. They found that
the lowest surface area carbon actually exhibited the highest Cd(II) adsorption, and that the
Langmuir value for each carbon increased with pH, a feature which was attributed to the
205
occurrence of a surface precipitation mechanism at increased solution pH values (where pH >
pH^). Their results for the two commercial carbons examined were similar to those reported in
this work for the Norit and Type C carbons, with Norit carbon (highest surface pH) exhibiting
much greater adsorption/precipitation of Cd(II) than Type C.
Kumar and Dara (1980) utilised various agricultural and food-derived wastes in metal adsorption
studies. They found that Cd([I) was adsorbed with varying degrees of success by wastes such as
bagasse, paddy husks, paddy straw, onion skins and garlic skins. Onion skins were most effective
at removing Cd(II), adsorbing almost 99% of the ions in solution, at pH 4.3. The Cd(II)
adsorption capacity of the onion skins was reported to be 7.9 mg/g.
Ferro-Garcfa (1988) examined the uptake of Cd(II) (and Cu(II) and Zn(II)) by activated carbons
prepared from almond shells (A), peach (P) or olive stones (H) by heating in CO . The carbons
were H-type, and exhibited a sharp increase in adsorption within the pH range 3-5, which was
assumed to indicate the region at which the pH^ occurred and the development of the negative
surface charge occurred. Consequently pH 5 was chosen for use during their adsorption
experiments. The carbons did not contain any sulphur, a feature which is unlike the carbons made
in this study, and is interesting because in the soil-carbons the presence of the sulphur is
considered to make a major contribution to the activation and the adsorption properties of the
carbons towards heavy metals, particularly Hg(II) (see section 5.4.4.5). These carbons exhibited
a very low affinity for Cd(II) (Q^^ values: A = 2.5 ppm, H = 5.91 ppm and P = 3.27 ppm)
which were the lowest for the three metals examined. The Cd(II) was expected to be adsorbed
more than the Zn(II), due to the smaller ionic radius (0.426 nm vs 0.430 nm respectively) but,
the lower affinity was explained in terms of the weaker polarising ability of Cd(II) compared to
Zn(II).
In summary, the literature results indicate that the carbon type (H or L), solution pH and
consequently the surface charge influence the extent of adsorption of Cd(II) by activated carbon.
Run 045 is an L type carbon, and although facilities were unavailable for determination of the
pH^ for this carbon, it clearly tends to affect solutions into which it is placed by turning them
acidic, so the pH^ is probably below pH 7. The adsorption mechanism operating for Cd(II)
probably comprises electrostatic interactions and surface complex formation but the type of
surface complexes formed will be dependent upon the surface species. Huang (1978) reported that
sulphonate-treated activated carbon exhibited improved heavy metal removal capacity, and since
the soil feedstock was rich in sulphur species, surface sulphur complexes would probably be a
major factor with respect to the non-electrostatic attraction of the Cd(II) to the surface, but, the
206
actual extent of the contribution of each mode of Cd(II) fixation is unknown.
5.4.4.3 Chromium adsorption isotherms
The first Cr solutions to be tested were not subject to pH adjustment. Figure 5.12 shows the pH
of the solutions after adsorption. The increased concentrations of Cr in the test solutions caused
the pH to drop to very acidic levels (pH 1). This in turn caused adverse effects with respect to
adsorption, with all of the carbons under test exhibiting minimal adsorption which could not be
modelled to either the Langmuir or Freundlich equations, as was the case with the analogous
Cd(II) adsorption results (section 5.4.4.2).
Figure 5.12 Variation of solution pH for Cr adsorption isotherms using solutions without pH adjustment
4
3.5
3
% 2.5 C3
•J 2
"5 1.5 ( / )
1
0.5
0 100 150 200 250
Initial nominal solution concentration (mg/l)
350
-Blank -"-Type C -<^Norit -»-Run045
Subjecting the solutions to pH adjustment prior to contact with the carbon resulted in the pH
behaviour shown in figure 5.13. This figure shows that the pH changes of the Cr solutions were
very similar to those for Cd(II) solutions (figure 5.8). The carbons caused the solution pH to vary
from the starting solution, with the H-type nature of Nor it increasing the pH to 8-9, but unlike
the Cd(II) solutions, the pH drop with increasing Cr concentration was not as pronounced as that
seen for the equivalent Cd(II) solution, with the highest concentration solution pH values
remaining above pH 8. The Type C carbon caused the pH to increase to approximately pH 6.5
whilst the run 045 solution pH did not vary greatly from that of the blank solution. Overall, the
pH behaviour of the Cr solutions were almost identical to the Cd(II) solutions (blank solutions
excluded). The increase in the solution pH values of Type C and Norit had an effect upon the
metal concentration in the test solutions. The Cr stock solutions were a purple/black colour which
was indicative of the solutions being predominantly composed of Cr([II) and not Cr(VI).
Chromium (III) is not as soluble as Cr(VI), a fact that has been illustrated by the pH/dissolved
207
ion graph in section 5.4.4.1 (figure 5.5) which showed that the Cr(III) in solution rapidly reduced
at pH > 6, due to precipitation effects. Further, the blank solution at a concentration of 300 ppm
tended to precipitate even at pH 5. Consequently, the 300 ppm solution data was not used in the
construction of the adsorption isotherms. The Langmuir and Freundlich plots for the Cr
adsorption are shown in figures 5.14 to 5.16 respectively and the isotherm constants calculated
from this data are given in table 5.6.
Figure 5.13
I J
Variation of solution pH for Cr adsorption using pH adjusted solutions
100 150 200 250 Initial nominal solution concentration (mg/l)
300 350
•Blank - -Type C •Nori t •»-Run045|
Figure 5.14 Langmuir isotherm for Cr adsorption by Run 045 from pH adjusted solutions
y c b
•Run 045
208
Figure 5.15 Langmuir isotherm for Cr adsorption by Type C and Norit SA4 from pH adjusted solutions
i
Tyf)e C -#"Nont SA4
Figure 5.16 Freundlich isotherms for Cr adsorption from pH adjusted solutions
Log Ce Type C -•-Norit SA4 Run 045
Table 5.6 Adsorption isotherm parameters for Cr
Carbon Langmuir parameters Freundlich parameters Comments
r • Q.« (mg/g)
B r2 Kp (mg/g)
1/n
run 045 0.98 3.08 0.72 0.94 0.48 0.49 pH adjusted to 5.5
Norit 0.95 45.46 2.20 0.86 23.77 0.56 pH adjusted to 5.5
Type C 0.98 6.14 6.04 0.96 3.60 0.39 pH adjusted to 5.5
The fit of the data to the adsorption equations was generally very good. The Q„^ of run 045 for
Cr is approximately one third the value of that for Cd(II), whilst for Norit and Type C
increases. The exceptionally high value of Norit (45.46 mg/g), when related to the high
209
solution pH, indicates that precipitation from solution was the major removal mechanism, rather
than adsorption. For Type C, it was possible that a mixture of adsorption and precipitation had
occurred.
The majority of previously published studies into Cr adsorption by activated carbons have been
based upon Cr(VI) removal. This is because Cr(VI) is very soluble and toxic to man and the
environment as a whole, being approximately 100 times more toxic than Cr(ni). Chromium is
an industrially important metal with many diverse uses including metal finishing, electroplating,
leather tanning and textile industries. The resulting waste waters are usually acidic and highly
oxidising. The method of Cr(VI) removal from solution by adsorption on to activated carbon
involves two mechanisms: the first is through physico-chemical adsorption and the second by
chemical reduction of the adsorbed species which produces Cr(III) (Kim and Zoltek, 1977; Alaerts
et al., 1989). Many workers have studied this phenomenon, such as Huang and Wu (1978) who
showed that Cr(VI) adsorption was affected by pH, with adsorption reaching a maximum at
approximately pH 5.5 and subsequently decreasing, so that at pH > 10 no adsorption occurred,
although reduction of Cr(VI) to Cr(III) did not occur at pH > 6. Srinivasan et al. (1988)
examined the uptake of Cr(VI) and Cr(III) by a carbon manufactured from rice husks using
H2SO4. They found that in the pH range of 2.0-3.0, maximum Cr(VI) removal occurred, but by
increasing the pH to > 6 the Cr(III) removal was >91%, a process which was probably due to
precipitation, although this was not stated. They also reported that in chromic acid solutions (pH
2.0-3.0) the carbon surface was slowly oxidised, resulting in the formation of Cr(III) from
Cr(VI). Jayson et al. (1993) reported that an activated charcoal cloth, which was microporous in
nature (95 % of surface area was associated with the micropores), exhibited higher adsorption
capacity for Cr(VI) than Cr(III). Maximum Cr(III) adsorption occurred at pH 4.5 and maximum
Cr(VI) adsorption occurred at pH 1, which is essentially in agreement with previously reported
work. The difference between the adsorption of Cr(III) and Cr(VI) by this carbon was attributed
to (i) the relative dimensions of the hydrated metal ions in solution and (ii) the nature of the
electrostatic forces between the charcoal cloth and the Cr ions. The hydrated Cr(III) ion has a
radius of 0.9 nm and the Cr(VI) a radius of 0.5 nm, consequently the Cr(VI) could penetrate the
micropores more readily than the Cr(III). They proposed that the increased Cr(III) adsorption with
increasing Cr(III) solution concentration was due to the loss of H O from the hydration shell
allowing greater entry of the Cr(III) to the micropores. The surface charges of the cloth
(pH^ = 2.7) varied with pH as did the pH of the hydrated Cr(III); for Cr(III) the positive charge
was neutralised by the inclusion of OH in the hydration shell in place of water, as shown below
(Jayson e? a/., 1993);
210
pH3.Z pm.O pHS.Q
[CriH^O)^^* ^ 5 [Cr(H^O),OH]^* ^ 5 ICKH^OUOID^r^^l^^CriH^OUOH)^-
H* H* 2H^*
The Cr(VI) ions possess a negative charge at all pH values hence the Cr(VI) ions were repulsed
at pH > 2.7, whilst the Cr(III) was attracted until the nett charge of the hydrated ion was
negative. Thus it was suggested that the major factor controlling adsorption of the Cr species from
solution by the charcoal cloth was the accessibility of the ions to the micropores. (Jayson et
of., 1993).
Most work on Cr adsorption has been performed using commercially available carbons, which
are predominately H-type. The H- and L-type carbons behave differently towards Cr. H-type
carbons possess a high pH^ (Huang and Bowers, 1978), their surfaces being positively charged
according to the following equilibrium; CxO + ZH O # CxOHj "" + 20H" (Huang and Wu,
1977). Consequently they are considered to be the most suitable carbons for adsorbing Cr(VI)
from solution by attracting the dominant Cr(VI) species, bichromate (HCrO/) and dichromate
(CrjOv ")- These are thought to react with the positively charged carbon surface and are
subsequently reduced as the carbon surface is oxidised, resulting in Cr(III) formation (Huang,
1978). The Cr(III) formation is especially rapid at pHs between 2.0 and 3.0 which are also the
most suitable for Cr(VI) adsorption. Hence, it is very difficult to prevent the reduction occurring
although Huang and Bowers (1978) proposed that if the reacting solution was maintained in an
oxidsing state the Cr(III) formation would be prevented. H-type carbons do adsorb some Cr(III)
below pH 6, with the maximum adsorption at pH 5 (Huang and Wu, 1977). Huang and Wu
(1977) noted that total Cr removal by activated carbon was not possible because of the reduction
step which released some Cr(III) back into solution, the re-solubilization being most intense at
the pH most effective in Cr(VI) adsorption (pH 2.5), although the Cr(III) could be removed at
a later stage by secondary precipitation. To overcome this problem, Huang and Bowers (1978)
proposed that two activated carbon beds could be operated in sequence, the first containing an H-
type carbon which would remove the Cr(VI) and also produce some Cr(III) by reduction. The
second bed would contain L-type carbon which would subsequently remove the Cr(III).
Srivastava et al. (1989) used a carbon made from the waste generated from fertilizer production
by oxidation in H^O; followed by heating to carbonise and finally air activation. The carbon
exhibited a strong affinity for Cr(VI) with maximum adsorption of approximately 260 mg/g
occurring at pH 2, but increasing the pH caused the adsorption to decrease, a result which has
been explained above and observed by many workers.
211
L-type carbons attract positive ions and two possible forms of the equilibrium were proposed by
Corapcioglu and Huang (1987) :
2(C0H) + # (C0)2\f+ + 2H+ or
2(C0) + # (COO M^
Consequently Cr(III) should be readily adsorbed by L-type carbons at pH values up to the point
of precipitation. The carbon surface could also be expected to act as a seed point for precipitation
once pH 6 is exceeded. Cr(III) solution behaviour is considered to be governed by the dominant
form of the Cr in solution, which has been reported to be Cr(OH) ^ in the pH range 2.8-6.0,
Above pH 6.0 Cr(IH) precipitates as Cr(0H)3 (Huang and Wu, 1977). McKay et al. (1985) found
that the adsorption of Cr(III) by Filtrasorb 400 commercial activated carbon at a pH > 5.7 was
affected by precipitation, which "masked" the true adsorption.
Bautista-Toledo et al. (1994) examined Cr(III) and Cr(VI) uptake by a carbon which was
subjected to surface oxidation by HNO3. They found that although the surface area of the carbon
decreased, as determined by N; adsorption, the capacity of the carbon for Cr(III) increased. This
was attributed to the acidic surface oxide groups formed by the oxidation, dissociating to provide
conjugate bases to which the Cr(III) would be attracted to. This implies that the oxidised carbon
was an L-type, compared to the un-oxidised carbon which was reported to have positively charged
surfaces and was an H-type. They also examined the effects of competing ions upon the
adsorption of Cr. Using the un-oxidised carbon, in the presence of NaCl, the Cr(III) adsorption
was relatively immume to ion effects, but the Cr(VI) adsorption decreased as the NaCl
concentration increased. This indicated that CI" was competing with the Chromate species for the
carbon surface. They also noted that for the oxidised carbon Cr(VI) adsorption increased, but
since the pH of that system was pH 2.6, the futher oxidiation of the carbon surface by Cr(VI) and
the production of Cr(III), as discussed above, was suggested as the cause of this phenomenon.
Although only "total" Cr was measured in this research, the main form in the solutions was
Cr(III). The Norit commercial carbon was an H-type carbon which, although it was observed to
removed very large amounts of Cr, the pH of the isotherm solutions (pH 8.14-9.37) suggest that
this removal was by precipitation. The Type C carbon exhibited mixed H- and L-type properties,
possessing a = 6.14 mg/g. The solution pH range was pH 5.85-6.56 which suggests that
this was composed partially of precipitation and partially adsorption, although precipitation
probably predominated. The run 045 carbon, an L-type carbon, was most suitable for Cr(III)
adsorption. The solution pH range was pH 4.63-5.90 which was within the expected solubility
range of Cr(III). The for run 045 was 3.08 mg/g, the lowest value of all the carbons, but
this value could at least be attributed entirely to adsorption of Cr and not another removal
212
mechanism. Thus, the dominant form of Cr in the test solutions was Cr(III). Run 045 adsorbed
Cr(III) from the solutions, whereas the commercial carbons reduced the concentration of Cr(III)
in solution but by precipitative means rather than adsorption.
5.4.4.4 Copper adsorption isotherms
The pH of Cu(II) solutions contacted with the carbon without pH adjustment are shown in figure
5.17. The observed solution pH behaviour was the same as that for Cr(III) and Cd(II) discussed
in sections 5.4.4.2 & 5.4.4.3 respectively. Only run 045 exhibited different adsorption behaviour,
with the adsorption increasing with concentration whilst the pH decreased, as shown in figure
5.18.
Figure 5.17 Variation of solution pH for Cu(II) adsorption isotherms using solutions without pH adjustment
100 150 200 250 Initial nominal solution concentration (mgyl)
300
• Blank TypeC •Nor i t -»-Run045
These data for run 045 were amenable to Langmuir and Freundlich analysis with the resulting
plots shown in figures 5.19 and 5.20 respectively and the calculated isotherm constants given in
table 5.7. The increased adsorption with pH was not the expected behaviour, because the
increased concentration of would have been expected to react with the available surface
functionalities, decreasing the Cu(II) uptake. This result thus implied that attractive forces other
than simple Van der Waals or electrostatic were operating, for example, surface complexation
may have been a factor.
213
Figure 5.18 Ce vs x/m for Cu(II) adsorption from solutions without pH adjustment.
20
-SO
150 200 Ce(ing/1)
300
-T^pe C -*-Norit SA4 -•-Run 045
Figure 5.19 Langmuir isotherm for Cu(II) adsorption f rom solutions without pH adjustment
1.4
Run 045
Figure 5.20 Freundlich isotherm for Cu(II) adsorption f rom solutions without pH adjustment
1.6
1.4
1.2
1
0.8
S 0.6
0.4
0.2
0
-0.2
Log Ce
Run 045
214
Figure 5.21 Variation of solution pH for Cu(II) adsorption isotherms using pH adjusted solutions
1
I (/>
100 150 200 250 Initial nominal solution concentration (mg/1)
-Blank •••TypeC •Norit Run 045
Figure 5.22 Langmuir isotherms for Cu(II) adsorption f rom pH adjusted solutions
1
T y p e C - » N o r i t S A 4 -* -Run045
200
Figure 5.23 Freundlich isotherms for Cu(II) adsorption from pH adjusted solutions
4
lypeC -a-Norit SA4 -*-Run 045
215
The pH behaviour of the pH adjusted Cu(II) solutions contacted with the carbons is shown in
figure 5.21. The pH of the Cu(II) solutions after each experiment exhibited similar behaviour to
those of Cr(III) and Cd(II), except that run 045 solutions were at a slightly higher pH (~ pH 6.6)
than observed for Cd(II) (pH 5) or Cr(III) (pH 5 - 6). From the study of the effect of pH upon
metal solution concentration (section 5.4.4.1; figure 5.5), Cu(II) starts to precipitate from solution
in the pH range 5 - 6 , thus indicating that for run 045, precipitation must be playing a part in the
removal mechanism. Norit, being an H-type carbon, exhibited characteristically high solution pH
values which would also imply that precipitation was the major removal mechanism for these
solutions. Only Type C exhibited pH values which were below those coinciding with the onset
of precipitation. These data were interpreted by the Langmuir and Freundlich equations, and the
resulting linear plots are shown in figures 5.22 and 5.23. The respective constants from the
equations are given in table 5.7.
The run 045 Cu(II) value for the pH adjusted solutions was much greater than the non-
adjusted, ie 20.00 mg/g compared with 7.04 mg/g respectively. This increased result was assumed
to be indicative of the extra influence of precipitation upon the reduction in the residual solution
concentration. The Norit result would have been similarly influenced by precipitative effects. The
mathematical nature of the Langmuir and Freundlich equations are such that they cannot separate
adsorptive and precipitative removal mechanisms when reductions in the adsorbate concentration
occur at pH values over the onset of precipitation. Only the reduction in Cu(II) concentration by
Type C, which exhibits the lowest value at 5.78 mg/g, can be unambiguously assigned to
removal by adsorption.
Table 5.7 Adsorption isotherm parameters for Cu(II)
Carbon Langmuir parameters Freundlich parameters Comments
r2 Q.« (mg/g)
B rz Kp (mg/g)
1/n
run 045 0.95 7.04 6.64 0.98 3.40 0.302 no pH adjustment
Norit - - - - - - no pH adjustment
Type C - - - - - - no pH adjustment
run 045 0.86 20.00 12.50 0.62 58.61 0.82 pH adjusted to 5.5
Norit 1.00 14.08 23.67 0.97 39.45 0.57 pH adjusted to 5.5
Type C 0.90 5.78 50.88 0.98 4.12 0.16 pH adjusted to 5.5
There are relatively few studies in the literature concerning the adsorption of Cu(II) by activated
carbons. There has been much more interest in the adsorption of Cu(II) by clays and minerals,
216
probably due to the importance of Cu(II) as both an essential component for growth and as a toxin
to plants. McLaren and Crawford (1973) in their study of 23 different soils concluded that the
major contributors to Cu(II) adsorption by soil were not considered to be the clay or Fe oxide
components, but the soil organic matter and the Mn oxides, and that the most suitable pH at
which to measure this adsorption, without interference by precipitation reactions, was pH 5.5.
Netzer and Hughes (1984) reported that Cu(II) adsorption on carbon commences in the region of
pH 3 with maximum adsorption occurring at pH 5 - 6. These observations confirm that the pH
chosen for this work, pH 5.5, was a suitable value to use to study the Cu(II) adsorption by the
soil-carbon and the other commercial carbons.
The relative contributions of each soil fraction towards the Cu(II) uptake is important, because
it implies that the Cu(II) adsorption exhibited by run 045 was probably wholly attributable to the
carbon component of the adsorbent rather than the mineral part. However, the adsorption
exhibited by run 045 in the low pH region (figures 5.16 and 5.17) is not as readily explained. The
pH of the solutions in contact with run 045 varied from pH 2.85 - 0.94, which were below the
pH at which adsorption is considered to commence. Corapcioglu and Huang (1987), in an
investigation of the adsorption of several metal species, noted for Cu(II) that L-type carbons
tended to adsorb more Cu(II) than the H-type carbons at pH values below the p l ^ . The optimum
pH for Cu(II) removal was reported to be between pH 3-5. Increasing the temperature of the
adsorption system from 25 °C to 100 °C caused an increase in the extent of adsorption. This
finding was opposite to that expected since adsorption is an exothermic process. Their explanation
was that the H-type carbon (Filtrasorb 400) was being oxidised as the temperature increased,
forming an L-type carbon which had increased metal ion adsorption capability due to electrostatic
attraction. However, they then further stated that electrostatic forces probably played a minor rOle
in the adsorption process, because they calculated that the free energy of bonding was 4 kcal/mol,
which was indicative of the formation of outer sphere complexes (hydrogen bonds) between the
hydrated metal species and the carbon surface oxide groups. This proposed adsorptive mechanism
may have been significant at the low pH values in this work, and consequently the influence of
the high H^ concentration (which would normally swamp the adsorption sites in acid solution),
would not have had an effect upon the adsorptive processes because the major interactions would
have been between the hydration shells of the carbon functional groups and the metal ions in
solution.
The effectiveness of other adsorbents for Cu(II) have been examined by Kumar and Dara (1980)
who, in their investigation, discussed previously in 5.4.4.2, reported that onion skins were the
most effective waste for removing Cu(II) from solution, with 83.4 % removal from an initial
217
solution of concentration of 40 ppm, at a final solution pH = 4.0. However the adsorption
capacity of the onion skins for Cu(II) was 3.76 mg/g, which was lower than some of the other
wastes used. However, the pH of the solutions in contact with the other waste materials which
exhibited superior adsorption properties compared to the onion skins were relatively high and
were all above pH 5.7. At this pH precipitation of the Cu(II) would be expected to commence,
a feature which could be excluded from the solutions treated by the onion skins, hence implying
that adsorption was the major process. Similar results were reported by Srivastava et al. (1989)
utilising their carbons made from the waste slurry from fertilizer manufacture. They reported that
Cu(II) was only poorly adsorbed by their carbon (2.8 mg/g) with maximum adsorption occurring
at approximately pH 6. They observed that the adsorption reaction was exothermic, as increasing
the temperature of the adsorption experiments decreased the Cu(II) adsorption.
Ferro-Garcfa et al. (1988) examined Cu(II) adsorption by activated carbons prepared from
agricultural wastes but the carbons all exhibited a much greater affinity for Cu(II) than Cd(II).
The values for Cu(II) for the three carbons were: A = 8.32 mg/g, H = 9.22 mg/g and P
= 7.21 mg/g. The reason for these larger values was attributed to the greater accessibility
of the Cu(II) ions to the pores because the Cu(II) ion (radius = 0.419 nm) is smaller than the
Cd(II) ion (radius = 0.426 nm). In terms of surface coverage of the carbons, however, the Cu(II)
only occupied 5 % of the total surface area. The possibility of metal-carbon surface complex
formation was considered too, with the consideration that Cu(II) forms more stable surface
complexes than Cd(II), which would enhance its uptake. Relating this supposition to this work,
the fact that Cu(II) ions were removed from the acidic solutions by run 045 and this adsorption
was amenable to interpretation by the isotherm equations, (whereas the Cd(II) results under
similar conditions were not), provides extra evidence that outer sphere complex formation was
probably operating.
Thus, in summary, the Cu(II) adsorption by run 045 and Type C were within the ranges typically
observed for other waste-derived carbons and commercial carbons, assuming that the solution pH
values were below the onset of precipitation. Above pH 6, precipitation was a major factor
affecting the residual solution concentration of Cu(II) which could not be differentiated from
adsorption. With Norit SA4 it was impossible to say if it had adsorbed any Cu(II) because
solution pH values were greater than the onset value for precipitation. Huang (1978) reported that
at pH 4-5, the optimum pH for Cu(II) adsorption, H-type carbons exhibit insignificant adsorption,
due to repulsion by the positively charged carbon surface. Consequently it is very unlikely that
Norit adsorbed any Cu(II). Run 045 exhibited substantial adsorption at acidic pHs, which was
considered to imply that outer sphere complexes were forming between the hydrated Cu(Il) ions
218
and the hydrated carbon surface complexes. Type C appears to have a low Cu(II) capacity
(5.78 mg/g), which indicates that Type C has L-type characteristics.
5.4.4.5 Mercury adsorption isotherms
The behaviour of the Hg(II) solutions without pH adjustment followed the same trend as for
Cd(ll), Cr(III) and Cu(II), as shown by figure 5.24. The adsorption isotherms for Hg(II) uptake
by all the carbons are shown in figure 5.25. Only Type C exhibits any levelling off of the
adsorption, with a maximum x/m value corresponding to 12.3 mg/g. Norit SA4 and run 045
exhibited adsorption capacities of 25.5 mg/g and 35.8 mg/g respectively. These were not,
however, their maximum adsorption capacities because the highest initial solution concentration
of 300 ppm was completely adsorbed by run 045, and Norit SA4 had not achieved an adsorption
saturation value, as shown by the slope of the Norit data line in figure 5.25.
Figure 5.24 Variation of solution pH for Hg(II) adsorption isotherms using solutions without pH adjustment
100 150 200 250 Initial nominal solution concentration (mg/l)
-Blank C -"-Norit -#-Run 045
It was decided that increasing the solution concentration beyond 300 ppm was not necessary
because at such elevated concentrations, more efficient methods of Hg(II) removal exist, such as
sulphide precipitation. The data for adsorption from solution without pH adjustment collected for
Norit and Type C were modelled to the Langmuir and Freundlich isotherm equations, the plots
of which are shown in figures 5.26 and 5.27 respectively. Equation constants are summarised in
table 5.8. The data fi-om run 045 was not used to calculate the constants due to the paucity of data
points and their very poor fit to the isotherm equations. Both the commercial carbons exhibited
good degrees of fit to the isotherm equations with values of 15.87 mg/g and 9.09 mg/g for
Norit SA4 and Type C respectively.
219
Figure 5.25 Hg(II) adsorption isotherms determined using solutions without pH adjustment
I
Ce(ing/I)
- lypeC -*-NontSA4 -0»-Run 045
Figure 5.26 Langmuir isotherm for Hg(II) adsorption from solutions without pH adjustment
2
lypeC -"-Nont -o-Run 045
Figure 5.27 Freundlich isotherm for Hg(II) adsorption from solutions without pH adjustment
Log Ce
lypeC -#-Nont SA4 -0-Run 045
220
Figure 5.28 pH of Hg(II) solutions after adsorption, initial pH adjusted to 5.5
10
100 150 200 250
Initial nominal solution concentration (mg/1) 300
-Blank -•-TypeC -•-Norit -#-Run 045
Adjusting the solution pH to 5.5 before adsorption did not result in any major changes in the
extent of adsorption by the carbons. Each carbon affected the solution pH slightly, although the
pH of both Type C and run 045 solutions followed the pH of the blank solution very closely, as
shown in figure 5.28. Norit maintained an alkaline pH for all the solutions. Due to the
insensitivity of the concentration of Hg(II) to changes in pH, the residual equilibrium solution
concentrations were virtually identical to those determined for the solutions with unadjusted pHs.
Consequently, there was a similarity between the shapes and the magnitudes of the adsorption
isotherm plots determined from solutions with and without pH adjustment, although Norit
exhibited x/m values which were depressed somewhat when pH adjustment had been applied, as
evident from a comparison of figures 5.25 and 5.29. The resulting Langmuir and Freundlich
isotherm plots for the pH adjusted solutions are shown in figures 5.30 and 5.31, and the constants
for these plots are summarised in table 5.8.
Figure 5.29 Hg(II) adsorption isotherms determined using solutions with pH adjustment
100
Ce(mg/1) 200
-TypeC Norit SA4 Run 045
221
Figure 5.30 Langmuir isotherm for Hg(II) adsorption from pH adjusted solutions
5
I
TypeC -«-NontSA4
The and Kp values for Type C (table 5.8) were very similar for both the adjusted and
unadjusted solutions (9.09 & 3.21 mg/g and 9.26 & 3.87 mg/g respectively). No comparable
or Kp values were available for run 045 and, as noted in table 5.8, the values reported
(28.57 mg/g and 18.24 mg/g) are very conservative. They are the largest and Kp values
reported for any of the metal uptakes by the carbons studied. Finally, as noted previously, Norit
SA4 exhibited different adsorption behaviour for the solutions with and without pH adjustment.
The Qmax and Kp values for the adjusted solutions were approximately five times lower than the
values reported for the unadjusted solutions. The reason for this is unclear, but, it was reported
by Fife et al., (1985) that if cation exchange is involved in the adsorption mechanism, adsorption
capacity decreases with increasing pH. Hence in this work, Norit may have been ion exchanging
Hg(II).
Figure 5.31 Freundlich isotherms for Hg(II) adsorption from pH adjusted solutions
Log Ce
Type C Norit SA4 Run 045
222
Table 5.8 Adsorption isotherm parameters for Hg(II)
Carbon Langmuir parameters Freundlich parameters Comments
rz (mg/g)
B r2 Kp (mg/g)
1/n
run 045# - - - - - - no pH adjustment
Norit 0.95 15.87 0.75 0.96 5.13 0.36 no pH adjustment
Type C 0.98 9.09 0.49 0.88 3.21 0.25 no pH adjustment
run 045* 0.93 28.57 2.69 0.62 18.24 0.33 pH adjusted to 5.5
Norit 0.80 3.56 0.39 0.82 0.84 0.42 pH adjusted to 5.5
Type C 0.91 9.26 1.26 0.98 3.87 0.31 pH adjusted to 5.5
solutions. Adsorption was also very intense. The isotherm had not reached its saturation point, thus
is very conservative
The major difference with Hg(II) and the other metals considered in this study was shown to be
its ability to remain in solution, even at high pH (section 5.4.4.1). This resistance to precipitation
by mercury, coupled with its very high toxicity, are the major reasons why alternative efficient
removal methods for this metal are required. It is usually separated from the aqueous phase by
precipitation as HgS, ion exchange, coagulation by aluminium or iron and adsorption on to
activated carbon (Huang and Blankenship, 1984). Schuster (1991) produced a review of the
behaviour of mercury in soil. He noted that mercury tends to form complexes with anions such
as CI', OH", S ", P, and N and the organic ligands which contain S. In fact, sulphur was the
preferred ligand, even in the presence of 0 ", and the soft acid nature of mercury (Pearson, 1967)
was considered to be the cause of this behaviour (Schuster, 1991). Covalent interaction are the
most important effects in mercury binding. In soil solutions, Hg(II) exists predominantly as
uncharged complexes, which implies that adsorption on to the solid phase is by formation of
insoluble complexes with inorganic or organic groups and not through CEC (Schuster, 1991). The
organic matter is of primary importance in the binding of mercury in soil as mercury is mostly
found associated with this fraction, due to the presence of S in the organic matter. However, the
type of binding which occurs is affected by the system pH. Lindberg et al. (Schuster, 1991)
reported that contaminated soil from near a mercury mine exhibited mercury distributions 200
times greater in the organic fraction than the cation exchangeable fraction. Andersen (Schuster,
1991) reported that inorganic Hg in acid conditions (pH < 4.5 - 5) was only bound by the
organic fraction, whilst as pH increased to neutral conditions, the iron oxides and clay fractions
became more important for uptake. This would indicate that the adsorption properties of the
223
carbons, especially run 045, for Hg(II) are linked with their surface species. As sulphur is a
dominant species of the soil carbon, complex formation will probably occur followed by
precipitation onto the carbon surface.
Ma et al. (1992), using HgClj as the Hg(II) source, found that the presence of CI' affected the
pH at which maximum adsorption was observed; pH 7 compared to pH 4-5 for other workers.
The means of Hg(II) retention was also considered to be different, with an extra mechanism
involving the physical trapping of species such as HgClj or Hg(0H)2 within the carbon
micropores. The fertilizer waste slurry carbon produced by Srivistava et al. (1989) exhibited very
high removal capacities for Hg(II) (as Hg(N03)2), with uptakes of over 600 mg/g at 27 °C.
Maximum uptake occurred at pH 2, although the adsorption decreased as the solution pH
increased.
It was reported in the review by Huang (1978) that ZnCl -activated carbon adsorbed HgCIj better
than HNO3 modified carbon, and that decreasing the solution pH increased the Hg(II) adsorption.
Their conclusion was that the method of activation caused greater influence upon the ability of
a carbon to adsorb Hg(II) than the initial feedstock. Hence, the use of ZnCl; in the preparation
of run 045 may have influenced the affinity of the product for Hg(II). Huang and Blankenship
(1984) studied 11 activated carbons for Hg(II) adsorption over a wide pH range (pH = 2.5-11).
They noted that most samples gave a maximum Hg(II) adsorption at pH 4-5 (similar to soil),
which decreased as the pH increased or decreased. Acidic conditions were the most favourable
for Hg(II) adsorption. The mechanism proposed for this by Huang and Blankenship involved an
adsorption/reduction system. They suggested that the Hg(II) diffused into the carbon where it
underwent reduction upon contact with the surface oxide groups, forming Hg° before diffusing
out of the carbon into the bulk solution. The type of carbon used affected the rate at which this
reaction occurred: L-type carbons were very effective at Hg(II) adsorption, whilst H-types were
not. However, their results were not directly comparable to the ones for this work because the
L-type carbon (run 045), displayed very high adsorption capacities which were outside the pH
range they reported to be optimum for Hg(II) adsorption. Also, the H-type Norit SA4 was very
effective at Hg(II) adsorption and even though the adsorption capacity decreased as the pH
increased, implying that cation exchange was occurring, the adsorption capacity was still greater
than the L-type, Type C carbon. These conflicting results would appear to support the discussion
of Schuster (1991), that surface complexes and functional groups play the most important r61e in
Hg(II) adsorption.
224
Reconsidering the previous findings reported by Huang (1978), concerning the effect of the mode
of activation upon Hg(Il) adsorption ability and relating this to the work of Huang and
Blankenship (1984), further discussion as to why activated carbons prepared from ZnC^ should
exhibit higher affinities for Hg(II) than carbons prepared by other activation routes is possible.
The ability of Hg° to form amalgams with other metals such as Zn may have been a factor in the
increased adsorption observed for run 045. The suggested mechanism of this reaction route is:
Zn° is produced in the reducing conditions experienced during carbonisation and activation of the
carbon feedstock. This metal is retained in the carbon after preparation (as observed in the XRF
plots discussed in 4.5.1), whilst the surface of carbon is susceptible to oxidation processes, as
discussed above and reported by Huang and Blankenship (1984). Hence, the production of Hgf
at the carbon surface followed by amalgamation of the Hg° formed with Zn° in the carbon lattice
would potentially result in an increased Hg(II) adsorption capacity being observed.
Thus, in summary, the carbon prepared from the contaminated soil exhibited extremely high
adsorption capacities for Hg(II) adsorbing in excess of 35 ppm of Hg(II) per gramme of carbon.
The proposed mechanisms for Hg(II) removal by run 045 involved two definite mechanisms: (i)
reduction of the Hg(II) to Hg° followed by amalgamation of the Hg° with the Zn° in the carbon
lattice produced by the ZnClj activation method and (ii) attraction of the Hg(II) to sulphur species
incorporated in the carbon surface functional groups.
5.4.4.6 Adsorption isotherm for a mixed metal solution
Figure 5.32 illustrates the pH behaviour of the metal solutions after contact with the activated
carbons. The Type C and run 045 carbons exhibited similar pH values which were almost the
same as the blank pH value for solutions of initial concentration > 100 ppm. The fact that all the
solutions after carbon contact exhibited pH values greater than pH 6 had important effects upon
this study of the adsorption from solutions of mixed metals. The presence of precipitates after
preparation of the stock solutions indicated that co-precipitation was a factor which was of
importance in these mixed solutions.
The concentration of each metal in the 8 different starting solutions was measured and are
presented in figure 5.33. The observed precipitation had severely affected the Cr(III) and Cu(II)
concentrations, with residual concentrations of below 10 ppm for all the solutions. The Cd(II) and
Hg(II) solution concentrations were also affected. The Cd(II) precipitation was no greater than
that recorded for the single element solutions considered in 5.4.4.1 (figures 5.5 and 5.6) at the
observed solution pH. However, the reduction in the Hg(II) concentrations was wholly
225
unexpected. The 10 ppm and 20 ppm solutions were unaffected by the mixed metal system, but
increasing the concentration to 30 ppm caused a reduction in the Hg(II) concentration of
approximately 25 %. Much greater decreases were observed as the required Hg(II) concentration
in the system increased to 300 ppm, where the measured Hg(II) concentration was only 88.5 ppm,
70 % lower than expected.
Figure 5.32 Variation of solution pH for the mixed metal solutions after adsorption
10
8
E A. 6 g •ri
•3 4 00
2
0
V
50 100 150 200 250 Initial metal concentration (mg/l)
300 350
- Blank pH •'lypeCpH -•-NoritpH -*-Run045pH
Figure 5.33 Initial concentrations in mixed metal solution
I I
I <n
200
150
100
Cu Cd Metals in solution
Nomina] metal concentration prior to pH adjustment (mg/l) s i o [1320 ^ 3 0 ra5o fflioo [110200 m s o o
The reason for these reductions was attributed to the co-precipitation of Hg(II) by the Cr(III) and
Cu(II) precipitates, which were of a colloidal nature and could be expected to trap the large,
hydrated Hg(II) ions (radius = 1.01 A ) within these precipitates as they formed. Ma et al. (1992)
reported that Fe(III) added to the carbon/Hg(II) system caused an enhanced Hg removal, but as
226
no separation of precipitation effects of the Fe(III) hydroxide gel was performed, it was probable
that co-precipitation of the Hg(II) by the Fe(III) occurred. Iron precipitation is a commonly used
industrial procedure, called "ferric floccing" for removing heavy metals from solution. The
consequences of these effects were that only Cd(II) and Hg(II) were present in solution at
sufficient concentrations to enable determination of the adsorption isotherm parameters for the
three carbons in the presence of each metal simultaneously.
Figure 5.34 Cd(II) and Hg(II) adsorption isotherms for the mixed metal solutions.
20
60 80
Ce(mg/1)
-TypeC(Cd) * Norit (Cd) -*-Rim045(Cd) -T^e C (Hg) Norit (Hg) -e- Run 045 (Hg)
140
The Cd(II) and Hg(II) adsorption isotherms for each carbon are shown in figure 5.34.
Comparison of these results with those obtained for the individual metals (Cd(II) - figure 5.9 and
Hg(II) - figure 5.29) shows that the Cd(II) adsorption by Norit and Type C were virtually
unchanged whilst run 045 exhibited a reduced adsorption, with the x/m plateau occurring at
approximately 7.5 mg/g, compared to approximately 15 mg/g in figure 5.9. The Hg(II) adsorption
by Norit and Type C shown by figure 5.34 exhibited a slight reduction compared to the single
element solutions given in figure 5.29. Run 045 still exhibited very intense adsorption
characteristics for Hg(II), irrespective of whether adsorption was occurring from single or mixed
solutions. The data for all the carbons adsorbing Cd(II) and Hg(II), except for Hg(II) adsorption
by run 045 due to insufficient points, were plotted according to the Langmuir and Freundlich
equations. These plots are shown below and the isotherm constants for each plot are summarised
in table 5.9.
227
Figure 5.35 Langmuir isotherms for C d ^ D and HgOO adsorption by Type C
8
Cadmium -^-Mercmy
Figure 5.36 Langmuir isotherms for Cd(II) and HgOO adsorption by Norit SA4
Cadmium •«-Mercuiy
Figure 5.37 Langmuir isotherms for Cd(II) and Hg(II) adsorption by run 045
Single data point
Cadmium Mercury
228
Figure 5.38 Freundlich isotherms for simultaneous Cd(II) and Hg(II) adsorption
25
s
-TypeC(Cd) •NoritSA4 (Cd) ^Rim045(Cd)
-Type C (Hg) Norit SA4 (Hg) Rim 045 (Hg)
Table 5.9
Carbon
Isotherm parameters for Cd(II) and Hg(II) adsorption from the mixed metal solution
Langmuir parameters Freundlich parameters Metal
r2 Q:« (mg/g)
B r2 Kp (mg/g)
1/n
run 045 0.99 6.21 0.92 0.79 2.41 0.28 Cd(II)
Norit 0.98 18.18 2.62 0.94 12.02 0.47 Cdai)
Type C 0.97 8.13 0.04 0.99 1.61 0.29 Cd(II)
run 045# - - - - - - Hgai)
Norit 0.52 2.47 1.27 0.80 1.11 0.29 Hgai)
Type C 0.97 9.01 0.64 0.92 3.52 0.34 Hg(II) # The Hg adsorption was so intense that Ce was zero for all but one of the data points.
The effect of competition for the adsorption sites by using Cd(II) and Hg(II) simultaneously has
been evaluated by using the values determined for adsorption of the individual metals by
each carbon from the pH adjusted solutions, which were given in table 5.5 for Cd(II) and table
5.8 for Hg(II) and table 5.9 for the mixed system. This data is summarised in table 5.10 for ease
of comparison.
From table 5.10, it can be observed that for the mixed metal solution the extent of Cd(II)
adsorption by run 045 and Norit SA4 decreased. However, the adsorption of Cd(II) by Type C
increased. The adsorption of Hg(II) by Norit SA4 and Type C from the mixed metal solution
decreased slightly, whilst the uptake of run 045 was still so great that the accurate determination
of isotherm constants was not possible. At the pH of the adsorptions, it is probable that
precipitation was the major removal mechanism for Cd(II) in contact with Norit SA4, whilst with
229
both run 045 and Type C the Cd(II) and Hg(II) appear to have utilised different locations upon
the carbon surfaces to adsorb. With Type C, the Cd(II) results suggested that it competes with
Hg(II) for some of the adsorption sites, or the Hg(II) may block neighbouring Cd(II) sites.
Table 5.10 values calculated for Cd(II) and Hg(II) adsorption from single and mixed metal systems
Carbon Metal Q m a x values (mg/g) Carbon Metal
Single metal solution Mixed metal solution
Run 045 Cd(II) 9.90 6.21
Type C Cd(n) 5.18 8.13
Norit Cd(II) 37.04 18.18
Run 045* Hgai) 28.57 -
Type C Hgai) 9.26 9.01
Norit Hgai) 3.56 2.47
ofHgai)
The discussions on the adsorption of the individual metals by each carbon considered the current
adsorption models for Cu(II), Cr(III), Cd(II) and Hg(II). The precipitation of Cu(II) and Cr(III)
simplified the adsorption system. Most other workers who have studied mixed metal systems have
tended to consider two or three metal systems. Netzer and Hughes (1984) found that in a mixed
Cu/Pb/Co solution, the metals appeared to compete for the same adsorption sites, with Cu(II)
being the least affected by the presence of the other metals. Arulanantham et al. (1989) examined
Cd(II) (and Pb) adsorption by a coconut-based carbon and a commercial coal-derived carbon. In
the presence of Cu(II), Zn(II) and Ni(II), the adsorption capacity of these carbons for Cd(II) (and
Pb) was extensively reduced due to competition by the other cations for the reactive sites on the
carbon surface. Srivastava et al. (1989) utilising a waste-derived carbon showed that competitive
adsorption did occur between different metals. They used a mixture of pairs of Cr(VI), Pb(II) or
Hg(II); with Pb(II) and Hg(II) paired, adsorption reduced for both metals, whilst Hg(Il)
adsorption decreased when Cr(VI) was present, but Cr(VI) adsorption was unchanged. Thus,
competition between metals for adsorption sites on carbons can be problematic, but from the
results presented above for run 045, the adsorption of Cd(II) and Hg(II) simultaneously by the
soil carbon did not appear to be much affected by competitive adsorption effects.
230
5.4.5 Analysis of the landfill leachate
Analysis of the two leachate samples are presented below. The first sample was taken from within
the landfill body (Deep well), the second from a central lagoon (Lagoon) where all site leachates
were collected prior to treatment and off-site discharge. The pH and the total organic carbon
content of each sample, determined within 24 hours of collection are summarised in table 5.11.
Table 5.11 pH and TOC value for the landfill leachate samples
Sample pH TOC (ppm)
Deep well 6.91 1609
Lagoon 8.09 621
The deep well sample contained over twice the TOC content of the lagoon. This was indicative
of the degradative mechanisms operating within the landfill, producing organic species as the
wastes were decomposed by bacterial and chemical actions. The lagoon sample had a much lower
TOC value due to the chemical, physical and biological attenuation effects taking place in the
landfill and perimeter collection systems. Comparing the TOC and pH values above with the
summarised literature data in table 1.13 (Section 1.6.2 (i)), the TOC and pH deep well values
shown above are typical for an established landfill (literature average values: TOC = 1788 ppm;
pH = 6.76).
The metal content of the filtered (0.45/im filters) and un-filtered leachates are reported in table
5.12. Filtering was used to evaluate the influence of the organic components upon the metal
species within the landfill. The metals most affected by the filtering were Ba, Cu, Mn, and Zn.
This result suggests that these metals were the most readily complexed by the organic components
in the leachate. The concentration values for the metals present in the leachate can be conveniently
compared with those from the literature, given in table 1.13 (section 1.6.2 (i)). The concentrations
of metals in the leachate were generally very low, approaching or below the detection limit of the
ICP spectrophotometer. The efficiency of the digestion procedure was also evaluated by use of
spike additions and calculation of the percentage recoveries. Each spike solution contained Cd,
Co, Cr, Cu, Mn, Ni, Pb and Zn at levels of 50 ppm (high spike) and 10 ppm (low spike). The
reagent blanks produced spike recoveries within 5 % of the dose, indicating tiiat the digestion
method was sufficiently efficient for the analyses of these metals in aqueous solution.
231
Table 5.12 Landfill leachate metal analysis
All values are in ppm
Metal Lagoon Deep well Metal
Filtered Unfiltered Filtered Unfiltered
Metal
Correction value to be added to the concentration shown in brackets
Ba 0.41 0.44 0.24 0.70
Cd < dl < dl < dl < dl
Co 0.04 (0.01) 0.04 (0.01) 0.04 (0.01) 0.04 (0.01)
Cr 0.07 (0.01) 0.07 (0.01) 0.07 (0.01) 0.08 (0.01)
Cu 0.01 (0.00) 0.10 (0.00) 0.01 (0.00) 0.11 (0.00)
Hg nd nd nd 0.03
Mn 0.34 (0.04) 0.35 (0.04) 0.43 (0.05) 1.51 (0.17)
Mo 0.03 0.03 < dl < dl
Ni 0.26 (0.05) 0.27 (0.05) 0.25 (0.04) 0.24 (0.04)
Pb < dl < dl < dl < dl
V 0.03 0.03 0.16 0.16
Zn < dl 0.13 (0.03) 0.07 (0.02) 1.07 (0.26) nd = not determined
<dl = below detection limits
In figures 5.39 and 5.40 the spike recoveries from the Lagoon samples and the Well samples,
respectively, are represented. All the metals showed a depression of their values, which varied
from an insignificant 0.19 % loss of spike (99.81 % recovery) for the low spike of Cu from the
unfiltered lagoon sample (LUSL), to 24.28 % loss of spike (75.78 % recovery) for the high spike
of Zn from the unfiltered well sample (WUSH). The spike recoveries decreased in the following
order of metals: Cu > Cr = Mn > Co > Pb > Ni > Cd > Zn. One whole set of low spikes
for the filtered Well sample (WFSL) were apparently lost in the digestion procedure, but after
comparison with the performance of the previous spikes, it was considered that by oversight the
spike had not been added to the sample. Overall, Cu was the metal which was most readily
extracted and analyzed, giving the best spike recovery in all of the different solutions. From the
inspection of the spike recoveries observed for the different samples, all of the concentrations
were depressed by a certain amount. By averaging the percentage recoveries calculated for
individual metals in each spike, correction factors for each of the metals used in the spike solution
were determined. These factors have subsequently been used to calculate correction values which,
when added to the metal concentration detected in the leachates, provided a metal concentration
which was considered to be representative of the actual concentrations present in the samples.
232
These values are shown in table 5.12, in brackets, beside the concentrations determined by ICP.
In general, the metal levels in these leachates were much lower than for those reported in the
literature.
Figure 5.39 Spike recoveries for Lagoon samples
120
i *o
a
N i Cu
\ i e t a ] in sp ike
SLFSH CmLFSL ISaUJSH raUJSL
Note. L Lagoon: U — Unfiltered: F — Filtered: SH = High concentration spike added: SL = Low concentration spike added.
Figure 5.40 Spike recoveries from Well samples
120
100
8 0
&• 60
u s 40 a
20
0
-20 Cr M n Co N i Cu
Metal in sp ike
Z n C d P b
iWFSH nmWFSL EWUSH SaWUSL
Note: W = Well: other letters as for Figure 5.40.
5.4.6 Adsorption testing with the aqueous wastes
5.4.6.1 Landfill leachate
The effectiveness of the commercial and the soil carbons in reducing the metal and organic levels
233
present in the lagoon landfill leachate are shown in figures 5.41 and 5.42. The concentrations of
metals in the leachate were generally very low when compared to those used in the previous
adsorption studies of this work. These low concentrations caused some problems with the
determination of the metals and the subsequent evaluation of their adsorption by the carbons. For
most of these elements their concentrations in solution were approaching the detection limits of
the ICP. When this occurs, the relative errors in the detection of the signal from the sample when
analyzed are much increased, a fact which can result in apparently increased metal concentrations
being reported in the leachate samples after exposure to the carbon. This situation arose with Co,
where both Norit and run 045 exhibited slightly increased residual concentrations.
Figure 5.41 Concentrations of metals and TOC present in the leachate sample before and after contact with the carbons
1000
I Q
1 8
1 Pe M n M b
Spec ie s adsorbed
Z n T O C
:Initial concentrat ion EDType C ( p p m ) E N o r i t S A 4 (pptn) M R u n 045 ( p p m )
Figure 5.42 Percentage reductions in the concentrations of metals and TOC in the leachate sample after adsorption by the carbons
1 0 0
8 0
« 60
• s 4 0
2 0
0 Ba C o Cr Cu Fe M n
Spec ie s adsorbed
N i V Z n T O C
n y p e C limiNorit S A 4 ^ R u n 045
Overall, the commercial carbons exhibited higher adsorptions of metal species than run 045 for
all the metals studied except Fe and V, which were reduced by 80.56 % and 46.88 %
234
respectively. Run 045 did exhibit some limited leaching of metals into the landfill leachate. The
levels of Cu, Mn and Zn all increased, with the increase in Zn being of most concern. The
increased Cu and Mn concentrations were very small at 0.11 and 0.85 ppm respectively. Both
of the final concentration values reported in figure 5.41 are well below those recommended by
the California State Authorities for metal concentrations in leachates ( < 5 ppm for Cu and 350
ppm for Mn) (Helsel and Groen, 1989). Leaching of Mn has been also observed by Helsel and
Groen (1989) who reported that after thermal treatment of soil from gaswork sites in air or N;
there were increased levels of Mn released by the soils, but the other metals studied (including
Cu) were not similarly affected. The enhanced Zn level can be wholly attributed to the method
of activation followed by insufficient washing of the carbons to remove the residual Zn. In section
4.4.7.2, the effect of successive washing of the carbon with HCl indicated that the first wash
removed most of the Zn but subsequent washing still resulted in further Zn removal. Thus the
Zn leaching should be remedied by a two or three stage washing procedure for the carbons.
The reduction in the TOC of the leachate by run 045 (13 %) was the same as the adsorption of
phenol by run 041 (13 %) which was prepared under exactly the same conditions as run 045,
(section 4.4). This result suggests that the active surface area of run 045 was accessible to
sufficient of the leachate organic species to become saturated. This result was of interest because
landfill leachate could be expected to contain much larger and more complex molecules than
phenol, and would not be expected to be adsorbed to the same extent. With the commercial
carbons, the expected lowering of adsorption was observed, with Norit SA4 removing only 68 %
of the TOC compared to 76 % for phenol whilst Type C exhibited a greater reduction in
adsorption ability by removing only 71 % of the TOC compared to 95 % of the phenol (table
4.5).
The literature data available for application studies of adsorbents to leachates is very scarce. Only
one reference was located which considered using adsorbents as in-situ treatment for landfill
systems. Chan et al., (1977) considered that rather than allowing landfill leachate to seep from
the depositories, after which it is collected and treated, incorporation of adsorbent systems into
the lining materials of the landfill should result in the removal of the toxic species thus negating
the need for leachate treatment. They manufactured their own leachates from industrial sludges;
a metal plating sludge, a CaFj sludge typical of those from electronics and aircraft manufacture
and a petroleum sludge. A range of adsorbents were studied: activated alumina, activated carbon,
bottom ash, cullite, fly ash, illite, kaolinite, Ottowa sand, vermiculite and natural zeolites.
Characteristics of the leachates which were measured included: TOC, anionic species and cationic
species. Two types of adsorption test were performed, a static test and a dynamic test. They found
235
that the same leachate under dynamic testing (laboratory scale lysimeters) exhibited better
removals than under static testing. As would be expected, all the adsorbents studied exhibited
variable affinities for the species present in the leachates. For example, the maximum organic
removals were observed with activated carbon and fly ash whilst activated alumina removed
Cu(II), Ca(II) and F most effectively. In conclusion they suggested that since no one adsorbent
removed all the species of concern from the leachates, the production of mixed liner material
which comprised fractions exhibiting each of the desired adsorption characteristics was the best
solution.
5.4.6.2 Electroplating effluent
The effect of carbon upon the residual metal levels in an effluent from a zinc plating company
was also examined, the results of which are shown in table 5.13. The effluent was produced by
the neutralisation and precipitation of the plating liquors from the acid baths of the plating
process. These particular liquors arose from vacuum filtration used to remove excess water from
the precipitated sludge prior to the disposal of the sludge to landfill, rather than the main flow
from the settling tanks.
Table 5.13 The treatment of an electroplating effluent with activated carbon
ID All concentrations are in ppm (% reduction in brackets) pH ID
Ca Cd Cr Cu Hg Zn
pH
Filter liquors initial conc.
23.90 0.00 1.28 0.01 0.00 0.24 10.13
Type C 6.2 0.08 1.20 0.00 0.00 0.02 9.20 (74.06) (6.10) (93.88)
Norit 23.4 0.05 1.23 0.00 0.00 0.00 9.89 (2.09) Ck52) (100)
run 045 283.00 0.04 0.63 0.00 0.00 0.00 7.31 (50.47) (100)
For comparative purposes, the analysis of the electroplating sludge was performed using EPA
digestion method 3050 (USEPA, 1986), the results of which are presented in table 5.14. The
effectiveness of the precipitation treatment for purification of the effluent was clearly
demonstrated. The levels of metals present in the effluent were very low, as evident in table 5.13.
Almost 100 % of the metals were retained in the sludge (table 5.14), which was very rich in Fe
and Zn, with minor amounts of Cr, Cu and Cd.
236
Table 5.14 Analysis of the metal plating sludge
Metal Ag Cd Cr Cu Fe Ni Pb Zn
Concentration (ppm or %)
5.5 7.6 8416 3265 45% 189 210 25%
The major metal species present in the effluent was Ca, which arose from the lime used as the
pH adjusting and precipitation agent. The next highest concentration of metal present was for Cr,
but the concentration of 1.28 ppm was very low. It is not possible to say which Cr species was
present, but considering the pH of the solution, it would suggest that the most likely species was
the Cr(VI) ion. However, with reference to the previous discussion of the adsorption behaviour
of carbons towards Cr(III) and Cr(VI) (Section 5.4.4.3), the Cr concentration was most
effectively reduced by the L-type carbon run 045, hence the species being adsorbed must have
had a positive charge ie, Cr(III) rather than HCrO/ or Cr O? ' and/or was being removed from
solution by precipitation methods. Further complications to this wastewater system may also have
been the presence of complexation agents such as EDTA, NTA, citrate, tartrate and ammonia,
which are considered to reduce the efficiency of the precipitation method employed to remove the
metals. However, the quality of the untreated effluent given in table 5.13 does not appear to
indicate that there were any significant effects attributable to the influence of anions causing a
deterioration of the efficiency of the precipitation mechanism. Ku and Peters (1987) reported that
anions can influence adsorption on H-type carbons by: (i) accumulating on the surface of H-type
carbons forming multiple layers; (ii) enhancing the approach of metal ions to the solid surface and
(iii) assisting in complex formation by metal ions with pre-adsorbed anions. Thus adsorption of
cations by H-type carbons can be enhanced by the presence of anions. In these experiments,
however, the removal of Cr from solution by the H-type commercial carbons was an order of
magnitude lower than for the soil carbons, which would imply that there were minimal extraneous
anions (excluding OH") present in the effluent. They also reported that anions such as , HS and
CN can adsorb on to and block the surface of H-type carbons to metals, although this reduction
in sulphide and cyanide concentration was considered beneficial since their removal from aqueous
solution can be problematic, and the threat of formation of HjS or HCN is minimised. Thus, it
appears that a specific interaction between the soil carbons and the Cr species in solution resulted
in its removal and that the species in solution was most probably Cr(III).
Other effects which were notable for the electroplating effluent were that all the carbons leached
Cd, which was worst for the commercial carbons, and that there was no Zn leaching from
run 045 when used to treat this effluent. This was particularly significant, because the Zn detected
in the filter-press liquors was being removed by run 045. This was in contrast to the liquors in
237
contact with Type C where there was still some Zn remaining in solution (Table 5.13). The only
problem with using run 045 to treat the electroplating effluent was that some Ca leached from run
045, which was probably due to the dissolution of the gypsum recorded in that sample by XRD
(Section 4.5.2, figure 4.54).
5.4.6.3 Summary
Different effects were observed for the use of the commercial and soil-derived carbons for the
treatment of these two aqueous effluents. This should be expected since the effluents had differing
compositions. It was noticeable, however, that these effluents contained much lower levels of
metal species than the literature data implied. Additionally, the composition of the landfill leachate
tended to cause some leaching of metal species from the carbon into the solution. This is in
contrast to the plating effluent, which did not exhibit any leaching potential. The results from
these experiments indicate that the soil-derived carbons are suitable adsorbents for treating
electroplating effluent.
5.4.7 Leach testing of the soils and carbons
The data from the two leach tests performed upon the soils and carbons examined in this work
are shown below. The three hour test mimicked the adsorption experiments, except that distilled
water was used as the liquid rather than leachate or metal solutions. Distilled water is a rather
aggressive medium because it is essentially free of salts, but does contain a large proportion of
dissolved CO;, which makes the water acidic. The pH and TOC data for this leach test are shown
in table 5.15. The blank comprised distilled water without solid addition and was used to correct
for pH/TOC changes due to the effect of the apparatus; these corrections are included in the
tables. The overall result of the test was that the carbons (Norit, Type C, run 045 and run 006)
did not exhibit any appreciable leaching of organic components into the distilled water and the
pH change was similar to that seen in the metal adsorption experiments discussed in Sections 5.4-
5.4.4.6 previously. The behaviour of the soils, however, clearly illustrated the benefit of the
carbonisation and activation procedure. Soil G1 was the source of run 045, and produced a very
acidic solution and a reasonably high TOC (19.79 ppm). Soil STA05, the source of run 006
discussed in chapter 4, gave an alkaline pH but leached far more organic components.
Consequently, carbonising and activating the contaminated soil eliminated completely the leaching
of organic components from the soil.
238
Table 5.15 pH and TOC results of the 3 hour leach test.
Sample pHf pHf - pHi Corrected TOC corrected pH (ppm) TOC
Blank 6.10 0.54 - 2.71 -
Norit 9.56 4.00 9.02 1.92 -0.79
Type C 7.00 1.44 6.46 1.56 -1.15
run 045 5.73 0.17 5.19 1.75 -0.96
run 006 6.61 1.05 6.07 1.89 -0.38
G1 2.94 -2.62 2.4 19.79 17.08
STA05 7.81 2.25 7.27 47.40 44.69
pHi = Solution pH before leach test: pHf = Solution pH after leach test
The metals leached by the commercial carbons are presented in figure 5.43 and by the raw soils
and the carbons prepared from them in figure 5.44. Metal leaching from the commercial carbons
was not a major problem. Calcium was the only metal leached in any quantity, and that was from
Norit. There was also minor leaching of Zn and Hg from Type C and Norit respectively.
Figure 5.43 Metals leached from Type C and Norit after the 3 hour leach test
10
& I I S
J 0.0001
Ca H g
Meta l s l eached
i i y p e C raNorit
For the soils and their associated carbons, Ca was the major element leached, but this was
reduced by converting the soil into carbon. The major source of the Ca was considered to be the
gypsum which was present in the XRD of G1 and run 045 discussed in section 4.5.2. Zinc was
excessively leached from run 045. The inadequacy of the post-production HCl wash on this
sample was to blame for this problem. In contrast, the Zn leaching from run 006 was several
orders of magnitude less than that for run 045, illustrating that thorough washing with acid
minimises this problem. Finally, increased leaching of Mn was observed for Gl/run 045, a
239
feature of thermally treated gas works soil which was reported by Helsel and Groen (1989). Run
006 does not exhibit this increase in leaching due to the apparently superior acid washing which
it received.
Figure 5.44 Metals leached from the raw soil samples and the soil-carbons after the 3 hour leach test
1 0 0 0
y 1 0 0
I 10
i a 1 o
'a
1 0.1
c! 0.01
0 .001 C o Cr Cu
Meta l s l eached
ISTAD5 S r u n 0 0 6 M G l fflRun 045
The TOC and pH data for the DIN leach test is given in table 5.16. It can be seen that the final
pH of these solutions, compared to the 3 hour test results in table 5.15, were different. However,
the greatest difference was between the TOC values, which were approximately five times greater
under the DIN test compared with the simple 3 hour test. This was explained by the different
water/sol id ratios used by these tests.
Table 5.16 pH and TOC results of the DIN leach test
Sample pHf pHf - pHi Corrected pH
TOC Corrected TOC
Blank 6.95 1.39 - 4.90 -
Norit 9.36 3.80 7.97 6.69 1.79
Type C 7.38 1.82 5.99 5.60 0.69
run 045 5.37 -0.19 3.98 7.77 2.87
run 006 6.31 0.75 4.92 8.43 3.53
G1 2.16 -3.40 0.77 103.60 98.70
STA05 6.31 0.75 4.92 216.78 211.87
The metal leaching data for the commercial carbons and the contaminated soils, presented in
figures 5.45 and 5.46 respectively, exhibit some variation compared to the 3 hour tests. With the
240
commercial carbons, there were lower concentrations of metals leached than under the 3 hour
test, and all the concentrations observed were not considered to be of concern.
Figure 5.45 Metals leached from Type C and Norit after the DIN leach test
10
I I § J 0.01
0.001 Ca
Metal leached
raType. C S N o r i t
Figure 5.46
1000
1 100
10
e 1 o
•-S
G 0.1
6 0.01
0.001
leach test
Cu H g Mti
Meta l s leached
S S T A D 5 • R u n 0 0 6 ffilGl ^ R u n 045
The two soils G1 and STA05 displayed slightly different leaching behaviour for the DIN test
compared to the 3 hour test. Under DIN testing STA05 showed greater amounts of Ca leaching
whilst G1 leached less Ca; in addition, the Zn leached from the carbons run 006 and run 045 also
increased. These changes in the amounts leached were not in proportion to the changed
water/solids ratios of the two leaching regimes (unlike the TOC), thus implying a different effect
was operating. The ability of the metals to enter the solution from the solid phase would be
influenced by the pH of the system and the effects of other dissolved ions in the solution. The pH
for the DIN tests were lower than for the 3 hour tests which means that the metals available for
241
leaching would have a higher solubility, as was seen with the increased leaching of Zn and Ca.
STA05 also exhibited slightly greater leaching of other metals when tested by the DIN leach test
compared to the 3 hour test, but overall the differences were not significant.
Thus, the limited leaching tests performed indicate that the organic components of the
contaminated soils were completely converted into non-leaching species, with the product
exhibiting leaching characteristics comparable to commercial activated carbons. The leaching of
metal species from the raw soils was very low, and apart from Ca and Zn the carbons made from
the soils also exhibited only minor leaching of other metal species. This indicates that the
activated carbons prepared from the contaminated soils would be suitable for most of the
applications which the commercial products are recommended for.
5.5 Summary
The ability of carbons manufactured from contaminated soil to remove metal species from
individual and mixed solutions has been demonstrated in this work. Application of activated
carbons to adsorption of metals from solution is an area of research which has received much less
attention than organic or gas adsorption studies. This is partially because the removal of metal
ions from solution is achievable, relatively cheaply, through several techniques such as ion
exchange, precipitation, electrolysis and adsorption by metal oxides (Corapcioglu and Huang,
1987). Adsorption of inorganic electrolytes by active carbon is affected by surface area, pore
structure, electrophoretic properties and surface acidity. Adsorption of heavy metals by carbon
is influenced by the metal-ligand charge distribution, the free metal and free ligand charges,
polarity of the ligand molecules and the pH^ of the carbon surface (Huang, 1978) and separation
of the individual contributions to the removal mechanisms is not readily performed. In soils,
metals can occur in various forms: (i) dissolved ions or complexes, (ii) electrostatically bound,
(iii) covalently bonded or coordinated, (iv) chelated or (v) precipitated. In the soil-derived
activated carbon, metal removal from solution could occur by (ii) - (iv).
pH is a very important factor when examining the uptake of the metal by adsorbents. Of the four
metals studied in this work, precipitation started for Cd(II), Cr(III) and Cu(II) at about pH 6, with
only Hg(II) being relatively insensitive to pH. Control of the pH with buffer solutions was not
very effective because the carbons possessed their own buffering capacity which tended to swamp
that of the main solution buffer. The two commercial carbons, Norit and Type C, were H-type
carbons although Type C carbon did exhibit some L-type characteristics. Norit tended to buffer
the solutions to approximately pH 9 and Type C to approximately pH 6.5. The soil carbons were
242
L-type, which tended to buffer the pH to approximately pH 5.5. The most effective way of
controlling the pH of the test solutions was found to be by adjusting the pH of the starting
solutions by addition of NaOH.
The adsorption processes were very rapid in their attainment of equilibrium. When precipitation
was not occurring, adsorption equilibrium between the carbon and metal species in solution was
attained within 15 minutes. The metal which was removed by carbon to the greatest extent was
Hg. The Hg(II) adsorption capacity of run 045 ( Q ^ = 28.57 mg/g; Kp = 18.24 mg/g) were the
greatest adsorption capacities reported for any of the four metals considered in depth by this
study. Higher values for Cd(II) (37.04 mg/g) and Cr(III) (45.46 mg/g) uptake by Norit SA4,
Kp values for Cr(III) (23.77 mg/g) and Cu(II) (39.45 mg/g) uptake by Norit SA4 and Kp value
for Cu(II) (58.61 mg/g) uptake by run 045 could not be unambiguously assigned to adsorption
processes, indicating that some or all of this apparent "adsorption capacity" was due to
precipitation at elevated pH values. Secondly, this adsorption capacity of run 045 for Hg(II) was
very conservative, due to the complete removal of Hg(Il) from solutions of concentrations up to
300 ppm.
The soil carbons were favoured for cation adsorption because of their negative surface charge,
but the mechanisms for adsorption of each metal species were considered to be slightly different.
For Cd(II) the carbon type (L- or H-) influenced uptake, whilst the ability of the metal to form
surface complexes with the carbon also affected the observed removals. The sulphur present in
the soil carbons was particularly significant in this respect. Chromium (III) adsorption was most
influenced by the surface charge of the carbon. Copper (II) removal was observed even from
acidic solutions, a feature which was explained by the interaction between the hydration spheres
of the carbon surface functional groups and the Cu(II) ions. Mercury (II) adsorption was probably
the most complex phenomena because even the mode of activation could influence the ability of
the carbon to adsorb Hg([I). Hence there were three possible removal mechanisms which may be
occurring: (i) the formation of surface species by the Hg(II) with carbon functional groups
(especially sulphur); (ii) the reduction of carbon surface oxides by adsorbed Hg(II), forming Hg°;
(iii) the amalgamation of Hg° with Zn° metal particles present in the carbon lattice from the
activation procedure. With mixed metal solutions the study indicated that there was some
competition between Cd(II) and Hg(II) for their adsorption sites, because the value for Cd(II)
was a third lower for the mixed solution (6.21 mg/g, table 5.9) compared to the single solutions
(9.90 mg/g, table 5.5).
The leach testing of the carbons and soils used in this research indicated that carbonisation and
243
activation of contaminated soil virtually eliminates leaching of organic species from the product
compared to the feedstock. Some metal leaching was observed, the presence of gypsum in the
feed soils contributed to Ca leaching whilst inefficient washing with HCl after ZnCl; activation
resulted in the presence of leachable Zn which was detected in the leach test liquors. It was noted
that the DIN test resulted in more acidic pHs than the 3 hour test, and this consequently resulted
in increased leaching of metal species. However, the overall leaching performance of the
commercial carbons compared to those made from contaminated soil was almost inseparable
indicating that the soil derived adsorbents would not be limited in their applicability compared to
the commercial products.
Both the commercial and soil-derived activated carbons demonstrated adsorption capability for
components of the aqueous waste systems examined. Better evaluation of their metal adsorption
properties was hampered by the lack of metals in the two aqueous waste systems available for
assessment. However, the results obtained indicated that the soil-derived carbons were capable
of reducing the TOC and the concentrations of some of the metals in the landfill leachate. To use
these materials as in-situ adsorbents in landfill system, it would probably be necessary to mix the
carbon with other adsorbents to produce a material which will exhibit the desired removals for
all the components of the leachate, thus achieving the required leachate quality without further
treatment. In contrast to this, the electroplating effluent was satisfactorily treated by the soil
derived carbon. It exhibited superior adsorption capability than the commercial carbons and the
resulting effluent contained concentrations of metals which were below the detection limits of the
ICP-AES. This result means that the soil-derived carbons are ideal for treating this type of
effluent.
244
CHAPTER SIX
CONCLUDING DISCUSSION
6.1 Introduction
The development of laws and procedures which promote the responsible management of waste
has been a slow process. The waste management system which operates in the UK has evolved
over many years as a result of public and political pressure in response to incidents caused by the
irresponsible disposal of waste. More recently, the major impetus for change and revision of this
system has come from Europe, with the various directives issued by the EC Commissioners,
which were discussed in Chapter 1. The long-held belief that the environment could cope with
the ever-increasing volumes of waste which were being placed into it were seen to be false when
it became evident that apparently resilient ecosystems such as the Great Lakes and the North Sea
were being poisoned by the accumulation of toxic man-made substances. These are just two
examples of the effects of reckless disposal practises which are still occurring, despite the best
attempts to limit or reverse the resulting environmental damage.
The majority of the concerns about our environment relate to the quality of the air and water
systems, and in that respect the industrial countries of the western world have been attempting
to halt, and in some cases, reverse the deterioration of the quality of these media, as discussed
in section 1.2.3. However, The removal or reduction in one pollution source is very often
counteracted by the growth in a new one. One example of this was the improvement in UK urban
air quality following the Clean Air Acts of 1956 and 1968, which has been progressively reversed
as private car ownership and usage increased during the 1980s and 1990s. Means of stabilising
and reversing this trend are still being sought.
The medium which has received the least attention in terms of legal protection has been land. This
would appear to be strange since it is the medium which has the greatest intrinsic value and is
considered to be a commodity. Indeed, the principle of contaminated land only really came into
existence when the Love Canal incident occurred in 1974. The UK has never suffered from an
incident of this severity, although there are many documented examples of contaminated material
appearing in gardens of houses built upon contaminated sites, of houses being subject to damage
due to contamination in the soil attacking the building, or of explosive gases from the ground
igniting in the house voids and destroying the house by detonation (ie. GLC, 1984, Anon.,
1991d). However, the regulation of contaminated land is not an area which has been terribly
successful in most of the countries where it has been attempted. The countries which took the lead
245
with regard to this aspect of contaminated land were Holland and the USA, as discussed in section
1.3.5.3, and both of these countries have needed to revise and change the original contaminated
land/soil protection laws due to both political and legal pressures. Perhaps the country with the
most notorious system is the USA. CERCLA (Superfund) was designed to ensure that
contaminated sites which posed a health risk were cleaned up quickly and completely, and the cost
of these actions would be borne by the polluter. Unfortunately, as the USA and USEPA have
discovered, the most noticeable benefit of the Superfund Act was that the lawyers fighting the
clean-up orders issued by the USEPA became very wealthy (Church and Nakamura, 1993).
In Holland, although the government does have a moral argument for cleaning up land with
regard to groundwater issues, the idea of multifunctionality and sites being fit for any use has
received much criticism and resulted in many court cases from unhappy land-owners disagreeing
with the actions and attitude of the Dutch authorities. The author is aware of one such site in the
Netherlands, where the ex-leaseholders were not consulted on the clean-up of the site after
vacating the area. Historical research had shown that the surrounding area had been liberally
polluted with gaswork wastes and harbour sediments which had been used by the site owners to
level the ground prior to leasing it. Despite this fact, the ex-tenants were to be wholly responsible
for all the clean-up charges. What made the clean-up even more unreasonable was that none of
the surrounding sites were to be cleaned, so the resulting multifunctional site would be akin to
an island in the middle of an ocean. The consequence of these acts has been similar to the USA
situation where more cases are being decided in court. To try and minimise these occurrences the
Dutch government revised the law and included a clause that made companies responsible for any
contamination found on the sites they occupied, unless they could prove that this is not the case.
This was a total reversal of the old law and makes the site occupier guilty until proven innocent!
(Simons Vinckx, 1994).
The UK, mindful of these aggressive laws and the adversarial situations which they create, has
resisted implementing laws which result in the clean-up of contaminated sites being decided in
court. The traditionally pragmatic attitude of the UK towards environmental issues is well
illustrated by the case of contaminated land. The abortive attempt to create the contaminated land
register resulted in the UK government consulting widely for the revised policy (DoE, 1994a)
with the consequent publication of the framework for contaminated land (DoE, 1994b). The
inclusion of the results of the public consultation on contaminated land issues into the 1995
Environment Bill meant that contaminated land was finally recognised as an environmental issue
by the UK. However, the USA and the Netherlands require, by law, sites to be cleaned up to a
set level or maintained and monitored after remediation. In contrast, the UK does not - market
246
forces are the preferred solution. In principle this is not unreasonable, particularly as one of the
most forceful arguments against the Dutch and USA regimes has been the un-sustainable nature
of the policies and laws. Supporters of this situation argue that the use of market forces means
also that companies which buy contaminated sites will restore them to a level which will minimise
their future liability, hence almost guaranteeing that the sites are fully decontaminated. However,
the flaw in this argument is that in the UK the cheapness of landfill space has prevented
contaminated soil technologies from developing to any great extent, which is almost the opposite
of that which has happened in the USA and Netherlands, where soil and contamination treatment
is a requirement for the site clean-up. Consequently, the usage of soil treatment techniques is
routine and new methods for addressing contaminants are continuously being developed,
particularly by the SITE programme in the USA, which was discussed in section 1.3.8.
However, as also stated in Chapter 1, most contaminated land treatments rely on destructive
methods for removing the contamination from the soil, and are costly to operate. For some of the
more intractable pollutants this is the only way in which they can be removed from the soil.
Chemicals in this group include PCBs, dioxins, cyanides, PAHs and heavy metals such as Pb,
Hg and As. In soil from gaswork sites, metals, cyanides and PAHs are ubiquitous contaminants.
Destructive treatment, unfortunately, results in a residue which does not resemble the soil in
composition or physical characteristics and has minimal worth. It is totally inert (sterile) and may
retain some of the pollutants (especially the heavy metals) in a concentrated form subsequently
requiring further treatment or disposal by a secure manner. The need for an effective treatment
of these particularly difficult materials, which results in a useable product is obvious. Some of
the novel processes which were detailed in section 1.3.8 have not received the usage which they
could have expected due to several factors which include: residues which require further treatment
or safe disposal, the cost of operation and, perhaps the most important factor; a lack of
acceptance by the regulators and the public. For example, the case for incineration, although
scientifically proven, has been poorly received by regulators, planners and most importantly the
public due to the historical occurrence of pollution by the poor operation of these plants.
Consequently, the use of accepted technologies for soil treatment, which not only destroy or
nullify the contamination whilst producing a useful, saleable product from the contaminated soil
is vital. To this end, the use of activated carbon manufacturing technology has been investigated
and shown to be eminently suitable for treating contaminated soil by converting it into a
carbonaceous adsorbent, with virtually no potential to pollute, and an inherent value.
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6.2 Assessment and treatment of the contaminated soils
6.2.1 Chemical analysis and activation agent selection for soil treatment
Contaminated soils are, by their nature, very heterogeneous and no soils are more so than those
from gaswork operations. The soils in this study were very representative of this type of
contaminated site, with much contamination and a great dis-similarity between samples taken from
the same site. The colour of the samples, shown in plates 3.1-3.6, was characteristic of the types
of contamination present, ie; STA02 was black, and very rich in carbonaceous materials, whilst
STA03 was dark blue, a confirmation of the high cyanide content of the soil. Finally, STA05 was
a pale brown colour and very odorous, implying VOC contamination, which was confirmed by
the thermal analysis data where weight-losses were detected at the lowest temperatures (refer to
figure 3.11 and table 4.2). The extent of contamination of the soils, given in tables 3.4 and 3.5,
varied from soil to soil, with some samples being grossly contaminated {ie, cyanide in STA03:
sulphur and Hg in STA04). The contaminant distribution between the different soil fractions
discussed in section 3.3.2 and shown in figure 3.7 indicated that the use of the BSI (1990)
guidance for analysis of soil samples, viz: using the <2mm fraction, did not provide a wholly
representative picture of the soil contamination, since analysis of the individual fractions up to
16mm showed that all size fractions contained similar levels of cyanide, sulphur and sulphate.
Although some of the analyses performed on the soils were of gross parameters such as total
cyanide and CHN, the data produced was of comparable use. These parameters, for example total
cyanide, are not of much benefit when a risk assessment is being performed upon the soils, but
were of great use when evaluating the effectiveness of the treatment procedure upon the soil
contaminants.
The correlation of this data with other analyses performed upon the soils, particularly the XRD
analyses which were discussed in section 4.5.2, enabled the identification of major crystalline
contaminant phases in the soils, which indicated that hydrated and anhydrous gypsum, quartz,
FejOj, Fe4(Fe(CN)6)3 and sulphur were present in these soils in significant amounts.
The lack of UK Governmental assessment values for contaminants meant that the interpretation
of the contamination found in the soil used in this work was performed using some reference data
which was not of UK origin nor necessarily valid for this task (tables AI.1-AI.3), which is a
common dilemma for commercial contaminated land remediation schemes. Most of the emphasis
on assessing contamination is performed with respect to the human health effects which may result
from interaction with the soil. In that respect, the Canadian and Dutch Assessment values
248
considered these interchanges, but they were derived using factors which were specific to the
country requiring these data. These could include the local geology, the source of the water and
population behaviour and density. Thus, the use of these data outside their country of origin could
be considered to be unreasonable. With the Dutch reference data, for example, this argument
would rest on the contention that the values are too cautious, but in reality, the extra safety factor
should be considered to be beneficial, although the severe cost implications to the site owners or
re-developers will nearly always result in the dilution of these standards.
The determination of the metals in the soils was performed using USEPA method 3050 (USEPA,
1986, 1991b) and the results discussed in section 3.3.3. Although this method was shown to be
suitable for these determinations, the analysis of some of the metallic contaminants was
complicated by interference effects from co-contaminants. In particular the presence of free
sulphur in the soils was linked to the decrease in the recoveries of analytical quality control
spikes. It was considered that the sulphur inhibited the solubility of the metals by forming metal
sulphides which co-precipitated, the effect of which was particularly evident with Hg, Zn and Pb.
In addition, the sulphur species also had a profound influence upon the activation procedure,
which is discussed later.
The proposed treatment scheme for producing activated carbon from gaswork contaminated soil
required the thermal characteristics of the soil to be established. Initially only one soil was
studied, STAGS, and the results of this thermal analysis were given in section 3.3.4. The initial
thermal analysis investigation evaluated the behaviour of the soil with and without the influence
of selected physical and chemical activating agents using just one of the soil samples. The results
of these investigations, shown in figures 3.11 and 3.13-3.17 and summarised by table 3.6 showed
that carbonisation occurred in the temperature range 150-175 °C. The activated samples also
exhibited reactions in the temperature range 450-675 °C. Only three activation agents were found
to influence the thermal reactions of the soil; CO;, H2SO4 and ZnCl;. Bulk samples for further
investigation were prepared utilising this data and fully analyzed for adsorption, porosity and
surface area characteristics.
Of the three potential activation routes used, the activant which produced the most effective
adsorbent was ZnCl;. This was clearly shown by the % adsorption of the carbons from aqueous
phenol and 4-nitrophenol solutions in table 3.8. Zinc chloride produced carbons with twice the
phenol adsorption capability of the alternative activation methods. The superiority of ZnCI; as an
activation agent has been confirmed by many other researchers (Chan et al., 1980: Tanin and
Giirgey, 1987 and Pollard etal., 1991a). The aqueous isotherm data, plotted in figures 3.18-3.24,
249
and summarised in tables 3.10 and 3.11 showed that both the Langmuir and Freundlich models
were, in general, applicable to the data. These plots also confirmed that ZnCL produced the most
adsorbent carbon from the contaminated soil, with Langmuir values of 0.12 mMol/g and
0.23 mMol/g for phenol and 4-nitrophenol respectively. These values were at least 25%
greater that those for the alternative activation agents.
The gaseous adsorption data, presented in table 3.12 and figure 3.25 confirmed the superior
surface area and microporosity which was developed by the sample activated by ZnCI;. The total
pore volume of the ZnClj sample was 69% and 83% greater than the CO2 and H2SO4 activated
samples respectively. ZnCl; is recognised as an activation agent which tends to develop
mesoporosity within samples, although the resulting carbons from this study also possessed a well
developed microporosity.
6.2.2 Production of the carbon with optimum adsorption properties and minimum
contaminant potential
Only soil STA05 had been used for the initial activation agent selection, but to demonstrate the
applicability of the process, a wider a range of samples required activation using ZnCl^ and full
characterisation for their adsorption ability, and the reduction or elimination of contamination to
the samples.
The thermal characteristics of the other soils (STA01-STA04 and Gl), with and without ZnCl;
addition were illustrated by figures 4.1 - 4.11 and summarised in table 4.2. The soils, although
chemically different with respect to their contamination and composition, exhibited carbonisation
reactions within the temperature range 150-275 °C and activation reactions within the range 450-
525 °C. These relatively narrow temperature ranges were reassuring, since it implied that use of
this technique for treating contaminated soils from disparate sources would not require very
different operating conditions for each fresh batch of soil. Investigation of the literature found that
much work had been performed on the study of the mechanisms of coal pyrolysis in the presence
of ZnClj (section 4.3.1.1). This work was considered to be very relevant to this research, because
the contaminated soils were polluted by residues produced from coal pyrolysis. The research by
Ibarra et al, 1991; Jolly et al, 1988; Kandiyoti et al., 1984; Neuberg et al., 1987 and O'Brien
et al., 1987 were particularly significant as they found that ZnClj dissociated upon addition to the
coal, with the Zn associating with any sulphur in the samples and the CI being dispersed
throughout the coal matrix. The ZnClj was thus considered to assist in the catalytic cracking of
the coal, eliminating Hj and other light aliphatics, whilst minimising tar formation. Since the soils
250
used were polluted with coal-derived residues, it could be expected that the ZnClj would have a
similar effect upon the soil. Confirmation of this supposition was provided by the XRF and XRD
analysis of the resulting soil-carbons (sections 4.5.1 and 4.5.2 respectively), which showed that
Zn was retained in all the soil-carbons and that ZnS was formed in some of the samples .
The sulphur species present in the soils had been noted to affect the accuracy of the metal
digestions performed upon the soils, but they were also noted to have a curious effect upon the
ability of the samples to develop adsorption properties. This was particularly noticeable with
sample STA04. When thermally treating the soils without ZnCl; addition, it was noted that the
sample still developed extensive adsorption characteristics. This self activation ability of the soil
was traced to the levels of sulphur species present initially in each soil, and figure 4.19 clearly
illustrated how the relative activity of the non-ZnCI^ samples was linked to the amount of sulphur
species in each soil. The use of FeSO^ as an activation agent further tested this hypothesis and
confirmed that sulphur species did have an activating effect upon these soils.
The effect of varying the amount of ZnCU added to each soil, upon the product yield and carbon
content were shown in table 4.3. Each soil sample behaved slightly differently when exposed to
similar amounts of ZnClz, with no one soil : ZnCl; ratio being optimum for each soil. Similarly,
the effects of the ZnCl; upon the contaminants present in the soils (cyanide, sulphate and sulphur),
shown in table 4.4, was not consistent. Of particularly interest was the observed increases in the
sulphur levels in some of the samples. This result was attributed to the reduction of sulphate into
sulphur during the processing, which was indicative of the reducing conditions which exist during
carbonisation/activation reactions, which were discussed in section 1.5.3.1.
The resulting carbons exhibited adsorption properties which were variable. All of the soil-carbons
were L-type, according to the classification of Steenberg (1944), as demonstrated by figures 4.12
and 4.13. The single point adsorption studies on each sample, given in table 4.5, showed that
these carbons when activated using a ZnCl; dosage > 33.33 % ^/w could remove between 17-
73 % of phenol and 29-96 % of 4-nitrophenol from 10 mM solutions (dose = 1 g
carbon / 100 ml adsorbate solution). This compares to 76-95 % and 94-99 % removal,
respectively, by the commercial carbons also studied. However, because of the variable
compositions of the soil carbons the most suitable means of comparing the adsorption of the
commercial carbons with that of the soil-carbons was by use of the relative activity parameter
(section 3.3.5.2). These values indicate that in many of the soil carbons, the carbon present in
the samples was more adsorbent than that of the commercial carbons. The main reason for the
lower adsorption was due to the smaller amount of carbon present in the soils initially. To assist
251
in making the decision for the optimal ZnCI^ addition for adsorption development, figures 4.14-
4.18 were constructed. The amount of ZnCU which was chosen was 33.33 % Although not
all of the samples studied exhibited maximum adsorption or relative activity values at this level
of ZnCl; addition, the extra adsorption capacity developed by using the supplementary ZnCl;
would not be warranted in large-scale treatment processes.
Adsorption isotherms were constructed for phenol and 4-nitrophenol adsorption by the soil-
carbons. These results were amenable to interpretation by both the Freundlich and Langmuir
models. The fit of the data (r^ to the equations was greater than 0.9 for almost all of the
adsorbents. The soil carbons also exhibited phenol values which were less than those from
the commercial carbons, and except for sample STA04 (run 001 and run 007) the 4-nitrophenol
data showed similar trends, as shown in tables 4.7 and 4.8 and Appendices IV and V. The phenol
Qmax values for the soil derived carbons were in the range 0.13-0.82 mMol/l/g for phenol, which
compared with 0.88 and 0.97 mMol/l/g for the commercial carbons. For 4-nitrophenol adsorption
by the soil carbons, values between 0.17 - 1.06 mMol/l/g were recorded which compared
to 0.93 and 1.32 mMol/l/g for the commercial carbons. The difference in the values between
the commercial and soil-carbons was not unexpected, nor was the dissimilarity between the phenol
and 4-nitrophenol values, as discussed in sections 4.3.4.3 and 4.3.4.4.
The gaseous adsorption study of the soil carbon showed that the samples were all mesoporous
solids, as defined by the type IV isotherm shape, and they all possessed mixed H3/H4 hysteresis
loops, which were discussed in section 1.4.2. However, there was a pronounced formation of
microporosity in many of the samples, a feature of ZnClj activation which has been noted by
other researchers (eg: Kadlec et ai, 1970; Rodrfgues-Reinoso and Molina-Sabio, 1982). The
influence of the ZnCl; upon the development of the surface area and porosity by the samples was
well demonstrated by the gas adsorption data, which was summarised in table 4.10. The curious
properties of the carbons made from soil STA04 were again highlighted, with run 015 exhibiting
a surface area which was greater even than Norit SA4 commercial carbon. Increasing the dosage
of ZnClz caused the mesopore pore-size distribution plots (figures 4.30-4.34) to show a shift
towards a larger mesoporosity, whilst the microporosity (Horvdth and Kawazoe plots; figures
4.36-4.40 and table 4.11) showed a con-current decrease. The adsorption capability of the soil
carbons was not as effective as the commercial products. This observation was consistent with
the earlier results, which indicated that factors such as feedstock composition and chemical
composition of the adsorbent can influence the adsorption ability of the material.
To simplify the initial stages of this research, constant residence times had been used throughout
252
the sample preparations, but to fully optimise the sample preparation, carbonisation and activation
residence times needed to be established (section 4.4). Carbonisation times of up to 3 hours and
activation times of up to 2 hours were used. By comparing the changes in the CHN ratio,
contaminant concentrations, phenol and 4-nitrophenol adsorption and gaseous adsorption data, the
sample which was considered to exhibit the most beneficial values was run 041 which received
a 33.33 % ^/w ZnClz loading and was carbonised and activated for 1 hour and 2 hours
respectively. The influence of the different carbonisation and activation times upon chemical
characteristics of the G1 soil samples was minimal, as was shown by the changes in the CHN
ratios and the contaminant levels for each of the samples prepared which are given in tables 4.13
and 4.14 respectively. However, the adsorption properties of the samples did exhibit sufficient
differences to enable the identification of the optimum carbonisation and activation time. In tables
4.15 and 4.16 the aqueous and gaseous results, respectively, show that run 041 possessed the
superior aqueous and gaseous adsorption ability.
To confirm that these parameters would not be subject to effects due to sample size, a large
sample (300 g) was prepared (compared to 50 g for the test samples). This sample, run 045, was
compared with run 041 for CHN and residual contamination levels in table 4.17 and aqueous
adsorption ability (Langmuir and Freundlich constants) in tables 4.19 and 4.20. It was found to
not differ significantly from the test sample with respect to its adsorption characteristics and
contamination reduction, indicating that the increased size of the processed sample did not unduly
affect the carbonisation and activation reactions.
The final stage of this part of the research work investigated the metallic contamination which is
unavoidably associated with gaswork soils. The digestions of the raw soils showed that most
heavy metals of concern were present in the soils, including elevated levels of Pb, Hg, Mn, Zn
and Cu in several of the samples (table 3.5). Since the soil-carbons were subjected to a dilute acid
wash to enable recovery of ZnCl^, both the wash liquors and solid product were analyzed for
metals. The data shown in table 4.18 demonstrated that, apart from Zn, metal removal by the HCl
wash was minimal. The leaching of Zn can be minimised by the extent and number of dilute acid
washes used to recover the ZnCl, after carbon preparation. The HCl wash is also beneficial as
it can ensure that any metal species which are not encapsulated by the carbonisation reactions are
removed before the materials are applied to adsorption applications. The XRF analysis of the soil
carbons (figures 4.47-4.52 and table 4.22) confirmed that much of the metallic contamination was
retained by the carbon product of the processing. These metals were considered to be actually
trapped in the carbon lattice in a non-leaching form. This was very important because if the
metals were to leach, the usefulness of the soil-carbons would be severely restricted. DIN leach
253
tests of the carbons confirmed that metal leaching was not of concern (section 5.4.7), nor was
leaching of organic species. Typically, between 2 and 3 ppm of TOC leached from the soil
carbons, which compares to 98-211 ppm for the raw soil and 0.7-1.8 ppm for commercial
carbons.
Considering that the contaminated soils used in this study contained complex, potentially
hazardous wastes, which tend to resist traditional treatment methodologies, or require complex
multi-stage treatment processes to facilitate the complete neutralisation of all the various
components, the use of the well established active carbon manufacturing technique of carbonising
and activating the material achieved a final product which was benign, possessed minimal
potential to pollute and had reuse potential.
6.2.3 Demonstrating potential applications for the soil-carbons
Although activated carbon has traditionally been used for removing organic pollutants from
various media, the use of carbon for metal adsorption is receiving much attention. Metal
containing effluents have been subject to increasing controls by the regulatory bodies, especially
after incidents such as Minimata Bay (section 1.1). The EC has recognised the need to limit the
release of soluble metal species into the aqueous environment, with many metals receiving
classifications according to their perceived toxicity (refer to table 5.1). For these investigations,
six metals were initially chosen for study; Cd, Cr, Cu, Hg, Ni, and Pb. Subsequent trial
investigations limited the study to 4 metals; Cd, Cr, Cu and Hg. The sources of metal-containing
effluents are many and varied, although from the literature considered and reported in section
1.6.1, effluents from metal plating facilities and landfill sites are considered to be particularly
problematic. These were chosen as suitable waste liquids to use to examine the metal adsorption
ability of the soil carbons.
The initial stages of this part of the work involved using standard solutions of the chosen metals
to examine the adsorption phenomena between individual metals and the soil-derived and
commercial carbons. Firstly, consideration of the kinetic aspects of the metal adsorption (Section
5.4.3) demonstrated that equilibrium was attained within 30 minutes for all the metals studied
(figure 5.1-5.4). However, a time of 3 hours was subsequently used for all adsorption
experiments, which was consistent with the contact time used during the organic adsorption
experiments and ensured that equilibrium had been established.
It had been recognised that there was a need to control the pH of the adsorbate solutions. Initial
254
studies used an NaOH/phosphate buffer system, which proved to be adequate for controlling the
solution pH when the metal concentrations were low (<10ppm). However, the use of the
buffered solutions for the construction of adsorption isotherms was found to be impracticable as
the starting metal concentration in the solutions increased above 10 ppm. This was because the
acid component in the metal solutions swamped the buffering capacity of the system, which
resulted in the solutions becoming very acidic with increasing metal concentrations, as shown in
table 5.4. This pH change affected the solubility of the metals, and hence prejudiced the
adsorption phenomena. The effect of the change in pH upon the solubility of the metals under
study was shown by figures 5.5 and 5.6. Only Hg was comparatively unaffected by pH induced
precipitation, whilst Cu, Cd and Cr all precipitated fairly readily. For precipitation effects to be
minimised, it was concluded that the pH needed to be at pH 5.5 or below, which was performed
using NaOH. This figure was just within the lower value which was reported in table 1.13 for
the chemical characteristics of landfill leachates.
Adsorption isotherms were constructed for solutions containing individual metals and a mixture
of the metals. For completeness, solutions which had not been subject to pH adjustment, as well
as those solution which had, were evaluated. The results from these un-adjusted solutions were
mostly as expected, with the solution pH values dropping rapidly as the metal concentration
increased (for example; figures 5.7, 5.12, 5.17 and 5.24). The effect of this pH decrease was to
minimise the adsorptive forces which may have been operating between the metal species and the
carbon surfaces making interpretation of the results by the Freundlich or Langmuir isotherm
models impossible. The only exceptions to this trend were the Cu(II) adsorption by run 045 and
Hg(II) adsorption by all the carbons, which were still adsorbed in the pronounced acidic
conditions, producing data which was amenable to interpretation by the two isotherm models, as
shown by figures 5.18 - 5.20 for Cu(II) and 5.25 - 5.27 for Hg(II).
The effect of the carbons upon the solution pH values were similar for all the metals. Norit
carbon caused the pH to increase to approximately pH 9, indicative of an H-type carbon. Type
C carbon caused a slight increase in pH, but the pH did not tend to rise above pH 7. Finally, the
soil carbon tended to cause the pH to fall slightly.
Considering the Hg(II) adsorption results, the values for the Hg(II) solutions without pH
adjustment were greater for Norit SA4 and the soil carbon than for the Hg(II) pH adjusted
solutions, whilst for Type C, the values were virtually the same. This high adsorption
capacity for Hg(II) was explained in terms of the organic functional groups associated with the
carbonaceous portion of the adsorbents. The sulphur species which are an unavoidable component
255
of the contaminated soil feedstock were of particular importance in this situation. Mercury (II)
has a particularly high affinity for sulphur species and readily forms complexes with the S
species, even in the presence of oxygen ligands, as discussed in section 5.4.4.5. Another potential
influence upon the adsorption mechanism for Hg(II) by carbon was proposed by Huang (1978)
who suggested that the mode of activation is more influential in the adsorption of Hg(II) than the
feedstock. Zinc chloride was found to be particularly effective in preparing carbons which
exhibited an affinity for Hg(II), a fact supported by this work. Finally, Huang and Blankenship
(1984) proposed that an adsorption/reduction mechanism may operate where Hg(II) is adsorbed
by carbons. They suggested that Hg(II) is reduced to Hg" within the pores of the carbon. If this
is the case, then the released Hg° may be available to amalgamate with any metal residues within
the carbon lattice, particularly Zn, which was observed in the XRF data discussed in Section
4.5.1.
For Cu(II) adsorption, which was discussed in section 5.4.4.4, the pH of the unadjusted solutions
was acidic, but Cu(II) uptake by the soil carbon was still observed. This suggested that the
mechanism of adsorption was not just reliant upon simple electrostatic forces, but that surface
complexation effects were also operating. The pH adjusted solutions were generally amenable to
interpretation by the isotherm models. In table 5.7, the isotherm constants were summarised, and
the values shown. Precipitation played an important r6Ie in the observed adsorption results,
and it was considered that because the final pH of the solutions was generally greater than pH
5.5, the optimum pH for adsorption, the majority of the observed metal removal was actually due
to precipitation on the carbon surfaces. When adsorption did occur, as with run 045 for the
unadjusted solutions, the carbon surface played an important part, as was reported by Corapcioglu
and Huang (1987), who suggested that the hydration shells of the carbon surface oxides and the
Cu(II) ions were interacting by hydrogen bond formation and causing the observed adsorption.
The adsorption of the remaining metals (Cd and Cr) was very dependent upon the pH of the
solutions, and the acid/base character of the carbons played an important r61e in determining the
final pH of the test solutions, as illustrated by figures 5.8 and 5.13. For Cd, discussed in section
5.4.4.2, adsorption was only observed where pH adjustment of the Cd solutions was performed
prior to carbon contact. The final pH of the carbons was shown in figure 5.8, with the L type
characteristics of the soil carbon being clearly shown. The Cd(II) adsorption isotherms for these
carbons were given in figures 5.09 - 5.11, and the values in table 5.5. It should be noted
that the large value exhibited by Norit was due to precipitation, and not adsorption. Where
adsorption was observed, the formation of surface complexes and the carbon type (H or L) were
major factors in influencing the Cd(II) uptake.
256
Chromium adsorption was discussed in section 5.4.4.3, using pH adjusted solutions. This metal
was the least adsorbed by run 045, as shown by the value in table 5.6, and was also the
metal which was subject to the most precipitation effects. The complex solution chemistry of Cr
was considered in depth in section 5.4.4.3, with the different possible oxidation states of Cr
existing as different charged species, with the charge of the Cr(III) ion changing with the solution
pH. It was noted that Cr(III), which was considered to be the dominant species in the test
solutions, was most readily adsorbed by L type carbons and rapidly precipitated at pH values
> 5.7. However, Cr(VI) is not readily precipitated, and if present was only adsorbed at low pH
(< 3) by H type carbons, but no evidence of Cr(VI) was found in these studies. Chromium(VI)
has also been reported to oxidise the carbon surface forming Cr(III) (Huang, 1978). Both Type
C and Norit SA4 removed much more Cr than run 045 from the test solutions (table 5.6), but
inspection of figure 5.13 shows that the solutions in contact with these carbons are at pH values
> 6, which means that it was very unlikely that adsorption was occurring, and precipitation was
dominating as the removal mechanism. Hence, Cr adsorption is influenced by the charge on the
Cr species in solution (the solution pH) and the carbon type, and that run 045 is very suitable for
Cr(III) adsorption, assuming that precipitation does not occur in the solutions under study.
Adsorption by the mixed metal solution showed that Cr(III) and Cu(II) precipitated from solution,
and Cd(II) and Hg(II) were subject to co-precipitation effects. Only Cd(II) and Hg(II) remained
in solution at sufficiently high concentrations to be studied for adsorption. It was noted that Cd(II)
adsorption decreased slightly with the commercial carbons, whilst the Hg(II) adsorption was
essentially unchanged. In summary, the adsorption of Cd(II) and Hg(II) by the soil carbon was
on different sites on the carbon surface, whilst for the commercial carbons, some adsorption site
competition was observed.
The samples of real landfill leachate and electroplating effluent obtained for adsorption testing of
the carbons were analyzed for their metal content, pH and in the case of the landfill leachate, the
TOC. The notable result from these analyses was that the levels of metal ions in each of these
waste liquors were of much less significance than the literature reports, as shown for the leachate
by comparing table 1.13 with table 5.12. Landfill leachate possesses a lower potential for
mobilisation of trace metals than distilled water due to the high dissolved ion concentration,
resulting in much lower metal levels in the leachate than expected. Leachate is a very complex
system which contains many species in an essentially aqueous medium. The ability of the leachate
to dissolve and carry species is dependent upon the concentration of each fraction. The leachate
was rich in organic species and the pH was relatively high, factors which will not favour the
mobilisation of metal species. The consequence of these low metal concentrations was that
257
accurate determination of the metal concentrations before and after adsorption was complicated
by the nearness of the concentrations to the detection limits of the ICP, hence affecting the
accuracy of the measurements, the effect of which was discussed in section 3.3.3. The pH and
TOC of the leachate were within the literature ranges reported in table 1.13. Treating the leachate
with carbon resulted in a noticeable drop in the TOC of the system, with a 13 % removal, which
was of a similar magnitude to that observed for phenol adsorption. The commercial carbons
exhibited greater metal adsorption than the soil carbon for all the metals except Fe and V, and
Zn leaching was observed to be a problem with the soil carbon, although this was attributed to
insufficient washing of the carbons with HCl after preparation.
The treatment of the electroplating effluent with the carbons resulted in different results to those
observed for the treatment of the leachate. The soil carbon adsorbed 50% of the residual Cr and
100% of the residual Zn from the effluent. It was also noticeable that the leaching of Zn from
the soil carbons was not a problem for the treatment of this effluent. This was a superior
performance compared to the commercial carbons, and demonstrated the potential applicability
of the soil carbons.
As a final stage to this research, several leach tests were performed to evaluate the benefit of the
conversion of contaminated soil into activated carbon with regard to the reduction of the leaching
potential of the pollutants originally associated with the soil. In tables 5.15 and 5.16, the virtual
elimination of the leaching of TOC from the soils after thermal conversion was clearly
demonstrated. For leaching by metals, both the commercial and soil carbons tested exhibited some
leaching, but for most of the metal species detected in the soil systems, there was a reduction in
the extent of leaching after carbonisation. However, the leaching of Zn from run 045 was
noticeable whereas it was not an issue from run 006, as shown by figures 5.44 and 5.46.
6.3 Summary
This research programme has conclusively demonstrated that the treatment of gasworks-derived
contaminated soil is viable by using it as a feedstock for manufacturing active carbon. This
treatment is unique in that it effectively decontaminates gaswork soil in a single process, whereas
apart from incineration, all other treatment schemes for soil of this type rely on separate processes
for each of the contaminating phases (metals, organics and anions treated separately). The
manufacturing technology is accepted and already complies with environmental emission
constraints of the regulatory bodies. There are several sites around the country dealing with its
production, most of which have long manufacturing histories. Finally, the product is usable, with
258
adsorption and leaching characteristics which are similar to commercial products.
259
CHAPTER SEVEN
CONCLUSIONS AND SUGGESTIONS FOR FURTHER WORK
7.1 Conclusions
7.1.1 Contaminated land is a major issue for both politicians and industry, which will not be
resolved in a simple or cheap manner. It is present in all industrialised countries and is
still being created. The characteristics of soil contamination are often such that it can
easily and quickly spread to both air and water.
7.1.2 The major impetus behind the development of the contaminated land issue in the UK has
been the need to reuse land due to restrictions upon the availability of new development
land. The need for more land has tended to coincide with the decline in the older
"traditional" industries which have left a legacy of industrial contamination, which has
meant that land for development is available, but requires attention before reuse.
7.1.3 In general, the public perception of contaminated land, which has been shaped and
developed by the national press, TV and environmental pressure groups is an ill-informed
one. The consequence of this has been much confusion with regard to contaminated land
and the actual risk to health which it presents. There is also a lack of informed guidance
with regard to the hazards of contaminated land, which still have not been addressed by
the UK government. The general ignorance about contaminated land has caused much
unease within industry with many companies owning land which is potentially of a greater
fiscal liability to the business concerned than the company's "book" value.
7.1.4 The treatment of contaminated soil is not generally considered as an economically viable
option within the UK due to the low cost of landfill space, despite the fact that
contaminated soils disposed of in this manner continue to be a source of pollution. There
is also the problem that most treatment technologies developed for contaminated soil
decontamination are new processes. These technologies are not only expected to fulfil the
requirements for minimum emissions to the environment, as set by the Government of
the day and enforced by the regulation bodies such as HMIP or the NRA, but they must
gain acceptance and trust from industry and the public as an effective and safe means of
treating contaminated soil. With this latter factor to consider, conversion of contaminated
soil into activated carbon would appear to fulfil these requirements since it is an
established industrial process.
260
7.1.5 Activated carbon can be considered to be a commodity. Its manufacture is a routine
industrial process, and it has many applications ranging from environmental emission
control to medical applications and chemical purification. The common feedstocks are
non-renewable, and especially where peat is concerned, the continued usage is of much
concern. The utilisation of particularly "difficult" waste materials for activated carbon
manufacture would not only assist in the protection of the non-renewable feedstock, but
would also reduce the pollution load on the environment from the disposal of these
wastes.
7.1.6 The composition of contaminated soils is very heterogeneous, especially those from
gaswork sites. In the samples studied, typical contaminants included cyanides, sulphur,
coal tars, heavy metals and asbestos and these were, in some instances, present as over
10% of the soil, by weight.
7.1.7 The potential for the conversion of contaminated soil into activated carbon was
successfully demonstrated with samples from a gasworks site in the UK, using several
activation agents. Chemical activation using ZnCU was the most able activation agent, and
produced reactions at the lowest temperatures, with the greatest carbon yields. This is
important because activated carbon manufacture is an energy intensive process, with any
achievable reductions in energy requirements whilst maximising the product yield being
of economic significance.
7.1.8 The organic content of the soil was converted into a porous, adsorbent solid char and oils
and gaseous products. Zinc chloride was the most effective activation agent at achieving
this conversion, based upon the N, adsorption surface area and phenol and 4-nitrophenol
aqueous adsorption ability of the resulting carbons. Leach testing of the raw and
converted soils further confirmed that the organic contamination was converted into a
benign, non-leaching form, as testified by the lack of TOC in the leachates.
7.1.9 The surface area of the carbons produced from the soils was very variable, with factors
such as activation agent dosage, carbon content, degree and types of contaminants present
in the soil affecting the porosity of the final product, hence influencing the surface area
development by each sample. One of the soils was converted into activated carbons which
exhibited adsorption and surface area properties similar to commercial products.
7.1.10 Sulphur species were discovered to play an important part in the chemistry of the
261
activation process. In the absence of ZnClz, the soils containing increased amounts of
these species displayed large increases in the N; surface area and adsorption of organic
aqueous pollutants exhibited by the sample. This self-activation was significant because
it means that activation agents would not always be necessary for producing activated
carbon from contaminated soil. Also, alternative, cheaper chemical activation agents could
possibly be used in the place of ZnCl .
7.1.11 The ZnCl; also affected the non-organic contamination in the soil. This was especially
noticeable for the cyanides, which were almost completely eliminated by the processing.
Similarly, sulphate and sulphur forms in the soils were reduced, although not to the same
extent as the cyanides.
7.1.12 The metals which were present initially in the soil were found to be associated with the
porous carbon product in non-leachable forms. ICP-AES analysis of the acid wash liquors
showed that only some of the metal species were removed in the washing stage, and leach
testing further illustrated that these metals were not in a leachable form. This implied that
they were associated with the solid carbon lattice. XRF analysis confirmed that many of
the metals from the soil were retained by the product.
7.1.13 The optimum process conditions for producing activated carbons from gaswork site soil
were established as; 0.5:1 /w ZnCU : soil ratio, carbonising for 1 hour and activating
for 2 hours. The temperatures used for these two stages were dependent upon the soil
composition, but from the different soil samples examined were observed to occur within
the temperature range 150-275 °C for carbonisation and 450-525 °C for activation.
7.1.14 Studies of the metal adsorption ability of the soil carbon established that these adsorbents
have affinities for cationic metal species which are superior to the commercial carbons
considered. This was proposed to be due to several complementary factors. The major
influence was that the ZnClj activated products appeared to be 'L' type carbons, so
possessing a negative surface charge which naturally attracted the cationic metals.
Secondly, the contamination of the soil contributed to the surface functionalities of these
carbons creating reactive groups to which the metals possessed reactive affinities, such
as sulphur reacting with Cd(II) and Hg(II). The adsorption of Hg(II) by the soil carbons
was extremely large, with the adsorption from solutions of concentrations approaching
300 ppm being almost total.
262
7.1.15 The study of the simultaneous adsorption of Cd(II) and Hg(II) from a mixed metal
solution showed that there was some competition between these metals for common
adsorption sites. Consequently, the adsorption capacities for both metals was reduced and
Cd(II) was adsorbed less strongly than Hg(II), since the extent of Cd(II) adsorption
reduced by approximately 50% compared to the single metal solutions.
7.1.16 Application of the products to treat landfill leachate and effluent streams from metal
plating plants showed that the soil carbons had an affinity for several heavy metals. It was
also observed that the TOC of the landfill leachate decreased simultaneously, indicating
con-current adsorption of organics and metals in the leachate. The metal plating effluent
was considered to be the most suitable effluent for treatment by the carbons. The metal
adsorption characteristics exhibited by the soil carbon were, in general, equivalent to
commercial carbons.
7.2 Suggestions for further work
During the course of this research, it became obvious that several features of the results warranted
further research, whose investigation was not possible within the scope of this thesis. These
research programmes are briefly discussed below:
7.2.1 The soils used in this research required separation from the rocks and rubble from the
site. The possible need for pretreatment of the soil to remove these large particles
followed by their decontamination must be considered. A soil washing facility may be
required to concentrate the contamination into the finer particle sizes whilst leaving the
larger material in a cleaned state suitable for reuse on the site. Disposal of this material
to landfill would not be an acceptable solution.
7.2.2 The carbonisation and activation stages also produced gases and oils which were not
considered, but off-gas analysis by TG-FTIR would provide absolute confirmation that
the assignments made to the weight and energy changes noted during thermal analysis
were correct. They could also provide an insight into the types of reactions occurring
within the soil, which may assist in the modification of the process to further enhance the
elimination and or entrapment of the soil contaminants within the product.
7.2.3 Similarly, by collecting the oils from the process, their organic composition could be
studied by FTIR. Any metals which evaporate from the system, such as Hg, Pb and As,
263
and might be a problem for the off-gas treatment system, could be analyzed for by
digestion of the oils followed by AAS or ICP-AES analysis. Knowing the composition
of these gases and oils would enable the design of the off-gas treatment system to be
tailored to the needs of this feedstock. Finally, the oils may have value as a secondary
liquid fuel, whether for the process or in incinerators or cement kilns. Thus establishment
of the calorific value of this oil, its combustion characteristics and how these are affected
by changes to the process operational factors would be important for commercial
reasons.
7.2.4 The sulphur species were thought to contribute significantly to the adsorption
characteristics of the carbons. The sulphur incorporated into the carbon lattice may be
present as inorganic and organic forms. The establishment of the type, location and form
of the sulphur present in the carbons would help to ascertain the rfile which sulphur plays
in the adsorption and activation characteristics exhibited by the samples.
7.2.5 In the metal adsorption study the effects of anions present in the metal plating liquors and
the landfill leach ate were not considered. These anions could affect the adsorption
properties of the carbon by binding to sites used by the metals. Similarly, the removal of
problem anions may be facilitated, such as CN" binding to positively-charged sites.
7.2.6 Removal of Hg(II) from solution is problematic, especially at low concentrations. The
carbon produced from the gaswork soil exhibited high adsorption affinity for Hg(II). The
nature of this affinity has been attributed to three possible mechanisms, but the exact
mechanism does require to be established. It is also important to know if the Hg(II)
attached to the carbon can be readily removed, or whether secondary treatment of the
Hg(II) loaded carbon may be necessary. A common method of disposal of Hg(II) bearing
wastes is to solidify them with cement and other additives, but it is recognised that Hg(II)
does not become involved in the reactions of the cement but is encapsulated in the matrix.
The Hg(II) can subsequently be released through simple leaching of the cement system.
Adsorption of the Hg(II) on to the soil carbons prior to solidification may surmount this
problem.
7.2.7 The soils both before and after conversion into carbonaceous adsorbents showed
reasonable resistance to leaching, but the problems of insufficiently washing the carbons
to remove the ZnCU used as the activation agent meant that Zn(II) was a major species
in the leaching solution. Further work is required to examine the effects of the washing
264
upon the soil carbons and their leaching behaviour, since excessive Zn(II) leaching will
limit their applicability in the treatment of waste waters.
7.2.8 Although the study of the suitability of the soil carbon as a treatment for landfill leachate
was not conclusive, the need exists to investigate the possibility of using the carbons as
an in-situ barrier for a landfill system, possibly by application as a daily coyer. From this
work, the effectiveness of activated carbon in the in-situ removal of organic and metal
species from landfill leachate could be evaluated. The movement toward engineered
containment landfill sites has meant that leachate collection and treatment is a major cost
and liability for landfill owners and operators, whilst the daily covering of the freshly laid
wastes is a UK legal requirement. Currently, soil is used, but this affords very little
additional benefit to the landfill owners, hence adsorbent daily cover would have distinct
advantages. Especial attention would be required to examine the ability of the carbons to
remove organic molecules. It would also be useful to study the function of the carbon as
a possible support for bacterial life within the landfill. This would serve the dual purpose
of slowing the movement of organic species through the landfill whilst also promoting
their degradation.
7.2.9 The final need for further research is to study the economics of the production of this
type of carbon. Without a full understanding of the costs involved in the production and
post-treatment of the products of this process, the potential of this method as a treatment
for contaminated soil cannot be fully assessed.
265
CHAPTER EIGHT
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Srivastava S. K., Tyagi R. and Pant N., 1989, Adsorption of heavy metal ions on carbonaceous material developed from the waste slurry generated in local fertilizer plants. Water Research, 23(9), pp. 1161-1165.
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284
APPENDIX I INTERNATIONAL GUIDANCE VALUES FOR CONTAMINATION IN SOIL
Shown below are the guidance values issued by various organisations and governments to assist in the assessment of the degree of contamination to a particular site.
Table Al l Canadian assessment and remediation criteria. (All values, except pH, in ppm)
Parameter Assessment Remediation Criteria Criteria
Agricultural Residential / Parkland
Commercial / Parkland
pH 6 to 8 6 to 8 6 to 8 6 to 8
As 5 20 30 50
Ba 200 750 500 2000
Cd 0.5 3 5 208
Co 10 40 50 300
Cr (6+) 2.5 8 8 n/a
Cr (total) 20 750 250 800
Cu 30 150 100 500
Hg 0.1 0.8 2 10
Ni 20 150 100 500
Pb 25 375 500 1000
Sb 20 20 20 40
Sn 5 5 50 300
Th 0.5 1 n/a n/a
V 25 200 200 n/a
Zn 60 600 500 1500
CN (free) 0.25 0.5 10 100
CN (total) 2.5 5 50 500
S (elemental) 250 500 n/a n/a Adapted from (CCME, 1991)
285
Table AI.2 Dutch A, B and C values for selected soil pollutants
Substance All concentrations are in mg/kg of dry matter
'A' 'B' ' C Reference value Further investigation is Clean-up is required
required
As - 30 50
Ba 200 400 2000
Cd - 5 20
Co 20 50 300
Cr - 250 800
Cu - 100 500
Hg - 2 10
Mo 10 40 200
Ni - 100 500
Pb - 150 600
Sn 20 50 300
Zn - 500 3000
CN (total- free) 1 10 100
CN" (total- complex) 5 50 500
S (total sulphides) 2 20 200
Total aromatics - 7 70
PAHs (total) 1 20 200
Revisions to the Dutch ABC soil assessment values
The ABC values mentioned above were considered to be flawed. They were not based upon any health or exposure criteria and with respect to the C values: the organic concentrations were over estimated and the metal values were under estimated. As mentioned in section 1.3.6.2, they have now been replaced with the passing of the Soil Protection Act (1994). The new parameters, derived by RIVM, have been established on an ecotoxicological basis. These replacements are classified as follows:
The A value has been replaced with the "Reference norm" The C value has been replaced with the "Intervention value"
The B values are no longer valid.
When applying these values to sites, there is a simple calculation which can be performed:
C*-=CK B
286
Where: A = Factor related to soil particle size (accounts for soil type) B = Factor related to natural organic levels in soil C = Concentration of pollutants detected in soil CV = Corrected value for contaminants in soil which should be compared to the
Reference norm or Intervention values.
This adjusted value is considered to represent the available contamination, accounting for its mobility within the particular site under study.
It is important to note that the Reference norm and Intervention values are regarded as absolute. There is a certain inflexibility in these new standards, when a site is cleaned up to between the two values, the site may be "multifunctional" but in the eyes of the authorities, it is not, because the pollution is not below the Reference norm.
Table AI.3 The revise Dutch reference and intervention values for selected soil pollutants
Substance All values are in mg/kg of dry matter
Substance Reference norm values
("A" values) Intervention values
("C" values)
As 29 55
Ba 200 625
Cd 0.8 12
Co 20 240
Cr 100 380
Cu 36 190
Hg 0.3 10
Mo 10 200
Ni 35 210
Pb 85 530
Sn - -
Zn 140 720
CN" (free) 1 20
CN" (complex pH<5) 5 650
CN (complex pH^5) 5 50
S (Total sulphides) - -
Total aromatics - -
PAHs (total) 1 40 (Adapted from Visser, 1994)
287
Table AI.4 Guidelines for contaminated soils, compiled by the GLC
Parameter Not contaminated
Typical values in mg/kg on air dried soils (except for pH)
Contaminated Slight contamination
Heavy contamination
Unusually heavy
contamination
pH (acid)
6 - 7 5 - 6 4 - 5 2 - 4 < 2
pH (alkali)
7 - 8 8 - 9 9 -10 10- 12 > 12
As 0 -30 30 - 5 0 50 - 100 100 - 500 > 500
Cd 0 - 1 1 - 3 3 - 10 1 0 - 5 0 > 50
Cr 0 - 100 100 -200 200 - 500 500 - 2500 > 2500
Cu 0 - 100 100 -200 200 - 500 500 - 2500 > 2500
Hg 0 - 1 1 - 3 3 - 10 1 0 - 5 0 > 50
Mn 0 - 500 500 • - 1000 100 - 2000 2000-1.0% > 1.0%
Ni 0 • -20 20 - 5 0 50 - 200 200 - 1000 > 1000
Pb 0 - 500 500 • - 1000 1000 - 2000 2000-1.0% > 1.0%
Sb 0 • - 30 30 - 5 0 50 - 100 100 - 500 > 500
V 0 - 100 100 -200 200 - 500 500 - 2500 > 2500
Zn 0 - 250 250 -500 500 - 1000 1000-5000 > 5000
CN (total)
0 - 5 5 • - 25 25 - 250 250 - 500 > 500
CN (free)
0 - 1 1 - 5 5 - 5 0 50 - 100 > 100
Coal Tar 0 - 500 500 - 1000 1000 - 2000 2000 - 1.0% > 1.0%
S (free) 0 - 100 100 -500 500 - 1000 1000-5000 > 5000
Sulphate 0 - 2000 2000-5000 5000-1.0% 1.0-5.0% > 5.0% Adapted from Kelly, 1980.
288
Table AI.5 Tentative 'trigger concentrations' for selected inorganic contaminants
Conditions 1. This table is invalid if reproduced without the conditions and footnotes.
2. All values are for concentrations determined on 'spot' samples based on an adequate site investigation carried out prior to development. They do not apply to analysis of averaged, bulked or composited samples, nor to sites which have already been developed. All proposed values are tentative.
3. The lower values in group A are similar to the limits for metal content of sewage sludge applied to agricultural land, the values in group B are those above which phytoxicity is possible.
4. If all sample values are below the threshold concentrations then the site may be regarded as uncontaminated as far as the hazards from these contaminants are concerned and development may proceed. Above these concentrations, remedial action will be required or the form of development changed.
Contaminants Planned Uses Trigger concentrations (mg/kg air dried soil)
GrouD A:Contaminants which mav nose hazards to health Threshold Action
As Domestic gardens, allotments 10 *
Parks, playing fields, open space. 40 *
Cd Domestic gardens, allotments 3 *
Parks, playing fields, open space. 15 *
Cr (VI) (1) Domestic gardens, allotments
Parks, playing fields, open space.
25 *
*
Cr (total) Domestic gardens, allotments 600 *
Parks, playing fields, open space. 1000 *
Pb Domestic gardens, allotments 500 *
Parks, playing fields, open space. 2000 *
Hg Domestic gardens, allotments 1 *
Parks, playing fields, open space. 20 *
Se Domestic gardens, allotments 3 *
Parks, playing fields, open space. 6 *
GrouD B: Contaminants which are ohvtotoxic but not normally hazards to health
Boron (water soluble) (3) 3 *
Copper (4,5) Any uses where plants are to be 130 *
Nickel (4,5) grown (2,6) 70 *
Zinc (4,5) 300 *
Notes continued over page.
289
Notes
* Action concentration will be specified in the next edition of ICRCL 59/83 1. Soluble hexavalent Chromium extracted by 0.1 M HCl at 37 °C; solution adjusted to pH 1.0 if alkaline substance
2. the soil pH is assumed to be about 6.5 and should be maintained at this value. If the pH falls, the toxic effects and the uptake of these elements will be increased.
3. Determined by standard ADAS method (soluble in hot water).
4. Total concentration (extractable by HNO3/HCIO4).
5. The phytotoxic effects of copper, nickel and zinc may be additive, the trigger values given here are those applicable to the 'worst-case': phytotoxic effects may occur at these concentrations in acid, sandy soils. In neutral or alkaline soils phytotoxic effects are unlikely at these concentrations
6. Grass is more resistant to phytotoxic effects than are most other plants and its growth may not be adversely affected at these concentrations
From ICRCL, 1987.
290
Table AI.6 Tentative 'trigger concentrations' for contaminants associated with former coal carbonisation sites
Conditions
1. This table is invalid if reproduced without the conditions and the footnotes.
2. All values are for concentrations determined on 'spot' samples based on an adequate site investigation carried out prior to development. They do not apply to analysis of averaged, bulked or composited samples, nor to sites which have already been developed.
3. Many of these values are preliminary and will require regular updating. They should not be applied without reference to the current edition of the report 'Problems arising from the development of gas works and similar sites'. (1)
4. If all sample values are below the threshold concentrations then the site may be regarded as uncontaminated as far as the hazards from these contaminants are concerned and development may proceed. Above these concentrations, remedial action will be required or the form of development changed.
Contaminants Proposed Uses Trigger Concentrations (mg/kg air-dried soil)
Threshold Action
Polyaromatic hydrocarbons (1,2)
Domestic gardens, allotments, play areas. 50 500
Landscaped areas, buildings, hard cover. 1000 10000
Phenols Domestic gardens, allotments. 5 200
Landscaped areas, buildings, hard cover. 5 1000
Free Cyanide Domestic gardens, allotments, landscaped areas. 25 500
Buildings, hard cover. 100 500
Complex cyanides Domestic gardens, allotments. 250 1000
Landscaped areas. 250 5000
Buildings, hard cover. 250 NL
Thiocyanate (2) All proposed uses. 50 NL
Sulphate Domestic gardens, allotments, landscaped areas 2000 10000
Buildings (3). 2000(3) 50000 (3)
Hard cover. 2000 NL
Sulpide All proposed uses. 250 1000
Sulphur All proposed uses. 5000 20000
Acidity (pH less than)
Domestic gardens, allotments, landscaped areas.
Buildings, hard cover.
pH5
NL
pH3
NL
See over for further notes.
291
Notes
NL: No limit set as the contaminant does not pose a particular nazard for this use. (1); Used here as a marker for coal tar, for analytical reasons. See 'Problems arising from the redevelopment of gas works and similar sites' Annex Al. (2): See 'Problems arising from the redevelopment of gas works and similar sites' for the details of analytical methods. (3): See also BRE Digest 250: Concrete in sulphate-bearing soils and groundwater.
From ICRCL, 1987.
292
APPENDIX n THE LANGMUIR EQUATION
(Adapted from Gregg and Sing, 1982 and Shaw, 1992)
This equation was proposed by Irving langmuir in 1916. It was recognised that intermolecular
forces decreased rapidly with distance, thus Langmuir assumed that molecules adsorbed onto a
surface would not be more than one molecular layer deep. This is the situation, generally, for
chemisorption or physical adsorption at low pressure and relatively high temperatures (>RTP)
(Shaw, 1992b). Consequently, Langmuir made three assumptions in the derivation of the
equation: (a) Only monolayer adsorption occurs
(b) Adsorption is localised
(c) The heat of adsorption is independent of surface coverage
Assumption (c) implies dynamic equilibrium, the rate at which molecules arriving from the gas
phase and condensing on bare sites is equal to the rate at which molecules evaporate from
occupied sites, (heat of adsorption is independent of surface coverage).
Assuming that the fraction of occupied sites = 6 and the fraction of bare sites =
8i+8,=l 1
Hence, the rate of condensation on unit area of surface equals:
a lq>% 2
where: p = pressure at equilibrium
ai = condensation coefficient (the fraction of molecules which condense on a surface)
k = constant from kinetic theory of gases:
k - J ^ 3
Evaporation of an adsorbed molecule from the surface is essentially an activated process in which
the energy of activation may be equated to the isosteric heat of adsorption q,. (Adsorption of
vapours on to solids is an exothermic process, the energy value of which is essentially equal to
the heat of condensation. As these energy values are very low, equilibrium is normally very
rapidly attained.) Therefore the rate of evaporation from unit area of surface equals:
z„ = number of sites per unit area
z^6 = number of adsorbed molecules (the fraction of occupied sites)
293
= frequency of oscillation of the molecules in a direction normal to the surface
At equilibrium:
since = 1 - d , by substitution into (5) gives (6):
Multiplying out (6) gives (7):
=z^ei Ui
Rearranging (7):
a^kp=a^lq}Q^ +Z^8iUie
Dividing by and rearranging (8):
ayJgy 01=
So, if n (in moles) is the amount adsorbed on Ig of adsorbent, then:
where n„ = monolayer capacity
Insertion of (9) into (8) gives (11), the Langmuir Equation:
n _ Bp 11
where:
12 V .
This is applicable to adsorption which is confined to a monolayer. B is an empirical constant and
cannot be evaluated from the relationship in (12) (Gregg and Sing, 1982).
A plot of p/V verses p should give a straight line, slope 1 / V„ and an intercept of 1 / BV„ on
the p/V axis.
294
However, the Langmuir equation is also used for assessing adsorption from aqueous solution, to
this end, the Langmuir equation can be adapted as follows:
If 01 equals:
n _ jx/m)
Where: x = Amount (concentration in moles) of adsorptive attached
to the adsorbent
m = Mass of adsorbent used
(x/m)^ = Monolayer capacity of adsorbent (Concentration/weight)
By substituting (13) into (11), and replacing p with c, the concentration of adsorptive remaining
in solution at equilibrium, gives:
;c _ m 1+Bc
An alternative form of the Langmuir Equation is:
1+BC e
Where: q, = (x/m)
Qmax = (x/m)^
C. = c
To plot the Langmuir Equation, the linear form is preferred:
: 1 16 W"") <3™
The monolayer capacity, Q ^ , can be estimated either directly from the isotherm (according to
the method of Giles and Nakhwa (1962)) or indirectly by applying the Langmuir equation. If the
effective area occupied by each adsorbed molecule is known, the specific surface area of the
solid can be calculated, using the relationship:
SSA-N/m)u^^, "
Where: SSA = Specific surface area of solid
Na = Avogadros number
= Surface area of adsorbate molecule
295
APPENDIX n i THE BET EQUATION
(After Gregg and Sing, 1982 and Castellan, 1983)
Originally proposed by Brunauer, Emmett and Teller in 1938, the BET equation is an extension
of the Langmuir equation, which allows for the occurrence of multilayer adsorption which
invariably occurs within most adsorption systems. Gregg and Sing (1982) report that Brunauer
Emmett and Teller made three simplifying assumptions in the derivation of this equation:
(a) the molar heat of adsorption equals the molar heat of condensation in all layers
except the first;
(b) identical evaporation-condensation conditions exist in all layers except the first;
(c) the number of adsorbed layers becomes infinite when P = P (the adsorbate
condenses/liquifies on the surface).
Assuming that the first adsorbed layer forms thus:
A(g) + S # A,S
Where: Kj = Equilibrium constant
01 = Fraction of sites covered by a single molecule
6 — Fraction of vacant sites
Additional layers form by sitting on top of one-another, giving a series of multilayers. Each layer
possessing its own equilibrium constant. Thus:
nr ^2 2 A(g) + AS # AzS 2 = - ^ ^
0 A(g) +A2 S # A3S ^3=
Q2P
fr ^4 A(g) + A3S # A4S
A(g) +A„., S # A,S 0 p ^
Where A4S is a surface site with 4 molecules stacked upon it.
The interaction A + S -*• AS (formation of layer 1) is unique, however, AS + A -* A^S is
effectively the interaction occurring between two A molecules in a liquid environment. The same
follows for formation of the 3rd, 4th layers and so forth up to the nth layer. Hence their
equilibrium constants "K^", for the transition A(g) # A(l), should be identical. The BET theory
thus assumes that Ko = K, = K4 = ... = K where:
296
K=— po
and P° = equilibrium vapour pressure for the liquid.
Using K, we can calculate the 6-, values by rearranging (2) - (5). Thus we get (7) - (10)
respectively:
Q2=QiKP 1
8
9
e . = e . . . B ' 10
Combining (7) and (8) gives:
Ultimately, we arrive at: 0 i=0i (^) ' ^
The sum of all the fractions must be equal to unity:
13 1=1
Substituting (12) into (13) gives: ' 14 I
Let KP = X, and by substituting into (14) we get (15): 1=0,,+0i(1+j:+jc^+j:^+j;''+...) 15
Assuming that this process can proceed indefinitely, as n -*• oo the series is the expansion of
l/(l-x). Hence, substituting this expression into (15) gives (16):
1=0 +— 16 ^ \-X
The equilibrium conditions for the first adsorption layer was given by equation (1), rearranging equation (1) gives equation (17):
0 = A 17 V KiP
If we let c = K; / K, =» K] = c K. Substituting for Kj in (17) gives (18):
! L = ! I 18 ^ cKP ac
297
Substituting (18) into (16) gives (19): l = ^ + J i _ = e , ( i - + - i - ) 19
cx (l-ar) * Of (1-j:)
Re-arranging gives (20): , 6 20
Let: N = Total number of molecules adsorbed per unit mass of adsorbent
c,= The total number of surface sites per unit mass
Therefore: c.0j = Number of sites carrying 1 molecule
c.@2 = Number of sites carrying 2 molecules
c,0„ = Number of sites carrying n molecules
So: ^ i^=c/iei+20j+303...)=cXi0f 21
i
Since KP = x, by substitution, (12) becomfij=0i(;t)'"^ 22
Substituting (21) into (22): =0,61(1 +2%+3%^+...) 23 i=i
It can be seen that (23) contains the derivative of the series shown in (15), as shown by (24):
l+2x+3j:^+...=-f(l+^+.'J^+J^^+.")~(T^)=^-T dx dx \-x (i_;r)Z
Substituting (24) into (23):
25
If complete monolayer coverage of the surface had occurred, then N„ molecules would be
adsorbed, where N„ = c„ thus: N - 5 1 l 26
Substituting (20) into (26):
(l-xY
J / - 27 ( l - ;c) [ l+(c-W
The amount of gas adsorbed is normally expressed as the volume at STP, which is proportional
to N, where N/N„ = v I v^, hence substituting for N in (27):
(I-A:)[l+(c-W
Substituting KP for x (see equation (15)) into (28) gives (29):
298
V- 28
v= 29
From (6), we know K = 1/P°, thus substituting for K into (29) gives the BET equation:
p
v=- ' ( l - ^ ) [ l . ( c - l ) ^ ]
Multiplying both sides of (30) by: {(1-P/P°)/(P/P°)} gives (31):
V.C
l+(c-l)— — P' P'
Taking reciprocals of both sides gives the linear form of the BET Equation:
J _ + ( E d ) * Z _ v ( l - L ) v„c v„c p '
po
Alternatively this can be written as:
p = J _ + ( E d ) . _ & viP^-P) v„c v„c po
P" 1 1
^ ( T - 1 )
30
31
32
33
1 _ 1 ^(c-1)^ P The Omnisorp uses equation (34): .p' ~ v c v c p°
This is derived by dividing the left hand side of equation (33) equation by P/P°:
35
A plot of (35) vs P/P° should be a straight line, slope (s) = (c - 1) / »/„ c
intercept (i) = i /
and c can be calculated thus: = l/(s + i)
c = (s/i) + 1
From v^, N„ can be calculated at STP:
299
^ 36 O.Q22414m^lmol
Since = the number of molecules required to cover a unit mass with a monolayer, if "a", the
area covered by one molecule is known, then the area of each unit of mass in the system of
investigation can be calculated (mVg = N„a).
Returning to the equilibrium constants K, and K; and considering them in terms of the standard
differences in Gibbs energies for the transformations gas-solid, gas-liquid.
38 1 - 3 7
AGi" = Standard Gibbs energy of adsorption of the first layer
AG°ii, = Standard Gibbs energy of liquefaction
Dividing (37) by (38) gives c: c=—=e K
Since: AG°i = AH", - TAS", and AG^, = AH°u, - TAS'g, (T = constant)
Assuming that entropy loss is similar, irrespective of which layer a molecule sits in, we can say
that: AS°i » A S \ . Thus, substituting AH" for AG" in (39) gives (40):
4 Q
c=e
The heat of liquefaction, AH°u,, is the negative of the heat of vaporisation, AH°^. Hence we can
say that A H \ = -AH%, so substituting into (40) gives (41):
41 c=<
Taking logs and re-arranging:
c=e
LH°=-LH°-RI\rc 42
AH°vap is a known adsorbate parameter, therefore, AH'i can be calculated from "c". Since "c"
must be greater than 2 for the BET treatment to be applicable (Gregg and Sing, 1982), then
AH°i < AH");,, and adsorption by the first layer is exothermic compared to liquefaction.
However, it is very important to be aware of the approximate relationship of equations (41) and
(42), since it is recognised that AH", for the first layer varies with surface coverage, especially
on the first layer (in deviation from assumption (c) used for the derivation of the Langmuir
equation (see Appendix II).
300
APPENDIX IV
FREUNDLICH AND LANGMUIR PHENOL ISOTHERM PLOTS FOR THE
SOIL AND COMMERCIAL CARBONS
The following figures represent the data used to calculate the Freundlich and Langmuir isotherm
constants, given in table 4.7, for phenol adsorption by the soil carbons prepared from STAOl-
STA05 and Gl, plus the two commercial activated carbons Nor it SA4 and Type C used for
comparison purposes.
Figure AIV.l Freundlich isotherms for phenol adsorption by STAOl carbons 0
I.
Run 002 "••Run Oil -*-Run 013
Figure AIV.2 Langmuir isotherms for phenol adsorption by STAOl carbons 100
S 40
l/Ce(l/mMol)
-Run 002 •••Run Oil Run 013
301
Figure AIV.3 Freundlicli isotherms for phenol adsorption by STA02 carbons 0
I
Run 005 -••Run 010 Run 014
Figure AIV.4 Langmuir isotherms for phenol adsorption by STA02 carbons 100
G 40
5 10
l/Ce(l/inMol)
-Run 005 • R u n 010 Run 014
Figure AIV.5 Freundlich isotherms for phenol adsorption by STA03 carbons 0
I,
LogCe
Run 004 -"-Run 008 Run 012
302
Figure AIV.6 Langmuir isotherms for phenol adsorption by STA03 carbons
a 80
10 15
1/Ce (1/mMbl)
-Run 004 * Run 008 -^-Run 012
Figure AIV.7 Freundlich isotherms for phenol adsorption by STA04 carbons 0.5
B -0.5
LogCe
Run 001 -"-Run 007 -*-Run 015
Figure AIV.8 Langmuir isotherms for phenol adsorption by STA04 carbons 250
S 100
100 130
1/Ce (l/mMol)
-Run 001 -"-Run 007 -*-Run 015
303
Figure AIV.9 Freundlich isotherms for phenol adsorption by STA05 carbons 0
I s
•Run025 -••Run027 Run006 -f-RunOOP -»-Run034
Figure AIV.IO Langmuir isotherms for phenol adsorption by STA05 carbons
I. J 3?. 40
4 6
1/Ce (l/mMol)
-Run 025 -"-Run 027 -*-Run 006 -*-Run009 -e-Run 034
Figure AIV.ll Freundlich isotherms for phenol adsorption by G1 carbons 0
LogCe
Run 041 • R u n 045
304
Figure AIV.12 Langmuir isotherms for phenol adsorption by G1 carbons
J 40
l/Ce(l/inMol)
Run 041 • R u n 045
Figure AIV.13 Freundlich isotherms for phenol adsorption by the commercial carbons
1
Log Ce
NoritSA4 Chemviron "^pe C
Figure AIV.14 Langmuir isotherms for phenol adsorption by the commercial carbons
200 400 600
1/Ce (1/mMol)
-NoritSA4 -^TypeC
1000
305
APPENDIX V FREUNDLICH AND LANGMUIR 4-NITROPHENOL ISOTHERM PLOTS FOR
THE SOIL AND COMMERCIAL CARBONS
The following figures represent the data used to calculate the Freundlich and Langmuir isotherm
constants, given in table 4.8, for 4-nitrophenol adsorption by the soil carbons prepared from
STA01-STA05 and Gl, plus the two commercial activated carbons Norit SA4 and Type C used
for comparison purposes.
Figure AV.l Freundlich isotherms for 4-nitrophenol adsorption by STAOl carbons 0^
Run 002 -"-RunOil Run013
Figure AV.2 Langmuir isotherms for 4-nitrophenol adsorption by STAOl carbons 80
60
» 40
100 150 l/Ce(]/mMol)
• Run 002 Run Oil Run 013
306
Figure AV.3 Freundlich isotherms for 4-nitrophenol adsorption by STA02 carbons 0
Run 005 -"-Run 010 -*-Run014
Figure AV.4 Langmuir isotherms for 4-nitrophenol adsorption by STA02 carbons 250
S 100
100 150 200 1/Ce (]/mMol)
-Rim 005 Rim 010 -»-Rim 014
Figure AV.5 Freundlich isotherms for 4-nitrophenol adsorption by STA03 carbons OJ
Run 008 Run 012
307
Figure AV.6 Langmuir isotherms for 4-nitrophenol adsorption by STA03 carbons 60
« 30
> 20
200 300 400 l/Ce(]AnMol)
-Run 008 -"-Run 012
Figure AV.7 Freundlich isotherms for 4-nitrophenol adsorption by STA04 carbons 1
5
0^
0
-0^
-1
-1^
- 2
•
2 ^
#
-3 - 2 -1
LogCe
•Run001 -"-Run007 Run015
Figure AV.8 Langmuir isotherms for 4-nitrophenol adsorption by STA04 carbons
> . 40
1000 1500 l/Ce(l/mMol)
2500
•Run 001 -"-RunOO? -*-Rim 015
308
Figure AV.9 Freundlich isotherms for 4-nitrophenol adsorption by STA05 carbons 0
•Run025 -•-Run027 -*-Run006 Run009 -e-Run034
Figure AV.IO Langmuir isotherms for 4-nitrophenol adsorption by STAGS carbons 250
I
100
l/Ce(l/mMol)
•Run 025 -"-Rim027 -*-Rim006 -f-Run009 -e-Run034
Figure AV.ll Freundlich isotherms for 4-nitrophenol adsorption by G1 carbons 0
-2 -IJ -0.5 0 0.5 LogCe
•Run 041 -"-Run 045
309
Figure AV.12 Langmuir isotherms for 4-nitrophenol adsorption by G1 carbons 80
^ 60
S 40
30 40
1/Ce (l/mMol)
•Rim 041 -••Run 045
figure AV.13 Freundlich isotherms for 4-nitrophenol adsorption by the commercial carbons
S
NontSA4 -*-lype C
Figure AV.14 Langmuir isotherms for 4-nitrophenol adsorption by the commercial carbons
% 40
2 4 6 8
1/Ce (Thousands) (l/mMol)
•NoritSA4 " • T ^ e C
310
APPENDIX VI PUBLICATIONS DERIVED F R O M THIS RESEARCH
Fowler G. D., Ouki S. K., Sollars C. J. and Perry R., 1993, The conversion of gasworks contaminated soil into a low-cost adsorbent using a novel application of thermal treatment technology. Contaminated Soil '93: Fourth International KfK/TNO Conference on Contaminated Soil, Berlin, 3-7th May 1993.
Fowler G. D., Ouki S. K., Sollars C. J. and Perry R., 1994, Thermal conversion of gasworks contaminated soil into carbonaceous adsorbents, Journal of Hazardous Materials, 39, pp. 281-300.
311