WASTE MATERIALS IN CONSTRUCTION Putting Theory into Practice
WASTE MATERIALS IN CONSTRUCTION Putting Theory into Practice
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Studies in Environmental Science 71
WASTE MATERIALS IN CONSTRUCTION Putting Theory into Practice Proceedings of the International Conference on the Environmental and Technical Implications of Construction with Alternative Materials, WASCON '97, Houthem St. Gerlach, The Netherlands, 4-6 June 1997
Edited by
J.J.J.M. Goumans ISCOWA The Netherlands
G.J. Senden ISCOWA The Netherlands
H.A. van der Sloot
Netherlands Energy Research Foundation (ECN) Petten, The Netherlands
1997 ELSEVIER Amsterdam
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Shannon
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ELSEVIER SCIENCE B.V. Sara Burgerhartstraat 25 P.O. Box 211, 1000 AE Amsterdam, The Netherlands
ISBN 0-444-82771-4 © 1997 ELSEVIER SCIENCE B.V. All rights reserved. No part of this publication may be reproduced, stored in a retrieval system or transmitted in any form or by any means, electronic, mechanical, photocopying, recording or otherwise, without the prior written permission of the publisher, Elsevier Science B.V., Copyright & Permissions Department, P.O; Box 521, 1000 AM Amsterdam, The Netherlands. Special regulations for readers in the U.S.A. - This publication has been registered with the Copyright Clearance Center Inc. (CCC), 222 Rosewood Drive Danvers, Ma 01923. Information can be obtained from the CCC about conditions under which photocopies of parts of this publication may be made in the U.S.A. All other copyright questions, including photocopying outside of the U.S.A., should be referred to the owner, Elsevier Science B.V., unless otherwise specified. No responsibility is assumed by the publisher for any injury and/or damage to persons or property as a matter of products liability, negligence or otherwise, or from any use or operation of any methods, products, instructions or ideas contained in the material herein. This book is printed on acid-free paper. Printed in The Netherlands
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Atmospheric Pollution 1978 edited by M.M. Benarie Air Pollution Reference Measurement Methods and Systems edited by T. Schneider, H.W. de Koning and L.J. Brasser Biogeochemical Cycling of Mineral-Forming Elements edited by P.A. Trudinger and D.J. Swaine Potential Industrial Carcinogens and Mutagens by L. Fishbein Industrial Waste Management by S.E. Jorgensen Trade and Environment: A Theoretical Enquiry by H. Siebert, J. Eichberger, R. Gronych and R. Pethig Field Worker Exposure during Pesticide Application edited by W.F. Tordoir and E.A.H. van Heemstra-Lequin Atmospheric Pollution1980 edited by M.M. Benarie Energetics and Technology of Biological Elimination of Wastes edited by G. Milazzo Bioengineering,Thermal Physiology and Comfort edited by K. Cena and J.A. Clark Atmospheric Chemistry. Fundamental Aspects by E. Meszaros Water Supply and Health edited by H. van Lelyveld and B.CoJ. Zoeteman Man under Vibration. Suffering and Protection edited by G. Bianchi, K.Vo Frolov and A. Oledzki Principles of Environmental Science and Technology by S.E. Jorgensen and I. Johnsen Disposal of Radioactive Wastes by Z. Dlouh~/ Mankind and Energy edited by A. Blanc-Lapierre Quality of Groundwater edited by W. van Duijvenbooden, P. Glasbergen and H. van Lelyveld Education and Safe Handling in Pesticide Application edited by E.A.H. van Heemstra-Lequin and W.F. Tordoir Physicochemical Methods for Water and Wastewater Treatment edited by L. Pawlowski Atmospheric Pollution 1982 edited by M.M. Benarie Air Pollution by Nitrogen Oxides edited by T. Schneider and L. Grant Environmental Radioanalysisby H.A. Das, A. Faanhof and H.A. van der Sloot Chemistry for Protection of the Environment edited by L. Pawlowski, A.J. Verdier and W.J. Lacy Determination and Assessment of Pesticide Exposure edited by M. Siewierski The Biosphere: Problems and Solutions edited by T.N. Veziro~lu Chemical Events in the Atmosphere and their Impact on the Environment edited by G.B. Marini-Bettolo Fluoride Research 1985 edited by H. Tsunoda and Ming-Ho Yu Algal Biofouling edited by L.V. Evans and K.D. Hoagland Chemistry for Protection of the Environment 1985 edited by L. Pawlowski, G. Alaerts and W.J. Lacy Acidification and its Policy Implications edited by T. Schneider Teratogens: Chemicals which Cause Birth Defects edited by V. Kolb Meyers Pesticide Chemistry by G. Matolcsy, M. Nadasy and Y. Andriska Principles of Environmental Science and Technology (second revised edition) by S.E. JQrgensen and I. Johnsen Chemistry for Protection of the Environment 1987 edited by L. Pawlowski, E. Mentasti, W.J. Lacy and C. Sarzanini Atmospheric Ozone Research and its Policy Implications edited by T. Schneider, S.D. Lee, G.J.R. Wolters and L.D. Grant
36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60 61 62 63 64 65 66 67 68 69 70
Valuation Methods and Policy Making in Environmental Economics edited by H. Folmer and E. van lerland Asbestos in Natural Environment by H. Schreier How to Conquer Air Pollution. A Japanese Experience edited by H. Nishimura Aquatic Bioenvironmental Studies: The Hanford Experience, 1944-1984 by C.D. Becker Radon in the Environment by M. Wilkening Evaluation of Environmental Data for Regulatory and Impact Assessment by S. Ramamoorthy and E. Baddaloo Environmental Biotechnology edited by A. Blazej and V. Privarova Applied Isotope Hydrogeology by F.J. Pearson Jr., W. Balderer, H.H. Loosli, B.E. Lehmann, A. Matter, Tj. Peters, H. Schmassmann and A. Gautschi Highway Pollution edited by R.S. Hamilton and R.M. Harrison Freight Transport and the Environment edited by M. Kroon, R. Smit and J.van Ham Acidification Research in The Netherlands edited by G.J. Heij and T. Schneider Handbook of Radioactive Contamination and Decontamination by J. Severa and J. BAr Waste Materials in Construction edited by J.J.J.M. Goumans, H.A. van der Sloot and Th.G. Aalbers Statistical Methods in Water Resources by D.R. Helsel and R.M. Hirsch Acidification Research: Evaluation and Policy Applications edited by T.Schneider Biotechniques for Air Pollution Abatement and Odour Control Policies edited by A.J. Dragt and J. van Ham Environmental Science Theory. Concepts and Methods in a One-World, Problem-Oriented Paradigm by W.T. de Groot Chemistry and Biology of Water, Air and Soil. Environmental Aspects edited by J. T61gyessy The Removal of Nitrogen Compounds from Wastewater by B. Halling-Sorensen and S.E. JQrgensen Environmental Contamination edited by J.-P. Vernet The Reclamation of Former Coal Mines and Steelworks by I.G. Richards, J.P. Palmer and P.A. Barratt Natural Analogue Studies in the Geological Disposal of Radioactive Wastes by W. Miller, R. Alexander, N. Chapman, I. McKinley and J. Smellie Water and Peace in the Middle East edited by J. Isaac and H. Shuval Environmental Oriented Electrochemistry edited by C.A.C. Sequeira Environmental Aspects of Construction with Waste Materials edited by J.J.J.M. Goumans, H.A. van der Sloot and Th. G. Aalbers. Caracterization and Control of Odours and VOC in the Process Industries edited by S. Vigneron, J. Hermia, J. Chaouki Nordic Radioecology. The Transfer of Radionuclides through Nordic Ecosystems to Man edited by H. Dahlgaard Atmospheric Deposition in Relation to Acidification and Eutrophication by J.W. Erisman and G.P.J. Draaijers Acid Rain Research: do we have enough answers? edited by G.J. Heij and J.W. Erisman Climate Change Research. Evaluation and Policy Implications edited by S. Zwerver, R.S.A.R. Rompaey, M.T.J. Kok and M.M. Berk Global Environmental Biotechnology edited by D.L. Wise Municipal Solid Waste Incinerator Residues by A.J. Chandler, T.T. Eighmy, J. Hartlen, O. Hjelmar, D.S. Kosson, S.E. Sawell, H.A. van der Sloot and J. Vehlow Freshwater and Estaurine Radioecology edited by G. Desmet, R.J. Blust, R.N.J. Comans, J.A. Fernandez, J.Hilton, and A. de Bettencourt Acid Atmospheric Deposition and its Effects on Terrestrial Ecosystems in The Netherlands edited by G.J. Heij and J.W. Erisman Harmonization of Leaching/Extraction Tests edited by H.A. van der Sloot, L. Heasman and Ph. Quevauviller
vii
Dear colleague, The international society ISCOWA herewith presents the pro ceedings of the international conference WASCON'97, which has been held from June 4 till June 6, 1997 in Valkenburg the Netherlands. SCOPE OF THE C O N F E R E N C E Many western countries are still facing the problem of a growing burden of waste materials, accompanied by a shortage of primary materials. Serious problems with cleaning-up old landfills and pollution of the groundwater are currently making disposal of waste very difficult in many countries. The protection of soil and water, the limitation of waste production and the re-use of waste materials are key items in policy concepts, generally stated "Sustainable Development". With respect to waste materials, extensive research has been carried out to find a market for these materials, e.g. powder coals fly ash in concrete and incinerator slag in road cons truction. Beneficial use of products derived from waste materials can in fact contribute to sustainable development. However, the market for such waste-derived products mostly involves their re-use as construction materials, implying close contact with the soil. If not properly managed, this may result in pollution of the soil, or even of the groundwater, due to the uncontrolled release of contaminants. In order to predict and control potential contamination, laboratory leaching tests have been developed in several countries, e.g. the USA, Canada, Germany and the Netherlands. The knowledge gained from this research can be used to contol or eliminate possible contamination. A problem is the fact that the various tests being used are not comparable, but harmonization is on its way. The second theme of the conference addresses the state of the art in technical solutions and procedures to use waste materials for the production of construction materials such as concrete, calcium silicate bricks, artificial gravel and other products. Solidification is discussed broadly, as is the treatment and application of MSWI by-products. Various contributions regarding environmental policy and legislation complete the conference. The organizing committee hopes that the conference indeed contributed to the solution of the environmental problems concerning the re-use of waste materials and thus to sustainable development in building practice.
On behalf of ISCOWA Dr. J.M. Goumans
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ix Committees
ISCOWA wishes to thank the members of the committees for their contribution to WASCON '97. Organizing Committee
Scientific Committee
G.J. Senden, ISCOWA, Chairman ir. J. de Castro, ISCOWA R. Fetlaar, Conference Manager L. Haverkort, ISCOWA ir. R.T. Eikelboom, Ministry of Housing, Spatial Planning and the Environment, The Netherlands dr.ir. N. Raemakers, University of Maastricht, The Netherlands
dr.J.J.J.M.Goumans,ISCOWA, Chairman prof.dr. J.Cabrera, University of Leeds, United Kingdom dr. H.A. van der Sloot, ECN, The Netherlands dr. J. Hartl6n, Sweden prof.dr.ing. P.Schieszl, IBAC, Germany Dr. D. Kosson, Rutgers University, USA prof.dr. Shin-ichi Sakai, Kyoto University, Japan
Finally ISCOWA wishes to thank the following organizations who gave financial support to WASCON '97: EC, DGXII, Brussels, Belgium Commission of the European Communities, Directorate General XII, Science, Research and Development, Directorate C : Industrial and Material Technologies, Measurements and Testing Ministry of Housing, Spatial Planning and the Environment.Director ate General for the Environment, The Netherlands GKE/Vliegasunie, De B ilt, The Netherlands Dutch Fly Ash Corporation CUR, Gouda, The Netherlands Center fir Civil Engineering, Research and Codes CROW, Ede, The Netherlands Center for Codes and Research in Civil Engineering Ministry of Transport and Watermanagement, Directorate General for Watermanagement, Delft, The Netherlands JWRF, Japan Waste Research Foundation Kyot 0 University, Kyoto, Japan Novem, Sittard, The Netherlands Netherlands Agency for Energy and the Environment
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Preface .......................................................................................................................................
VII
Overview of MSWI residue recycling by thermal processes ...................................................... 1 Kyoto University Tokio, Kyoto, Japan S. Sakai and M. Hiraoka, Quality improvement of MSW bottom ash by enhanced ageing, washing and combination processes ................................................................................................................. Tauw Milieu bv and Waste Processing Association, The Netherlands J.J. Steketee, R.F. Duzijn and J.G.P. Born
13
Construction materials manufacturing by the technology of melting .................................... 25 Kubuta Corporation, Osaka, Japan S. Abe Producing permeable blocks and pavement bricks from molten slag ................................... 31 Takuma Corporation Ltd., Hyogo, Japan M. Nishigaki Investigation of sintering processes in bottom ash to promote their reuse in civil construction(Part 1) Element balance and leaching ...................................................................................... 41 (Part 2) Long term behaviour .................................................................................................... 51 ABB Corporate Research Ltd, Switzerland and Forschungszentrum Karlsruhe, Germany A. Selinger, V. Schmidt, B. Bergfeldt, H. Seifert, J. Vehlov and F.G. Simon
The acid extraction process ......................................................................................................... 59 H. Kawabata, Kobe Steel Ltd., Hyogo, Japan T. Inoue, Unitika Ltd, Osaka, Japan Pre-treatment of MSWI Fly ash for useful application ............................................................ 67 TNO Waste Technology Division, Apeldoorn, The Netherlands E. Mulder and R.K. Zijlstra Direct melting process for MSW recycling ................................................................................ 73 Nippon Steel Corporation, Tokyo, Japan M. Osada
The ABB dry ash concept: INRECTM ...................................................................................... 79 ABB Corporate Research Ltd, Switzerland A. Selinger and V. Schmidt Municipal Solid Waste Incinerator Bottom Ashes as Granular Base Material in Road Construction ....................................................................................................................... 85 Institut fur Technische Chemie, Institut fur Strassen - und Eisenbahnwesen, Germany G. Pfrang-Stotz, J. Reichelt Test Project Crushed Masonry 50/150 mm in the Ventjagersplaat River Dam .................. 91 Ministry of Transport, Delft, The Netherlands ' H. A. Rijnsburger
xii
Evaluation of treatment of gas cleaning residues from M S W I with chemical agents ........... 95
Environment Preservation Center, Kyoto University, Japan, ECN, The Netherlands S. Mizutani, H.A. van der Sloot, S. Sakai Recycling for road improvement .............................................................................................. 105
OECD, USA Ch.J. Nemmers Quarries reinforcement with stabilised bottom ashes ............................................................ 115
INERTEC and ADEME, France A. Bouchelaghem, M-C. Magnie and V. Mayeux The influence of monolith physical properties on diffusional leaching behaviour of asphaltic pavements constructed with M W S combustion bottom ash .... :............................. 125
University of New Hampshire, Durham, USA T. Taylor Eighmy, D. Crimi, I.E. Whitehead, X. Zhang and D.L. Gress Design and construction of a road pavement using fly ash in hot rolled asphalt ................. 149
University of Leeds, CEMU, Dept. Civil Engineering, England J. Cabrera Engineering properties of the coal ashes stored in the Valdeserrana Lagoon Andorra power plant .................................................................................................................................
167
Polytechnical University of Valencia, Spain P.A. Calderon Garcia, E. Peris Mora and J. Parrila Juste Mine tailings - practical experiences in rifling up harbours ................................................. 175
Public Works, Engineering Division, Rotterdam, The Netherlands J. van Leeuwen and K. Ratsma Upgrading the use of recycled material - UK demonstration project ................................... 185
Building Research Establishment, Watford, England Dr. R.J. Collins Beneficial use of contaminated sediments within the Meuse river system ........................... 193
IWACO and Ministry of Transport, Public Works and Water management, The Netherlands A.L. Hakstege, J.J.M. Heynen and H.P. Versteeg Integration of Testing Protocols for Evaluation of Contaminant Release from Monolithic and Granular Wastes ............................................................................................. 201
gutgers University, Dept. Chem. Biochem. Eng., USA ECN, The Netherlands D.S. Kosson and H.A. van der Sloot Development of a leaching protocol for concrete .................................................................... 217
ECN, IBAC, NMi, Research Institute of the Cement Industry I. Hohberg, G.J. de Groot, A.M.H. van der Veen and W. Wassing Use of a Chelating Agent to Determine the Metal Availability for Leaching From Soils and Wastes .........................................................................................................................
Rutgers University, Dept. Chem. Biochem. Eng., USA A.C. Garrabrants and D.S. Kosson
229
xiii
Leaching Characteristics of Communal and Industrial Sludges ........................................... 247 ECN, The Netherlands P. A. J. P. Cnubben and H. A. van der Sloot Influence of Concrete Technical Parameters on the Leaching behaviour of Mortar and Concrete ...................................................................................................................................... 253 IBAC, Germany I. Hohberg and P. Schiessl By-products management and quality control ........................................................................ 259 Dutch Fly Ash Corporation, The Netherlands J.W. van de Berg and A. Boorsma Maasvlakte Fly-ash processing plant ....................................................................................... 269 Dutch Fly Ash Corporation, The Netherlands J.B.M. Moret and J.W. van den Berg Fly ash as binder in concrete ..................................................................................................... 279 KEMA, The Netherlands L.J.L. Vissers Upgrading and quality improvement of PFA .......................................................................... 289 KEMA, The Netherlands H.A.W. Cornelissen The effect of the Dutch building materials decree on the by-products from coal fired power stations .................................................................................................................... 301 Dutch Electricity Generating Board, The Netherlands M.P. van der Poel Prediction of environmental quality of by-products from coal fired power plants ............. 311 KEMA, The Netherlands R. Meij Short leaching test compared to a column leaching test for internal quality control of coal bottom ash ........................................................................................................................... 327 KEMA, The Netherlands E.E. van der Hoek and F.J.M. Lamers Retention in mortars of trace metals in Portland clinckers ................................................... 339 LAEPSI - INSA Lyon, France I. Serclerat and P. Moszkowicz Study of cement-based mortars containing Spanish ground sewage sludge ash .................. 349 Polytechnical University, Valencia, Spain J. Monzo, J. Pay, M.V. Borrachero, A. Bellver and E. Peris-Mora Fly ash - useful material for preventing concrete corrosion .................................................. 355 IMS, Beograd and Faculty of Technology, Novi Sad, Yugoslavia S. Mileti, M. Ili, J. Ranogajec and M. Djuri
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Fly ash as the basic material for inorganic binders production ............................................ Institute for Materials Testing, Belgrade, Yugoslavia M. Iliac, S. Miletic and R. Djuricic
365
A study of Potential of Utilising Electric Arc Furnace Slag as Filling Material in Concrete ...................................................................................................... 373 Royal Institute of Technology, Sweden C. B~iverman and F. Aran Aran. Properties of portland Cement Mortars Incorporating High Amounts of Oil -Fuel Ashes ............................................................................................................................ Universidad Polytecnica de Valencia, Spain J. Pay, M.V. Borrachero, J. Monz¢, M.J. Blanquer and E. Gonz lez-L¢pez
377
The use of fly ash to improve the chloride resistance of cement mortars ............................. 387 University of Leeds, CEMU, Dept. Civil Engineering, England J. Cabrera and G. Woolley
Low lime binders based on fluidized bed ash .......................................................................... 401 Moravia-Silesian Power Plant and Technical University of BRNO, Czech Republic J. Drottner and J. Havlica Structural performance of reinforced concrete made with sintered ash aggregates ........... 411 University of Leeds and Maunsell & Partners Consulting Engineers, England P.J. Wainwright and P. Robery Investigating waste/binder interactions by neural network analysis ................................... 421 Imperial College of Science, Technology and Medicine, London, England C. D. Hills, J.A. Stegemann and N.R. Buenfeld The use of MSWI bottom ash in hollow construction materials ........................................... 431 Net Brussel, Brussel, Belgium E. Jansegers
Using Chemfronts, a geochemical transport program, to simulate leaching from waste materials ........................................................................................................................... Royal Institute of Technology, Stockholm, Sweden C. B~iverman, L. Moreno and I. Neretnieks
437
Overview of geochemical processes controlling leaching characteristics of MSWI bottom ash ...................................................................................................................... 447 ECN, The Netherlands J. Meima and R.N.J. Comans.
Heavy metal binding mechanisms in cement based waste materials .................................... 459 Swiss Federal Institute of Environmental Science and Technology, Switzerland C. Ludwig, F. Ziegler and C. A. Johnson
ICPMS, Hydride-generation ICP-MS and CZE for the study of solidification/stabilisation of industrial waste containing Arsenic ........................................ 469 University of Leuven, Dept. Chem. Engineering, Belgium C. Vandecasteele, K. van den Broeck and V. Dutr,
xv
Application of computer modelling to predict the leaching behaviour of heavy metals from M S W I fly ash and comparison with a sequential extraction method .............. 481 Katholieke Universiteit Leuven, Belgium P. Van Herck, B. van der Bruggen, G. Vogels and C. Vandecasteele Models for leaching of porous materials ..................................................................................491 Polden, Insavalor and LAEPSI, INSA Lyon, France P. Moskowicz, R. Barna, F. Sanchez, Hae Ryong Bae and J. Mehu A generalized model for the assessment of long-term leaching in combustion residue landfills ...........................................................................................................................501 Royal Institute of Technology, Sweden J.N. Crawford, I. Neretnieks and L. Moreno Influence of the Type of Cement used on the Leaching of Contaminants Leached from Solidified Waste Containing Arsenic ........................................................................................513 Depart. Chem. Engineering, Kath. Universiteit Leuven, Belgie V. Dutr6 and C. Vandecasteele Verification of laboratory-field leaching behaviour of coal fly ash and M S W I bottom ash as a roadbase material ..............................................................................519 INTRON, ECN, Technical University Delft, The Netherlands J.P.G.M. Schreurs, H.A. van der Sloot and Ch. F. Hendriks Leaching of chromium from steel slag in laboratory and field tests -a solubility controlled process ? ....................................................................................................................531 Swedish Geotechnical Institute, Sweden A.-M. F~illman The application of incinerator bottom ash in road construction ...........................................541 Danish Road Institute, Denmark K. A. Phil Acid resistance of different monolithic binders and solidified wastes .................................. 551 Wastewater Technology Center .Corp., Canada J.A. Stegemann and C. Shi Research and Standardization Programme for Determination of Leaching Behaviour of Construction Materials and Wastes in the Netherlands ................................. 563 Van Heijningen Energie en Milieuadvies B.V. and ECN, The Netherlands R.J.J van Heijningen and H.A. van der Sloot Utilisation of flue gas desulphurisation by-products in the cellular concrete technology ....................................................................................................................571 University of Cracow, Dept. Mining and Metallurgy, Poland W. Brylicki and A. Lagosz State of the art of gypsum recovery for a Spanish power plant ............................................. 581 Polytechnical University, Valencia, Spain E. Peris-Mora, J. Monz¢, J. Paya and M.V. Borrachero
xvi
Fine grinding of hard ceramic waste in the rotary-vibration mill ........................................ 591 Technical University of Mining and Metallurgy, Cracow, Poland J. Sidor, A. Mariusz W6jcik and J. Kordek Influence of the Ca content on the Coal Fly Ash Features in Some Innovative Applications ............................................................................................................. 599 Universita di Messina, Universita di Reggio Calabria, Italy P. Catalfamo, S. Di Pasquale, F. Corigliano, L Mavilia Processing and application of phosphoric gypsum ................................................................. Intron, Kemira Agro, Hydro Agri, The Netherlands R. van Selst, L. Penders an W. Bos
603
Valorization of Lead-Zinc Primary Smelter Slags .................................................................. 617 Metaleurop Recherche, France, ECN, The Netherlands, Polden INSA-Lyon, France D. Mandin, H.A. van der Sloot, C. Cervais, R. Barna, J. Mehu The long term acid neutralizing capacity of steel slag ............................................................ 631 Royal Institute of Technology, Stockholm, Sweden J. Yan, C. B~iverman, L. Moreno and I. Neretnieks Reusing water treatment plant sludge as secondary raw material in brick manufacturing .................................................................................................................. 641 TNO, Reststoffenunie Waterleidingbedrijven, Boral Industry bv, The Netherlands L. Feenstra, J.G. ten Wolde and C.M. Eenstroom Assessment of chemical sensibility of Waelz slag .................................................................... 647 Polden, Insavalor, Laepsi, INSA, France, ECN, The Netherlands, Metaleurop Recherche, France Hae-Ryon Bae, R. Barna, J. M,hu, H.A. van der Sloot, P. Moskowicz and C. Desnoyers Immobilisation of heavy metals in contaminated soils by thermal treatment at intermediate temperatures ....................................................................................................... 661 IWACO, SCG, ECN, The Netherlands C. Zevenbergen, A. Honders, A.J. Orbons, W. Viane, R. Swennen R.N.J. Comans and H.J. van Hasselt Investigation strategies for contaminated soils in Finland ..................................................... 673 Geological Survey of Finland, Espoo, Finland H.L. Jarvinen Development of fast testing procedures for determining the leachability of soils contaminated by heavy metals .......................................................................................... 679 lWACO, ECN, SCG, The Netherlands J.J.M. Heynen, R.N.J. Comans, A. Honders, G. Frapporti, J. Keijzer and C. Zevenbergen Electrokinetic transport in natural soil cores .......................................................................... University of Leeds, England D.I. Stewart, L.J. West, S.R. Johnston and A.M. Binley
689
xvii
Re-use of sieve sand from demolition waste ............................................................................699 TNO Waste Technology Division, Apeldoorn, The Netherlands E. Mulder Organic substances in leachates from combustion residues ..................................................705 Link6ping University and Swedish Geotechnical Institute, Sweden I. Pavasars, A-M. F~illman, B. Allard and H. Bor6n
Leaching behaviour of PCDD/Fs and PCBs from some waste materials ............................ 715 Environment Preservation Center, Kyoto University, Japan S. Sakai, S. Urano and H. Takatsuki Environmental quality assurance system for use of crpshed mineral demolition waste to use in earth constructions ...........................................................................................725 VTT Chemical Technology, Finland M. Wahlstr6m, J. Laine-Ylijoki, A. M~ia~itt~inen, T. Luotoj~rvi and L. Kivek~s Environmental certification of bottom ashes from coal fired power plants and of bottom ashes from municipal waste incineration ....................................................................735 KEMA, Dutch Fly Ash Corporation and Waste Processing Association, The Netherlands F.J.M. Lamers, J.W. van den Berg and J.G.P. Born Quality assurance in the laboratory analysis of contaminated soils ..................................... 749 M.J. Carter Associates, England L. Heasman Dutch policy as incentive to environmentally controlled re-use of waste materials ............ 755 Ministry of Housing Spatial Planning and the Environment W.M.A.J. Willart, The Netherlands Evolution of regulations and standards for stabilized hazardous industrial waste management in France ..............................................................................................................757 SPDP Ministerede rEnvironnement, POLDEN-INSA Lyon Developpement, ADEME, France A.-F. Didier, J. M , h u , Valerie Mayeux Test methods and criteria for the assessment of immobilized waste ..................................... 765 INTRON, The Netherlands G.J.L. van der Wegen Inorganic immobilisation of waste materials ...........................................................................769 Delft University of Technology Faculty of Civil Engineering F. Felix, A.L.A. Fraaij and Ch. F. Hendriks Physical properties and long term stability of stabilized contaminated soil ........................ 781 Tampere University of Technology, Finland P. Kuula-V~iis~inen, K. Kumila and H-L. J~irvinen Evaluation of contaminant release mechanisms for soils and solidified / stabilized wastes .......................................................................................................787 Rutgers University, Dept. Chem. Biochem. Eng., USA A.C. Garrabrants
xviii
Response of various solidification systems to acid addition ................................................... 803 Wastewater Technology Centre Burlington Canada J.A. Stegemann, C. Shi and R.J. Caldwell Contaminated soil - cement stabilization in a demonstration project .................................. 815 Public Works, Engineering Division, Rotterdam, The Netherlands J. van Leeuwen, A. Pepels and G. van Ernst Stabilization of a galvanic sludge by means of calcium sulphoaluminate cement ............... 823 Univ. of Napels, Frederico II, Italy R. Cioffi, M. Lavorgna, M. Moarroccoli and L. Santoro Reuse of secondary building materials in road constructions ............................................... 831 Public Works, Environmental Engineering Department, Rotterdam, The Netherlands T. Berendsen MSWI residues in The Netherlands, Putting Policy into Practice ........................................ 841 Service Centre MSWI Residues and Waste Processing Association, The Netherlands J.G.P. Born and R.A.L. Veelenturf The Materials and Energy Potential method for the quantitative distinction between waste valorization and elimination in the cement industry ....................................................851 TNO Institute of Environmental Sciences, The Netherlands J.A. Zeevalkink Using environmental economics in decision making and policy formulation for sustainable construction ............................................................................................................859 University of East Anglia (CSERGE) and University College, England A.L. Craighill and J.C. Powell Application of secondary materials : a success now, a success in the future ........................ 869 Ministry of Transport, Public Works and Watermanagement, The Netherlands J. Th. van der Zwan Sustainable Building and the Use of Raw Materials in the Civil Engineering Sector ......... 883 RWS-DWW, The Netherlands H. Wever.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
Overview of MSWI residue Recycling by Thermal Processes Shin-ichi Sakai a and M a s a k a t s u Hiraoka b Environment Preservation Center, Kyoto University, Kyoto 606-01, Japan b Institute of Systems Engineering Research for Global Environment (ISERGE), Kyoto 600, Japan a
Abstract The melting technology reduces the volume of incinerator residues, bottom ash and fly ash, making the melted slag stable and non-toxic. Moreover, this type of treatment allows the melted slag to be used as a resource again. In Japan, the melting process was developed in the 1980's and has been in practical operation at around 24 municipal solid waste (MSVV) incineration facilities including scheduled ones. By the melting process, PCDDs/PCDFs in residues are decomposed at temperature of approximately 1,400~ in the furnace and heavy metals are concentrated in the fly ash of melting process. The drafting of an 'effective reuse manual' is introduced, aiming at promoting the safe reuse of incinerator residues, by setting reprocessing technologies, reuse standards and their evaluation methods.
1. Introduction The gross amount of municipal solid waste (MSVV) generated annually in Japan is approximately 50 million tons. Approximately 71.2% of this MSW is incinerated, producing approximately 6 million tons of residue which is then landfilled, with leachate control. Recently it has become more and more difficult to secure landfill locations, particularly in urban areas. Consequently, reducing the volume of incinerated MSW ash and looking for ways in which to reuse residues, are urgent targets to be developed. Fly ash produced during MSW incineration is classified as "general wastes requiring special controls." One of the following four treatment methods must be applied to the generated fly ash: 1) melting and solidification, 2) solidification with cement, 3) stabilization using chemical agents or 4) extraction using acid or other solvents. The melting technology reduces the volume of incinerator residues, bottom ash and fly ash, making the melted slag stable and non-toxic. Moreover, this type of treatment allows melted slag to be used as a resource again. The melting operation works by keeping the temperature at approximately 1,400~ in a hightemperature furnace by electricity or by the combustion of fuel. After the residues' physical and chemical state changes, they are cooled in order to solidify it again. In this way, the mass and volume of the residues is greatly reduced, producing a high-density melted product. By melting the residues at such a high temperature and with the change in physical and chemical state, it is possible to produce a melted slag with high stability. However, this technology needs to be improved in certain areas, e.g. reducing the rate of repairing refractory materials, and improving control technology to ensure stable operation of high-temperature melting. The melted-solidified slag can be used as construction material, such as for roads, and is also a useful material in land reclamation, since the bulk of the material is reduced by half to one-third of the original incinerator ash. Another advantage of this method relates to the fact that incinerator fly ash contain hazardous substances such as heavy metals, which can cause problems when they leach out into waterways. By this process of melting and solidification, metallic compounds are stabilized in the 'molecular' structure of the waste product, thereby preventing them from leaching out and dispersing into the surrounding environment. 2.
Melting Technology
2.1 Present Status of Melting Process Development 1,2) In Japan the sewage sludge melting process was developed in the 1980's and has been in practical operation at around 10 full-scale plants. 3).4) In some plants being operated MSW fly ash, along with bottom ash, is melted. The first melting plants used thermal surface melting furnaces,
electric arc-type and coke-bed type melting furnaces. Since then new melting technologies such as plasma melting furnaces, electric resistance melting furnaces and low frequency induction furnaces have been developed and put into practice. At present, 24 municipal solid waste incineration melting-treatment facilities (including scheduled ones) which use the system are shown in Table 1. Some of the systems are still at the trial stage of operation. Each company is, however, making efforts to proceed in their research and development and to bring their technology to the marketplace. Melting technology is almost at a feasible stage. Fusion or vitrification of MSW incinerator residues is not practiced in Europe and North America 5), but detoxification of thermal filter ash has been under development 6)
Table 1
Full-Scale Melting Plants of MSW Incinerator Residues in Japan
Municipalities 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12. 13. 14. 15. 16. 17. 18. 19. 20. 21. 22. 23. 24.
2.2
Numadzu City Kashima Town Eastern Saitama 2 Eastern Saitama 1 Isahaya City Sayama City Tokyo Ota Anan City Handa City Omiya City Matsuyama City Sakado City Shirane Regional Center Tokai City Abiko City Eastern Saitama, New 1 Kinuura regional center Sayama City Mima regional center Hachioji City Tamagawa regional Togane City Kamo regional center Yokohama City
Completion 08/1979 0611981 0311985 0311986 03/1987 0311991 0411991 10/1991 02/1993 0311993 03/1994 07/1994 10/1994 0311995 0311995 09/1995 09/1995 03/1996 03/1997 0311998 0311998 0311998 0311999 0312001
Capacity ton/d 20 6.5/8h 14.4 15 12.3 15 250 4.8 24 75 52 9.6 7/16h 15 10 80 15 15 5/16h 18 25 26 30 60
Unit No. 1 1 2 2 1 1 2 2 1 1 1 1 1 1 1 2 2 1 1 2 2 1 2 1
Manufacturer Kubota Takuma Takuma Takuma Kubota Kubota Daido Takuma Ebara Daido Ebara Takuma Kubota Nippon Steel Hitachi Zosen Daido I.H.I Takuma Kobe S t e e l NKK Daido Takuma Hitachi Zosen NKK
Furnace type Rotating surface Surface melting Surface melting Surface melting Rotating surface Rotating surface Electric arc Surface melting Plasma Electdc arc Plasma Surface melting Rotating surface Coke bed Surface melting Electric Arc Coke bed Surface melting Plasma Electric Joule Electric Arc Surface melting Plasma Electric Joule
Principles of the Melting Systems
At present there are a variety of furnace melting systems that have been developed and are being put into practice. These systems can be divided roughly into two categories: one uses fuels as an energy source and the other uses electricity. The systems can be further classified as follows: (1)
o Surface melting furnaces o Swirling-flow melting furnaces o Coke-bed melting furnaces o Rotary kiln melting furnaces o Internal melting furnaces (2) <Electric melting systems> o Electric-arc melting furnaces o Electric resistance melting furnaces o Plasma melting furnaces o Induction melting furnace (High-frequency, Low-frequency) Some of the fuel-burning melting systems, e.g., coke-bed melting and rotary kiln melting, can not only melt the incineration residues, but can also directly melt MSW. Each of the nine kinds of systems listed has its own particular characteristics. At this stage it is
not possible to say clearly which system is the best. It is most important to use the most appropriate system for the particular conditions of each municipality, or to select a system according to a priority setting. Generally, in the case of a large incinerator with a power generation facility, the electric melting system, which can make use of the recovered electric power, can be selected. In case of a comparatively small incinerator without power a generation facility, the fuel-burning melting system will be selected. 2.3 Surface Melting This is one of the fuel burning-type melting systems. It uses heavy oil, kerosene or gas as the fuel. The structure of the furnace consists of an ash feeding device, main body and burner, as indicated in Fig. 1 7.8,9). One type of design has a pair of furnaces with the two systems facing each other. In another design, the furnace itself consists of an outer body and an inner body, with the outer body rotating. The surface melting furnace works in such a way that continuously-supplied incinerator residues melt from the surface by the heat of the fuel burning. It is then discharged via the outlet port. In this way, the melted slag hardly touches the furnace body directly, and the incinerator residues themselves act as an insulator to protect the furnace body. This type of furnace has a rather large exhaust gas volume and is more suitable for the comparatively small capacity range. INCINERATIONASH ' ~
BURNER
:::::~~.:::.:.:
ASH SUPPLY NELTING SLAG
~
SLAGCONVEYER
(a) Fixedbed type I A~ HOPPER
I NCI HEi~TI ON ASH BURliER R BODY
_._,
CONBUSTIONAlP, UTERBODY
Jl, ~J~ I ~
(b) Rotatingtype Fig. 1
_..,.., EXHAUSTGAS
(c) Fixed bed type I[ Structures of Surface Melting Furnace 7,B)
2.4 Electric Arc Melting The structure of an electric-arc melting furnace is shown in Fig. 2 lo). it consists of the furnace body, lined with refractory lining, an artificial graphite electrode which penetrates to the inside of the furnace, a power supply to feed electricity, an inlet for the entry of the residues, an exhaust and an outlet port. This type of melting furnace works by the application of alternating current to the electrode, which is arranged so as to generate an arc discharged inside the furnace. The heat produced by arcing causes the residues on the metal base to melt. The arcing generates such a high temperature that even residues containing metal can be melted evenly within a short time. The melted slag is removed continuously via the outlet port. It is quenched with water and taken out by conveyor. Any components in the incinerator residues are burnt completely in this type of furnace, and are then removed by the exhaust gas. The atmosphere in this furnace is oxidative. This technology has been applied in the field of steel making.
POWER SOURCE EOUIPI~ENT
I
'
~
.
'~J/
L---,EL.:rING slAGLAYE, Fig. 2
Structure of Electric Arc Melting Furnace lo)
2.5 Plasma Melting This is another type of furnace that uses electricity to melt the incinerator residues. The structure of the furnace is shown in Fig. 3 11, 12). It consists of the surface body, with refractory lining, plasma torches, and a power supply system. There are a variety of plasma torch designs in use, made by the different manufacturers, and each has its specific character. This type of furnace works as follows: first it makes an arc discharge at the electrode inside the plasma torches. This is then passed through the plasma formation gas (air or inert gas) to produce a high-temperature plasma. This plasma is then directed to the incinerator residues by being continuously supplied into the melting furnace. In this furnace there are two types of atmosphere, oxidation and reduction. The melted slag is continuously removed through the outlet port. .. I ~{ClIIERAT I ON ASfl
~ i . 9 ----14 r.,.. 9 PL~SH/,FORgATIO" GAS I I "
. ~ llOeffR~
~1" C(~BUSTIOfi(;8~SER /
i
:-i
I
!
PLAS~ I"OR~
,,~.p~ ~o~,.~ F.J
POffi~R SOURCE EOUI P~IENT
(a) Single torch .type
IdFITI ~ $1.,~G
(b) Twin torch type Fig. 3
Structures of Plasma Melting Furnace 11,12)
3.
Behavior of Heavy Metals and PCDDs/PCDFs in the Melting Process
3.1 Standard Leaching Tests Leaching of heavy metals from the slag was evaluated using the standard leaching tests defined in Notifications No.13 and No.46 of the Environment Agency (JLT13, JLT46) in Table 2. Some points about the standard leaching tests are discussed in the next section. A typical analytical result is shown in Table 3. All of the specified substances in the slag leachate were either nondetectable or below the detection limit, demonstrating that the slag satisfies the environmental standards. In addition, the very low leaching of lead, which has recently become a problem in the effective utilization of recycled materials, is one of the remarkable features of this process.
Table 3
Leaching Test Results for MSW Incinerator Residues and Melted Slag 13)
Sample Item pH Cadmium, Cd Lead, Pb Hexavalent chromium, Cr 6. Arsenic, As Mercury, Hg Cyanogen, CN Selenium, Se Alkylmercury, R-Hg Organophosphorus, Org-P PCB Thiram Simazine Thiobencarb Trichloroethylene Tetrachloroethylene Dichloromethane Carbon tetrachloride 1,2-dichloroethane 1,2-dichloroethylene Cis-1,2-dichloroethylene 1,1,1-trichloroethane 1,1,2-trichloroethane 1,3-dichloropropene Benzene Zinc, Zn Copper, Cu Chloride ion, CI Electric conductivity mS/m
Fluidized bed Stoker furnace furnace fly ash fly ash mg/I mg/kg mg/I mg/kg 12.3 6.8 0.01> 0.1> 33.5 335 28.3 283 10 100 0.04> 0.4> 0.2> 2> 0.01> 0.1> 0.01> 0.1> 0.0005> 0 . 0 0 5 > 0.0005> 0 . 0 0 5 > 0.1> 1> 0.1> 1> 0.01> 0.1> 0.01> 0.1> 0.0005> 0 . 0 0 5 > 0.0005> 0 . 0 0 5 > 0.01> 0.1> 0.01> 0.1> 0.0005> 0 . 0 0 5 > 0.0005> 0 . 0 0 5 > 0.006> 0.06> 0.006> 0.06> 0.003> 0.03> 0.003> 0.03> 0.02> 0.2> 0.02> 0.2> 9 * 0.03> 0.3> 9 9 0.01> 0.1> 9 9 0.02> 0.2> 9 9 0.002> 0.02> 9 9 0.004> 0.04> 9 9 0.002> 0.02> 9 9 0.004> 0.04> 9 9 0.03> 0.3> 9 9 0.006> 0.06> 9 9 0.002> 0.02> 9 9 0.001> 0.01> 5 50 850 8500 1> 10> 1> 10> 9530 95300 9500 95000 3580 3630 -
3.2
Behavior of Heavy Metals
(1)
Behavior and Mass Balance of Metals 1,s)
Minimum limit of Molten slag determination (hydropulping) mg/I mg/I mg/kg 9.3 0.01> 0.1> 0.01 0.01> 0.1> 0.01 0.02> 0.2> 0.02 0.01> 0.1> 0.01 0.0005> 0.005> 0.0005 0.01> 0.1> 0.01 0.01> 0.1> 0.01 0.0005> 0 . 0 0 5 > 0.0005 0.01> 0.1> 0.01 0.0005> 0 . 0 0 5 > 0.0005 0 . 0 0 6 > 0.06> 0.006 0 . 0 0 3 > 0.03> 0.003 0.02> 0.2> 0.02 0.03> 0.3> 0.03 0.01> 0.1> 0.01 0.02> 0.2> 0.02 0 . 0 0 2 > 0.02> 0.002 0 . 0 0 4 > 0.04> 0.004 0.02> 0.2> 0.02(0.002) 0.04> 0.4> 0.04(0.004) 0.03> 0.3> 0.03 0 . 0 0 6 > 0.06> 0.006 0 . 0 0 2 > 0.02> 0.002 0.01> 0.1> 0.01(0.001) 0.1 > 1> 0.1 0.1> 1> 0.1 2 20 1 4.48 -
Inorganic compounds like metallic elements, especially in fly ash, are redistributed after the melting treatment according to the boiling temperature. It is considered that metals with high boiling points like Si, AI and Ca, are converted into slag and substances with low boiling points like Cd and Pb are converted into fly ash or melting furnace exhaust gas. As shown in Table 4, the concentrations of heavy metals with low boiling points like Cd and Pb in ESP ash from melting furnaces are 5 to 10 times higher than those of fly ash. Based on this analysis, the flow rate of flue gas and the quantity of solids formation, the mass balance and the transfer rate of each constituent are shown in Table 5, assuming the input to be 100. SiO2, AI203, CaO, Fe, Mg, Mn, T-P, TiO2, T-Cr and Cu indicate high transfer rates into slag. In contrast, Cd and Pb volatilize into flue gas and are finally concentrated into ESP ash from melting furnace. Na, K, T - S , T-CI, As and
Table 2. Test name
Env~ronmental Agency Notification No. 13 (Note 1)
Environmental Agency Notification No. 46 (Note 1)
Leaching Test Methods
Leachmg vessel
Unspecified
Unspecified
Ministry of Health and Welfare Tentative draft of slag test (Note 2) Airtight bottle (CO2 method) or beaker (pH-static method), (1L polyethylenebottle or 1L glass beaker)
Sample
< 5 mm
< 2 mm
Sample mass
> 50 g
'509
Solvent
Distilled water (Adjusting Distilled water 1) pH 4. CO?saturated water (C02 pH 4 through the way of 20 minto pH 5.8-6.3 by HCl or (Adjustingto pH 5 8-63 3 method) bubbllng of deionized water by NaOH) by HCI) 2) Adding HNOl to deionized water, C02 gas. and keeping the 1st eldon pH 7 and the 2nd one pH 4. (pH-static method) 10. 1 10 1 10.1 ( 5 . 1 ~ 2 ) l0:l 1 1 1 (COZmethod). 1
US ratlo Leaching frequency
Horizontalshaklng (200 t~meslm. ampltude 4-5 cm)
Duration
6 hours
6 hours
Filtration
Ipm GFP
Temperature Ordinary (approx 20°C) Note 1
Note 2 Note 3 Note 4 Note 5
TCLP pH dependency test (Author et at's commonly (EPA Method 1311) used method) 1L beaker at this test Any mater~al compatible with waste, zero-head space conta~ner
Unspecified (1L glass beaker at this test)
1L beaker
10-30 mm
20-50 mm (C50 mm: uncrushed)
c 125 pm
Uncrushed (fly ash, hydropulped slag)
< 9 5 mm
> 50 g
'509
169
50 g at this test
100 g
1) Acetic ac~dbuffer Adding HNOsto deisonized - A t this test, distilled water, and keeping the 1st water and HNO3 or NaOH solution (pH 4 93) Using solvent different in 2) Acetic acid solution elub'on pH 7 and the 2nd (pH 2 88) (Note 4) acidity (alkalinity) or one pH 4. keeping the leachate a certain pH 10.1 20 : 1 1 0 0 : l (50:1x2) 1 1 2 (Note 5)
-
2 (pH-static method) (Note 5) HorizonMl shaking (200 timeslm. amplitude: 4-5 cm)
Agitation
Availability test (NEN 7341)
Ministry of Construction Tentative draft of C02 method
- Horizontal shaking, 200 timeslm.
Stirring and splashing (200 rpm)
Stirrer
Stirrer
Rotating and shak~ng (30 + 2 rpm)
24 hrs (C02 method). 3 hrs x 2 (pH-static method) After 20 mln centr~fugal 0 45 pm MF separation at 3000 rpm. 0.45 pm MF
24 hours
3 hours x 2
23 hours at thls test
18 hours
At this test, 0.45 pm MF
0 6 - 0.8 pm GFF
Ordinary (approx 20°C)
Ordinary (approx. 20°C)
Ordinary
22.3 i 39:
amplitude: 4-5 cm (C02 method) - Stlrrlng and splashing (pH-stabc method)
Ordinary (approx. 20°C)
0.45 pm MF After 20 mincentrrfugal separation at 3000 rpm. 0.45 IIIT MF Ordinary
dichloromethane,carbon tetrachloride, 1.2-di-, 1,l.l-tri- and 1, I ,2-trichlorwthane. 1,3dichloropropeneand benzene (volatile matters), an Erlenmeyer flask For trl-, tetra-, 1.241- and ~i~-1,2-di-chloroethylene. wlth screw cap (500 ml) was used. As for agitation. 4h-stirring by stirrer was implemented. Regarding filtration of elubon, the filtrate was extracted by syringe and filter paper was attached to the syringe. This IS the same method that is applied to the examinahon of volatile substances in sludge (Environmental Agency, nothication no.13) or soil (Environmental Agency, notification no 46) Test In C02 method or in pH-static method is selected. pH targek at this test were 2. 4. 6, 8, 10. 12 and 13. Dlstllled water is added to the sample of 5 g and they are shaken for 5 min, pH IS measured, the solvent of 1) is chosen if pH is over 5 If pH IS below 5, 1.0 N HCI of 3.5 ml IS added and 10 mln-shaklng 1s done at 50°C. If pH 5>, 1) is selected. and ~fpH 5c. 2) IS used. New solvent is added to filter residues and the leaching operation is repeated.
Table 4
Compositions of Solid Materials
Fly ash
Item* Moisture Heating Value Ash Combustible C H N Volatile-S Volatile-CI O TotaI-S TotaI-CI Si Ca AI Na K Mg Fe P TiO2 Mn Cd Pb Zn Cu As TotaI-Cr TotaI-Hg
0.56 58O 89.3 10.7 5.9 0.28 <0.01 <0.1 <0.05 4.5 0.56 10.3 11.3 13.7 5.4 2.1 2.1 1.5 1.5 0.67 1.3 0.06 46 1600 4900 440 13 400 2.7
% kcai/kg % % % % % % % % % % % % % % % % % % % % mg/kg mg/kg mg/kg mg/kg mg/kg mg/kg
Slag
ESP ash from melting furnace
<0.1 <0.1 <0.01
1.80 <0.1 <0.01
0.33 0.32 14.2 19.2 7.5 0.7 0.3 2.1 2.1 0.96 1.7 0.09 3.0 110 1200 13O0 5.0 1000 0.13
1.8 40.5 <0.01 0.20 <0.01 17.8 13.8 0.01 0.38 <0.01 <0.01 0.02 420 5000 5000 2500 64 42 1.6
* All data are on dry-solid basis except for moisture.
Table 5 Item
SiO2 AI203 CaO Fe Mg Mn Na K T-P T-S T-CI TiO2 T-Cr Cu Cd Pb As Zn T-Hg
T r a n s f e r Rate of Each Metal C o m p o n e n t
Inputs Fly ash from MSWl 100 100 100 100 100 100 87 100 100 100 100 100 100 100 100 100 100 100 100
Outputs NaOH
13
Slag 100 110 110 110 110 120 21 11 110 46 2.5 110 200 230 5.2 5.4 30 19 3.8
Dust in ducts 0.030 0.063 0.33 0.50 0.091 0.000 29 17 0.000 2O 10 0.054 0.30 17 20 7.3 9.7 1.3
ESP ash from MF 0.000 0.000 1.11 1.9 0.050 2.5 55 49 0.000 24 29 0.000 0.78 42 68 7O 37 38 4.4
Flue gas
0.2O 0.17 0.51 8.2 0.00 0.16 0.24 0.24 0.00 0.12 22
Total 100 110 110 110 110 120 110 77 110 91 5O 110 2OO 29O 93 93 74 67 31
Zn take the positions in the intermediate, which convert some parts into slag and some parts into dust, depending on their chemical forms. For the slag utilization and the resource recovery from the melting furnace fly ash, heavy metals have to be highly concentrated into melting furnace fly ash. (2) Recycling of Melting Furnace Fly Ash as Resources Melting furnace fly ash from melting process is produced about 4% of the total input in case of melted bottom ash only, and 6% to 10% in case of melted bottom ash with fly ash. Melting furnace fly ash contains a considerable amount of Pb and Zn. It is therefore necessary to take care of heavy metal stabilization and control in landfill. In the future the recycling of melting furnace fly ash should be chosen instead of its disposal. We are trying to use it as a non-ferrous smelting material. In any smelter all elements except target metal are treated as impurities, so it is not allowed to apply heavy duty to smelter for removing impurities. Halogens such as chlorine must be restricted to very low levels to prevent equipment corrosion problems. Therefore, resource recovering process should concentrate Pb and Zn separately and reduce the CI content in the concentrated cake to the upper limit. Following is a kind of resource recovery process, that is the combination process of fly ash melting with bottom ash by furnace, and wet treatment system to recover Pb and Zn. The 1st stage of the melting process adopts promotive evaporation in the furnace, chloridizing and vaporizing heavy metals by chloride in fly ash, cooling the exhaust gas, and collecting the condensate dust by using a bag filter. The 2rid stage of wet treatment is shown in Fig. 410). Dust is mixed with acid solution, transferring Zn, Cu into a liquid state by separating non soluble Pb compounds by filter. The Zn, Cu are changed into hydroxide by neutralization, and then sulfured. After that they are divided to solid and liquid forms. Thus, heavy metals scarcely converted into a liquid state. By the test plant operation, the following results were obtained, as shown in Table 6. About 35% Pb and 31% Zn contained cake are estimated to be usable as raw materials by mixing it with natural concentrate (Pb: 50~70%, Zn: 45-~60%), and waste water, satisfying the regulation limits.
3.3 Behavior of PCDDs/PCDFs Residues from MSW incinerator contains somewhat dioxins (PCDDs/PCDFs). Table 7 shows analysis results on the dioxin content in melted fly ash with surface melting furnace, containing a rather high concentration of dioxins. Dioxins of 3500 ng/g (10 ng TEQ/g) were contained in fly ash from MSW incineration. However, only 0.063 ng/g (0.00 ng TEQ/g) is detected in slag, 0.74 ng/g (0.00 ng TEQ/g) in fly ash from melting furnace and 23 ng/N (0.25 ng TEQ/Nm 3) in flue gas. The decomposition rate is 99.99% in PCDDs/PCDFs and 99.98% in TEQ 1,9) The level of dioxins contained in incinerator residues can be reduced by decomposition due to the heat of the melting process, at the high temperature of approximately 1,400%. It is suggested, therefore, that waste produced from the melting process will be environmentally compatible. 4. Reutilization of Melted Slag The technology to turn slag produced from the melting of MSW residues into a reusable resource has been investigated. Some ideas are, to use the slag as a fill for road surfaces, as a component in asphalt mixtures, in concrete structures and in secondary products (e.g. interlocking blocks, tiles, and bricks ). There are already established standards in place for these types of natural materials such as crushed stones. In order to develop similar standards for such materials that contain slag, some research has been done. The results have shown that slag can be used in different ways, depending on the particular application and conditions of use, as shown in Table 8 14) This does not mean that the use of slag in such existing materials will immediately become widespread. At present, using melted slag as a resource is still at the stage where possibilities are being investigated and the feasibility of operations is being studied. From now on research and development is must concentrate on producing slag that will be acceptable to user's needs. For this reason some standardization of slag products is needed urgently. To evaluate the possibility of using slag as a resource, economic considerations such as marketing and pricing need to be studied, and the quality of slag-containing products needs to be brought in
condensate dust
NaOH
neutralization I
Naris
I sulfurationl
H20
extract,onI $
HCI
I filtr~ ti~
Table 6
zinc~cake
Zn 6.86 12.28
Zn cake [%] Waste water [mg/I]
2.61 0.03
31.09 0.19
Regulation
0.1
[mg/I]
Table 7
Fly ash
P5CDDs H6CDDs H7CDDs O8CDD PCDDs T4CDFs P5CDFs H6CDFs H7CDFs O8CDF PCDFs PCDDs+PCDFs TEQ
5
Cu 1.34
Fe 1.43
Si 0.72
AI 0.30
Cd 0.073
As 0.007
0.19 6.93
16.43 0.95
2.60 2.72
1.44 1.02
0.0001 0.383
0.18
2.85
4.93
0.020 0.027 0.009
-
-
0.05 3
10
<0.005 0.1
o.1
Hg 2.0 ppm
Cr 0.005
1.6 ppm 0.038 9.5 ppm 0.013 <0.0005 <0.01
o.oo5
0.05 (cr e§
C o n c e n t r a t i o n s of PCDDsiPCDFs and TEQ in Solid and Flue Gas Samples
Item
T4CDDs
waste water
Chemical Composition of Pb, Zn Cake and Waste W a t e r 10) Pb
[%] [%]
drain I
Pb, Zn Separation and Recovering Process lo)
3.14 34.71
Dust Pb cake
1
I
t Fig. 4
I pHcontroll
I filtrationl
I
lead ~cake
$
Solids, ng/g Slag
from MSWI 1600 910 640 110 88 3300 34 43 49 16 7.4 150 3500 10
Flue gases, ng/Nm 3 @O2=12% ESP ash
Inlet of air
Inlet of gas
Inlet of
Outlet of
<0.004
from MF 0.032
preheater (Sl) 0.45
cooler (S2) 0.45
ESP (S3) 1.2
ESP ($4) 0.23
0.010 0.008 0.004 0.020 0.042 <0.004 <0.004
0.010 0.013 0.021 0.25 0.33 <0.004 0.007
0.41 0.91 0.52 7.0 9.3 1.8
1.3 3.8 9.6 15 31 8.2
0.31 0.97 1.8 3.5 6.8 1.3
<0.004 0.004 0.017
0.022 0.050 0.3
0.021 0.063
0.41 0.74
1.9 2.5 0.91 4.1 11 20
7.0 8.5 10 14 48 79
0.0
0.00
0.63 1.7 1.7 3.5 8.O 3.5 5.1 4.7 3.3 5.5 22 30 0.40
0.13
0.54
2.0 3.3 3.4 5.9 16 23 0.25
]0 line with that of similar existing materials. Table 8 Property
Properties of a Melted Slag and Commercial Roadbed Materials, and Various Specified Values 14,15) Specified values for road aggregate Crushed Guidelinefor asphalt stone for paving road construction Lower Upper Surface Air Water Crusher Reclaimed Class 1 subsub- layer/ cooled cooled run (C) crushed grade grade base stone layer l a y e r layer (RC-40) >2.45 2.65 2.65 2.66 2.45 Slag
Specific dry gravity ~urface 0.12 0.75 Moisture % absorption rate 30-35 50-60 Abrasion loss % % <1.0 <1.0 Stability % 2.8 3.6 Optimum moisture content 98-100 15-20 Compensated CBR 5.
Typical commercial products
1.34
4.43
20.7 10.9 4.5
23.0
123.0
98.8
8.5
Crushed stone for concrete >2.5 <3.0
<3.0 <35
Specified values for concrete aggregate
<50 <20
<50 <20
>20
>80
<30 <12
<40 <12
'Effective Reuse Manual' of MSW Melted Slag (draft)
As I mentioned above, high-temperature treatment technologies for incinerator residues (melting and solidification methods) have been developed, and it is becoming possible to reuse the residues effectively by melting. It is important to reuse them as road construction materials or concrete aggregate and so prevent environmental disruption caused by quarrying. Some basic ideas for the effective utilization of incinerator residue products are as follows: (1) Incinerator residues should be reduced through waste discharge control and recycling, (2) Incinerator residue products should be positively promoted for reuse, in order to reduce the final landfill volume, and (3) Incinerator residue products have to be stable in order not to cause environmental pollution, such as soil contamination and ground water pollution. With this in mind, the Ministry of Health and Welfare is discussing the drafting of an 'effective reuse manual', aiming at promoting the safe reuse of incinerator residues, by setting reprocessing technologies, reuse standards and their evaluation methods. I would like to introduce an outline of the manual here. It deals with the following: (1) Slag produced by melting MSW incinerator residues under the temperature of 1200 ~ 1400~ or higher, (2) Slag produced by gasification/melting of MSW under the temperature of 1200 ~ 1400~ or higher, and (3) Solid substances produced by sintering MSW incinerator residues under the temperature of 1000 ~ 1300~ The Ministry of Construction is also considering applying standards for the reuse of melted slag from sewage sludge, and the Environment Agency is examining standards for various recycling flows such as iron slag from steel production or coal ash. The effective reuse standards have to incorporate physical standards such as strength and durability and quality standards focusing on environmental impact. Within the quality standards, hazardous substances exposure routes to human body should be considered. The routes are roughly classified into three, or food, air, and drinking water. Among them, the most important exposure route is drinking water, that is, hazardous substances leaching from residue products into ground water through rain. Therefore, leaching tests were adopted for quality confirmation and drinking water standards were applied for slag standards. At present, 29 items of drinking water
1] standards are concerned with human health. 23 items including cyanogen and other organic compounds are considered to decompose because melted slag and sintered products are manufactured under the temperature of a thousand and some hundreds degrees centigrade (refer to Table 2). Thus, they were left out of the effective reuse standards. For the following six heavy metals, standard values were set; Cadmium: 0.01 mg/I, Lead: 0.05 mg/I, Hexavalent chromium: 0.05 mg/I, Arsenic: 0.01 mg/I, Total mercury: 0.0005 mg/I and Selenium: 0.01 mg/l. Non-processed melted slag is used as filling and road-bed materials, concrete and asphalt aggregate and cement materials. Processed-melted slag, secondary products, are used for concrete, asphalt and interlocking bricks. As they are used in many ways it is difficult to identify the places where they are used. Therefore the standards were established to be applied in all cases regardless of place or purpose. Looking at these utilization, it is considered that melted slag products would be more useful than shapeless melted slag. Products like bricks seem to leach smaller amounts of hazardous substances than shapeless melted slag because of the decrease of the rate from surface area, the less permeability and the encapsulation of hazardous substances by solidification. Present leaching tests, however, (JTL13 and JTL46) can be applied only to crushed samples (less than 5 mm on JLT13 and less than 2 mm on JLT46). These tests can not evaluate the effects of solidification, stabilization and formation. Considering the difference in form between shapeless materials and solid products, adequate leaching test methods should be applied for the evaluation. Test methods under discussion are shown in Table 2, and the basic ideas are drawn in Fig. 5. The test samples were regulated 10---30 mm in size because a certain strength and high reproducibility are required. Taking into account changes of chemical properties, including acidification by acid rain or carbon dioxide, acid solvents have to be used. Reprocessed melted slag is now being used close to our living environment, as opposed to landfill sites. Evaluation of the environmental impact should be carded out as close as possible to our living environment and prior to its use. The further development is expected. Incinerator residue products (materials and products) I
q~
Shapeless products (below 10 mm in size)
Shaped products (over 10 mm in size) Compressive strength (below 5 N/mm 2)
Compressive strength (over 5 N/mm2) m
I Not clearing the tests
(EndurancWetests) , g
r
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Notification No. 46 test of Environment Agency
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Clearing the tests
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Leaching tests for stronger products Size of test samples: 10-30 mm, Acid solvent (HNO3 or CO2), Shaking for 24 hours
Frame Work of Leaching Tests by the F o r m of MSWI Residue Products (draft)
Conclusion
The following points should be considered if MSW incinerator residue melting facilities are to share a major role as terminal facilities in municipal waste management and recycling in general. Firstly, the appropriate technology should be established so that slag should be considered as a recyclable resource. It is necessary to make the quality of slag-containing products equal to that of equivalent existing materials, in order to promote their use. Secondly, economic aspects need to be carefully considered. Using new processes such as melting will add considerable costs to the existing MSW treatment system. However, even
]2 without further treatment after incineration, fly ash still has to be treated, since it has been designated to be 'specially controlled waste'. So currently, the final cost of treatment is higher than it appears. Considering this point, efforts are needed to decrease costs by improving the melting technology to make it more suitable. However, we should always be prepared to pay cost of recycling. From now on, the treatment of incinerator residues should be thought of in terms of the recycling of resources aimed at promoting a more recycle-conscious society. Incinerator residues contain hazardous substances such as heavy metals and dioxins. To use the material efficiently, safety aspects must be well thought out, so as not to leave a negative heritage to future generations. Considering all these points above, we need to proceed further in development of this technology.
References 1. Sakai, S., Tejima, H. & Kimura, T.: Cycle Technologies and Strategies on MSW Incineration 2.
3. 4. 5. 6. 7. 8. 9. 10. 11. 12. 13. 14. 15.
Residue, Air & Waste Management Association, VIP-53, pp.737-749 (1996) Hiraoka, M., Sakai, S: The Properties of Fly Ash from Municipal Waste Incineration and its Future Treatment Technologies, Waste Management Research of the Japan Society of Waste Management Experts, 511], pp.3-17 (1994) Takeda, N., Niraoka, M., Sakai, S., Kitani, K. and Tsunemi, T.: Water Science Technology, 1989, 21, pp.925-935 Sakai, S., Hiraoka, M., Takeda, N. and Tsunemi, T.: Water Science Technology, 1990, 22, pp.392-338 International Ash Working Group: An International perspective on Characterization and Management of residues from Municipal Solid Waste Incineration (1995) Hirth, M., Wieckert, C.H., Jochum, J. and Jodeit, H.: A Thermal Process for the Detoxification of Filter Ash from Waste Incinerators, Recycling International, pp. 1561-1566, Vol.2 (1989) Nishigaki, M ; Reflecting Surface-Melt Furnace and Utilization of the Slag, Seminar on Cycle and Stabilization Technologies of MSW Incineration Residue, pp.55-65 (1996) Ishida,M ; The Demonstration Test of Burner Type Ash Melting System, The Hitachi Zosen Technical Review, Vo1.56, No.3, pp.56-61 (1995) Abe, S ; Ash Melting Treatment by Rotating Type Surface Melting Furnace, Seminar on Cycle and Stabilization Technologies of MSW Incineration Residue, pp.41-54 (1996) Kinto, K ; Ash Melting System and Re-use of Products by Arc Processing, Seminar on Cycle and Stabilization Technologies of MSW Incineration Residue, pp.31-40 (1996) Jimbo, H ; Plasma Melting and Useful Application of Molten Slag, Seminar on Cycle and Stabilization Technologies of MSW Incineration Residue, pp.22-30 (1996) Ishida,M ; Twin Torch Type Plasma Arc Ash Melting of Municipal Solid Waste Incinerator, The Hitachi Zosen Technical Review, Vol.56, No.2, pp.57-62(1995) Japan Waste Research Foundation: Treatment and Utilization of MSW Incinerator Residues (1996), in Japanese Kouda, M ; Experimental Pavement Using Household Waste Slag Sand, Seminar on Cycle and Stabilization Technologies of MSW Incineration Residue, pp.92-106 (1996) Yamagishi, K.; Research and Development on the NKK Electric-Resistance Furnace for Melting Ash from Municipal Waste Incineration and on Effective Use of the Slag, Seminar on Cycle and Stabilization Technologies of MSW Incineration Residue, pp.76-91 (1996)
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
13
QUALITY IMPROVEMENT OF MSWl BOTTOM ASH BY ENHANCED AGING, WASHING AND COMBINATION PROCESSES J.J. Steketee a, R.F. Duzijn a & J.G.P. Born b aTauw Milieu by, PO Box 133, 7400 AC Deventer, the Netherlands bWaste Processing Association, PO Box 19300, 3501 DH Utrecht, the Netherlands
Abstract Natural aging of MSWl bottom ash is known to have the effect that the leaching of most constituents decreases, especially with respect to heavy metals. Tauw Milieu has developed a process to speed up the aging process by controlling storage conditions and leading a CO 2 containing gas mixture through the material. Aging process conditions were optimized and the process was then tested on bottom ash samples from Dutch incinerators with a volume of up to 500 kg. Results showed that within several weeks a quality improvement is achieved that corresponds to one or more years of natural aging. Leaching of heavy metals, DOC and most macro-elements decreased by max. 90%. Leaching of salts did not decrease. This means that generally speaking, the requirements of the Dutch Building Materials Decree are met for the material after this w a y of processing, although this may not always be true for bromide. Washing processes are especially suitable for the removal of soluble salts like chloride- and bromide salts. The removal of sulphate is more difficult owing to its poor solubility. Research into washing processes has focused on minimization of the amount of water and optimization of conditions such as pH, percolation rate etc. Although a quality improvement was achieved, relatively strongly leaching bottom ash did not meet the requirements of the Building Materials Decree. With a combination of washing and aging, the leaching of salts as well as heavy metals can be reduced to low levels. Based on a preliminary design of an installation for the treatment of 100,000 tonnes of bottom ash per year, the exploitation costs of aging, washing and combination processes were calculated.
1
INTRODUCTION
In the Netherlands, the environmental quality of building materials to be applied on or in soil is mainly assessed on the basis of its leachability. The limit values for the immission of inorganic contaminants into the soil are listed in the Building Materials Decree [1], in mg/m 2 for a period of 100 years. These limit values are based on the principle of 'marginal burdening of the soil': the quality of the soil or groundwater below the used material may only be affected marginally by the material's leaching. This has been translated into a contaminant concentration increase of 1% of the target value over a period of 100 years, in a 1-metre-thick soil layer directly below the used material. This leaching standard results in a classification of materials into t w o categories: 1) materials that hardly leach can be freely used; no special demands are set on their application other than that it must be possible to remove the materials, and that therefore they should not be mixed with the soil (category 1 ). 2) materials that leach more than that may be applied on the condition that a watertight top lining be applied (infiltration _< 6 mm/year) and that the material be used only at a height of > 0.5 m above the mean highest groundwater level (category 2).
14 Materials that do not meet the immission limit values at an infiltration rate of _< 6 ram/year are not allowed to be applied. There are a few exceptions to this. One of these is MSWl bottom ash, for which however extra stringent demands are imposed on the top lining of the construction. The immission requirements laid down in the Building Materials Decree can be translated into emission values as measured in column tests, with the help of a calculation model. Table 1 presents the emission requirements for some components. Table 1
Maximal leaching of building materials, as measured by means of a column test, at an application height of 0.7 m. Emissions in mg/kg d.w. at a cumulative Liquid/Solid value of 10.
Parameter
Category 1
Category 2
Antimony" Arsenic Cadmium Chromium Copper* Lead Molybdenum" Nickel Zinc
0.045 0.88 0.032 1.3 0.72 1.9 0.28 1.1 3.8
0.43 7.0 0.066 12 3.5 8.7 0.91 3.7 14
Bromide" Chloride Sulphate
2.9 600 1,140
4.1 8,800 22,000
* potentially critical parameters for MSWl bottom ash Although the application of MSWI bottom ash can continue for the time being (against higher costs, as stricter demands are set towards the top sealing), the waste processing branch is striving for a quality improvement of up to category 2 level. In the long run, category 1 should be attainable. The quality improvement of MSWl bottom ash can be attained by input measures (such as the MSWl refusal to accept certain waste materials for processing), by process measures and by an end-of-pipe treatment of the bottom ash. The article at hand concentrates on quality improvements achieved by the after-treatment of slags, a process in which the leaching behaviour is of major importance. In order to improve its leaching behaviour, MSWl bottom ash can be subjected to several processes, that amount to either the fixation or the removal of contaminants. Contaminants can be fixed by e.g. a melting process, adding agents, or aging the material. Tauw Milieu has developed an enhanced aging process, in which the leachate quality rapidly improves within a short period of time [2]. This method is particularly successful when applied to heavy metals. On the other hand, the quality of the material can be improved by washing processes; these are particularly effective for easily soluble salts. At the request of the Waste Processing Association, Tauw Milieu has optimized the aging process and scaled it up to 500 kg. At the same time, research was carried out into optimizing the washing process, and into combined washing and enhanced aging processes. The research objectives were the process-technical upgrading of the above processes and finding the most cost-efficient way of making MSWl bottom ash meet the category 2 demands of the Building Materials Decree.
15
2
THEORETICAL BACKGROUND
It is a generally known fact that when MSWl bottom ash is stored outdoors, the quality of its leachate improves considerably after some months to some years [24]. This improvement is the result of several different reactions, namely hydration, carbonation, the (microbiological) oxidation of organic substances, the oxidation of metals, various precipitation reactions (including the formation of proto-clays) and the weathering of glass phases. Most of these reactions help to improve the leaching characteristics. Through carbonation, a pH of about 8.5 is obtained, at which the solubility of many metals in the form of cations (copper, lead, zinc, etc.) is minimal. The oxidation of organic substances results in decreasing emissions of COD and metals that are complexed by organic substances. The oxidation of metallic iron and aluminium, various precipitation reactions and the formation of clay minerals improve the MSWl sorption behaviour. In addition, metals can be incorporated within the newly-formed matrices. One unfavourable reaction during aging is the mobilization of sulphate, which is believed to be the result of the slow hydration of anhydrite [5]. The aging process can be enhanced by bringing the material into contact with air containing a raised CO 2 concentration (enhancement of carbonation), raising the temperature, optimizing the moisture content, and occulation with microorganisms (which play an important role in the degradation of organic substances). In the washing processes, readily and poorly soluble components can be distinguished. Readily soluble components, e.g. chloride and bromide, can be easily removed with little water. Poorly soluble components, e.g. heavy metals, can only be partially removed by washing. Even if only the leachable fractions of these elements are to be removed, more washing water will be needed than for such a substance as chloride. As these elements are usually subsequently still supplied from the matrix, a greater quantity of contaminants should be removed in the washing process than one would expect based on the leaching test. As the treatment of the washing water is one of the major cost items, the volumes of washing water should be kept to a minimum; in other words, the solubility of the contaminants should be maximal. In order to obtain maximal solubility there are various treatment methods, e.g. pH correction, the addition of complexers, or raising the temperature. Earlier research [6] has shown that the use of such complexers as EDTA is less effective than pH correction. One of the reasons may be that complexers partly adsorb to the bottom ash and are released again during the next leaching test. However, pH correction has the drawback that salts remain behind in the material (chloride, if hydrochloric acid is used for pH correction). These remainders can be removed by again washing with water.
3
SETUP OF THE I N V E S T I G A T I O N
As t w o of the most critical elements are copper and molybdenum, the tests were predominantly carried out on the bottom ash of t w o Dutch MSWl plants with the highest copper and molybdenum leaching concentrations. At a later stage, a third MSWl, the bottom ash of which showed more average leaching results, was involved in the investigation. In total, batches from 12 different production periods were investigated. The aging process was optimized by varying the process parameters at a small scale (1 - 2 kg). The range of the investigated conditions is presented in Table 2. In total, 26 experiments were carried out. Next, the kinetics of the process were investigated at a larger scale (50 - 100 kg) by regularly taking samples during the treatment. These tests were carried out at optimal conditions with only the CO2
]6 concentrations being varied (between 4 and 40%). Finally, the established optimal process conditions were verified at a scale of 500 kg, paying attention to such technical aspects as pressure drop and possible quality differences as a function of the layer thickness. Table 2
Ranges of investigated parameters for optimization of enhanced aging
Parameter
range
Temperature (~ Moisture content (% of d.w.) Gas flow rate (dm3/kg, h) Microbiology
25 - 80 1 5 - 30 0.5-5 inoculation, addition of nutrients 4 - 40 9 - 20 3 - 56
CO2-content (vol.percent.) O2-content (vol.percent.) Residence time (days)
Two kinds of washing processes can be distinguished: processes intended to remove salts (metals are fixated through aging, therefore it is in fact a combined process) and processes aimed at removing both salts and heavy metals. For both process types, the process conditions were optimized by varying the different parameters at a scale of 10 kg. Table 3 presents the parameter variation ranges. Table 3
Ranges of investigated conditions for optimization of washing processes
Parameter
Washing processes for salts L/S-value (I/kg) Percolation-rate (I/kg,day) Intensity of water contact Temperature (~
Washing processes for metals and salts pH L/S-value (I/kg)
range 0.5-5 0.2-2 percolation - shaking 20 - 60
3-4 1 - 20
Finally, combined processes were tested, in which the sequential order of the processes, the L/S-ratios and the scale sizes were varied. At a small scale (1-2 kg) and L/S values of 1 - 5, three process sequences were tested: washing-aging, aging-washing, and aging-washing-aging. At a scale of 100 kg and L/S value of 1, t w o process sequences were tested, namely washing-aging and aging-washing. Based on the test results, global designs were drawn up for practice installations for some of the tested process configurations, and cost estimates were made of both the investment and exploitation costs.
17
4
WORKING ROUTINE
The bottom ash was sampled by employees of the concerned MSWls, in conformity with the common quality assurance procedures. The samples were mixed samples taken over a 2-week production period. All aging tests were performed in the form of columns through which a gas mixture was blown upwards. The gas mixture was composed of air, CO 2 and N2, at the desired ratios. The mixture was furthermore saturated with water vapour. The temperature in the columns was created by supplying heat from an external source. Samples were taken during (intermediate samples) or at the end of the tests by emptying the columns and composing mixed samples. The large-scale test material was also sampled per layer of ca. 0.5 m; the total bed height was 1.5m. For the small-scale tests (up to 2 kg), crushed material (< 4 mm) was used; the other tests were performed with material of the usual size. Any released condensate or drainage water was recirculated. The leached concentrations were determined at the beginning and at the end of the experiments, by means of shaking tests (L/S 20) and column tests (L/S 10), in accordance with NEN 7343 and NEN 7349, respectively. Intermediated leaching tests only involved shaking tests. The leachate was analyzed for As, Cd, Cu, Cr, Mo, Ni, Pb, Sb, Zn, Ca, CI, SO4, COD, pH and E.C. The treatment performance was expressed in terms of the decrease (in %) of the leached quantities in relation to the initial values. The washing tests were performed by leading tap water upwards through the columns filled with bottom ash. The columns were flushed until the desired L/S ratio was obtained. Afterwards, leakage water was added to the washing water. pH-control experiments were conducted as stirring tests after the addition of concentrated hydrochloric acid. In combination tests, columns were both vented and flushed in certain sequential orders. The analysis package was the same as that for the aging tests. As the composition of the material remained more or less the same, it was extensively analyzed before the test, but after the test only dry weight and organic substance were determined. Furthermore, civil-technical parameters were determined before and after the large-scale tests, namely grain size distribution, crushing factor, and iron- and unburned material content.
5
RESULTS
5.1
Enhanced aging
The optimal process conditions for enhanced aging are: a temperature of 60~ a moisture content of 15-20%, gas composition of 4-8% CO2 and 18-20% 02, a gas flow rate of 0.5 m3/tonne/hour, and a maximum residence time of 4 weeks. The required residence time depends on the initial quality of the bottom ash. Where category 2 values are only slightly exceeded (copper up to 200%, molybdenum up to 50%), a residence time of a week will suffice. Stimulation of the microflora is not required.
]8
Table 4
Leached emissions (in mg/kg d.w.) for MSWl 1 before and after enhanced aging (four weeks), and performance (% decrease leaching). Shaking test L/S 20, averages of 4 tests at a scale of 100 - 500 kg. fresh material emission
pH
.
SO4
.
.
3170
Sb (2)
0.23
Cu
6.1
,,,
.
class
performance (%)
8.1
-
913 1430. .
emission
class
10.7
E.C. (1) CI
aged material
-
905
Cat. 2 .
1430
Cat. 2
0
6790
Cat. 2
-114
Cat. 2
.
.
Cat. 2
0.24
Cat. 2
-4
>Cat. 2
0.63
Cat 91
90
Zn
0.82
Cat. 1
<0 9
Cat 91
>84
Pb
0.78
Cat. 1
< 0.06
Cat. 1
> 92
0.72
Cat. 2
64
2.00
Mo
>Cat. 2
(1) electrical conductivity, in pS/cm (2) Sb is column test emission (L/S 10)
Table 4 summarizes the results of the large-scale tests with material from MSWl 1. These s h o w that good results ( 8 0 - 9 0 % decrease in leaching) can be obtained for cationic metals and COD. After treatment, the material meets the category 1 requirements for these parameters. The performance w i t h regard to molybdenum is less effective (over 60%), whilst for antimony and chloride little or no improvement is measured, and sulphate leaching increases, as expected. In spite of this, the material as a whole meets the requirements for classification as a category 2 material 9 The results of tests on bottom ash of other investigated MSWls are largely the same, w i t h higher performances for antimony, (58% for MSWl 2, and 4 2 % for MSWl 3). A general conclusion is that although metals and metalloids can be fixed to a sufficient degree, enhanced aging is not effective for salts. If bromide is critical, or the category 1 level should be obtained, aging will have to be combined with a washing process.
5.2
Washing
The tests showed that chloride can be removed up to category 1 level, using little washing w a t e r (L/S = 0.5) at short retention times (0.5 day, shorter retention times were not tried). As bromide has similar chemical properties, these conditions probably also apply to this element. Sulphate is much harder to remove 9The category 1 value was not even obtained after washing at L/S 5 and a residence time of 5 days. The removal performance improves only slightly ( > 10%) when raising the temperature from 20 ~ to 60 ~149Lowering the percolation rate w i t h a factor 10 is more effective, w i t h 2 8 % more sulphate being removed (refer to Table 5). However, also at these conditions - that are financially not very attractive - the category 1 value is not attained.
]9
Table 5
Effect of percolation rates during washing of M S W l bottom ash. Washing up to L/S = 5, T = 20 ~
Parameter
percolationrate 0.2 I/kg,d (T 25 d)
percolation rate 2 I/kg,d (T = 2.5 d)
Load (mg/kg dm)
Performance (%) (1)
Load (mg/kg dm)
Performance (%) (1)
Removal Chloride Sulphate
1390 2390
102 68
1000 1235
74 35
Leaching Chloride Sulphate
100 (2) 1020 (2)
93 71
260 (2) 1660 (2)
81 53
=
(1) removal (in per cent) with the washing water, or decrease (in per cent) in leachate of washed material in relation to the initial leaching values, as measured during shake test L/S 20. (2) leaching of washed material, measured during shake test L/S 20.
Table 5 shows that kinetic aspects play a role in the washing process. This may both be related to transport (diffusion from the pores of larger ash particles) and to mineral conversions. It is remarkable that even for the readily soluble component chloride, percolation rate/residence time differences have an effect. This effect is however significantly more noticeable for sulphate: at low percolation rates the removal performance is 3 0 % higher. The performance on the basis of leaching is only 18% higher; in absolute quantities, an additional 1,1 15 mg of sulphate has been removed, but the leaching has decreased by 640 mg/kg dw. This illustrates that the relation between removal and residual leaching is not linear - ' n e w ' sulphate becomes available through subsequent supply. Tests during which the material was shaken while being washed did not result in higher removal efficiencies in comparison to the percolation washing method. Conditions at which metals always meet the category 2 standards were not found. Lowering the pH to values of < 4, or raising the L/S ratio to 20 had an unsatisfactory and sometimes even negative influence on the residual leached quantities. At L/S 1 and pH 4 the best results are obtained. It is however possible that in the case of less strongly leaching bottom ash, the category 2 level can be attained by washing. In summary, it can be concluded that chloride and bromide can easily be removed (little washing water, short residence time), but that there are no easy methods to remove sulphate to category 1 level. The bottleneck probably is the mineral form in which the sulphate is present. As regards metals, the category 2 level could not be reached in any of the tested samples by means of washing at acid pH and variable L/S. Washing may hold good perspectives for bottom ash in which metals only slightly exceed the category 2 limits.
5.3
Combinations of w a s h i n g and aging
The sequential order of washing and ageing can be important for several reasons: 1) there may exist positive or negative interactions between both processes; 2) washing after aging will result in less contaminated washing water; 3) after aging, sulphate is more mobile, and so less washing water might be required in order to attain the category 1 value.
20 Three variants were investigated in small-scale tests, namely washing (L/S 5) followed by aging (6 weeks); 3 weeks aging followed by washing (L/S 5) followed by 3 weeks aging; and aging (6 weeks) followed by washing (L/S 1). The investigation showed that the end quality of all samples was almost the same, always well under category 2 limits. For many components, category 1 values were attained, but not for antimony. Sulphate only approached this value after washing up to L/S 5. Table 6 presents the contaminant loads that were removed during the washing process. Metal and metalloid loads in the washing water prove to decrease if aging of the material precedes the washing process. This is not so for salts although sulphate loads are lower at the aging/washing sequence, the leaching results show that washing at L/S 1 is not sufficient for obtaining the desired quality improvement. Interpolation of the measuring data shows that for washing after aging, 20 - 25% less washing water is needed in order to attain category 1 sulphate values. It is however obvious that this process order does not result in the desired low washing water volume of about 1 m3/tonne (L/S 1). The solubility of sulphate remains a limiting factor, also after aging has taken place, so that relatively much washing water (L/S 3 - 7, depending on the sulphate content in the bottom ash) will still be required.
Table 6
Loads in washing water in mg/kg ds in relation to process sequences of washing and ageing MSWI 1
MSWl Process (1)"
MSWl 2
IJ
m
WILLS 5)/ A(6w.)
A(3 w.)/WIL/S 5)/A(3 w.) ,,
Salts Chloride Sulphate
1000 4500
1175 4700
940 1950 .....
1325 6250
1175 6250
1450 2350
Metals Antimony Copper Molybdenum Zinc
0.28 1.65 1.35 0.25
0.07 0.65 0.11 1.6
0.05 0.19 0.09 0.85
1.30 5.50 0.85 0.75
0.11 2.15 0.11 1.20
0.11 1.30 0.14 0.83
(1) W = W a s h i n g
A(6 w.)/ J WILLS 11 ~,',
WILLS 5)/ A(6 w.)
....
I A(3 w.)/WILIS J 5)/A(3 w.)
A (6w.)l WILLS 1 )
at s t a t e d L/S v a l u e ; A = aging at stated r e s i d e n c e t i m e
Tests with bottom ash from three MSWIs, at a scale of 100 kg, during which washing at L/S 1 was applied before or after aging, proved that: 1) when the fresh material does not leach too much (values of about 50% above category 2), washing may suffice in order to reach category 2 levels; 2) washing in combination with aging is useful when chloride or bromide must be removed, or the molybdenum removal efficiency increased. The Mo-performance of the combination process (70 - 95%) lies on average about 20% above the performance of just aging. Therefore, washing is a good option for increasing the Mo-performance; 3) washing could furthermore be used to reduce the required aging residence time. Tests however show that for strongly leaching bottom ash, a far-reaching reduction of the residence time to 1 week is not feasible. Bringing down the residence time from 4 to 3 weeks however might be possible. Economic considerations also play a role in this.
2] 6
C O S T S OF Q U A L I T Y I M P R O V E M E N T S
The costs of MSWl bottom ash treatment were estimated on the basis of a conceptual design of an installation with a capacity of 100,000 tonnes per year. The installation comprises a number of roofed, concrete basins. The process evolves in batches, the same basins can be used for aging and washing by leading either water or gas through the ash. On the basis of MSWl bottom ash quality differences, four different treatment options were worked out: 1) enhanced aging during 1 week; 2) enhanced aging during 1 week, followed by washing up to L/S 1; 3) enhanced aging during 4 weeks; 4) enhanced aging during 4 weeks, followed by washing up to L/S 1. Longer residence times of course lead to larger installations, whereas for the washing water an extra buffer tank is needed in order not to overload the water treatment installation. The estimation of the investment costs includes the costs of construction of the concrete units, including roofing, foundation, costs of the land, ventilators, gas mixture feed and outlet pipes, transport belts for the supply of the material, and a switch box. In first instance it was assumed that purified flue gas was to be used for carbonation, and as an alternative, the purchase of C02 was included. A schematic drawing of the installation is presented in Figure 1. Flue gas or CO=
Gas out(• l
Water
recycling
,
iiiiiiii
recycle(_.*80~1
I,
MSWI bottom ash
; -G
I
I I I I I I I I
I
~ A i r-humi-
Hild~r
~'~I (open)
Heaer (open)
I I Washwa~r
(|
Figure 1 Schematic representation of the installation for enhanced aging and washing of MSWl bottom ash
22
For an assessment of the exploitation costs, the f o l l o w i n g items w e r e taken into account: * interest and repayment (annuity depreciation, instalment of 10 - 20 years, 8 % interest); * maintenance: 2 % of construction and 6% of mechanical-electrical investments per year; * personnel: 0.5 - 0.8 man year for operation; * use of shovel (and personnel) for emptying the installation; * electricity, insurances; * costs of w a t e r treatment: all in costs of NLG 3 - 6 per m 3 (provided that it is allowed to discharge such salts as chloride and sulphate); * costs of purchase of CO2: NLG 0 . 2 5 per kg. The results of the calculations are summarized in Table 7. The inaccuracy of these estimations is + / - 4 0 % . Variations in depreciation periods and higher costs for CO 2 purchase are included in this margin. Table 7 s h o w s that enhanced aging is the cheapest w a y of improving the quality of M S W l b o t t o m ash. If bromide and chloride are critical components, and w a s h i n g is therefore required, applying only the washing process is an alternative provided that the leachate does not contain too high concentrations of e.g. copper. If strongly leaching b o t t o m ash is the basic material, a combination of w a s h ing and aging will be the most attractive option from an economic v i e w p o i n t . As w a t e r treatment costs are the predominant cost factor in the w a s h i n g process, higher w a s h i n g w a t e r volumes will sooner lead to higher costs than longer residence times do.
Table 7
Costs of M S W l bottom ash quality improvement, in NLG per ton
Process
Costs in NLG per ton bottom ash
1 week aging washing up to L/S 1 1 week aging + washing up to L/S 1 4 weeks aging 4 weeks aging + washing up to L/S 1
7
DISCUSSION
5~
7. = - 10.50 9.50 - 12.50 10.50 14.50- 17.50
AND CONCLUSIONS
Enhanced aging is an effective technique for reducing the leaching of metals and organic substances within a short period of time (1 - 4 weeks). A reduction can be obtained of 80 - 9 0 % , and for m o l y b d e n u m of over 6 0 % . Aging does not result in reduced leaching of salts - in order to achieve this the aging process should be combined w i t h washing. Chloride and most probably also bromide salts can easily be removed by washing at an L/S value of 0.5. In order to remove sulphate, considerably higher L/S values are required, namely 3 - 7. From the v i e w p o i n t of the desired final quality and costs, the f o l l o w i n g (combinations of) processes are the most attractive: Objective
Bottom ash quality (1)
process choice
Category 2
metals critical, salts not salts critical, metals moderately raised salts critical, metals strongly raised
aging washing aging + washing
Category 1
difference not of importance ...............
aging + washing
1) quality in relation to category 2 limit values
23 The costs of aging are between NLG 5 - 10. = per ton bottom ash (depending on the required residence time), of washing NLG 7 - 1 1 . = , and of combination processes NLG 10 - 18. =. In order to reach the category 1 level, costs of about NLG 25, = per ton should be taken into account. At the moment, upscaling and optimization of the process is being investigated on pilot scale (50 ton). During this investigation, both aging and combinations of aging and washing are examined.
8
ACKNOWLEDGEMENT
The financial support (contract number 3 5 3 6 9 3 / 0 5 1 0 ) of the National Research Program for the Recycling of Waste Substances, which is jointly managed by the Netherlands Agency for Energy and the Environment, and by the National Institute of Public Health and Environmental Protection, is gratefully acknowledged.
9
REFERENCES
1. Bouwstoffenbesluit bodem- en oppervlaktewaterbescherming. her Koninkrijk der Nederlanden, 567 (1995).
Staatsblad van
2. Steketee, J.J. & Urlings, L.G.C.M. Enhanced natural stabilization of MSWl bottom ash: a method for minimization of leaching. In: Goumans, J.J.J.M et al (eds.) Environmental aspects of construction with waste materials. Elsevier, Amsterdam (1994). 3. Simon, F.G., Schmidt, V. & Carcer, B. Alterungsverhalten von MVA-Schlacken. M(~II und Abfall 11:759-764 (1995). 4. Zevenbergen, C. Natural Weathering of MSWl Bottom Ash. Thesis Universiteit Utrecht (1994). 5. Lahl, U. Schlackeaufbereitung durch Alterung und Laugung. In: D.O. Reimann (Hrsg.): Entsorgung von Schlacken und sonstigen Reststoffen. Beihefte zu MUll und Abfall. Heft 31, Erich Schmidt Verlag, Berlin (1995). 6. Buijtenhek, H.S., Steketee, J.J. & de Zeeuw, J.H. Praxisbezogene Entwicklungen und Techniken zur gezielten Schlackenw&sche. In: D.O. Reimann (Hrsg.): Entsorgung von Schlacken und sonstigen Reststoffen. Beihefte zu MUll und Abfall. Heft 31, Erich Schmidt Verlag, Berlin (1995).
This Page Intentionally Left Blank
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
CONSTRACTION MATERIALS MANUFACTUARING BY THE TECHNOLOGY OF MELTING Seiichi Abe, Kubota Corporation, Japan Namba AK Bldg, 3-1-4 Motmachi, Naniwa-ku Osaka 556, Japan 1.
Preface Most municipal waste is incinerated first ; then the residue is buried at landfill sites equipped with waste water t r e a t m e n t facilities. This is the general waste disposal way in Japan. 50,304 thousand tons of wastes are g e n e r a t e d every year, and 73 % of t h e m are burnt. Subsequently, 6,O13 t h o u s a n d tons of incineration residue is left to be buried. Obviously, refuse incineration, which reduces the bulk of waste, is an indispensable t r e a t m e n t m e t h o d for J a p a n in view of the small land area and the need for effective land use. However, it is also true t h a t m u c h dioxin is emitted to the e n v i r o n m e n t during the incineration process. In J a n u a r y 1997, the M H W (Ministry of H e a l t h and Welfare) announced its policy to reduce dioxin e m i t t e d to the e n v i r o n m e n t to 5 ]z g or less per 1 ton of waste incinerated. In order to a t t a i n this target, not only dioxin in flue gas generated through incineration is reduced, but also dioxin contained in bottom ash and fly ash m u s t be decomposed t h r o u g h exposure to a high t e m p e r a t u r e atmosphere for example. Because the most effective m e t h o d for this is a melting furnace system, the MHW recommends the introduction of the melting furnace facility to each municipal government. The m e l t i n g furnace system inevitably g e n e r a t e s slag, but only dumping slag will m a k e no contribution to prolonging the lifetime of landfill sites. On the contrary, this slag can be utilized as a m a t e r i a l resource for products used in daily life, because it contains almost no dioxin and extremely small quantities of heavy metals. Substitution for sand will be the simplest and easiest utilization of the slag as material resources. In addition, it (,'an be used as a raw material for sintered products such as tiles, taking a d v a n t a g e s of its t h e r m a l characteristics. The paper presents the e x p e r i m e n t a l results of tile production from the slag.
25
26
2.
Test Method ( 1 ) Test Samples As test samples, we used the slag obtained after melting bottom ash of municipal waste and the slag after melting fly ash generated during incineration. The analytical results of constituents are shown in the Table-1. Furthermore, the constituents of the (:lays and frits used as additives are also listed.
Table 1
Analytical 9 Results of Constituents of Slag & Others
Constituents Si02 Al20:~
_
(%) _(%).
CaO
_
(%)_
Fe203
_(%)_
_
_
(%)
Na20 K20 T-S -T-C1 Cu -Cd -Pb As Zn
__ _
(%)
M~O
-T-Hg Cr
Aggregate Molten Sla_gz Botto~n-Ash , Fly Ash
Units
-
(%) (%) -(%)(mg/kg) -(mg/kg) -(mg/kg) (mg/kg) (mg/kg) (mg/kg) (mg/kg)
-
38.4 19.1___ 25.38 3.3 3.1 2.8 1.1 0.25
35.51 _
15.35
31.05 3.3 3.95 1.87 0.65 0.41
N.D.
N.D.
1600 N.D. 140 1.7 1600
410() N.D. 170 1.2 830
_
N.D.
N.D.
520
230
Molding Mat'l Ordinary Clay 60.16 26.96 0.18 0.88 0.13 0.21 2.8
Additive Frit
m
m
_
_
50.28 5.8 0.49 N.D. 0.8 7.89 0.11
27
( 2 ) Trial M a n u f a c t u r i n g of S i n t e r e d P r o d u c t s Fig. 1 shows the trial manufacturing procedures. First of all, clay, frits, and water were added to the slag pulverized by a ball mill. Then, they were mixed and kneaded to make a test piece through press molding. After drying, we sintered it in an electric" furnace at 1000~ Table 2 shows the m a n u f a c t u r i n g conditions for the test piece. The test pieces used for the Bending Test are shown in Photo 1.
Slag Pulverized
I Raw Mat'ls Compounded
I Mixed and Kneaded
Molded
Dried up
Sintered
L Tested / Inspected Fig. 1 : Trial Manufacturing Procedures of Sintered Product
Table 2 : Manufacturing Conditions of Sintered Products
Compound Ratio of Each Material Water Content during Mixing & Kneading Test Piece Dimensions Molding P r e s s u r e Heating-up Speed Sintering T e m p e r a t u r e Retention
Slag : Clay : F r i t = 60 : 20 : 20 (%) 8% 100mm • 30mm (for bending test,) 100mm • 100mm 500 kgf/cm 2 100~ 1000~ 1 hr
28
Photo 1
The 9 Test Pieces Used for The Bending Test
( 3 ) Measuring
Methods D i m e n s i o n a l accuracy, bending strength, and w a t e r absorbance were m e a s u r e d respectively to find the c h a r a c t e r i s t i c s of the sintered products. (T~ Dimensional Accuracy We m e a s u r e d the side of 10 sample pieces to compute an averaged value. The difference between the a c t u a l length of the sides and the average value was defined to be dimensional tolerance. I,l+L2+
"-- L I ) + L I ( )
Dimensional = Tolerance (mm)
-
Ln
1() Ln 9Side Length o f n th sampe (mm)
('.'2) Bending Test We applied the testing method of :~-points bending. Cah:ulation was made as follows" 3 Pl,
Bending Strength (kgf/c:m 2)
,3:
P : Maximum l)estruc.tion Loading (kgf) L :Span l,ength (cm) w: Width ()f Tes|. Piece (cm) t :Thickness of Test Pio.ce (cm)
= 2 wl 2
Wale, r Abs()rl)ance W
W a t e r A/)s()rban('e ( % )
--
-
W()
x
1()()
W()
W 9Weight, aher Absorption (g) Wo 9 Dried Weight (g)
29
3.
Test Results
( 1)
D i m e n s i o n a l Accuracy Table 3 shows test results.
Table 3
2
3
4
5
6
7
8
9
10
93.30
92.60
93.40
93.20
93.65
93.30
92.80
92.90
93.05
93.25
93.145
+0.15
-0.55
+0.25
+0.05
+0.50
+0.15
-0.35
-0.25
-0.10
+0.10
• 0.55
91.70
92.15
91.40
92.20
92.35
91.20
91.90
91.65
92.25
92.15
-o.2o
+0.25
-0.50
+0.30
+0.45
-0.70
0.00
-0.25
+0.35
+0.25
1
Sintered P r o d u c t (~) * 1 Difference from Average Sintered P r o d u c t (~) * 2 Difference from Average
Test 9 Results of Dimensional Accuracy (mm) Average
91.895 • 0.70
, 1 9S i n t e r e d p r o d u c t s made of bottom ash slag , 2 9S i n t e r e d p r o d u c t s made of fly ash slag As a t e s t result, it is found t h a t the d i m e n s i o n a l tolerance of the sintered m a t e r i a l made of bottom ash is -+0.55mm and t h a t of the one made of fly ash slag is -+0.70ram. Both t o l e r a n c e s stay within +-2mm as those specified by JIS S t a n d a r d s .
(2)
Bending Strength Tables 4 and 5 show the test r e s u l t s of each sample.
Table 4
Measurement 9 Result of Bending Strength
29.85 8.20 100.9 538.7
2 29.35 7.95 98.3 556.4
3 29.30 7.90 95.1 546.1
29.10 8.35 170.5 882.4
29.05 7.50 147.5 947.8
29.00 7.45 135.5 883.9
1
Sintered Product made of bottom ash slag
Sintered Product made of fly ash slag
Width(mm) Thickness(mm) Load(kgf) Bending ,1 Strength(kgf/cm2) Width(ram) Thickness(mm) Load(kgf) Bending ,1 Strength (kgf/cm 2)
9 Bending S t r e n g t h was m e a s u r e d a t 70ram span length.
4
5
29.35 8.25 109.7 576.6
29.30 8.05 114.2 631.5
29.15 7.90 156.1 900.9
29.05 8.05 177.0 987.2
30
Table 5 : Bending Strength of Slag-made Sintered Products (kgflcm 2) l
Type of Tile Sintered Product made of bottom ash slag Sintered Product made of fly ash slag ,Presently m a r k e t e d sample tiles
Averaged Bending S t r e n g t h 570 920 330
As shown in the above table, the sintered product has a strength of 570kgf/cm 2 or more. Slag-made sintered product is in no way inferior to the tiles presently marketed. Therefore, it is considered to be an appropriate material not only for walls and roofs but also for floors which require higher strength.
(3)
Water Absorbance Table 6 shows the test results of water absorbance.
Table 6 : Water Absorbance of Sintered Product ( % ) Type of Tile Sintered Product made of bottom ash slag Sintered Product made of fly ash slag Presently m a r k e t e d sample tiles
Water Absorbance 10.5 0.2 0.4
As shown in the above, water absorbance of sintered product made of incinerated ash slag and sintered product made of fly ash slag is 10.5% and 0.2%, respectively.
4.
Conclusion
The fifth)wing advantages of slag-made sintered product were found through these tests ; (1) It is possible to sinter even at relatively low t e m p e r a t u r e of I()()()~ (12_)The strength is comparable to ordinary tiles. (i"3)Dimensional accuracy is superior due t() small shrinkage. Meanwhile, it is ascerlained through other tests that s()me crystals such as an()rthile (Ca() 9A}~ 2():~ 92Si()2), diopside(Ca() 9MgO 92Si()2), gehlenite(2Ca() 9A~ 2():~ " St()2) are fi)rmed in lhe slag. (]enerali()n ()f lhese crystals are probable factors which increase the strength ()f slag-ma(le sintered l)r()(lu(:Is, but this phen()men()n has not been fully eluci(lated yet. In any (:vent, it is a fact that low teml)eralure sintering is possible. ()bviously this will bring al)()ul new l)ractical uses for the slag such as tiles as a constructi()n material.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997Elsevier Science B.V. All rights reserved.
31
Producing Permeable Blocks and Pavement Bricks from Molten Slag Masahide Nishigaki Technology Development, Takuma Co., Ltd. 2-33, Kinrakuji-cho 2-chome Amagasaki, Hyogo, Japan
Abstract Studies of the technology of melting and solidifying are underway as a means of detoxification and volume-reduction of municipal solid waste incineration residue. The molten slag does not leach heavy metals, and can be used as concrete gravel and roadbed aggregate. We have produced water-permeable blocks from slags made by a surface-melting melting furnace and a plasma-melting furnace. Also, we have fabricated and installed pavement bricks using slag from a commercial surface-melting furnace. Both attempts have been successful, as the end results satisfied the required product standards, and showed no heavy-metal leaching.
1. Introduction The treatment and disposal of municipal solid waste (MSW) come under the jurisdiction of the municipality where it is generated, and it is the common practice to incinerate the waste and landfill the subsequent residue. However, acquiring suitable sites for landfilling is becoming more difficult every year, and the reduction of waste itself as well as the volume destined for landfill, has become an urgent issue. As new regulations controlling waste disposal methods and recycling are issued, vast efforts are made in recycling wastes and developing means of effective re-use of the residues. A new technology has been developed by which the residue is melted at a high temperature and the molten ash is turned into slag by quenching, reducing the waste volume and detoxifying simultaneously. There are 16 melting facilities in operation today, and experiments are underway to utilize the slag coming from these plants. A standard specification for its application is being prepared for the promotion of its use. Due to its characteristics, use of slag is mainly concentrated in the areas of roadbed materials and concrete aggregates. In this article, we report the test results of making water-permeable blocks made of slag coming from two different systems of melting, and also the results of fabricating and installing pavement bricks made from slag produced by a commercial melting plant.
2 Melting Systems The methods of melting incineration residue can be divided roughly into two major categories, fuel-burning system and electric melting system. Takuma has already built seven operational facilities in the fuel-burning category, which are Reflective Surface-Melting Furnaces. In the electric melting system, Takuma is currently developing and experimenting Graphite-Electrode
32 Plasma Melting Furnace. Both types of melting furnaces are installed adjacent to MSW incineration plants. The residual ash is first processed by magnetic separator and sorted according to the particle size before going into the melting furnace. The slag is either quenched in water, or air-cooled gradually. Figure I shows the incineration residue melting system. In c i n e r a t e r
Baghouse
ExhaustOas
Cleaning Unit
1[
Stack
[ Exhasut Gas I
Melting Furnace Surface Melting F
Water-quenched Slag
Plasma Melting Furnace Slag Conveyor
Air-cooled Slag
Figure 1 Refuse Incineration Residue Melting System 2.1. Surface-melting Furnace Figure 2 illustrates the structure of a surface-melting furnace. The ash which was stored in the ash storage bin is fed into the furnace by the ash extruder. It forms a melting slope at roughly the equilibrium pitch of the ash. Burners pointing approximately perpendicular to such slopes are installed on the roof of the furnace. The ash is melted from its surface by the heat of these burners and the radiant heat from the refractory materials of the roof. The molten slag flows down along the melting slopes which arc located diametrically around the slag tap at the center of the furnace. The slag flows further down along with the combustion gas into the quenching bath, and chilled quickly, it turns into black
Ash Hopper
Ash Feeder
----L
I-
Melting Furnace
Slag Conveyor
Figure 2
Surface-melting Furnace
Burner
33 granular slag. The furnace exterior is cooled by the water jacket, and its inner wall consists of refractory materials of superior heat and erosion resistance.
2.2. Plasma Melting Furnace Figure 3 shows the schematics of plasma melting furnace. It is lined with refractory materials which are protected on the lid and side walls by water jackets. The bottom has aircooling box that protects its refractory linings. MainElectrode The incineration residue is fed into the furnace at a pre-determined quantity by ExhaustGas means of a screw feeder. It is melted into slag Starting Electrode ~] A by the conductive heat from the hightemperature plasma arc. It then overflows through the slag tap continuously. The exhaust gas from the melting furnace goes out through the slag tap where the Ash unburnt gas generated by the melting process is burnt, preventing the molten slag discharge from cooling and solidifying. Most metal contents in incineration residue such as iron and copper are reduced during melting, and accumulate at the bottom of the furnace due to their higher specific gravity. They are discharged periodically by opening the tap hole located at the bottom of the furnace.
Figure 3 Plasma Melting Furnace
3. The Characteristics of Molten Slag We have melted bottom ash and fly ash from a stoker-type incinerator using various surface-melting fumaces and plasma fumaces of 5t/day capacity, and obtained data on slag characteristics. The bottom ash was wet, and it came through 30mm mesh, then de-watered down to 5% moisture. Its ferrous content was removed by the magnetic separator. The fly ash came solidified with cement, and was tested as it was. Tests were carried out using (1) bottom ash alone, and (2) a combination of three parts bottom and one part fly ash.
3.1. Analysis Table 1 shows the results of analysis of the slags The main ingredients of the slag are SiO2. A1203 and CaO. While surface-melting is done in an oxidizing atmosphere, it is in a reducing atmosphere with the plasma-melting furnace. Therefore, Fe203 is reduced into metallic iron in the plasma furnace, decreasing the iron content in the slag. This also gives more grayish tinge to the slag. The same can be said about other metals.
34
Table 1
Result of Analysis of the Slags
Item
Surface-Melting. Furnace B o t t o m Ash i Mixed Ash Slag Slag*
Unit
Plasma Melting . Furnace B o t t o m Ash i Mixed Ash Slag ~ Slag*
.................... .s!.o.~ ................................ . ~ .............................. .3...9.:0.......... i .................... .3..7.:~.......................... 4..3.:.3.............. i ................ .41:.4 .........
TiO2 % 1.3 ~ 1.1 1.2 ! ................... .A1...2.O................................ 3 .,~ .............................. 2.3=3............................... 2.3=9.......................... 2.8:7.............. Fe203 % 10.2 " 9.2 2.9 i CaO % 19.2 i 22.3 18.9 i ..................................................
, .........................
, ....................................
~ .....................................
..................................................
i .........................
b ....................................
.,. .....................................
: ..................................
r
M~.O................................
!
9 .....................................
.i. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1.2 26.0 2.9 24.0
, ..................................
. ~ ................................ 2 : 9 .......... i ...................... .2.:7 ............................ ..3.:O............. i ................... .2.:.8.......... .................... .~..O. ................................ ..~.o................................. 1..1 .......... i ...................... O.:8 ............................. 1.:0. ............ i................... O:.4 .........
Na20 P205 Softening Point Melting Point Flowing Point
..................................................
% % ~ ~ ~
, .........................
.....................................
, ....................................
3.9 2.4 1,170 1,180 1,200
9 .....................................
.: ~ .....................................
3.0 2.3 1,200 1,200 1,220
3.8 0.9 1,200 1,210 1,280
~ ~ i ~ i
4. .....................................
2.5 1.0 1,220 1,230 1,260
, ..................................
.,. .....................................
i ..................................
* M i x i n g Ratio: B o t t o m A s h 9F l y A s h = 3 91
Table 2
Physical Test Result
.................. M e l t i n g . . F u m a c e
of the Slag
............................. S u r ( a 2 e
No.
F1-Li
Specimen
II
Bottom Ash
%
..... i ~ i y 3 ~ f i ............................................
~e!tinz
F1-S
F . u m a c e ................ P ! a s m a F2-L
F2-S
P1-L
~e!t!ng
100i
i
100
75
75
100
100 i
~ ....... /I ............ i ~ ! ............ 6 .............. 2 ~ ............ ~ ................ 6 ................ 6 i
Particles Clearing: : ,..53..mm.M.e.sh..Si.z..e............................ .~ ......... 100.Oi 100.0 37.5mm % 94.9i 96.1 31.5 % 93.0i 94.5 26.5 % 90.1 i 93.8 19 % 87.4i 91.7 "'13"~2.................................................................................... % 84.1 :i................................................................ 90.0 9.5 % 82.41 100.0 88.5 100.0 4.75 % 75.3 i 94.6 83.9 88.2 2.36 % 55.1 i 77.8 64.2 64.5 ............................................................................
.,, .................
~ ......................................................
.......................................................
4 .......................................
r ....................................
.......................................................
r .......................................
r .................
............................................................................
~
r .................
4 ..................
~ ..................................
.......................................................
r ......................................
r .................
4 ..................
b ..................................
.......................................................
.i ....................
9. . . . . . . . . . . . . . . . .
a ..................
, .................
.................
. 100.0, 97.7 9",~12".................. i ..................... 91.9 100.0i 100.0 83.3 94.2i 95.1 69.3 74.5i 60.1
~ .........................................................................
.,, ........................................................................
,,
~ .....................
r .................
4 ..................
r ......................................
75
........ ~ ........
i
o, ...................................
~ ...................................
.......................................................
.,i
....
P2-S
i
~1
,Furna~
P1-S
.~ .....................
.........................................................
.................
r .................
4 .....................
r .................
4 ....................
9.................
~ ....................
..~:1..8. ..................................................... . ~ .................. .2.5:7i ....... 3.8.=3............ 2.8.7 .......... .2.9.=2 .......... 3.3,3. .......... 28:5i ....... 2.5..6 .... ..60.0.~.m. ............................................... % .................... .8.:.!.i ....... 1.4:1 .............. 8.7. ............. 7:8 .......... 1..3.#.. .......... 1.4=8.] .......... 7:1..s... 3 0 O .......................................................
150 Specific Gravity Saturated Dry Surface
......................
.~ . . . . . .
r ........................
% .................... 3..8 i ......... 6 :~1 ........... 3:0 ............. 3:4 ............. 6 . 2 ............. 3.3 i .......... 2 . 9 ....
% ,~ . . . . . . . . . . . . . . . . . . . . .
1.9i 2.064
.................
2.94i 2.744
r ..................
1.2 2.294
~. . . . . . . . . . . . . . . . . .
1.4 2.657
t ..................
2.8 2.423
.................
2.0i 2.644
r .................
1.5 2.665
4 ....................
..Bul..k..S~.e.c.i.0c..~r.a.v!t:t. ..................... - ......... ....1.:.87.6. ....... ..2.#..1...8........ .?.:.1...3..8......... .2..#..8.4 .... ...2:.25..1 ........ 2..60.1. ......... 2.:.619. ....
Apparent S.G. Water Absorption Rate Friction Loss Stability Wash Loss Weight/Volume Solid Volume Ratio Optimum Compression Moisture Maximum Dry Density Modified CBR Liquid/Plastic Limit
%
2.307 2.793 2.534 9.958 0.990 7.306 70.6 17.0i 64.1 23.9 27.3 6.4i 1.8 1.3 1.0 1.426 1.540 1.489 75.4! 58.7i 69.7
2.787 2.801 15.4 8.0 1.0 1.516 58.7
2.720 7.644 61.1 17.6 1.6 1.441 63.7
59.5 i
1.582 60.4
~cm 3 % %
15.0i 15.0! 10.6 1.801 i 1.801 i 1.860 29.4i 29.41 29.5 NP i NP i NP
10.6 1.860 29.5 NP
11.7 i 11.7i 1.692 ! 1 . 6 9 2 1 22.7 i 22.7i NP i NP i
1.786 25.9 NP
ke,/l
i '
2.718 1.661 16.2 4.8 1.5
1.548
2.746 1.774 14.7 10.7 1.7
11.o
35
3.2
Results of the Physical Tests Table 2 shows the results of physical tests of the slag. The letter 'L' in the table represents the results of tests according to the coarse aggregates standard (particle size distribution:10---40mm) whereas the letter'S' indicates the results according to the fine aggregates standard (particle distribution:0--10mm). The latter slag was shifted through 10mm mesh. The average particle sizes for surface-melt slag and plasma slag are both within the range of 1--1.5mm, but surface-melt slag has larger portion of coarse particles. Water-quenched slag is glassy, which, if containing coarse particles of more than 10mm, is fragile, and is inferior in water-absorption, friction loss and stability. It does not meet the specification of coarse aggregate. Once coarse particles over 10mm are removed, the test results in each item are improved. The particle size distribution indicates higher percentage of 5--2.5mm particles and lower percentage of lmm and smaller particles. This water-quenched slag can be crushed, and its particle distribution can be adjusted to satisfy the specification for fine aggregate. Its modified CBR value is between 20% and 30%, but it can be used as road aggregate without ill effect when mixed with other roadbed materials available commercially. 3.3
Safety The result of leaching test, as the prime safety benchmark for the slag, shows that the leachings of heavy metals are within the guideline, and satisfies the soil standards specified by the publication No13 and No46 of the environmental agency. 4.
Producing W a t e r - P e r m e a b l e Blocks
The results of producing water-permeable blocks (size: 200mmxl00mmx60mm) are described below. As mentioned earlier, bottom ash by itself, and its mixture with fly ash, were used in surface-melting furnace and plasma furnace,
4.1
Method of Manufacturing Figure 4 shows the manufacturing process of the water-permeable blocks. The slag went through mesh size 5mm, and the iron content that causes bubbling is removed by the magnetic separator. Two kinds of blocks were made, one with a single layer of slag, and another with two layers, the surface layer being added for appearance using shard. Table 3 shows the mixing ratio for the components of these blocks. The base layer consists of slag and sintering accelerator that sinters the slag at temperatures below the melting point of the slag, at the ratio of 91:9. Water solution of CMC(an organic binder) was added for molding. Since the slag from the surface-melting furnace showed melting point about 30 ° to 50°C lower than that of slag coming from plasma furnace, the sintering temperature was kept lower than the melting point. The period of sintering for the slag from surface-melting furnace is kept at 15 hours of rising temperature, 2 hours maintaining at 900°C, while they are 17 hours, and 2 hours at 1000°C, respectively. 4.2 The Result of Production The results of various physical tests of sintered water-permeable blocks are listed in Table 4.
36 The permeation index for both types satisfies 0.01cm/sec defined by the Interlocking Association Specification. On average, single-layer block turns out 0.046cm/sec whereas two-
Pre.Treatment '
'~
I Shifting through 5mm mesh; Magnetic Separation
Table 3
Mixing Materials ....
,-Sintering Accelerator ,-Shard
[ KneadingjBlendin ] ----CMC Water Solution I
Molding
Mixing Ratio
Raw Material Slag Base Layer Accelerator Slag Surface Layer Shard Accelerator
. . . .
Single-layer Brick
! Compression @60kg/cm 2
[[ Ratio(%) 91.0 9.0 45.5 45.5 9.0
Double-Layer Brick Surface Layer
[
D~,,il~,
,,,
] 48 Hours at 110~
I
1900~176 Figure 4
Production Process of Water Permeating Blocks
Base Layer
layer block betters it by 37% at 0.063cm/sec. This is attributed to the fact that the permeability of the shard is so high that the difference c o m e s from the base layer. Relative bending strength is lower with the two-layer block that can be presumed as caused by the low affinity between the base and top layers. The heavy metals leaching test proved that it is within the soil standard specified by the Publication No13,No46 of the Environmental Agency.
Table 4
Physical Test Results of Water-Permeable Blocks Furnace
Surface-Melting Furnace
........................~~iMatenal~~................................t~ottom~ ................ Asn~i.......i ! " B ~ 1 7 6 i Fly Ash ...M...o..1..d...e..d....D....e..n...s..i.t.y. .........~...K/..c..m... 3" Si.nte:reaDe.nsity
........ i.
cm::L
1.88 i
1.80i
Plasma Melting Furnace
I! []ILB ] ..Sp.ec.....t .........tmttom~ ................,,,~sn~i .......rBoitomAsia+lli,,It Value ] . i FlyAsh ]1
1.90 i 1.87
1.86i
1.81 i
1.89i
1.85
:::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::::
..........- .......
...S...h...r..i..n...k..a.2_,.e.:....I~.....ng.t..h.....i ........%........ 0.54i 0.39i 0.58 ~ 0.48 0.731 0 . 7 2 i 0.05j 0.11 -Width i % 0.55! 0.47! 0.68 i 0.68 0.87i 0.89 i 0.13i 0.34 .........'r......... ..........................-...H....e.i~.h.t.....i........%...............0.,..1..9...i.........0.,..2...5..i.....L.0...0......~......1.,..2...9...............1.,..1...3...i.......1..,.4..9....L....-Q..,..2...1...i .........-0....0..~ Permeation Index i c..m_2.se 0.038 !0.051 i0.042 i0.065 0.061 i0.082 i0.039 i0.052 0.01 Bendi.n.g.Strength [ k~/.c 46.7i 40.4i 51.7 i 57.1 38.3 i 32.0 ~ 40.2 i 35.0 30 Comoression i k~f/c 184.7i 278.3i 300.0i 321.6 474.3i 521.1i 319.51 274.1 170 ..................
.................................................
." . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
................................................
.. .................
,.. ..................
. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
,~ . . . . . . . . . . . . . . . . . .
? ...................
, .....................................
+ ..................
.
r ..................
-. ...................
." . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
~ ..................
-r ..................
~ ..................
.. ..................
~ ...................
. ..................
.: ..................
-
-. ......................................
t ..................
? ...................
9 .....................................
i! ..................
....................................
..................
+ ..........................................................................
37
4.3 X-ray Diffraction of Water-Permeable Block In order to determine the mineral composition of the product, X-ray diffraction was done, and diffraction values are calculated to show their mineral composition, as shown in Table 5. It reveals that the main components are SiOz, AlzO3 and CaO, causing gehlenite and anorthite to precipitate, and also more augite precipitated in the surface-melting slag block than that of plasma furnace slag, due to the influence of iron. The exterior view of waterpermeable block is shown in Photograph 1.
Table 5 Mineral Composition of Water-Permeable Blocks Furnace
Surface Melting
Bottom Materials
Gehlenite Anorthite Augite Wollastonite
ASh 1-Layer Block +
i Bottom Ash
i + Fly ASh i 2-Layer i Block i
+ +++
i
Plasma Melting Bottom i Bottom Ash i +Fly Ash Ash 1-Layer i 2-Layer i Block Block
+++
+++
+++
+
++
++
+++
+
+
Precipitation +++>++>+>o 1. Gehlenite : 2CaO.A1203.SiO2 2. Anorthite : CaO.A1203.2SiO2 3. Augite : Ca(Mg,Fe)Si206 4. Wollastonite : CaO.SiOz
Photograph 1 Water-Permeable Block
5. Production of Pavement Brick
We produced approximately 11,000 pieces of pavement brick using slag from a surfacemelting furnace we installed. The bricks were used to pave the footpath in our new head office premises, covering roughly 240 m 2.
5.1 Preliminary Testing Before the production, a preliminary test was conducted. The sintering was to be done using ordinary tunnel kiln. Sintering temperature was set within the 1200--1230~ range, and the suitable material mix and the granular sizes were determined accordingly. Small test lots of 2 tons were formed and ~Slag Brick sintered in order to locate problem areas. The ~--~, ' " " ',~,-~.............. Shelf following discoveries were made. l - S u p p o r t Brick (1) Slag size was to be limited to 1.0mm and ,l[ ~ ~ ~ ]-~~_______---Brick Base smaller, and the slag ratio was set at 20%. (2) Sintering was done of shelves, and multiple ~Truck layers were sintered to the extent that no deformation occurred. (3) Sintering Temperature was set at 1200-.1230~
t_q
Figure 5
Brick Sintering Shelf Assembly
38 (4) The color variance of the product was such that the top layer became dark while the bottom layer did not get sufficient sintering, and turned grayish. This was avoided by placing bricks with larger spacing. (Refer to Figure 5 Brick Placement on Sintering Shelves) (5) For coloring, 2% of chromite pigment was added after checking the relationship between the amount of pigment and product coloring.
5.2 Production Process Figure 6 shows the production process of pavement bricks. After having its ferrous metals removed by the magnetic separator, the slag was crushed, and went through 1.0mm sieve. The recovery rate was approximately 95%. The brick material ratio was: Slag(1.0mm<): Grog (3.0mm<): Ceramic Gravel (1.0mm<): Clay (1.0mm<) at 20 : 35 : 25 : 20. To this mixture, 2% of pigment were added. Kneading and blending was done by a Mtiller mixer for 15 minutes. Molding was done by a 200 ton friction press, and the bricks were loaded onto the sintering truck. Drying for about 48 hours at 80 - 150~ was followed by the sintering process in a tunnel kiln for about 80 hours at 1200~176
(~) From Slag to Material i
Slag
Ii "iJ Drying ~-*l Crushing 200~3000c Fret Mill
~-q Sieve H
Bagging
1.0ram Mesh
Storage I
l
25kg Paper Bag
Brick Manufacturing Process Slag
J
~[
Size: 1.0mm>
20%
Crushing ~-~1 Sieve
Grog ,.
Fret Mill
Matrix
Weighing[
3.0mm Mesh
~[ Crushing ~-*l Sieve
Ceramic Clay
Pigment
Fret Mill
1.0mm Mesh
Mix-MI 15 mind,
l ~l Weighing [ q B!endin ~ [ Molding 35%
Weighing Ceramic 25% Clay 20%
[Weighing [
~l Weighing[
l
,,,
Water& Binder
2%
Packing ~ ['H Final . . Product . . ~-~ Inspection ~-~ Sintering II Shrink Wrapping
~-
2OOtonF.P.
~._.._l
Tunnel Kiln 1200~176 about 80 hours
Figure 6 PavementBrick Manufacturing Process
[I Drying I~ 80~150~ abt.48 hours
39
5.3. Quality of the Products (1) Color The color is brownish, with delicate variances due to the location of the brick on the truck as it received different temperatures and oxygen of varied densities. Black dots appear on the surface caused by the oxides of metals remaining in the slag. Photograph 1 displays the appearance of the bricks as they are laid own on the floor.
Photograph 2
Pavement Brick Photograph 3
Pavement Brick
(2) Quality Table 6 shows the physical properties. These are the average values derived from 10 samples drawn at random from the products. The water absorption ratio at 4.1% and the compression strength at 1,278kg/m 2 meet the JIS standard of <13% and >200kg/m 2, respectively.
Table 6
Physical Properties of Pavement Brick Item
Unit
..A.p.parent Pore Ratio
%
Avcrag e 9.4
.......
! i i
Standard Deviation 1.06
~
i JIS R 1250 Standard i Brick No.3 i :!
i Ordinary ~ Pavement Brick i
-
t
-
Water Absorp.t.!o.n..Rate.................%...............:..5:.1........i .............9...5.............[.............!e.s.s..t..h.an...1..3..............i .....................-...................... 9
Apparent Specific Gravity
-
2.5
~
0.027
i
-
i
2.35
Bulk Sp..ecific Gravity.
-
2.26
i
0.018
i
-
i
2.29
......................................................
....C.omp.res.s.i.o.n...S.t.r.engt..h. Bending Strength
" .....................
.....k.~cm 2
.1.,..2..7..8.. ......i 134
!
; ..................................................
128.1
i
21.7
i
T ............................................
more than 200
'
-
-
!
-
mm
226.5 i
0.78
i
Standard Size
i
230 m m
i Width
mm
112.51
0.58
i
Standard Size
9
Measurement
T ...............................
i Length
9
kg/cm 2
~' .....................
~
i
:
114 m m
40
(3) Leaching Test In order The The
to determine
the safety
of the brick,
test specimens
were
bricks
test methods
were:
Publication
three
pH Method(controlled Table
to pH4
7 shows
Table 7
crushed
using
the leaching
elution
to sizes
tests have
smaller
No13,No46
than
been
carried
5mm,
out.
and the whole
of the Environmental
Agency
bricks. and
Low
HNO3).
test results.
Leaching Test Results of Pavement Bricks
Specimen
Crushed
Whole
Environmental Soil
' i t e m ............. U n i t ............. Nol3" ................ No4"6 .......... "pH4iHN'O3i ........... N o l 3 ................. No46" ........... p'H'4iH'N'O3i"
Standard
.~..~ .......................................... .7:.3 ..................... 7:.8 ....................... 4..1 ........................ 6..4. .................... .6:.4 ....................... .4..2 ......................... .-. ............ ..~-..rI.z. ........... m . ~ ................. - .................. .<.9:.000..5. .......... ..<.0:0.09.5. ..................... - ................ ..<.0:00.9.5............... .<..o:.ooo..5.................. ..S..one ........ ..T-.~.~ ........... m . ~ .......... .<..0.-0.00.5............ .<.9..000..5. .......... ..<.0.00.0.5............... .<.0.00.0.5. ........... ..<.0:00.0.5............... .<..0-.000..5.............. ..<.0:000.5.. ...... ....C.~................. m . ~ .......... .<..0.0.1................ ..<.0..01. ............... ..<.0.01 ................... ..<.0.0.1................. ..<.0.01 ................... .<..0..0.1................... .<.0.0.1............. ..g.b................... m . ~ .......... ..<.0.0.1................. .<..0..0.1................ ..<.0.0.1.................... ..<.0.0.1................. ..<.0.01 ................... .<..0..0.1................... ..<.0.0.1............. ..O.-~ ............. . m . ~ .................. - ................. .<.0.0.2................. ..<.0.0.2............................ .-. ................ ..<.0.0.2.................... .<.0.0..2........................ ..S..o.ne........ ....C.r6..+............... m . ~ ........... .<..0..0..5................ ..<..0..0.5................ .<.0.0.5.................... ..<.9.:0..5................. ..<.0.0.5................... .<..0..0.5.................... ..<.0.0..5............. . . ~ . ................ m . ~ ........... .<..0..0.05............. ..<.0..00..5.............. ..<.0.0.0..5................. .<.0:0.0.5............... ..<.0:00..5................ .<..0.90.5................. ..<.0.01 ............ ..T-...C...S............ . m . ~ .................. - ................. .<.0.0.2................. .<.0.0.2.. ........................... -. ............... ..<.0.0.2.................... <.0.0..2. ....................... ..S..o.ne....... ..~...C.~............. . m . ~ ................. - .................. .<..0..000..5........... .<.0.000.5.. ...................................... ..<.0.00.0.5.. ............ <..0..000.5. .................. ..S..o.ne....... ..S.e ................. . m . ~ .......... .<..0..0.0.5............... .<.9..00.5.............. ..<.0.09..5................. .<.0.0.0.5............... ..<.0.00..5................. <..0.90.5. ................ ..<.0.0.1.............
CI
mg/1
6.
Conclusion With
slag.
this report,
Further
by the recycling
0.2
study
industry society.
0.6
we have
covered
on the control,
as well
as the
1.2
0.9
the examples
disposal
governments,
and
1.4
of effective
re-use
re-use
of this material
a s it is t h e
key
to the
0.8
-
of incineration should realization
residue
be undertaken of resource-
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction:PuttingTheory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
41
I n v e s t i g a t i o n o f S i n t e r i n g P r o c e s s e s in B o t t o m A s h to P r o m o t e the R e u s e in Civil C o n s t r u c t i o n (Part 1) - E l e m e n t B a l a n c e a n d L e a c h i n g A. Selinger 1), V. Schmidt 1), B. Bergfeldt 2), J. Vehlow 2) and F.-G. Simon l) 1)ABB Corporate Research Ltd., CH-5405 Baden, Switzerland; 2~Forschungszentrum Karlsruhe GmbH, P.O. Box 3640, D-76021 Karlsruhe, Germany
Abstract Two types of bottom ash from MSWI, wet and dry discharged, were annealed in a temperature range from 700 ~ to 1065 ~ under reducing and oxidizing conditions, using a small rotary kiln. The products of these experiments were characterized in respect to their composition and their performance in Swiss and German leaching tests. The results lead to an improved understanding of the ash formation process on the incineration grate. Increasing ash temperatures on the grate result in a better ash burnout, transfer of some components to the flue gas and reduced leaching in the German DEV-S4 test. However, improvements in the Swiss leaching test are not to be expected.
1 Introduction Bottom ash from municipal solid waste incineration (MSWI) has proven to be a suitable civil construction material in many applications [1-3]. Recently, even a quality certificate was established in Germany [4]. However, bottom ash is by no means a chemically inert material, and its composition and properties may vary in a wide range [5]. To prevent any contamination of ground and surface waters, an ash material usually has to pass a standard leaching test prior to its use. Different opinions on the long term risks of such ash applications unfortunately led to a great variety of tests and leaching limits in the various countries [6]. In an overall view, the most critical ash components are heavy metals, namely lead, zinc and copper, further chloride and organic carbon. Besides the composition of the municipal waste itself, four major processes define the final quality of the bottom ash: - the incineration process - the ash discharge (dry or wet) - treatment of the ash (sorting) - aging
42
The benefits of aging [7, 8] are established and therefore many regulations require a storage of the ash for three month or so [9-11 ]. Ash discharge and treatment is discussed somewhere else [12,13]. Concerning the incineration process, little quantitative work has been published on the ash properties as a function of the ash temperature during incineration [14, 15]. This is not surprising, as mass burning of waste, i. e. on a grate system, is a quite variable process due to permanent changes of the waste fuel. In addition, the temperature profile of a given ash volume is very difficult to track. In an effort to fill this gap, a rotary kiln was used to treat bottom ash at different temperatures similar to conditions present on the incineration grate. Annealing at rising temperatures should lead to microscopic and macroscopic sintering of the ash which is reflected in changes of the mineralogical composition and the leaching behavior. "Part 1" of this work discusses the experiment and the effect of the annealing temperature on the composition and the leaching behavior in the German standard leaching test DEV-S4 and the Swiss TVA test. In "Part 2", the influence of those sintering processes on the mineralogy, alkalinity and availability of harmful elements are described.
2 Experimental Setup The experiments were performed in 1996 during two weeks, kindly hosted by a German incineration plant. A total of 30 ash samples of some 15 kg each were treated.
2.1 A s h material Two types of bottom ash were used for the annealing and sintering experiments. Both originated from combustion of municipal solid waste on a modern grate system. The first material was sampled from a German plant with a standard wet ash discharge system (chain conveyor). It will be referred to as "wet ash". The second material was produced at a pilot installation in Switzerland, where ABB demonstrated the dry extraction of bottom ash as part of the InRec T M process. It will be referred to as "dry ash". Ash particles of larger than 32 mm diameter were not considered. The ash was not further pre-treated, except that it was carefully mixed to get similar samples for each experiment. Both ashes were not allowed to age. The wet ash was taken just before the experiments, while the dry ash does not age due to the lack of water.
2.2 Rotary kiln Ash on an incineration grate passes a temperature profile, while being transported and slowly mixed by the movement of the grate bars. To offer a somewhat equivalent environment, a rotary kiln was used for the experiments. If not agitated, the ash would bake at temperatures below 1000 ~ as seen in earlier lab experiments. This is not representative for the real process and can be prevented by moving the ash slowly. We used an electrically heated laboratory rotary kiln of Mannesmann Demag (formerly PLEQ). The model HT 11 had an inner diameter of 0.4 m and a volume of 0.11 m 3, was rotated twice per minute and could be heated up to 1100 ~ The kiln was charged and
43 discharged at experiment temperatures and could accommodate up to 20 kg of material. Gas mixtures could be added into the hot zone. Treatment time and rotating speed were kept low enough to prevent substantial grinding of the material.
2.3 Treatment conditions 2.3.1 Temperature The aim of the study was to investigate changes of the bottom ash properties below the melting temperature. Thus, melting of the ash was not desired. Therefore, the temperature was increased until partial melting led to a rapid growing of the particles. The particle size distribution plot in Fig. 1 shows, that 50% of the wet ash had formed large particles after treatment at 1065 ~ Both materials were treated at temperatures from below 700 ~ to 1065 ~ with emphasis of the region between 900 ~ and 1000 ~
100% ,-
90% .-~--Wet Ash
80% t~
=
__._, = , o o o
70%
I"/
oc
-//
_. T = , O , , o c
60%
/ /
50% t~
40% 30%
o9 2o% ~.
10% 0%
_-0.01
0.1
-
~ 1
10
100
Particle size in m m
Fig. 1
Ash is forming larger agglomerates at 1000 ~ and above
The temperature was measured in the middle of the kiln and verified by measurements directly in the ash and by adding probes with silver and copper shots (melting points of 962 ~ and 1085 ~ respectively). The given temperatures are accurate within +20 ~
2.3.2 Time Each experiment lasted approx. 1 hour, thus the material was kept some 30 min at the desired temperature. This corresponds roughly to the time ash will be at the temperature maximum on the grate. Fig. 2 displays the temperature profile of a wet and a dry sample.
44
1000
900
O
800
o
700 .... ,__ 9
~
/
600 500 -20
~
~
I
0
20
40
Time
Fig. 2
60
in m i n
Temperaturein the rotary kiln during sample treatment
2.3.3 Atmosphere It was seen, that the gas composition of the kiln atmosphere had great influence on the annealing and sintering process. To keep as many parameters constant as possible, only two concepts were applied: The "wet ash" was kept in its own atmosphere. The water vapor displaced the air from the kiln, so that a reducing wet atmosphere resulted due to reducing ash components like carbon and aluminum. Carbon monoxide (CO) was formed, which was not burned until the kiln was discharged. The "dry ash" was treated in an oxidizing gas mixture of 60% air, 30% H 2 0 and 10% CO2 (percentage by volume), according to an estimation of the incinerator gas composition after the main burning zone.
3 Composition and Element Balance 3.1 Composition o f the untreated ashes The following table summarizes the composition of the two ashes, compared to average values from the literature [5] and our own results. The particle size distribution of the wet ash was already shown in Fig. 1. The dry ash distribution is very similar.
45
Table 1
Ash composition =.
Ash
material
Wet
Ash
!
Dry Ash
Literature
Ave.
Water C o n t e n t
17.4 %
9
0 %
15 - 20 %
Melting Temp.
1205 ~
~
1200 ~
1150 - 1200 ~
All values given in mg/g Total Carbon (TC)
16.4
12.9
Organic Carbon (TOC)
11.0
9.3
Lead (Pb)
1.5
1.89
Zinc (Zn)
2.7
3.2
Copper (Cu)
3.46
4.83
28 i
10
i
~
2
.
4.7
i
-
2.1
Chromium (Cr)
0.253
0.197
~
1.2
Nickel (Ni)
0.19
0.097
i
0.21
Cadmium (Cd)
0.079
0.037
i
0.021
Calcium (Ca)
93.0
130.0
":
76.8
Potassium (K)
4.73
6.07
Sodium (Na)
7.89
10.63
i
23.4
Chloride (CI)
6.41
4.9
Sulfate (SO42-)
8.11
"
9.6
2.8
11.19
~
10
3.2 Element Balance after Treatment After thermal treatment, a weight loss is observed. For the dry ash, this is mainly due to organic carbon and carbonate destruction plus evaporation of volatile compounds. The same applies to wet ash, in addition to the loss of the water content. For the highest treatment temperatures, the weight loss value was an average 2% higher than for the lowest ones.
Total Carbon (TC)
1 1 1 1 0
,
0
g"; ~P
0 0 0 500
600
700
800
Temperature I
900
1000
~
Fig. 4 Total carbon content vs. treatment temperature. TC o f untreated ash is plotted on the ordinate ("500 ~
1100
46
Carbon In Figure 4, the total carbon content of 19 treated samples is summarized. Up to 700 ~ mainly the organic carbon (TOC) is destroyed while above 800 ~ also the carbonates are decomposed. Oxidizing conditions (dry ash) yielded the lowest TOC values of less than 0.05% above 800 ~
Metals Lead, cadmium and potassium contents decreased by some 20% in the tested temperature range. Those metals are found enriched in fly ashes due to the higher vapor pressure of their compounds (Chlorides). For the other investigated metals (see Fig. 3) no pronounced change was found. Some heavy metals even showed somewhat higher values after annealing, which we explain by an improved availability of the new phases in the analytical procedure (grinding plus digestion).
Anions Above 800 ~ about half of the Sulfate content was destroyed, with or without oxidizing conditions. Chloride content, in contrast, remained constant at reducing conditions (wet ash), while decreasing by some 30% if air was added at high temperatures.
4 L e a c h i n g Behavior 4.1 German Standard Leaching Test DEV-S4 In this leaching test, as described in DIN 31414-$4, the sample is agitated with a water/sample ratio of 10 for 24 hours. Thus, the pH value of the leachate is controlled by the amount of available and soluble alkali, mainly calcium oxide, in the sample. As a consequence, the leachate concentration of a heavy metal is rather independent from its total content. More important is the availability of the metal, which is influenced by the particle surface area, the mineralogical environment of the metal and the mechanical stability of the particles. However, the most important factor is the pH value of the leachate, as the solubility of many metals, namely the heavy metals, depends strongly on the pH. If the pH value is lower than 12, the formation of soluble lead hydroxo complexes is usually low enough to pass, i. e. the requirements for reuse according to the German LAGA Z2 class [8, 9].
pH Value and Conductivity For both types of ash, the pH was lowered by an average of 1.2 units after treatment, independent from the actual treatment temperature. With a few exceptions, the pH values resulting from the annealed samples were found to be below 11.5, which led to the expected low lead and zinc concentrations. Obviously, the decomposition product of CaCO3 is not free CaO, but is directly integrated in less soluble mineral matrices (see part 2). The electrolytic conductivity of the leachates is basically caused by the anions hydroxide, chloride and sulfate and is therefore directly correlated to the pH value.
47 Table 2
Leaching in the DEV-S4 Test (Summary) Wet Ash
Treatment
>900 ~
<900 ~
[ raw
pH
Dry Ash
12.4 [
11.0
i
1.8
i
1.2
11.2
i raw
LAGA Z2
<900 ~
12.7 i
>900 ~
(Re-use)
11.6
13
11.6
i
9.9 ..-[
1.7
[
1.8
6
0.21 [
0.08
[
0.02
0.05
Conduct.
mS/cm
5.9 [_.
Pb
mg/1
1.08 !
0.03
i
0.01
Zn
mg/1
4.24 [
0.003 i
0.01
9.65 [
0.04
i
0.11
0.30
Cu
mg/1
0.43[
o.oo3 i
0.004
1.81 i
0.005 i
0.067
0.3
mg/1
0.03 i
1.20
0.45
0.07 i.:
0.29
0.01
0.3
C1
mg/1
351 [
231
i
82
261 i
66
250
SO4
mg/1 ....
152 i
314
i
170
87
36
600
Cr
....
.:
i~ :"
108
i
208
i.:
:
!i
i: 9
Heavy Metals Annealing of both types of ash at any temperature significantly reduced the leaching of the main heavy metals lead, zinc and copper. Except for one lead value, even the low LAGA Z2 leaching limits are met - without aging of the ashes. As copper, different from lead and zinc, is not sensitive to high pH-values, the reason for the low leaching after treatment will be the formation of new stable phases. Lead leaching values showed a continuous decreasing trend towards high temperatures, which the others, surprisingly, did not. Chromium showed a different behavior. When there was enough oxygen available, the strongly bound chromate(III) was partially oxidized to chromate(VI), which is soluble and more toxic than its precursor.
Anions Chloride, while not significantly evaporated, showed nevertheless a rapid decreasing solubility with increasing treatment temperature. In contrast, sulfate became less soluble with air, but quite soluble under reducing conditions (not distinguished in Table 2), despite the decomposition of up to 60% of the sulfate (see 4.2). Again, the chemical context seems more important than the total content. 4.2 Swiss TVA L e a c h i n g Test The Swiss Leaching Test, according to the "Technische Verordmmg Abf~ille", is performed with carbon dioxide saturated water. At the resulting pH value of about 6, the solubility of lead (-carbonate) is very low, but zinc carbonate is soluble. Annealing at any temperature had no definitive effect in respect to the Swiss test. The leaching values varied statistically, but no trend was found. Only for nickel, an increased elution was found. Summary table 3 gives an impression of the achieved values.
48
Table 3
Summary of leaching in the Sw&s TVA leaching test .:
CH-TVA Tests pH
:
...................................................................
Conduct.
.'.
[
i Untreated Ashi Treated Ash~ CH-TVA Inert 6.06
,. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
mS/cm i
2.77
i
6.04
~.................................
. ...........................................
~
"
2.29
Pb mg/1 ! 0.015 i 0.005 i O. 1 ...................................................................i......................................~.................................~........................................... Zn mg/l i 3.09 i 4.72 i 1 ..................................................................i......................................!.................................!........................................... Cu mg/1 i 0.21 ~ 0.20 ~ 0.2 Cr
mg/l
Ni
mg/l
~
1.96
~., .
i 9
0.09
1.88
.:
0.05
~.. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
~...........................................
~
~
,
0.32
i
,
0.2
5 Conclusions Although ash on an incineration grate may presently reach temperatures as high as applied in our rotary kiln experiments, this is never true for the whole mass stream. We showed, that annealing changes the ash properties already at temperatures around 700 ~ with a significant step at around 900 ~ N e w compounds are formed, which feature a higher leaching stability in a basic environment (DEV-S4). Improved burnout and mass transfer to the flue gas are additional advantages, if a homogenous high temperature can be achieved for the bottom ash. We recommend to aim at an ash temperature of at least 800 ~ better 900 ~ for some 20 min. Excess oxygen must be avoided to prevent chromate elution. Good mixing of the ash on the grate is mandatory. Improvements in the Swiss leaching tests are only to be expected for ash heated up to the melting point. However, an ash quality which is sufficient for construction purposes can be achieved at much lower temperature.
References [1] Mesters, K., Untersuchungen zur Mobilisierung von leichtl6slichen Salzen aus MVAsche. VGB Kraftwerkstechnik 73(12), 1058 (1993) [2] Schoppmeier, W., Mechanische Aufbereitung von Schlacke aus Mt~llverbrennungsanlagen mit dem Schwerpunkt Schrott. VGB Kraftwerkstechnik 73(12), 1055- 1057 (1993) [3] Schoppmeier, W., Gezielte Aufbereitung fester Verbrennungs~ckst~xlde aus der Sicht eines privaten, fiberregionalen Schlackeaufbereiters und -verwerters. M#ll und Abfall (Supplement) 31, 117- 123 (1994) [4] RAL, Gtite- und Prfifbestimmungen f~r Mtillverbrennungsaschen (MV-Asche). RAL, Deutsches Institut ffir Gtitesicherung und Kennzeichnung e. V., Gtite- und Prfifbestimmungen, RAL-RG 501/3, Sankt Augustin (1996) [5] Faulstich, M., Schmelzen von Rfickst~inden aus der Mfillverbrennung -Integrieren oder Nachschalten?-, Reaktoren zur thermischen Abfallbehandlung, Thom6-Kozmiensky, K.J. (Editor), EF-Verlag fi~r Energie und Umwelttechnik GmbH, Berlin, 175-188 (1993)
49 [6] IAWG, An International Perspective on Characterisation and Management of Residues from Municipal Solid Waste Incineration. International Ash Working Group, Final Document (1995) [7] Lahl, U., Schlackeaufbereitung durch Alterung und Laugung. M~ll und Abfall (Supplement) 31, 86 - 91 (1994) [8] Simon, F.G., Schmidt, V. and Carcer, B., Alterungsverhalten von MVA-SchlackerL M~ll und Abfall(11 ), 95 (1995) [9] LAGA, Anforderungen an die stoffliche Verwertung yon mineralischen Reststoffen/Abfdllen-Technische Regeln-. Mitteilungen der LAGA, L~inderarbeitsgemeinschaft Abfall, Erich Schmidt Verlag, Berlin (1994) [10] Minister of Environment (F), Disposal of bottom ash from municipal solid waste incineration. Board for pollution and risc prevention, Circulaire avec annexes, DPPR/SEI/BPSIED/FC/FC No. 94-IV- 1, Paris (1994) [ 11 ]
Schweizerischer Bundesrat, Technische Verordnung ftir Abf~ille (TVA) (1990)
[12] Simon, F.G. and Andersson, K.H., InRec T M process for recovering materials from solid waste incineration residues. ABB Review(9), 15-20 (1995) [13] Selinger, A., The ABB dry ash concept: InRec TM, WASCON "97, Maastricht, The Netherlands, ISCOWA (Publisher) (1997) [14] Schneider, J., Vehlow, J. and Vogg, H., Improving the MSWI Bottom Ash Quality by Simple In-Plant Measures, Environmental Aspects of Construction with Waste Materials, Goumans, J.J.J.M., van der Sloot, H.A. and Aalbers, T.G. (Editors), Elsevier, Maastricht, 1. 3.6. 1994, 605 - 620 (1994) [15] N~iBlein,F., Wunsch, P., Rampp, F. and Kettrup, A., Influence of Combustion Bed Temperature on Concentration and Leachability of Metals in Slags form an Incinerating Plant. Chemosphere 28(2), 349 - 356 (1994)
This Page Intentionally Left Blank
Goumans/Senden/vander Sloot, Editors Waste Materials in Construction:PuttingTheory into Practice 9 1997 Elsevier Science B.V. All rightsreserved.
51
Investigation of Sintering Processes in Bottom Ash to Promote the Reuse in Civil Construction (Part 2) - Long T e r m Behavior B. Bergfeldt 1), V. Schmidt 2), A. Selinger 2), H.
Seifert 1), and J. Vehlow 1)
1)Forschungszentrum Karlsruhe GmbH, P.O.Box, 3640, D-76021 Karlsruhe, Germany, 2)ABB Corporate Research Ltd., CH-5405 Baden, Switzerland
Abstract Bottom ashes from two MSWI characterized by different discharge technologies were annealed at temperatures of about 700 ~ up to 1065 ~ under reducing and oxidizing atmosphere, respectively. All products were tested for their leaching stability. Investigations of the mineralogical composition were carried out additionally. The following main conclusion can be drawn from these experiments. Mineral composition is influenced by sintering with respect to the decomposition of Ca(OH)2 and CaCO3 at temperatures above 950 ~ and formation of less soluble Ca-silicates. This results in different pH values and acid neutralization capacity compared to the original materials. The effect of sintering on the availability and release of heavy metals under the applied conditions is weak.
1
Introduction
Former investigations carried out in our laboratory [Schneider 1994, Vehlow 1995] showed that sintering of bottom ash is a promising way to receive a construction material with leaching behavior required in the German technical guideline residential waste [TA Siedlungsabfall, Bundesministerium 1993]. In the recent investigation sintering experiments were conducted in a rotary kiln. The material used was either bottom ash from a municipal solid waste incinerator with a wet discharge system or material received from the InRec TMprOcess developed by ABB. The first material, named "wet ash", was annealed at different temperatures under a slightly reducing atmosphere while the second material, named "dry ash", was annealed under oxidizing atmosphere. "Part 1" of this work discussed the experiments and the effect on the quality of the products. In "Part 2" the influence of sintering processes on mineralogy, alkalinity and availability of harmful elements are summarized. These parameters help to evaluate the behavior of bottom ash with respect to long term leaching. Additionally the chemical reactions during sintering can be estimated by considering the variations of all those parameters.
2
Methods
2.1
Sample Preparation
Since all analytical methods applied use liquid samples and all leaching tests ask for different grain sizes all solid materials require some pretreatment. At first they were sub-
52
divided to an appropriate sample size using a riffle box. One subsample was sieved and the material > 3 mm crushed in a jawbreaker for the column tests. Another subsample was crushed and milled for analytical purposes and for the availability test.
2.2
Mineralogical Investigations
The ground samples were examined by X-ray diffractometry to identify the mineral composition. In two sintering tests samples were obtained with larger grainsizes than the original material. These larger pieces were used to get thin sections for microscopy.
2.3
Leaching Tests and Acid Neutralization Capacity (ANC)
The German DEV $4 test [DIN 38414] has been applied to judge the quality in accordance to the German regulations (see part 1). These data do not allow any conclusion in which way the sintering process influenced the leaching mechanism due to modifications of crystal structures and mineralogy. To obtain more detailed information about these parameters the Dutch test NEN 7341, which includes a column leaching test and availability test, was conducted. The alkalinity and the acid neutralization capacity (ANC) are also important factors which influence the leaching behavior. Additionally the titration curves of the bottom ash samples allow an estimation of the chemical species of some elements. For most of the samples a titration with HNO3 was carried out with an automated titration setup. Following the method described in Johnson et al. (1995) a liquid to solid ratio of 100 was chosen in order to avoid the supersaturation of Ca with respect to gypsum. The titration was carried out under nitrogen gas to prevent CO2 entering the system.
2.4
Digestion and Analytical Method
For the analyses of heavy metals about 100-300 mg of each ground sample was digested using a HNO3/HC1/HF mixture in a teflon bomb heated in a microwave oven. The halogenides were extracted by superheated steam in a glass set-up. Total Reflecting X-Ray Fluorescence Analysis (TRFA) was used for metal analysis. The anions were analyzed by Ion Chromatography (IC).
3
Test Results
3.1
Mineralogy Table 1 depicts the phase composition of some samples analyzed by X-Ray diffracto-
metry. Quartz is the main constituent in all samples. The mineralogical composition of the original materials varies mainly relating to the calcite content - in the materials "wet ash 1" and "dry ash" calcite exists only as an accessory constituent. The effect of sintering can be seen by the decreasing of the calcite content. In the samples "wet ash" which had been annealed by higher temperatures - above 850 ~ - calcite is present in lower concentrations than in the other ones. In most samples derived from original materials "wet ash 1" and "dry ash" it could not be determined by X-ray diffractometry. These results are related to the TIC values mentioned in part 1 of this work which proof the destruction of carbonates at temperatures of about 900 ~ But even at high temperatures in the range of 1000 to 1065 ~ - the destruction of calcite is not complete, because in the thin section prepared of samples "wet ash 1, 1000~ '' and "wet ash 1, 1065~ '' calcite can be found in the matrix as a minor component.
53
Table 1"
Phase composition of some original materials and sinter products
Sample
wet ash
Temperature ~
wet ash 1
untr. 950 a
i
Quartz
xxx
untr. 700
955
1065
XXX
XXX
XXX
XXX
XXX
XXX
XXX
XXX
X
X
X
X
XX
XX
~XX
i
xxx
dry ash
950 I
1065 1
untr. 680 I
I
1
SiO2 9
|
|
Magnetite Fe304
x
xx
x
XX
X
Gehlenite Ca2A12SiO7 Akermannite CazMgSi207
xx
xx
x
XXX
XX
i
Calcite CaCO3
'xx
x
Dolomite CaMg(CO3)2 Feldspar K[A1Si308]
XXX
X
XX
XXX
X
X
X
X
X
X
X
X
X
(x)
x
xx x
x
X
X
Na[A1Si308] Ca[AlzSi208]
i i
Diopside Ca(Mg,Fe)[Si206]
ix
X
Anhydrite CaSO4
(x)
(x)
x
untr. = untreated, xxx constituent (5-10 %), x = ac9 = main constituent ' (>10 . %), xx . = minor . . . . cessories (<5 %), (x) = traces The formation of gehlenite (CazAI[(SiA1)207] and akermannite Ca2Mg[Si207] gives evidence of mineral reactions, too. These minerals can be found in each sample even in the original material, which itself is a product of combustion in the formation temperature range of these minerals, 700-800 ~ (Pfrang-Stotz, 1995). While the X-ray-diffractometry shows only a slight increase of gehlenite and akermannite the thin sections indicate an increase with increasing temperatures. 3.2
Leaching Test Results
3.2.1 Acid Neutralization Capacity Alkalinity and acid neutralization capacity of bottom ashes depend mainly on the chemical form of the main constituents, i.e. alkali metals, earth alkali metals, silica, aluminum, and iron. Calcium, for example, occurs as hydroxide, carbonate, and bound in silicates. In leachates of bottom ashes the presence of calcium hydroxide and -silicates as well as of alkali metal hydroxides is responsible for pH values of about 11 up to 12.5. pH values of about 10 to 8 are mainly related to Ca CO3 [Johnson 1994]. The initial pH values measured after ten minutes and the pH obtained by means of the DEV $4 Test of the original materials and the sinter products are shown in table 2. The initial pH is a hint for the more soluble pH forming fraction. The original material "wet ash 1" and respective sinter products show pH values after 10 minutes in the range of 11.0- 11.6 while the initial pH values of "dry ash" and its sinter products range between 12.0- 12.4. This indicates that pH controlling components like Ca(OH)2 of the original materials are responsible for the pH in the sinter products as well. After equilibration the pH values
54 of the original materials and of their sinter products show more variation. All leachates remain in the pH level above 11.0 where Ca-hydroxides and -silicates are responsible for the pH. The equilibrated leachates of original materials are saturated with respect to Ca(OH)2. The leachates of the sinter products of "wet ash 1" still remain undersaturated with respect to this phase. Since Ca(OH)2 was saturated in the leachates of the "dry ash" sinter products after ten minutes, but not after equilibration, some precipitation reactions must have taken place without solution of additional hydroxides. It can be supposed that during annealing parts of the Ca(OH)2 have been decomposed and components with a more "acid" character have been formed, for example CaCO3 or less soluble Ca-silicates. Which reactions have taken place can be investigated by evaluation of the acid titration curves and the acid neutralization capacity. Figure 1 compiles titration curves of the original materials "wet ash 1" and "dry ash" and respective products annealed at 1065 ~ Table 2"
Initial pH and pH after 24 hours (DEV $4 Test) and ANC at pH 4 (in meq/g)
wet ash
untreated
700 ~
850 ~
915~
950 ~
1000 ~
1065 ~
initial pH
11.4
11.0
11.2
11.4
11.7
11.5
11.6
pH after 12.4 24 h
11.4
11.0
11.0
11.2
11.0
11.6
ANC
2.164
2.244
1.810
1.610
1.982
1.823
1.560
dry ash
untreated
initial pH
12.4
12.4
12.1
pH after 12.7 24h
11.5
ANC
3.907
790 ~
955 ~
965 ~
1000 ~
1065~
(02, COz)
(02, CO2)
(02, CO2)
12.4
12.0
12.3
11.4
12.3
11.2
11.5
4.090
3.917
2.799
3.155
(02, CO2)
3.782
The "dry ash" materials have a higher acid consumption than the "wet ash 1" materials. The slope of the original materials" curves varies only slightly with decreasing pH down to pH 5. The constituents which control the respective pH seem to exist in similar amounts in both untreated samples. The slope of the curves of the materials annealed at temperaures higher than 1000 ~ is steeper up to a pH of about 5. This fact points out that Ca(OH)2 had disappeared but no carbonate had been formed. At these temperatures Ca has obviously been used up for the formation of silicates like gehlenite, akermannite, and diopside. Another value derived from the titration curves is the ANC which is depicted in table 2, too. This is an important parameter for the estimation of long term leaching behavior. There is a slight decomposition of buffering substances in the sinter products of "wet ash 1". A temperature dependence in this case can not be deduced. The sinter products of the "dry ash" obtained at low temperatures show a similar or even slightly increased ANC whereas products of higher temperatures are characterized by lower ANC. The raise of ANC due to annealing processes can be explained by the artificial atmosphere with an excess of oxygen and carbon dioxide. Carbonates are more stable due to higher partial pressure of CO2 which raises the decomposition temperatures of carbonates [Warne, 1991]. Moreover, the oxidizing atmosphere favors the formation of metal oxides. The raise of ANC can be observed
55 at a pH range of 5-3 where neutralization by metal hydroxides (AI(OH)3, Fe(OH)3) is effective [Scheffer, 1989].
12,0 10,0
-~- "wet "wet ~ "dry . . .
~'x.
8,0
1" untreated" 1" 1065 ~
ash ash ash" .
untreated
1000 gC (O2,CO2)
=ca. 6,0 4,0 2,0 0,0 0,000
I
I
I
I
I
1,000
2,000
3,000
4,000
5,000
6,000
ANC in m e q / g
Fig.l" Titration curves of original material from "wet ash 1" and "dry ash" and their products annealed at 1065 ~ and 1000 ~
3.2.2 Leaching Test and Availability Column leaching tests provide information about the time dependent leaching behavior due to varying liquid to solid ratios (0.1 - 10). The pH of the eluent (pH 4) is quickly neutralized in the leachate due to reaction with the solid material (grain size > 3mm). Leaching under these circumstances is evidently controlled by diffusion and sorption processes.
I--e-Untreated -*-700 *C -e- 950 oC -,-1065 oC i
10000
9 _
1000 0 -I
2 0
~
9
9
IConcent-.'onl =1
I Availabilit~
100
r
A --
I
r 9
r 9
[]
I = = = ==t=== = = = 1 " = = = = = = = q
10
leluti~ I ~
N
=
0,1 0,01
I 0,1
1
1(
LIS
Fig.2: Total concentration, availability, and elution of Zn of "wet ash 1" and sinter products
56
To evaluate the leaching behavior, which is only controlled by chemical speciation, the Dutch availability test was developed. The test models severe environmental conditions by establishing a L/S 100 and a pH of 4. The release during time and the highest available - but in most cases never reached - amounts of harmful species can be estimated by combining both leaching tests according to NEN 7341. Figure 2 compiles the elution, availability, and the total concentration of Zn in the "wet ash 1" and in some of the sinter products. The slope of the elution curves of the original material and the material sintered at 700 ~ correspond closely though the release of the annealed material is lower. But it seems that leaching of Zn is not complete at a L/S of 10. At contrary the samples annealed at 950 ~ and 1065 ~ show no further release of Zn after L/S = 1. These similarities between the slopes are not true for other heavy metals, but with the exception of Ni it can be taken as fact that the release from original material is higher than in any annealed material.
100
I
9 untreated
[] 700 ~
[] 950 ~
10
[] 1065 ~
I
1
la) "wet ash 1"
0,1 0,01 0,001 0,0001
. 0
.
.
'7"
.
.
I--
.
i
N
0
availabilty
,,
a.
100
I b,, U.
0
[I...
,
,
,
.~
0 N release (IJS = 10)
9untreated
[] 790 gC
,1
10 --
~I~J
-
-Ib)"dryash"l-
0,1 0,01 0,001 0,0001
v
o
i
| i.i..
i
~
availabilty
i
= o
i
= N
i
-, a.
, i
, i
m 0
i
| ~
i
~
i
",
0
|
i
,-
N
release (L/S = 10)
i
,,
I~.
Fig.3" Availability and release of some elements normalized to total concentration a) "wet ash 1", b) "dry ash"
57
Differences in the availability are not noticeable. Results of former experiments carried out in a laboratory oven without motion or control of atmosphere show apparently greater effects caused by annealing of bottom ashes (Schneider, 1994). They show decrease of the availability between the original and the annealed (1000 ~ material in a range of a factor of 10. In order to compare availability and release at L/S 10 of the lithophilic elements Ca, Fe, Ni and the volatile constituents Cu, Zn, and Pb as an effect of different annealing temperatures figure 3 and 4 compile these values. The differences of total concentrations between the samples are taken into account by normalizing the availability and the release to the corresponding total concentration. The six elements show little modified availability due to the annealing of the "wet ash 1" (fig. 3a). Only Ni shows a decreasing availability after treatment at 1065 ~ The availability of Fe is increased. The effect of annealing on the release at L/S 10 in these samples for the lithophilic elements is not very strong, either. The finding for Ni, which is only detectable in the leachate of the product annealed at 1065 ~ can be explained by the fact that the Ni concentrations in all leachates are near the detection limit of the TRFA. Cu, Zn, and Pb are less leachable after treatment. The explanation for Pb can be found in the lower pH range of the elutes. The lack of differences in the availability show that a incorporation into low soluble compounds did not take place during the annealing process. The difference in release at L/S 10 elucidates the formation of sulfides in the slightly reducing atmosphere at the surface. These compounds are insoluble at this L/S and pH (at about 11) prevent the heavy metals from being leached. The availability of Ca and Ni in the sinter products of the "dry ash" is nearly the same as in the untreated material (fig 3b). The availability of Cu, Zn, and Pb is slightly more influenced in the "dry ash" sinter products than in the "wet ash 1" products. But again this effect is not very significant. However, the release of Zn and Pb at L/S 10 is less decreased in the "dry ash" sinter products. This can be explained by the slightly higher pH at about pH 11.5 [Schneider 1994]. The release of Ca and Fe is even increased after annealing of "dry ash" due to formation of oxides on the surface in the oxidizing atmosphere. Annealing of "dry ash" was most effective for release of Cu and Ni. Cu could not be measured in column leachates after thermal treatment, the release of Ni decreases more than 10 % of the release in the untreated "dry ash".
4
Evaluation and Conclusions
In order to compare former investigations of the influence of thermal treatment on the quality and leaching behavior of bottom ashes, carried out in a laboratory oven, to a more realistic scale, with regard to full scale combustion chambers, experiments in a rotary kiln were conducted. Wet bottom ash from an MSWI and dry bottom ash derived from the InRec process was annealed at different temperatures without controlling of atmosphere and under additional O2 and CO2, respectively. One goal of this investigation was to find out the best conditions for combustion of waste. The effect of annealing in the rotary kiln can be summarized as follows: 9 Analyses of the mineralogical composition proof the formation of Ca-silicates and the decomposition of calcite at high temperatures. 9 The pH value found in leachates of the DEV $4 Test is slightly decreased after annealing. 9 The ANC after thermal treatment is slightly decreased in products of the "wet ash" and similar to the original material or even higher in products of the "dry ash". 9 Considering pH and ANC decomposition of Ca(OH)2 and formation of carbonates (under artificial atmosphere) up to a temperature of 900 ~ and Ca-silicates, respectively, has to be supposed.
58 9 The influence of sintering on the availability is insignificant. 9 Annealed samples show a decreased release of elements during column tests. Compared to the former laboratory tests the effect on the release is less in products of the rotary kiln sintering. Evaluation of these results requires that additional modifications of temperatures or atmosphere by supply of oxygen or CO2 during combustion of waste are not economically due to their weak effect. The focus for running a MSWI plant should be directed towards a sufficient residence time of the waste at the back end of the grate at temperatures high enough to guarantee a full burnout and some sintering effects on the bottom ash. This should result in a leaching stability sufficient to the German law for deposition.
5
References
Bundesministerium ftir Umwelt, Naturschutz und Reaktorsicherheit (1993), Dritte Allgemeine Verwaltungsvorschrift zum Abfallgesetz (TA Siedlungsabfall), Bundesanzeiger Jahrgang 45, Nr 99a. Johnson, C.A (1994), Das Langzeitverhalten von Mtillschlacke im Hinblick auf die Mobilit~it von Schwermetallen und Salzen.- Entsorgung von Schlacken und sonstigen Reststoffen, Beihefte zu MUll und Abfall 31, 92-95. Johnson, C.A., Brandenberger, S., and Baccini, P. (1995), Acid Neutralizing Capacity of Municipal Waste Incinerator Bottom Ash.- Environ. Sci. Technol., 29, 142-147. NEN 7341 (1993), Determination of leaching characteristics of inorganic components from granular (waste) materials. Netherlands Standardization Institute (NNI), Delft. Pfrang-Stotz, G. and Schneider, J. (1995), Comparative Studies of Waste Incineration Bottom Ashes from Various Grate and Firing Systems, Conducted with Respect to Mineralogical and Geochemical Methods of Examination. -Waste Management & Research, 13,273-292. Scheffer / Schachtschabel (1989), Lehrbuch der Bodenkunde, Stuttgart: Enke-Verlag, 118. Schneider, J., Vehlow, J., and Vogg, H. (1994), Improving the MSWI Bottom Ash Quality by Simple In-Plant Measures, Environmental Aspects of Construction with Waste Materials, (Goumans, J.J.J.M., v.d. Sloot, H.A., and Aalbers, Th.G., eds.) Amsterdam, London, New York, Tokyo: Elsevier, 605-620. Vehlow, J. (1995), Reststoffbehandlung - Schadstoffsenke "Thermische Abfallbehandlung"-,in Fakten- Die Thermische Abfallverwertung der Zukunft, (Fachverband Dampfkessel-, Beh~ilter- und Rohrleitungsbau e.V., FDBR ed.), DUsseldorf, 45-66. Warne, S.St.J. (1991), Variable Atmosphere Thermal Analysis - Methods, Gas Atmospheres and Applications to Geoscience Materials.- in Thermal Analyses in Geosciences, (SmykatzKloss, W. and Warne, S. St. J., eds.) Berlin Heidelberg, New York: Springer-Verlag, 62-83.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
THE ACID EXTRACTION
59
PROCESS
T.INOUE,Unitika LTD. Kyutaro-cho,Chuo-ku,Osaka,541 ,Japan H.KAWABATA,Kobe Steel LTD. IwayaNakamachi 4-chome,Nada, Hyogo,657,Japan Abstract Considering from a point of view of the recycling of resources that melting fly ash produced by melting the fly ash and incineration residue discharged from incinerators of municipal refuse is useful resources of concentrated heavy metals, this paper presents acid extraction processes developed for using the useful heavy metals and salts produced from the fly ash and melting fly ash as the resources. This paper reports the acid extraction processes from the fly ash in terms of the operational results of AES Processes operated at present and the problems to be solved in the future. This paper also reports the application of the acid extraction processes to the melting fly ash in terms of the description of the separating recovery process of heavy metals operated at present in a bench scale, the experimental results and the problems to be solved in the future.
1. INTRODUCTION Approximately 50 million tons of the municipal refuse is discharged every year in Japan and 78% of it is incinerated. The bottom ash and the fly ash are produced as the incineration residue in the incinerators and the fly ash containing low-boiling heavy metals are treated and landfilled by any of four techniques (melting-solidification, cementing-solidification, chemical stabilization and acid extraction) designated for "Specially Controlled Municipal Wastes" by "the Waste Disposal and Public Cleansing Law". In particular, the melting-solidification technique gains attention as a method for reducing the volume of the bottom ash and the fly ash, making them harmless and using them as resources. Melting fly ash mainly containing low-boiling heavy metals is, however, produced also by the melting-solidification technique. The fly ash and melting fly ash mainly containing harmful low-boiling heavy metals arouse an environmental problem by the landfill, while viewing from the standpoint of the circulation of resources, they are regarded as useful resources containing concentrated heavy metals.This paper presents two types of the acid extraction processes for alleviating the environmental load and utilizing the heavy metals and salts as the resources.
2. ACID EXTRACTION SULFIDE PROCESS OF FLY ASH (AES PROCESS) The description of the acid extraction and sulfide stabilization process (AES process) developed for stabilizing the heavy metals contained in the fly ash and for recovering the salts contained in it and the operational results of a commercial plant installed in O Municipal Incineration Plant are described hereafter.
60
Table 1.
Outline of Waste Incineration System at O municipal incineration plant Item
Description
Type of fumace ......
Full continuous combustion
Type of incinerator
Stoker
Incinerator capacity
115t/d • 2 furnaces
Gas cooling system
Boiler
Precipitation system
Dry type electric precipiter
Exhaust gas treating system
Wet type NaOH- scrubbing
The outline of the facilities installed in O Municipal Incineration Plant is listed in Table 1. The plant has been operated for approximately 12 years since the startup of the operation in 1986 and approximately 3 to 4 tons of electrostatic precipitator (EP) fly ash per day and approximately 20 to 30 tons of the waste water from the gas scrubbing process (scrubbing waste water)per day are produced by incinerating approximately 160 to 180 tons of municipal refuse per day. The EP fly ash and scrubbing waste water are treated together. Approximately 3 to 4 tons of EP fly ash is landfilled as a harmless cake. Approximately 2 to 3 tons of solid salts per day and approximately 18 to 27 m 3 of water per day are recovered by the evaporating crystallization of the scrubbing waste water treated by the waste water treatment method to reuse as the feed salt for the soda-production plant and the cooling water for the whole AES plant without discharging, respectively.
2.1 Outline of AES p r o c e s s The AES process is based on the following principle, 1)after the acid extraction of the fly ash, the mixture of the fly ash and scrubbing waste water is neutralized with NariS, 2) the heavy metals contained in the fly ash are converted into stabilized cake consisting of insoluble heavy metal sulfides, and 3) the salts and water are recovered from the filtrate. The flow sheet of AES process in O Municipal Incineration Plant is illustrated in Fig. 1. O Municipal Incineration Plant mainly comprises the EP ash-stabilizing equipment, waste water-treating equipment, salt-recovering equipment. The main stages of AES process are as follows. 1)Mixing stage: EP fly ash slurry is prepared by mixing scrubbing waste water with EP fly ash, from which soluble salts including KC1, NaC1 and CaC12 and a part of heavy metals are extracted. 2) Stabilization reaction stage: The pH value of the EP slurry sent from the previous state is adjusted to 6 with HC1 to extract the heavy metals. Then the pH value of the extract is adjusted to 8 by adding NariS to convert into heavy metal sulfides. 3)Dewatering stage: Stabilized slurry from the previous stage is coagulated with polymer flocculant and then is dewatered. Thus the dewatered cake is landfilled.
61
EP f l y ash I s t a b i I, z a t i on ! eClU i pment I
g
3 ~4t/d (Dry)
~" | ~ .
I.rl I,*
|
r
~
=~>
..nd~,,,i~
isl-~! 3~4t/d (water o o n t e n t :50 %)
20,,.,30m3/d
Wastewater
Salt
equipment
recovery
~ u i pment
~g
~
To sodium
~ t
i n d u s t r y as
2~3t/d
"
~
Recycling
as
cooling
water
18 ~ 27ma/d
Fig 1.
Flow diagram of fly ash and wastewater treatment processes at O Municipal Incineration Plant Table 2.
disolved salts (%)
Results of treatint~ b~r AES Process Leaching test Chemical analysis T-Hg Cd Pb T-Hg Cd Pb PCDDs/ PCDFs (mg/L) (mg/L) (ms/L) (rig/L) (me,~) (me/L) (me,/L)
Scrubbing wastewater Ep fly ash Stabilized cake
5.25
0.92
0.015
0.45
46.6
1.75
90
1600
Dewatered filtrate
8.42
0.013
1.0
2.2
5.0 0.65 ng/g 3.6 ng/g
0.018 <0.0005
7.5 0.014
27.0 <0.02
2.1
Treated water
5.64
<0.0005 <0.001
<0.02
<0.01
Recovered solid salt
98.9
<0.0005 <0.001
<0.02
<0.01ng/g
Impurities contained in the filtrate are removed by the waste water treatment method and the solid salts recovered from the water by evaporating crystallization are reused as the raw material for the soda production plant. Meanwhile, the whole evaporated water is condensed and reused as the cooling water for the salt-recovering plant and others.
62
2.2 Operational results An example of the operational results obtained from AES process in O Municipal Incineration Plant is listed in Table 2 and the following can be confirmed. 1)T-Hg in the scrubbing waste water is 0.92 mg/1 which exceeds the standard of 0.005 mg/1 or under. In addition, it contains 5.25% of soluble salts. 2)The EP fly ash contains 46.6% of soluble salts and high-concentration of heavy metals. The leaching concentration of the heavy metals including T-Hg, Cd and Pb are 0.018, 7.5 and 27.0 mg/1, respectively, which exceed the standard of respective heavy metals for landfill. 3)The contents ofT-Hg, Cd and Pb in the stabilized cake are below the standard for landfill of each heavy metal. Since those useful heavy metals are concentrated in it, reusing them as the resources are being studied. 4)The solid salt is pure white crystalline and the purity is approximately 99%. It is, therefore, used as the feed salt for the soda-production process because T-Fe as an impurity of heavy metal is contained as little as 3.92 mg/kg. 5)No PCDDs/PCDFs (dioxins) are contained in the solid salt for the soda-production process. The stabilized cake contains, however, as much as 3.6 ng/g, whereupon it must be treated by any method such as high-temperature dechlorination process for reusing as the resources.
3. SEPARATING RECOVERY PROCESS OF HEAVY METALS MELTING FLY ASH
FROM
The separating recovery process of heavy metals developed for separating and recovering useful heavy metals from the melting fly ash produced by melting the incineration residue is described hereafter.
Naris NaOH --1
Dust
S~176 (Melting) ~
i NariS , ! -I NaOH-.-q
Sul Sulfide 1 Effluent " ~ Coagulation l~'Coagulation 2 T Treatment'~ Residue [~ I~ Residue Washing Primary Sufide CuS
Figure 2
Washing ~
(Melting)
SecondarySulfide PbS + ZnS
Flow diagram for recovery of heavy metals
63 3.1 Outline of process This process proceeds with dissolving the melting fly ash as containing low-boiling heavy metals concentrated by the melting process in hydrochloric acid, adjusting the pH value to that specified to each heavy metal concerned and separating and recovering each heavy metal as the heavy metal sulfide produced by adding NariS to the solution. Since the theoretical calculation of the solubilitiy product reveals that the solubility products of heavy metal sulfides are much different from each other in a low-pH range, the process utilizes the phenomena. The outline of this process based on the principle is illustrated in Fig. 2. This process belongs to the wet refining process and comprises the acid dissolving stage, the production and separation stage of sulfide and waste water treatment stage. These stages are described hereafter. 1)Acid dissolving stage: the melting fly ash is dissolved in hydrochloric acid to extract heavy metals and the concentrations of heavy metals concerned in the solution are analyzed to decide the quantities of chemicals to be used in the following stage. Meanwhile, the dissolution residue is separated from the supernatant liquid, dewatered and melted again. 2)Production and separation stage of sulfide: The pH value of the supernatant liquid sent from the previous stage is adjusted to pH specified to each heavy metal concerned and NariS of the quantity determined in the previous stage is added to the liquid to produce, precipitate and recover each heavy metal sulfide. The recovered heavy metal sulfide slurry is washed and dewatered to prepare the recovered materials. 3)Waste water treatment stage: The supernatant liquid separated in the previous stage undergoes the pH adjustment and coagulating separation, the separation from the waste water sludge and the discharging. The waste water sludge is dewatered and melted again with the dissolution residue.
3.2
Separating recovery experiment of heavy metals
An experiment for separating and recovering the heavy metals was made using the sample of 60 kg/batch. Two samples of melting fly ash of A- and B-Ashes listed in Table 3 were used for the experiment. The compositions of both samples are listed in Table 4. Table. 3 Sample
Properties of melting fly ash
Charactristics of melted materials
Type
Gas treatment
,
....
NaOH blowed Stoker incineration residue Surface melting (Bottom ash + Fly ash,NaOH blowed) Electrostatic presipitator ....
Fluidized bed incineration residue Plazma melting (Bottom ash+Fly ash,Ca(OH)2 blowed) ....
Ca(OH)2 blowed ......
Fabric filter
64
A- and B-Ash samples contain as high as 21.42% of Na and as high as 29.6% of Ca, r e s p e c t i v e l y , because both ash s a m p l e s are affected by alkaline components added to the off-gas produced by melting the incineration residue. The higher contents of heavy metals concemed with recovery were selected as follows: 1.56% of Pb and 2.72% of Zn in A-Ash and 0.75% of Cu, 0.80% of Pb and 0.46% of Zn in B-Ash. The procedure for precipitating-separating recovery of the heavy metals in the sulfide production process is conducted by selectively separating Pb and Zn from AAsh and CuS and (PbS & Z n S ) f r o m BAsh, respectively. 1)Properties of recovered materials The properties of the materials recovered from A- and B-Ash are listed in Table 5. A sample of A-1 of the Pb-containing material and that of A-2 of the Zncontaining material recovered from AAsh are judged to be applicable to the raw materials for the refinings of lead and zinc, respectively. Another sample of B-1 of the Cucontaining material and that of B-2 of the (PbS&ZnS)-containing material recovered from B-Ash are judged to be applicable to
Table.4 Properties of melting fly ash Item Unit A B SiO2 % <0.05 0.40 ! Ae 203 % 0.04 0.38 Fe203 % 0.12 0.38 Ca % 0.66 29.60 K % 20.51 7.83 Na % 21.42 5.93 PO4 % 0.23 0.11 Mg % 0.01 0.12 Ti % < 0.01 0.02 T-S % 0.52 2.50 SO4 % 3.28 7.48 T-C e % 44.15 22.46 Pb % 1.56 0.80 Sn % 0.18 0.01 Cu % 0.17 0.75 Zn % 2.72 0.46 Ni % <0.01 <0.01 T-Cr ppm 72.4 80 Mn ppm 52.6 100 T-Hg ppm 0.21 0.39 Cd ppm 21.8 80 As ppm 27.1 5.65 CN nnm <0.5 <0.5 Table.5 Sample A- 1 A-2 B- 1 B-2
Properties of recovered materials Pb 40.27 2.26 4.56 17.57
Zn 15.86 61.28 0.11 34.57
Cu 51.56 0.66
Ce 1.29 0.66 0.87 0.60
the raw materials for the refinings of scrap copper and ISP, respectively. 2)Properties of remelted material The dissolution residue produced from the acid-dissolving process and the waste water sludge produced from the waste water treatment process are remelted Fig. 3 illustrates the migration rates of the main elements contained in the melting fly ash of B-Ash to each process. The figure reveals that such elements as Si, Ca, P, A1 and Fe with the rate of migration of slag is high migrant to the remelted substance. It is, therefore, confirmed that a problem causing the circulating concentration of the main components is solved by remelting the residue.
65
It is inferred from the data that the circulating concentration o f trace components also does not arouse the problem. The variations of the properties of the ash and the behavior of the trace components will be, however, investigated by a demonstration test. 3)Properties of effluent The effluent produced by treating B-Ash satisfies the standard of waste water. An example of the analytical result of it is listed in Table 6. Fig. 3 reveals that the rates of migration of Na, K and C1 in the treating water of the effluent to the effluent side are high.
Item Cd CN Org-P Pb Cr6+ As T-Hg Se PCB BOD COD SS oH
Table.6 Properties of effluent Unit Effluent Standard for Effluent mg/~ 0.03 0.1 mR/e <0.01 1 mR/~ <0.01 1 mR/~ <0.01 0.1 mR~ e <0.02 0.5 mR/ ~ 0.00 0.1 mg/~ <0.0005 0.005 mR/~ 0.02 0.1 mR/e <0.0005 0.003 mR~ ~ 24.40 160 mg/e 84.80 160 mR~ e 17.00 200 5.90 5.8---8.6
Accordingly, it is able to recover those salts as solid salts by evaporating crystallization and reuse them as the feed salt for the soda-production plant, and condense the evaporated water and reuse it for the whole process. 4)PCDDs/PCDFs (dioxins) The analytical results of dioxins contained in the melting fly ash, dissolution residue, recovered materials, effluent sludge and effluent are illustrated in Fig. 4. The concentrations of dioxins are indicated by TEQ in the figure. Since 97% or more of the dioxins contained in the melting fly ash migrate to the dissolution residue and effluent sludge, they are thermally decomposed by remelting to make harmless. The concentrations of dioxins in the recovered materials containing Cu and (Pb plus Zn) are approximately 0.03 ng-TEQ/g in total and those in the untreated effluent are below 0.1 pgTEQ/~ . If necessary, an optical decomposition process by ozone treatment, etc. is applied.
/
0.35 A m~ 0 . 3 0 0.25 J=
~,
~
40%
~
. . . . . . ......... .. .. .. . . . . . . . . . . . . . . .
.
Fo
AI
Figure.3
Cr
Cu
Pb
Zn
.
2
0
0.15 O.lO
.. . .
0.05
I~ ~I~I ~I~I , Si
0
Cd
, S
o.oo P
Ca
Na
K
Distribution of elements in molten fly ash
molten
fly ash
CI
Fig.4
recovered substance
remelted substance
effluent substance
Balance of Dioxins for the process of heavy metals
66
4. C O N C L U S I O N S 1)Acid extraction sulfide process (AES Process) of fly ash The stabilization of fly ash, the recovery of salts and the circulation of water without discharge of the scrubbing waste water were demonstrated by treating the fly ash and scrubbing waste water together in a full-scale plant. Not only stabilizing the heavy metals contained in the fly ash but also utilizing the heavy metals contained in the stabilized cake as well as the salts to use as the resources will be investigated in the future. 2)Separating recovery process of heavy metals from melting fly ash It was confirmed that the each useful heavy metal is separately recovered from the melting fly ash and these recovered materials have so high quality levels that they are able to be used as the raw materials for nonferrous metals. The effects of the properties of the melting fly ash as on the process conditions and the quality levels of the recovered materials and the behaviors of trace components in the melting fly ash will be investigated in the future.
Acknowledgements The authors acknowledge the very special assistance received from Unitika LTD. in the development of the acid extraction process of fly ash (AES Process) and Kobe Steel LTD. in the development of the separating recovery process of heavy metals from the melting fly ash. Also, special thanks are given to the officers and staffs concerned in the guidance and cooperation for those developments. 5. P R E F E R E N C E S 1. M.Fujiwara,Treatment manual for special controlled general waste,fly ash,Japan Waste Research Foundation,Vol.4,No. 1,P68,1993 2. H.Katsuura,T.Inoue,H.Hiraoka,S.Sakai,Full-scale Plant Study on Hy Ash Treatment by the Acid Extraction Process,Seminar on Cycle and Stabilization Technologies of MSW Incineration Residue,P 137,1996 3. M.Hiraoka,S.Sakai,Review on properties of fly ash from incineration plants and its treatment technologies,Waste Management Research,Vol.5,No. 1,P9,1994 4. H.Kawabata,H.Kinari,M.Katayama, Seperating Recovery Process of Heavy Metals From Melting Hy Ash ,Waste Management Research,Vol.7,No.1 ,P488,1996 Japan Waste Research Foundation,Annual report of WI21C"Slim-Waste" Research, P67,1996 T.Itou,Vitrification of Fly Ash by Swirling-How Furnace,Seminar on Cycle and Stabilization Technologies of MSW Incineration Residue,P66,1994 .
Goumans/Senden/vander Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997Elsevier Science B.V. All rights reserved.
67
Pre-treatment of MSWI fly ash for useful application Evert Mulder and Renze K. Zijlstra T N O - Waste Technology Division P.O.Box 342, 7300 AH Apeldoorn The Netherlands Phone: +31 55 549 3919, Fax: +31 55 549 3287 e-mail: [email protected]
Abstract At TNO a feasibility study has been carried out into the possibilities of removing a number of easily leachable elements from MSWI fly ash. The aim of this study was to find a way to usefully apply MSWI fly ash (after treatment) as a road base construction material. A combination of a slight washing step and a stabilisation/solidification-step with cement and other additives appeared to be convenient to meet the severe standards of the Dutch Building Materials Decree. The slight washing step removes the cadmium and chloride that originally was present for more than 90%. Zinc and sulphate are removed for more than 50%. The remaining solid material (approximately 70% of the original quantity) can be easily processed into a bound road foundation layer, adding some 20% cement and other additives. The leaching characteristics of the stabilised material meet the most severe standards of the Dutch Building Materials Decree in the Netherlands. This means, that the material may be applied without any provisions. To the contrary, stabilised MSWI fly ash (that was not washed) could not even meet the less severe standards. Some preliminary experiments with a Pilot Plant Installation show promising results. Chloride in particular can be removed to the same extent as was found in the batch experiment. The removal of some metals seem to be more complicated. This is subject for further research. A global cost estimate shows that, for the Dutch situation anyway, the processing costs of the combination (washing, processing of the washing liquor, and stabilisation) equals the costs for disposal of MSWI fly ash.
1.
Introduction
In the Netherlands the fly ash from Municipal Solid Waste Incinerators (MSWI) is considered hazardous waste. The Building Materials Decree (BMD) nevertheless allows the useful application of such wastes. Therefore the waste should meet de standards for leaching set in the Decree. Several studies have been performed to develop processes for making applications with MSWI fly ash. In those processes the fly ash was stabilised or solidified with for instance cement. All these attempts have been unsuccessful so far due to high leaching
68 values for anions in particular. Together with metals like cadmium, molybdenum and zinc, anions tend to be easy leachable. Other processes are based on thermal treatment of the MSWI fly ash. Most of these processes achieve sufficiently stabilisation/solidification of the metals, but not for the anions. For this reason TNO tried to pre-treat the MSWI fly ash before stabilisation or solidification. For this purpose the possibilities of a slight washing step were studied. The aim was to develop a technique with which MSWI fly ash is transformed into a secondary building material. The washing should be sufficient to remove the easy leachable components. This aim is in contrast with for instance the 3R-process, where it is the intention to leach as much of the heavy metals as possible [1]. In the 3R-process the removal of the metals is the main goal, whereas in the TNO development the production of a secondary raw material is the main goal.
@
W a s h i n g of M S W I fly a s h ( b a t c h e x p e r i m e n t )
A batch experiment was carried out on some forty kilograms of MSWI fly ash. The fly ash was mixed with water in a liquid solid ratio of 101/kg. Under intensive stirring the pH was maintained at 4 by adding nitric acid during the experiment. The acid consumption was some 3 mole acid per kg fly ash. This is a little less than the amount Laethem e.a. needed in their experiments [3]. After 4 hours the mixture was filtered over a paper filter. The cake was subsequently washed twice with fresh water in a lower L / S ratio of about 2. After the experiment the cake was dried and analysed right after filtration. The moisture content of the cake was approximately 30%. During the experiment some 30% of the fly ash was dissolved. Also the wash-water was analysed. The results of the experiment are presented in Table 1.
Table 1: Results of the washing experiment Elements
Original concentrations (mg/kg)
Concentrations after washing (mg/kg) 1)
Decrease in concentration 2)
Cd
220
32
90 %
Cu
660
920
3%
Mo
17
23
5%
Pb
6000
7800
9%
Zn
14000
6900
66 %
C1-
53000
5800
92 %
8042-
64000
79000
14%
1) = Measured in the solid mass after washing and drying 2) - - After correction for mass reduction of about 30%
69 The conclusions of this washing experiment are: 1. The acid consumption of washing MSWI fly ash at a constant pH of 4 is approximately 3 mole acid per kg fly ash. 2. A reduction in weight of about 30% occurs, mainly caused by the dissolving of the salts. 3. Most of the cadmium, chloride and zinc are removed. 4. Sulphate is also removed, although to a lesser extent,. 5. Copper, lead and molybdenum are not substantially removed.
@
U s e f u l a p p l i c a t i o n of t h e w a s h e d fly a s h
The washed fly ash from the batch experiment was stabilised with cement. The stabilisation was conducted at a road construction firm, specialised in applying waste materials as secondary raw material. The application aimed at was a road base foundation layer. The stabilisation was performed on the washed fly ash as well as on the unwashed fly ash. Of both types of fly ash proctor cylinders were made. In total some 20% of additives were needed to make proper cylinders with the fly ash. The additives comprised mostly of cement, stabilising additives and water. Both types of proctor cylinders hardened well. The cylinder made with washed fly ash had a low porosity in particular. This was proven by the results of the diffusion test. One of these results is the tortuosity factor, which stands for the porosity of the tested material. A high tortuosity factor means that the porosity is low. The tortuosity of the proctor cylinder made with washed fly ash was 20,000, whereas the tortuosity of the proctor cylinder made with unwashed fly ash was 320. The tortuosity of normal concrete is approximately 1,000. The proctor cylinders were tested with the diffusion test (or tank leaching test) according to the Dutch standard NEN 7345. In this test the intact (solidified) product is immersed in acidified, demineralised water. The water is renewed at seven times, up to 64 days. Afterwards the eluates are analysed and diffusion coefficients are calculated from the results. Also emissions are calculated, expressed in m g / m 2. Table 2: Results of diffusion leaching tests after stabilisation/solidification:
Element
Stabilisation/solidification product of washed fly ash ( m g / m 2)
Stabilisation/solidification product of unwashed fly ash (mg/ m 2)
Standards for Category I applications 51
Cu
14
Mo
48
Pb
64
61
120
Zn
88
14
200
C1-
420000
680
18000
70 Table 2 presents the emissions form both the cylinders as well as the standards according BMD. The standards presented are meant for category I materials. Category I materials may be applied without any restrictions. This means that these standards are the most severe. On the other hand, category 2 material may be applied only if provisions are taken that prohibit or prevent leaching. Table 2 presents clearly that washing of MSWI fly ash decreases the leaching of chloride, zinc and molybdenum. The decrease in leachability of zinc and chloride can be explained from the results of the washing step; the greater part of the zinc and chloride, present in the original fly ash, were washed out. The decrease of the leachability of molybdenum cannot be explained from the results of the washing experiment. Molybdenum was washed out for only 5%. Most probably the serious leaching of molybdenum from the unwashed material is caused by the increasing porosity of the material during the leaching test. During the leaching test 24,000 m g / k g chloride is leached; more than half of the quantity, available for leaching. This causes the formation of a reasonably porous structure during the performance of the test. The different fractions of the leaching test indeed show increasing molybdenum concentrations, with the course of the test. Also the difference in tortuosity points into that direction (as is mentioned before). From the comparison of the emissions with the standards of the Building Materials Decree (BMD) it can concluded that the washed and stabilised fly ash may be applied as category I building material. The stabilised unwashed fly ash, on the other hand, may not even be applied as category 2 building material due to high emissions of chloride and molybdenum.
4.
Pilot plant
Consequently TNO developed a Pilot Plant Installation for treating MSWI fly ash with a slight washing step. The installation comprises a three counter current washing units. Figure I shows the TNO Pilot Plant for washing granular materials. With this installation experiments of different conditions for slightly washing fly ash are currently being performed. Preliminary experiments with this installation show that the results of the batch experiment can be reproduced for the anions at even lower liquid/solid ratios and higher pH. This means a significant reduction in the use of both fresh water and acid. Chloride in particular was removed to a large extent. However the metals cannot be removed as good as was found in the batch experiments. Possibly due to big differences in pH values in the successive washing steps the metals are partly precipitated again. This is subject to further research at the moment. Furthermore running a counter current process is proven feasible. Results of experiments performed with this installation are not yet available.
71
Figure 1: Pilot plant for continuous washing of fly-ash. The most feasible application of a slight washing process at a municipal waste incinerator (MSWI), is to integrate it in the scrubbing system of the incinerator. For the scrubbing of MSWI flue gases, in the Netherlands most of the time a wet process is used comprising of two steps and produces a waste water stream. This stream is neutralised, precipitated and filtered to remove the heavy metals as a hydroxide sludge. The remaining salt water is either drained away or evaporated. As is shown in figure 1, the washing process yields a waste water stream as well. This stream can be fed into the waste water treatment process of the incinerator. Part of the filtrate may be recycled in the washing process to increase the metal concentration. In particular a high zinc concentration could make recovery of this metal economical feasible. 5.
Process Costs
The costs of the washing process, taking into account the extra costs of waste water treatment and a negative price of the washed fly ash, and based on an installed capacity of 10,000 tons/year, is roughly estimated at Dfl 200,-. These relatively high costs consist mainly of acid used, waste water treatment costs, disposal costs of the waste water treatment filter cake and a negative price of the washed fly ash of Dfl 50,per ton.
72 The total costs of Dfl 200,- per ton fly ash is lower than the current disposal costs of MSWI fly ash as a hazardous waste. However, in The Netherlands disposal of untreated MSWI fly ash will be prohibited in the near future. In that situation Dfl 200,per ton is considerably lower than that of stabilization/solidification, followed by disposal as a 'less' hazardous waste or of vitrification, followed by useful application of the product. The process costs may be lowered by using acid from the first flue gas scrubbing step as leaching acid or by using less acid. Also the disposal costs of the filter cake may be reduced, by means of stabilization/solidification, followed by disposal as a 'less' hazardous waste.
6.
Conclusions
The following conclusions can be drawn from the results of the work that has been done: 1. A 'slight' washing step is a promising way of pre-treating MSWI fly ash before useful application. After washing the fly ash can be applied as for instance a secondary raw material in road construction. Even application in concrete seems to be possible. 2. Slight washing of fly ash removes easily leachable components as cadmium, zinc and chloride for 70 to 90%. The bulk of the fly ash remains as it was. 3. Useful application of the washed fly ash after stabilisation/solidification is environmentally acceptable. The leaching values of some key elements from a test proctor cylinder made with washed fly ash are below the limit values of the Dutch Building Materials Decree. 4. The most simple, and therefore most feasible way of implementing the slight washing process, is integrating the process in the scrubbing system of the incinerator. 5. The costs of the washing fly ash are estimated to be about the same or even lower than the current disposal costs of fly ash. Of more importance is the fact that after washing the fly ash can be usefully applied. 6. A counter current washing process of fly ash has proven to be technically feasible on pilot scale.
References [1] J. Vehlow, e.a., Semi Technical demonstration of the 3R-process, Waste Management & Research No. 8 (1990), 461 - 472. [2] R.K. Zijlstra and E. Mulder, Feasibility study of washing and useful application of MSWI fly ashes, TNO-report, Apeldoorn, september 1994, (in Dutch). [3] B. Laethem, e.a., Integrated treatment of MSWI-residues; treatment of fly ash in view of metal recovery, Studies in Environmental Science 60, Environmental aspects of construction with waste materials, Elsevier (1994), 525 - 537.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
73
D I R E C T M E L T I N G P R O C E S S FOR M S W RECYCLING
Morihiro
OSADA
Nippon Steel Corporation 6-3, O t e m a c h i 2 - c h o m e , C h i y o d a - k u , T o k y o 100-71, J a p a n
Abstract "Direct Melting Process for MSW Recycling" is a system where various kinds of wastes are directly melted in a single step, and their ashes are converted into slag and metal for recycling. In this process coke and limestone are charged into the melting furnace with the waste for high temperature melting and the adjustment of basicity. And high-temperature reducing atmosphere acts to facillitate the volatization of alkali salts and heavy metals contained in the waste, suppress the entry of heavy metals into the slag. Therefore the slag is non-noxious and can be utilized effectively as civil engineering and construction materials. 1. I N T R O D U C T I O N In Japan approximately 50 million tons of municipal wastes are produced a year and it is becoming increasingly difficult to secure sites available for the final disposal of the incineration residue of these wastes. To overcome this difficulty, various methods for reducing the quantity of wastes have been sought to lengthen the useful life of each of the final disposal sites presently available. More than a decade ago the first commercial plant of MSW direct melting process was constructed and has been in operation still now. The direct melting process permits recycling of wastes because the slag, one of the final products of the process, is non-noxious and can be utilized as a civil engineering and construction material. This paper describes the characteristic features and adveantages of the direct melting process as well as the progress thus far attained in recycling.
2. PROCESS DESCRIPTION 2.1 Two routes of waste melting The waste melting technology can be classified largely into two types: the one is a two-step system in which wastes are incinerated first and the ashes produced from this incineration step are then melted, and the other is a system where wastes are directly melted in a single step. Compared with the former system, the latter single-step route provides a simpler equipment configuration and requires less manpower for its operation. 2.2 Features of direct melting process "Direct Melting Process for MSW Resycling" is shown in Fig. 1 and has the following special features: (1) The system provides a high degree of flexibility in adjusting to different types of wastes, combustible or incombustible, including even wastes that cannot be recycled at the collection level and those unsuitable for incineration, e.g., sludge and landfill wastes. (2) All slag and metal produced as the final output of the waste melting process can be reused effectively as a resource. Therefore, the dust collector ash is the only waste this system discharges, helping lengthen the useful lives of final disposal sites.
74
Wasteheatutilization Electric-generator ! L ~
C~on
Coke.Limestone J~!
~
: ~ ~
~
Spaceheatingetc.
t_er~jpera, t_ur~ !'.21:?-..~:-;! =
nduced-drurt Stack fun
(~Combustionair fun 0 ~ t i c ~.. seoarater
"~ O
| J
[:..: ] Flyash
~<~::::~'[10x:n generater Forceddraftfun
:':.:~.::~::i:i~;.:;~:;i:~:
~
equipment ~ to Landfill
Fly a s h treatment
Fig. 1 FundamentalFlowof Direct MeltingProcessfor MSWRecycling (3) The energy recovered as heat from the process can be effectively utilized for electricity generation. (4) The process provides superior control of toxic gas emissions, and is, itself, environmentally Coke
sound. (5) All these features mean substantial cutbacks in
Waste
Limestone
the expenses incurred at all stages of waste treatment from collection to final disposal. 2.3 Outline of direct melting process The melting furnace proper used for this process is a shaft furnace and receives the waste, coke and ag/ eating zone
limestone from the top. The shaft furnace consists of three zones- the drying/preheating zone (maintained around 300~ 300% to 1000~
the thermal decomposition zone (at
mal mposition
and the combustion/melting zone (at
1700~ to 1800~ In the drying/preheating zone the waste charged is dried on heating. The waste thus dried gradually
bustion/ ng zone
descends through the furnace and is fed through the thermal
decomposition
zone,
where
organic
substances thereof are gasified. The gases produced in this way are discharged from the furnace top and c o m p l e t e l y b u r n t in the subsequent combustion chamber. The hot exhaust gas from the combustion chamber is sent then to the
~lag Fig.2 Direct melting furnace
75
waste-heat boiler to recover steam for electricity generation and other energy-efficient applications. Ashes and inorganic matter that survive the above thermal decomposition descend with coke to the combustion/melting zone. The coke burns in the presence of the air blown into the furnace through tuyeres, evolving a high temperature and intense heat, with which the ashes and inorganic substances are melted completely. The hot melt produced in this way is, following the adjustment of its basicity by CaO contained in the limestone charged, discharged from the system through the taphole and sent through the granulating cage, where it is rapidly cooled and solidified as granular mixture of slag and metal. The granular mixture is separated with the magnetic separator into slag and metal for recycling purposes.
3. W A S T E R E C Y C L I N G T H R O U G H M E L T I N G P R O C E S S 3.1 Recycling of melt If the waste melting technology is to be viable, environmentally and economically, the melt as the byproduct of the process must be not only harmless and compact but also recyclable as a useful resource. To meet this requirement, the following are imperative. (1) The melt is completely molten and non-noxious. (2) The melt can be separated into slag and metal. (3) The slag thus separated is highly consistent in quality. To meet all these requirements, both software and hardware need to be established to ensure a highly stable flow of melting unaffected by any change in the properties of the charge into the system. The coke-bed type melting system, as noted earlier in the process description, fully satisfies these requirements. More specifically, the system is so configured that the melt can be discharged from the furnace bottom only after it has moved clear of the hot coke bed, an important feature effectively 100
I I m" i
N 40 20 i
~ Pb
" Zn
Na
K
S
Fig.3 Distribution ratio of each element
CI
5OOl 400' o E E 300
Slag & Metal Fly ash
oO
200
nJ3
100 0~
91750~ o 1700~ [] 1600~
' ~
Experimental
atm Oxygenpartial pressure
Fig. 4 Relationship between Pb content of slag and oxygen partial pressure
preventing the melt from being discharged while still containing solids. And the high-temperature reducing atmosphere acts to facilitate the volatization of alkali salts and heavy metals contained in wastes, suppress the entry of heavy metals into the slag and increase the content of heavy metals in the fly ash. Fig.3 shows the example of the ash melting test result under the reducing atmosphere condition. [ 1] Fig.4 shows the result of simulation analysis of ash melting process. According to this result Pb content of slag decreases in the high temperature and reducing atmosphere. [2] Meanwhile, ferrous materials, if any, in the slag may impair the recyclability of the slag because they can cause red rusting and volume expansion. To avoid this, a wet type magnetic separator has been developed as a means for separating the melt into slag and metal with high efficiency.
76
Table 1 Leaching Test Results of Slag(mg/1)
Table 2 Chemical composition of Slag Direct Blast-furnace Slag Melting Slag (reference) =~ SiO, 37--42 30--41 .~ CaO 33--45 35--45 AhO3 12-- 18 12-- 20 MgO 1.2--1.8 3--7 Na~O 3.5--6.3 0.23 K~O 0.4--0.6 0.24 ~ s 0.6--1,6 0,2--0,3 FeO 0.1--0.8 0.3--1.7 M-Fe 0.1--0.4
Direct Eiviromental Lower Melting Slag standard limit of for s o i l analysis Mercury alkyl
N.D.
N.D.
0.0005
Total mercury
N.D.
0.0005
0.0005
Cadmium
N.D.
0.01
0.001
Lead
N.D.
0.01
0.005
Organic phosphorus
N.D.
N.D.
0.01
Hexachrome
N.D.
0.05
0.005
Arsenic
N.D.
0.01
0.005
Cyanogen
N.D.
N.D.
0.001
PCB
N.D.
N.D.
0.0005
Trichloroethlene
N.D.
0.03
0.002
Tetrachloroethelene
N.D.
0.01
0.0005
Table 3 Physical properties of Slag 16.0 100.0 99.0 9.5 4.75 94.5 59.2 2.36 Grading (%) 1.18 23.8 0.60 8.4 2.6 0.30 0.15 0.9 1.590 Unit Volume standard 1,431 weight (kg/e ) light-duty 2.74 Absolute specififc gravity 0.806 Absorption (%)
With the proprietary, wet type magnetic separator, it is today possible to keep the metal content of the slag at 0.5% or lower. Such a high efficiency of magnetic separation comes from the unique process feature that the slag and metal solidify separately as the melt is granulated, which owes much to the excellent fluidity of the slag- a benefit of the combination of high-temperature heating by coke and the basicity adjustment by limestone. 3.2 Physical properties and uses of slag and metal As evident from Table 1 which indicates the results of the leaching tests conducted on the slag, all measurements stand below the enviromental standard for soil. Table 2 shows the typical chemical compositions of the direct melting slag and blast-furnace slag as a reference. For the contents of the three main constituents- SiO2, CaO and AhOa, the slag is roughly comparable to the blast-furnace slag. In terms of the alkali content, however, the former is a little higher than the latter. The physical properties of the slag are indicated in Table 3. Photo 1 shows the appearance and use of the slag. Tne slag can be utilized effectively as civil engineering and construction materials, especially as base concrete aggregates for interlocking blocks. [3] In addition,
Concrele block
Appearance of slag
Interlocking block Phoot I Uses of slag
77
. . . . . . . .
~::~::~:~:~:~:~:~:~:!~:~!~! :~!~!~i~i~ .....:.:~.:.::~i:!i~iiiiiii`!iii!~i!~iii!~iii~i~i!iii!~!i:!ii!~i!i~ii!ii!ii!iii!ii~i~ii!~ii!i::,:.~ii~!!iii~!!~i~
ii~!i,~i~i,ii~i!~ii:iiiii~ii~i~iiiiiiiilii!iii!iiiiii!ii!iiiiiiiiiiiiii!!!~ii~i ,::................ ~iii~ii~ii~!!i~i!i~i!~ii!ii~ii~i~i~;!ii!iiii~i~i~i~iiiii!i~i~i;!;ii!!;i~iii:ii~i~ii~!~i~ii~!~ii~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~iii~iiii)i!ii~iii~!il)!!i i i~i4i Appearance of metal
Counterweight for construction machine Photo 2 Uses of metal
the results of the past paving tests have indicated that it can be put into commercial use also as aggregates for asphalt paving. Photo 2 indicates the appearance and typical use of the metal. The metal is easy to handle and is already in commercial use as counterweights for construction machinery. 3.3 Recycling of fly ash The fly ash is naturally expected to be cleaned of harmful substances and sent to final disposal sites. Noting that heavy metals can be separated from the melt and enriched during the melting process, however, it would be a matter of great significance if the enriched heavy metals could be recovered from the system and returned to mines. Fly ash treatment process using an existing pellet production facility is expected to overcome this difficulty. The chlorination and volatilization process involves addition of calcium chloride to the baked pellet feed to a rotary kiln and recovery of usable metals and good quality pellets from the kiln. Since there are upper limits set for the Zn, Na, K and Cu contents of the pellet for blast furnace use, Na and K which are contained in large amounts in the fly ash need to be eliminated in advance. To determine the feasibility of the above treatment process, preliminary tests using fly ash were carried out. It has been found possible, at least technically, to recover heavy metals from the fly ash as one of the final products of the waste melting process. But there are several restrictions that need to be overcome including the limited availability of suitable locations and still high treatment costs. This technology, however, has received widening interest as a great breakthrough into a 100% resource recycling system based on the waste melting process. 4. C o u s t r u c t i o n R e c o r d The history of R&D for the direct melting process dates back more than 20 years, and two commercial plants of its kind were delivered 17--18 years ago to Kamaishi City and Ibaraki City each. These direct melting process have satisfactorily been in operation. They have fully proved compatible with the changing natures of wastes and anti-pollution regulations that are becoming increasingly exacting. Noting this, Ibaraki City constructed a new plant for the capacity addition and renewal; this plant went on stream early in 1996. Besides, construction of similar plants are now under way at four sites in Japan. (See Table 4 )
78
Table 4 Construction Record Start-up
Capacity
Kamaishi City, Iwate Pref.
1979
50t/d )< 2 furnaces =100t/d
Ibaraki City, Osaka Pref.
1980
150t/d X 3 furnaces =450t/d
1600kw X 2units
Ibaraki City, Osaka Pref.
1996
150t/d X 2 furnaces =300t/d
5000 kwX 1 unit
Iryu-kumiai, Hyogo Pref.
1997
60t/d X 2 furnaces =120t/d
1100 kwX 1 unit
Kagawatobu-Kumiai Kagawa Pref.
1997
65t/d X 2 furnaces =130t/d
1600 kwX 1 unit
Iizuka City Fukuoka Pref.
1998
90t/d X 2 furnaces =180t/d
1200 kwX 1 unit
Ibaraki City Osaka Pref.
1999
150t/d X 1 furnace =150t/d
3500 kwX 1 unit
Power Plant Not Provided
5. C O N C L U S I O N S A revolution is in process in the area of waste treatment in Japan. What is demanded most is a system that can deal with types of wastes with greater flexibility and enable total recycling of resources. From this point of view, the direct melting process may provide a suitable answer to the above requirement. The high-temperature reducing atmosphere inherent in the coke-bed melting furnace helps efficiently separate the charge into the system altogether and recycle the by-products of the process. Gradually, Progress has been achieved in the commercial utilization of the waste melting process. [4] But, to reshape our society into a recycling-oriented one, We feel that much more effort should be channeled into this area and hope that the direct melting process will contribute to total recycling of MSW
6. R E F E R E N C E S [1] M. Osada et al. : Proceedings of the 6th Annual Conference of the Japan Management Experts, PP.381-383 (1995) [2] S. Osada et al. : Proceedings of the 7th Annual Conference of the Japan Management Experts, PP.467-469. (1996) [3] Y. Nakagawa : J o u r n a l of Japan Waste M a n a g e m e n t Association, PP.471-475 (1995) [4] T. Shiraishi et al. : Proccedings of the 5th Annual Conference of the Japan Management Experts PP. 343-345 (1994)
Society of waste Society of waste Vol.48, No.208, Society of Waste
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
THE ABB DRY ASH CONCEPT:
INREC
79
TM
A. Selinger and V. Schmidt ABB Corporate Research Ltd., CH-5405 Baden, Switzerland
Abstract
ABB has developed and demonstrated a new way to sort and recycle ash from waste incineration. In this process, the bottom ash is kept dry and can therefore be directly separated into clean metals, construction material and a fraction suitable for smelting to yield a clinker substitute.
Where does bottom ash originate from?
In developed economies, individuals and businesses are encouraged to separate a number of waste materials for special treatment and recycling. At the same time, dumping of untreated waste is not longer seen as a sustainable solution and thermal treatment is favored. The waste which finally ends up in such a treatment plant is only the "remaining waste", a very special and rather unpredictable fuel, as the separation of waste fractions is subject to rapid change, both in space and in time. Fortunately, many efforts to collect waste separately affect several waste fractions at the same time. If, for example, green waste, paper and glass are separated, than water, fuel and inert material are removed and the overall composition of the remaining waste may not change too much. Thermal waste treatment in a modem incineration plant with efficient flue gas cleaning combines several benefits: The organic fraction is destroyed and the energy content is recovered. A small amount of flue gas cleaning residues contains the toxic components like cadmium and mercury. About 10% of the waste volume, some 25% of weight remains as a mineralized bottom ash. A cost-efficient process which produces marketable products from this bottom ash would provide the final missing link for a sustainable solution to the waste problem. ABB named this patented process InRec TM [Schwyter 1992], [Btirgin 1995].
80
720 kg Clean Off-Gas (.20.co2) 1000 kg Waste
45 kg Flue Gas Resid. 235 kg Bottom Ash__-', 8 kg Salt for Deposition 50 kg ! 15 kg 97 kg 10 kg HCI and Metals Artificta Fly Ash and uypsum Gravel Fine Fraction
Re-use
Fig. 1
Mass balance of municipal solid waste (example)
Wh at is bottom ash?
Half of the ash consists of well-defined objects which did not change much during incineration: ceramics, bricks, glass, ferrous metals, brass, copper, sand, stones and so forth. The other half is real ash, mostly oxides and carbonates of the dominant elements of the earth' crust: Silicon, aluminum, iron, calcium and the alkali metals. Potential problems arise from the heavy metal content of typically 1%, namely copper, lead and zinc. Additionally, chlorides may be a problem due to their solubility in water. Chemically, bottom ash is very reactive due to the content of free calcium oxide (lime). With water, the strong base calcium hydroxide is formed, which results in a pH-value of more than 12 in the aqueous phase. Under these conditions, many elementary metals (e. g. iron and aluminum) are oxidized and can no longer be recycled: 2 AI + 6 H20 -> 2 AI(OH)3 + 3 H 2
Additionally, the hydrated lime causes hydraulic reactions with other oxides [Simon 1995], similar to those occurring during the formation of concrete, i. e. Ca(OH)2 + n SiO2 -> CaO
9n SiO2 + H20
This leads to mineral attachments on the scrap iron. Furthermore, it has been shown that the composition of bottom ash changes dramatically as a function of particle size (see Fig. 2). After hydratation, a separation of the more toxic fine material is impossible.
81
1.0
.........
~X~ 9
0.9
9
C
.
.
.
.
.
.
.
.
.
.
.
,,
'X
'l 0.8 x 0.7
E _ r
.
".
x. ""x
..-q,
0.6
%,
',
"X
0.5
o 0.4 O 9 0.3
> .,-
0.2
%
"
~- 0.1 ..... .....................
0.0
0.01
Fig. 2
I
0.1
.....................
I
.....................
1
[. . . . . . . . . . .
10
Particle size [rnm]
100
Composition of bottom ash as a function of particle size (dry ash from a Swiss plant)
W h a t is t h e I n R e c T M - P r o c e s s ?
The InRec TM process is based on the idea to sort the dry bottom ash before any reaction with water takes place. Up to four modules may be combined to take full advantage of the process.
Cement Melting ~ll~Additive Solidification Metal ~l)PConcentrate Toxic Fine Fraction Dry Ash Discharge
~ ) ~ A r t i f . Gravel Ash Sorting l ~ ) ~ l r o n ~ j l P A I , Cu, Zn
Fig. 3
The InRec T M Process - General Scheme
1. Ash formation The W+E grate with Advanced Combustion Control provides the best possible ash quality: Repetitive roll-over of the waste fuel on the double-motion overthrust grate combined with a controlled supply of combustion air results in complete burnout of the waste and in a maximum transfer of the volatile components into the flue gas.
82 2. Dry discharge The traditional water bath is replaced by a dry ash discharge system. Depending on the subsequent ash sorting, the ash can be discharged through a double-flap lock, which provides sealing to the atmosphere. In this case, the sorting process includes handling of the coarse fraction. Alternatively, a roller grate can be used to separate the coarse material within the ash discharge system. In this case, only the main fraction with particle sizes of up to 40 mm is discharged dry, while the coarse fraction is dumped into a water bath as usual. Both options have been installed in a Swiss incineration plant and were successfully tested for several month in 1995 and 1996.
3. Sorting The sorting of the dry ash is performed with well proven mechanical components. In the first step, the fine fraction (i. e. particles smaller than 2mm) is separated with a Liwellscreen or by wind sifting. This way, the dust load for the following steps is minimized. Iron is removed by means of a magnet drum and an overhead magnet, while an eddy current separator is used to extract the non-ferrous metals from the bottom ash. To separate metals which are sintered to mineral material and to crush larger aggregates down to gravel size, a hammer mill is used.
.
.
.
.
ZigZagSifterI L~/~,~~ ,netjcSeparator
FrU~aht:nret ~
"
~
i
~u
Fine Fraction Construction Material Metals Scrap Iron Oversize
Fig. 4
The InRec T M Process." Dry Sorting of Bottom Ash
Details of the sorting process depend strongly on the desired use of the end products. However, usually 5 fractions are obtained: 950-60%
main mineral fraction (particle size i. e. 2-40 mm) for construction purposes
91 5 - 2 0 %
ferrous metals
91-2%
non-ferrous metals, of which typically 2/3 is aluminum and 1/3 is copper and brass
9
some coarse fraction (amount and composition depends individually on the incinerator)
91 5 - 3 0 %
fine fraction for further treatment
83 All the mechanical components of the sorting process have been tested in full size with InRec T M material. The metals are recovered in an unsurpassed high quality, so that they achieve good revenues on the recycling market. The mineral fraction passes the common leaching tests, so that the use in road construction is possible in most countries. Test of an independent lab have proven, that also the mechanical properties are satisfactory for such reuse.
4. Smelting of the fine fraction with AshArc TM As most of the toxic components of bottom ash are concentrated in the fine fraction, further treatment is required. If reuse of this fraction is desired, is can be treated in the AshArc T M process. Afterburner / Quench Ash Fly Ash Additives Fan Fabric Filter AshArc Furnace Discharging (glass-like slag)
Fig. 5
Discharging (Metal concentrate)
The AshArc T M Process
AshArc T M is a modification of the ABB DC arc furnace which is currently operated in more than 30 installations in the metallurgical industry. The ash is fed into the furnace through a hollow electrode. It is melted by an electrical arc which burns between the graphite electrode and the molten metal anode on the bottom of the furnace. Depending on the furnace throughput, the molten ash is discharged batch-wise or continuously. Most of the salt content, zinc and lead is evaporated. The graphite electrode provides a reducing atmosphere in the oven, which further helps to reduce the metal content of the ash as elemental metals are collected separately on the bottom. To prevent formation of toxic organic compounds, the exhaust gas is first passed through a CO afterburner and then cooled rapidly with air. The evaporated chlorides and metals resublime and are removed by bag filters. The filter cake, a heavy metal concentrate, can be recycled in the metallurgical industry. The flue gas is further cleaned with standard methods to remove SO2 and HC1. Additional benefit can be achieved by treating the waste incineration fly ash together with the fine fraction.
84 The AshArc T M process has been tested in full scale to determine its suitability for melting bottom ash. 50 tons of bottom ash and fly ash were melted to products which even met the tough requirements of the cement industry. Leaching requirements for use in various construction applications are easily met (DEV-S4, Swiss TVA for ,,Inertstoffqualit/~t" etc.). If vitrification is not desired, to save costs and energy, the fine fraction with its high lime content can simply be stabilized or solidified by mixing with water and clay or cement. The suitability of this fraction for flue gas cleaning purposes is currently being tested.
Conclusions
InRec T M enables the total recycling of waste incineration bottom ash, without the effort of melting the whole material. Therefore, it provides the potential to recover the energy and the metal content of the mixed household waste, plus producing artificial gravel for construction. Continued efforts to reduce the heavy metal input into the waste will additionally increase the reuse potential of the ash. The treatment cost of such a process is comparatively low, especially if the AshArc T M melting process is omitted. Compared to a modem treatment process for regular (wet) ash, considerable savings can be expected. This is due to the simplified ash discharge, easy storage and better efficiency of the sorting plant combined with an increased market value of the resulting products.
Btirgin, M., Schmidt, V. and Simon, F. (1995), Verfahren zur Rtickgewinnung von Wertstoffen aus Mtillverbrennungsschlacke, European Patent Office, EP 0 0691 160 A1 Schwyter, L. (1992), Verfahren und Vorrichtung zur Aufbereiung von Schlacke aus Abfallverbrennungs~fen, European Patent Office, EP 0 372 039 B 1 Simon, F. G., Schmidt, V. and Carcer, B. (1995), Alterungsverhalten von MVA-Schlacken, Mtill und Abfall (11), 95
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
85
Municipal Solid Waste Incineration (MSWI) Bottom Ashes as Granular Base Material in Road Construction G. Pfrang-Stotz Institut for Technische Chemie/Thermische Abfallbehandlung (ITC/TAB) Forschungszentrum Karlsruhe Postfach 3640, D-76021 Karlsruhe, Germany Phone: +49 (0) 7247 82 2952 FAX: +49 (0) 7247 82 4373
J. Reichelt Institut fQr Straiten- und Eisenbahnwesen (ISE) University of Karlsruhe Posffach 3640, D-76128 Kadsruhe, Germany Phone: +49 (0)721 608 3869 / FAX: +49 (0)721 60 76 10 Introduction In the Federal Republic of Germany wastes should, in the first line, be avoided and, in the second line, be recycled or utilized for energy production. When MSWl bottom ash is utilized in road construction as granular base material it has to satisfy the structural engineering requirements, and its environmental compatibility has to be ensured. With regard to the mineralogical composition [1] four types of bottom ashes can be distinguished: 1. grate ash: 2. raw bottom ash: 3. bottom ash:
material discharged from the end of the grate grate ash discharged via a quench tank, containing grate siftings raw bottom ash which had been processed and stored for at least three months 4. aged bottom ash: raw bottom ash which had been stored for several years In order to evaluate these requirements, mineralogical, chemical and structural engineering methods were applied to raw bottom ashes as well as to processed bottom ashes stored for three months, all originating from 12 municipal solid waste incineration plants operated on different process technologies.
Elution Behaviour (DEV-S4 Test) The elution behaviour of MSWI bottom ashes is characterized by the German DEV-S4 test [2]. The DEV-S4 test is used for the comperative evaluation of the environmentally relevant properties of industrial by-products based on the crucial limits of national regulations (TL HMVA-StB 95) [3]. The statistical interpretation of a quality controlled MSWI bottom ash shows (Table 1): The largest fraction of cations in the solution is made up by calcium, sodium and potassium. The anions are sulfate and chloride. The largest standard deviations are refered to the elements calcium, sodium, potassium, chloride and sulfate. These elements build up the soluble salts halite (NaCI) and sylvite (KCI), the sulfates anhydrite (CaSO4) and bassanite (CaSO~0,5H20) as well as carbonates, e.g. Calcit (CaCO3). The high leaching rates of these elements are attributable to these mineral phases, which react very sensitive of modifications of the physical-chemical environment.
86 The arithmetic means ofthe concentrations of the heavy metals, chloride, sulfate and cyanide are clearly below the crucial limits by TL HMVA-StB 95 [3]. A comparison of the elution data (DEV-S4) of MSWl bottom ashes produced in 12 MSWl plants operated on different process technologies shows (Table 2): The crucial limits laid down by TL HMVA-StB 95 [3] are conform to the specifications of the values assigned.
Structural Engineering Properties The statistical interpretation of a quality controlled MSWI bottom ash shows (Table 3): > The crucial limits laid down in TL HMVA-StB 95 [3] are observed by the MSWI bottom ashes investigated, with exception of the parameter ,,Resistance to Freezing and Thawing". The low resistance to freeze-thaw-changes can be attributed to the high contribution of low resistant mineralogical attributes such as high porosity, specific mineral composition and relatively weak carbonate bond between the bottom ash constituants. Based on the results of investigations, the following summarizing conclusions can be drawn for the comparison of MSW incineration bottom ashes produced in 12 MSWl plants operated on different process technologies (Table 4, Fig. 1 and 2): The crucial limits laid down in TL HMVA-StB 95 [3] are observed by the MSWl bottom ashes investigated, with exception of the parameter ,,Resistance to Freezing and Thawing". The impact destruction values (SZ-~12) of the Impact Test determined as a measure of resistance to mechanical fracmentation are within the limits laid down in TL HMVAStB 95 [3]. However, a comparison with the impact destruction values of natural mineral substances, i. e. basalt (SZ_~12:9 - 20 wt.%) makes evident that similar stabilities cannot be achieved by MSWl bottom ashes. This is attributable to the relatively weak carbonate bond between the bottom ash constituents and to the specific composition of mineral substances. Thus, a high content of stable silicates and oxides exerts a positive influence on the stability properties whereas high fractions of salts and sulfates reduce the stability. At present time energy- and cost-effective processes were tested to optimize the mineralogy, the leaching behaviour and the structural engineering properties of MSWl bottom ashes without influencing the present process technology expense for MSW Incineration decisively. [1] PFRANG-STOTZ, G. & REICHELT, J. (1996): Mineralogical aspects of environmentally relevant and structural engineering properties of municipal solid waste incineration (MSWl) bottom ashes. International Conference on Incineration and Thermal Treatment Technologies, May 6-10, 1996, Savannah, Georgia, U.S.A., 271-277. [2] DIN 38 414, Teil 4: Deutsche Einheitsverfahren zur Wasser-, Abwasser- und Schlammuntersuchung: Schlamm und Sedimente (Gruppe S), Bestimmung der Eluierbarkeit mit Wasser ($4), Beuth-Verlag, Berlin, (1984) [3] Technische Lieferbedingungen for HausmQIIverbrennungsasche im Stra6enbau (TL HMVA-StB), Forschungsgesellschaft fQr Stra6en- und Verkehrswesen, K01n, (1995).
W G n c e to Fragmentation
PH el. Conductivity
I
I
I
%bymass1
16
1
38,9
1
28
1
10,4 127
mSlcm
I
'Crucial Linmits: Technical Supply Terms for MSWl Bottom Ashes (TL HMVA-StB), 1995
26,l
1
35,5
1
41
1
1
9,5 62,9
1
11.4 175
1
1
1
max. 40
1
17bis13 600 00
4
Crucial Linmits: Technical Supply Terms for MSWl Bottom Ashes (TL HMVA-StB), 1995
0,063
0,09
0,25
0,71
Sieve Size (mm)
2
Fig 1: Grading Curves of MSWl Bottom Ashes
5
8
11
1622,431,5
w
w
Proctor Curves of MSWl Bottom Ashes
I
I
8
10
12
I
14 16 18 Water Content by mass percentage Fig. 2 Proctor Curves of MSWl Bottom Ashes
I
20
22
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
91
TEST PROJECT "CRUSHED MASONRY 50/150 mm IN THE VENTJAGERSPLAAT RIVER DAM"
Mrs. H.A.Rijnsburger, Road and Hydraulic Engineering Division, Ministry of Transport, Public Works and Water Management, Delft, The Netherlands. Abstract: The suitability of crushed masonry as a core material in hydraulic engineering structures has been examined during the Ventjagersplaat test project. A rock-fill dam was built to protect the Ventjagersplaat in September 1992. The South Holland Directorate of the Ministry of Transport, Public Works and Water Management (RWS) gave permission for approximately 150 metres of the core of this dam to be built with 50/150 mm crushed masonry instead of rock rubble. The use of crushed rubble for the core material in darns of this type is considered acceptable based upon the results of the Ventjagersplaat test project. Darns of this type are subject to relatively less stringent requirements. For possible application in comparable projects a number of recommendations are made in this article concerning the requirements to be imposed. 1. Introduction To comply with the terms of the Building and Demolition Waste implementation plan of the Ministry of Housing, Planning and the Environment the reuse of building and demolition waste should rise from 6.8 million tons in 1990 to approximately 12.2 million tons in 2000. To achieve this, "new" markets must be developed. This applies particularly to the largest component of building and demolition waste, masonry waste. An application of crushed masonry, for which there are many opportunities, is as the core material in dams. To demonstrate this, the South Holland Directorate of the RWS made a test location in the Haringvliet available in 1992. 2. Aims of the test project The test project had the following aims: enhancement of the opportunities for the use of crushed masonry. demonstration of the structural and practical feasibility of the application as the core material in dams. gaining an understanding of the effects upon costs when crushed masonry is used instead of rock rubble. 3. Ventjagersplaat test project The test project involved the application of crushed masonry in a dam to be newly built on the western side of the Ventjagersplaat. For this application a 150-metre length of the core of the dam was built using 50/150 mm crushed masonry instead of 50/150 mm rock rubble. All of the dam exterior surface was covered by 10-60 kg rock rubble.
The dam structure is shown in Fig 1. The work was necessary to protect Rack n.d:YaLeIO-BO ktj 600 kl:j/m'2 Rock rubbLe 10-80 kg 500 kg/rnA2 the shoal, used as a feeding ground by birds, from attack by currents and waves. A batch of _~.P.~_~!~ ~ _ ' ~ - o ~ ........ I approximately 720 tons of 50/150 mm crushed masonry was produced in three days at a building and demolition waste processing .'.'.'.'.'.'.'.'.'.SO/1SO mm crushed masord-y \ ~ ballast \ (dumped on geolextiLe) establishment. This batch consisted of approximately 60% brick, 40% concrete and 2% of other components. There were almost no Fig. 1: Cross-section of the dam at the test site pieces retained on the 180 mm sieve from the batch of 50/150 mm crushed masonry. In addition, the content of flat pieces was very high. This was caused by the basic material used and the limited facilities at the crusher for crushing coarsely enough. It is advisable, by discussion with the clients and the building waste processor, to reach a proper balance between, on the one hand, the requirements and wishes of the clients and on the other, the facilities of the crushing plant.
L
L
_1oo.
92
4. Aspects of the execution
Execution of the project took place in September 1992. From the first results of the test project it was concluded that the application of 50/150 mm crushed masonry was a good alternative for the core material in the building of dams. During the execution of the work using crushed masonry as the core material no significant differences were found with rock rubble concerning ease of processing, suitability for vehicular traffic and stability. The loads imposed by the action of waves and currents have not produced any visually-observable damage. It was, however, necessary to remove any floating wood and aerated concrete from the structure manually. There was, as a result of the washing out of brick dust, a red colouration of the water, but this disappeared after one day. Schematically, the loads imposed during execution can be seen as loading/dumping six times and driving over four times. A degradation of the material occurred as a result of the loading during execution: the Ds0 (median sieved size) fell by approximately 20% which corresponds with a "here to there factor" of 0.8. The degradation that occurred was, for the situation examined, acceptable in every respect. It appeared that, when driven over a thin layer of crushed masonry on a hard surface (in this case a hardened foundation), there was clearly powdering. Degradation of the crushed masonry was simulated by drop tests. These showed that the extent of degradation, particularly at the start, was determined markedly by the presence of flat pieces in the crushed masonry. Degradation of the crushed masonry can be limited by restricting the destructive transfer movements, preventing vehicular traffic over it and keeping the content of flat pieces low.
5. Structural suitability Los Angeles Abrasion As a measure of the material Ventjagersplaat Crushed Masonry strength the Los Angeles Abrasion LAA (% w/w degradation) value (LAA value) of the basic material (composition approximately 60% brick/40% concrete/2% other ,,2,60 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . components) was determined. The 50 . . . . . . . . . . . . . . . . . . . . . . . . . ~ . . . . . . . . . . . LAA values were also determined for samples of the three separated types of material (concrete, brick and calcium silicate brick) taken before and after processing and '20 . . . . . . . . . . . . . . . . . . . after a period of three years in the -'110 . . . . . . . . . . . . . . . . . . . dam core. Because the winters during the three-year monitoring 0 concrete brick calciumsilicatebrick period were mild, LAA values were main r masonry r also determined for pieces removed Jl ~ before positioning ~--~ after positioning I from the core and subjected to 25 fromthe core alter 3 years [ ] after freezing/thawing test freezing/thawing cycles according Fig. 2: LAA values for the main constituents of crushed masonry to NEN 5184. LAA values from 45 to 55% were found for the crushed masonry used. The LAA values measured for the three separated material types were fairly constant and were not changed demonstrably by placing in position, after three years practical loading and after freezing/thawing loading (see Fig. 2).
i ~I
Resistance to freezing and thawing The resistance to freezing and thawing of the crushed masonry used, as a total for the three main components: concrete, brick and calcium silicate brick, is of the order of a 1.5 to 2% weight loss when tested according to NEN 5184. The winters during the three-year test period were relatively mild. It has been established that the freezing/thawing resistance did not fall during this period. The fall to be expected for the D~0 as a result of freezing/thawing is small compared with the "here to there factor" found of 0.8.
93
Settling of the dam
No significantly different subsidence of the upper part of the dam was measured for the part where crushed masonry was used as the core material than for the part where rock rubble was used. Settling measurements were in the range from approximately 1 to 10 mm with a mean of between 3 and 3.5 ram. It should be noted that the mean subsidence with time (during the 46 months after construction), for both the part of the dam with a rock rubble core and the part with a core of crushed masonry, has hardly increased (see Fig. 3).
SETTLING OF THE VENTJAGERSPLAAT DAM 2
E -9 ~1 (--
rock rubble and crushed masonrycores
"
I
-
--i-
,
I
t
*
I
__1_
-l: TYPE
_}_
-J--
~
--
r162 - 1 0 J n,'
-12
_ 14
_ 22
_ 33
46
] R o c k rubble o Mean ] Crushed masonry o Mean
Time in months Fig. 3: Settling of the Ventjagersplaat dam
6.
Recommendations for the requirements to be imposed upon crushed masonry as a core material for dams Based upon the Ventjagersplaat trial project the application of crushed masonry as the core material for dams of this type, where the requirements can be relatively flexible, is considered to be acceptable. For a permissible fall in the Ds0 to 75 to 80% the following requirements are recommended and involve the mean properties of the batch: - composition: (according to NEN 5942) calcium silicate brick: 10% w / w maximum brickwork + calcium silicate brick: 60% w / w maximum The crushed masonry may not contain any parts that will remain floating in water. - flat pieces content: (according to DWW-MAW-R81054)
40% w / w maximum
- LAA value of the separate constituents: (according to ASTM C-535) concrete: brickwork: calcium silicate brick:
45% w / w maximum 60% w / w maximum 70% w / w maximum
Note that the permissible fall of the Ds0 must first be determined, so that after construction the filter requirements included in the design can still be met. In most cases a reduction by approximately 20% will cause no design problems. In special cases where only a lesser reduction is permissible it is advisable to make the flat pieces content requirement more stringent and to limit the number of destructive transfer movements. 7. Environmental aspects The crushed masonry used came from a rubble crushing installation certificated by the Stichting Kwaliteitsborging Korrelmix [Aggregate Quality Guarantee Foundation]. The crushed masonry complied with the requirements in mid 1992 of the Building Materials Order that was then being drafted. Thus the crushed masonry could be used without restrictions in surface water. The material must be taken back by the owner if the works for which it has been used no longer have a function and/or they are no longer maintained. This applies in general to all Category 1 materials that have not been shaped into a form and for example, to rock rubble.
94 After the Building Materials Order comes fully into force in mid 1998 the producer will have to show that the granular material complies with the requirements of the Order. Before then the rules according to the IPO (Interprovincial Consultative Body) policy will apply. 8. Economic aspects The cost of the crushed masonry delivered to the site (in 1992) was f 31 /ton compared to f 29/ton for rock rubble. Nevertheless the use of crushed masonry based on m 1 of the dam to be built, is about 20% cheaper. This difference is caused by the fact that each metre of the dam's length needs fewer tons of crushed masonry (approximately 4.5 tons compared to 6 tons of rock rubble). This is a consequence of the lower density of crushed masonry (1900 kg/m 3) compared with the density of rock rubble (approximately 2600 kg/m3). It should be noted here that about 45% of the cost of supply on the site of the crushed rubble arises from the higher transport costs as a result of an additional transfer. 9. Conclusions The Ventjagersplaat test project has shown that the application of 50/150 mm crushed masonry in bulk quantities can be a good alternative structurally, environmentally and during construction to the traditional rock rubble as the core material in dams. In the period of examination of 3 years until now no indications have been found of a significant deterioration of the crushed masonry. It is considered acceptable, based on the Ventjagersplaat test project, to use crushed masonry as the core material in darns of this type for which the requirements can be relatively less stringent. Recommendations have been made, in connection with possible applications in other comparable projects, concerning the requirements to be imposed. Parties involved The test project was carried out under the responsibility of the CUR (Centre for Civil Engineering Construction, Research and Regulation) research committee B37 "Application of alternative materials in hydraulic engineering", working party 8 (applications) and supervised by the Ventjagersplaat project group. After the dissolution of committee B37 the responsibility for the test section was taken over by the Road and Hydraulic Engineering Department (DWW) of the RWS. During the execution of the project, in addition to the Ventjagersplaat project group, the following parties were represented: Aannemingsbedrijf Spaans en Zn of Werkendam (contractor), Brekerij Julianahaven of Dordrecht (masonry waste crusher), the South Holland Directorate of the RWS of Rotterdam and Fugro Consulting Engineers of Arnhem.
Literature 1. Implementatieplan bouw en sloop afval, herziene versie mei 1995. Publicatiereeks afvalstoffen, 1995, nr. 23 [Implementation plan for building and demolition waste, revised version May 1995. Publications series on waste materials, 1995, No. 23.] Proefproject Metselwerkgranulaat Ventjagersplaat - Haringvliet. Fugro rapport M-6022, d.d. 12 augustus 1993. [Ventjagersplaat - Haringvliet crushed masonry test project. Fugro report M-6022 of 12 August 1993.] Proefproject in Haringvliet: metselwerkgranulaat veelbelovend. Land en Water, Jaargang 33, nr.lO (oktober 1993), p.38-41, [Test project in the Haringvliet: crushed masonry very promising. Land and Water, 33, No. 10 (October 1993) pp.38-41] M.T. van der Meer, H.A.Rijnsburger, J.M.Arnst, M.Geense. Proefproject rnetselwerkgranulaat darn Ventjagersplaat. Eindconclusie. DWW rapport W-DWW96-053, Publicatiereeks Grondstoffen Nr 1996/04. [Crushed masonry test project for Ventjagersplaat dam. Final conclusion. DWW report No. W-DWW-96-053, Raw materials publications series No. 1966/04.] Bouwstoffenbesluit 1995; Staatsblad van het Koninkrijk der Nederlanden; Besluit van 23 november 1995. [Building Materials Order; Official Gazette of the Kingdom of The Netherlands; Order of 23 November 1995] containing regulations concerning the use of building materials on or in the ground or in surface water. IPO-interimbeleid "Werken met secundaire grondstoffen". [IPO interim policy "Working with secondary raw materials".]
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
95
Evaluation of treatment of gas cleaning residues from MSWI with chemical agents.
S. Mizutani*, H.A. van der Sloot**, S. Sakai* * Environment Preservation Center, Kyoto University, Japan ** Soil & Waste Research, Netherlands Energy Research Foundation, Petten Abstract In Japan as in many other countries treatment of Municipal Solid Waste Incinerator (MSWI) fly ash is necessary to obtain a less hazardous material for landfilling (in the context of this paper MSWI fly ash consists of combined electrostatic precipitator ash and air pollution control residues). Ultimately, treatment may lead to potential use of materials. Different treatment processes based on the addition of chemical agents are currently in use or in development in Japan. In this paper, the leaching behaviour of untreated MSWI fly ash and chemically treated fly ash has been studied. The treatment processes involved are: treatment with a chelating agent, phosphate treatment, ferfite treatment. In the characterization of the leaching behaviour of treated and untreated fly ash, which all fall in the category of granular materials, the Japanese leaching test (JLT-13), the Availability Test, a pH dependent leaching test, a redox property test, and the serial batch test under open condition and closed condition from L/S = 2 to 100 are performed. The tortuosity of the compacted materials was tested separately in diffusion tubes using Na-22-tracer. The data obtained from the different tests are placed in perspective to derive conclusions on the environmental properties of the treated wastes under confined conditions and the management control measures that are needed to prevent undesired release of contaminants.
Samples One untreated MSWI fly ash as commonly generated in Japan and three different chemically treated materials of the same residue have been studied. The treatment processes involved are: treatment with chelating agent, phosphate treatment, ferrite treatment. In the rest of the paper, the materials are coded as "original", "ch", "ph" and "re" respectively. The original fly ash is a mixture of fly ash and APC (air pollution control) residues from an electric precipitator in the stoker incinerator for municipal solid waste. The flue gas is treated with injection of dry Ca(OH)2 into the gas stream, therefore the residue is highly alkaline material. This type of MSWI fly ash is very common in Japan. Chelating agents are very commonly applied treatment agents in Japan. It is based on complexation of Table I Contents of metals in samples heavy metals with pHn ANC': Contents (mg/kg) .... an organic sulfide. meq/g Cd Pb Cu Zn Na K Phosphate original 12.6 4.44 74 1400 570 6000 29000 38000 ch 12.4 0.84 45 694 330 4220 30400 29100 treatment is based ph 11.0 2.49 43 680 320 4120 25200 32100 on the formation of fe 10.6 0.74 64 884 310 6300 26400 21700 insoluble metal * 1 final pH contacting with 10 times volun~ distilled water. phosphates, such as *2 caloKated from added HNO3up to pH 7 at Avail___abilityTest. Pbs(PO4)3CI, ....
96
Cd3(PO4)2. Ferrite process is a special treatment method. It is based on the formation of a crystalline form of Fe203. By adding FeSO4 with water into the fly ash and subsequently adjusting the pH to 9.5 to 10.5 by addition of NaOH, the material is heated to about 60 - 70 ~ A crystalline phase of Fe203 is formed and some metals are included in the structure and thus stabilized. Contents of some metals and ANC (acid neutralizing capacity) of the samples are shown in Table 1. Test Method
Exp. 1 JLT-I3 (Japanese leaching test No. 13) 1) Weigh 50 g of samples and add 500 mL of distilled water into a polyethlene bottle to reach an L/S of 10. Then the bottle is closed with a cap and shaken horizontally for 6 hours. After shaking, the leachate is filtered over a glass fiber filter (1 um). Exp. 2 pH dependent test Weigh 25 g of sample and add 250 mL of distilled water into the polyethylene bottle to reach an L/S of 10. The liquid is stirred continuously while measuring the pH. Acid or base (reagent is HNO3 or NaOH) are added in order to keep the pH at the preset pH value for 6 hours. After 6 hrs, pH and redox potential of the leachate is measured. The leachate is filtered over a membrane filter (0.45 urn). Exp. 3 Serial batch test under open condition and closed condition open test; Weigh 20 g of sample and add 40 mL of distilled water into a glass beaker (L/S =2). The liquid is stirred continuously for 23 hours. After 23 hrs, pH and redox potential of the leachate are measured. The leachate is filtered over a membrane filter (0.45 urn). The residue on the filter was used as a sample for next step. To the residue, 160 mL of distilled water are added (L/S = 8, total L/S is now 10) to the beaker. It is stirred for 23 hrs. After stirring, the leachate is filtered over a membrane filter. These operations are repeated 5 times. Quantity of leachant was respectively 40 mL, 160 mL, 200 mL, 600 mL and 1000 mL to reach in individual steps L/S 2, 8, 10, 30, 50 respectively. Therefore the cumulative L/S values reached are 2, 10, 20, 50, 100. closed test; Weigh 20 g of sample and add 40 mL of distilled water into a plastic bottle which can be closed (L/S =2). Then N2 gas is flushed in the head space of the bottle in order to purge the air. Other operations are performed in the same way as the open test. The first and second step are the same as in the recent CEN test 2). Exp. 4 Availabifity Test This test was performed pursuant to NEN 73413). Exp. 5 Redox property test4) Weigh 50 g of material and add 100 mL of distilled water in a bottle. Shake the bottle for 24 hours. After shaking, pH and Eh of the leachate are measured and compared to the Eh value of the distilled water, whose pH is adjusted to the same pH value as that of the leachate. When the Eh value of the leachate proves to be lower than that of the destilled water by more than 50 mV, then the material can be considered to exhibit reducing properties and the reducing capacity of the solid material needs to be determined. The reducing capacity is determined as follows: weigh 2 g of material in a plastic bottle and add 20 mL of 0.1 M Cerium (IV) sulfate. After shaking for 2 hours, the materials were titrated with 0.1 M Fe 2§ - ferro sulfate. The reaction is:
97 Ve 2+ + C e ~+ . . . > F e ~+ + C e ~+
The results are expressed in mMol 02 per gram solid material. Exp. 6 Diffusion tube tests) The diffusion tube is filled with untreated, wetted material to a length of 25 mm with a consistency similar to that of the labelled material. The Na-22-tracer labelled material is added up to a total length of 50 ram, and the second piston is put in position. The tube is stored in a saturated environment to avoid drying out during the experiment. After one day, the combined segments are cut into slices and are transferred to pre-weighed counting tubes, dry at 85 ~ weigh, and count in a sodium iodide crystal connected to a two-channel analyzer. The mass of the slices and the total mass of the tube contents are used to calculate the axial length of each slice. The effective diffusion coefficient were calculated based on the mobility of the Na-22-tracer, sample weight and the moisture content. From the pDe of Na and its free mobility in water the tortuosity of the granular matrix is derived. Results and discussions
Leaching concentration for JL 7"-13 Table 2 pH and metal concentration for JLT-13 The results of JLT- 13 are shown (unit: mg/L except pl-I) in Table 2. As the final pH is high, Pb pH Cd Pb Cu Zn . is leached from original fly ash in high original 12.6 ND 7.44 0.04 2.54 concentration due to its amphoteric ch 12.4 ND 0.01 ND 0.04 ph 11.0 ND ND ND ND character (Japanese standard of Pb for fe 10.6 ND ND ND ND waste disposal is 0.3 mg/L). All of the treated materials meet the requirements, because the release of the metals is controlled. However, it is a result of only one single batch test and other evaluations are needed to ensure that there are no long term environmental impacts. Availability of metals Comparing the availability of the Table 3 Release of metals in Availability Test (unit: mg/kg) materials, we can see a significant Cd Pb Cu Zn difference between the three treated original 66 168 139 4490 materials (Table 3), though there is ch 55 166 10 3500 almost no differences in JLT-13. For ph 14 0 23 1590 example, "ph" treatment decreases the fe 54 160 82 3520 availability of Pb significantly, though "ch" and "re" material show little decreasing of the availability. For the other metals similar changes are observed. When availability is reduced, this implies that the metals are incorporated in insoluble mineral phases. The availability reflects a leaching potential and as such is an important factor for evaluating long term environmental impact. pH and pe pe of the leachates are plotted versus pH (Fig. 1). Here pe is a value of negative logalithm of electron activity and there is a relationship pe = EH/59.2 (En means the standard hydrogen potential). For normal oxygenated water with pH adjusted to cover the range pH 4 - 13, there is an almost linear relation (slope is about 1) between pH and pe 6~. For leachates of fly ash or treated residues, there is a similar relation far from the
98 I level observed in oxygenated water. This points DW ( 0 2 = 0 . 2 a t m ) - - , - - o r i g i n a l - = - c h - - - - - p h - - ~ - f e I 18 at strong reducing properties for all. The intercept of y-axis is indicative of the redox behaviour of the materials. As "ch" treated material has the smallest intercept value, we can see that chelating agent has a high reducing capacity. This is in agreement with the results of 6 redox property test. The redox capacity values 2 derived from redox property test are shown in Table 4. The lower value for ph and fe treatment relative to the original material may 1 4 7 10 13 oH be due to some kinds of oxidation during the treatment process. On the contrary, chelating Fig. 1 Relationships between pH and pe agent have reduced the original Table 4 Redox capacity of samples (unit: mmol O2/kg) material in the treatment Sample original ch ph fe i process, redox capacity 387.8 448.8 235.0 188.1 I
14
"~-
-2
n
i
i
t
Release of metals Release of Cd, Pb, Cu, Zn versus pH from each materials are plotted (Fig. 2). As for cadmium, the leachability is reduced for all treatments. Particularly, the treatment :
original
. . . . Availability
-
ch
=
= fe ]
ph
1000
",,.
100
\ |
10
0.1
UD
E
2
5
8
11
2
14
1000
5
8
11
14
10000
_m
Q) n,
Pb
.
100
Cu
. ~
Zn
1000 100
10
10
1
1
0.1 2
5
8
11
14
0.1 2
5
pH Fig.2 Metal release vs. pH of lcachate
8
11
14
99 with chelating agent results in a significant reduction in release in the pH range 6 to 10 ( > 500). For the two other treatments ph and fe the reduction is not as large, although still very significant. In the pH range below 4 the difference between the different treatment options is not as large (factor 2 to 10). As for lead, there are two pH domains to be discussed in relation to treatment effects: one is the behaviour of under alkaline conditions and the other is the behaviour under acidic conditions. In the neutral pH range leachability of Pb is small for all materials including the original material. All three treatments methods reduce Pb leachability significantly in the pH range 5 to 12. The "ph" reduces release even down to pH 4. In the acidic range below pH 4 the difference becomes less until at pH 2 the difference amounts to a factor of 2 - 3 only. As for copper, "oh" treated material has a strong effect over the entire pH range from 2 to 13. Although "ph" and "fe" materials show similar behaviour, there is a difference when both change in release and change of ANC (acid neutralising capacity) are taken into account. This aspect is discussed in the next section. As for zinc, there is small difference among three treated materials, though they are all lower than 10 ~ original the untreated fly ash in the pH range 4 to 10. -=- ch 8
ANC and release of metals The ANC [meq/g] versus pH are plotted E4 in Fig. 3. There is a significant difference =~ < between the four materials. In the neutral pH o range 5 to 9, there is a significant difference -2 between original fly ash and treated materials. It 1 4 7 10 13 means that it is harder to decrease the leachate pH pH for original fly ash than treated materials. Fig. 3 ANC value vs. pH The sensitivity to externally imposed pH changes as reflected by the -----Cd-rebase . . . . . w a s t e standard . - A d d e d acid change of ANC in 1 100 . 10 100 10 relation to pH is an inal 8 important factor for 8 evaluation of treatment methods. In Fig. 4, the 4~ Cd release curve and added acidity curve q %t~f \ versus final pH are 0.1 , , .... "%....A 0 0.1 0 1 4 7 10 13 4 7 10 13 plotted. Japanese 10 100 10 standard for waste ,9 ch fe disposal for Cd (0.3 '| mg/L, it is equivalent 10 to 3 mg/kg of release.) are also drawn as a dot-dash .9. . . . i line in the same figure. ". _ -0.1 However, the pH3 1 4 7 10 13 oH #I release behaviour is Fig. 4 Added acidity and release of Cd not the only aspect to "& .
.
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100 be addressed here. The 25 25 amount of acid required to odginal ch 20 20 make a change in pH is 15 15 crucial in the sensitivity ===Ill 9 10 III 10 analysis. The buffeting 5 5 F 9 9 capacity that reflects the I I i i mj | | i i l~ iiii i 0 0 resistance of the material to 3 5 7 9 11 1 v 1 3 5 7 9 11 a change in pH due to 25 25 o acidification is higher for ,, fe 20 ph 20 "ph" than for "fe". In Fig. 5 15 15 this is illustrated by the 10 10 concentration change of Cd 9 a 5 II II 5 9 9 I I tlll relative to the amount of i i i l 0 ' ' ' " 0 acid required to make the 1 3 5 7 9 11 I 3 5 7 9 11 change. A larger value pH indicates that less acid is needed to bring about a Fig. 5 pH vs. dC/d(acid) of Cd for 4 materials significant change in concentration. In the graph for "ph" and "re" two maxima are observed; one at pH 7.5 8 and one at pH 4. In comparison with the untreated material the sensitivity for pH change has shifted from pH 8 to pH 4 for all treated materials. In the "ch" treated material, the peak at pH 8 is eliminated entirely. A pH of 4 will not be reached under landfill conditions easily provided there is no mixing with material containing degradable organic matter. Similarly, the role of changing the reducing conditions can be discussed. The information needed for such an evaluation is currently insufficient. However, as a general statement it is clear that oxidation of the sulfide treated materials must be avoided as significant changes in metal leachability can be expected upon oxidation. In a management option to be selected for disposing of the waste, this aspect must be addressed.
pH change of leachate under open condition and closed condition The pH change of the leachate versus L/S value of open and closed test are plotted (Fig. 6). There is a significant difference between open test and closed test. In open test, pH of the leachate decrease continuously. In the closed test the pH increases first and then decreases or open closed remains at the same pH 14 value. The final pH of the open test is always lower 12 12 ~ than that in the closed 11 11 vessel for all samples by 1 10 lO ~ o~in= to 2 units. It can be 9 -o-ch 9 explained by CO2 in the -'-r:' 8 i , ,J. l atmosphere" and it is an 8 0 20 40 60 80 100 0 20 40 60 80 100 important to consider the p.effect of CO2 during US Fig. 6 Change of pH of the leachate leaching test operations.
101
Redox potential of leachate under open condition and closed condition T h e EH (standard 800 8OO hydrogen potential) Chelate '- . odginal 6O0 60O values are plotted versus pH in Fig 7. Solid marks 400 4OO (QO l, etc.) mean 200 2OO closed tests' results, and 0 0 open marks ( O Q I--1, etc.) shows the open q , -2OO -200 9 10 11 12 13 14 8 9 10 11 12 13 14 tests' results. As the ORP "~ 800 is a function of the pH, it - .. ferrite phosphate is meaningless to 60O 6OO compare the ORP values 400 4OO directly. In order to 2O0 2OO compare the ORP values in a meaningful way, the 0 0 measured ORP values for -200 -20O the treated and untreated 8 9 10 11 12 13 14 8 9 10 11 12 13 14 material must be pH compared relative to the Fig.7 E. vs. pH for serial batch test dotted lines reflecting the ORP for oxidised conditions as a function of pH.In general, EH values of the leachate in the open test are higher than the closed ones. This is a result of oxidation by 02 in the air and exhalation of Hz from the reaction of the material with alkaline water. For original ash, the lowest points are the result of the closed test at L/S = 2. In the test the swelling of the closed bottle was obserbed. It may be the effect of the H2 gas generation from the contacting with distilled water. This gas generation can be explained by following reaction: 2 AI + 3 H20 + 8 OH" ---> 2 AI(OH)4" + 3 H2 The formation of hydrogen gas from original ash during the leaching test was confirmed by gas chromatography. The swelling of bottle was not observed for treated materials. The treated materials were mixed with water in treatment process at which state part of the H2 may escape to the atmosphere, whereas, the leaching test is the first opportunity to contact with water for untreated fly ash. The large difference between the original fly ash in the open test and closed test is attributed to the degassing of H2 from the open vessel. The "ch" treated materials are generally reducing and their leachate under open conditions are lower than the dosed test results of other materials. Furthermore, for chelating material, EH v a l u e s increase at the same pH. It means the material is oxidized during the serial batch test. This is attributed to the higher reducing capacity of this material. On the contrary, the plots of ph and fe materials are almost pararell to the dotted line. It indicates these materials are neither oxidized nor reduced during the serial batch test. 41"
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102
Release of metals under open condition and closed condition As for the I 9 sedalbatch Cor~ert ....... A~ilability 9 JLT-13 I metal concentration and release, almost 100001000I Pb(close) 100001000t" Pb(open) i Pb, Cd, Cu, Zn are 100 ""i ................................... 100 ""i ............. ~'~................ ," not detected from 1 0 ~ " treated materials, pH 1 of the leachate is one 0 20 40 60 80 100 0 20 40 60 80 1O0 of the most important 10000 t" ...................... factors. Therefore, ~ 1000 ............... 7_n(close) 100001000100t10...................................... 9 Zn(open) only metal release E lOO from original ash are h,.~ shown in the figures 1 I, , , , 1 (Fig. 8). Pb, Zn and " o 20 40 60 80 100 0 2o 40 60 80 1oo Cu in closed test 1000 tOO0 / ! leached out more 100 ..................................... 9 t 100 .... Cu(close) 1 _, ' . ' Cpen)u(o than open test. Na 10 10 and K leached out in 1 highly concentration o~ "-,---'-"~, , , :t 0.1 0 20 40 60 80 100 0 20 40 60 80 1O0 from all materials, and there is small US differences between Fig.8 Metal release vs. L/S in the serial batch test open and closed condition. The I -*-- original ---- ch --*- ph --~ fe release of salts from __~ 20 2O original fly ash and lS J 15 z~ treated materials is 1 1o 10 plotted in the figures 5 K(open) I 5 K(closed) (Fig. 9). Na a n d K 0 from treated 0 20 40 60 80 100 0 20 40 60 80 100 materials leached out t~ 20 20 more than original fly 15 ,'---~--~ )' t 15 ash. It implies that cr 1 0 .5 '~ --f-' " i . with the additives Na Na(ope.) S and K are added. The 0 salt release from 0 20 40 60 80 100 0 20 40 60 80 100 untreated and treated US materials is high. In Fig. 9 K and Na release vs. L/S in the serial batch test evaluating environmental impact, it is important to focus on a broader range of elements than only heavy metals. Several studies have shown that also oxyanions can be important as they often show a maximum leachability at neutral pH 8). 1
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Effective diffusion coefficient and tortuosity of materials In order to be able to quantify release under confined conditions, where diffusion may be the release controlling mechanism, it is important to determine if the treatment influences the physical behaviour of the material. This is done by measuring the
103 tortuosity of the treated materials by carrying out diffusion tube tests with a Table 5 pDe value and tortuosity of materials constituent considered not to interact with material od~linal ch ph fe the matrix (here Ha is chosen). The pDe pDe 9.59 9.85 9.84 10.02 value for Ha, which is an negative tortuosity 10~ 10~ 10~ 10T M logarithm of an effective diffusion coefficient value, for each material are shown in the Table 5. The tortuosity is caluculated from Do/De. Do is the free mobility of Na in water and the value is pD0 = 8.88 at 22 ~ All treatments show a pDe value larger than the untreated. This is attributed to the formation of a denser matrix in part caused by additional precipitate formed. This is consistent with the observation that "re" treated material the effect most significant. As was shown in the batch extractions this positive physical effect does not necessarily mean a lower release as other factor come into play. Conclusion
In order to evaluate the treatment of MSWI residues with chemical agents, the performance of three kinds of treated materials are compared with the behaviour of the untreated MSWI fly ash. Each treated material shows good results for JLT-13 test, however the evlauation based on this single extraction test is too limited to assess the potential environmental impact of the materials in different disposal scenarios. The treatment using a chelating agent is based on complexation of metals with an organic sulfide and the treated material showed a very high reducing capacity and strong retention for metals over a wide pH range. The treatment with phosphate showed a significant decrease in availability, especially for lead, and a very low leachability over the entire pH range from 4 to 13. The treatment with ferrite showed an increased high physical retention as well as a good retention in the pH domain 5 - 12. In order to evaluate the performance of the treatment by chemical agents, the changes in leaching behaviour of elements as a function of pH were compared with the buffeting properties of the treated materials (Acid Neutralizing Capacity), that can be derived from the pH controlled leach test. Besides, the final ORP value of leachate are plotted versus pH of the leachate, the linear relationship are recognized for each material. In order to know the effect of the air during the leaching test, the serial batch test under open conditon and closed condition are performed. There are clear differences between open test and closed test in view of pH and ORP. It is a result of carbonation and oxidation by respectively CO2 and 02 in the air. Furthermore, the H2 generation from untreated ash explains the low ORP in a closed test. In terms of waste management the following conclusions can be drawn: - When materials are produced with very strong reducing properties either in-plant or afterwards through treatment, the management of the materials requires that measures are taken to ensure that the material will never be exposed to the atmosphere, as oxidation will ultimately lead to loss of retaining potential and could result in uncontrolled release. - The generation of H2 in treatment plants is potentially dangerous, if this aspect is not sufficiently recognised in the design and operation of such plants. - Based on the information generated in the more elaborate tests, better predictions of
104 the behaviour of materials can be made. Particularly, in relation to quantifying the changes resulting from changes in exposure conditions (failure of lining, top cover or exposure to the atmosphere). This information can be used to better manage these wastes.
Acknowledgement This paper is based on the results of the study during Mizutani's visit to ECN in the Netherlands. Patrick Cnubben, Dirk Hoede, Petra Bonouvrie and Marco Geuzebroek are gratefully acknowledged.
References 1) Environment Agency in Japan: Japanese Environment Agency Notificaition of No. 13 (1973) [in Japanese] 2) CEN TC 292 Working Group 2: Compliance test for Leaching of Granular Waste Materials and Sludges, Tenth Draft, (1994) 3) NEN 7341. Leaching characteristics of building and solid waste materials Leaching tests - Determination of the availability of inorganic components for leaching. Draft 1993 (previously part of NVN 2508), Netherlands Normalisation Institute, the Netherlands 4) Drait NVN 7348 Determination of reducing properties of materials. 5) van der Sloot,H.A., de Groot, G.J., and Wijkstra, J." Leaching characteristics of construction materials and stabilization products containing waste materials. In: Environmental aspects of stabilization and solidification of hazardous and radioactive wastes ASTM STP 1033, Cote, P.L. and Gilliam, T.M. Eds., American Society for Testing and Materials, Philadelphia, 1989,pp. 125-149 (1987) 6) Lindsay, W. L. : Chemical Equilibria in Soils, A Wiley-Intersicence publication (1979) 7) Schramke, J.A. : Neutralization of alkaline coal fly ash leachates by CO2 (g), Applied Geochemistry, vol.7, pp. 481-492 (1992) 8) H.A. van der Sloot: Developments in evaluating environmental impact from utilization of bulk inert wastes using laboratory leaching tests and field verification. 1996.Waste Management, 16 (1-3), 65-81. 9) IAWG: An International Perspective on Characterisation and Management of Municipal Solid Waste Incineration Residues, final document, Chap. 20 (1995)
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
105
RECYCLING FOR ROAD IMPROVEMENTS Presentation by Charles J. Nemmers, P.E. at the W A S C O N '97 Conference The Netherlands, June 1997
Einstein once said "The significant problems we face cannot be solved by the same level of thinking that created them." These words challenge us to think anew how we use and reuse materials for the construction and reconstruction of transportation facilit;es, especially highways and other roads. The Road Transport Research Program Steering Committee of the Organization for Economic Cooperation and Development (OECD) headquartered in Paris, formed a Scientific Expert Group to review the issue of recycling in road building and I will be reporting on what we did. But it is the need to do our work at a different level (i.e., Einstein quote) that made our collective work both challenging and productive. Fifteen nations plus two Asphalt associations were involved in our study group, a survey of recycling in methods, products, laws, and procedures in these countries was conducted and a state-of-the-practice report was prepared. Several of my colleagues on this report are with us here today and I trust that what I say will be close enough to what we wrote that they will be able to identify it. Recycling is now a well proven technology, often a preferred choice for construction, a backbone for an entire equipment industry, and a requirement, if not a necessity, in many countries. Since 1977 - the year of publication of the first OECD report on the Use o f Waste Materials and By-Products in Road Construction - both the volume and quality of the recycling of road by-products in the road sector have increased significantly in OECD countries. However, the generation of waste remains high as shown in the chart below.
106
Amounts of Municipal Trash in the World The world's most industrialized nations produced an estimated 450,000,000 metric tons of municipal trash in 1992. This figure shows kilograms of trash produced per person per year in the urban areas around the world in selected countries
Kilograms of Trash Produced Per Person In Municipalities of Selected Countries
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107 In 1995, the OECD Road Transport Research Program Steering Committee determined that the advances in recycling called for the Scientific Expert Group on Recyclingfor Road Improvements to take a new look at this important subject area. The primary goal of the Group was to generate information that would describe the state-of-the-art in recycling and promote recycling of waste and by-product materials, especially those generated in the road sector, in road construction. It was anticipated that the Group's efforts would help to: establish or change policy in Member countries; support improvements in current specifications; identify workable technologies; and identify needed research and knowledge transfer initiatives with the expectation of increasing recycling efforts and discovering more innovative and efficient solutions. SUCCESSFUL RECYCLING REQUIRES L E A D E R S H I P While addressing the recycling of both road by-products and non-road by-products in this report, it is agreed that recycling of road by-products takes precedence in order for t'_-z road industry as a whole to take responsibility and demonstrate leadership to the fullest extent possible. The principle of "cleaning up own house first" must apply. The responsibility for road by-products belongs to the road sector. The following general hierarchy for waste management was adopted (applicable to any industry; in this case, applied to the road construction industry): 9 9 9 9
Minimise waste production. Recycle in parent industry. Recycle in other industries. Incinerate. 9 Incinerate with energy recovery; 9 Incinerate to reduce volume. 9 Dispose of in a landfill. Following this hierarchy, an industry (i.e., road construction) not only sets an example for recycling in general, but takes responsibility for reducing its own by-product disposal problems by first attempting to recycle within the industry itself. In the case of the highway industry, it certainly makes sense to recycle road by-products in roads. Not only will this reduce costs of disposal during construction, but the material similarities suggest that such uses would be technically feasible as well. It also makes sense to use by-products from other industries, when appropriate. In fact, the report shows that in some cases (e.g., fly ash in cement concrete) certain by-products may actually benefit road construction by improving the materials.
108
From information supplied by the participating countries, the Group identified three areas that must be jointly considered for successful recycling efforts. They are: 9 engineering factors; 9 environmental factors; 9 economic factors; Each of these areas contain elements - ranging from the political context to specific engineering risks - which have a significant bearing on the potential application of by-products. SURVEYS ARE BASIS OF STATE OF THE PRACTICE Two major surveys were conducted as part of this study. 1) Survey to assess the extent of current use of various by-product materials in Member countries. This assessment was accomplished by distributing survey questionnaires to OECD Member countries. Materials generated by road construction were considered separately from those generated by other industries. a) For road materials, the information requested in the survey included extent of use in various road applications, amounts available, a summary of important properties, test methods and guidelines and description of any new techniques being developed. b) For non-road materials, the information requested also included extent of use in
109 various road applications, additional information was requested on amounts, material tests and acceptance criteria, construction equipment and procedures, quality control tests, standard specifications and factors used in evaluating environmental andeconomic suitability. 2) Survey to assess waste management and recycling policy in Member countries, especially in regard to road building. Information requested here dealt with official policy, organisation(s) responsible for policy, economic issues affecting policy, regulations, obstructions to implementation and technology transfer. These surveys gave us engineering environmental and economic input into our three circle model. The first survey showed that in the 20 years since the last OECD report on recycling in road construction more c-.untries are now recyclin~ many different materials and much more of these materials as part of road construction. Recycling of asphalt pavement is being done in all of the countries surveyed. Nearly 100 million tons of RAP are being produced annually. Other waste materials from non-road industries are also being recycled into road construction and over twenty of these by-products were identified. These products ranged from old tires to various slags and ashes, to glass, paper and plastic. It is clear that the engineering properties, the environmental consequences and the economic possibilities of these non-road by-products need to intersect for them to be a useful recycled material. Through our study we identified several road and non-road by-products that were clearly winners and these were: 9 reclaimed asphalt pavement in new asphalt pavements; 9 reclaimed concrete pavement in new concrete pavement; 9 blast furnace slag as supplementary cementing material in Concrete or stabilised base and subbase; 9 steel slag used in base courses and asphalt pavements; and 9 coal fly ash used in a highway embankment. Recycling materials considering the engineering properties and the environmental consequences is good business but this is not the only way to encourage recycling. This is where our second survey comes into play. Here we asked for the policies that were being utilized to increase recycling. It is clear that many countries are using policies (that is incentives and disincentives) to effect an outcome favoring recycling. It is possible for governments to influence the free market by tax enforcement or subsidy strategies that are designed to promote recycling and the use of recycled by-products. All market parties are encouraged to act in a market-conforming mariner by actions such as: providing sufficient information on the long-term performance of by-products, stimulating test or research projects, implementing a waste tax, providing recommendations and requirements for the use of by-products or subsidising recycling and reproduction facilities. Similarly "Restrictive"
110 regulations can be used to reduce the production of waste and to control its disposal. These regulations must often be balanced, however, with additional policies that encourage sorting, recycling, and reuse.
In several countries, government takes the responsibility for increasing recycling and reuse of by-products. It also works to develop a system of laws and regulations that will restrict the construction industry in how they deal with wastes and encourages the use of by-products. In order to monitor results, most countries set recycling goals for the future. Before a new byproduct is accepted for use, it may be necessary for this by-product to undergo research or a period of testing and evaluation that validates its quality and reliability. A combination of government incentives (subsidies, research) and disincentives (taxes, dumping fees) done co-operatively with industry seems to offer a good formula for success. To further the advancement of recycling world-wide, all countries - regardless ot their current good practices in recycling - should continue to explore new recycling opportunities, advance technical knowledge, and design increasingly effective programs. Keeping in mind that Einstein is calling us to move to a higher level of thinking, our group identified an Implementation Plan for Recycling and provided a model partnering agreement that illustrates the importance of co-operation and outlines the philosophies, roles, and broad-based implementation responsibilities of each partner in carrying out recycling.
111
Individual Partner
Implementation Item ,
By-Product Supplier
Environmental Authority
Engineering Authority (Road Administration Agency)
Owner Agency (Road Administration Agency) Contractor
,,
Specifies product names and characteristics. Develops quality plan to ensure product consistency. Produces product to meet specifications and quality plan requirements. Defines cost and sale parameters. Specifies general environmental regulations. Dcfines pollutants and risks of pollution. Defines appropriate contamination limits. Defines methods to study pollution and specifies associated tests. Proposes taxes or specific incentives to facilitate the use of by-products. Determines suitability of by-products which can be used in road construction and rehabilitation. Undertakes documentary research reviews and data base and supplier information. Considers technical/economical feasibility for each application. Proposes combination of by-product use, technical processes and cost considerations to owner agency and co-operating partners. Conducts any necessary research and demonstration sites on selected options with owner agency and co-operating partners. Develops specifications for project applications and proposes standards. Defines performance requirements or technical needs. Selects by-products and techniques according to program needs/economies. Reviews research results and selects demonstration sites in conjunction with engineering and environmental authorities. Adopts technical regulations and standard documents for contracts. Proposes specific techniques on projects in conjunction with suppliers. Carries out construction activities using by-products according to standards and specifications set forth by owner agency, environmental and engineering authorities. Defines techniques as incorporated in technical proposals and requirements for future applications.
112 SIX RECOMMENDATIONS This Recycling for Road Improvements report recognises that the use of by-products in road construction is contingent upon the technical, engineering, value of the material. However, this contingency is viewed not as an excuse so as to not recycle, but rather as a need to disseminate good information and examples as well as a larger call for continued research into the prudent use of by-products. There is a definite governmental role in this sector. It is suggested that measured public sector involvement can accelerate this adoption of much recycling technology. From the investigation and analysis of the current state of recycling, a series of recommendations are identified that could help to keep pace with increasing global stores of byproducts and decreasing space for landfills.
1. Test materials before recycling. The results of the survey on the use of recycled road materials, shows a significant growth in the quantity, diversity and quality of recycling over the past 20 years. The most commonly used road by-products continue to result from recycled asphalt pavements. The scale of using recycled concrete pavement by-products is less significant than that for asphalt concrete but is developing. The effectiveness of recycling clearly requires one to test the road material
before
recycling so as to: 9 9 9 9
Identify the engineering factors that are critical. Determine the highest level of recycling that is possible. Assure that recycling this time will not foreclose recycling options in the future. Avoid recycling materials that have serious health consequences - i.e., pavements, containing coal tar - or accommodating other materials that have health and safety consequences, i.e., lead-based paints, asbestos in demolition waste.
2. Ensure that recycled by-products are used wisely. Recycling is underway in all countries and will continue to increase as research and new technologies expand the opportunities. The OECD countries responding to the Group's survey clearly show that they are, in many ways, "cleaning up their own house first." It is clear that sometimes recycling (i.e., fly ash, slag aggregates) may offer a product that is better than virgin materials. But it is extremely important to review previous, related research and to test and evaluate as necessary to be sure that the by-product does perform acceptably. There are limitations in nearly all areas of by-product technologies and it is only worthwhile to recycle when you stay within the performance range of the material.
113 In using non-road by-products one must fully consider the acceptable boundaries of use. For instance, while steel slag is a good material in asphalt, it must not be used in concrete. Similarly, many of the applications of unbound by-products require that they not be used in areas close to a water table. Other by-products, such as scrap tires, offer some recycling potential. However, many countries are still researching the value of incorporating scrap tires into hot mix asphalt. Although the road offers good opportunities to accept by-products from non-road sources, it should not serve as a "longitudinal landfill." In other words, the road must first and foremost serve its transport function. When compatible with the recycling of non-road by-products, the road can offer good, reasonable opportunities for governments and industry to recycle byproducts from industries outside the road sector. Continued research into recycling non-road by-products, by both the producers and users, holds promise to reduce global waste disposal and also to reduce the resulting stress that this places on the environment.
3. Promote the increased use of proven recycling solutions. To show how recycling of road by-products and non-road by-products really works, "winners" were identified. Asphalt pavement and concrete pavement recycling are technologies that work, have significant environmental advantages and are economically attractive. The research, demonstration projects and equipment developed in response to these technologies shows that the road industry has done much to "clean up its own house first." These technologies are remarkably similar in all countries and their results are consistent. Similarly the recycling of non-road by-products, such as slags and coal fly ash~ottom ash, were explained through examples from several different countries. Interested individuals can now see a range of by-products from "very successful" to "offering good potential" and use this shared information to improve recycling techniques. We must share.
4. Support policies that foster recycling and discourage dumping. As important as the engineering factors are to the use and recycling of road and non-road materials, other non-technical factors are also significant. Governmental laws, policies, and regulations can set restrictions so as to reduce the production of waste, control disposal and limit the use of new materials - or governmental regulations can be promotional and subsidise the use of recycled by-products, fund research, testing, evaluation and demonstration of recycled materials. What becomes clear as one reviews the survey responses is that a balance of regulations and policies is paramount to the success of a recycling program. Within this regulatory environment there is a stronger emphasis on controls in those countries with denser populations and fewer available disposal options. Promotional options seem more effective in those environments where more options to recycling are available. It appears that increased support of research into better use of by-products was needed in all countries. This is especially true in countries that have established landfill restrictions. For
114 example, the European Community has strong landfill restrictions with a year 2002 compliance date. Most countries reported recycling goals, with many being over 50 percent recycling by the turn of the century. In addition to the need to balance restrictive and promotional policies, it is clear from the information received that the success of any recycling program also depends on the involvement of the public and private sectors. The construction industry, the academic community, the owners of the roads, the recycling industries, equipment manufacturers and others can solve many more recycling problems by working together rather than in isolation. Pilot studies and test sections are strongly recommended. 5. Balance engineering, environmental and economic factors.
The central theme of this entire report is captured in the three overlapping circles where the best potential for recycling is defined by that area of overlap that occurs when engineering (technical), environmental and economic factors are properly balanced. If these factors are not in balance, the result can be environmental requirements to use a recycled by-product that has not been proven, roads being built using structurally sound byproducts but creating unacceptable side effects, or a serious underestimate of the economic effects of any of these strategies. The best solutions are those that balance all three factors in developing an informed judgement. Trade-offs will always need to be made. What is important is that these trade-offs be made with an understanding of all three factors and a concern for "the big picture." 6. Increase research and knowledge transfer.
The lack of adequate information on the long-term performance of by-products, standard specifications and testing requirements will continue to slow the recycling movement. The study team encourages: an increase in recycling research and demonstration projects; an increase in sharing recycling technology among countries; an increased use of incentives for recycling; creation of better devices to encourage recycling; and an effort to convince sceptics (through documentation and publication) of effective recycling. This report should help in this regard.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
115
Abdelkrim BOUCHELAGHEM, INERTEC, Nanterre, France Marie-Claire MAGNIE, INERTEC, Nanterre, France Val~rie MAYEUX, ADEME, Angers, France
SUMMARY
With the new developments of the Beaulieu District in Caen (France), the reinforcement of the underground mine workings appears to be necessary : these mines have been worked since the Xlth century and are now beginning to fall in. In place of traditional material (scouring sand for example), they are filled in with a specific mortar, prepared with inertized bottom ashes, developed for this use by INERTEC. Iron has first to be removed from the bottom ashes, coming directly from the MSW incinerator. Bottom ashes are then passed through a sieve, in order to obtain an homogeneous material, and stored on a watertight area. Sieved bottom ashes are mixed with water and reagents, as defined in INERTEC process in order to have a pumpable product, and then pumped to mine workings. Final material features are tested by inner and outer controls, according to the french regulation. Besides, a specific study has been begun by INERTEC and ADEME on the long term behavior of inertized bottom ashes in underground mine workings, in application of the X 30-407 french methodology, which first results are given in this paper.
116 principles for waste management are integrated :
Introduction In 1995, France published its first standard about long term
the 9 ultimate waste (i.e. waste which cannot be reused,
leaching behavior. It is a guideline which describes the
recycled or treated in the current technical and economical
methodology to assess the long term behavior of waste in a
conditions) alone has to be stored,
given scenario.
reduction 9 of waste production and / or waste noxiousness,
Concerned about environmental protection, INERTEC, in
waste 9 valorization, limitation 9 of waste transport in terms of distance as much
collaboration with ADEME, has decided to adopt this new step to the innovating operation of inertized incineration
as duration.
bottom ashes utilization as mine backfill material.
On December 18, 1992, specific regulations about industrial
Before presenting the first results of this operation, we would
waste landfill dumping confirm for industrial waste the great
like to introduce the French regulation context.
principle of the 1992 act and modify noticeably these landfilling conditions.
1. First laws about w a s t e 1.1 How to manage waste disposal in general
1.2 Particular case o f bottom ashes
With its first law about wastes (1975) and a specific law
On May, 1994, a French Ministry Circular specifies criteria
about classified plants (1976), France owns an efficient
regarding incinerator bottom ashes for landfill dumping and
legislative systems based on the principle of producer's
reused in road basement.
responsibility and best available technology to manage their
Three classes of bottom ashes are defined : 9Class V, low leachable fraction : this class of bottom ashes
waste from production to disposal. Rapidly, a complete network of collective disposal plants for
can be reused directly but are subject to a few restriction
municipal and industrial waste is created.
concerning contact with water, 9Class M : intermediate bottom ashes, they are reusable
Judged heavy at first, these new laws contribute to the development of a new industry whose efforts are increasing.
after
maturation
or treatment
(12
months
maximum)
On July 13, 1992, the 1975 act is updated and new
provided after this the criteria for class V are respected.
Solubles (%) Parameter
Class V limit
Class M limit
Solubles % TOC (mg/kg) S04 (mg/kg) Cr 6 (mg/kg) As (mg/kg) Cd (_mg/kg) Pb (mg/kg) Hg (mg/kg)
5 1500 10000 1.5 2 1 10 0.2
10 2000 15000 3 4 2 50 0.4
.
.
.
.
Toe (mg/kg)
1
t[g (mg/kg)
..
/ .
S04 (rag?k-g)
Cr6 (mg/kg)
>---....
Pb (mg/kg)
.d (m~/kg)
Figure 1- Regulation Limits for Domestic Refuse Incinerator bottom Ash
[
(]lass M lirnit:]
117 9Class S : High leachable fraction, their disposal must be in
each mine was worked without controls, frequently to the
class II landfills.
detriment of the most elementary safety rules. Pillar widths
The test actually used for pollution potential is a compliance
may vary from 3m to lm on a side, from one room to the
test. It consists in three successive leachings after grinding
next. Each mine grew differently over time with different
the ash to 4 mm size, using standard X 31-210 test
pillar numbers and sizes and different faulting affecting the
procedure. Leachate pollutant contents are compared with
roof. Pillars that were too few in number or too small in size
the limits specified in the ministry circular (see fig. 1).
cracked and sometimes collapsed, bringing the roof down with them. Periodical
inspections by the Caen mines
2. An avoidable evolution
department indicates that the competent overlying strata of
These new conditions of waste disposal are not defined on a
massive limestone have so far prevented ground level
notion of impact on environment. It is probably because of
subsidence.
the lack of scientific data in this field.
The land was mainly agricultural, i.e. without building loads
This lack of scientific data was often made up for by the
and only intermittently occupied, and raised no safety
development
problems.
of
technologies
improving
always
more
However, plans to develop the Beaulieu area (22 hectares)
disposal waste. However it seems today that this approach of best available
made it necessary to provide support to all the workings to
technology can not constitute the only answer in term of
prevent future damage to buildings,
roads and buried
objectives and level to reach as far as environmental
services, since static loads applied
by buildings and
production is concerned.
static/dynamic loading from roads would put added stress on
In
this
content,
ADEME
(the
French
Agency
for
mine
roofs.
The
support
work
was
awarded
to
Environment) has been committed for over 5 years with
SOLETANCHE, with a range of support designs to suit pillar
several
at the
cracking and ground loads. The workings under the planned
integration of <~ in regulations. This means
Georges Pompidou road will be the most exposed area
that criteria fixed to choose the best disposal means will be
because of severe static and dynamic loading : it had also to
based on reliable tools of measurement and assessment of
be backfilled. The same method has been used for parts
the real impact of waste on environment.
that are too dangerous for men to enter. At the very
It was therefore for ADEME the obvious thing to accompany
beginning o f the workings, SOI_ETANCHE suggested to use
the innovating operation of Caen.
inertized bottom ash mortar for backfilling.
partners,
in many
researches
aiming
Inertization
process design, quality control specifications and control 3. Application in backfilling BEAULIEU mine workings
tests have been conducted by INERTEC.
3.1 General description The CAEN underground workings (FRANCE), several million
3.2 Development of bottom ash mortar
cubic meters in extent, have been in use since the 11th for
The two categories of pollutant found in incinerator bottom
lime carbonate rock. There was no precise legislation and
ashes are salts and heavy metals.
118 Salts consist mostly of soluble sulfates which must be
Then compliance tests were required to validate the
changed to insoluble forms within the mortar of inertized
laboratory formulation in terms of pumpability, strength and
bottom ashes.
chemical retention of pollutants as required by the Owner
The most important of the heavy metal pollutants is lead,
(City of Caen) and the supervisory agencies. No work was
some of which may enter ash leachate in dissolved form.
allowed to begin before validation of the results obtained
Developing incinerator bottom ash as a stopping material
from these tests by the monitoring panel set up by the
involves:
supervisory agencies. The tests consisted of producing the
- formulating the correct mix of reagents to fix pollutants,
inertized bottom ash mortar in full-size plant and checking
- obtaining the required rheology.
(mainly from samples) that it complied with the expected
For the specific application of the BEAULIEU mine working
treatment : suitable rheology to pump the inertized bottom
backfilling, the basic INERTEC formulation for producing a
ash mortar from the stationary inertization unit to the most
inertized mortar from incinerator bottom ash had to be
remote backfill site (about 600 m) and respect of class V
adjusted to obtain a pumpable mortar able to be grouted
limits.
from the stationary inertization unit to the mine working. The bottom ashes aimed for this application were produced by
3.3 Full-Scale P r o d u c t i o n
the municipal solid waste incinerator of Caen and were
A full-scale production plant was set up on the Beaulieu
classified as class M because of solubles, lead and total
development site over the workings to be supported. It is
organic carbon in the leachate. The mix formulation was also
shown schematically in the following figure.
defined in laboratory to have sufficient strength (and pumpability) and comply with the class V limits.
~rap rlmov/Bid ~ ~==h
~ P
J~
J
/,,_ ~
IIII -"
un~
Mortarpump
|^
Soluble= (s) ]
I
('lass v hmJt i Class
X hm)t
9
Sample I
9
Sample2
9
Sample4
As (mu/ku)
Figure 2 : Leaching test results before treatment Solubles
SOLETANCHE instigated a Quality Assurance Plan for the duration of the contract (one year from february 1996) to
(%)
TOC (mg/kli)
ne~zedboeoma=h
Figure 4 : Treatment plant
control impacts from the work on the environment.
Pig (mg/kg)
l
~
lnertlzed bottom ash Class v hmit
(ma/ka)
Figure 3 : Leaching test results after treatment
Untreated bottom ash were delivered directly from the incinerator. Each lorry load was sampled, and, as well as recording color, smell and other visual aspects, the site laboratory ran a quick leaching test to determine pH and solubles content. This test ensures that the load is not
119 category S material : during all the work, measured solubles
4. Assessment program : application of the X30-407
were between 4% and 7%, well below the class M 10% limit.
methodology
Besides these quick tests, other tests were performed
4.1 Description of the study
separately, internally and externally, on a sample taken by
The aim of the study is also to evaluate the leaching
the county laboratory.
behavior of inertized bottom ashes in the Beaulieu mine
After removing metal scrap, the bottom ash was screened to
workings. According to the X 30-407 methodology guideline,
produce an even grain size and stockpiled on a watertight
the program consists of different steps :
floor. Rain and wash water were collected in a pond and recycled as process water in the ash inertization unit. No
Description of the scenario
water has been discharged off-site. The water table (10 m
Previous
studies
have
shown
that
mechanical
and
below the lowest level of the mine workings) was monitored
geotechnical conditions can be neglected in the first step of
by taking samples from three piezometers around the site.
the methodology application : changes may principally have
The backfilling work was then made by phases.
a influence on the exchange with the leachant. Biological
The sorted and screened ash were fed into the mixer with the mix water and inertization reagents_ After a few minutes
alterations will probably be very limited because of pH -
conditions
(pH
~. 12,
generally
harmful
for
bacteria
mixing, the mortar was discharged and pumped directly into
development) and air lack. Hydrogeological and climatic
the mine. The mortar set into a low-permeability solid that
conditions will also be the most important factors in the
will not react with the environment. The quality of the final
leaching behavior evaluation. The scenario of inertized
material was monitored internally (Quality Assurance Plan)
bottom ash disposal can be described by the following
and by an outside body for each separate section of the
diagram:
work. The Quality Assurance Plan also covered noise and
Seepage, rain water
odors, which may seriously inconvenience nearby residents ~[p
throughout the duration of the job. As with the analytical tests, these nuisances were monitored separately, by site and by outside bodies. The Beaulieu mine stopping job can be summarized in a few figures: Daily capacity of screening unit : 120 t/day Daily capacity of inertization unit : 400 t/day Duration : 12 months (february 1996 to february 1997)
Mine working limits = exchange surface
~
~ , . ~.~.....-~, . ; ,
i" " ~,.
ater in equilibrium with lime carbonate rock
;
,-"=-.-
!'i!".i .... ~ : ~.", . . ~ v ' . ~ ,
, ,,,.. ~ . . , . .
---,~,-r.': ~ ' - . - , . ,
" " .",i ..... :.,
f,-
Water after contact with inertized bottom ashes
Water table feeding
Figure 5 : disposal scenario
Volume of workings to be backfilled (on the Beaulieu area): :
General informations have been collected from previous
40,000 m 3 (equivalent to the annual ash production from
studies (from mine working for example) and meteorological
Caen municipal solid waste incinerator plant).
office. Seepage rate can be estimated from rain and temperature measures. The soil above the inertized mortar
120 will be considered as saturated so that seepage water
monolithic sample to a continuous flow of distilled water,
comes directly on the inertized mortar.
made by evaporation of a constant water reserve. The original equipment has been modified so that water in
Description of the waste
contact with the sample is at room temperature.
Some properties of the inertized bottom ashes have been already studied in the laboratory test program before the
Modeling of leaching behavior, thanks to long term leaching
beginning of the full-scale work. They needed to be
tests.
completed for the long term leaching evaluation :
Water steam ~ C water
9characterization of bottom ashes : chemical variability according
to
quick
leaching
test
results,
grain
Distilled
size
water
distribution, total chemical composition...
Column
9characterization of inertized mortar : chemical variability
Cooling olumn
Siph
Tested sample
according to leaching tests results, strength variability to confirm the structural durability (already demonstrated on a
~
few samples), permeability and porosity to determine the
--Jl " Siph~ t -- -~ Intermediat flask
-1
/ \ '~ reserve 11 / ~ / Heatingdevice
water transport main regime, total chemical composition...
J
Figure 6 : modified "soxhlet" equipment
Experimental determination of leaching behavior This step consists of determining the relevant parameters influencing the leaching behavior:
results. A
9nature of the leachant in the specified scenario :water after contact with carbonate lime rock (through analysis of water collected in the Beaulieu mine workings and pH stat tests in laboratory on carbonate lime rock samples) 9physical parameters : they will be neglected in the first step of the methodology application because temperature and moisture variations are very limited in the working mine area (10 meters below the ground level). 9 mechanical
and geological
Experimental validation of leaching behavior with in situ
parameters will also be
neglected (structural integrity of the mortar is already checked in a previous step of the study). The leaching behavior will also be studied through long term leaching tests (several months) with modified "soxhlet" equipment (see fig. 6). This test consists of putting a waste
mine working
room
has been
chosen
in
agreement with the city of Caen and equipped in order to follow the potential release of pollutants from inertized bottom ashes during several years : 9horizontally laying of a drainage material on a first layer of inertized bottom ash (the chosen room was already partly backfilled and the set mortar laid about 2 meters from the roof of the mine working) to collect percolation water, 9vertically laying of a drainage material, along the wall, to collect runoff water. The room was then completely backfilled so that the drainage material is now in contact with the inertized bottom ash mortar. The drainage material collects water and directs it to collecting pipes (one for each system) which run across the room walls and arrive in the next room.
121 An additional piezometer with taking of cores has been dug
always below the class M limits and even sometimes below
after the setting of the mortar in the chosen room. Four
the class V limits.
piezometers will also be sampled at regular intervals to detect a potential release from the mortar. A
8
~6 =4
~
2
| ,%
[] []
i
[]
[]
[] O
[]
02-oct 22-oct 11-nov 01-ddc 21-d6c 10-jan 30-jan Sampling date Figure 9 : Solubles variability
I comple Watercollector Figure 7 : in situ instrumentation
A
~0
Planned Georges Pompidou Road -
, ~-
30
~- -..-:.:.~.
n
10
[]
o
o
[ m
0
[] O 0
[]
[]
[]
02-oct 22-oct 11-nov 01-d6c 21-d6c 10-jan 30-jan Sampling date
~l
'p
Figure 10 : Lead leachate content
r
An average leachate composition would be the following, Instrumented room (9A)
with arsenic, chromium, cadmium and mercury under the ICP detection limits :
Figure 8 : site geography
Parameter
Leachate content (X 31210)
4.2 F i r s t r e s u l t s
Water content (%)
81.1%
Solubles (%)
4.8%
SO4 (mg/kg)
1988
COT (mg/kg)
1187
Pb (mg/kg)
11.4
4.2.1 Laboratory tests
Non treated bottom ashes have been sampled weekly during several months, in order to follow the leachate variability and to determine an average leachate concentration. Arsenic, chromium, mercury and cadmium are seldom detected. Variability of pH of leachates and water content is not very important. Lead, total organic carbon and solubles are
Table 1 : Average leachate composition
Grain size analysis shows the apparent heterogeneity of the material. Largest grains are mostly glass particles and unburnt
residues
(paper...).
Glass
particles
will
be
considered as chemically inert for the leaching behavior
122 study. Chemical analysis do not allow to determine any clear
Calcium comes partly from bottom ashes (with a pH value
correlation between heavy metal content and particle size.
higher than 12).
Calcium seems yet slightly more concentrated in smallest particles.
After two months of long term leaching test (total water flow about 170 liters on a sample of 130 cm 2 surface, e.g. ratio
Concerning the mechanical properties of inertized bottom
liquid/solid ~ 22 m3/m2), the total extracted concentrations of
ashes, two parameters have been parallel followed with
heavy metals are very low (maximum 1.5% of total content),
curing time : compressive strength and water permeability.
so that analysis results are often below the detection limit.
The compressive strength increases between 28 and 90
A part of extracted heavy metals precipitates in the water
days while the water permeability decreases. After 90 days,
reserve (probably as calcium silicate components) and can
water permeability is about 7.10.9 m/s and the compressive
be only dissolved by fluoridric attack.
strength seems to reach an asymptote (see fig. 11).
The lead extraction diagram shows the total extracted
E :E ==
concentration and the cumulative concentrations in the clear
2
reserve water (before precipitate attack). In all leachate
1.5
samples, concentrations are yet very low, close to the ICP detection limits.
= 90.5
E
8
5
.....
0 20
40
60
Curing time (days)
80
100
}_
~
;
o
~r "/=
Figure 11 : compressive strength e v o l u t i o n
......
According to the french leaching test, heavy metals are not
0
10
20
30
40
i
50
Test day number
on a soluble form in inertized bottom ashes : they are not Figure 12 : lead extraction diagram
detected in leachates. The total content in the mortar is also very low: Heavy metals
Total content (mglkg)
As
< 17
Cd
6.6
Cr
147
Pb
535
T a b l e 2 : i n e r t i z e d b o t t o m a s h analysis
10.0 8.0 .. m
....-i
6.0 . . . . .
~
4.0 "~ 0
.......i "~
_~1"'"
2.0 0.0
0
10
20
30
40
Test day number
Lead has the highest concentration among heavy metals. Major components are silicium and calcium. A part of silicium content comes from glass pieces, which represent the major phase of largest particles in bottom ashes.
Figure 13 : calcium extraction diagram
50
123 Characteristic parameter of hydraulic binders matrix (like
the low permeability measured on inertized bottom ash
calcium) are parallel analyzed to determine the leaching
mortar samples. Two water samples have been collected
behavior of the material.
from the vertical drainage material. After contact with inertized mortar, water has a more basic pH and a higher
The extraction curves can't be assimilated to straight line
salt
with a 1/2 slope in logarithmic scales, which would be
chloride), but heavy metals concentrations are close to
characteristic of simple molecular diffusion. The leaching
detection limits. Two samples are however not sufficient to
mechanisms
dissolution-precipitation
make an accurate interpretation : other samples will be
(interactions with the matrix components) and diffusion
analyzed at regular intervals, when rainfall will be high
coupling, which are more difficult to model.
enough to have seepage (unlike in january and march).
result
from
concentration
(especially
sodium
and
potassium
Long term leaching tests are not yet stopped and studies are now going on to realize a modelisation
of leaching
mechanisms.
5. Conclusion The community benefits from the backfilling operations described, in two ways.
4.2.2 In situ results
First, there is an economic benefit. The saving on purchase
Water has been collected before the complete backfilling of
and haulage of borrow material to backfill the workings, and
the chosen room for experimental validation and analyzed to
disposal of class II ash from the Caen domestic refuse
determine the composition of water in contact with inertized
incinerator plant in landfills, offsets the costs involved in ash
bottom ashes. Main components are calcium, sodium and
re-use for scrap removal, screening and inertization.
magnesium, with a slightly basic pH :
There is also a direct benefit for the environment. Re-use of
Sample I
Sample 2
Sample 3
bottom ash frees a commensurate capacity at the Caen
pH
8,29
8,02
7,91
landfill for municipal waste that is not fit for re-use. At the
Conductivity (mS/cm)
0,807
0,839
0,842
same time, it avoids the need to quarry borrow materials
AI (mgll)
0,33
0,34
0,35
Ca (mg/I)
134,4
133,7
135,3
The assessment study, in application of X 30-407 guideline,
resources from the environment.
K (mg/I)
0,34
3,68
3,54
will allow to support this original valorization mean, that
Mg (mgll)
7,35
7,66
7,67
seems at first a safety recycling way for inertized bottom
Na (mgll)
13,72
18,64
18,31
ashes, according to the leaching exposure.
Si (mgll)
1,17
1,21
1,2
Table 3 : site water composition
Water
table
samples
thanks
to
piezometers
has
approximately the same composition. Concerning the drainage system, no water has been collected by the horizontal drainage material, which suits to
This Page Intentionally Left Blank
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
125
The Influence of Monolith Physical Properties and Integrity on Diffusional Leaching Behavior of Asphaltic Pavements Constructed with MSW Combustion Bottom Ash T. Taylor Eighmy, Douglas Crimi, Ingrid E. Whitehead, Xishun Zhang, David L. Gress Environmental Research Group A 115 Kingsbury Hall University of New Hampshire Durham, NH 03824 USA ABSTRACT The leaching of two municipal solid waste bottom ash constituents (relatively mobile C1 and relatively immobile Ca) was used to evaluate the effects of bituminous pavement physical properties (bottom ash substitution, asphalt cement content, voids, microcracking, bituminous polymer aging) and monolith integrity on diffusional leaching parameters such as cumulative release (Bt,cl, Bt,ca), tortuosity (,), chemical retention (Rcl, Rca), and effective diffusion coefficients (pDe,cl, PDe,c~). For intact laboratory monoliths, mix design parameters such as increasing asphalt cement content and increasing bottom ash substitution both were directly related to increases in Bt,cl and Bt,ca , and to increased ,. While increasing the bottom ash loading increased the driving gradient for chloride and calcium, the concomitant increase in asphalt cement content increased ,. For laboratory monoliths subjected to accelerated aging so as to predict future environmental performance, 9was found to be influenced only by monolith microcracking. For pavements subjected to field aging in a test road, no differences could be seen between cores evaluated just after road construction and cores evaluated after 1.25 years of service life. During the 1.25 year study period, some beneficial increases in tortuosity, probably from densification, occurred. For compacted granularized pavements where monolith integrity is lost, the presence of asphalt cement in the bottom ash pore structure and on the particle surface still limited diffusional release. The tortuous, hydrophobic nature of the intact monolith system controls diffusion at the macroscale level. The presence of asphalt cement on the particle surface and in the internal pore structure of the bottom ash particles also controls diffusion at the microscale level, even in compacted granularized specimens. INTRODUCTION The life cycle of a waste material used in the construction of a pavement may initially involve a primary use (e.g. as an aggregate substitute in a bituminous wearing or binder course monolithic pavement) as well as a secondary use (e.g. as reclaimed asphalt pavement use in compacted granular base course). Knowledge as to the environmental behavior of the product as well as the leaching of waste constituents is necessary. The physical properties and integrity of the pavement can play an important role in the type of leaching (wash off, dissolution, diffusion) that can occur. These properties also control the magnitude of the fluxes of waste constituents from the pavement. Fluxes are typically small in asphaltic pavement monoliths 1,2,3. Flux
126 measurements can be useful for source term estimates for health risk assessments, waste utilization application evaluations, and regulatory approval. Figure 1 shows possible leaching scenarios related to waste utilization in pavement construction 4. The very low specific surface area, low permeability, and high structural integrity of the monolith or densely compacted granular material results in leaching regimes where diffusion is the dominant mass transfer mechanism. This differs dramatically from loose, uncompacted granular materials where solvent percolation and advection produce contaminant fluxes that are orders of magnitude larger than diffusional fluxes. Monolith diffusional leaching (Figure 2a) involves diffusion into the pavement of solvents such as H20 and diffusion out of the pavement of dissolving solutes such as cationic and anionic species. Diffusion occurs by tortuous pathways within the product, frequently at bituminous polymer/waste particle interfaces, within the bituminous polymer, and within the waste particles themselves. However, the integrity of the product provides a macroscale control on diffusional fluxes 2. At the macroscale level, the pavement bulk density, air voids, waste substitution rate, asphalt cement content, and monolith integrity will all impact monolithic diffusional leaching and contaminant fluxes from the monolith. For compacted granular diffusional leaching scenarios (Figure 2b), the relatively higher surface area, higher permeability, and lack of structural integrity result in leaching regimes where diffusion is the dominant mass transfer mechanism, but diffusion pathways are much less tortuous. Diffusion occurs by tortuous pathways within the waste particles, and between the waste particles and the residual asphalt cement within the pore structure of the waste particles and coating the particles. However, the particle and asphalt cement coatings provide a more microscale control on diffusional fluxes i. At the microscale level, the residual asphalt cement content, waste particle surface area, and waste particle surface chemical speciation will all influence granular diffusion. For bituminous asphalt pavements constructed as wearing, binder or base course, the inplace hydraulic regime is not well defined. For intact pavements, percolation regimes are unlikely. Rather, contact with water is at monolith exterior surfaces, either by periodic precipitation, localized intermittent groundwater, or periodic condensation of soil water vapor. Assumptions can be made about the types of surfaces that are wetted and the wetting/dying cycles that the surfaces are exposed to 3,5. The hydrophobic nature of the asphalt cement complicates the interpretation of water uptake during wetting, water absorption, and reaction with the waste particle 6. Nevertheless, from a conceptual approach, diffusional leaching and cumulative release have been verified for coal fly ash constituents in bituminous shoreline protection monoliths 7, nuclear wastes in bituminous asphalt-stabilized monoliths 6, and pavement material made with MSW bottom ash 3,5. As part of our efforts to understand the environmental performance of bituminous pavements made with MSW bottom ash, we have examined aspects of monolith properties (e.g. waste substitution, asphalt cement content) and monolith aging (voids, microcracking, bituminous polymer aging, field aging) on monolithic diffusional leaching as well as granular properties (particle size) and asphalt/particle integrity (presence of asphalt cement) on compacted granular diffusional leaching. These scenarios constitute major primary and secondary uses for bottom ash in bituminous pavement. Within each scenario, a continuum between early and later stages of the life history of the product are also provided. Finally, evaluations are provided for
127
both bottom ash (consisting largely of grate ash and small quantities of grate siftings) as well as grate ash; the environmental properties and performance of both these materials being synonymous 3. M O N O L I T H DIFFUSIONAL L E A C H I N G FUNDAMENTALS Figure 2a depicts the fundamental processes involved in leaching from monolithic specimens. The figure depicts the diffusive release of constituents via pathways of varying tortuosity, long term dissolution of solid phase material at the monolith surface, and initial washoff of material from monolith surface. These phenomena exhibit different cumulative release behaviors, as evidenced by plot slopes, on log-log cumulative release versus time plots. Diffusion occurs by the random movement of individual molecules or ions. It is driven by the difference in chemical potential between the solid and the leachant 8. In a slightly porous solid, the ion flux of a soluble contaminant in the pore water system is defined by Fick's Second Law, which relates the concentration of a diffusing substance to both space and time: dC dt
- L
d 2C dx 2
(I)
where C is the concentration of the diffusing ion, and L is the leach constant, with units of a diffusion coefficient (mE/s). In the case of one dimensional diffusion, the leach constant, L, is obtained by applying the following relationship: L
f2.D ~
R'I;
(2)
where f is the available leachable fraction of the element in the material, D o is the mobility of the element in water (mE/s), R is the chemical retention factor of the element (unitless), and z is the physical retardation or tortuosity factor (unitless). When the leachable amount of the element equals the total amount of the element present in the material (i.e., f= 1), the leach constant L equals the effective diffusion coefficient, which is modified for retention and tortuosity 9. If Equation 2 is transformed to logarithmic values and using the relationship pDo = -log Do, then the new relationship is: pL = POo+ 2pf- pR - px
(3)
Since the available leachable fraction of the element may be obtained from analytical data and remains as a constant, Equation 3 may be simplified to: pD = pD~ - pR - pz where D e is the effective diffusion coefficient (mE/s) 9. This equation allows for the determination of the relative contributions of D o, R, and z to the magnitude of D e. As will be
(4)
128 seen, the relative contributions of R and x play the dominant role in the magnitude of D e as well as in the cumulative flux or extent of the source term leaching from the monolith. The relative magnitude of R and x, and their change as a function of monolith properties and aging is crucial to understanding long term environmental behavior. The mobility of an element within the monolithic pore space may be compared to its free mobility in water by determining the physical and chemical retardation factors of the product and the element, respectively. To calculate the physical retardation factor, it is necessary to choose an ion which does not chemically interact with the matrix (i.e., R= 1). In most studies, sodium is chosen and the physical retardation is calculated with the formula 9. :
DN/D,,N.
(5)
where DNa is the diffusion coefficient of sodium in water (PDNa-- 8.88 m2/sec at 25 ~ and De,Na is the effective diffusion coefficient of sodium in the monolith (m2/s). The chemical retardation factor for the element is determined using the following formula: R : D/(D,~ x)
(6)
where D• is the diffusion coefficient for the element in water (m2/s) and De,x is the effective diffusion coefficient for the element in the monolith (mE/s). In monolith leach testing, the material is considered homogeneous and is immersed in a leachant which is renewed at regular intervals. The concentration of the element is uniformly distributed and the surface is maintained at a constant concentration. The solution to Equation 1 under such conditions is presented by Crank 10. C - C, Co -
C,
x
-- err
2~f~-~
(7)
where C is the element's concentration in the monolithic material as a function of place and time, C1 is the constant concentration at the surface of the monolith, Co is the initial concentration of the element in the material, D is the diffusion coefficient and err is the standard error function. Since the leachant is regularly renewed during the leaching experiment, the surface concentration, C~, is assumed to be zero 9, and the solution to Equation 7 under this assumption is: O :
9 Bt 2
4t(U~jd) ~
where B t is the cumulative release of the element at time t (mg/m2), U m a x is the maximum leachable quality of the element from the monolith (mg/kg) and d is the bulk density of the monolithic material (kg/m3).
(8)
129 By plotting the cumulative release of an element leaching from the monolith as a function of time, it can be determined whether matrix diffusion or other leaching mechanisms, such as dissolution or surface wash-off, more common to granular materials, are occurring. The solubility of certain elements within the solid material can be significantly high, such that longer term dissolution of the element at the surface proceeds faster than diffusion through the pores of the solid matrix. Additionally, due to process conditions, a material may be covered with a soluble surface coating that is readily leached with initial leachant contact. To determine which leaching mechanism is controlling release of the element from the monolith, Equation 8 is rearranged to yield the following 9: Bt =
Ureax dv/(aD/7~)V/I
(9)
After log transformation, Equation 9 becomes: log(Bt) = 1/2 log(t)+ l o g [ U d~/(4D/r~)]
(10)
From the monolith leach test results, the release of the element per time interval may be calculated using the formula 9: c~ V~ A
(11)
Bi = 1 0 0 0
where B i is the release of the element per unit area in period i (mg/m2), c i is the concentration of the element in the ith period (~tg/L), V i is the volume of the leachant (L), and A is the surface area of the monolithic material (m2). The cumulative release of the element for all N periods (N=8) is calculated from: Bti-- Bi
V/t
for i = 1 to N
(12)
where Bt, i is the cumulative release of the element for all periods (rag/m2), t i is the contact time after period i (seconds), and ti.1 is the contact time after i-1 periods (seconds) 9. Eight leaching periods are usually used (0.25, 1, 2, 4, 8, 16, 32, and 64 days). This conforms to a time series relationship (renewal time of the nth period is equal to the square of the period sequence number times the renewal time of the first period) that is based on both leaching behavior and on data distribution in log(Bo) - log(ti) plots. If the logarithm of the cumulative release, Bt, i, is plotted versus the logarithm of time, t i, for the eight periods, the slope of the resulting graph indicates the mechanism controlling the release of the element. Since the slope may change over different time intervals, the slope is examined over the ranges 9. The effective diffusion coefficient for the element is then calculated
130 from each period for the release per period (Bi) using the data points where the slope is 0.50 + 0.15 with a deviation of less than 50% and the slope of the final range is smaller than 0.65 by: 71: Bi 2
Do,~x --
4(U~,~d)2" (~i - t~i_~)2
(13)
where De,i, x is the effective diffusion coefficient of element x calculated from the release in the ith period (m2/s), and the other terms are as previously described. COMPACTED GRANULAR DIFFUSIONAL LEACHING Figure 2b depicts the fundamental processes involved in diffusional leaching from compacted granular specimens. The figure depicts the diffusive release of constituents via pathways of varying tortuosity, long term dissolution of solid phase material at the particle surface, and initial wash-off of material from the particle surface. The same principals for examining release mechanisms, cumulative fluxes, effective diffusion coefficients, tortuosity, and chemical retention in monoliths can be applied to compacted granular specimens provided that they are leached in static, renewal-based leaching tests. This modification to the monolith leaching test was developed by Kosson et al. 1~. MATERIALS AND METHODS The focus of this paper is on the use of municipal solid waste combustion bottom ash from the Concord, New Hampshire waste-to-energy facility as an aggregate substitute in a primary application in bituminous binder course pavement and in a secondary application as stabilized base course pavement. A great deal of information has already been presented on the use of bottom ash from the Concord facility in asphalt pavements ~,12,13,14,as well as the Laconia, New Hampshire bottom ash utilization demonstration 3,15,16. The purpose of this research was to identify factors controlling diffusional leaching behavior in monolithic and compacted granular applications, with focus on two elements that exhibit different diffusive leaching behaviors from the MSW bottom ash pavement products; chloride and calcium. The former is considered to be relatively diffusive and mobile because it will not readily sorb or precipitate. The later is less diffusive and more immobile because it can readily sorb or precipitate. The research was conducted in four parts: (i) the use of laboratory monoliths to evaluate the effects of bottom ash substitution rates and asphalt cement content on monolith diffusional leaching, (ii) the use of laboratory monoliths to study the impacts of air void content, microcracking, and accelerated bituminous polymer aging on monolith diffusional leaching, (iii) the use of field cores from the Laconia, New Hampshire bottom ash utilization field demonstration to examine short term field aging on monolith diffusional leaching, and (iv) the use of laboratory granularized monoliths and granular materials on compacted granular diffusional leaching. For the first set of experiments, the Dutch monolith leach test (NEN 7345) was used to evaluate the effects of bottom ash substitution and asphalt cement content on monolithic diffusional leaching. Cumulative release (Bt,ci , Bt, fa), monolith tortuosity (~), chemical retention
131
(Rca , Rcl ), and effective diffusion coefficients (pDe,o, pDe,ca ) were used as response variables. Monoliths were made using Marshall mix design procedures 12,13. Three bottom ash substitution ratios were used (25,50 and 75%). The 50% substitution was made in duplicate. A variety of AC20 asphalt cement contents were evaluated (e.g. 5 to 12%) as part of the mix design procedure. In all, six monolithic cylinders (10 cm diameter) were made. Monoliths at optimum asphalt cement contents were tested for monolithic leaching properties. For the second set of experiments, the Dutch monolith leach test (NEN 7345) was again used to evaluate the effects of monolith void content, aging, and microcracking on monolithic diffusional leaching 2. Cumulative release (Bt,cl, Bt,ca), monolith tortuosity (~), chemical retention (Rca , RCI ), and effective diffusion coefficients (PDe,cl , PDe,ca ) were used as response variables. A partial 33 + 6 factorial design experiment was conducted. A gyratory test method (GTM) mix design procedure was used 2. Grate ash was substituted 50% for natural aggregate. An asphalt cement content of 7% AC-20 was used. Three degrees of air voids (low, medium, high), aging (none, medium, high), and microcracking (none, medium, high) were integrated into the experimental design. Air voids (5%, 7.5%, 10%) were created during the compaction of the specimens using the GTM. The 10%, 7.5% and 5% air voids were produced using 60, 110 and 300 cycles, respectively on the GTM. Aging was simulated by heat treatment in the presence of forced hot air (107~ for 5 days for moderate aging, 107~ for 10 days for severe aging). The regimen was based on methods compiled by von Quintus et al. 17. An Instron machine was utilized to subject some of the samples to cyclical loading parallel to the direction of compaction to produce microcracking (i.e. moderate and high cracking levels). A load of 8.22 MPa at 2 cycles per second was used. Moderate microcracking was produced after 1,250 + 250 cycles. High microcracking was produced after 1,750 + 250 cycles. A total of 14 10 cm cylindrical specimens were made. For the third set of experiments, 10 cm road cores from the test and control sections of the Laconia, New Hampshire bottom ash paving demonstration were collected and subjected to the same Dutch monolith leaching procedure (NEN 7345) to evaluate the effects of short term field aging on monoliths leaching properties 3. Cores taken just after road construction and after 1.25 years of service were evaluated. The test section binder course was constructed using a GTM mix design procedure. A 50% grate ash substitution was used with an asphalt cement content of 7%. The binder course was subjected to field compaction methods. A total of four cores were collected from each section at each sampling time. After coring, the binder course was manually excised from each core for testing. For the fourth set of experiments, a compacted granular diffusion leaching test was conducted in an identical manner to the NEN 7345 except that compacted granular specimens were used. The method was developed by Kosson et al. 11. Specimens were compacted in HDPE beakers to bulk densities of 1,500 to 1,675 kg/m 3 the beakers were then suspended in the leachant at a L/S of 20 to facilitate setup and avoid reverse gradient effects. Three sets of samples were run in duplicate: <300 um ground grate ash, <1.9 cm ground grate ash, and < 1.9 cm ground pavement monoliths made with 50% grate ash and 75 AC-20 asphalt cement. These specimens were designed to mimic the use of granularized binder course in a secondary use as a compacted base course 3. The measure of surface area available for leaching was the area of the open (but submerged) top of the beaker. A 2 cm thick layer of glass beads (6 mm diameter) was used to minimize agitation.
132 For the NEN 7345 monolithic and granular procedures, the leachant was Nanopure| ASTM type II (double-deionized) water reduced to pH 4 using Baker Analyzed| Ultrex II| Ultrapure nitric acid. The leachate was filtered and analyzed periodically (0.25, 1, 2, 4, 8, 16, 32, 64 days) to determine specified element concentrations leached per indicated time frame. After each filtration, new contact solution was added to the sample. Samples were analyzed using ion chromatography (IC), graphite furnace atomic absorption spectrophotometry (GF-AAS) and inductively coupled argon plasma atomic emission spectrometry (ICAP-AES). Monolith specimens were also ground to less than 300 ~tm and subjected to the Dutch total availability leaching procedure (NEN 7341) to determine Umax,the fraction available for leaching. RESULTS AND DISCUSSION There are approximately 45 elements in the Concord grate or bottom ash 13. Although 45 elements are present, only 16 consistently leached from specimens ground to < 300 ~tm when subjected to the total availability leach test: C1, Ca, Zn, Cd, Mg, Cu, Mn, Pb, Sr, Si, Fe, A1, Na, K, Ba, Cr 13. Lysimeter leaching data for the granular (< 1.9 cm) bottom ash lysimeter show that the bottom ash leaches C1, SO42-, Ca, K, Mg, Na, Fe, Mn, Si and Sr i. Similar constituents leached from a lysimeter containing pavement rubble containing bottom ash, but at much lower levels i. Only seven constituents (Na, Cl, SO42, Ca, Si, Mg, and Zn) routinely leached from the grate or bottom ash test specimens during the monolith leach test 3.13.
Influence of Bottom Ash Substitution and Asphalt Cement Content Marshall mix design procedures were used to identify optimum asphalt cement contents for various levels of bottom ash substitution in binder course. The environmental performance of these monoliths were evaluated. This evaluation is relevant to understanding the possible performance of binder course based on mix design formulation. For instance, increasing the ash substitution percentage may increase the ash surface area available for chemical retention and increase the need for asphalt cement 12,14 which increases tortuosity; both of which are beneficial. Chloride cumulative release (Bt,o), calcium cumulative release (Bt, ca), and tortuosity were influenced by asphalt cement content (Figure 3); though confounding with bottom ash substitution is possible (see below). Generally, as asphalt cement content increased, cumulative fluxes increased even though tortuosity increased. The observed increases in cumulative release are explained by the higher percent bottom ash substitutions that occurred concomitant with increased asphalt content (Figure 4). This is intuitive given the fact that increased ash loadings increase the driving force promoting diffusion of chloride and calcium. However, these increased fluxes are not considered deleterious given their only modest increase over the range in ash loadings that were used. Influence of Accelerated Aging: Voids, Microcracking, and Bituminous Polymer Aging An experimental design was used to study the effects of accelerated aging (decrease in air voids from traffic densification over time, increase in microcracking from cyclical mechanical stress, bituminous polymer aging from oxidation) on the environmental performance of binder course pavement monoliths. Such an evaluation is useful in predicting the long term behavior of the application and in providing source term estimates.
131
(Rca , Rcl ), and effective diffusion coefficients (pDe,o, pDe,ca ) were used as response variables. Monoliths were made using Marshall mix design procedures 12,13. Three bottom ash substitution ratios were used (25,50 and 75%). The 50% substitution was made in duplicate. A variety of AC20 asphalt cement contents were evaluated (e.g. 5 to 12%) as part of the mix design procedure. In all, six monolithic cylinders (10 cm diameter) were made. Monoliths at optimum asphalt cement contents were tested for monolithic leaching properties. For the second set of experiments, the Dutch monolith leach test (NEN 7345) was again used to evaluate the effects of monolith void content, aging, and microcracking on monolithic diffusional leaching 2. Cumulative release (Bt,cl, Bt,ca), monolith tortuosity (~), chemical retention (Rca , RCI ), and effective diffusion coefficients (PDe,cl , PDe,ca ) were used as response variables. A partial 33 + 6 factorial design experiment was conducted. A gyratory test method (GTM) mix design procedure was used 2. Grate ash was substituted 50% for natural aggregate. An asphalt cement content of 7% AC-20 was used. Three degrees of air voids (low, medium, high), aging (none, medium, high), and microcracking (none, medium, high) were integrated into the experimental design. Air voids (5%, 7.5%, 10%) were created during the compaction of the specimens using the GTM. The 10%, 7.5% and 5% air voids were produced using 60, 110 and 300 cycles, respectively on the GTM. Aging was simulated by heat treatment in the presence of forced hot air (107~ for 5 days for moderate aging, 107~ for 10 days for severe aging). The regimen was based on methods compiled by von Quintus et al. 17. An Instron machine was utilized to subject some of the samples to cyclical loading parallel to the direction of compaction to produce microcracking (i.e. moderate and high cracking levels). A load of 8.22 MPa at 2 cycles per second was used. Moderate microcracking was produced after 1,250 + 250 cycles. High microcracking was produced after 1,750 + 250 cycles. A total of 14 10 cm cylindrical specimens were made. For the third set of experiments, 10 cm road cores from the test and control sections of the Laconia, New Hampshire bottom ash paving demonstration were collected and subjected to the same Dutch monolith leaching procedure (NEN 7345) to evaluate the effects of short term field aging on monoliths leaching properties 3. Cores taken just after road construction and after 1.25 years of service were evaluated. The test section binder course was constructed using a GTM mix design procedure. A 50% grate ash substitution was used with an asphalt cement content of 7%. The binder course was subjected to field compaction methods. A total of four cores were collected from each section at each sampling time. After coring, the binder course was manually excised from each core for testing. For the fourth set of experiments, a compacted granular diffusion leaching test was conducted in an identical manner to the NEN 7345 except that compacted granular specimens were used. The method was developed by Kosson et al. 11. Specimens were compacted in HDPE beakers to bulk densities of 1,500 to 1,675 kg/m 3 the beakers were then suspended in the leachant at a L/S of 20 to facilitate setup and avoid reverse gradient effects. Three sets of samples were run in duplicate: <300 um ground grate ash, <1.9 cm ground grate ash, and < 1.9 cm ground pavement monoliths made with 50% grate ash and 75 AC-20 asphalt cement. These specimens were designed to mimic the use of granularized binder course in a secondary use as a compacted base course 3. The measure of surface area available for leaching was the area of the open (but submerged) top of the beaker. A 2 cm thick layer of glass beads (6 mm diameter) was used to minimize agitation.
134 and after 1.25 years of service life (time = 1.25 years). Cores were taken from slabs removed at the pavement margin where traffic loads were light and where any aging would be related to the natural densification or aging. The purpose of this effort was to see if any shorter term aging phenomena could be observed. The data for chloride leaching is presented in Table 1. Generally the test section monoliths containing the 50% grate ash substitution and the 7.0 % asphalt cement content behaved like intact laboratory monoliths with low cumulative releases, low chemical retentions, high tortuosities, and low diffusivities. Student's t-test (95% confidence level) was used to statistically evaluate aging. No significant differences were seen with respect to short term aging in the field and chloride diffusional behavior. There were differences between the test and control sections; largely because there is no significant level of leachable chloride in the natural aggregates used in the control pavements. Tortuosity in both the test and control sections significantly increased over time (Student's-t, one tail, 95%); probably from natural densification processes. The data for calcium leaching is presented in Table 2. Calcium leaching in the test section behaved like intact laboratory monoliths. There were significant differences (Student's-t, one tail, 95%) in the effective diffusion coefficient (PDe,ca); calcium becoming less diffusive with time; largely from statistically significant increases in tortuosity. While cumulative fluxes were significantly different between the test and control sections, effective diffusion coefficients (pDe,c~) and chemical retention values (Rc~) were not significantly different. Overall, there were no deleterious significant changes in the test or control monoliths over the 1.25 year study period. Modest increases in tortuosity from natural densification explain the observed changes. Influence of Monolith Integrity The compacted granular leaching test was used to evaluate the effect of loss of monolith integrity (via granularization) on environmental behavior. Three increasing levels of loss of monolith integrity were investigated; low granularization of binder course monoliths (<1.9 cm) so that the resulting coarsely ground ash particles are coated both within the particle pore structure and on the particle exterior with asphalt cement binder, medium granularization using coarsely ground bottom ash particles (<1.9 cm) without asphalt cement, and high granularization using finely ground bottom ash particles (<300 um). This evaluation is useful in better understanding the possible environmental performance of a recycled binder pavement containing ash in a secondary application as a compacted granular base course. Such an evaluation can also provide source term estimates. As shown in Figure 8, the cumulative release of chloride (Bt,c0 and the cumulative release of calcium (Bt,ca) both dramatically increased as granularization increased. This would be expected as tortuosity would likely decrease as diffusion path lengths are drastically shortened. Chloride cumulative releases ranged from 70 to 40,000 mg/m 2. Calcium cumulative releases ranged from 10,000 to 50,000 mg/m 2. These cumulative releases from granular diffusional leaching are orders of magnitude larger than monolithic diffusional leaching. The level of granularization dramatically affected tortuosity (Figure 9a). As granularization increased, tortuosity values decreased from rather low levels (80) down to very low levels (3). These tortuosity values are much less than those seen for intact monoliths. This is
135 expected as granularization would remove macroscale control and shorten diffusion path lengths. Chloride, which is not very reactive chemically with respect to sorption or precipitation, did not exhibit any correlation (Figure 9b) between microcracking and chemical retention (Rcl). The effective diffusion coefficient for chloride (PDe,cl) modestly increased with increasing granularization (Figure 9c); largely because of the decrease in tortuosity (Figure 9a) and the link between tortuosity and diffusivity (see Eq. 4). The diffusivity of the chloride is still higher than those seen in the intact monolith; however, the values are still much less diffusive than a freely mobile anion. The chemical retention of calcium (Rca), which is reactive chemically with respect to sorption or precipitation, was correlated with increased granularization (Figure 10b). At high levels of granularization, Rca values increased from 50 to about 250. These values are significantly higher than those seen for calcium in intact monoliths. This increase from granularization is attributed to either sorption on the ash particle surfaces or precipitation on the ash particles surfaces. Despite significant increases in Rca, the effective diffusion coefficient for calcium (pDe,ca) increased with increasing granularization (Figure 10c); largely because of the decrease in tortuosity (Figure 7a) and the link between tortuosity and diffusivity (see Eq. 4). Even those these values are low relative to the monolith scenario, they are still significantly less diffusive than a freely mobile cation. Influence of Ionic Radii and Chemical Reactivity of the Diffusing Constituents There are some interesting observations that can be made between the relative diffusional behaviors of chloride and calcium in each of the four experiments. Chloride has a larger ionic radius ( 1.81 A) and diffusivity or free mobility in water ( 2.03 x 10.9 m2/s, pDo = 8.69) than calcium (1.00 A, 7.93 x 10-10 m2/s, pD o = 9.10) 19,20.Chloride is considered to be a conservative constituent; it is highly mobile and does not readily participate in sorption, ion exchange or precipitation reaction 21. It will form strong complexes with certain metals (e.g. Zn, Cd, Hg, Pb) 21; however, these are not prevalent in bottom ash leachates 1,3.13.Calcium is considered to be less conservative and more reactive. It is less mobile and readily participates in sorption, ion exchange, and precipitation reactions though these are pH-dependent 21. It also forms large complexes such as ion pairs like CaSO4 0, CaeO4 , and CaCO3 ~ 21. More mobile and less reactive chloride will not be as responsive to 9 and have lower R values. Less mobile and more reactive calcium will be more responsive to 9 and have higher R values. These differences highlight the contributions of 9 and R to the effective diffusion coefficient as shown in Equation 4. Generally, the chloride anion was more diffusive and less reactive than the calcium cation in the four experiments. This is expected given the properties of the two ions diffusing in a nonreactive, hydrophobic bituminous polymer matrix. However, with increasing loss of monolith integrity via microcracking or granularization, the loss of the tortuosity was made up by increased chemical reactivities within or upon the ash particle surfaces. Chloride did not show this behavior, because it is less reactive than calcium. These data suggest that other more reactive ash constituents (Si, Mg, Zn ) would behave similarly to calcium and less reactive constituents (Na, SO42) would behave similarly to chloride. Generally, this has been observed 3.
136 CONCLUSIONS In terms of pavement formulation, increased asphalt cement content in the mix design is beneficial with respect to increasing monolith tortuosity at both the macroscale and microscale levels. Higher levels of asphalt substitution, requiring additional asphalt cement content to fill internal bottom ash pore structures, also increase the driving force and therefore the cumulative flux of constituents leaching from monoliths, but not in a deleterious manner. In terms of accelerated monolith aging, microcracking was the only experimental variable influencing diffusional leaching behavior. Increased microcracking decreased diffusional path lengths whereby release occurs and therefore reduced tortuosity. Short term field aging in a bottom ash pavements did not produce any marked changes in the behavior of chloride or calcium in field cores. In fact, some beneficial increases in tortuosity occurred during the 1.25 year study period. If monoliths are granularized, the presence of asphalt cement in the internal pore structure of the bottom ash particles and on the particle surfaces still constitutes an effective control mechanism. The results suggest that MSW bottom ash constituent leaching behavior is controlled by the tortuous hydrophobic nature of the bituminous polymer system in the pavement. ACKNOWLEDGEMENTS This research was supported by Wheelabrator Environmental Systems Inc., the Concord Regional Solid Waste/Resource Recovery Cooperative, the U.S. EPA via Rutgers University, and the U.S. DOE via the National Renewable Energy Laboratory. Doug Crimi presently works for the Bureau of Land Management in Juneau, Alaska. Ingrid Whitehead works for Roy F. Weston Inc. of Concord, New Hampshire. Xishun Zhang works for Heritage Corp. in Indianapolis, Indiana. REFERENCES 1. Whitehead, I.E., Eighmy, T.T., Gress, D.L. and Zhang, X. An Environmental Evaluation of Bottom Ash Substitution in Pavement Materials. In Municipal Waste Combustion, Air and Waste Management Association, Pittsburgh, Penn., p. 356 (1993). 2. Eighmy, T.T., Crimi, D., Hasan, S., Zhang, X. and Gress, D.L. Influence of void change, cracking, and bitumen aging on diffusional leaching behavior of pavement monoliths constructed with MSW combustion bottom ash. Trans. Res. Rec. 1486:42 (1995). 3. Eighmy, T., Gress, D., Crimi, D., Hasan, S., and Karpinski, S. The Laconia, New Hampshire Bottom Ash Paving Report, NREL/TP-430-20959, National Renewable Energy Laboratory, Golden, CO (1996). International Ash Working Group. An International Perspective on Characterization and Management of Residues from Municipal Solid Waste Incineration. ECN, Petten, the Netherlands, (1994). Kosson, D.S., van der Sloot, H.A., and Eighmy, T.T. An approach for estimation of contaminant release during utilization and disposal of municipal waste combustion residues. J. Haz. Matl. 47:43 (1996). Fuhrmann, M., Pietrzak, R.F., Franz, E.M., Heiser, J.H., and Colombo, P. Optimization of Factors That Affect Leaching, BNL-52204, Brookhaven National Laboratory, Upton, New York (1989). .
137
10. 11.
12.
13.
14.
15.
16.
17.
18.
19. 20. 21.
van der Wegen, G. and van der Plas, C. Validation of leaching model on actual structures. In (J.J.J.R. Goumons, H.A. van der Sloot, and Th.G. Albers, eds.) Waste Materials in Construction, Elsevier, Amsterdam, the Netherlands, p. 55 (1991). Conner, J.R. Chemical Fixation and Solidification of Hazardous Wastes. Van Norstrand Reinhold, N.Y., (1990). de Groot, G.J. and van der Sloot, H.A. Determination of Leaching Characteristics of Waste Materials Leading to Environmental Product Certification. In (T.M. Gilliam and C.C. Wiles, eds.) Solidification~Stabilization of Hazardous, Radioactive and Mixed Wastes, ASTM, Philadelphia, Penn., p. 149 (1992). Crank, J. The Mathematics of Diffusion, Oxford University Press, Oxford, England, (1975). Kosson, D.S., Kosson, T.T., and van der Sloot, H. A. Evaluation of Solidification~Stabilization Treatment Processes for Municipal Waste Combustion Residues. NTIS PB93-229 870/AS (1993). Gress, D.L., Zhang, X. Tarr, S., Pazienza, I. and Eighmy, T.T. Municipal Solid Waste Combustion Ash as an Aggregate substitute in Asphaltic Concrete. In (J.J.J.R. Goumons, H.A. van der Sloot, and Th.G. Albers, eds.) Waste Materials in Construction, Elsevier, Amsterdam, the Netherlands, p. 161 (1991). Eighmy, T.T., Gress, D.L., Zhang, X., Tarr, S. and Whitehead, I. Bottom Ash Utilization Evaluation for the Concord, New Hampshire Waste-to-Energy Facility. University of New Hampshire, Durham, N.H., (1992). Zhang, X., Gress, D. and Eighmy, T. Bottom Ash Utilization as an Aggregate Substitute in Hot Mix Asphalt. Proc. 2nd Annual Great Lakes Geotechnical/ Geoenvironmental Conference, p. 132 (1994). Musselman, C.N., Killeen, M.P., Crimi, D., Hasan, S., Zhang, X., Gress, D.L. and Eighmy, T.T. The Laconia, New Hampshire Bottom Ash Paving Project. In (J.J.J.R. Goumons, H.A. van der Sloot, and Th.G. Aalbers, eds.) Waste Materials in Construction, Elsevier, Amsterdam, the Netherlands, p. 315 (1994). Musselman, C., Eighmy, T., Gress, D., Killeen, M., Presher, J., and Sills, M. The New Hampshire Bottom Ash Paving Demonstration, US Route 3, Laconia, New Hampshire. Proc. 1994 ASME National Waste Processing Conference, 16th Biennial Conference, ASME, N.Y., N.Y., p. 83 (1994). von Quintus, H.L., Scherocman, J.A., Hughes, C.S., and Kennedy, T.W. AsphaltAggregate Mixture Analysis System. National Cooperative Highway Research Program Report 338, Transportation Research Board, National Research Council, Washington, D.C. (1991). de Groot, G.J., van der Sloot, H.A., Bonouvrie, P. and Wijkstra, J. Karacterisering van het Uitlooggedrag, van Intacte Producten. MAMMOET deelrapport 09, ECN-C-90-007, Petten, the Netherlands (1990). Shannon, R.D. Revised effective ionic radii and systematic studies of inter atomic distances in halides and chalcogenides. Acta Crys. A32, 751 (1976). Li, Y.H., and Gregory, S. Diffusion of ions in sea water and deep-sea sediments. Geochim. Cosmochim. Acta 38, 703 (1974). EPRI. Inorganic and Organic Constituents in Fossil Fuel Combustion Residues, EPRI EA-5176, EPRI, Palo Alto, California (1987).
138
Table 1:
Chloride Leaching From Laconia, New Hampshire Bottom Ash Utilization Demonstration Road Cores
Pavement Type and Age
Bt, cl
RCI
1;
pDe.cl
mg/m 2
-
_
m2/s
930-2,540
1.1-7.4
2,2008,800
12.1-13.4
Test Section (Time = 1.25 years)
230-520
0.66
13,800
12.6
Control Section (Time= 0 years)
no release
-
6,000
30-160
-
13,800
Test Section (Time = 0 years)
Control Section (Time= 1.25 years)
Table 2:
Calcium Leaching From Laconia, New Hampshire Bottom Ash Utilization Demonstration Road Cores
Pavement Type and Age
Bt,ca
Rca
1:
pDe,ca
mg/m 2
-
.
m2/s
Test Section (Time= 0 years)
2,9005,300
6.8-44.6
2,2008,800
13.8-14.3
Test Section (Time= 1.25 years)
1,2003,400
6.0-46.4
13,800
14.0-14.9
Control Section (Time= 0 years)
600-730
8.4-15.9
6,000
13.8-14.1
Control Section (Time = 1.25 years)
400-650
4.8-25.9
13,800
13.9-14.6
139
ADVECTIVE TRANSPORT
Infiltration
Infiltration
m
Percolation and Solubility Control DIFFUSIVE TRANSPORT
Diffusional Flux Granular Diffusion
Monolithic Diffusion
Figure 1" Schematic Depicting Advective Versus Diffusive Transport
BituminousPolymeric Matrix Very ~ ~ ~ Tortuous [-'"J~J--Z]~~~_.~~~h,~ L.. :~ ~ Diffusion LessTortuous 1.~ N , / ~ " ~ D i f f u s i o n Path Diffusion ~ [ i i ~ i ~ ~ ~ ( ' ~ , , ~ . J 6 a ,... I "..-,~Jf ~,~Y'E\~',J~,/1 ll--Jl~. Dissolution ~ ~ S l , g h t l y ,., q . AsphalCement t ~ " ~ Mncrocrac~, CoatedAshand~ ~ ~_~~ ~ LessTortuous A g g r e g a t e P a r t i c l e s ~ ~ ~ "" 'Wl o,,,us,on
NaturalAggregate I~ I -f v
,[
AshParticle ~ ~I~ ~ II1~ ~.
k,~. ....... AsphaltCementCoating~ r . ~ wasn-un ~ ~"~'~'~"~ [ ( ~ ~ . ~ ~ "
a) MonolithicAsphalt PavementBinder,Course
~
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This Page Intentionally Left Blank
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
149
DESIGN AND CONSTRUCTION OF A ROAD PAVEMENT USING FLY-ASH IN HOT ROLLED ASPHALT
by S.E. Zoorob and J.G. Cabrera C.E.M.U., Department of Civil Engineering, University of Leeds, Leeds LS2 9JT, U.K.
ABSTRACT
Using fly ash to substitute the filler in bituminous mixtures is not only a way of disposing of this waste in a safe manner but it is a way of reducing the energy requirements for the preparation and placement of this composite in road pavements. This paper describes briefly the background work carried out in the laboratories of CEMU to develop the mix design and to evaluate the reductions of energy requirements for the production of a bituminous mixture, i.e., hot rolled asphalt (HRA) which is used in the UK to surface motorways and heavily trafficked roads. Hot rolled asphalt is prepared in site plants heating the aggregates and bitumen at 160 ~ C and is placed and compacted normally at not less than 125~ Thus the energy requirements for the preparation, placement and compaction of HRA is very high. The HRA designed at CEMU is a low energy mix which contains fly ash. It can be prepared at a temperature of i25 ~ C and placed and compacted at a temperature as low as 85 ~ This mix was used for the construction of the overlay of a vew heavily trafficked road. the A689 in the north of England. The construction of the overlay for the pavement is described and the results of the in situ pavement are assessed with results obtained during four years of monitoring the performance of the section paved with the new low energy HRA. The paper also discusses the design and development of a semiautomatic rut measuring devise which was especially designed to monitor the performance of the pavement. The environmental and cost implications of using fly ash in bituminous mixes is quantified using the results obtained during the laboratory design and construction of the trial HRA overlay.
150 INTRODUCTION One of the most important requirements of bituminous mixtures is that their compaction characteristics in the laboratory can be repeated during construction in the field. Poor performance of bituminous mixtures in road pavements is in many cases attributed to poor mixing and inadequate compaction. Mixes that can be mixed, handled and placed without difficulties are said to be workable. Workability is a parameter which indicates these attributes in a bituminous mix. Most bituminous mixes can be made workable if enough high temperature of compaction is maintained during the process, this is obtained by heating mineral aggregates, filler and bitumen to relatively high temperatures, and transporting and laying the mixes in short periods to avoid loss of temperature. Many mixes become unworkable when they reach temperatures of + i20 ~ The main objectives of the study reported in this paper were: To assess the effect of fly ash (FA) on the engineering and performance properties of hot rolled asphalt. To assess the influence of changes on the temperature of mixing and compaction in conventional and FA hot rolled asphalt. To validate any findings using a wide range of mineral aggregates and fillers. To conduct a full scale trial using one of the design mixes. -
-
The project on the design of low energy hot rolled asphalt (LEHRA) using FA was supported by the Energy Efficiency Office of the Department of the Environment U.K., Cleveland County Council U.K., National Power U.K. and Tilcon North Limited U.K. MATERIALS USED IN THE INVESTIGATION AND THEIR PROPERTIES The materials used in the laboratory investigation were representative of those used in road pavement construction in the North of England. Four coarse aggregates, four sands, three limestone powders and four fly ashes were selected for the study. The code used to distinguish them are : Coarse Aggregate (CA) Limestone Filler (L)
Sand (S) Fly Ash Filler (FA)
The symbol is followed by a number from 1 to 4, this distinguishes the type and origin of each component. Details of the origin of these materials have been reported in reference [1 ]. Relevant properties are shown in Tables 1, 2 and 3. The particle size distribution of the limestone and fly ash fillers were similar mainly on the silt size range. There is a marked difference between the two types of filler. Limestone fillers are on average finer than FA fillers and have a higher specific gravity. The shape factor number [2] which is a measure of the sphericity of a particle shows that FA is predominantly spherical in shape while limestone is not. This characteristic of FA allows it to function as a filler in a solidliquid or solid-plastic composite without unduly increasing the viscosity of the composite.
151
Table 1 IVlaterial
S1 $2 $3 $4 CA1 CA2 CA3 CA4
Table 2 Filler FA1 FA2 FA3 FA4 L5 L6 L7 Spec. limits
Table 3
Sand and Aggregate Properties Relative Density on Oven dried basis 2.489 2.383 2.534 2.579 2.861 2.893 2.728 2.753
Relative Density on a saturated and Surface dried basis. 2.494 2.402 2.547 2.595 2.910 2.923 2.756 2.763
Filler Properties Voids of Dry Relative Compacted filler Density 0.211 2.179 2.412 0.275 2.249 0.281 2.384 0.235 2.773 0.255 2.824 0.243 2.725 0.248 -
Apparent Relative Density 2.502 2.430 2.568 2.621 3.009 2.982 2.807 2.780
Bulk Density in Toluene (g/ml) 0.450 0.638 0.344 0.422 0.612 0.625 0.638 0.5 - 0.9
Water Absorption (% of dry mass) 0.210 0.808 0.512 0.611 1.719 1.032 1.030 0.353
i
% retained on 75gm sieve 14.64 12.77 9.16 11.81 3.30 5.72 8.48 < 15.00
Filler Mean Diameters, Specific Gravities and Surface Areas
Filler Type
Drax PFA Blyth PFA Thorpe Marsh PFA West Burton PFA Ballidon Filler Marsden Limestone Scottish Limestone
Mean Diameter (micron) 14 10 11 9 7 7.4 5.4
Specific Gravity g/cc 2.179 2.412 2.249 2.384 2.773 2.824 2.725
Surface Area (m2/g)
Calculated Surface Area (m2/g)
Shape Factor
0.196 0.248 0.242 0.279 0.309 0.287 0.407
0.215 0.213 0.218 0.233 0.217 0.171 0.180
2 3 3 3 3 3
The bitumen used was a straight run nominal 50 pen grade bitumen. The bitumen properties measured are shown in Table 4.
152
Table 4
Properties of Bitumen
Penetration 25 ~ (• Softening Point (R&B) ~ Relative Density Penetration Index
52 53 1.029 -0.20
P R E P A R A T I O N OF HOT R O L L E D ASPHALT MIXES The proportions of coarse aggregate, sand and filler required to produce size distributions within the specifications given in BS 594" Part 1"1985, were 9 Coarse aggregate" Fine Aggregate" Filler 34% 9 56% 9 10% The eight aggregate combinations labelled M1 to M8 were used to produce HRA mixes at various mixing and compacting temperatures. An example of the resultant particle size distribution for mix 2 together with BS 594 specification limits are shown in Figure 1. Specimens were compacted in the laboratory, using the gyratory testing machine (GTM) [3] The main characteristic of the GTM compactor is that it allows the application of an axial static pressure at the same time that the specimen is subjected to a dynamic 'kneeding motion' which resembles the mode of energy applied in the field by construction plant. For each combination of mixing-compaction temperature, four samples were prepared using the GTM for compaction and measurement of the Workability Index. The compaction conditions in the GTM : Vertical pressure 0.7 MPa, Angle of Gyration 1~ No. of revolutions 30. These conditions give an energy of compaction of the same order as the energy of compaction applied by 50 blows per side with the Marshall hammer. The mixing and compaction temperatures used in the project are given in Table 5.
Table 5 Code TI T2 T3 T4
Mixing and Compaction Temperatures Mixing Temp. ~ 140 140 130 130
Compaction Temp. ~ 125 115 115 105
Code T5 T6 T7 T8
Mixing Temp. ~ 120 120 110' 110
Compaction Temp. ~ 105 95 95 85
153 RESULTS F R O M THE L A B O R A T O R Y STUDY
Workability In this study the method used to assess workability as developed by Cabrera [3, 4], consists of monitoring the specimen height and hence volume reduction during compaction. Knowing the specific gravity of the mix, for any specimen, the porosity can be plotted against the number of compactive revolutions. The experimental line should approximate a linear relation of the form: P, = A - B log to (i) A and B are constants. where A - intercept with the y axis. B = slope of the line. i = number of revolutions. The Workability expressed by the "Workability Index" (W.I.) is defined as the inverse of the constant A. i.e. the porosity at zero revolutions multiplied by 100. W.I.=
/ l) ~-
x 100
As expected W.I. values decrease as temperature decreases due to the increase in bitumen viscosity as the softening point is approached. But most importantly, all FA mixes at all mixing and compacting temperatures exhibit higher W.I. values than conventional HRA mixes, an example is shown in Figure 2. This can be attributed to the fact the FA particles have more rounded and less angular texture aiding workability. There is clear evidence that FA mixes will compact better than limestone mixes even at the lowest temperature of compaction used in the laboratory.
Stability. and Flow Values Densities. Marshall Stabilities and Flows were obtained according to BS 598 [5]. In general, Stability values decrease as mixing and compacting temperatures decrease. In all cases, the Stability of the conventional hot rolled asphalt mixes were only slightly higher than their counterpart FA mixes. Nevertheless, the stability of all mixes satisfy the criteria for roads carrying medium traffic (up to 6000 vehicles/lane/day). See Tables 6 and 7 for the stability and flow design parameters.
Criteria for the Stability of laboratory designed asphalt. Table 6 BS 594" Part 1 91985 Marshall Stability of complete Traffic (colmnercial vehicles mix (kN). per lane per day) 2to8 Less than 1500 1500 to 6000 4 to 8 Over 6000 6 to 10 Table 7
Asphalt Institute Design Criteria Light Traffic Medium Traffic Compaction 2 x 35 2 x 50 Stability (kN) 3.33 5.33 Flow,(0.25mm) 8 - 18 8 - 16 Porosity(%) 3 - 5 3 - 5
Heavy Traffic 2x75 8.00 8 - 14 3-5
154 FA mixes exhibit consistently lower flow values than their counterpart conventional mixes at all mixing and compacting temperatures, nonetheless almost all the flow values measured were less than 4 mm.
Porosity and Voids in Mineral Aggregate (VMA) There is no marked change in VMA values as the temperature of mixing and compaction decreases. Also for each aggregate type, both conventional and FA mixes do not exhibit a great change in VMA values. Porosity values for all the mixes at all temperatures were below 6%.
Creep Stiffness The creep test is carried out on duplicate specimens at 40~ The test lasts two hours, and gives results which allow the characterisation of the mixes in terms of their long term deformation behaviour [6]. Analysis of creep test results carried out on the hot rolled asphalt containing FA filler show the normal variation in stiffness of mix values with respect to stiffness of bitumen. The stiffness of the bitumen was obtained from a Van der Poel nomograph. The nomograph gives values of stiffness as a function of the time of loading, the temperature difference between test conditions, the Softening Point temperature, and the Penetration Index. From the Smix - Sbit experimental values, regression lines were obtained. These regression equations are of the form" Log Smix = X Log Sbit + C and were used to obtain the Smi x values at one hour loading time. Figure 3 shows an example for mixes 1 and 2.
Determination of the Optimum Bitumen Content (o.b.c.) The Leeds Design Method [4], recommends that the optimum binder content should be obtained by averaging the binder contents corresponding to the following parameters" Maximum Stability, Maximum Density, Minimum voids in the mineral aggregate, Maximum compacted aggregate density, Maximum Stiffness. The optimum value obtained should lie within 3 - 5% porosity and below 4 mm Flow. The o.b.c's for the mixes prepared at different temperatures of mixing and compaction were then averaged and the results used as the o.b.c, for each mix combination independently of the temperature of mixing and compaction. On average the o.b.c's of all mixes were very close to 7% for all temperatures [ 1].
155
Fatigue Testing Programme In this part of the programme, an Instron 8033 Servo Hydraulic Dynamic Testing machine was used to produce a sinusoidally shaped loading pulse on the beam specimens. The machine was also equipped with an oscilloscope to aid in monitoring the shape of the applied load and the response pattern. The output data. consisting of the magnitude of the applied load, the approximate piston head position and the number of cycles, were updated and displayed on the visual display unit of the Instron. To measure the resultant strains on a beam, a PL-60 wire resistance strain gauge having 60 mm length was glued at two locations of one beam side. The central strain gauge was fixed at a location approximately 25 mm above the bottom of the beam. The strains produced by the repeated loading at the center of the beams were amplified by a Differential Strain Amplifier and then transmitted to a data logger which was connected to a computer. The fatigue test was performed by placing the rectangular beams on a 50 mm thick rubber foundation which had a modulus of elasticity value of 4.3 MPa. The dynamic load was applied to the centre of the beam via a rubber loading block 60 mm wide, see Figure 4. The whole assembly consists of the beam and rubber pad which rests on a very stiff 25 mm thick steel plate which is supported directly on the base of the loading frame.
Method of Analysis During fatigue testing, sudden failure due to detect propagation in the detected as the mode of failure in the test configuration, hence brittle toughness calculations are not applicable. Thermal fatigue or ductile deformation in the region of rupture due to the build up of heat is the mechanism.
brittle mode was not fracture and fracture t'ailure where plastic more realistic failure
By employing a dissipated energy approach the results of different types of dynamic tests, carried out under different sets of conditions and with several types of asphalt mixes, can be described by a single, mix-specific relation :the number of cycles to fatigue is related mainly to the amount of energy dissipated during the test. The occurrence of rest periods, the use of controlled-stress or controlled-strain tests, the effects of frequency and temperature do not significantly influence the dissipated energy relation. During a controlled strain test the stress amplitude and the phase angle change. This means that for the calculation of the total dissipated energy it is necessary to integrate the functions of stress and phase angle over the number of loading cycles concerned. This integration is approximated by a summation of the energy into fixed intervals of "constant" cycles, i.e. cycles in which it can be assumed that the stress and phase angle in that interval are nearly constant. From the energy relation it follows that at a higher fatigue life of the mix more total energy per unit volume can be dissipated [7].
156 By using the original strain versus the number of cycles, the cumulative energy dissipated for each beam was calculated up to the point where the test was terminated. Figure 5 shows the relation between the number of cycles to the cumulative dissipated energy value of 60x 106 J/m 3 for the initial strain at the four main load levels (95 N, 150 N, 260 N and 1400 N). It is clear that FA and ordinary mixes behave in the same manner and there is no distinction between their fatigue properties.
FULL SCALE ROAD T R I A L Following the successful outcome of the laboratory investigation, a full scale road construction trial to assess the performance of Low Energy Hot Rolled Asphalt (LEHRA) under real intense traffic loading was carried out. The selection of the mix for the construction of the road trial was indirectly determined by the geographical position in the U.K. of the road selected by Cleveland County Council. The road selected served high traffic volumes which provided realistic conditions for the evaluation of the performance of the LEHRA. It was also a requirement that the road structure should not have reached a service life requiring reconstruction, but one where a strengthening surface layer should be the most acceptable engineering improvement. A section of the A689, Eastbound slow lane, one mile West of Trunk Road A19, was made available for the construction of the trial section. The most convenient materials were those designated Mix 1 and Mix 2 of the laboratory study. Materials and Location of the Road Trial
The trial pavement was constructed on April 1991. The plant for the preparation of the bituminous mixes was made available by Tilcon North Ltd., and consisted of a Miller Batch Asphalt Plant of 240 ton/h capacity with delivery of 3 ton mix per batch, located in Blaydon, North Yorkshire. Two Hot Rolled Asphalt mixes were produced 9A control mix containing the conventional limestone filler and an experimental mix containing FA filler designed according to the Leeds Design Method. Both mixes conform to BS594, Part 1" 1985. Designation 30/14. Analysis of the mix composition produced in the batching plant showed that the mix was identical to the mix designed in the laboratory and described earlier. Production Sequence
The trial consisted of the production and laying of 70 tonnes of LEHRA and 80 tonnes of conventional HRA in the slow lane, Eastbound carriageway of the A689 Wolviston to Billingham road [8]. The length of the trial road was 320 m and the thickness of the layer was 40mm. The total work for the morning was scheduled to be 150 tonnes. Evaluation of the production process highlighted the following"
157 1. The PFA would not flow properly through the silo screw-feed system, flow was maintained by manually rodding the base of the silo. The use of aeration and vibration which is common in the concrete industry will solve this problem. 2. Attempts to match the slow rate of filler feed with an equivalent sand feed rate and the reduction of temperature, caused the drier burner control to become unstable below about 135oc. The final production temperature of the mixes were as indicated in Table 8.
Table 8 a) b) c) d)
Production Temperatures
Material Control HRA LEHRA Mix LEHRA Mix Control HRA
Target Temp. (oc) 160 130 140 160
Actual Temp. (~ 140 130 137.5 160
Quantity (Tonnes) 30 30 40 50
Mix temperatures at the plant were taken using electronic thermocouple type thermometers in accordance with the recommendations set out in BS 598: Part 109: 1990. Because of the long loading time and breezy conditions, there was a drop of around 8o(2 in the temperature of the material which was first loaded. A hand held thermal anemometer was used for the measurement of ambient temperature and wind velocity at the site. The results indicated a range of speeds varying between 5 and 7 m/s. Wind gusts of up to 10 m/s were also recorded. The temperature was 20oc + 1~
Ener~, Use During The Trial Analysis of gas meter readings on the drum dryer during the plant mixing stage indicated the following: Control Mix 257 Therms for 80 Tonnes - 3.21 Therms / Tonne. (338.6 MJ / Tonne.) LEHRA Mix 200 Therms for 70 Tonnes - 2.86 Therms / Tonne. (301.7 MJ / Tonne.) Data for the asphalt mixing plant to December 1990 show an energy cost of production (at a target efficiency of 85%) of s 1.01 p for asphalts, and 58p for macadams per tonne. Of the 43p difference, 70% is a consequence of the additional moisture content of asphalt sands when compared with crushed rock used in macadams. The other 30% is 'heating' energy; a consequence of the higher mixing temperatures needed for 50 pen. grade bitumens usually used with the asphalts. The substitution of limestone by PFA allowed a reduction in the temperature of the raw materials in the asphalt plant at the heating/mixing stage of production. This reduction in heating energy is equivalent to 30%. Therefore using the difference of heating requirements for HRA and macadams, it can be said that the savings at the heating / mixing stage are 30% of 43 p or 12.9 p / tonne of LEHRA produced. Table 9 shows the details of the trial road construction, including materials and temperatures of mixing and compaction.
Tnl)le 9
Details o f the T r i a l Ro21cl Construction.
A 689 Wolviston to Billingham HRA Trials, laid 11th April 91. Distance (m) 0 - 80 81 100 101 - 145 146 - 200 201 - 240 240 - 320
-
Materials Control HRA (Limestone Filler). HRA with FA Section I. HRA with FA Section 2. HRA with FA Section 3. IIRA with FA Section 4. Control HRA (Limestone Filler).
R.O.S. = Rate o f spread o f chipping (kg/m2). S.M.T.D. = Texture depth measuremerits by Laser Texture Depth Meter (mm). Note; the two rows o f data indicate the near side and o f f side measurements respectively.
159 A S S E S S M E N T OF T H E P E R F O R M A N C E OF T H E T R I A L P A V E M E N T The following control tests were carried out in the field 9 1- Traffic Control. 3- Surface texture measurements.
2- Rut Depth measurements. 4- Dynamic deflection measurement.
Cores were obtained from the different sections of the pavement and measurements of porosity, density and stability were carried out.
laboratory
Traffic Count for A689 Eastbound Cleveland County Council and Leeds University carried out measurements of vehicle flows, these were : Total number of Vehicles in both directions - 8877 Total HGV's in both directions - 2832 % of HGV' s - 31.9 % Total number of Vehicles Eastbound - 4371 % o f H G V ' s - 32 % Total number of Vehicles Westbound - 4506 % of HGV's - 31.7 %
Rut Depth Measurements Initially rut depth measurements for the entire trial length were taken on 5 dates initially using a Straight Edge (3 m in length). The average rut depth values showed that, is was not possible to draw any trends, since there is no evidence of road deterioration. Realizing the need for a more accurate means of monitoring rut depth profiles, which does not entail a worker having to lie on a wet road surface on a cold day taking readings using a small meter rule to an unrealistic accuracy of 1 mm, a rut depth measuring beam was developed at the laboratories of CEMU, University of Leeds. The Leeds Rut Depth Measuring Beam (LRDMB) automates the process of rut depth measurement. A measuring arm, follows the pavement rut depth profile, this is connected to a carriage that runs along the beam length. A hand held microcomputer then digitizes and stores the electrical signals sent off from the measuring arm at every 100 mm traveling distance along the beam. The signals are generated via an angular variable differential transformer that measures the change in angle that the arm makes as it follows the irregularities of the pavement surface. The digital values are then transferred to a template on a spreadsheet and this converts the readings to actual deflections in mm and plots the resulting profile. A longitudinal-section of the beam constructed is shown schematically in Figure 6. Figure 7 shows the longitudinal rut depth profiles along the entire trial obtained using the LRDMB. The experiment confirmed that the road sections made with LEHRA showed rut depth values which were comparable with those of sections containing the conventional HRA.
160
Surface Texture Measurements Texture depth measurements by Laser Texture depth meter, measured in accordance with the specifications for Roads and Bridges clause 929 [9],were also taken. An example of the results is shown in Table 9, and the second, a more recent set is presented in Table 10. D y n a m i c Deflection M e a s u r e m e n t s Deflection measurements taken on the trial road using a Deflectograph [10,11,12] are shown in Figure 8. The trends shown in this Figure lead to the clear conclusion tht L E H R A and H R A deflect to the same extent confirming the L E H R A is as good a material as the conventional HRA. T a b l e 10 Distance (m) 0-10 10-20 20- 30 30 - 40 40- 50 50 - 60 70 - 80 80- 90 90 - I00 100-110 110- 120 120- 130 130- 140 140 - 150 150- 160 160- 170
T e x t u r e Depth M e a s u r e m e n t s Near Side Off Side Distance (m) Wheel Track Wheel Track 0.87 0.94 170- 180 1.15 0.96 180- 190 1.14 0.75 190 - 200 1.12 0.83 200- 210 1.11 1.25 210- 220 1.27 1.2 220- 230 1.08 1.17 230- 240 1.27 1.24 240- 250 1.26 1.15 250- 260 1.13 1.23 260- 270 0.95 1.28 270- 280 1.07 1.23 280- 290 1.27 0.96 290- 300 1.09 0.93 30O- 310 1.11 0.98 310 - 320 1.16 1.06 320- 330 330- 340
Near Side Wheel Track 1.15 1.31 1.22 1.18 1.33 1.25 1.16 1.07 1.08 1.22 1.06 1.11 1.22 1.21 1.11 1.0 0.84
Off Side Wheel Track 1.19 1.04 1.15 1.02 1.02 1.07 1.07 1.15 1.12 1.02 0.87 0.82 0.9 1.01 1.2 1.02 0.95
Laboratory. Tests Cores taken at various times during the monitoring period of 4 years showed that the FA and conventional HRA performed satisfactorily. The details of measurements carried out on these cores have been reported in references [8 and 13].
161 CONCLUSIONS From the results obtained in the Laboratory the following conclusions are offerred 9
9 Fly ash is known to possess predominantly spherical particles when observed under an electron microscope, this enables them to improve the packing properties of HRA. FA particles tend to occupy more bulk volume per unit weight as compared with the more familiar irregular surfaced limestone powder particles. This can be observed from the lower bulk density in the Toluene test. 9 FA filler hot rolled asphalt has far higher workability index than conventional hot rolled asphalt for any of the aggregate combinations used. This finding implies that hot rolled asphalt containing FA can be mixed and compacted at temperatures as low as 110 ~ 85 ~ respectively without impairing its engineering and performance properties. 9 The savings in energy input are considerable and thus FA-HRA can be classed as a low energy material. 9 Replacement of limestone filler with FA does not affect the optimum bitumen content of hot rolled asphalt. 9 The Stability and Flow of the FA mixes satisfy the criteria for medium traffic (up to 6000 CVd), laid down by the Ministry of Transport and the Asphalt Institute of U.S.A. for the range of temperatures tested. 9 Using the Energy approach to eliminate the effect of temperature, it is shown conclusively that mixes containing F A replacement behaved in a similar manner to conventional HRA mixes under repetitive flexural loads. Under a particular stress level, both types of mixes needed the same number of cycles to dissipate a chosen quantity of energy. F r o m the results of full scale trials, the following conclusions are offerred Laying Operations It was clear that using FA as a filler has allowed the temperature of mixing and compaction to be lowered without affecting the optimum binder content. This not only has a significant benefit in terms of heating energy, but also encourages laying under adverse winter conditions. The net energy saving amounts to 12.9 p / tonne of the total energy expenditure in the preparation of the mix. Other savings regarding laying at low temperature and avoidance of waste due to cooling have not been quantified. Laboratory. Results Hot Rolled Asphalt with pfa fillers conformed to BS 594:1985 specifications and were able to perform as well as the materials containing the conventional Limestone fillers. Results from cores tested in the Laboratory were within the specifications for roads carrying up to 6000 CV's per day. Road Monitoring Monitoring of the road has been continuosly carried out : rut depths, surface texture and deflectogragh measurements have been analysed. There was no sign of any form of distress, and the trial sections have shown excellent performance.
162 Hot Rolled Asphalt with PFA as a filler has per~brrned as well as the conventional HRA containing Limestone filler. Cleveland County Council will continue to monitor the performance of the trial section so that information regarding its long term performance can be confirmed.
REFERENCES
l0 ll
12
13
Cabrera J.G. and Zoorob S.E., 'Design of Low Energy Hot Rolled Asphalt', Proc. Performance and Durability of Bituminous Materials, Ed. J G Cabrera and J R Dixon, pp. 289-308, Spon, 1996. Cabrera J.G and Hopkins C.J. 'The influence of PFA shape on the properties of concrete. Ash Tech 84, Second International Conference on Ash Technology, pp. .~v_~-398, London. Cabrera J.G. 'A new methd for the assessment of the workability of bituminous mixtures'. Journal of the Institution of Highways and Transportation No. 11, pp 17 to 23, 1991. Cabrera J.G. 'Hot bituminous mixtures: Design for performance'. 1st National Conference on Bituminous Mixtures and Flexible Pavements. University of Thessaloniki, Greece, pp 1-12, 1992. BS 598: Part 107: 1990, Sampling and examination of bituminous mixtures for roads and other paved areas. Cabrera J.G. and Nikolaides A.F., 'Creep performance of cold dense bituminous mixtures'. Highways and Transportation. Vol. 35, no. 10, pp. 7-15, 1988. Zoorob S.E. and Cabrera J.G., 'Laboratory investigation on the fatigue properties of low energy hot rolled asphalt', 2nd National Conference on Bituminous Mixtures and Flexible Pavements. University of Thessaloniki, Greece 1996, pp. 1 to 13. Rockliff D. 'The use of pulverised fuel ash as a filler in hot rolled asphalt mixturespractical aspects'. Proc. Performance and Durability of Bituminous Materials. Ed. J G Cabrera and J R Dixon, pp. 309-315, Spon, 1996. Department of Transport. Specification for Highway Works, Clause 929, HMSO, London. Kennedy C.K.and Lister N.W., 'Prediction of Pavement Performance and the Design of Overlays', TRRL LR 833. Department of Transport, Roads and Local Transport Directorate, Departmental Advice note: HA/25/83. Deflection Measurement of Flexible Pavements. Analysis, Interpretation and Application of Deflection Measurement. Lister N.W., Kennedy C.K. and Ferne B.W., 'The TRRL Method for Planning and Design of Structural Maintenance', Proceedings Fifth Intemational Conference on the Structural Design of Asphalt Pavements, University of Michigan Vol. 1, Delft 1982, pp 709 to 724. Cabrera J.G. and Zoorob S.E., 'Field performance of low energy hot rolled asphalt'. 2nd National Conference on Bituminous Mixtures and Flexible Pavements. University of Thessaloniki, Greece 1996, pp. 129 to 150.
163 Figure 1 Particle Size Distribution for Mix 2. Coarse & Fine aggregates - Birtley, Filler = Birtley Limestone. % Passing. 100
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Figure 3 Stiffness of mix (MPa) v.s. bitumen c o n t e n t
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164
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Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
167
Engineering Properties of the Coal Ashes stored in the "Valdeserrana" Lagoon. Andorra Power Plant (Spain). CALDERON GARCIA, PEDRO A. Associate Professor. Dept. of Engineering Construction, Universidad Polit6cnica de Valencia. PERIS MORA, EDUARDO Professor. Department of Engineering Construction, Universidad Polit6cnica de Valencia. PARRILLA JUSTE, JESUS Research Assistant. Dept. of Engineering Construction, Universidad Polit6cnica de Valencia. ABSTRACT: Andorra Coal-fired Power Plant produces annually over 1,000.000 tons of coal ashes (15 % bottom ash - 85 % fly ash) that are stored in the Valdeserrana Dam (a 18 Hm 3 lagoon) .The chemical composition of the ashes placed in the lagoon is similar to that of dry ashes. However, the mineralogical analysis show a decrease of the vitreous components, which would account for a decrease of its pozzolanic properties. The ashes show no cementation while in the lagoon and behave very much like a natural soil. The gradation of the particles varies with the distance from the pouring point, being the coarser particles (bottom ash) closer to it. The coarser materials can be classified as a loose silty sand. The finer materials (fly ash) can be classified as weak silt with traces of sand. Both materials have high moisture content and high void ratios and are normally (or even under) consolidated. Still, they have low compressibility, good friction angles and high consolidation coefficients. Therefore" these ashes can be efficiently improved by different methods like precharge or mixing with lime or cement if they are to be used as foundation soil. 1. INTRODUCTION
total area (about 1/3 of its volume) is filled with ashes (see figure 1). A research program is being carried out to investigate the potential uses of the ashes long time after its disposal in the lagoon. Its first part has assessed the chemical and geomechanical properties of the ashes, determining their aptitude as a foundation soil. The site investigation included field and laboratory testing. The paper presents their main results.
Andorra Coal-fired Power Plant (Teruel, Spain) produces annually over 1,000.000 tons of residual coal ashes. These ashes are composed approximately by 15 % of bottom ash and 85% of fly ash. They are mixed together with water and conducted through a pipe to the Valdeserrana Dam where the sluice is poured and the ashes settle. The lagoon has a total capacity of 18 Hm 3 and about half of its 1
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168 2. EXPERIMENTAL PROGRAM
Figure 2 shows the main results of the field investigation. The ashes are placed directly over the natural soil and the deposit reaches a maximum thickness of 21 meters. The investigation showed that the size of the ash particles of the deposit vary with the distance from the pouring point. Close to it (Bore Hole 1), where the water flows faster, the deposit contains mainly the coarser particles (mostly bottom ash). As the distance from the end of the pipe increases, the proportion of finer particles in the deposit increases, becoming a mixture of bottom and fly ash. At a far distance from the pouring point the deposit contains mostly fly ash (finer particles). The underlying natural soil is formed by a Terciary sedimentary formation, composed by layers of very hard clay and cemented gravel. They can be considered as a rigid and impervious stratum.
2.1. Bore Holes and Dynamic Penetration Tests A total of 3 bore holes, reaching to a maximum depth of 21 meters, and 8 Dynamic Penetration Tests (DPT, Borros type) were carried out to investigate the general properties of the deposit. The bore holes were placed along a small track that goes inside the lagoon, at different distances from the pouring point. The Dynamic Penetration Tests (DPT) were done with a very ligth machine that was able to circulate over the soft surface of ashes. A D PT was done beside each bore hole in order to correlate their results with the soil observed at the bore hole. The rest of the penetration tests (up to a total of 10) were placed at different points of the lagoon (gee figure 1).
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Figure 2.- Section A-A'. Soil profile
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169
2.2. Chemical analyses The chemical composition of ashes located at different depths and places (and therefore, of different ages) were determined. Table I shows the results of these analyses compared to the average composition of fly and bottom ashes before its disposal. As can be seen in Table I both fly and bottom ashes have very similar chemical
composition and they do not change much due to the disposal in the lagoon. Only a small descent of the calcium oxide (CaO) content can be observed. Also, it can be seen that Andorra's ashes are of the silico-aluminous type (opposed to the sulpho-calcic type), with small CaO content, and almost no free calcium oxide, so they are not self-cementitious. Actually the ashes did not exhibit any cementation at the lagoon.
TABLE I .- Chemical composition of fly & bottom ashes before and after its disposal in the lagoon
Percentby
weight
::iiiii~~i:iiii!il Bottom ash .iiii!!i!!fl~~iiiiiiii!!il (freshly
.iiii.i..[a~ag~ :!i!!i! disposed) ,,,
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41,0
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Still, these ashes are pozzolanic [1]. The pozzolanic properties of a coal ash depend mainly on the vitreous components in its mineralogical structure. To investigate the pozzolanic potential of the ashes after its disposal, a mineralogical analysis was carried out on 3 samples of different ages from the lagoon and 1 sample of non-disposed fly ash. The results are shown in figure 3. They show that the ashes lose some vitreous components after its disposal, losing some of its pozzolanic properties. Still, simple tests show that the ashes still harden in the presence of calcium, keeping some-but not all- its pozzolanity.
iiii!ii:i:::i:!::?!i:::::!!i:i:'::::::'i:'?:ii::i'ii'i::::::'i:?':...i:;..
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o" (1) MOSTLY BOTTOM ASH (Recent disposal) (2) FLY ASH MIXED WITH BOTTOM ASH (Old disposal) (3) MOSTLY BOTTOM ASH (Old disposal) (4) FLY ASH (Before disposal)
Figure 3.- Minerological analysis of Andorra's coal ashes
170
2.3. Index properties The index properties of the different types of ashes were determined. The particle size distributions were obtained combining sieving and sedimentation procedures. Figure 4 shows the gradation curves. For the sake of comparison, the figure includes the typical ranges of fly and bottom ash [2], [3]. All the samples were "non plastic". It can be seen in figure 4 that there are three different types of gradation curves. Each of them correspond to a type of ash deposit. The deposits identified as "fly ashes" fit very well inside the typical range of fly ashes. They can be classified according to the Unified Soil Classificatin System (USCS) as silt with sand. Looking at the finer sizes in the curve, it can be seen that the percentage of clay (% finer 0,002 mm) varies from 11 to 18 %.
The ashes identified as "mostly bottom ash" have a high content of fine particles (% finer than 0,08 mm betweeen 28 and 36), much higher than what is usual for bottom ashes. This is because they are not "pure" bottom ashes, they are "contaminated" with fly ash. According to the USCS they can be classified as a silty sand with some gravel. The third type of gradation curves comes from a variable mixture of fly and bottom ashes. These curves are intermediate between the other two. Their USCS classification varies between "sandy silt" and "silty sand with traces of gravel". The specific gravities of the ash particles (G,) are very high, between 2.49 and 2.85, higher than what is usual in these materials. The high iron-oxide content and the neglectable carbon content in their composition may be the reason of these high G,.
GRADATION CURVES 100 80 9~
60
4o
20
0
0,001
X
FLY ASH
0,01
0,1
1
Particle Size (mm) BOTTOM ASH
Figure 4.- Particle size distribution of Valdeserrana ashes.
---t---
10
100
MIX FLY ASH - BOTTOM ASH
171
2.4. Density and porosity
These values show that the state of the ashes in the lagoon is very loose.
The ashes have been placed in the lagoon by hydraulic filling and they have not been consolidated or compacted. Therefore, they present high water contents and low densities. Table II summarizes the properties of the 3 types of ashes identified. It can be seen that the ashes in the lagoon have high void ratios (from 0.64 to 1.07) and dry densities (Yd) from 12.8 to 18.8 KN/m 3. Even though these dry densities values seem high for a fly or a bottom ash (see [2], [3], [4h they represent low relative densities (about 35 %) for Andorra ashes due to their high specific gravity.
Table
2.5. Shear strength The ashes stored at Valdeserrana Dam behave mostly as a frictional material. None of the ashes displayed any cementation. Only the fly ashes (finer particles), had a small undrained cohesion, similar to that of a non-plastic silt. The friction angles were obtained by direct shear tests on undisturbed samples. The values obtained are high, typical in this materials (see
[31, [41, [51).
II.- Geotechnical properties of Valdeserrana coal ashes Fly-bottom
Mostly
ash mixture
bottom ash
32
19-38
26-29
e0 (void ratio)
0.85
0.64-1.07
0.72
~'d (KN/m3) (dry density)
14.2
12.8-18.4
15.8-17.7
Nsrr
0-3
7-29
12-13
N20(Borros, average)
2.8
12
17
32
33-35
39-50
Cc (compression index)
0.03
0.06-0.06
0.04
C, (recompression index)
0.009
0.007-0.009
0.006
C. (secondary compression index)
0.0017
0.0015-0.0038
0.0017-0.0025
0.06
0.025-0.07
0.028-0.04
1.0"10 "6 - 5.0"10 "~
5.0"10"~
6.0.10 "6. 1.2.10-5
0.03-0.15
0.3
0.3-0.4
Geotechnical Property
Mostly fly ash
w, (%) (natural water content)
c. (KPa)
15-25
tp' (o) (effective angle of shearing resistance)
C/C. k (cm/s) (permeability)
Cv(cm2/s) (coefficient of consolidation)
The friction angle of the fly ash is 32 ~ , higher than what is usual for a natural soil of the same USCS classification. However, the undrained cohesion and the penetration tests performed in these ashes gave low resistances, what shows that the fly ash is normally or even under consolidated in the lagoon. Bottom ash friction angles were higher, between 39 and 50 ~. Even though their
relative density were low they showed values of tp' higher than what is usual for a natural soil of the same USCS classification. These values of the friction angle did not fit well with those obtained by empirical relations with dynamic penetration tests. The vesicular shape of the particles, different to the typical shapes of natural soils, may be the reason for this bad fitting.
172 The friction angles of the mixture of bottom and fly ash (33 to 35 o) are between the angles of both materials. They are somewhat closer to the fly ash values, probably due to the high percentage of fine particles (50 to 61%), that become the ones governing the behavior. In all cases, the ashes have higher friction angles than what be expected for natural soils of the same USCS classification and the same relative density.
2.6. Compressibility
(fly ashes) to 10-5 cm/s (mostly bottom ashes), typical values for natural soils of the same USCS classification. The coefficients of consolidation could not be obtained by the oedometer tests. The samples completed the primary consolidation very fast (less than 15 seconds) so it could not be observed. Therefore, the coefficients of consolidation were estimated using the permeability values and the volume coefficient of compressibility (m0, related by : Cv
--
k
Y,mv The compressibility of the ashes has been estimated by one-dimensional consolidation tests. The values of the compression index (Co) varied between 0,03 and 0,06, being very similar for all the ashes. These indexes of compression show that the ashes have a compressibility similar to that of a medium dense sand. This would mean a very low compressibility (high stiffness) for a fine soil like fly ashes, and would be in the lower part of the usual range for fly ash [2]. As for the bottom ash, it would mean that their compressibility is similar to that of a natural soil with the same USCS classification and the same relative density. The recompression indexes (Cr) were between 1/3 and 1/6 of the virgin compression indexes (C~), what shows that consolidation or compaction improves dramatically the stiffness of the deposits. The ashes displayed very low secondary compression, almost neglectable for practical purposes.
2.7. Permeability, consolidation The permeability of the deposits has been estimated by variable-head Lefranc tests performed inside the bore holes. The values of permeability fitted very well the values obtained by Hazen's empirical relation [6] : k = 100 (DIo) 2
where k is the permeability in cm/s and Dlo is the effective size of the soil in cm. This good fitting is probably due to the spherical shape of the finer particles (fly ash). The permeability values ranged from 106 cm/s
The values obtained were very high, between 0.03 cm/s 2 (fly ash) and 0.4 crh/s (bottom ash). Therefore, consolidation of these deposits will take place very rapidly. 3. CONCLUSIONS Andorra Coal-fired Power Plant produces annually over 1,000.000 tons of coal ashes. These ashes are composed approximately by 15 % of bottom ash and 85 % of fly ash. They are mixed together with water and conducted through a pipe to the Valdeserrana Dam (a 18 Hm 3 lagoon) where the ashes settle. The engineering properties of the ashes placed have been studied and are presented in the paper. The chemical composition of the ashes placed in the lagoon is similar to that of dry ashes. However, the mineralogical analysis shows a decrease of the vitreous components, which would account for a decrease of its pozzolanic properties. Still, they keep some pozzolanity after its disposal. The ashes are not self-cementitious. They show no cementation while in the lagoon and behave very much like a natural soil. The specific gravities of the soil solids are very high for a coal ash (average 2.8). The gradation of the particles varies with the distance from the pouring point. The coarser particles (bottom ash) settle close to it while the finer particles (fly ash) settle at a higher distance. The coarser materials can be classified as a loose silty sand. The finer materials can be classified as weak silt with traces of sand. Both materials have high moisture content and high void ratios and are normally (or even under) consolidated. Still,
173 they have low compressibility, good friction angles and high consolidation coefficients. Therefore these ashes can be efficiently improved by different methods like precharge or mixing with lime or cement if they are to be used as foundation soil. 4. ACKNOWLEDGEMENTS The authors want to express their gratitude to ENDESA for the funding of the study. REFERENCES 1. Pay~i, J.; Borrachero, M.V.; Peris, E.; Aliaga, A.; Monz6, J. (1994) "Improvement of Portland Cement~Fly Ash Mortar Strength ussing Classified Fly Ashes". Environmental Aspects of Construction with Waste Materials. Elsevier Pub. ISBN 0-444-81853-7 Amsterdam 1994. 2. Pardo de Santayana, F. (1992) "Comportamiento geot~cnico de cenizas volantes en reUenos compactados y su evoluci6n a lo largo del tiempo" Centro de Estudios y Experimentaci6n de Obras Pt]blicas. MOPTMA. Pags. 19-27. 3. Huang, W.H. and Lovell, W: (1990) "Bottom Ash as Embankment Material" Geotechnics of Waste Fills-Theory and Practice, ASTM STP 1070 4. Dawson, A.R., Bullen, F. (1991) "Furnace Bottom Ash : its Engineering Properties and its Use as a Sub-base Material" Proc. Inst. Civil Engrs. Part 1, Vol 90, Oct. 1991, pp. 993-1009 5. Toth, P.S.; Chan, H.T. and Cragg, C.B. (1988) "Coal ash as structuralfill, with special reference to Ontario experience" Canadian Geotechnical Journal, Vol. 25, no. 4, Nov. 1988, pp. 694-704 6. Hazen, A. (1911) Discussion of "Dams on Sand Foundation" by A.C. Koenig, Trans. ASCE, Vol. 73, p. 199.
This Page Intentionally Left Blank
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
175
Mine tailings - practical experiences in f i l l i n g up harbours J. van Leeuwen and K. Ratsma Gemeentewerken Rotterdam (Public Works), Engineering Division, P.O. Box 6633 3002 AP Rotterdam, the Netherlands.
Abstract
In the port of Rotterdam, old harbours are being filled up to create new sites for industrial activities. Large quantities of materials are required for these projects. Because of the mechanical, environmental and economical advantages, mine tailings are therefore used. In the past some 1.000.000 tons of this natural, inorganic residue of coal mining was used in several filling up projects in the port of Rotterdam. The muncipality of Rotterdam has carried out an environmental impact assesment on the use of mine tailings. Conclusion was that the oxydation of the mineral pyrite is the critical aspect of the mine tailings. To prevent oxydation, it's only used below the groundwater table, in anoxic zones. In the laboratory various experiments have been carried out on the contents and leaching of components, according to the legislation and the licences. Besides, the oxydation of pyrite is controlled in the field. At one harbour site (Beatrixhaven) the oxydation of pyrite is being monitored by measuring the pH and redoxpotential with an in-situ cone. As the mine tailings are produced in coal mines in Germany, the environmental survey starts with sampling at the mine. In the meantime, the mine tailings are shipped to Rotterdam. At the site, the delivered material is then again sampled and examined. The mine tailings are put in the harbour and the results of the survey have to meet the standards. The conclusion of the paper, after several years of experience with the application of mine tailings, is that the use of mine tailings is environmentally safe and nleet the standards. The boundary is to use this material only in an anoxic environment (below the groundwater table). The major bottle neck in the benificial use of mine tailings is the long time, required for a licence (6 months) for a waste material, compared to those of raw materials (few days). 1.
INTRODUCTION
The policy of the Municipality of Rotterdam is to stimulate the beneficial use of secondary materials. For this reason the Port of Rotterdam has adopted mine tailings in large scale projects for filling up harbours (ref. 1). Mine tailings are the anorganic residues of coal mining. They are not mined as such (they are not a primary raw material). Due to their favourable geological properties mine tailings have been used in hydraulic engineering projects in Rotterdam since the nineteensixties (ref. 2). As a result of their size-distribution and their high friction angle they have a good filtering action, are resistant to erosion, and can be used at a steep slope (under water). An application less specific to mine tailings is the filling of banks and harbours - for which they are also very suitable. The annual amount of mine tailings used in the harbour area of Rotterdam depends greatly on the ongoing major projects, in particular filling up projects.
176
r'~E'~4,t~NEMAA$
~
~
I
PLANNED FILLING-UP PROJECTS
Figure 1. The Waalhaven / Eemhaven port area. Each project requires an amount of between several thousand tons and several hundred thousand tons. In the nineteen-nineties the average annual use was about 400,000 tonnes. Both unclassified mine tailings (0/70 mm) and classified mine tailings (10/125 mm) are used in Rotterdam. In many cases classified mine tailings are used in protective structures since these have a greater resistance to erosion in comparison with. unclassified mine tailings. The mine tailings are a by-product obtained during the mining of coal, and are therefore designated as a secondary construction material. Consequently from the early nineteen-nineties onwards a licence is required for their use, in accordance with the Pollution of Surface Waters Act of the Netherlands (Wvo). Although mine tailings are a naturally-occurring material (comprised of clay and sandstone), contamination of (stored) mine tailings could not be precluded in the past due to the manner in which the material was mined and stored. Until the nineteen-seventies hydraulic oils containing PCBs were used during the mining of coal, and as a consequence there was a risk of an incidental marginal contamination of mine tailings with PCBs. There is now a (European) regulation prohibiting the use of hydraulic oils containing PCBs, and the Municipality of Rotterdam uses only (washed) mine tailings from current production from working mines. Due to the large scale of the projects in the Waal-/Eemhaven port-area (see figure 1), the authorities asked for an Environmental Impact Assesment (EIA). In the Mine Tailings EIA (ref. 3), drawn up by the Municipality of Rotterdam in 1992, it was concluded that the use of mine tailings obtained from current production from working mines in the Ruhr region in Germany is environmentally safe in anoxic or anaerobic conditions. The Pollution of Surface Waters Act stipulates that a licence is required for the use of mine tailings. In addition the Province of Zuid-Holland considers mine tailings to be a waste product, and that consequently a licence is also required pursuant to the Environmental Management Act (WM) of the Netherlands. The Municipality of Rotterdam is challenging this standpoint in legal proceedings. In the past Public Works have carried out several studies on the technical and environmental aspects of the benificial use of mine tailings, which are presented under chapter 2 and 3.
177
2. ENVIRONMENTAL ASPECTS 2.1 General
The environmental quality of mine tailings may constitute a restriction to their beneficial use. There are three decisive factors involved: the absence of PCBs and PCDMs the oxidation of pyrite the leaching of chloride and sulphate Oil containing PCBs, PCDMs and/or other chlorinated hydrocarbons used in old mining techniques has been spilled in small amounts. This contaminated oil may be mixed with the coal and the residue, the mine tailings. When the coal is separated the oil could be adsorbed on to the mine tailings. Oil containing PCBs or PCDMs may no longer be used in mining. For this reason the authorities stipulate that tests be made to demonstrate that oil and PCBs or PCDMs are not present. The environmental licences contain an article which prohibits the use of batches "of mine tailings containing any detectable PCBs. The mineral pyrite is of natural origin, and is present in the mine tailings. In an aerobic environment, the redox potential is positive (>0mV) and oxygen is available. On contact with oxygen the pyrite may oxidize, ultimately producing sulphuric acid: FeS 2 (s) + 3.5 02 + H20 ---> Fe 2§ + 2 SO42- + 2H §
[1]
This may result in a decrease of the pH and a serious leaching of sulphate and heavy metals, available in mine tailings and/or the surrounding soil. For this reason it is important that oxidation of the pyrite be prevented. One of the conditions attached to the benificial use of mine tailing in infrastructural works is that exposure to air be avoided and that the redox potential is negative. The material is therefore only used below the water table.
2.2 Laboratory tests
Many laboratory tests have been carried out in the past in order to examine the environmental impact of the reuse of mine tailings. These tests consisted of the following methods: * contents (extraction with concentrated acid) * leaching-column test (NEN 7343) -diffusion test (NEN 7345) - shaking test (NEN 7349) The results have been compared to standards for leaching with an height of application of 10 meters. These standards are given in legislation for constructional materials in the Netherlands ('Bouwstoffenbesluit', the Building Materials Decree). The Pollution of Surface Waters Act of the Netherlands also contains standards for leaching. The results are shown in the following table.
Parameter As Ba Cd Cr Cu Hg Pb Mn Ni Zn SO, CI PAH 6 borneff PAH 10 (VROM) PAH-EPA PCBs PCDMs
Content (n = 76) 12 382 0.5 82 47 0.3 31 408 40 118 716 846 0.8 2.1 2.8 below detection level below detection level
Leaching (n = 12) 0.371 0.303 0 0.023 0.038 0 0.007 0.067 0.026 0.135 269 733
Leaching standard h=lOm
0.83 1.5 0.022 0.23 0.37 0.017 0.99 0.71 2.3 557 227
Table 1: Average contents and leaching of trace-elements and salts (in mglkg dry weight) of mine tailings from the Auguste Victoria Mine. Table 1 shows that the amounts of trace elements leached from the mine tailings meet the standards for non-isolated use in infrastructural works. Since mine tailings are mainly used in marine environments the value for chloride, which exceeds the standard, should not be a problem. However this might be a problem when the mine tailings are used in other environments.
2.3 Field experiments As mentioned above the possible oxidation of pyrite is a critical aspect in the reuse of mine tailings. To avoid contact with oxygen or air the Municipality of Rotterdam has committed itself to use mine tailings below the watertable only. Several hydraulic projects have been carried out in this manner in the Port of Rotterdam. The question was whether the oxydation of pyrite really occurs in practice. Since the use of monitoring wells would promote the contact between oxygen and mine tailings other possible methods of sampling figure 2: the Envirocone
179
and monitoring have been investigated. A survey has been carried out in which the in-situ pH and redox potential were measured in order to determine whether the environmental conditions would allow any oxidation of pyrite (ref. 4). Measurements were made with apparatus known as an 'Envirocone' (see figure 2), a cone which is pressed into the layer of mine tailings to carry out in-situ measurements. The cone was used at the 'Beatrixhaven' site which was constructed with a large amount of mine tailings, a few years ago. The cone was pressed into a layer of mine tailings. The temperature, pH and redox potential were measured at various times and depths in three locations. The results were recorded on-line. Sampling cones ('Pleistosonde, ref. 5) were also pressed into the mine tailing layers in order to sample the (low-oxygen) groundwater. The samples were analysed in a laboratory for the following parameters: pH - redox potential - oxygen -
ground
level
=
3.5
sand E1
d --" ii p l
E2
--
1.5 m -NAP, groundwater
P2 mine
tailings 12
m
-NAP
Figure 3: Measurements at one location (Beatrixhaven-site) key: El, E2: P1, P2: NAP:
in-situ measurements with Envirocone sampling points of the Pleistosonde reference level
table
m
+NAP
180
Results The redox potential as a function of depth is shown in the figure and the table below. 300 w
I
200
100
~
g
-100
,
-200
.
-300
,
-500
I 0
2
4
6 D e p t h (in
8
10
12
meters)
figure 3 example of the results of in-situ measurements at one location. R e d o x potential
Location
Envirocone (mY)
Laboratory (mY)
Oxygen (mg/I)
1-1 (5.0 m-bgl)
-193
1-2 (11.0 m-bgl)
-330
-12
2.6
2-1 (5.0 m-bgl)
-330
-212
0.27
2-2 (9.0 m-bgl)
-570
-146
0.07
3-1 (5.0 m-bgl) 1)
-420
-64
3-1 (7.0 m-bgl)
-370
3-2 (11.0 m-bgl)
-360
-163
0.96
measured in the covering layer of sand bgl = below ground level Table 2 Results of measurements of the redox potential using the Envirocone compared to laboratory analyses of the redox potential and the oxygen concentration in groundwater samples. Table 2 shows that the redox potential is negative in both the Envirocone measurements and in the groundwater samples. The negative redox potential is lower in the samples, possibly due to an interaction with oxygen during transport and analyses.
181 The results of in-situ measurements show that the level of the negative redox potential in a layer of mine tailings is such that the oxidation of pyrite will not occur, according to equation [1]. The pH values lay between 6.5 and 9, indicating that sulphuric acid was not present. This demonstrates that mine tailings are environmentally safe when used below the (ground-)water table, thereby avoiding contact with oxygen.
3.
LOGISTICS AND ORGANIZATION
3.1 Logistics
In the last few years most mine tailings used by the Port of Rotterdam have been supplied by just one producer, the Auguste Victoria Mine in Germany. Mine tailings for use outside the Marl region are transported to the Marl docks immediately after they are separated from the coal. The mine tailings are transported using a covered conveyor belt, 1.8 km in length. A sieving installation at the end of the conveyor belt is used to classify the mine tailings when this is required. The mine tailings are then loaded into cargo ships using smaller adjustable conveyor belts. The ships usually sail directly to Rotterdam, where the mine tailings are in effect a return cargo. The ships are unloaded immediately on arrival in Rotterdam, i.e. after a short delay of a maximum of one day. The mine tailings are usually dumped directly in their final location (see figure 4), using a hydraulic crane on a pontoon. In some cases the mine tailings are to be dosed using a stone classifier, and they are then discharged at an intermediate point. These methods allow the use of the mine tailings without intermediate storage ("one-to-one production and use"). This reduces the contact with oxygen.
EXISTING MINE
~
i
TAILING
WAiERL,NE~
-
SITE
i
i
/
~
.
~
~. \
figure 4: Cross-section of a filling-up project
3.2 Organization
Once the engineering of a hydraulic engineering project has been commissioned, usually by the Port of Rotterdam, then a decision has to be made whether the use of mine tailings is a serious possibility.
182
Factors influencing the choice of the material are: material/geophysical properties; environmental quality; costs; availability; planning. With the exception of the planning these factors have already been mentioned above. In the planning, the time needed for the licensing procedures stipulated by the environmental legislation will determine whether mine tailings will be used. A period of 7 89 months between the application for an environmental licence (Environmental Management Act) and the date that this comes into effect should be taken into account. The planning should also take account of the time needed.to draw up the application for the licence, and for invitations to tender. In practice a period of about a year between the first plans to use mine tailings and their actual use should be borne in mind. The time delay involved may mean that it is necessary to abandon the principle of using secondary construction materials and opt for primary materials, not requiring a licence.
3.3 Monitoring
In accordance with licences issued under the Environmental Management Act and the Pollution of Surface Waters Act a wide range of tests must be performed before and during a construction project which uses mine tailings. This means that detailed plans for the organization of the work should be drawn up in good time: prior to the execution of the construction, project information must be available from extensive and lengthy tests carried out on samples of mine tailings sampled at the mine; during the execution of the construction, project samples must be taken regularly for the purposes of a limited number of rapid tests. This involves the manager (on behalf of the licence holder), the competent authority, the supplier, and the laboratory. Prior to the commencement of the construction project, a monitoring programme is drawn up indicating which samples will be taken. However, in practice the original monitoring programme often has to be amended continually during the execution of the work due to changes in the progress, i.e. the supply of the mine tailings. The manager plays a vital central role in the monitoring programme. He is also responsible for the administration and provision of information (amounts used, results from the monitoring programme) to the competent authority. In other words the use of a secondary material such as mine tailings requires a lot of extra knowledge, organization and time from the manager.
183
Examination programme for mine tailings (w 2.2) admittance examination (sampling at the mine): a minimum of 3 sets of analytical data per year, at most 1 year old, composition (complete) tank leaching test column test cascade test preliminary examination (sampling at the mine): every 100,000 tonnes composition (complete) tank leaching test check examination / monitoring (sampling at the work): every 25,000 tonnes composition abridged tank leaching test column test
4. C O N C L U S I O N S In the last decade several million tons of mine tailings have been used in Rotterdam. The Ministry of Housing, Spatial Planning and the Environment of the Netherlands has laid down uniform regulations with regard to the use of raw and secondary building materials, which contain simple procedures (a statement is sufficient). The Department of Public Works have carried out many studies to assess the environmental impact of mine tailings. From these studies and our experiences, the following conclusions can be drawn: -It has been shown that neither the content and amount of trace elements and salts leaching into the water should constitute a problem. This means that, in practice, mine tailings can be used in an environmentally safe way, provided that the oxidation of pyrite is prevented. -In spite of jurisprudence and the results from all the monitoring programmes within the scope of licences the authorities still consider mine tailings to be a waste material. This means that a licence has to be obtained, involving a procedure which takes six months. The bottlenecks in the implementation of a policy for the beneficial use of secondary materials might then be the time required to follow the necessary procedures as compared to those for traditional materials (2 days), and the number of laboratory analyses involved (composition and leaching).
184
-The restriction of the use of mine tailings below (ground-)water level has been shown to be satisfactory. Attention should be given to the leaching of chloride from the mine tailings. An extra washing stage in the coal mining process might reduce this problem. Two processes could improve the chances for the beneficial use of mine tailings: the introduction of national certification of the quality (including simple procedures) the use of adequate monitoring techniques (e.g. the Environcone instead of monitoring wells)
Acknowledgements
We are very grateful to Paul van Eijk of the Port of Rotterdam and the staff of De Beyer, ECT and AP vd Berg for their contribution to this study.
REFERENCES 1: 2: 3:
4: 5:
Dassen W.G. et al Reuse of waste materials in constructional works, experiences in the city of Rotterdam; in "Environmental Aspects of Construction with Waste materials" (WASCON-proceedings), (1991) Gijt, J.G de et al, Results of pile load tests in mine tailings, proceedings of 'Piling, European practice and worldwide trends' (1992). Gemeentewerken Rotterdam (Public Works), 'MER-Mijnsteen', (Mine Tailings Environmental Impact Assessment with regard to the use of mine tailings for large scale filling of harbours in the WaaI-Eem haven area of Rotterdam) (in Dutch), (1992). Gemeentewerken Rotterdam & AP vd Berg Machinefabriek, Report of the investigation into the operation of the Envirocone (in Dutch), (1994). Zandbergen R.R. and J.L.M. van Leeuwen, The sampling of anaerobic groundwater in proceedings of TNO/KfK-congres Contaminated Soil '95 (1995).
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
185
UPGRADING THE USE OF RECYCLED MATERIALUK DEMONSTRATION PROJECT Dr R J Collins Building Research Establishment Watford WD2 7JR UK
ABSTRACT
This paper describes full-scale demonstration projects on recycling and the use of reclaimed materials in construction work at BRE. Included are both demolition and construction phases and, in particular, the first-ever use in the UK of recycled aggregates in ready-mixed concrete. Reclaimed bricks and reclainr.A flooring materials were also used. The projects, a new office and seminar building at Watford and a 500m 3 strong floor facility at Cardington near Bedford, also enabled the practicalities of reuse and recycling to be studied in terms of commercial, operational and contractual issues. At Watford, 96% of waste from the demolition of old buildings was recycled, including the resale of roofing sheets and timber boarding, and the salvage of other materials for internal use or donation to charities. Difficulties in the supply of suitable reclaimed and recycled materials for new construction highlighted the need for improved coordination and quality control. Overall, this increased costs, but savings were made in the demolition contract and in the use of reclainrxl flooring. Future savings of costs may be made by pre-planning for the use of waste streams for input to the design process plus a sensible approach to risk management.
186 INTRODUCTION In recent years there has been considerable interest and research into reuse and recycling of materials during the demolition and construction of buildings. In the UK, however, little progress has been made in the practical implementation of research and government recommendations. The construction of the Energy Efficient Office of the Future on the BRE Watford site was an ideal opportunity to carry out a demonstration project to identify and study the practicalities of reuse and recycling and to show the commitment of the UK Department of the Environment and of BRE to sustainable construction. As part of this research programme, concrete containing recycled aggregate was used both in the new building at Watford, and in a new strong floor facility at the BRE Cardington Laboratory at Bedford.
T H E E N E R G Y - E F F I C I E N T OFFICE OF THE FUTURE AT BRE A new office and seminar facility at the heart of the BRE Garston site - the energy-efficient office of the future - will act as a model for low energy and environmentally aware office buildings of the next century. J Sisk & Son Ltd have constructed the building to the designs of a team led by architects Feilden Clegg in consultation with BRE staff. Recycled aggregates have been used under the supervision of structural engineers Buro Happold and staff of BRE Inorganic Materials Division as well as reclaimed bricks and other reclaimed materials.
READY MIXED C O N C R E T E W I T H CRUSHED C O N C R E T E A G G R E G A T E This building incorporates the first-ever use in the UK of recycled aggregates in ready-mixed concrete. Crushed concrete from Suffolk house, a 12-storey office block being demolished in central London, has been used as coarse aggregate in over 1500m 3 of concrete supplied for foundations, floor slabs and structural columns and waffle floors. The new building has been used as a full-scale demonstration project to show how new and higher-grade uses of waste and recycled materials can be introduced in support of UK commitment to the principles of sustainable development and to assist the attainment of Government targets for the contribution of these secondary materials to aggregate supply.
SITE CLEARANCE Old buildings on the site required demolition before new construction could start. The majority of the materials were either reclaimed for reuse or recycled. Concrete and masonry from the old building were crushed on site by the demolition contractors, G J Gaywood Ltd, using portable plant. This material has been used for hardcore during construction of the new building. Gaywoods also reclaimed all metals for recycling and timber roof boards for reuse in the manufacture of pine furniture. Before final demolition of the old buildings on the site, many of the abandoned fittings were removed and donated to Hertfordshire schools through a recycling scheme run by Rotary and SATRO (Science and Technology Regional Organisation). Materials included light fittings, sockets, fire alarm equipment, office blinds, racking, filing cabinets, wood offcuts etc. An inventory of all materials is given in table 1 and indicates that 96% by volume were recycled and only 4% went to landfill.
187
Table 1 Summary of demolition waste management Matedals
Estimated Quantity
Disposal option
Bricks
500m 3
hardcore on site
Roofing sheets
300m 2
sold for reuse
Roofing timber
300m 2
sold for furniture
Slate cladding
40m 2
removed for reuse
Iron and steel
90 tonnes
sold to metals dealer
Lead and copper
1.3 tonnes
sold to metals dealer
Concrete
600m 3
hardcore on site
Fixtures, fittings, furniture
6m 3
reused
Cast iron drainpipes
10
reused
Remainder of building
50m 3
landfdl
C O N C R E T E IN THE ENERGY-EFFICIENT OFFICE OF T H E FUTURE For the foundations, a C25 mix (75mm slump) for Class 2 ground conditions was specified. According to BRE Digest 363, a minimum OPC-based cement content of 330kg/m 3 and a maximum free water/cement ratio of 0.50 are required. For floor slabs etc, a C35 mix, also with 75mm slump was specified. RMC trial mixes (all with water-reducing admixture Fosroc Conplast P250) gave the alternatives in table 2 below. (e) - estimated by interpolation
Table 2 Trial Mixes Description
OPC + ggbfs
OPC / ggbfs
28-day strength
C25 Class 2 Mix 1 C25 Class 2 Mix 2 C25 Class 2 Mix 3
375 kg/m 3 375 kg/m 3 375 kg/m 3
100/0 50/50 30/70
47N/mm 2 45N/mm 2 36N/mm 2 (e)
C35 Mix 1 C35 Mix 2
385 kg/m 3 400 kg/m 3
50/50 30/70
46N/mm z 46N/mm 2 (e)
C25 Mix 3 and C35 Mix 1 were chosen for use. A high content of ground granulated blastfumace slag was chosen for the C25 mix to maximise chemical resistance, but a slightly lower level was chosen for the C35 mix for protection against carbonation. All mixes contained 985 kg/m 3 of crushed concrete coarse aggregate apart from mixes for pumping in which this was reduced by 50 kg/m3 and the cement content increased by 10 kg]m 3. With the aid of a portable laboratory, RMC made frequent tests on the concrete delivered to site and showed that the maintenance of quality was adequate (figure 1).
188 Figure 1 Concrete mix quality control data (compressive strength in N/mm 2)
12-
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STRONG FLOOR AT BRE CARDINGTON LABORATORY A new strong floor at the BRE Cardington Laboratory was chosen to demonstrate the use of mixed concrete/masonry recycled aggregate derived from general demolition waste. The strong floor is a heavily reinforced slab 0.5m thick on which to place full-scale buildings up to eight storeys high for testing to destruction. Although the tender documents suggested that preference would be given to tenders incorporating the use of recycled aggregates, none was offered as supplies of suitable material were not readily available. Since a rapid start to the work was required, one of the tenders which was the most suitable in all other aspects (but of course quoting the use of natural aggregates) was accepted, leaving about 3 weeks to arrange and certify a suitable source. 500m 3 of concrete was to be placed in one day, 23 February 1996, and about 100 tonnes of RCA (containing up to 50% brick) would be required to replace 20% of the natural coarse aggregate The nearest source that could be found in the short time available was King's Cross, but the tonnage required in the 20-5mm size range could not be produced until a few days before casting. Another source of material was located on the other side of London near Dartford. Samples of both materials were taken for trial mixes at RMC. Both contained in the region of 40% brick masonry. From 7 day strength data it was assessed that there would be no problem in achieving a strength of at least 35N/mm 2 at 56 days (as required in the project specification) in any of the test concretes with 20% replacement of natural coarse aggregate by RCA. The mix chosen for use contained RCA from King's Cross and 385 kg/m 3 cementitious material (70% ggbfs / 30% ordinary Portland cement). No problems were encountered with production and placement, and strength data indicated that strength requirements were exceed by a wide margin.
189
SPECIFICATION AND RISK Despite a large body of laboratory research data showing that recycled aggregates can give excellent performance in concrete, practical use in the concrete industry has been restricted by questions of specification, risk, availability and cost. These issues have been highlighted in the demonstration project at BRE. In a recent study ~on the relationship between specifications and the use of recycled materials and other wastes as aggregates, it was concluded that British Standards in themselves (and in particular, BS5328 for concrete) did not prevent the use of recycled aggregates in concrete. The main reason for exclusion of such material is the wording of contract specifications which in addition demand compliance with BS882 (Natural aggregates for concrete). Only BS 1047 (Blastfurnace slag aggregate) or sometimes BS3797 (Lightweight aggregate) are allowed as alternatives. To use any other type of aggregate, evidence of suitability for purpose (as required in BS5328) is not in itself sufficient for specifiers. Whether or not the substitution of recycled aggregates for natural aggregates in concrete increases the risk of failure, modifying the specification to allow little-used non-BS materials will tend to place all of risks pertaining to its use on the specifier. This will result in unacceptable extra costs which will probably be difficult (and costly) to quantify. The only practical course, until such use of recycled aggregates becomes more common is for the client to take on this part of the risk. The client may be prepared to do this either to save costs or, more likely, to gain "green" credentials. It is possible, however, that the client may be prevented by fundholders in their own specifications from taking extra risks or using recycled aggregates. BRE as client has gone down this path both to make a closer assessment of such practical matters, and to demonstrate its confidence in the use of recycled aggregates in higher grade applications. The Department of the Environment might be regarded as "fundholder", not in this instance applying additional specifications, but having objectives similar to BRE, and in the promotion of an increase in the use of recycled materials in construction.
SPECIFICATION OF RECYCLED MATERIAL AND AVAILABILITY With all building projects, a wide range of different materials and services are used and all must be on site at the correct time to avoid delays and causing an escalation of costs. Recycled aggregates prepared for use in concrete are not generally available in the UK and thus it is difficult at the present time to specify them for use in concrete construction. Apart from general statements to the effect that BRE expected the main contractor to assist in the incorporation of the latest of environmentally aware practices, no absolute commitment was made in the BRE project to the use of recycled aggregates in concrete. This aspect was "retrofitted" to the job specification, and accepted by the contractors as long as there was no extra cost to them and no hold-up in the supply of materials. Any difficulty in this respect would result in reversion to the use of natural materials. Further aspects of risk and specification are considered in references 2-4.
COSTS OF USING RECYCLED AGGREGATE The need for trial mixes, quality control etc. has cost implications. Obviously, as a market develops, such costs will be reduced but are likely to remain higher than for natural materials. Costs will also be incurred for the higher cement contents often required by recycled aggregates. Such costs may eventually be offset by the increasing cost for landfill. An extra cost in the BRE
190 project was for increased transport requirements. This was accepted in order to obtain the most secure supply lines of material and ensure that this "one-off.' demonstration project had the greatest possible chance of success.
BRICKS Around 80,000 reclaimed facing bricks were used for the external cladding of the BRE Energy Efficient Office of the Future. Whole bricks could not be salvaged from the old buildings on site because of the strength of the mortar bonding them together. The material was thus crushed on site as hardcore for the new building. Insufficient time was available to test these materials for use as aggregate in concrete for the new building. It was proposed to purchase the bricks from the demolition of a hospital situated about 2km away. These bricks would have been considerably cheaper than new ones. Unfortunately, the demolition was delayed beyond the time that the bricks needed to be on site and bricks were obtained instead from a quality assured reclamation company. Greater costs were thus incurred in addition to extra charges from the both architect and main contractor resulting from the use of imperial size bricks in a building designed for metric bricks. Buiklings more than about 60 years old contain hydraulic lime mortars which assist considerably in the reclamation of whole bricks. Consideration was given to the use of hydraulic lime mortar for the new building but this would not give sufficient lateral strength to the modern thin-walled cavity construction. Lime mortars were generally used before cavity wall construction was introduced; sufficient lateral strength was given by the thickness of the walls. Lime mortars could have been used in the outer leaf of the new building if the thickness of the internal load-bearing blockwork were increased from 140 to 190 mm. This could not be justified in terms of the efficient use of material resources.
FLOORING Hardwood parquet block flooring removed during refurbishment work at the former County Hall building in London has been relaid in the new BRE building at Watford. Only 300m 2 of the 18,000m 2 available were required. Despite the extra costs of cleaning and laying, an overall saving of 30% was made compared with the costs of equivalent new material.
DISCUSSION AND CONCLUSIONS The construction of the new Energy Efficient Office of the Future at BRE has proved to be an excellent test bed both for demonstrating new recycling methods and for highlighting the dit~ulties encountered. The main problem was in sourcing suitable materials in sufficient time. This caused the main worry both for the graded recycled aggregate and also for a suitable source of reclain-ed bricks to be used in the external cladding of the building. There is an argument that materials should be collected together in advance, but this needs a considerable amount of preplanning, and clients and contractors alike do not wish to have large stockpiles lying idle for long periods. Such pre-planning, however, could allow clients and their architects to specify recycled material from the outset rather than substituting recycled materials afterwards as in this project. The "substitution" aspect of the BRE project, however, has been quite instructive. This aspect
191 of the project could well have been lost if a building had been designed with recycled and reclaimed materials in mind. The construction of the new strong floor at Cardington has illustrated how recycled materials might be incorporated in concrete on a routine basis. No significant effect on concrete handling, maintenance of high production rates, placement properties, strength development etc. appears to be caused by replacing 20% of the natural coarse aggregate by RCA. Also the RCA in this case should potentially be more widely available, being produced from general demolition waste containing a mixture of concrete and brick. Because of the high production rates required for the Cardington project, a readymix plant was totally dedicated to the one project and a silo was available for the exclusive use of the RCA. Provision of extra storage capacity for RCA on a more general basis would require a fairly general use of RCA to be cost-effective at current price levels. The next hurdles are (a) to pave the way for a greater availability of RCA of the correct grading and sufficient quality for use in concrete eg by a pilot quality control scheme, and (b) to establish acceptability of the product on the market. Some clients may be seeking "green" credentials, but at some stage in the future cost savings may also be obtainable. Job specifications for projects may cause some problems but these are not insuperable. BRE was able for its own construction projects, to use its wide range of expertise in discussing and overcoming specification problems where possible. The experience from this should in new projects assist the development of appropriate action in relation to standards and specifications and other barriers to utilisation.
REFERENCES Collins, R. J. and Sherwood, P. The Use of Waste and Recycled Materials as Aggregates: Standards and Specifications. HMSO, London, 60pp. (1995) 0
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Collins, R. J. Increasing the Use of Recycled Aggregates in Construction in Concrete for Environment Enhancement and Protection, pp73-80. Concrete in the Service of Mankind Symposium, University of Dundee, 24-28 June 1996. Collins, R. J. Recycled Aggregates in Ready-Mixed Concrete in Sustainable Use of Materials. BRE/RILEM international seminar, Watford, 23-24 Sept 1996. Hobbs G. and Collins, R. J. Demonstration of reuse and recycling of materials: BRE Energy Efficient Office of the Future. BRE Information Paper IP3/97, BRE, Wafford (1997).
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193
Beneficial use of contaminated sediments within the Meuse river-system 1
A.L. Hakstege a, J.J.M. Heynen b, H.P. Versteeg c Ministry of Transport, Public Works and Water Management, Civil Engineering Division, P.O. Box 20000, 3502 LA Utrecht, The Netherlands IWACO B.V., Consultants for Water and Environment, P.O. Box 8520, 3009 AM Rotterdam, The Netherlands Ministry of Transport, Public Works and Water Management, Road and Hydraulic Engineering Division, P.O. Box 5044, 2600 AA Delft, The Netherlands a,b,c
Ministry of Transport, Public Works and Water Management, Project office Zandmaas/Maasroute, P.O. Box 1593, 6201 BN Maastricht, The Netherlands
Abstract Sediments in the Meuse valley are contaminated on a large scale by diffuse sources. The project "Zandmaas/Maasroute" aims at the enlargement of the discharge capacity. When realizing the project, a full-scale clean-up operation would not be realistic and not effective because of recontamination and the enormous scale and costs involved. However, there will be a unique opportunity to break this deadlock, if the concept of dynamic soil management is applied. This approach to remediation means putting back contaminated sediments of indigenous quality within the river system. Preconditions are determined by risk assessment and local conditions. The first aim is improvement of the environmental quality of the river system. Other objectives such as the development of natural areas, the mitigation of geohydrological effects and the exploitation of sand and gravel can be realized by several forms of beneficial use of contaminated sediments. Dynamic soil management is a pragmatic and cost-effective solution for the problem of contaminated sediments.
1. INTRODUCTION Large parts of the Dutch Meuse valley were flooded in December 1993 and January 1995. Consequently, a governmental advisory committee ("Boertien II") was ordered to assess the situation and to give recommendations on measures to reduce the risk of flooding. Among others, the commission recommended to enlarge the main channel of the Meuse over a stretch of about 170 km called "Zandmaas". Apart from that, the river Meuse should be adapted to improve the river for navigation. Flood risk reduction and
lDisclaimer: this article represents the personal views of the authors and does not necessarily represent the official views of competent authorities.
194 navigation improvement are the main targets of the project "Zandmaas/Maasroute" A third objective is development of more natural areas in the Meuse valley. Several alternatives to reach these three targets are being developed and will be elaborated in an environmental impact assessment (EIA) to be accomplished in 1998. The following measures to enlarge the river discharge capacity are being considered: deepening of the main channel; widening of the main channel; lowering of floodplains; excavation of flood channels.
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The realization of these measures will result in the removal and handling of the following materials: useful and marketable mineral resources (gravel,sand); clean non-marketable soil; contaminated sediments and dredged materials. -
The management of contaminated sediments is of crucial importance for the project in terms of environmental, logistical and financial aspects. For the project Zandmaas/Maasroute the amount of contaminated sediments and dredged sediments to be removed are estimated at ca. 5 to 11 million cubic metres.
2. DYNAMIC SOIL MANAGEMENT Since the Roman Ages and more intensively during the last two centuries, the Meuse river system has been diffusely contaminated by upstream industrial activities [1]. The predominant contaminant is zinc. Other contaminants to be found are cadmium, PAH, PCB, DDT. Zinc levels in the river sediments frequently exceed Dutch intervention levels. The pollution is mainly confined to the top soils in the floodplains and banks of the river Meuse which are contaminated on a large scale. Dutch regulations prescribe that severely contaminated soils should be cleaned or properly disposed of in controlled large disposal sites. However, a full-scale clean-up operation of the river system Meuse is not feasible and not effective, because of: recontamination because of the ongoing industrial pollution upstream as well as erosion and redeposition of older contaminated sediments; the huge scale and enormous costs involved; general shortage of nearby existing disposal sites and public opposition against new locations for disposal (NIMBY). -
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Many river widening and nature development projects stagnate because of these problems. Faced with these issues and the necessity to realize megaprojects in river systems, Dutch authorities formulated guidelines for a different approach to remediation [2]. This approach is called dynamic soil management. The main objective of dynamic soil management is to improve the actual environmental situation in the most efficient way. Where improvement is not possible, e.g. because of
195 recontamination, the situation should at least not deteriorate (stand still principle). The central issue is that the main part of the contaminated sediments, which are excavated due to widening of the river, are not transported to a large disposal facility but are stored nearby in the river system. The project Zandmaas/Maasroute offers the opportunity to elaborate this new approach into a design for a sustainable solution of the contaminated sediment problem. Preconditions for a design based on dynamic soil management are: Environmental improvement should be achieved by concentrating and, if necessary, insulating the replaced contaminated sediments in such a way that the emission of contaminants to surface and ground water as well as dispersion by erosion is reduced. Furthermore contact possibilities for men and organisms with contaminants have to be diminished. The possible solutions depend on local conditions. The environmental risks for men and ecology should be acceptable and risk assessment should be related to land use. This means different standards for agriculture, recreation, nature etc. An important finding in this respect is that zinc concentrations have no apparent relation with enviro-toxicological effects [3]. Acidity and cation binding capacity of the soil are the main parameters controlling behaviour of zinc. Leaching tests (column test NEN 7343) on zinc contaminated sediments and soils in this project and in other studies [4,5] show that even when zinc concentration levels are very high, zinc leachability can be very low. Consequently, zinc availability for organisms are likely to be low. For so-called "Hot-Spots" the traditional remediation approach should be followed. In case of severely contaminated sites this means removal and transportation to a regular disposal site. Hot spots may either be contaminated from local sources and contain other contaminants than indigenous to the region, or represent much higher contaminant levels than regional background value. The assessment of environmental soil quality in relation to local back-ground levels to distinguish "HotSpots" needs to be further elaborated. Within the river system zones with more or less homogeneous soil qualities are distinguished, based on frequency of flooding, soil type and land use. Displacement of contaminated soil of indigenous quality should be possible within the same zone. The design should lead to beneficial use dependant on local conditions (see next section). -
The measures should fit within the Dutch legal framework.
196 3. BENEFICIAL USE The concept of dynamic soil management is optimized by combining environmental improvement with other benefits such as the upgrading of nature and landscape, ecological improvement and revenues from minerals. These possibilities depend on local conditions: e.g. geohydrological, hydraulic and river morphological aspects and the occurrence and depth of mineral deposits. Possible options for replacement of contaminated sediments in combination with beneficial use are: re-use as topsoil for the new banks. After excavation and dredging activities, needed for widening and/or deepening the river, the former contaminated top soil is replaced as the new topsoil. In this way the function of the soil remains the same its new location. When reconstructing the riverbank, the development of natural habitats is promoted by a gradual and non-protected slope: a so-called ecological riverbank. hydrological sealing of flood channels. When flood channels are constructed, clayey materials are needed to seal off the channel from groundwater flows (preventing drainage). Contaminated soils released in reconstructing the river are usually quite suitable for this. hydrological barriers. The contaminated top soils, which are mostly clays with a low water permeability, are replaced in a vertical shield near the riverbanks. This clay shield functions as a geohydrological barrier for groundwater flows towards the river which causes a rise of groundwater level in the adjacent area. In this way further aridization of ecological valuable natural areas is prevented. exchange of mineral resources with contaminated sediments in the river channel. After widening of the river the contaminated sediments are replaced in a pit in the river channel, which is made for this purpose. The location of these pits is determined by the occurrence and depth of valuable gravel and sand deposits and a favourable geohydrological situation, where the groundwater flow is directed towards the river. ecological improvement of existing deep gravel pits When existing deep gravel exploitation pits can be made shallower, this may lead to an improvement of the ecosystem. Dredged materials (clean or contaminated) may very well be used for this purpose [6]. On some locations several types of beneficial use can be combined e.g. hydrological barrier and exchange with minerals.
197 4. PILOT P R O J E C T A pilot project is planned to be carried out in 1998 in a small river section of seven kilometres. The design is based on dynamic soil management for the widening of the main river channel, which is partly completed with ecological riverbanks. The main objective is to gain experience for the project Zandmaas/Maasroute on the aspects of engineering, a design based on replacement of contaminated sediments, the acquisition of environmental permits and the prediction of effects on the river system and environment. In suitable parts of the river section four methods of replacement of contaminated sediments in combination with beneficial use are planned: re-use as top soil for the new river banks; a clay shield as hydrological barrier; exchange of contaminated sediments with sand and gravel; ecological improvement by the filling of an abandoned gravel pit. -
Three methods are displayed in figure 1. An extensive monitoring programme will be carried out on environmental, hydrographic and geohydrological effects. Environmental monitoring comprises chemical and ecological parameters before, during and after the reconstruction. An estimation is made of potential emissions from local storage of contaminated sediments based on data from detailed soil investigations. This model will be validated by the monitoring results.
5. O U T L O O K Dynamic soil management is a practical and cost-effective remediation approach for the project Zandmaas/Maasroute. This approach will soon be formalized as the remediation policy for the floodplains of large rivers in The Netherlands. The concept of dynamic soil management is now elaborated by the competent authorities into a regional remediation policy for the river Meuse. The pilot project will serve as a first test case for this new approach. Several river reconstruction alternatives in which options for beneficial use are included will be considered in the environmental impact assessment.
198
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Figure 1. Options for beneficial use of contaminated sediments.
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199 6. REFERENCES Hakstege, A.L., Kroonenberg, S.B. and van Wijck, H., Geochemistry of Holocene clays of the Rhine and Meuse rivers in the central-eastern Netherlands, Geologie en Mijnbouw 71: 301-315 (1993). Weggemans, J., Dynamic soil management in river systems (in Dutch, Actief Bodembeheer Rivierbed), unpublished policy work document, (Dec. 1996). Kok, A.P., Beneficial use of contaminated sediments in flood plains (in Dutch, Toepassing klasse 4 in natuurontwikkeling uiterwaarden), unpublished internal report, Bouwdienst Rijkswaterstaat (1996). Heynen J.J.M., Comans, R.N.J, Honders, A., Frapporti, G., Keijzer, J. and Zevenbergen, C., Development of fast testing procedures for the determination of leachability of heavy metals contaminated soils, these proceedings. Comans, R.J.N., Modelling processes controlling metal leaching from contaminated and remediated soils, these proceedings. Koethe, H. and Bertsch, W., Legislative and technological requirements in Germany for the subaquatic disposal of contaminated dredged material in gravel pits. Proceedings CATS III p. 69-79 (1996).
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Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
201
I N T E G R A T I O N OF TESTING P R O T O C O L S FOR EVALUATION OF CONTAMINANT RELEASE FROM M O N O L I T H I C AND GRANULAR WASTES D a v i d S. K o s s o n 1 and H a n s A. van der S l o o t 2
~Rutgers, The State University of New Jersey Dept. of Chemical and Biochemical Engineering P.O. Box 909 Piscataway, NJ 08855-0909 USA
2Netherlands Energy Research Foundation Westerduinweg 3 P.O. Box 1 Petten, N.H. 17 55 ZG The Netherlands
ABSTRACT Testing for evaluation of contaminant release from soils and wastes is necessary to compare management options and project potential environmental impacts. Frequently, the available management options include disposal as either an untreated or treated granular or monolithic waste form with varying extents of water contact. In addition, some wastes may be suitable for utilization as fill material or aggregate replacements in construction applications. Selection amongst these choices requires substantially more intensive leaching testing and evaluation then is provided in most regulatory frameworks. Ideal testing and evaluation protocols would permit the necessary comparisons, provide realistic estimates of contaminant release under the scenarios evaluated, and be accomplished through use of a minimum number of a~mlyses carried out over a very short time interval. A series of testing protocols has been under development for granular and monolithic wastes. The goal has been a series of tests that can be carried out in less than approximately one week with approximately six samples for chemical analysis that provide measurement of fundamental leaching parameters. These fundamental parameters then are used to estimate contaminant release under different management scenarios. Substantial reductions in testing requirements can be achieved through integration of the individual test methods. This paper will present the testing evaluation framework, using data obtained from several wastes to illustrate the benefits of the framework. INTRODUCTION Emphasis on sustainable development has focused increased attention on management of process residuals to attain the maximum practical use from materials previously considered only suitable for disposal. Thus, the societal definition of a "waste" must be redefined to reflect new alternative management options, including both utilization and disposal. The following definitions of "waste" have been offered to reflect this change: "waste" shall mean any substance or object which the holder discards or imends or is required to discard (EEC, 1991). "waste" is not an indication of a specific property. It merely describes a situation, a transition in the life of a given product/material namely from being useful to a given purpose to being no longer needed by the owner (EEC, 1996). "waste" might still be quite useful for somebody else or for other purposes (EEC, 1996). Whether a product/material is harmful or not is determined by its inherem hazardous properties (e.g., toxic, explosive, radioactive, etc.) and the way the product/material is handled in the environment both during its useful life and waste life (EEC, 1996). "wastes" per se are not inherently harnlful or benign (EEC, 1996). These definitions necessitate defining the characteristics of a waste in the context of both its constituents which potentially may be harmful and the manner in which it is managed. Thus, tools are needed to infer the behavior of waste constituents during a variety of management options. These management options
202
may include utilization as a replacement material in a variety of applications (e.g., as aggregate in road construction or brick manufacturing), or disposal under varying conditions (e.g., monofill, landfilling with other wastes under defined conditions). Leaching tests are used as a tool to estimate the release of constituents from solid wastes during utilization and after disposal, assess efficacy of waste treatment processes, and develop endpoints for remediation of contaminated soils. Usually the goal of the testing is to answer the question "Is the selected management option for the waste environmentally acceptable?" However, the answer to this simple question requires consideration of several inter-related aspects, including (i) the release rate and extent of potentially hazardous constituents from the waste, (ii) attenuation processes associated with the constituents of concern as they migrate from the waste to the receptor being considered, and (iii) the inherent toxicity of each specific constituent. Considerable effort has resulted in accurate assessment techniques and data for evaluating contaminant attenuation and toxicity for several cases. In contrast, estimation of constituent release by leaching most often assumes either (i) the total content present is available for release, (ii) the contaminant concentration in the leachate will be equal to that measured during a single batch extraction and is constant with time, or (iii) the fraction of the contaminant extracted during a batch extraction is equal to the fraction that will leach (USEPA, 1990; Goumans et al., 1991; Goumans et al., 1994). These approaches frequently result in grossly inaccurate (both over- and underestimation) of actual release, which in turn, forces disposal of materials which are suitable for beneficial use, forces remediation of soils to levels beyond that necessary for environmental protection, and unnecessarily depletes disposal capacity. Erroneous estimation of leaching behavior also results in treatment processes that "pass the test" rather than improving waste characteristics. For example, a treatment process for a particular waste stream has been shown to reduce the extracted concentration for a regulatory test (TCLP) but result in increased release compared to management scenarios without the treatment (Garrabrants, et al., 1996). Thus, methodologies which result in a more accurate estimate of contaminant leaching may both improve environmental protection through more efficient use of resources and be economically beneficial. Motivating concerns against adoption of testing methodologies that provide more realistic estimates of leaching include (i) increased waste management costs would result from more detailed waste characterization, (ii) detailed characterization would result in processing delays, and (iii) more stringent waste treatment methods would be required. The objective of this paper is to present an integrated framework for solid waste testing that provides more realistic estimates of contaminant release, greater flexibility for waste management, and is economically reasonable. This framework was developed based on current understanding of release of inorganic waste constituents; however, an analogous approach would most likely be applicable to organic species as well. Emphasis is placed on testing that is in support of waste utilization. W H A T R E L E A S E I N F O R M A T I O N IS NEEDED? Development of a set of protocols to evaluate environmental acceptability of waste management options must begin by defining the specific questions to be answered and overarching characteristics of the particular waste. Typically an answer to one of the following three questions is being sought: 1. Is this waste suitable for a particular utilization, treatment, or disposal option? 2. Which specific utilization, treatment, or disposal options are appropriate for this waste? 3. Which wastes are most appropriate for a specific utilization, treatment, or disposal option? The first question is the most frequently asked by the generator of limited quantities of waste and by a regulatory agency trying to enforce environmental protection. The comparison is limited to a single waste (type or fixed quantity) and a single selected management option. The second question is appropriate for a generator of a specific waste type on a continuing basis. Here, evaluation of a variety of management options for the waste may be motivated by economic, regulator3, or long-term sustainability considerations. Example wastes include coal combustion ash and municipal waste incinerator ash that may be either
203
landfilled or beneficially utilized. The third question is asked by the waste recipient. This is in response to multiple wastes competing for a specific utilization option (e.g., use as aggregate in road base) or disposal facility. Overarching characteristics of a waste that must be considered include quantity and production constancy (both quantity and quality as a function of time) 1, physical characteristics (e.g., particle size, water content), inherent extreme characteristics 2 and public acceptability 3. The goal of waste testing should be to provide information about contaminant release from a waste in the context of the anticipated disposal or utilization conditions. Thus, testing should reflect the conditions (e.g., pH, water contact, etc.) that will be present in the waste and at its interface with its surroundings during the long-term, which may be significantly different than the properties of the material immediately following production 4. Three levels of testing can be defined to efficiently address the above waste management questions. In all cases, the material to be tested should be the material as it is going to be used or in its final form, not an untreated form if treatment is going to occur prior to utilization or disposal. Detailed characterization (Level I) is required for initial characterization of a high volume waste stream which may be either disposed on a continuing basis or utilized. Level I testing is used to define the primary waste characteristics (e.g., composition, physical properties, and leaching parameters) and variability associated with critical parameters. After Level I testing is completed, quality control testing (Level III), which requires the least amount of testing of the three levels, can be used to insure constancy of critical waste characteristics within previously established limits of variability. The specific measurements and frequency of testing needed for Level III testing are selected based on both the specific properties of the waste and the critical requirements of management option. Quality control data should be evaluated in the context of the detailed characterization data and be used as an indicator that a major change in critical waste properties has not occurred. Thus, it is anticipated that only specific elements of concern previously identified would be measured using very simple test procedures. Random testing of a more complete set of species with concise testing (Level II) can be used as a verification and a regulatory enforcement tool without the added expense of constant testing. On-site verification testing may not be relevant to utilization because more stringent quality assurance would be required. Level II testing also is used for determining if low volume waste is suitable for a specific disposal option or more detailed assessment of a waste that fails quality control testing. A decision flow diagram for application of the three testing levels is presented in Figure 1.
1 The limited quantity or high degree of variability associated with a waste may make utilization technically or economically impractical, or not expeditious. 2 The origin or composition of a waste may result in characteristics which prima facie preclude certain management options (e.g., high inherent toxicity, extreme acidity or corrosivity). 3 Public acceptance of certain waste management options may be limited by specific waste characteristics (e.g., noxious odor) or political considerations. 4 Examples where the material as produced has different constituent release behavior then during utilization are: (i) concrete pillars immersed in surface water where release reflects the neutral pH of surface water rather than the alkali pH of Portland cement concrete, and (ii) use of steel slag in coastal protection applications where V and Cr leaching is reduced by the natural formation of ferric oxide coatings in the utilization environment (Hockley and van der Sloot, 1991).
204 WASTE (FOR DISPOSAL OR BENEFICIAL USE)
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Figure 1. A decision flow diagram for use of different levels of leaching tests to evaluate a waste. Levels I, II, and III are detailed characterization, concise testing, and quality control testing, respectively.
For low volume wastes, Level II testing provides a decision between two types of disposal (nominally, "non-hazardous" or "hazardous") or provides the option of more detailed characterization to facilitate selection of treatment or alternative management options. For high volume wastes, the first decision point is whether or not the generic waste type under consideration has been evaluated in detail for the management option being considered. If this has not occurred, Level I testing is required. If testing of the generic waste type for the specific application has occurred previously, the second decision point is whether or not the waste from the specific generating facility has been characterized in detail. If Level I testing on the specific waste had been carried out, then quality control testing (Level III) can be implemented. If not, then Level II testing is used to verify that critical properties of waste produced from the specific facility are consistent with the previously characterized generic waste type. In summary, the questions to be answered by each level of testing are as follows: Level I What actions are required to make the material acceptable for disposal or utilization? What options are available to make desired changes in material properties? Is the material acceptable for utilization? Level II Is the material (low volume) acceptable for disposal (controlled environment)? Level III Is this material the same as the material previously characterized?
205
For bulk wastes which are managed largely through a small set of options, field testing should be carried out at selected locations and time intervals aider utilization or disposal to allow verification laboratory based release estimates. In this manner, laboratory and field measurements can be used in a iterative manner to improve predictive capabilities. One clear advantage of this decision scheme is that it permits waste management decisions to be based on detailed knowledge of the waste characteristics and it's expected behavior under the selected management option rather than based on a single screening test that may have little relevancy to the actual waste behavior. Detailed leaching characterization of a waste requires measurement of intrinsic parameters which control constituent release. Two modes of constituent release, local equilibrium and diffusion controlled release, can be defined to establish which parameters are important for a particular combination of waste type and management option. Local equilibrium at the boundary between the waste and infiltrating water can be assumed when the Peclet number (Pe) is less than 1 (Massry, 1997) where Pe = R = Dobs =
L v
= =
(RZ/Dobs)/(L/v), median particle radius, observed diffusivity of the element through the porous solid matrix, length or depth of fill through which infiltration percolates, and seepage velocity.
This ease occurs with granular wastes and seepage velocities are relatively slow. Conversely, diffusion controlled release can be assumed when Pe is greater than 10. This case occurs with monolithic wastes, granular wastes compacted to low permeability, or granular wastes overlain by low permeability covers or surrounded by materials with much greater permeability 5. For all diffusion controlled release scenarios, water flow is predominately around, rather than through, the majority of the waste. Fortunately, most utilization and disposal scenarios result in one of the above limit cases. Natural infiltration usually results in slow percolation rates and when infiltration is controlled by capping, flow rates are reduced even further. Thus, local equilibrium is a good assumption in these cases. Caution must be used when assuming either of these limit cases if large pH or redox gradients exist at the interface between the waste and it's surrounding environment. Large pH or redox gradients can result in precipitation or rapid dissolution phenomena for some elements as gradient boundaries redistribute over long time intervals (Sanchez, 1996; van der Sloot et al., 1994). For most alkaline wastes, the most prevalent interface reaction is absorption of carbon dioxide, resulting in the formation of carbonate species and pH decreases towards 8. An appropriate assessment of these conditions requires evaluation of the change in leaching in response to external factors. Thus, testing under conditions imposed by the external factors rather than by the waste may be necessary. Table 1 presents the parameters suggested for each level of testing. There are multiple test methods available for measurement of several of the parameters. Integrated concise testing protocols are presented in more detail in sections of this paper which follow. For the purposes described here, availability is defined as the quantity of a particular element or species which is not tightly bound (e.g., in an amorphous silica matrix) in the waste matrix and potentially may leach. Thus, it serves as the thermodynamic driving force for release. Most often solubility controls aqueous concentrations at local equilibrium because substantially more of a constituent of concern is present in the solid phase than required to saturate the solution. The solubility of many potentially hazardous elements is strongly a function of pH, thus the titration of pH and solubility is required. Integration of measured parameters to estimate long-term release for the cases of local equilibrium and diffusion controlled release has been presented previously in detail (Kosson, Sloot and Eighmy, 1996; Schreurs et al, this conference). Both cases require site-specific information about the management scenario (Table 2) to provide release estimates.
5 For the case where 1
206
Table 1. Suggested parameters to evaluate leachin~ for different levels of waste testing. Release Mode 1 Parameter L.E. D.C. Test Methods Level I I - Detailed Characterization 2
Availability
+
Solubility vs. pH (4
+
Acid neutralization capacity (4
+ +
NEN 7341 (1995); EDTA extraction (Garrabrants, et al., 1997) Static pH (IAWG, 1997; Comans et al, 1993.); Acid neutralization capacity (Kosson et al., 1993) Same as for solubility above NEN 7345 (1996) or draft NVN 7347 (compacted granular; van der Sloot et al, 1997; Kosson et al, 1993) NEN 7343 (1996) column leaching Draft NVN 7348 (1994)
Matrix interactions + Redox capacity ?7 Particle size distribution + Porosity + Hydraulic conductivity + Level I I - Scree. ning Evaluation 3 Concise protocols as described in text. Availability + + Acidity or alkalinity, + + Release at LS 2 & 10 at pH controlled by + PrEN 12457 (CEN TC 292, 1996) waste Release at pH 8, LS 10 + Diffusivity + Level III - Quality Control 4 Availability 98 98 Release at LS 10 (controlled pH) 5 + Release from monolith or compacted + granular sample 6 Random use of Level II testing for + + Concise protocols as described in text. resulatory verification (ca. 4/yr) ~Local equilibrium (L.E.) or diffusion controlled (D.C.). 2Elements and species analyzed in extracts to include potentially hazardous elements (e.g., Pb, Hg, Cr, etc.), principal cations (e.g., Na, Ca), principal anions (CI, SO4, CO3) and dissolved organic carbon. 3Elements and species analyzed in extracts to include potentially hazardous elements (e.g., Pb, Hg, Cr, etc.), total dissolved solids, and dissoh,ed organic carbon. 4Elements and species analyzed in extracts lilnited to minimum necessary to verify constancy of waste. 5Single extraction at pH controlled to constant pH consistent with natural pH of waste or pH =8.0. 6Leachant refreshed at 1, 8 and 24 hrs with extracts combined for single cumulative release estimate. 7Only if waste origin or initial testing suggests strongly reducing matrix. 8Used if Level I testing indicates wide variability in availability of potentially hazardous elements.
207
Table 2. Intrinsic waste properties management option characteristics needed to estimate constituent release for a specific manasement option. Intrinsic Waste Properties Management Option Characteristics (local equilibrium or diffusion controlled release) (disposal or utilization alternatives) Availability Application Geometry (length, width, depth) Multi-phase Partitioning Application Boundary Conditions Solubility as a function of pH Low permeability caps or liners Desorption partitioning 1 Drainage layers Diffusivity Water Balance Observed diffusivity Precipitation frequency Net infiltration and seepage velocity Tortuosity Waste Controlled Chemical Influences External Chemical Influences Reducing or oxidizing conditions Alkalinity or acidity Reducing or oxidizing potential and capacity2 pH range Constituent matrix interactions3 Dissolved organic carbon or complexing agents Physical Characteristics Temperature Variations Annual mean temperature8 Particle size distribution4 Monolith dimensions5 Exposure to Frost or freezing9 Hydraulic conductivity6 Porosity7 ~Required only when constituent is present as a sorbed species rather than a solid phase. This distinction can be obtained from the slope of the extraction data as a function of LS in the concise protocol. 2Required only for wastes that exhibit extreme reducing potential or oxidizing potential during initial testing (e.g., blast furnace slag). 3preliminary indication of this condition is obtained from the slope of the extraction data as a function of LS in the concise protocol. Includes dissolved organic carbon as potential chelating agent and and depletion of solubility limiting species such as sulphate. 4Required for granular materials only. 5Required for monolithic materials only. 6Usedto estimate seepage velocity and infiltration rates. 7Can be used to estimate diffusion resistance or tortuosity (see Schaefer, et al., 1995). 8Used to calculate constituent diffusivity in water. 9Needed to evaluate requirements for physical durability of waste form.
CONCISE TESTING APPROACH Understanding of the specific testing objectives and the controlling variables allows for integration of testing. Use of integrated, or concise, testing protocols permits coverage of a broad range of leachability controlling aspects with a minimum of experimental work. The goal is to include as many controlling factors as possible with a minimum number of extracts and analyses. Application of concise testing protocols is targeted towards Level II testing and can be used for screening as an initial step for Level I testing. An example of a concise testing protocol for granular wastes, along with test results for a lead and zinc smelting slag, is presented in Figure 2. The test method can be summarized as follows (van der Sloot et al., 1994): P.art A. 2-Step Serial Batch Test Deionized water as extractant, no pH control (waste controls extract pH) Particle size reduction with 95% to less than 4 mm Extraction 1 - LS6=2 ml/g, contact period of 6 hr, well mixed, closed vessel Extraction 2 - LS=8 ml/g (cumulative LS= 10 ml/g), contact period of 18 hr, well mixed, closed vessel Measurements (each filtered extract): pH, Eh, TDS 7, DOC 8, conductivity, relevant major 9, minor or trace elements 61iquid'to-solid ratio (LS) 7 total dissolved solids (TDS)
208 Part B. pH Controlled Serial Batch T.est Deionized water as extractant with specified pH control Particle size reduction 95% to less than 300 lam Extraction 1 - LS=10 ml/g, pH=8 or 12, contact period of 24 hr, well mixed ~~ Extraction 2 - LS=50 ml/g (cumulative LS=60 ml/g), pH=4, contact period of 24 hr, well mixed Measurements for each extraction: acid or base consumption, DOC, Eh, relevant major, minor or trace elements (filtered extracts) From the above combination of 4 extractions several leaching considerations can be evaluated. Part A of this procedure facilitates distinction between solubility, availability, desorption and matrix interaction controlled release ~. Comparing release at pH 8 (LS=10) to that at the material's own pH (cumulative release after second extraction) permits evaluation of the extent of constituent retention in the matrix relative to very mobile species (e.g. Na, CI). EH measurement (relative to pH) provides indicates if the matrix has reducing properties. DOC measurement proivdes an indication of the potential for chelation effects for mobilization of metals. TDS provides for evaluation of the potential impact from total soluble salts. Part B of this procedure permits evaluation of leachability changes in crucial pH domains (pH 4 , 8 , own pH) and estimation of availability. In addition, acid/base neutralization capacity is determined from acid or base consumption used to control pH at the designated levels. This information is relevant for determining how long a material can impose its own pH conditions on a leachate. DOC measurement during these extractions is used to evaluate the potential for future mobilization of DOC. Finally, the combination of leachability and pH data from the four extractions shows the potential sensitivity of leaching to changes in pH domain relevant to conditions observed in practice. It also provides an indication of the likely changes in leachability when leaching conditions change in the long term (e.g., pH changes resulting from uptake of biologically generated carbon dioxide, remineralization, or acidification). Results for a primary slag from a lead and zinc smelting process are illustrated in Figure 2, where concise results are compared to more detailed characterization testing. Solubility controlled release and oxidized conditions is indicated for lead (Mandin et al., this conference). Concise results are consistent with the more extensive test methods.
8 dissolved organic carbon (DOC) 9 Relevancy is based on either regulatory significance or l~aown presence of significant content in the waste being evaluated. 10The objective of this protocol is to obtain leaching data for three pH values: (i) the pH of the pore solution of the material, (ii) pH 4, and (iii) pH 8. If the pH of the material is approximately 8, the pH 8 extraction (Part B, Extraction 1) is replaced by using the same conditions but controlling the extraction atpH 12. ~1 Cumulative release plotted as a function of LS can be interpreted as follows: slope= 1, solubility control; slope< 1 and cumulative release approximately equal to availability, availability control; slope 1, matrix interaction control (e.g., depletion of sulfate during initial extraction resulting in greater solubility during second extraction).
209
TIME D E P E N D E N T RELEASE AND RETENTION 1000
Availability C
100
pH 5.3 -6
o
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SOLUBILITY C O N T R O L L E D C O N D I T I O N S
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l'
SERIAL BATCH TEST LS=2; 6 hrs; c l o s e d v e s s e l LS=2-10; 18 hrs; c l o s e d v e s s e l pH C O N T R O L L E D TEST LS=10; pH=12 control; 24 h r s LS=10; pH=4 control; 24 hrs R e c o r d pH, E
n, DOC, TDS
and Conductivity.
C O N S I S T E N T pH BEHAVIOUR
R E D U C I N G PROPERTIES O N L Y AT pH < 9 900 ~ . ~ _
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Acid Neutralization Capacity is low ( < 0.1 mol/kg): sensitive to external controlling factors. Conductivity low. DOC not relevant in this material.
Figure 2. A concise testing protocol for granular materials and concise testing results for a lead and zinc smelting slag in comparison with results from detailed characterization. The leaching of pulverized coal ash is presented as a second example of application of the concise protocols to a granular material in Figure 3 (van der Sloot et al., 1989). Two sets of graphs are presented for calcium (a major constituent), molybdenum (an oxyanion) and zinc (a trace metal) in the figure: (i) leaching as a function of pH, and (ii) leaching as a function of LS. Both Level I characterization data and Level II concise data are included to illustrate consistency between the data sets. The concise leaching results at cumulative LS 2 and 10 follow the same pattern as observed in the more elaborate column leaching test. These data comparisons for granular materials reflect a wide range of practical conditions encountered in field scenarios. In addition to the leaching behavior as a function of pH or LS, the availability and total composition data are provided. The availability corresponds to the greatest release observed in the pH controlled test. All leaching and composition data are expressed in the same units
210
(mg/kg), to allow direct comparison of data between the two graphs. The impact of a treatment process or changes in pH on constituent release can be quantified from the combined graphs. Release as a function of time is related to LS using the infiltration rate. These results show that while characterisation cannot be replaced by concise testing, concise testing can be used for rapid evaluation of material behavior in the context of prior characterization for management decisions. MEASUREDTIMEDEPENDENTRELEASE
EVALUATION OF RELEASE CONTROL Iso-release lines (mg/m 2) at t=lOOyr 19
Total Potential leachability
Zn ,,..,
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TANK LEACH TEST L / A = 1.5; 8 r e n e w a l c y c l e s ; c l o s e d v e s s e l
analysis of 3 combined leachates pH CONTROLLED
TEST ( c r u s h e d )
L S = 10; pH---8 control; 24 hrs L S = 5 0 ; p H = 4 control; 24 hrs
R e c o r d p H , E H, D O C , T D S and C o n d u c t i v i t y .
SENSITIVITY OF LEACttlNG TO pH CHANGE 1ooo . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1oo _._~_. ~,
PREDICTION OF RELEASE WITH TIME 1.20 Block 10xlOxl0 cm 100 ....... 0.80
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Tile Acid Neutralization Capacity amounts to 7 mol/kg. Buffering of pH will persist for considerable time DOC not relevant. TDS of limited importance.
Figure 4. A concise testing protocol for monolithic materials (including compacted granular materials) and concise testing results for zinc in a cement stabilized municipal solid waste fly ash in comparison with results from detailed characterization. An example of a concise testing protocol for a monolithic waste, cement stabilized M S W I fly ash (van der Sloot et al, 1995), is illustrated in Figure 4. The same approach may be applied for compacted granular
211
materials which exhibit diffusion controlled release in the field. as follows:
The test method can be summarized
Part A. Tank Leaching Test Monolith size: 4x4x4 cm cube (compacted granular ca. 10 cm dia. x 4 cm min. height) Liquid to surface area 1.33 ml/cm2 (monolith -128 ml; compacted granular - 105 ml) Deionized water as leachant, leachant renewal at 1, 8 and 24 hr. Closed vessel Measurements (each filtered extract): pH, Eh, DOC, TDS and conductivity, relevant major, minor and trace elements Part B. pH Contr.olled Serial Batch Test Deionized water as extractant with specified pH control Particle size reduction with 95% to less than 300 lam Extraction 1 - LS--10 ml/g, pH=8, contact period of 24 hr, well mixed Extraction 2 - LS=50 ml/g (cumulative LS=60 ml/g), pH=4, contact period of 24 hr, well mixed Measurements for each extraction: acid or base consumption, DOC, Eh, relevant major, minor or trace elements (filtered extracts) The pH controlled serial batch test provides constituent release at a pH typical of carbonate buffering at an LS=10, which is typical of longer term liquid-to-solid ratios attained in the field from infiltration. This condition also provides an indication of availability for oxyanions, acid neutralization capacity and potential for reducing conditions. The second step extraction at pH 4 provides an estimate of availability for heavs., metals and acid neutralization capacity. The tank leaching test provides an estimate of observed diffusivity for elements of interest and effective diffusivity or tortuosity based on release of a non-interactive species (e.g., C1). Combined interpretation of availability and observed diffusivity permits evaluation of long-term release by comparison with an iso-release lines on a pD vs. availability nomograph (IAWG, 1997; Kosson et al, 1996). Measurement of TDS on all extracts allows evaluation of the potential for rapid release of very soluble salts. Results of the concise testing presented in Figure 4 indicate that good agreement was attained with the results of more extensive testing carried out on the same material.
212 j
20000 10000
10000
r,~
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Total
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Figure 3. Concise testing results for pulverized coal asia in comparison with results from detailed characterization.
Two additional sets of graphs (Figure 5) are presented on leaching of calcium, molybdenum, and zinc from the same cement stabilized MSWI fly ash (van der Sloot et al, 1995): (i) leaching as a function of pH, and (ii) leaching as a function of time. In addition to the pH or time dependent release, the availability and total composition data are provided. Both Level I characterization data and Level II concise data are included to illustrate consistency between the data sets. The data are consistent for the elements representing the behaviour of a major element (Ca), an oxyanion (Mo) and a metal (Zn). The release during the short time interval of concise testing (< 24 hours) follows the same pattern as observed in the longer duration tank leaching test. These data presentations for monolithic materials reflect a wide range of practical conditions to be encountered in field scenarios. In these graphs not all leaching and composition data can be expressed in the same units, which limits the direct comparison of data between the two graphs. However, the consequences of changes in pH with time or aider treatment on the release as
213
a function of time can be evaluated qualitatively from the combined graphs. These results also show that while characterisation cannot be replaced by concise testing, concise testing can be used for rapid evaluation of material behaviour in the context of prior characterization for management decisions.
21XX~
le+007 +
10000
+
+
+
+
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. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
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Figure 5. Concise testing results for calcium, molybdenum and zinc in a cement stabilized municipal solid waste fly ash in comparison with results from detailed characterization. The arrows in the graphs of concentration as a function of pH (left side) indicate the pH of the extracts from the tank leaching test for the monolithic materials.
ECONOMIC CONSIDERATIONS Testing costs nmst be considered for an evaluation protocol based on initially detailed characterization (Level I) followed by quality control testing (Level III) and random application of Level II testing for validation. A basis for economic comparison was developed assuming that typical current evaluation protocols for bulk wastes incorporate a single step batch extraction on a bi-weekly basis with a full elemental analysis (ca. 20 elements). The alternative approach considered here would require initially
214
Level I characterization followed by Level III on a monthly basis using a composite sample with a reduced analytical set (ca. 5 elements). The altemative approach also would include Level II testing on a random basis four times per year. While initial costs are greater for the alternative approach, long-term costs are reduced. Based on these assumptions, the payback period for the alternative approach would be approximately two years. CONCLUSIONS An alternative approach for evaluation of the environmental impact of waste utilization and disposal has been presented. The alternative approach requires initially detailed waste characterization followed by greatly reduced testing requirements to achieve quality control and integrated concise testing to screen new wastes that are similar in origin to previously characterized wastes and low volume wastes being evaluated for disposal. The advantages of this approach include: (i) flexibility of waste evaluation and management based on improved understanding of waste characteristics, (ii) estimation of contaminant release in the context of the waste management scenario and anticipated environmental conditions, (iii) quality control in the context of a detailed understanding of waste characteristics, and (iv) long-term reduction in costs associated with waste testing. ACKNOWLEDGMENTS The part of this work carried out at Rutgers University was supported in part through funding from the Hazardous Substances Management Research Center, an Advanced Technology Center of the New Jersey Commission on Science and Technology and a National Science Foundation Industry/University Cooperative Research Center. The part of the work carried out at ECN was supported through basic funding of ECN in relation to national and European standardisation. REFERENCES EEC, 1991, Council Directive 91/156/EEC article l(a) EEC, 1996 Energy and Waste Forum Consensus paper final draft - June 1996. J.M. Bemtgen DG XII. IAWG (International Ash Working Group; A.J.Chandler, T.T.Eighmy, J.Hartlen, O.Hjelmar, D.S.Kosson, S.E.Sawell, H.A.van der Sloot, J.Vehlow). 1997. Municipal Solid Waste Incinerator Residues. Studies in Environmental Science 67, Elsevier Science, Amsterdam, 974 pp. A.C. Garrabrants, T.T. Kosson, H.A. van der Sloot and D.S. Kosson. 1996. "The Determination of Systematic Leaching Behavior of a Petroleum Catalyst Waste," Waste Solidification - Stabilization Process - International Congress, November 28-December 1, 1995, Nancy, France. A.C. Garrabrants and D.S. Kosson. 1997. Use of a chelating agent to determine the metal availability from soils and wastes. WASCON 1997. Maastricht, The Netherlands. J.J.J.M. Goumans, H.A. van der Sloot and Th. G Aalbers, eds. 1991. Waste materials in construction. WASCON 1991. Elsevier, Amsterdam. J.J.J.M. Goumans, H.A. van der Sloot and Th. G Aalbers, eds. 1994. Environmental aspects of construction with waste materials. WASCON 1994. Elsevier, Amsterdam. D. Hockley and H.A. van der Sloot. 1991. Long-term processes in a stabilized waste block exposed to seawater. Environ. Sci. & Technol., 25:1408-1414. D.S.Kosson, T.T.Kosson, H.A. van der Sloot. 1993. Evaluation of Solidification/STabilization Treatment Processes for Municipal Waste Combustion Residues. USEPA, NTIS PB93-229-870/AS. D.S. Kosson, H.A. van der Sloot and T.T. Eighmy. 1996. An approach for estimation of contaminant release during utilization and disposal of municipal waste combustion residues. J. Hazard. Mat., 47:43-75. I. Massry. 1997. The Impact of Micropore Diffusion on Contaminant Transport and Biodegradation Rates in Soils and Aquifer Materials. Ph.D. Dissertation, Rutgers, The State University of New Jersey (New Brunswick, NJ). NEN 7341. 1995. Leaching characteristics of solid (earthy and stony) building and waste materials. Leaching tests. Detemaination of the aavailability of inorganic components for leaching. First edition. Netherlands Normalization Institute, Delft.
215
NEN 7343. 1995. Leaching characteristics of solid (earth and stony) building and waste materials. Leaching tests. Determination of the leaching of inorganic constituents from granular materials with the column test. First edition. Netherlands Normalization Institute. NEN 7345. 1994. Determination of leaching from monolithic contruction materials and waste materials by means of a diffusion test. Netherlands Normalization Institute, Delt't. NVN 7347 NNI (DRAFT). 1994. Determination of the maximum leachable quantity and the emission of inorganic contaminants from granular construction materials and waste materials - The compacted granular leach test. Concept Dutch pre-standard, preliminary edition. NVN 7348 (DRAFT). 1994. Determination of reducing properties and reducing capacity of waste materials and construction materials. F. Sanchez. 1996. Etude de la Lixiviation de Milieux Poreux Contenant des Especes Solubles: Application au Cas des Dechets Solidifies par Liants Hydrauliques. These, L'Institut National des Sciences Appliquees de Lyon. C.E. Schaefer, R.R. Arands, H.A. van der Sloot and D.S. Kosson. 1995. Prediction and experimental verification of liquid phase diffusion resistance in unsaturated soils. J. of Contaminant Hydrology, 20:145-166, 1995. USEPA 1990. Part 261, Appendix II - Method 1311 Toxicity Characteristic Leaching Procedure (TCLP), Federal register, Vol. 55, No. 61, March 29, 1990, Rules and Regulations, pp. 11863-11877. J.P.G.M. Schreurs, H.A. van der Sloot en Ch.F. Hendriks. 1996. Verification of laboratory-field leaching behaviour of coal fly ash and MSWI bottom ash as roadbase material. This conference. D. Mandin, H.A. van der Sloot, C. Cervais, R. Barna and J. Mehu. Long term leaching behaviour of Pb and Zn slags in relation to beneficial application. This conference. H.A. van der Sloot, O. Hjelmar and G.J. de Groot. 1989. Waste/soil interaction studies - The leaching of molybdenum from pulverized coal ash. In: Flue gas and fly ash, Eds. Sens, P.F. and Wilkinson, J.K., Commission of the European communities, Elsevier applied science, London. H.A. van der Sloot, D. Hoede and R.N.J. Comans. 1994a. The influence of reducing properties on leaching of elements from waste materials and construction materials. In J.J.J. Goumans, H.A. van der Sloot and Th.G. Aalbers (Eds.): Environmental aspects of construction with waste materials. Elsevier Science, Amsterdam, pp. 483-490. H.A. van der Sloot, D. Hoede and R.N.J Comans. 1994. The influence of reducing properties on leaching of elements from waste materials. In: WASCON 1994: Environmental aspects of construction with waste materials Eds. J.J.J.M. Goumans, H.A. van der Sloot and Th. G Aalbers, Elsevier, Amsterdam p. 483 - 490. H.A. van der Sloot, D.S. Kosson, T.T. Eighmy, R.N.J. Comans and O. Hjelmar. 1994. An approach towards international standardization: A concise scheme for testing of granular waste leachability. In: Environmental Aspects of Construction with Waste Materials. Eds. J.J.J.M. Goumans, H.A. van der Sloot, Th.G. Aalbers, Elsevier Science Publishers, Amsterdam, 453-466. H.A van der Sloot, G.J.L. van der Wegen, D. Hoede, G.J de Groot and Ph. Quevauviller. 1995. Intercomparison of leaching tests for stabilized waste. Commission of the European Communities, EUR 16133 EN.
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Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
217
Development of a Leaching Protocol for Concrete I. Hohberg l, G.J. de Groot 2, A.M.H. van der Veen 3, W. Wassing 4 1Institute for Building Materials Research (ibac), Aachen University of Technology, Germany 2Energy Research Foundation ECN-BCM, Petten, The Netherlands 3NMi - Van Swinden Laboratorium B.V., Eygelshoven, The Netherlands 4Research Institute of the Cement Industry, Dtisseldorf, Germany
Abstract Based on two pre-validation studies and a final validation study a characterisation test for the determination of the leaching behaviour of concrete has been developed. The test protocol includes a tank test and an availability test. Generally, the emissions from the concrete specimen investigated were very low and often close to or well below the detection limits. Therefore, the development of a short-time characterisation test proved to be difficult. From the leaching results it could be deduced, that the leaching from concrete is in principle diffusion controlled. The results from both the tank leaching test and the availability tests indicated that the precision in terms of repeatability of the tests is good and the reproducibility is acceptable. 1 Introduction For cement-based materials, a main aspect in respect to environmental compatibility is the leaching of heavy metals and salts/1/. Up to now, no uniform and accepted leaching test exists to evaluate the leaching of environmentally relevant compounds from concrete, either with and without the application of industrial by-products. In this paper, the development of a procedure for the basic characterisation of the leaching behaviour of inorganic species from concrete is described. The protocol thus developed aims to be a tool that enables the assessment of the environmental quality of concrete. The paper summarises the research work carried out during a research project funded by the European commission under the Measurement and Testing Programme/2/. The experimental work was carried out in three stages. The first stage was implemented as a small interlaboratory study, together with the investigation of several relevant variables in a leaching test. Although only a validating interlaboratory study was planned after this first stage, it was decided to organise a second interlaboratory study, that served as a pre-validation of the test procedure. The final stage in experimental work was the validation of the test procedure by means of an interlaboratory study.
2 Materials and sample preparation For the preparation of the concretes for the first and second interlaboratory study an ordinary Portland cement (PC1) in compliance with the requirements of European standard ENV 197
218
Table 1: Total contents of some minor elements of the Portland cement and of the fly ashes used in the investigations/2/
Composition (mg/kg) PC1 FA1 FA2
Component
I
Antimony, Arsenic, Barium, Cadmium, Cobalt, Chromium, !Copper, Nickel, Lead, Zinc, Vanadium,
has been selected. For the validation study a Portland cement with higher alkali content (PC2) was used, since it was assumed it would lead to a somewhat higher leachability. Two types of bituminous coal fly ashes were chosen as concrete additions. FA1 was from a conventional dry bottom boiler and FA2 from a wet bottom boiler. FA2 showed considerably higher concentrations of some relevant minor elements than FA1. Sand and gravel from the river Rhine were used as aggregate. The total contents of minor elements of the cement and the fly ashes are reported in table 1 (the element contents of PC2 are assumed to be in the same range as for PC1)/1/.
Sb As Ba Cd Co Cr Cu Ni Pb Zn V
2 8.9 317 0.3 7.7 86 28 32 22 311 61
I
6.7 27 803 0.65 51 118 118 111 142 334 244
21 471 1330 11 74 360 178 238 1870 2190 269
The composition of the concrete mixture for the first interlaboratory study (C1) can be taken from table 3. The water-binder ratio (w/(c+0.4f)) was 0.55. The concentration levels of most of the species in the leachates obtained from the concrete samples of the first interlaboratory study were too low to be suitable for a validation of the test procedure. Therefore, a second interlaboratory study was performed with a concrete of higher water-binder ratio (higher permeability) and addition of FA2 (with higher element concentrations). The composition of this concrete mixture is given in table 3. The water-binder ratio (w/(c+0.4f)) was 0.75. Table 3- Composition of the concretes for the two interlaboratory studies and the validation study
Table 2: Amounts of compounds added as spike to the concrete mixture
component
Substance
unit
concrete
Amount g/g mixing water
Portland cement, PC1 Portland cement, PC2 Fly ash, FA1 kg/m 3 Fly ash, FA2 water solution of heavy metals (see table 2) aggregate (grading curve A/B 16)
302 60.5 -
181
-
1799
200 1 0 0
180 -
1819
270 60
Na2HAsO4. 7H20
6.25 (=2.5 g As/g)
Cd(NO3)2.4H20
2.74 (=1.0 g Cd/g)
KzCr207
4.24 (=1.5 g Cr/g)
NH4VO3
4.69 (=2.0 g V/g)
-
-
162 1730
The composition of the concrete of the validation study (C3) is reported in table 3. This concrete resembled very closely that of the first interlaboratory study. In order to guarantee reasonable element concentrations in the leachates, the concrete was spiked with As-, Cd-, Cr-, and V-compounds. These metal compounds have been introduced into the samples by the mixing water (see table 2).
219
Table 4: Curing of the concrete cubes for the two inter-
laboratory studies and the validation study
Concrete curing in the mould fog room 20~ rel. humidity climate chamber 20~ rel. humidity age of the samples at the beginning of the tank test 3
First interlaboratory
3.1
General
C1
C2 C3 duration in days
1 6
1 6
1 6
56
42
56
69
52
75
From each concrete mixture concrete cubes (0.1 m by 0.1 m by 0.1 m) were produced. The curing of the concrete samples as well as the age at the beginning of the leaching tests is given in table 4. Some of the concrete cubes were grounded for the availability tests. The total element concentrations of the three concretes are summarised in table 9.
study
From literature it followed, that the leaching of environmentally relevant compounds from concrete samples is mainly a diffusion controlled process/1/. Therefore, a standard leaching procedure should include a diffusion test (tank test). For further characterisation of the samples and to be able to determine effective diffusion coefficients (in order to assess long-term leaching behaviour), it is necessary to determine the amount that is available for leaching (availability) /1, 2/. As a result, following leaching tests were selected for the first interlaboratory study: 9 Availability tests: -
leaching test according to Dutch NEN-ISO 7341 /3/ (first step at constant p H - 7 and second step constant at pH-4; test A),
-
leaching test similar to Dutch NEN-ISO 7341 (first step without pH-control and second step at constant pH - 4; test B),
-
leaching test according to DIN 38 414 T4 (DEV-S4)/4/, with demineralised water (test C).
9 Diffusion test (tank test): The tank test procedure developed adopts the principles of the Dutch standard NEN 7345 /5/. For application of the leaching test in the construction producing industry, it was felt that the time necessary for obtaining the results was too long. A protocol was developed that implements the following conditions (test D): - Samples: concrete cubes, 0.1 m by 0.1 m by 0.1 m - leachant: demineralised water -
L/S ( V / V ) :
- leaching periods: - temperature: - leachant refreshments: - stirring action:
5 : 1
6, 24, 54, 96, 168 hours 20~ 5 no stirring
The influence of various parameters was investigated, in addition to the intercomparison tests in order to be able to find suitable parameters for the leaching of concrete. The following
220 variations of the standard conditions were selected: flow of CO2/air (test E), test duration: 14 days (test F); L/S (V/V): 2.5, 25 (test G); stirring during the test (test H); sea water as leachant (test I). Five laboratories were involved in the first interlaboratory study. The concrete cubes from the mixture C1 (see table 1) were used for the leaching tests.
3.2
Data evaluation
The interlaboratory studies in this project have been evaluated in accordance to ISO 57252:1994. According to the NEN-ISO standard 5725 the precision is determined using two parameters, repeatability and reproducibility: • repeatability: precision under conditions where independent test results are obtained with the same method on identical test material in the same laboratory by the same operator using the same equipment within short intervals of time. • reproducibility: precision under conditions where independent test results are obtained with the same method on identical test material in different laboratories with different operators using different equipment. The precision can be expressed as the standard deviation of the test results obtained under repeatability or reproducibility conditions, indicated by 'Sr' and 'SR' respectively. It should be noted that the precision does not relate to the true value of test result, but depends only on the distribution of random errors.
3.3
Results and Discussion
3.3.1
Availability tests
The average results from the three availability tests are summarised in table 9. Some results from the two availability tests performed similar to NEN-ISO 7341 (test A and test B) are presented in Fig. 1.
10000-
No significant difference is found between results from test A and B. The leached amounts of the concrete samples yielded with the DEV-S4-procedure (test C) are generally lower than with test A and B. Most concentrations in the leachates were below the detection limits (see table 9).
amount leached in mg/kg Concrete CI
1000" 10o 10 I
0.1
Na
Cd
CI
Cr
Cu
Pb
S
Zn
Fig. 1: Comparison of two p H regimes in the availability test (test A: 1st step at pH= 7, 2 nd step at pH=4; test B: I st step without control, 2 nd step at pH=4)
The presented results show, that the availability test according to NEN 7341 is suited for an estimation of the leaching potential needed for calculation of effective diffusion coefficients. For concrete, no major influence is found when using a pH higher than 7 for one of the pH-static leaching steps. Using the DEV-S4
221
procedure, the potential is not correctly estimated (the leachable amounts would be estimated too low). 3.3.2
T a n k test
Evaluating the tank leaching test data according to NEN 7345, the results shown in table 5 were obtained. In this table the pDe-values (negative logarithm of effective diffusion coefficient) are combined for those elements, for which results indicate that the leaching is controlled by diffusion. The pDe-value is a measure of the mobility of the elements (the lower the pDe-value, the higher the mobility). Table 5: Results o f tank leaching tests f o r concrete C1 evaluated according to NEN-ISO 7345/2/
Na
labora-
As
Cr
C1 pDe
ty
pDe
DTL
-
DTL
-
DTL DTL
tory
pDe
pDe
2
n.d.
DTL
n.d.
DTL
n.d.
DTL
n.d.
DTL
n,d,
13.8 0.11 12.6 0.16 13.8 0.10 12.6 0.06 13.3 0.6 DTL 13.6 0.4 DTL -
n.d. 11.9 n.d.
cr -
n.d.
-
n.d.
Cu dr
14.0 0.33 13.4 0.26 13.8 0.21 14.5 0.19
pDe
ty
pDe
DTL
-
DTL
DTL
-
DTL
-
DTL
DTL DTL
dr
DTL
14.5 0.50 n.d. 14.6 0.16 16.1 0.24
14.5 0.08 DTL 14.2 0.12 DTL
n.d.
-
16.4
0.19
pDe.." negativelogarithmof the effectivediffusioncoefficient DTL: onlydetectionlimit values measured
n.d.: no diffusion control found I n f l u e n c e o f the f l o w o f air or C 0 2
PDe
leaching (test E)
D Cr
12
x Zn
13
CO2
[] D•
14
i~
xr-1 []
15 16
[]
AIR.
X
normal
, 5
d u r i n g tank
6
7
8
9
10
pH
11
Fig. 2: Relation between mobility (expressed as the negative logarithm o f the diffusion coefficient) and p H
The influence of the flow of air or CO2 (test F) could be translated into a pH effect. This is summarised in Fig. 2 for some metals. The mobility, expressed as pDe (negative logarithm of the diffusion coefficient) changes as function of pH corresponding with the solubility curves for the metals. Strictly speaking this effect can be interpreted by differences in availability with one c o n s t a n t pDe or one availability value with different pDe values. I n f l u e n c e o f the tank leachinv, p e r i o d (test F)
No significant influence was found from the difference in total leaching time (7 and 14 days) on the leaching rates. Each of the five leaching periods was doubled in the 14 days leaching test compared to the 7 days test. This results in a theoretical increase in
222 the concentrations by a factor of 1.4 if the leaching is diffusion controlled. This factor is generally too small for a significant improvement of the possibilities for interpretation.
Influence of liquid~solid ratio during tank leaching (test G) Larger leachant volumes (test G, L/S (V/V) - 25) have had only the effect that the fractions were extra diluted so the measured concentration levels were even lower. The interpretation of the effect of the smaller leachant volume (L/S (V/V) - 2.5) was limited because concentration levels were still near DTL.
Influence of stirring during tank leaching (test H) The diffusion model used assumes virtually zero concentration levels in the leachate. If this model is correct stirring of the leachant during the experiment should not influence the concentration levels in the leachates. In the experiments H no influence was found although interpretation was hampered by low concentration levels.
Usage Of sea water as leachant during tank leaching (test I) Due to the low concentration levels in the leachates and the presence of, relative to these levels, rather high concentrations in sea water for several elements, only general indications for differences in leaching could be identified. Leaching levels from sea water seemed to be somewhat lower for the elements Cr and Zn/2/. )De
Influence of temperature during tank leaching
----O-- Na
(test J)
CI
12 -
-
- O r
- - o - Zn
~ . I - A
13 -
14 -
16
.~"
,
,
,
,
,
,
10
20
30
40
50
60
Temperature
in ~
Fig. 3: Relation between mobility (expressed as the negative logarithm of the diffusion coefficient) and temperature.
The temperature is expected to influence the leaching behaviour (higher temperature causes higher concentration levels in the leachates) /6/. For some elements, whose concentrations in the leachates were at reasonable levels, the mobility (pDe) is plotted in Fig. 3 against the temperature. The results indicate that for temperatures higher than 20 ~ indeed the mobility increases with temperature. At lower temperature (< 20 ~ the concentration levels in the leachates become too low to give a reliable result. 3.3.3
Conclusions
Although the majority of the factors investigated has an influence on the test results, the concentration levels obtained during the tank leaching test were very low, often well below the detection limits. None of the variations on the test protocol led to a relevant increase in concentration levels measured, so it was decided to keep the tank test procedure as simple as possible. For the further evaluation of the leaching protocol for concrete the availability test according to Dutch standard NEN 7341 and a modified diffusion test similar to Dutch standard NEN 7345 were adopted.
223
4
Second interlaboratory study
4.1
Generals
Due to the fact, that the concentration levels obtained in the first interlaboratory study were too low to be suitable for a proper validation of the test protocol, an additional pre-validation study was necessary. In this second interlaboratory study specimen from concrete mixture C2 (see table 1) with higher water-binder ratio (higher permeability) and addition of a fly ash with higher concentrations of relevant elements (FA2) were investigated. The experimental conditions for the tank leaching test read as follows: - leachant: - L/S (V/V): - leaching periods: - stirring action: - temperature:
demineralised water 5:1 6, 24, 78, 168,336 hours no stirring 20~
The only difference in experimental conditions with the preliminary experiment was the time intervals. The parameters to be analysed were sodium, arsenic, calcium, chloride, chromium, copper, potassium, and sulphate. 4.2
Availability test
The results of the availability test are summarised in table 9. The total concentration and the availability of C1 and C2 are represented for some examples in Fig. 4. The results illustrate the fact that using a fly ash with higher element concentrations does not necessarily results in higher leaching levels.
200
4.3
a m o u n t in m g / k g
Diffusion test
219,/ 102
100
-
14 r-j_o_.2
!i
; ,7
Asl ~ l c r
i2.7 r..._~
~i
22
iil .7
. .
_
7
4.2
. ~.N
9
1.2
6
.
Icu I z~, [As I Pb I cr I Cu [ Zn
Although the fly ash used for the second round of experiments generally had a higher content of heavy metals and the tortuosity of the concrete produced with this fly ash was about 3 times lower/2/ than that for concrete C1, the leached amounts of heavy metals from this concrete were again very low.
.
The results of the tank test from the second interlaboratory study are sumFig. 4: Comparison of concrete C1 and C2 in respect to marised in table 6. Note the limited total element contents and availabilty number of results for which diffusion control was found according to NEN 7345 procedure. Most measured concentrations were at DTL-levels. This presents a problem for the evaluation of the leaching test itself concerning these elements. The levels for Na and K were high enough for adequate analysis, but the leaching mechanism could not be deduced successfully from the leaching data, due to S-shaped release curves/2/. c o n c r e t e with FA1 ( C 1 )
c o n c r e t e with F A 2 ( C 2 )
224
Table 6: Results o f the tank test from the second interlaboratory study Lab
Na pDe
As cr
pDe
Ca a
pDe
C1 a
DTL DTL
1
2
_
_
4
15.85
0.13
17.59
o.22
pDe
Cu a
K cr
pDe
S cr
12.13
o.28
DTL _
_
6 _
DTL
DTL DTL
_
15.39
0.21
> 13.35
DTL DTL DTL DTL DTL
15.34
0.14
> 13.32 13.92
0.29 > 13.45
10.07
0.23
11.68
0.29
cr
> 14.5
0.07 > 13.58
_
pDe
> 13.5 > 13.24
10.62
pDe
DTL DTL 10.91
0.16 > 14.71 o.15 > 14.88 o.16
_
_
5
Cr cr
DTL DTL DTL DTL
_
3
pDe
>14.55
14.26
0.56
13.58
0.16
DTL DTL
DTL DTL DTL
11.13
0.18
10.90
0.08
15.76
0.15
15.60
0.17
DTL DTL
DTL: concentration levels below detection limit pDe: effective diffusion coefficient; values with >: concentration levels close to detection limit - 9 not determined
The results indicated that a validation study with concrete specimen used in the second interlaboratory study was not useful. There seemed to be no alternative than to use 'spiked concrete' for this evaluation. The spiking of a concrete with additional amounts of several elements seemed to present a suitable option to measure concentrations distinctly above the DTL and thereby to be able to validate the leaching test procedure. 5
Validation study
5.1
Generals
The test conditions for the tank tests in the validation study were the same as during the second interlaboratory study. A spiked concrete mixture (C3, see table 1) was used as a test sample. In the validation study 18 laboratories from the following countries participated: the Netherlands, Germany, Belgium, France, United Kingdom, Sweden, Poland, Hungary, Portugal, and Austria. In CEN TC51/WG 12/TG6 a discussion had taken place whether demineralised or mineralised water should be used as a leachant for cement-based materials. The main reason for this discussion is that the use of demineralised water may lead to dissolution of the surface of concrete. On this background, in this validation study, the use of a mineralised (low hardness water) was included. 5.2
Availability
The average results of the availability test and the result of the determination of precision is shown in table 7.
225
Table 7 : Average results and precision o f the availability test
Element Availability i mg/kg Na As C1 Cr K S V
5.3
307.3 9.7 779.7 75.7 1820.0 1507.0 20.6
Nr/NR Rejected
Sr
SR
%
%
-
laboratories
7.2
28.4
3/4
-
24.2 27.6
64.2 104.9
14/15
I 1.9 6.4 3.0 14.4
23.3 16.7 18.8 29.7
11/12 14/15 14/15 13/14 11/12
1 1 1 2 3
The very limited precision for As and C1 can be explained by a higher analytical error caused by relatively low concentration levels relative to DTL. When 10 times DTL as the minimum concentration level is used, the average level of precision expressed as Sr and SR is 8.6% and 23.4%, respectively. The average value of SR is relatively high.
Diffusion test
The precision results are listed in table 8 for two types of leachants, demineralised (standard) and mineralised water. The precision is lower at concentration levels below 10 times DTL. D e m i n e r a l i s e d water
The evaluation of the leaching data is done by determining E64d (emission after 64 days, assuming diffusion as dominating mechanism) from the experiments with demineralised water according to NEN 7345. Using only values larger than 10 times DTL, the average values of Sr and SR are 14% and 28% respectively. These values are about the same as for the precision values found for the measured release with demineralised water in 14 days (E14~-measured), 12% and 22% respectively. In the E14a-measured values, data for elements are included for which diffusion control according to NEN 7345 could not be identified. The rather low values for Nr and NR in table 8 (maximal value is 16 which is the number of laboratories that performed the test) were therefore not caused by rejected values in the outlier procedure but by the result of the evaluation procedure in the leaching test protocol that no diffusion control could be established. Consequently no diffusion coefficient could be determined for these elements and therefore the test result E64, which is calculated from this diffusion coefficient, could not be obtained. M i n e r a l i s e d water as a l e a c h a n t
In order to investigate the influence of the usage of the so-called low hardness' water on the leaching behaviour of concrete, water with 0.25mM CaClz.2H20 + 0.5mM NaHCO3 was also used as a leachant. The initial concentrations in the leachate for Na, Ca and C1 are a factor 4 to 10 higher than the amount leached from the test piece. These elements can therefore not be used for the comparison. For all elements measured (see table 8) except for Na and C1, no significant difference in leaching was found.
226
Table 8: Precision o f diffusion test for demineralised and mineralised water as leachant Leachant
Demineralised water
Mineralised water Demineralised water
Element
E64 acc.to NEN 7345 Na As CI Cr K E64 acc.to NEN 7345 K As El4 measured Na As Cd C1 Cr K
Mineralised water
V S E14 measured Na I) As Cd CI~) Cr
Value mg/m 2
Sr
SR %
Nr/NR
%
rejected laboratories
17.8 11.3 11.8 14.1
26.8 25 50.4 32.0
6/6 5/10 3/8 14/15
14.1 10.2
16 18.8
13/13 3/5
9.1 9.7 15.8 11.3 15.3 11.9 7.6 17.1
24.8 26.8 41 56 20.7 21.7 19.6 17.4
8/9 13/13 2/2 10/10 16/16 16/16 15/15 15/15
1
3.6 13.9 4.1 12.8 6.5 9.6 26.8
34.8 24.5 57.9 16.7 14.3 15.7 29.9
9/9 9/9 1 8/8 12/13 12/12 13/13
1 -
release 64 d
3139 13 455 339 31505 release 64 d
34795 11 r e l e a s e 14 d
1489 8.5 0.05 211 113 13442 78 1012 r e l e a s e 14 d
4944 7.6 0.02 5366 111
K
14068
V S
75 1030
-
1
~) present in blank
Effect o f spiking T o create larger concentration levels in the leachates, w h i c h were needed to validate the l e a c h i n g p r o c e d u r e , the concrete was spiked with As, Cd, Cr and V. The use o f a spike was o n l y successful to a certain extent. T h e concentration levels o b t a i n e d in the leachates were increased, but the e m i s s i o n b e h a v i o u r o f the spiked e l e m e n t s differed from e l e m e n t s that were p r e s e n t in the constituents o f the concrete. As a result, diffusion control could be clearly dem o n s t r a t e d for the matrix using m a j o r e l e m e n t s but only for a limited n u m b e r o f m i n o r elem e n t s due to l o w leaching levels and the spiking a r t e f a c t / 2 / . 6 Conclusions T h e aim o f this research project was the d e v e l o p m e n t o f a leaching standard for the determination o f the e n v i r o n m e n t a l quality o f concrete. The test p r o c e d u r e should be a short time c h a r a c t e r i s a t i o n test. G e n e r a l l y , the e m i s s i o n s f r o m the investigated concrete s p e c i m e n were v e r y l o w and often close to or well b e l o w the detection limits. F r o m the results o f the three i n t e r l a b o r a t o r y studies
227
it could be deduced, that following major elements are suitable to identify diffusion control during leaching from concrete: K, As, Cr and Cl. If these elements show diffusion control, it can be concluded, that the leaching from the matrix is in principle diffusion controlled. The results from both the tank leaching test and the availability tests indicated that the precision in terms of repeatability and reproducibility of the tests is good. No significant differences were observed between the tank leaching test with demineralised water and mineralised water. The test protocol with demineralised water has following advantages: demineralised water is easily to harmonise and the problem that species that are used to 'mineralise' the water cannot be included in the evaluation of the tank leaching test does not occur. The development of a short time characterisation test procedure for standardised concrete with usual constituents was difficult. The results show that the two types of concrete that were tested in this project do not have significant leaching of environmentally relevant elements. The concentrations of released environmentally relevant elements were mostly below the DTL. This is consistent with results in literature regarding the leaching behaviour of concrete with usual ingredients/1, 6/. This could indicate that further testing with a characterisation test is unnecessary for concrete with usual ingredients with known leaching behaviour. However the leaching procedure described is useful for the determination of the leaching behaviour of concrete with unknown ingredients. Table 9: Element contents and availability results of the concretes investigated/2/
c1
C2
C31)
FA1
FA2
FA1
concrete ......
addition method
total content
I
availI availability; A ability; B
availability; C
parameter
total content
availability; A
total content
I
availability; A
mg/kg ,,,
Nlmean
Nlmean
N [ mean N
As
5.2
7
0.16
10
<0.1
8
<0.01
9
22
1
< 0.9
12
95
1
12.3
Cadmium, Cd
0.2
7
0.48
9
0.2
8
< 0.01
9
1.3
1
0.4
1
63
1
42
6
Chloride,
104
5
95
7
142
5
< 35
9
108
1
73
8
n.a.
1
985
26
12
124
1
82
32
7
1
n.a.
mean Nlmean Arsenic,
C1
Nlmean
N I mean Nlmean
C h r o m i u m , Cr
48
7
2.7
10
4.1
8
0.12
8
46
1
1.2
Cobalt,
Co
4
7
1.2
8
1.7
8
< 0.03
7
3
1
n.a.
30
Copper,
Cu
11
7
3.7
9
3.6
8
<0.03
9
11
1
3.6
12
12
1
n.a.
Lead,
Pb
14
7
0.7
9
1.3
8
<0.04
9
87
1
4.2
1
24
1
n.a.
Mercury,
Hg
n.a.
-
< 0.1
8
< 0.1
8
< 0.01
9
0.03
1
n.a.
-
n.a.
1
n.a.
Nickel,
Ni
121
3
12
2
n.a.
n.a.
-
17
1
n.a.
-
26
1
n.a.
Potassium, K
3046
2
835
4
890
2
666
2
4080
1
1047
12
n.a.
1
2085
Sodium,
Na
398
7
154
10
161
8
76
9
760
1
199
12
n.a.
1
312
8
Sulphur,
S
1217
7
1206
9
1115
7
308
9
1112
1
1141
10
n.a.
1
1543
10
Thallium,
T1
32
< 1.2
6
0.1
8
0.1
8
< 0.01
9
0.4
1
0.11
1
n.a.
1
n.a.
Vanadium, V
18
4
n.a.
2
n.a.
-
n.a.
-
16
1
n.a.
-
114
1
29.9
30
Zinc,
81
7
55
9
52
8
< 0.8
9
219
1
102
1
80
1
n.a.
-
Zn
~): this concrete was spiked with arsenic, cadmium, chromium and vanadium (see tables 3, 2) A: procedure according to NEN-ISO 7341" 1st step 3 h constant pH -- 7, 2 nd step 3 h constant pH --- 4 B: procedure similar to NEN-ISO 7341" 1st step 3 h without pH-control, 2 nd step 3 h constant pH : 4 C: procedure according to DIN 38414 Teil 4 (DEV-S4-procedure) n. a.: not analysed
228 Characterisation of the leaching mechanism of elements that leach in sufficiently high concentrations can be hampered by initial effects like surface wash-off and an S-shaped curve as occurred for most spiked elements. In order to be able to calculate the long-term leaching behaviour of concrete for these elements accurately the leaching time intervals of the tested leaching procedure have to be prolonged. But this would turn the 'short time' characterisation test that was aimed at in this project back into the standardised Dutch leaching test procedure NEN 7345. 7
References
1. Hohberg, I.; Miiller, Ch.; SchieB1, P.; Volland, G.: Sachstandsbericht ,,Umweltvertr/iglichkeit zementgebundener Baustoffe. In: Schriftenreihe des Deutschen Ausschusses for Stahlbeton (1996), Nr. 458 2. Groot, G.J. de; Hohberg, I.; Lamers, F.J.M., Veen, A.M.H. van der; Wassing, W.: Development of a leaching protocol for the determination of the environmental quality of concrete. MAT1 CT930026, Final report. March 1997. 3. NEN 7341 02.95: Leaching characteristics of solid earthy and stony building and waste materials. Leaching tests. Determination of the leaching availability of inorganic compounds 4. DIN 38 414 Teil 4: Deutsche Einheitsverfahren zur Wasser-, Abwasser- und Schlammuntersuchung; Schlamm und Sedimente (Gruppe S) 5. NEN 7345 03.95: Leaching characteristics of solid earthy and stony building and waste materials. Leaching tests. Determination of the leaching behaviour of inorganic components from building monolithic waste materials with the diffusion test 6. Groot, G.J de; Sloot, H.A. van der; Bonouvrie P.; Wijkstra J.: Karakterisering van het uitlooggedrag van intakte produkten. Mammoet deelrapport 09. March 1990
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
229
USE OF A C H E L A T I N G A G E N T TO D E T E R M I N E THE M E T A L A V A I L A B I L I T Y FOR L E A C H I N G F R O M SOILS A N D W A S T E S Andrew C. Garrabrants and David S. Kosson
Department of Chemical and Biochemical Engineering Rutgers, the State University of New Jersey Piscataway, NJ, USA 08855-0909
ABSTRACT Use of the available content of elements, or availability, as the thermodynamic driving force for release can provide a more realistic assessment of leaching potential than total elemental content. Availability is often measured by an extraction procedure using a high liquid-solid (LS) ratio and pH control at levels which maximize solubility within the limits of a reasonable release scenario (e.g., pH 4.0 for cations and pH 8.0 for oxyanions).
An alternative extraction procedure using ethylene diamine tetraacetic acid
(EDTA) as a chelating agent was evaluated to eliminate the need for rigorous pH control and provide a single extraction step method. The availability of As, Cd, Cu, Mn and Pb in four materials, representing both treated and untreated soils and wastes, were tested using the EDTA procedure. The approach of a Dutch availability test, NEN 7341, and total elemental content were used as a basis for comparison of results.
INTRODUCTION Accurate determination of inorganic contaminant leachability from soils and waste is essential for risk assessment and hazardous waste management decision-making. An often misleading estimate of inorganic leachability from wastes and soils is to measure the total constituent concentration in the solid phase and assume that mass release continues until depletion. A protocol that accurately determines the mobile, or available, fraction of a constituent in a solid matrix is important for assessment and release models.
The resulting measure of "availability" can be viewed as (i) the
potentially mobile ~content of an element or (ii) the element content which serves as the thermodynamic source or driving force for release.
In the latter context, availability
has been used in conjunction with solubility data and observed diffusivities from monolith leach tests to estimate contaminant release for a variety of environmental scenarios
(1-3).
230 In past experimentation, these fixed and mobile fractions have been referred to as "detrial" and "non-detrial" fractions, respectively, and have been quantified by a number of sequential extraction techniques with increasingly aggressive extractants (46). The non-detrial, or available, fraction has been observed to include water soluble, exchangeable, and specifically or physically absorbed constituents (7). The available fraction of the total constituent concentration, therefore, refers to that amount that is leachable from the solid phase, excluding that which is immobilized by incorporation into glassy, amorphous, or geologically stable mineral forms (8).
Theoretically,
measurement of the available fraction should be carried out under conditions that permit complete solubilization of the elements of concern in mineral phases which could solubilize under credible extreme environmental scenarios.
For example, a standardized
availability test, NEN 7341 (9), uses pH 4.0 and 7.0 as conditions for solubilization of potential toxic cations and oxyanions, respectively. A large fraction of solubilization is anticipated by use of reduced particle size (e.g., < 300 llm) and high liquid-solid ratio (e.g., 50 ml/g) for each extraction step. This approach has been criticized because the specified time intervals can be insufficient to approach complete solubilization or equilibrium and, despite high LS ratios, mass release of some low solubility constituents may be limited by saturation of the aqueous phase (1, 10). Suggested alternative approaches to measurement of the available fraction have included smaller particle sizes, longer extraction intervals, and long-term leaching to depletion (1, 10, 11). Development of a readily implemented test to measure availability requires that (i) fractional solubilization be maximized over the range of environmentally significant pH values, (ii) extractions be completed in a short time frame (e.g., 48 hours) and (iii) sample preparation, manipulations and specialized equipment needs should be minimized. The approach explored in this paper is the use of a chelating agent at neutral pH in a single step extraction to achieve the above objectives. By using a chelating agent during extraction to relax metal solubility limitations in the extraction fluid, it is hypothesized that a rigorous estimate of the maximum constituent fraction available for release can be measured.
O B J E C T I V E S AND E X P E R I M E N T A L DESIGN
The specific objectives of the work presented here were to (i) optimize a chelation extraction procedure to determine the availability of inorganic constituents from wastes and soils using a single waste, (ii) evaluate the application of the optimized test to several waste types, and (iii) compare the results of the chelation procedure to
231 the total elemental content and the availability measured using the approach described by NEN 7341 (9). The experimental factors evaluated during test method optimization were (i) liquid-solid ratio at two levels (50 and 100 ml/g), (ii) chelating agent concentration at three levels (50, 100 and 150 mM) and (iii) contact time at three levels (18, 24 and 48 hours). An extraction pH of 7.0 + 0.5 was selected to maximize solubility of anions simultaneously with increasing solubility of cations through chelation. The availability of arsenic, lead, cadmium, copper and manganese was determined in this study. Arsenic and manganese represent those elements that are capable of speciation as oxy-anions. Lead, arsenic and cadmium are common toxic cations in many soils and waste. Manganese and copper are important plant nutrients that may be toxic to plant life at high soil concentrations. The release of a constituent with time is an asymptotic process, and for a given LS ratio the system can be considered at equilibrium when no significant time effect on the liquid phase concentration is observed. The required contact time to reach equilibrium is dependent on the particle radius due to mass transfer rate limitations in larger particles.
Since short testing times are desirable, the diameter of the solid material
particles must be small enough to minimize mass transfer rate constraints.
However,
particle size reduction usually requires a grinding or milling process that can be cumbersome or impractical.
NEN 7341 suggests that three to four hour contact times
are adequate to achieve equilibrium for particles of <300 ~m maximum diameter (9). F&llman et al. (10) have shown that some low solubility metals may be mass transfer constrained within this time interval and recommended that a minimum 24 hour contact time with size reduction to < 125 ~m diameter particles.
For this study, a maximum
particle size of 300 ~m was used based on preliminary testing and system modeling which indicated that equilibrium conditions could be achieved in less than 48 hours (10). Ethylene diamine tetraacetic acid (EDTA) was chosen as the chelating agent because of a high affinity for a wide range of cationic metals (12). EDTA is a tetraprotic acid (H,Y) having four dissociation constants [pKa of 2.00, 2.67, 6.16 and 10.26, yielding H3Y-, H2Y2, HY 3- and Y"- respectively (13)]. At pH 7.0 + 0.5, the primary form of EDTA will be the singly protonated acid (HYa). Below a pH of 6.16, the doubly protonated form of EDTA exists with stability constants that differ from the above form by several orders of magnitude. Table 1 presents the metal-complex stability constants for many common metals with the protonated forms of EDTA prevalent at near-neutral pH along with solubility constants for each metal hydroxide form.
The large stability
232 constants illustrate that the presence of EDTA can significantly increase metal solubility, often by many orders of magnitude, by formation of soluble metal-chelates. The extraction potential of EDTA concentrations at or less than 100 mM have been studied in releation to remediation of contaminated soils (7, 13-16).
For this
optimization, EDTA concentrations higher than those previously reported in literature also were investigated Initial optimization of the test procedures was carried out using municipal solid waste incinerator (MSWl) combined air pollution control residue and bottom ash (17). This waste type offers measurable quantities of many metal contaminants, and therefore the effectiveness of the test parameters can be evaluated for a number of constituent behaviors.
For the comparison portion of this study, four wastes were studied that are
representative of common waste types with differing physical and chemical properties. In addition to the MSWl combined ash, two different field contaminated soils and a contaminated soil treated by solidification/stabilization (S/S) were tested.
The total
concentration of principle contaminants for each waste is summarized in Table 2.
EXPERIMENTAL
METHODS
Prior to sample extraction for measurement of availability, each sample was titrated with sodium hydroxide solution in the presence of the desired EDTA concentration to determine the quantity of sodium hydroxide required to achieve a final pH of 7.0_+0.5. Aliquots of base were added to 2 g samples of particle size reduced material and agitated for 24 hours in a reciprocating shaker.
Separate titrations at I_S
ratios of 50 ml/g dry and 100 ml/kg dry were conducted to mimic test conditions.
For
the optimization study, eighteen extraction conditions were conducted in duplicate. A 4 g sample of MSWl combined ash was put into each of thirty-six 500 ml HDPE extraction vessels. One of three EDTA concentration leachants was added to each extraction vessel at the proper volume for each I_S ratio (i.e., 50 and 100 ml/kg dry).
The required
equivalents of sodium hydroxide was added to achieve a final solution pH of 7.0_+0.5. When all of the extractions were setup, the 36 bottles were arranged in a random order on a reciprocal shaker for the required contact time applicable for each extraction (i.e., 18, 24, or 48 hours). After the appropriate contact time, six extractions representing the two desired LS ratios at three EDTA concentrations were removed from the tumbler.
Each extraction
was allowed to settle for 10 minutes to segregate solid and liquid phases. Approximately 5 ml of the supernatant was removed from each extract so that the pH of the leachate
233 could be measured. The remainder of each the leachate was filtered by vacuum through a 0.45 l~m pore size polypropylene filtration membrane held in a polycarbonate filtration apparatus. The samples were not acid-preserved prior to analysis to prevent precipitation of EDTA salts at very low pH. All aqueous solutions were prepared with de-ionized water. Aqueous solutions of EDTA were mixed at the three concentrations (e.g., 50, 100, 150 mM) using ACS certified disodium, dihydrate salt of EDTA (Sigma Chemical Co., St. Louis, MO). 10 N sodium hydroxide solution (Fisher Scientific, Bridgewater, NJ) was used to adjust for solution pH as necessary. All extractions were carried out in 500 ml high density polyethylene (HDPE) bottles.
Extraction slurries were filtered through polycarbonate
filter membrane holders and 0.45 ~m pore size polypropylene filtration membranes (Fisher Scientific, Bridgewater, NJ). A sequential extraction approach similar to NEN 7341 was used as a comparison to the EDTA procedure. A 10 g test sample was sequentially extracted in DI water at a LS ratio of 50 ml/g under static pH conditions.
For the two extractions, the solution pH was
maintained at pH values of 7.0 and 4.0, respectively, with a pH controller that delivered 0.5 N nitric acid into the reaction vessel.
After the first extraction contact time of four
hours at a pH of 7.0, the sample was filtered through a 0.45 l~m pore size polypropylene membrane and an analytical sample preserved. The solid was returned to the reaction vessel and extracted with an additional 50 ml/g of DI water at a pH of 4.0 for three hours. At the end of the second interval, the extract was filtered and another analytical sample was preserved for chemical analysis. The samples were analyzed separately and the released constituent masses for each extract were combined to determine the availability.
For materials without sufficient acid neutralization capacity, a parallel
batch procedure was used to yield one extract at each solution pH. Since these were parallel extractions, the released masses were not combined to determine constituent availability. All of the extraction samples were analyzed within one week after generation. A Varian Model 640 flame atomic absorption spectrophotometer was used to analyzed the preserved samples for As, Cd, Cu, Pb, and Mn. All analytical samples were analyzed using the same bulk standard and performance check solution.
DATA REDUCTION AND STATISTICAL DESIGN
Availability was calculated from the analytical data by multiplying the leachate concentration by the appropriate LS ratio,
234 AVL~,x = (C~,x)(LS~)
Equation 1
where: 9 AVL~,x is the availability of constituent "x" for treatment "i" [mg/kg dry], 9 C~,x is the concentration of constituent "x" for treatment "i" [mg/I], and 9 L S~ is the liquid-solid ratio for treatment "i" [ml/g dry].
For the statistical design of the procedure optimization, the primary experimental factors were (i) the concentration of leachant EDTA at three levels (50, 100, and 150 mM), (ii) the LS ratio at two levels (50 and 100 ml/g dry) and (iii) the contact time for each extraction at three levels (18, 24, 48 hours). Figure l a shows a schematic representation of the experimental design with shaded blocks indicating the treatment combinations that were conducted. The availability values of Cd, Cu, Pb and Mn were optimized examining leverage and prediction plots using the statistical software JMP
(18). The MSWl combined ash samples were particle size reduced and homogenized,
then stored in air-tight containers for over five years (17).
Since the consistency of
homogenization, handling and storage of the replicates could not be verified over this time interval, a Split-Split Plot Blocking Design using replication as a blocking factor was used to minimize significant blocking effects. The whole plot factor for the design was EDTA concentration, the split plot factor was the LS ratio, and contact time was the split-split plot factor. Thirty-six extractions were needed to complete the full factorial. For the comparison study, the applicability of the optimized treatment combination for a number of wastes types was evaluated and compared to values obtained using the NEN 7341 approach. The applicability of the optimized EDTA procedure was evaluated by comparison to additional extractions performed in parallel by increasing each experimental factor separately. Also, a combined increase of all factors was performed to show if there was significant change in availability due to interactions amongst the factors.
Figure l b shows a schematic representation of this experimental
design. To validate that a maximum mass release was extracted, the solid material recovered from the optimized extractions was extracted a second time at the same conditions and the additional mass release was determined. In order to compare the EDTA Availability values derived from each of the five extractions shown in Figure lb, a one-
235 way ANOVA was conducted using JMP (18) The hypothesis tested was the equality of replicate mean values between extraction conditions within some experimental error.
RESULTS AND DISCUSSION
The objective of the optimization was to statistically determine the optimal combination of experimental factors (i.e., EDTA concentration, LS ratio, and contact time) to maximize the measured availability of Pb, Cu, Cd, and Mn in MSWI combined ash. Once the availability was evaluated for each treatment and response variable, the data residuals were analyzed to determine the significant factors for the availability response of each constituent. There was no significant effect of EDTA concentrations greater than 50 mM on the availability of Pb, Cu, Cd, and Mn.
This is consistent with previous work (13, 15)
where EDTA concentrations greater than approximately 10 mM did not result in an increased extraction of Pb, Cu, Mn, and Zn from contaminated soils. The availability data from the optimization study showed an obvious replicate effect for Pb, Cd, and Mn while the availability of Cu did not display a trend due to replication.
Since the replication
effect was strong enough to outweigh variance due to the experimental factors, the significant factors had to be identified for each constituent response within each replicate.
Figure 2 shows two leverage plots for LS ratio on Cu availability that
illustrate the difference between a significant factor and a factor that does not have a significant effect on the sample mean. Each leverage plot illustrates how the standard least square model changes when a factor is removed from the model fit. The tested factor is significant if the 95% confidence curves (shown as dashed lines) intersect the sample mean shown on the leverage plot. The main effects (i.e., EDTA concentration, LS ratio and contact time) were found to influence only the availability of copper in one replicate of the MSWI ash. All other constituents showed no significant main effect or interactions other than replication. Although the main effects were not found to be significant, an optimized combination of experimental factors was found by maximizing the predicted availability responses for all four constituents.
Figure 3 presents the prediction profiles from the
screening model showing two sets of trends in availability (response variable) as a function of the three experimental factors.
In each trend set, the response value
extremes for each constituent are shown on the vertical axis while high and low factor values (e.g., EDTA concentration, LS ratio and contact time) are shown on the horizontal axis. The current value for the predicted availability response and factors are shown
236 respectively as gray horizontal and vertical lines.
The lines and markers within the plot
show how the trend in the predicted response value changes when the current factor values are changed.
Error bars represent the 95% confidence interval around the
predicted values. Figure 3 also shows the interaction between LS ratio and EDTA concentration.
As the LS ratio is increased from 50 ml/g to 100 ml/g at 50 mM EDTA
(i.e., shifting from left to right in the figure) the trend in Cd availability changes from nearly flat to a negative trend as a function of EDTA concentration.
From these prediction
plots, it is clear that availability increases for all constituents when both LS ratio and contact time are maximized. Since the only observed effect of increasing the concentration of EDTA within the tested range was on the copper availability for only one replicate, the EDTA concentration was minimized to 50 mM in order to address practical considerations. Excess EDTA in analytical solutions hindered chemical analysis and required additional analytical expense. Thus, an overall optimized combination of treatment factors was determined to consist of EDTA concentration of 50 mM, LS ratio of 100 ml/g, and contact time of 48 hrs. For the comparison study, the objectives were (i) to compared EDTA extraction availability to that determined by NEN 7341 and (ii) to check that the combination of experimental factors optimized for the MSWl combined ash also was applicable for other waste types. There was no statistical difference in extraction means as a result of increasing the experimental factors. significant.
Only a replicate effect for Soil 2 was observed to be
Therefore, it can be concluded that the variance in availability brought
about by sampling, handling, analytical and other experimental errors outweigh the variance as a result of the experimental factors. In general, the mean EDTA extraction availability values were higher for all material than NEN 7341 availability values.
This implies that either the liquid phase of
NEN 7341 extractions was solubility-constrained or dilute EDTA solution was able to dissolve a large fraction of the solid phase mineral. Chelators are much less aggressive towards solid phase mineralization than strong acids or caustics (19).
For example, the
dissolution of Fe oxide phase with EDTA is kinetically controlled and can take up to three months (14).
Thus, it was concluded that NEN 7341 was solubility-limited for all
constituents in the tested materials. Figures 4a-4e compare the mass releases measured by NEN 7341 and the EDTA extraction to the total concentration for each constituent and all materials.
The fraction
of the total constituent that was removed with each technique is presented in Table 3.
237 Arsenic:
Figure 4a shows that only 13% and 8% of the total arsenic for the untreated
and S/S treated Soil 2, respectively, was removed under NEN 7341 conditions.
In the
EDTA extractions, 100% of the arsenic in both materials was found to be available. There was no apparent decrease in the fraction of As that was available as a result of the treatment process of Soil 2 and only the dilution effect on the total concentration by the addition of Portland cement was observed. Copper: The releases of copper from the tested materials were all significantly less than the total Cu concentration regardless of the extraction technique (Figure 4b). This shows that some of the copper may be incorporated in detrial mineral phases.
In general
EDTA was able to remove approximately 40-60% of the total copper whereas only 20% was found to be available by the NEN 7341 approach. Lead: Figure 4c shows that 100% of the total concentration 1 was available by the EDTA extraction for all but the S/S treated soil. The fixation effect due to treatment of this material reduced the available fraction of Pb to 64% over the untreated sample.
When
the NEN 7341 approach was used to determine available Pb in the treated and untreated Soil 2, the lead concentration in the extract was less than the analytical detection limit. NEN 7341 data would indicate that all of the lead present in Soil 2 is unavailable for release. The discrepancy between NEN 7341 and EDTA availability values may be explained by the speciation of Pb in the soil. If the speciation of lead in the solid is such that lead solubility at pH 4 is negligible, there would be no detected lead availability following the NEN 7341 approach. Manganese: The increase in Mn availability determined by EDTA extraction over NEN 7341 is shown in Figure 4d. The chelation extraction was able to remove approximately 50% of the total manganese concentration from each sample while NEN 7341 availability values were between 15 and 25%. The available fractions of Mn that were determined show that a dignificant fraction of Mn is speciated in detrial phases in the tested materials. Cadmium: The total cadmium concentrations for both soils were below the detection limit, therefore the total content value was reported as the detection limit.
EDTA
availability values for Cd in these samples, as well as the S/S Treated Soil, were greater than 100% of the reported total concentration (Figure 4e).
In the MSWl combined ash,
Method 3050 dissolution ()and atomic absorption spectrophotometry were used to measure the total cadmium concentration.
1 as measured by x-ray fuorescence (20)
The ability of NEN 7341 to predict Cd
238 availability was variable amongst the tested materials with the best parity to EDTA availability in the MSWl combined ash and S/S Treated Soil 2 (e.g., 100% of the reported total Cd by EDTA versus 65-70% by NEN 7341) and the least parity for the untreated Soil 2 (e.g., 100% of the reported total Cd by EDTA versus only 13% of the total Cd detected). The fixation effect of S/S treatment on Soil 2 is shown by the EDTA availability values that decrease with treatment by more than the dilution factor.
The
NEN 7341 approach availability value shows that Cd becomes more available when the soil is treated. As a protocol, a single batch extraction (such as the EDTA extraction) under conditions that can remove the available fraction of many material constituents is more desirable than a two-step sequential extraction. The EDTA extraction technique yields a more rigorous availability value than NEN 7341 for the constituents from each of the four tested wastes. Examination of constituent mass release stemming from the second challenge of the solid phase from the optimized EDTA extraction test conditions show that the available fraction was quantified in one extraction.
More than 94% of the available
As, Cd and Cu was removed in the first extraction as well as greater than 88% of the available Pb and Mn. The mass of each constituent that was removed in the subsequent optimal extraction were found to be on the order of the confidence interval of the mean availability values.
CONCLUSION
Since a fraction of inorganic constituents in the solid phase (e.g., Mn, Cu, etc.) may be immobilized in amorphous or geologically stable mineral forms, a rigorous determination of the available fraction is important. One approach to measure the available constituent fraction involves static pH extractions in a two-step sequence to solubilize both cations and anions. One such standardized method, NEN 7341, has been criticized for physical limitations that can result in a saturated liquid phase.
In this
study, an alternative protocol using chelating agents to relax the solubility limitation for some low solubility constituents was developed.
Ethylene diamine tetraacetic acid
(EDTA) is a common chelating agent that forms water soluble complexes, or chelates, with many metals resulting in an increased release from the solid phase. Optimization of the experimental factors (e.g., EDTA concentration, liquid to solid, or LS, ratio, and the extraction contact time) showed that no significant difference in the availability of each lead, copper, cadmium and manganese from a mixture of municipal solid waste incinerator bottom ash and APC residue.
Examination of leverage
239 and prediction plots showed that the overall maximum availability of the four constituents, however, could be measured at long time intervals and high LS ratios. The concentration of EDTA did not have a significant effect on the maximizing the availability values. An optimized EDTA extraction procedure was developed using 50 mM EDTA at a LS ratio of 100 ml/g and a contact time of 48 hours. The applicability EDTA technique as a protocol was tested on four materials representing common waste types. These wastes included contaminated soils and S/S treated wastes. For all of the tested wastes, the availability of As, Cd, Cu, Pb and Mn as determined by EDTA extraction were found to be significantly higher than the constituent availability following the NEN 7341 approach. The constituents that showed the greatest difference between the extraction techniques were As, Cd and Pb. These constituents were completely available as measured by extraction with EDTA in all the tested materials, whereas the NEN 7341 approach showed that a large fraction of each constituent was unavailable for release.
In fact, the available fraction of lead in one soil
for both untreated and solidification/stabilization forms was found to be undetectable by the NEN 7341 approach, illustrating that solid phase speciation can influence the results of pH dominated tests. Manganese and copper were found to be considerably less available in the tested materials than As, Pb and Cd by both techniques. By the increasing each of optimization parameter (e.g., EDTA concentration, LS ratio and contact time) the constituent availability values were found to be insensitive to deviations from the optimized EDTA procedure. A second sequential extraction at the optimized conditions proved that approcimately 90% of the available fraction was removed in a single batch extraction for most constituents. Thus, the available fraction of the tested constituents was measured in a single extraction at near-neutral pH.
240 Table 1. Metal-complex stability constant with EDTA for common metals (13) compared to solubility constants for metal hydroxides (21 I. Metal (Cation) log KMFDTA Cadmium (Cd 2+) 1 6.5 Chromium (Cr 2+) 23.0 Copper (Cu 2+) 1 8.8 Iron (Fe 2§ 14.3 Iron (Fe 3§ 25.1 Lead (Pb 2+) 1 8.0 Manganese (Mn 2+) 1 4.0 Nickel (Ni 2+) 18.6 Zinc {Zn2* / 16.5 na - information was not available
log KM.~DT^ 2.9 2.3 3.0 2.8 1.4 2.8 3.1 3.2 3.0
log KMOH - 1 4.3 na - 1 3.8 -16.3 -38.6 - 1 9.8 - 1 2.7 -1 5 . 3 -16.2
Table 2. Mass concentrations measured by total constituent analysis, EDTA extraction and NEN 7341 approach. As x+95%c.i,
Cd x+95%c.i,
Availability [mg/cg] Cu Pb x+95%c.i, x+95%c.i,
MSWI Ash 1,220 m 2,140 m Total Conc. nt 31 1,480 1,460 EDTA nt 28 465 NEN 7341 <25 ~ 20 403 Field S0il 1 <5 d 690x+3 54+1 Total Conc. 5.8+0.5 <10 d EDTA nt 690+20 23+3 <5 d NEN 7341 nt 274 <12.5 Field $0i! 2 20,000+1500 <104 d 1,520x+5 Total Conc. 14,300+10 24,000+4000 EDTA 180+8 8,790+300 1,800+100 NEN 7341 2,560 13 1,550 <25 ~ S/S Soil 2 <60 c 14,250 c 1,100 c 10,230 c Total Conc. 14,000+900 EDTA 83+8 3,780+40 700+40 41 NEN 7341 1 ,100 2,140 <25 ~ All total concentrations by NAA (20) except where noted: d - Value below detection limit, availability value at detection limit reported nt - not tested m - Method 3050 (22) c- calculated from Soil 2 and OPC NAA data x - X ray fluorescence method
Mn x+95%c.i. 2,130 1,180 375 930+30 130+10 23 357+15 170+10 49 480 c 250+20 110
241
Table 3. Available constituent extracted by EDTA method and NEN 7341 compared to total concentration Avail./Total MSWl Comb. Field Soil 1 Field Soil 2 S/S Treated [% ] Ash Soil 2 Arsenic EDTA Avail. nt nt 100 99 NEN 7341 nt nt 13 8 Cadmium 90 na 1 O0 t 1 O0 t EDTA Avail. 65 na 13 t 70 t NEN 7341 CODDer . . EDTA Avail. 69 43 61 37 NEN 7341 19 23 a 11 21 Lead EDTA Avail. 100 100 100 64 NEN 7341 38 40 2a 2a Manganese EDTA Avail. 55 14 48 52 NEN 7341 18 2 14 23 nt - not tested na - not applicable, all values below detection limits a - availability value below detection limit, availability value of detection limit reported t - total concentration value below detection limit, detection limit reported
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242
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Figure 2. JMP (18) leverage plots for LS ratio on Cu availability from MSWl combined ash replicate B and C, respectively. Significant factors are represented by 95% confidence intervals intersecting sample mean values and Prob>F of less than 0.05.
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Figure 3. JMP (18) screening model profiles for the availability values of Pb, Cu, Cd, and Mn as a function of experimental factors LS ratio, EDTA concentration, and contact time.
243
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Figure 4. R e l e a s e of (a) arsenic, (b) copper, ( c ) l e a d , (d) m a n g a n e s e and (e) c a d m i u m from f o u r m a t e r i a l s as d e t e r m i n e d by E D T A Availability and N E N 7341 Availability. V a l u e s c o m p a r e d to total c o n c e n t r a t i o n m e a s u r e d by N A A or X R F (20). Unfilled items indicate that d e t e c t i o n limit v a l u e s w e r e reported.
244 ACKNOWLEDGMENT
The authors would like to acknowledge William Strawderman, Ph.D. for his statistical contributions to the project.
Research for this project was supported in part
through funding from the Hazardous Substances Management Research Center, an Advanced Technology Center of the New Jersey Commission on Science and Technology and a National Science Foundation Industry/University Cooperative Research Center. REFERENCES
Garrabrants, A.C., Kosson, T.T., and Kosson, D.S. "The determination of systematic leaching behavior of a spent petroleum catalyst." Paper presented at the Proc~d~s de Solidification/Stabilization des D~chets - Congr~s International, Nancy, France, Nov. 28 - Dec. 1 (1995). .
.
,
Kosson, D.S., and van der Sloot, H.A. "Selection of leaching tests for evaluation of treatment processes and waste management options." Paper presented at the Proc~d~s de Stabilization/Solidification des D~chets - Congr~s International, Nancy, France, Nov. 28 - Dec. 1 (1995). van der Sloot, H.A., Kosson, D.S., Eighmy, T.T., Comans, R.N.J., and Hjelmar, O. "Approach towards international standardization: A concise scheme for testing of granular waste leachability." Environmental Aspects of Construction with Waste Materials, edited by J.J.J.M. Goumans, H.A. van der Sloot and Th.G. Aalbers, 453. Elsevier Science B.V., Amsterdam, The Netherlands (1994). Gad, M.A., and LeRiche, H.H. A method for separating the detrial and non-detrial fractions of trace elements in reduced sediments. Geochim. Cosmochim. Acta. 30: 841 (1966).
5.
Tessier, A., Campbell, P.G.C., and Bisson, M. Sequential extraction procedure for the speciation of particulate trace metals. Anal. Chem. 51" 844 (1979).
6.
Miller, W.P., Martens, D.C., and Zelazny, L.W. Effect of sequence in extraction of trace metals from soils. Soil Sci. Soc. Am. J. 3 0 : 5 9 8 (1986).
7.
Linn, J.H., and Elliott, H.A. Mobilization of Cu and Zn in contaminated soil by nitrilotriacetic acid. Water, Air, and Soft Pollution. 3 7 : 4 4 9 (1988).
,
.
10.
Kosson, D.S., van der Sloot, H.A., and Eighmy, T.T. An approach for estimating of contaminant release during utilization and disposal of municipal waste combustion residues. J. Hazard. Mater. 4 7 : 4 3 (1996). NNI NEN 7341. Leaching characteristics of soil-, construction materials and wastes - Leaching tests - Determination of the availability of inorganic constituents for leaching from construction materials and waste materials. NNI (Dutch Standardization Institute), Delft, the Netherlands (1994). F&llman, A.-M. Aspects of the performance and design of the availability test for measurement of potentially leachable amounts from ash. submitted for review in Environmental Science and Technology (4/29/96): (1996).
245 1 1.
Barna, R. "Etude de la diffusion des polluants dans les d~chets solidifies par liants hydrauliques." Doctoral Thesis: Institut National des Sciences Appliqu~es de Lyon (1994).
1 2.
Chen, T.C., Macauley, E., and Hong, A. Selection and test of effective chelators for removal of heavy metals from contaminated soils. Canadian Journal Civil Engineering. 2 2 : 1 1 8 5 (1995).
1 3.
Allen, H.E., and Chen, P.-H. Remediation of metal contaminated soil by EDTA incorporating electrochemical recovery of metal and EDTA. Envir. Progr. 1 2 : 2 8 4 (1993).
1 4.
Elliott, H.A., and Brown, G.A. Comparative evaluation of NTA and EDTA for extractive decontamination of Pb-polluted soils. Water, Air, and Soil Pollution. 45-361 (1989).
15.
Yu, J., and Klarup, D. Extraction kinetics of copper, zinc, iron, and manganese from contaminated sediment using disodium ethylenediaminetetraacetate. Water, Air, and Soil Pollution. 7 5 : 2 0 5 (1994).
16.
Brown, G.A., and Elliott, H.A. Influence of electrolytes on EDTA extraction of Pb from polluted soil. Water, Air, and Soil Pollution. 62" 157 (1992).
1 7.
Kosson, D.S., Kosson, T.T., and Sloot, H.v.d. Evaluation of solidification/stabilization treatment processes for municipal waste combustion residues. NTIS PB93-229 870/AS, US Environmental Protection Agency, (1993).
18.
JMP Version 3.5.1. SAS Institute, Inc., Cary, NC, USA.
19.
Macauley, E., and Hong, A. Chelation extraction of lead from soil using pyridine2,6-dicarboxylic acid. J. Hazard. Mater. 40- 257 (1995).
20.
Landsberger, S. Nuclear techniques and the disposal of non-radioactive solid waste. Inter. Atomic Energy Agency bull. 35" 14 (1993).
21.
CRC Handbook of Chemistry and Physics. Edited by David R. Lide. 71st ed. CRC Press, inc., Boca Raton, FL).
22.
USEPA Test Methods for Evaluating Solid Waste, SW-846, Acid Digestion of Sediments, Sludges, and Soils, Method 3050. US Environmental Protection Agency, Washington, DC (1996).
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247
LEACHING CHARACTERISTICS OF COMMUNAL AND INDUSTRIAL SLUDGES
Patrick A.J.P. Cnubben and Hans A. van der Sloot ECN, Soil and Waste Research Petten, The Netherlands
Introduction
In view of regulatory requirements and difficulties in judging treatment options for communal and industrial sludges, leaching properties of treated sludges are becoming increasingly important [ 1]. The leaching behaviour of sludges is largely dictated by the nature of the sludge. The main factor in this respect is the level of organic degradable matter, which may lead to the generation of Dissolved Organic Carbon (DOC) capable of metal complexation [2,3,4]. In addition, sludges may exhibit reducing properties, either by addition of sulphides as in the case of stabilization of industrial sludges or as a result of anaerobic degradation. The consequences of differences in chemical conditions on the release from sludge can be observed from pH controlled leaching experiments performed in a pH-static test. Examples are provided for a range of sludges; sewage sludge, gas production sludge, dredging sludge and galvano-sludge. Sewage sludge and gas production sludge can be clasified as sludges containing organic matter whilst galvano-sludge has a typically inorganic signature. The aim of this presentation is to demonstrate similarities in leaching behaviour from different types of sludge. Similar leaching behaviour is demonstrated for the major elements: A1, Fe, Mg, Mn, S, P and the minor elements: Zn, Pb, Cu, Ba, Cr and Mo.
Materials and Methods
pH-static test In the pH-static leach test, the pH is continuously adjusted by acid (1 M HNO3) or base (1 M NaOH) addition to selected preset values. The general characteristics describing the tested sludges are given in table 1. The extraction is continued for 24 hours [2]. After finishing the pH-static leach test the leachate samples are filtered over a 0.20 pan filter. The leachate samples are analysed for components using ICP-AES. The results from the pH-static experiment are given for the major compounds: A1, Fe, Mg, Mn, S, P and the minor elements: Zn, Pb, Cu, Ba, Cr and Mo.
Communal sewage sludge The communal sewage sludge used in this evaluation is obtained from the communal wastewater treatment plant located in the vicinity of ECN, The Netherlands. The communal sewage sludge was leached as received (L/S -- 37.5 1/kg) at 4 pH's (4.1, 6.6, 9.4, 12.0).
Communal sewage sludge with addition of Ca(OH)2 The sewage sludge is identical to previously described sludge. Before the pH-static leach test at a presetted pH value was carried out sufficient Ca(OH)2 was added to the reach the preset pH value. Fine tuning of the pH is achieved by adjustment via addition of acid (1 M HNO3) or base liquid (1 M NaOH). The aim of this procedure is to asses the behaviour of the sludge after lime addition. The leaching experiment were carried out on the sludge as received at 4 pH's (7.1, 9.0, 9.9, 11.9).
Gas production sludge
248 The production sludge used in this evaluation is obtained from a gas production facility located in the North of the Netherlands. The pH-stat experiment were performed at 4 pH's (7.6, 8.9, 9.9, 11.9) on the sludge as received with a L/S-ratio of 4.2 1/kg.
Dredge sludge The dredge sludges is obtained from a harbor sanitation project in the South-Western part of the Netherlands. The pH-stat experiment was performed at 4 pH's (4.0, 5.5, 7.0, 8.5) on the sludge as received.
Galvano-sludge The used galvano-sludge is obtained from a galvanic industry located in the South of the Netherlands. The sludge was diluted before the pH-stat experiment was started. The L/S ratio as received was .... l/kg. During the leaching experiments a L/S - 32.6 1/kg was used. The leaching was performed at 8 different pH's (4.1, 5.6, 6.6, 7.6, 8.5, 9.5, 10.5, 12.0). Both communal sludges and dredging sludge are reducing. The general characteristics of the sludges used in this evaluation are given in table 1. Table 1. General characteristics of the slud~ es under investigation. Property communal galvano-sludge gas production sewage sludge sludge Dry weight [g/kg] 2.4 32.7 69.3 L/S [1/kg] 37.5 ~ 32.6* (2.0 #) 4.2* (0.5 #) pH as [-log H § 6.5 7.5 7.3 received 29 EH as [mV] -66 276 received The L/S-ratio applied during leaching test. # L/S-ratio of the material as received
dredge sludge
_-2.5# 7.6 -202
Results and Discussion
The aim is to point out similarities in leaching behaviour of different sludge types demonstrated for some selected elements. The general leaching behaviour will be briefly discussed per element. When exceptional leaching behaviour is observed this will be discussed. Aluminium. Figure a. The general leaching characteristic is determined by amorphous or crystalline (gibbsite) AI(OH)3 and can be observed in most materials. Iron. Figure b. In the pH domain < 7 Fe shows behaviour as Fe(III). Comparance of the Fe leaching from sewage sludge and sewage sludge with Ca(OH)2 addition shows that Ca precipitates DOC, thereby reducing the DOC mobilisation from sewage sludge. Magnesium. Figure c. In the high pH-region leaching of Mg is usually controlled by (brucite) Mg(OH)> The relatively high leaching of Mg at pH > 10 can be explained by DOC complexation. Manganese. Figure d. Mn shows typical V-shaped leaching behaviour. Comparance of the sewage sludge and the sewage sludge with Ca(OH)2 addition again shows the effect of DOC immobilisation. Copper. Figure g. The solubility of Copper in inorganic systems, such as galvano-sludge, is likely to be controlled by (tenorite) CuO. The sludges with high organic matter content also leach Dissolved Organic Carbon (DOC) capable of forming organocopper complexes which are very soluble. The presence of these compounds can explain the strongly enhanced leaching of Copper for these sludges against the leaching behaviour of galvano-sludge representing a inorganic situation.
249 Lead. Figure f. Leaching of Pb from communal sewage sludge is controlled by Dissolved Organic Carbon which forms complexes with enhanced leachability. Leaching from gas production sludge also shows DOC controlled leaching. Leaching of the communal sewage sludge with the use of Ca(OH)2 shows a decrease in Pb leaching. Addition of Ca(OH)2 precipitates the DOC thereby reducing the DOC effect. Zinc. Figure e. The behaviour of Zn is identical to the leaching behaviour of Pb. The effect of the DOC immobilisation using Ca(OH)2 shows the same results as is demonstrated with Pb and Fe. Barium. Figure h. Leaching from communal sewage sludge and gas production sludge, at high pH, is likely to be controlled by BaSO4. Communal sewage sludge with the use of Ca(OH)2 shows that in the high pH-region leaching of the sludge is controlled by both BaSO4 and BaCO3. Chromium. Figure i. There are distinct differences in leaching behaviour of the galvano sludge and the sludges containing organic matter. The behaviour of these is typical for chromate leaching. Galvano-sludge demonstrates a V-shaped pattern with minimum leaching between pH 6-7. The pattern by which Cr is leaching is matching with the leaching behaviour of Fe. It is suggested that Fe controls Cr leaching. In this case Cr can be present as Cr (III). Molybdenum. Figure j. Molybdenum exhibits typical oxyanion leaching behaviour. Above pH 5 the mobile species MoOn 2- is formed. Sulphur. Figure k. The leaching pattern of Sulphur is equal for all the sludges but differs in magnitude. The communal sewage sludge shows a high leachable amount of sulphur due to anaerobic degradation by which the sulphur is generated. Phosphorous. Figure 1. Leaching behaviour of P from the galvano-sludge is correlated with that of Fe (possibly as phosphates). High amounts of P are leached from sewage sludges because the material contains a large amount of P due to the mainly organic composition of the material. After addition of lime to the sewage sludge the amount of leachable P decreases with increasing pH.
250
Figure 1. Leaching characteristics of the different sludges, spheres (black)" communal sewage sludge, diamonds (green): communal sewage sludge with Ca(OH)2, squares (blue): gas production sludge, triangles (red): galvano-sludge, cross (purple): dredge sludge. 10000
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251 Conclusion There is a distinction in leaching behaviour between inorganic and organic matter containing sludges. The main effect that can be noticed is the contribution of DOC to metalmobilisation and leaching. DOC mobilisation in sludges, especially in sludges with high organic matter content like communal sewage sludge, is a main controlling factor responsible for high leachability in the pH domain 7 -12. By studying the general leaching characteristics of sludges and materials the behaviour of these type of materials can be identified and management options chosen according to optimal conditions [1,5]. References [1 ] H.A. van der Sloot, Developments in evaluating environmental impact from utilization of bulk inert wastes using laboratory leaching tests and field verification, Waste Management, vol. 16, pp 65-81, 1996. [2] R.N.J. Comans, H.A. van der Sloot, P.A. Bonouvrie, Speciatie van contaminanten tijdens uitloging van bodemas, december 1993, ECN-C-93-090. [3] R.N.J. Comans, P.A. Geelhoed, Speciatie onderzoek aan verontreinigde grond en baggerspecie, oktober 1996, ECN-C--96-084. [4] J. Buffle, Complexation reactions in aquatic systems, 1988, Ellis Horwood Limited, Chichester. [5] H.A. van der Sloot, L. Heasman, Ph. Quevauvillier, Harmonisation of leaching/extractions tests, Elsevier, in press. [6] The International Ash Working Group (IAWG), Municipal Solid Waste Incinerator Residues, Studies in Environmental Science 67, Elsevier 1997.
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Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997Elsevier Science B.V. All rights reserved.
253
Influence of C o n c r e t e Technical P a r a m e t e r s on the L e a c h i n g B e h a v i o u r of M o r t a r and C o n c r e t e I. Hohberg and P. Schiessl Institute for Building Materials Research (ibac), Aachen University of Technology, Germany
Abstract The main aspect in respect to environmental compatibility of cement-based materials, is the leaching environmentally relevant substances. Realistic leaching tests were carried out in order to examine the influence of mortar and concrete technical parameters on leaching. It could be shown that the leaching of heavy metals from cement-based building materials, with and without the addition of fly ash, is independent of the total amount of these components in the building material. Especially, the age of the samples has an influence on the leaching rates. Generally, the emissions from the mortar and concrete samples investigated were very low.
1
Introduction
The main aspect in respect to environmental compatibility of cement-based materials, is the release of inorganic compounds (e.g. heavy metals salts) due to the contact with water e.g. rain or groundwater. The leaching of environmentally relevant substances from cement-based building materials is often a diffusion dominated process/2/. The release due to diffusion, the diffusion rate for the individual ions/substances can be determined in practice-related leaching tests (tank tests) by investigating the time-dependent leaching behaviour. The diffusion rates allow an assessment of the success of the immobilisation of environmentally relevant substances by a cement matrix. A detailed description on leaching mechanisms and test methods is given in/2, 3/. The following influencing parameters are to be considered with respect to concrete: 9 the porosity and pore structure/density of the concrete 9 the degree of hydration and the age of the concrete 9 the carbonation The type of cement and the use of concrete additions like e. g. fly ash and silica fume as well as the water/cement-ratio should have an influence on the leaching behaviour, since they lead to a distinct change in the pore structure of the concrete/3, 4, 5/. In addition to this, the compaction, the curing, as well as the achieved degree of hydration are also of importance. The carbonation (the reaction of calcium hydroxide with the carbon dioxide from the air) of the cement matrix lowers the pH value of the pore solution. Environmentally relevant substances which had been immobilised can probably dissolve, and thus become available for the leaching processes/2/.
254 2
Investigations
2.1
Materials and methods
The investigations described in this paper were carried out in order to investigate the influence of concrete technical parameters on the leaching behaviour of heavy metals from cementbased materials with and without the application of industrial by-products. The following concrete technological influences were investigated: 9 type of cement (normal Portland cement (PC) and blast-furnace slag cement (BFSC)), 9 type of addition (addition of fly ashes with different heavy metal contents), 9 w/b-ratio (w/(c+0.5f)=0.4, 0.5, 0.6), 9 age of the samples (90 days, 180 days and 360 days). A normal portland cement (PC) and a blast-furnace slag cement (BFSC) in compliance with the requirements of German standard DIN 1164 were selected for the investigations. Four bituminous coal fly ashes (FA) with different heavy metal contents were chosen as concrete addition. For the investigations mortar mixtures with and without addition of FA were prepared according to European standard EN 196-1. The fly ash content (f) in the mixtures with FA was 20 mass.% in relation to the total binder content (c+0.5f). Mortars with different w/b were produced. Additionally, concrete mixtures with and without addition of FA were prepared. The grain size distribution of the concrete aggregates corresponded to grading curve A/B 16 in accordance with the German standard DIN 1045. The w/b was 0.5 for all concretes. Mortar specimen (40 by 40 by 160 mm) and concrete cubes (100 by 100 by 100 mm) were produced from the mixtures. After one day in the mould, the specimen were cured until the investigations in a climate chamber at 20 ~ and 95 % relative humidity. Tank leaching tests similar to that described in the Dutch standard NEN 7345 with a duration of 56 days and 8 leachant renewals were performed with the mortar specimen and the concrete cubes. 2.2
Results and discussion
The results from the tank leaching tests with the mortar and concrete specimen are summarised in table 1. Generally, the amounts of heavy metals leached are very low, often near the detection limits, due to the immobilisation of the heavy metals in the highly alkaline, dense cement matrix. The addition of fly ashes with relatively high contents of heavy metals does not result in higher emissions. Due to the pozzolanic reaction and the filler effect of the fly ashes the emissions from mortar with fly ashes are often lower than those for the mortars without fly ash addition (especially for zinc). The leaching results from the mortars are represented for zinc in dependence on the total zinc content in Fig. 1. The total zinc content was each divided into the fraction from the aggregate, the cement and the fly ash that were calculated based on the composition and the contents of the ingredients.
255
Table 1:
Tank leaching results from the investigations of mortars and concrete in dependence on cement and fly ash addition (w/b =0.5; f/c=0.25)/2, 3/ mortar
material
concrete
PC
cement
BFSC
PC
- levi
addition
I
I
total amount leached in mg/kg 1)
parameter As
< 0.01
Cadmium, Cd
< 0.01
Arsenic,
[BFSC
Chromium, Cr Copper, Cu Lead, Pb
0.09 0.15 0.03 < 0.01 < 0.01 0.01 0.01 0.03 <0.01
0.03 0.02 0.02
0.07 0.01 0.02
<0.02
10.02 10.02 < 0.02
0.08 0.09 < 0.01 <0.01 0.03 <0.01
0.02 0.10 < 0.01 0.02 <0.01 <0.02
0.09 0.02 <0.02
0.04 0.02 <0.02
0.05 < 0.02 <0.02
Nickel,
Ni
< 0.01
n.d.
Zinc,
Zn
0.18 I 0.09 I 0.07 I 0.09 I 0.11 I 0.02 I 0.02 1<0.02
0.13 [ 0.04 [ 0.04 I 0.04
1) leached amounts after 56 days and 8 leachant renewals PC: normal portland cement; FA 9bituminous coal fly ash; BFSC: blast-furnace slag cement Zinc content in mg/kg 0.20
leached zinc [mg/kg] o
0.15,
14s
I
II
t3s
/
0.10
0.05
0.0 PC
o
hout addition of FA
PC/FA1PC/FA2 PCIFA3PC/FA4 BFSC BFSC/FA1BFSCIFA2 cement/addition
without addition..~f FA
; 0
20
40 60 leaching time [days]
Fig. 1." Tank leaching test results in relation to total zinc contents for different mortars in dependence on cement and addition (PC: portland cement; FA."fly ash; BFSC: blast-furnace slag cement," w/(c+O.5J)=O.5for all mortars)/1/
The amounts of zinc leached in the tank tests are, for mixtures with addition of fly ash, inspite of higher total contents, less than those of the corresponding mixtures without fly ash. This applies to the mortars as well as to the concretes. This effect can be attributed to the immobilisation of zinc due to the formation of insoluble complex salts. The diffusion of zinc in the leachant is, in addition to this, impaired due to the compacting effect of the fly ash. The amounts of leached zinc are especially low for blast-furnace slag cement mortars. The zinc is presumed to be fixed in the glassy matrix of the blast-furnace slag and thus cannot be mobilised. Fig. 2 shows the tank leaching results from mortar specimen for chromium. The influence of the fly ash addition is not as distinct as for zinc. The chromium amounts leached for mortar with fly ash FA1 lie above the value for the mortar without the addition of fly ash. Fly ash FA1 has a high fraction of soluble chromium compared to the other fly ashes/2/. A decrease
256 in the amount of leached chromium can be recognised in the other mortar mixtures with fly ash. This can also be attributed to the compacting effect of the fly ash.
60T~miu"~m3c~ II ~
PC
!n mg/----kg I R*=~""='~'~t"/~... I] ,Is
4,
I I'] :
Fc~' J'~
]/
PC/FA1PC/FA2 PC/FA3PC/FA4 BFSC BFSC/F^IBFSC/FA2 cement/addition
t[
reachedchr~ ~
10"121
I /
0
I
~
m [ g/kg] BFSC with and
~
without a d d i _ t ~
20
40
l,
[
60
leaching time [days]
Fig. 2: Tank leaching test results in relation to total chromium contents for different mortars in dependence on cement and addition (PC: portland cement; FA: fly ash; BFSC: blast-furnace slag cement; w/(c+O.5J)=O.5for all mortars)
The amounts of leached chromium from the mixtures with blast-furnace slag cement are rather high compared to the total and available chromium content. The reason may be a reduction in the pH value of the pore solution due to the large content of blast furnace slag. This effect does not occur with zinc, since zinc has its minimum solubility at pH values around pH-10. leached chromium in mglm2
4.0 3.o-
age
=
M1with-
M2 with
28 d
~
-----0-
90~
--o
38od
out FA
. . . .
n--.
...o ...... o...
2.0
1.0I ~
-""c>----~-~
~~'~
--..-..-0
0
o co,~,~,,..~,.-~-,~ -,---,-~---~-- -----'~ 0
In Fig. 3, the leaching behaviour of chromium for mortars with and without addition of fly ash is shown in dependence on the age of the samples. As expected, the leached amounts decrease with increasing age of the samples. The influence of the hydration age occurred in the same manner for the other elements/2/.
FA1,
20
40
60
80
100
2.3
Determination of emissions
leaching tim in days
The release of heavy metals from cementbased building materials is mainly controlled by diffusion. In order to be able to compare and assess the results from tank tests, the heavy metal amounts which were released due to diffusion (emission after 365 days) were calculated from the test results (details of the calculation see/2, 3/). Fig. 3: Results from tank leaching tests; leaching behaviour in dependence on the age of the samples
The results of the calculations for the investigated mortars and concretes with the examples of zinc and chromium are represented in Fig. 4. The following can be taken from the results:
257
calculated emission after 365 days in m g / m 2 . ~
I
concrete
mortar
Zinc
_
i..:~;
.
~
. . . . . PC . .FA .....
1 .
I
I
i;-
PC with FA
. . . . BFSdC t~h;At
PC ~,Fh~U PC
"
BFSC SFA w|l;AU I
Fig. 4: Chromium and zinc emissions from mortars and concretes in dependence on type of cement and addition (PC: cement; BFSC: blast-furnace slag cement; FA: fly ash; w/(c+O.5J)=0.5)
9 Smaller emissions due to diffusion were yielded for the PC mortars with addition of fly ash in relation to that without fly ash addition. This can be attributed to the filler effect of the fly ash and to a compaction in the pore system due to the pozzolanic reaction. This leads to a hindering in the diffusion process. All fly ashes had a diffusion hindering influence on the metals zinc, copper and lead/2, 6/. The influence was not so clear in the case of chromium.
9 The chromium emissions calculated for the mortars made of blast-fumace slag cement are larger than those for portland cement mortars, whereas only very small amounts are released for the heavy metals zinc, copper and lead. The addition of fly ash does not have a strong influence on the release rates. This can be explained by the fact that the use of blast-furnace slag cement strongly reduces the release rates. The additional effect of the fly ash is not noticeable.
9 The emissions calculated are generally larger for concretes than for mortars. This could be attributed to the higher porosity of the hardened cement paste in the contact zone areas between aggregate and hardened cement paste. This effect is more distinct for concrete, because of its higher fraction of contact zones, than it is for mortar. Another possible explanation is the smaller fraction of cement matrix resulting in lower pH values in the leachates. calculated emissions after 365 days in m g l m 2 without F A - - - with FA O
0.3
zinc
[]
chromium
zx
copper
) .......
0.4
o -" "" " " " "" " "
0.5
0.6
0.7
w/b-ratio ( w l ( c § 0 . 5 f ) )
Fig. 5: Emissions from mortar with and without addition of fly ash in dependence on water~binder-ratio
In Fig. 5, the emission in dependence on water/binder-ratio are represented for mortar with and without addition of fly ash. Fig. 5 shows no distinct influence of the w/b-ratio on the emissions. The emission is partly reduced with higher w/b ratios (contrary to what was expected). This is attributed to higher dissolution of other ions (like calcium) which lead to a precipitation of the heavy metals. Generally the leached amounts are very low, therefore, the influence of the w/b-ratio could not be deduced. For main elements like sodium or potassium, the emission increase with increasing w/b-ratios/2/.
258 3
SUMMARY
The results represented here have shown, that the leaching of heavy metals from cement-based building materials, with and without the use of fly ash is largely independent of the total amount of these constituents in the building material. The immobilisation of the substances considered and the density of the cement matrix are decisive in the leaching. Generally the emission of heavy metals from the mortars and concretes investigated were very low. The addition of fly ashes reduces in most cases the leaching rates, especially for zinc. The w/b has no distinct influence on the leaching of heavy metals. Higher ages of the cement-based materials and thus the hydration grade lead to lower leaching rates. 4
References
1.
SchieBl, P.; Hohberg, I.: Umweltvertr~iglichkeit yon zementgebundenen Baustoffen. In: Forschung, Vortr~ige der DBV-Arbeitstagung (1996), S. 65-72
2.
Schie/31, P.; Hohberg, I.: Umweltvertr~iglichkeit von zementgebundenen Baustoffen: Untersuchungen zum Auslaugverhalten von sekundaren Baustoffen. Aachen, Institut ftir Bauforschung 1995. Forschungsbericht Nr. F 414
3.
Hohberg, I.; Mialler, Ch.; SchieBl, P.; Volland, G.: Umweltvertr~iglichkeit zementgebundener Baustoffe. In: Schriftenreihe des Deutschen Ausschusses fiir Stahlbeton (1996), Nr. 458
4.
Hardtl, R.: Ver~'aderung des Betongefiiges durch die Wirkung von Steinkohlenflugasche und ihr Einflul3 auf die Betoneigenschaften. In: Schriftenreihe des Deutschen Ausschusses ftir Stahlbeton (1995), Nr. 448
5.
SchiefS1, P.; Alfes, Ch.: Verwendung von kiinstlichen Puzzolanen als Zusatzstoff im Beton- Grunds~itze der Wirksamkeit. Teil B: Silicastaub. Aachen, Institut fiir Bauforschung 1993. Forschungsbericht Nr. F 344 - 1B
6.
Hohberg, I.; SchiefSl, P.: Einfluf5 betontechnischer Parameter auf das Auslaugverhalten von M6rteln und Betonen. Freiburg : AEDIFICATIO, 1996. In: Werkstoffwissenchaften und Bausanierung. Berichtsband zum vierten Internationalen Kolloquium, Band III, S. 1505-1519
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
259
Construction Raw Materials from Coal Fired Powerstations By-products management and quality control Authors: J.W. van den Berg A. Boorsma Vliegasunie PO box 301 3730 AH De Bilt The Netherlands phone: +31 30 2209109 fax: +31 30 2204444 Abstract It is a challenge to use the potentials of the by-products of coal combustion. Because of environmental, economical and technical reasons full utilization has to be stimulated. In the Netherlands, by-products from coal-fired power stations are fully accepted and utilized for 100% as construction raw materials in the building industry. Vliegasunie has the task to find and encouragemsponsible (and economically attractive) uses for the by-products. This is only possible when a great deal of attention is paid to quality control. Quality control is needed in order to be able to provide the market with good predictions of the technical quality of the fly ash to be supplied as construction raw material.
Introduction Vliegasunie (Dutch Fly Ash Corporation) was founded in 1982 by the Dutch electricity generating companies with the task of marketing all the by-products which are produced when electricity is generated using coal as fuel. As of January 1995 the activities of Vliegasunie were integrated with those of Gemeenschappelijk Kolenbureau Elektriciteitsproduktiebedrijven ("GKE"), the "sister" company which is responsible for the supply of coal to the Dutch power stations. This allows for a better grip on the actual production process.
260
In the Netherlands there are 8 coal fuelled electricity production units, divided over 6 locations, which used approximately 9 million tons of coal fuel in 1995. These locations are: EPZ
Amercentrale Geertruidenberg
EPZ
Centrale Borssele
Unit A8
645 Mwe
Unit A9
600 Mwe
Unit BS12
403 Mwe
EPON Centrale Gelderland Nijmegen
Unit G 13
602 Mwe
EZH
Unit MV1
518 Mwe
Unit MV2
518 Mwe
Unit HW8
600 Mwe
Unit MC7
250 Mwe (ICCG)
UNA
Centrale Maasvlakte Hemwegcentrale Amsterdam
Demkolec Buggenum
The total coal fuelled electricity production capacity in operation at this moment is 4,136 MWs. The total installed capacity of the Dutch electricity generating companies was approx. 15,000 MW. The share of coal as fuel for the generation of electricity is about 45 % The fuel input is as follows: 9 natural gas 9 uranium 9 coal
47% 8% 45 %
GKE buys 8-10 million tons of thermal coal annually, accounting for 5% of the world trade in thermal coal. GKE imports coal from a large number of mines spread over different parts of the world. The geographic origin of the coal used in the Netherlands was: 1995
1996
Australia
22
4
%
U.S.A.
27
25
%
South Africa
16
21
%
Columbia
16
27
%
Poland
6
5
%
Indonesia
8
16
%
Others
5
2
%
261
The composition of the blend of coal which is purchased and blended for the power stations is crucially important to the quantity and quality of the by-products produced by electricity generation. The way in which the coal is supplied and stored also affects the characteristics of the fly ash produced. Therefore the coal is blended before conveying it to the power plants in order to ensure a constant quality. The coal combustion by-products from the coal fired power stations per annum are: Pulverized fuel ash (fly ash)
800.000 - 900.000 ton
Bottom ash
100.000 - 120.000 ton
Fluegas desulphurization gypsum (FGD gypsum).
375.000 - 425.000 ton
Coal gasification slag Coal gasification fly ash
40.000 - 50.000 ton 5.000
-
10.000 ton
The total amount of the coal combustion by-products from coal-fired power stations is between 1.3 and 1.5 million tons.
Fly ash
Pulverized fuel ash is formed during the process of combustion of pulverised coal in the furnace of the power station's boiler. The coal is ground to powder-coal in coal mills, before being blown by compressed air into the furnace and burned at temperatures ranging between 1300 and 1600 ~
depending on
the type of boiler in question. Coal contains 10 to 15% non-combustible mineral material. Most of this material melts in the furnace. The majority (approximately 90%) of the non-combustible material is transported into the stack by means of the flue gasses, where it solidifies into fine grained ash particles (fly ash). These ash particles are extracted from the fluegases by electro-static precipitators which have an efficiency of almost 100%.
The fly ash extracted from the precipitators is then transported to silos for dry storage. The fly ash which can't be marketed immediately is moistened and stored in open air depots.
262 Fly ash is primarily used as construction raw material in the building industry. The applications in the various market areas in 1996 has been: Cement industry
58 %
- cement production 22 % - clinker production
36 %
Artificial gravel
18 %
Asphalt filler
10 %
Concrete
10 %
Others
4%
Cement One of the applications in the cement industry is as raw material for Portland clinker. Portland clinker is manufactured by cindering a homogeneous mixture of ground lime stone and claylike materials. Fly ash can be used as a substitute for these claylike materials because it has practically the same chemical composition. An other application of fly ash in cement is as raw material for Portland Fly Ash Cement. The cement industry manufactures class A Portland Fly Ash Cement which has the same characteristic properties as normal class A Portland Cement. This is achieved by using a finer ground, high quality Portland clinker and adding approximately 25% high quality fly ash.
Artificial gravel Fly-ash is used for the manufacture of artificial gravel (Lytag). Lytag is manufactured by cindering spherical fly ash pellets. The unburned coal particles in the fly ash are used as fuel in the production process. Lytag is produced by VASIM in Nijmegen. Lytag can be used as replacement material for conventional mineral aggregate such as gravel or crushed rock.
263
Asphalt filler Fly ash is also supplied to the asphalt industry as filler in asphalt mixures for road constructions.
Concrete Finally there is the market segment concrete. In this segment the fly ash is delivered directly to the users, for application in concrete mixures as partial replacement of cement, or as a filler. For these applications the fly ash is delivered under certification.
Bottom ash Bottom ash is formed by the slagging of ash particles in the lower parts of the furnace. The amalgamated ash falls through open grids in the furnace floor where it is collected in hoppers. The bottom ash is then moistened and stored in open air depots.
The largest market segment here is the application as light weight sub-base material in road construction and other civil constructions, which accounts for 59% of the bottom ash production. The remaining 41% goes into the manufacture of concrete blocks.
FGD-Gypsum Fluegas desulphurisation gypsum is formed during the process of removing sulphur dioxide from the fluegases. The fluegases are led through a scrubber and oxidation tower, where they are brought into contact with a lime or limestone suspension. After an interaction with an excess of air, gypsum is formed. Up until 1992 most of the FGD-gypsum was used for the manufacture of plaster board. Since the middle of 1993, approximately 25% of our FGD-gypsum has been upgraded to anhydride to be used as an additive to influence the setting time of mixes for self levelling floors. The other 75 % is still used for the manufacture of plaster board.
264
Coal gasification by-products Integrated Coal Gasification Combined Cycle is a new technology for the generation of electricity. Gasification of coal is a process in which coal is partially oxidated by air, oxygen, steam or carbon dioxide under controlled conditions to produce a fuel gas. The hot fuel gas is cooled in heat exchangers, with the production of steam, and cleaned before combustion in a gas turbine. The offgases from the turbine are used in a boiler to produce additional steam for a steam turbine. The electrical efficiency can be around 45% with minimal impact on the environment. A demonstration-unit of 250 MWe has been constructed in Buggenum, the Netherlands, based on the Shell Coal Gasification Process.
The by-products coal gasification slag and fly ash are formed in the gasifier reaction vessel and differ clearly from bottom ash and fly ash from pulverized coal-fired boilers. This is due to the reducing conditions and the higher operating temperatures in the gasifier. The slag which runs from the reactor wall is quenched in a water-bath at the bottom of the reactor and subsequently locked out off the system. The fly ash is separated with a cyclone. Part of the fly ash is recirculated to the gasifler to obtain higher carbon conversion levels. The composition and properties of the fly ash will depend on the degree of recycling.
There will also be a significant difference with respect to the types of ash produced. In a pulverized coal-fired boiler about 90% of the ash is fly ash and about 10% bottom ash, whereas in the Buggenum gasifier approximately 80% of the ash will end up in the slag and 20% in the fly ash. Characteristic for the Buggenum unit will be that the sulphur present in the coal is converted to elemental sulphur instead of gypsum as is normal practice in Dutch pulverized coal-fired units. This pure form of sulphur is a basic raw material for the chemical industry.
265
A research programme has been started in order to identify timely industrial applications for coal gasification slag and other coal gasification by-products. A steering committee consisting of representatives of Demkolec, KEMA, Novem and the Vliegasunie supervises the research- and development work in accordance with the Industrial Development Program Coal Gasification (IOKV).
The research programme was initially focused on the characterization of coal gasification slag and fly ash. The suitability of coal gasification slag for civil-technical applications was initially investigated on a small scale by means of specimens in the laboratory. Subsequently larger quantities have been produced industrially on pilot plant scale and during the first year of the demonstration period. The properties of these applications have been investigated by laboratory testing (such as leaching behaviour and pressure strength) The slag was studied as a replacement for sand in concrete and concrete products, in asphalt concrete and civil engineering applications and as an alternative in foundation and filling work. The fly ash was studied for the use in concrete.
Marketing For the marketing of all the by-products Vliegasunie has a co-operation agreement with a subsidiary of Cementbouw, Vulstof Combinatie Nederland BV ('~/CN"). VCN buys all the byproducts, fly ash, bottom ash and FGD-gypsum, which are not already reserved either for delivery to existing customers on the basis of long-term contracts or for processing into artificial gravel in the factory of Vasim. VCN is also responsible for carrying out the logistics for Vliegasunie's existing iong-term contracts.
266 Quality
System
General The Dutch power stations are fuelled with bituminous coal. As was mentioned earlier GKE imports coal from different parts of the world and from a large number of mines on behalf of the Dutch power stations. The composition of the blend of coal which is purchased and blended for the power stations is crucially important to the quantity and quality of the by-products produced by electricity generation. The way in which the coal is supplied and stored also affects the characteristics of the fly ash produced. Vliegasunie's integration with GKE is also expected to yield benefits in the area of quality control. Better insight will be possible into the relationship between the quality of coal and that of the construction materials. This is needed in order to be able to provide the market with good predictions of the technical quality of the fly ash to be supplied as construction raw material.
The diversity in geographical origin of the coal in combination with the variety of boiler types, can lead to a great variation in the composition of the fly ash. Parameter
Average
St. dev.
SiO2 AI203 Fe203 CaO MgO Na20 K20 TiO2 P205 pH L.O.I. < 32 mu < 45 mu
55.5 26.3 7.5 4.3 1.6 0.6 1.7 1.4 0.5 10.7 4.5 73.1 80.9
4.2 2.8 2.0 2.2 0.5 0.4 0.5 0.2 0.4 1.6 2.3 7.2 6.3
! % % % % % % % % % % % %
Average quality of the fly ash produced in the Netherlands
267
Each application in the various market areas in the building industry has its own specific quality requirements. These requirements are partially dictated in standard specifications as EN-450, ENV 197 and ENV 206, which are being drafted by the CEN, as part of the European Regulations. Some of the quality requirements are dictated by the relevant industries themselves.
Information System (ISRA) Vliegasunie's policy is based on the assumption that the construction raw materials which are delivered to its customers will meet quality criteria agreed upon in advance with the customer. In order to meet these criteria, good co-ordination between quality and logistics is essential. In order to achieve this, an Information System for the Marketing of Coal Residues ("ISRA") was set up which provides access to logistical and quality data, and allows the relationships between quantities and qualities of construction materials to be traced. This system is an important element in the quality system. Following the introduction of the system in 1994, extensive work was carried out on data communication with the power stations. Every attempt was made to align these systems as closely as possible with the information systems of the power stations, and to take advantage of new developments such as links with the weighbridge computers and the Laboratory Information Management Systems (LIMS).
Quality control Upon delivery to the Netherlands the coal is analysed. Based on the results of this analysis, and taking into account the unit and the burner in which the coal will be fired, it is possible to predict the composition and the quality of the fly ash that will ultimately be produced. This gives the great advantage of being able to select the possible applications and thus the potential markets for the fly ash before it has been produced. The quality control is performed at power stations, which have modern and adequately equipped laboratories at their disposal.
268 Each day samples are taken for chemical and physical analysis. The test results are forwarded to the Information System (ISRA) where they are compared with the results of the aforementioned prognosis. If necessary the logistics concerning the destination of the fly ash can be adjusted. Depending on the client's quality requirements, extra samples can be taken and analysed before and during the loading of the fly ash. In this way the client can be given the assurance that the particular delivery of fly ash meets his specific requirements.
Certification Certification entails guaranteeing the quality that is provided by Vliegasunie, both in terms of the materials supplied and in terms of environmental factors. The procedures which are laid down for this purpose cover the whole area of quality control. The requirements for the utilisation of fly ash in concrete are dictated in the EN 450. Based on these requirements a procedure for the certification of fly ash has been agreed upon with a certification institute. This procedure consists of a set of guide lines for the assessment of fly ash and incorporates an internal quality monitoring scheme, which is performed under the auspices of the certification institute. The quality control of the fly ash production is part of the internal quality monitoring scheme.
As part of the quality system attention will be paid to the following elements of quality control: - Quality control procedures -
Quality manual
- Round robin tests - Sampling procedures
Concluding
remarks
In the Netherlands the by-products of the coal fired powerstations are fully accepted as construction raw materials. An adequate quality system is essential for such a measure of market acceptance.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
269
MAASVLAKTE FLY ASH PROCESSING PLANT
Authors: Jos. B.M. Moret, Marketing Manager Jan W. van den Berg, Quality Manager Vliegasunie PO box 301 3730 AH DE BILT The Netherlands phone: + 31 30 2209 109 fax: + 31 30 2204 444 Abstract
In 1995 Vliegasunie put the Maasvlakte Fly Ash Processing Plant in Rotterdam, The Netherlands, in operation. With a capacity of 250.000 tonnes per annum, this installation has an important role in maintaining a record of 100% use of Dutch fly ash by continuing to meet the customer's need for fly ash of constant quality and quantity. The main objectives of the installation are storage and upgrading of fly ash. The fully computer controlled installation is located adjacent to the site of EZH's Maasvlakte Power Plant in Rotterdam. The Plant consist of an intake silo, an upgrading plant and an end-product silo. In the upgrading plant the carbon content of the fly ash is reduced to less than 5% and the fineness is improved to over 70% less than 45 pm. The Plant has a blending installation in order to produce a desired constant product quality. The total storage capacity for incoming fly ash is 9.000 tonnes and for end-product is 32.000 tonnes. Most of the fly ash the Plant receives from the adjacent Maasvlakte Power Plant by pipeline. The rest of the fly ash is transported to the Plan both by bulk truck and by barge. The upgraded product is transported to the customers both by bulk trucks and by barges.
270
For internal transport airslides are installed for horizontal transport and a pneumatic transport system for vertical transport. To meet with strict environmental regulations, all air used for transporting fly ash is filtered in order to reduce the possible dust emissions to 10 mg dust per m 3 max. The first year of production shows that the installation is functioning as expected.
Introduction In thespring of 1995, construction of Vliegasunie's Maasvlakte Fly Ash Processing Plant in Rotterdam, The Netherlands was completed. In the course of 1995, an extensive programme of testing and alignment took place as this plant is the first in the world to upgrade fly ash produced by coal fired power plants on such large scale. With a capacity of 250.000 tonnes per annum, this plant is intended for the upgrading of approximately 40% of the total production of fly ash of the Dutch power generating companies and represents a major investment in the future on the part of Vliegasunie and its shareholders. Only through such continued investment does Vliegasunie expect to maintain its record of 100% use of its by-products by continuing to meet the customer's needs for a fly ash of constant quality and quantity. Objectives of the installation As known, the quality of fly ash depends on several factors. The main factors are the type of coal that is used for power generation, the completeness of the coal combustion, the type of furnace and the type of flue gas scrubbers. In 1988, Vliegasunie foresaw that the quality of the fly ash produced by the Dutch power generating companies would diminish and the demand for high quality fly ash for the cement industry and as a filler and binding agent in concrete would increase.
271
One of the major factors for the diminishing quality of fly ash is the agreement between the Dutch government and the combined power generating corporations to reduce NOx and CO2 emissions. For this reason, the coal fired power plants installed LowNOX installations. These installations have a negative effect on the quality of fly ash. Another problem Vliegasunie foresaw was how to match production and demand. In the winter when production of fly ash is high demand is low; in the summer demand for fly ash is high while production is low. These two problems resulted in the design of the Maasvlakte Fly Ash Processing Plant. The plant has two objectives: (dry) storage and upgrading of fly ash. The storage facility allows matching of production and demand; the upgrading guarantees a constant high quality to our customers who need a high and constant quality of fly ash: the cement and concrete industries. Through sieving and blending of different qualities of fly ash, a constant quality can be guaranteed. Table 1 shows the quality of the fly ash produced in this plant. Table 1 Quality parameters Code description
unit
minimum
maximum
SiO 2 Silicon oxide
%
50
60
AI203 Aluminium oxide
%
20
30
AI/Si
-
0.45
Fe203 Iron oxide
%
0
10
CaO Calcium oxide
%
0
5
MgO Magnesium oxide
%
0
4
C*
%
3
4.5
pH
-
10
14
451Jm % less than 45 IJm
%
75
90
Loss on ignition
After completing several studies, engineering and obtaining the necessary permits, Vliegasunie commenced construction of the installation at the end of 1993.
2?2
Lay out The installation is located adjacent to the site of EZH's Maasvlakte power plant at Rotterdam, next to the existing fly ash silo bins belonging to the power plant. Figure 1 shows an overview of the plant.
Figure 1 Overview of the Maasvlakte Fly Ash Processing Plant ince the costs of transportation of fly ash from the production site to the client form a considerable part of total costs, the logistical costs of supply and delivery need to be minimised. Aspects of cost minimisation which led to the choice of this location are: the 9 availability of a site next to a power plant to allow direct transport to the installation; a9 power plant that produces enough fly ash of a quality that can be upgraded; a9 location with excellent hinterland connections both by barge and bulk truck.
2?3 The Maasvlakte Fly Ash Processing Plant consists of an intake silo, an upgrading plant and an end-product silo. The square intake silo consists of 9 slip-formed concrete bins of 1,000 tonnes content each. The fly ash content of the concrete used for the silo is up to 70 kg/m 3 All intake of fly ash by bulk trucks, barges or from the existing EZH fly ash silos, is via this intake silo. Underneath the bins of the intake silo, a blending installation is located. Up to three different flows of fly ash can be blended in a continuous product flow. Figure 2 shows part of the transport system underneath the bins.
Figure 2 Transport installation underneath intake silo bins Also underneath the bins, the control room is located. From this control room all operations (intake, outtake, internal transport, blending and sieving) are controlled. For this purpose, the installation has a fully computerised control system that can be operated by using a mouse-controlled computer screen. All transport routes are shown on a mimic panel.
274
To observe loading and unloading of barges and trucks and the sieving in the adjacent upgrading installation, the operators have three video screens with cameras on various locations. Next to the intake silo is located the upgrading plant. The upgrading plant has six sieves to reduce the carbon content of the fly ash to less than 5% and to improve the fineness to over 70% less than 45 l~m. For quick analysis of fly ash samples a small laboratory is set up in the upgrading plant. More complete analyses of the samples are carried out in the laboratory of the Maasvlakte power plant. The end product silo is located next to the intake silo. This silo consist of 4 slipformed concrete bins of 8,000 tonnes content each. Each end-product bin has a loading spout for loading into bulk trucks. The fly ash and the fly ash residues from the upgrading process can be also loaded into bulk trucks underneath the intake silo. Barges can be loaded via the ship loading bin and two buffer bins in the intake silo using the existing barge loading installation of the Maasvlakte power plant. Next to the end-product silo is found the weighbridge house. The complete computerised administration of ingoing and outgoing flows of fly ash is found in this building.
Operations Intake of fly ash, into the plant, takes place in three different ways: from 9 trucks (approximately 6,250 t/a = 210 trucks/a); from 9 ships (approximately 56,250 t/a = 100 ships/a); directly 9 from the Maasvlakte power plant (approximately 187,500 t/a). Figure 3 shows the simplified flow chart of the processing plant.
275
Upgrading BulkTruc&
~BulkTruck
q
BulkShip ~BulkShip
Pipeline --!
Blending
Maasvlakte Fly Ash Processing Plant
Figure 3 Flow chart Maasvlakte Fly Ash Processing Plant The quality characteristics of intake fly ash product are obtained from the information sent by the supplier. Random product tests are made in order to confirm these characteristics. The confirmed product characteristics are finally entered into the Fly Ash Information System. A planner then makes a planning in the Information System for the storage and internal upgrading of the fly ash, according to the demands of each customer. Different types of weighing of the intake fly ash takes place, depending on the type of intake: by 9 means of a truck weighbridge for intake from trucks; by 9 means of draught survey for intake from ships; by 9 means of a continuous flow-meter for intake directly from the Maasvlakte power plant.
276
The weighing result is entered into the Fly Ash Information System by an administrator, and is linked to the storage and/or upgrading planning for that fly ash batch. The Maasvlakte Fly Ash Processing Plant makes use of self-unloading trucks and ships. For trucks, the unloading equipment is mounted on the trucks. Using compressed air to build up the pressure, this equipment discharges the fly ash pneumatically, through conveying pipes, to the selected storage bin. In the case of ships, a booster pump transfers the fly ash, via conveying pipes, into one of the silo bins of the intake silo. Fly ash coming directly from the Maasvlakte power plant is transported via airslides and the booster that is also used for ship unloading conveyance to the selected intake bin in the intake silo. After weighing and unloading, the fly ash is conveyed to one of the bins in the intake silo. The exact bin chosen for storage is determined by a plant operator, who bases his decision on the planning made in the Fly Ash Information System, the type of intake and the quality characteristics of the batch. Internal processing of fly ash takes place in one or more of three different processes: upgrading; 9 9 internal 9 transport. Fly ash that is to be upgraded can only be extracted from one storage bin. This means that blending during the sieving operations is not possible. Upgrading of fly ash results in two fly ash types: the upgraded type which will be transhipped to one of the intake bins in order to be blended with other flows, and the residue type that is always transhipped to a determined bin for residues for temporary storage. After upgrading, every batch is sampled. The samples are tested on various product characteristics. The fly ash is stored until the test results are known and have been entered into the Fly Ash Information System. The System couples the new quality characteristics of the fly ash batch to the already existing internal planning. Through blending, two or even three batches of fly ash with different compositions and/or varying particle sizes can be mixed together in order to produce one batch of fly ash of a desired constant product quality.
277
The blender is fed from several bins in the intake silo. The blended fly ash is pneumatically transported back to a bin in the intake silo. After blending, a sample is taken from every batch and the samples are tested on various product characteristics. The fly ash is stored until the test results are known and have been entered into the Fly Ash Information System. After that, the fly ash is transported to the end-product silo for outtake. For internal transport between mutual bins and the upgrading plant, from bins to the blending installation or from bins to truck loading equipment, airslides are installed for horizontal transport and a pneumatic transport system for vertical transport. Fly ash stored in an intake or end product bin can be recirculated through the Silo complex, the starting and end bin being the same. This is done in order to guarantee the quality of the fly ash. Each internal transport is planned by the logistic planner with the help of the Fly Ash Information System. Outtake of fly ash can take place in one of two ways: by 9 truck (approximately 100,000 t/a = 3,400 trucks/a); by 9 barge/ship (approximately 150,000 t/a = 250 ships/a). Empty trucks are weighed (tare weight) when they arrive on the site. The operator selects the storage bin from which fly ash will be loaded on the basis of the information in the Fly Ash Information System. Via aspirated loading spouts the trucks are loaded. After loading, the trucks are again weighed (gross weight). The results are again entered into the Information System. For each incoming barge or ship, a draught survey is made. The operator again selects the storage bin from which fly ash is to be loaded on basis of the information in the Information System. Fly ash is transported via an aerobelt to the ship loading installation. Via an aspirated loading spout the ship is loaded. After loading, a draught survey once again takes place (gross weight). Also, the results are entered into the Information System. To meet with strict environmental regulations, all air used for transporting fly ash is filtered in order to reduce the dust emissions 10 mg dust per m3 maximum.
278
Sampling Power plants that supply fly ash to the Maasvlakte Fly Ash Processing Plant also send information on product characteristics for that specific batch of fly ash. These are stored in the Fly Ash Information System. The plant manager occasionally authorises random sampling and laboratory testing of intake fly ash in order to confirm the fly ash quality specifications sent by the supplier. After all types of internal processing product sampling and laboratory testing takes place. Through testing, the product quality after processing and therefore the quality of stored fly ash can be determined. A sample from each outtake batch of fly ash is also laboratory tested in order to be able to determine the quality characteristics of the fly ash being shipped to a customer. For internal processing use, the fly ash samples are tested on only max. four parameters: carbon content, fineness, colour and pH. For intake and outtake testing, up to 25 product quality characteristics are examined. Most of these samples have to be analysed in external laboratories. These laboratories are required to have the test results ready within 16 hours after the sample was taken. All the test results are entered into and stored in the Fly ash Information System which can therefore, when requested, calculate the quantity and quality of fly ash sent to each customer on each day. In this way, invoices are easily produced, and storage and upgrading records are easily kept.
Conclusion In the year that has installation passed since commissioning the installation, the Fly Ash Processing Plant has proven its capability of producing a constant quality of fly ash supported by the Vliegasunie Fly Ash Processing Plant Quality system. With this quality ash several tests were made to produce concrete. The outcome of these tests resulted in a Product Certificate on the fly ash, controlled by the Dutch Certification Institute Kiwa. With this product certificate Vliegasunie guarantees her customers in the concrete industry a quality of fly ash suitable to produce excellent concrete. With this highly sophisticated installation Vliegasunie expects to be able to match production and demand, both in quality and quantity, for the coming 15 years.
Goumans/Senden/van der Sloot, Editors Waste.Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
279
FLY ASH AS BINDER IN CONCRETE
Leo J.L. VISSERS KEMA P.O. BOX 9035 6800 ET Arnhem, the Netherlands
ABSTRACT
Research efforts over the past years towards an effective utilisation of fly ash in concrete have lead to a quantitative understanding of the efficiency factor (k-value) of fly ash. For that KEMA developed a uniform determination method, based on starting points from previous research work. Measurements of the k-value of Dutch fly ashes during the last 7 years show that the mean k-value increased from 0.42 up to 0.75 with a decreasing mean standard deviation from 0.20 to 0.11. This is the result of quality control, blending of coal mixtures and better firing techniques. 1
INTRODUCTION
The use of fly ash in concrete is becoming more and more important, because of economical and technical reasons. Cement (the expensive part of concrete) can partly be replaced by fly ash. The quality of fly ash used in this market segment must be guarded very strictly. In the Netherlands several standards are used for the quality control of fly ash concrete with fly ash as admixture and cement. These standards are: -
NEN 5950 for concrete
-
NEN 3550 for cement
-
CUR 26 and NEN EN 450 for fly ash
For an effective utilisation of fly ash in concrete a quantitive understanding of the efficiency of fly ash is essential. A lot of research has been performed over the last decades. This has lead to an efficiency factor (k-value) in the Dutch standards of 0.2 for fly ash in combination with CEM I cement. KEMA has performed k-value measurements of Dutch fly ash over the last eight years. Review of research work from other laboratories showed that it is necessary to develop a uniform determination method for measuring the k-value. In this present paper an evaluation of k-value measurements and method is presented.
280 FLY ASH IN CONCRETE
In the Netherlands it is nowadays common to use fly ash as mineral admixture in concrete. Therefore the properties of fly ash must satisfy the requirements in EN 450. Concrete must meet the requirements in accordance with NEN 5950. To fly ash in combination with CEM I cement (in all environmental classes) and CEM III (in environmental class 1) an efficiency factor (k-value) of 0.2 may be adjudged in respite of the minimal cement content. The total binder content (cement plus fly ash) is determined with the formula (c+k.f) where c=cement content, k=k-value, f=fly ash content. Evaluation of research work of investigators in other countries showed that in some cases higher k-values are adopted. In the German standards k-values up to 0.7 are allowed. A very important aspect in the evaluated research work showed to be the determination method of the k-value. There is no uniform determination method. Because of the influence of many parameters like water to cement ratio and concrete age, it is necessary to develop a uniform method. PRINCIPLE OF THE K-VALUE
The cementing efficiency (k-value) of fly ash is defined as that portion of fly ash that can be used as cement without effecting the properties of the concrete in relation to concrete that does not contain fly ash. In common the compressive strength is used as a criterion for the determination of the k-value. However, also the durability properties like carbonation or permeability can be used as criterion. In figure 1 the principle of the k-value calculation is given. The k-value is calculated as follows: o~,p= ~/(l+k.q))
k=(e)/o),~-1).1/(p
where: co =
water/cement ratio fly ash concrete
e),~=
water/cement ratio of concrete without fly ash
(p =
ratio fly ash/cement
281 The relation between water/cement to ratio and compressive strength of concrete has been described many times in the literature. Abrams, Ferret, Walz, Weber and Wesche have established an empirical relation between compressive strength, cement standard strength, concrete age, water to cement ratio and fly ash to cement ratio. The calculation method for determining the k-value, based on the empirical relationship, is very complicated because of seven parameters in the equation. KEMA developed a simplified calculation method with two parameters based on the linear relationship between the compressive strength and reciproke water/cement ratio dependent on the cement type, water/cement ratio, concrete age and binder content (figure 3). The formula for this relationship is for reference concrete without fly ash: f'c = A+B/%p where f'c =
the compressive strength and c% is the water/cement of ratio of the reference.
and for fly ash concrete: f'c = P+Q/~ where f'c =
is the compressive strength and co is the water/cement of ratio fly ash concrete.
A, B, P and Q are parameters which can be calculated from data of reference and fly ash concrete mixtures. FACTORS INFLUENCING THE K-VALUE The efficiency factor (k-value) of fly ash is influenced by numerous variables in terms of the characteristics of fly ash and cement as well as the parameters influencing the concrete mix design itself. The influence on the cementing efficiency of fly ash depends on the following characteristics: For fly ash:
-
physical properties like particle shape, size and distribution
-
chemical properties like composition, glass content and pozzolanity.
F o r cement: -
physical properties like particle size and distribution
-
chemical properties like composition, slag content and alkali content
282 For concrete mix: -
total binder content
-
fly ash content
-
curing method and time
-
temperature
-
water to cement ratio.
Because of the influence of the fly ash type as well as the cement type it is not possible to define an efficiency factor (k-value) for only one fly ash type. The k-value of a certain fly ash type should be determined in combination with a cement type.
Binder content Minor compressive strength variations have much influence on the k-value. The compressive strength decreases when the cement content increases, because cement stone has a lower elasticy modulus than aggregate. Therefore, it is important to define the total content (c+f). In figure 2 the compressive strength versus the total binder content (c+f) for a constant water to cement ratio is given for fly ash concrete and concrete without fly ash (reference). The difference between the compressive strength of fly ash concrete (260 kg/m 3 - point A) and the reference concrete (360 kg/m 3 - point D) is smaller than between point B (fly ash concrete 360 kg/m 3) and point C (reference 260 kg/m3). Therefore it is very important to select a reference and fly ash concrete mixture with constant total binder content.
Curing method and age The consequence of use of fly-ash in concrete could be an increase in sensibility for curing, because of the delayed puzzolanic activity of fly ash. Research showed that only at the first three days, wet curing was necessary. The influence of the puzzolanic activity of fly ash in combination with CEM I cement is evident. In combination with CEM III cement this puzzolanic activity is less. The conclusions that the k-value increases with increasing age of the concrete.
Temperature The influence of a lower temperature (10 ~
on the k-value is negligible.
283 Fly ash content A fly ash content in the range f/c = 0.18 until 0.33 has a negligible influence on the k-value for CEM I and CEM III cements.
Water to cement ratio The k-value is highly dependent on the water/cement ratio of the fly ash concrete. If the water/cement ratio increases, the k-value decreases. This is caused by the delayed puzzolanic activity of the fly ash. In table 1 and 2 and figure 3 the linear relationship between water/cement ratio and compressive strength are given for a reference and fly ash concrete. According to the standards a maximum water/cement ratio of 0.65 is allowed. This agrees with a water/cement ratio for fly ash concrete of 0.72 and a water/cement + fly ash ratio of 0.58. The difference in regression coefficients between reference and fly ash concrete causes the dependence of the k-value on the water/cement ratio. If the starting point for a k-value determination is a "worst case" approach, then kvalues should be determined at water/cement ratios 0.72 (water/binder ratio 0.58) for the fly ash concrete. DEFINITION OF STARTING POINTS OF THE K-VALUE DETERMINATION The performed research on the influence factors of the k-value has lead to formulating starting points for a uniform determination method. This method has to be performed for each fly ash/cement combination. K-values should always be determined at prescribed water/cement ratios. With a general k-value of a fly ash type in combination with a given cement type is meant a k-value which has been determined according to the following rules: -
water/cement + fly ash ratios for fly ash from 0,60 until 0,45
-
constant cement + fly ash content of reference and fly ash concrete
-
a curing time of 28 days; humidity > 95%.
Because of the effect of sensitivity of the k-value for slight differences in compressive strength, it is important to determine every measuring point accurately (three times).
284 K-VALUES OF DUTCH FLY ASHES KEMA performs the quality control of Dutch fly ashes by order of the Dutch Fly ash Corporation since 1988. Since 1989 also measurements of the k-value of fly ashes in combination with CEM I 32.5R are carried out at water/cement + fly ash ratio 0.60. In figure 4, k-values of Dutch fly ashes from several power plants after 28 days curing are given. As can be seen in figure 4 the efficiency factor k of Dutch fly ashes has increased since 1995 with about 35% to a value of 0.75. Also the standard deviation decreased from 0.20 to 0.11. This is the result of quality control of the Dutch fly ashes, the use of coal mixtures (blends) and better firing techniques.
7
CONCLUSIONS
The k-value of Dutch fly ashes, measured in combination with CEM I cement, has increased over the years. At this moment the k-value has a mean value of 0.75 with a standard deviation of 0.11 at a water/cement + fly ash ratio of 0.60. This is the result of quality control, the use of blend coal mixtures and better firing techniques. A review of research work performed by German and Dutch researchers showed that a uniform determination is needed. KEMA developed a determination method based on previous research being: -
the k-value applies to a given fly ash type in combination with a given cement type
-
the k-value is dependent on the water/cement ratio
-
the k-value should be measured at a constant binder content for fly ash and reference concrete
-
the k-value should be measured at a curing time of 28 days and a humidity of > 95%.
REFERENCES
IBAC, 1987. Untersuchungen an Mbrteln und Betonen mit Steinkohlenflugasche fQr eine erweitere Anrechenbarkeit der Steinkohlenflugasche. INTRON, 1989. De k-waarde van vliegas.
285 KEMA, 1988. K-waarde van vliegas in beton. KEMA, 1992a. Spreiding van de k-waarde van vliegas in beton; invloed van partij cement en monstergrootte. KEMA, 1992b. Schatting van een algemene k-waarde van vliegas in beton. KEMA, 1992c. Effect van de analysemethode op de k-waarde van vliegas in beton. KEMA, 1992d. K-waarde van gezeefde vliegassen. KEMA, 1994. Model voor het effect van de granulometrie van vliegas op de eigenschappen van mortel en beton. KEMA, 1996. Vergelijking van de berekeningsmethoden van de k-waarde van vliegas. NEVILLE. Properties of concrete.
286
Appendix A page 1
te" L-
.> t,..
E O O
co
co water/cement ratio
Figure 1
Principle of the k-value
(W/C+F) constant +
f/c=0.25
A
f/c=O
45 E E zr 40
-
C
"o
o0
D
t-
e
A
.>
.=3o Q.
E
8
B
25 240
270
I
I
I
3OO
33O
36O
binder content (c+f) kg/m3 Figure 2
Influence binder content concrete
3,9O
287 Appendix B page 1 K-VALUE DETERMINATION Table 1
Compressive strength reference at different water/binder ratios
water/binder ratios
compressive strength N/mm 2
WBF 0.40
59.8 - 59.7 - 61.2
WBF 0.50
46.1 - 43.8 - 45.6
WBF 0.55
3 7 . 2 - 36.3
WBF 0.60
3 2 . 9 - 3 1 . 7 - 32.4
Table 2
Compressive strength fly ash concrete at different water/binder ratios WBR 0.45
WBR 0.48 - 0.49
WBR 0.65
va-1
45.1
40.4
21.7
va-2
44.1
41.4
18.8
va-3
44.7
40.1
19.3
va-4
44.2
41.3
18.7
fly ash sample
288
28 days curing [!i!i!!i!!!!i!i!i!!i!i~i!ilil plant 1
plant 2
mean value
1.00 f 0.90 0.80 0.70 r
0.60
i
,> 0.50 0.40 0.30 0.20 0.10 0.00
1989
1990
1992
1993
1995
year Figure 4
K-values Dutch fly ashes
+
reference
+
fly-ash
concrete
70 tM 60 E E z 50 -
/
c-
E:
|
40
-
(/)
> 30 -
,m (/) (/) (I)
~. 20
-
I
F
E
o o lOo o.oo
I
0.50
,/" /
I
1.00
I I I I
II
1.50 WCR
Figure 3
K-value of fly ash 28 days curing
I
I
2.00
2.50
3.00
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
289
UPGRADING AND QUALITY IMPROVEMENT OF PFA
H.A.W. Cornelissen KEMA P.O. Box 9035 6800 ET Arnhem, the Netherlands
ABSTRACT An example of upgrading and quality improvement is micronization of fly ash (PFA). The usual fly ash particle size range is up to 200 micrometer. For the improvement of concrete properties, however, fine fly ash particles in the range up to 10 micrometer are preferable. Micronized fly ash is produced with a mean diameter of less than 5 micron. This very fine powder is an excellent high performance type II filler for concrete. 1
INTRODUCTION
In the Netherlands about 30% of the production of electricity is realised by coal-firing, which results in a collection of about 900.000 tons/year fly ash in the electrostatic precipitators. Nowadays, fly ash is fully accepted as a raw material. It is used for the production of cement (62%), light weight aggregate (20%), concrete, asphalt and other applications (18%). With respect to by-products management the approach is more and more focused on the reduction of costs and further economical optimization. The value of a product like fly ash is mainly determined by the value of the products that will be partly substituted, or by new attractive possibilities of the end products. If for instance cheap regular sand is replaced by fly ash, its value is relatively low. If fly ash can be used as cement, the value is higher. If, moreover, the resulting concrete properties are improved, its value will raise further. In general, quality control measures and upgrading will turn by-products into valuable resources as schematically indicated in figure 1. To improve the quality of fly ash, methods for benification are being introduced. The methods are focussed on various fly ash parameters like grain size and carboncontent (1).
290
A main reason for upgrading fly ash is to enhance its added value, which makes it more profitable. For the use in concrete, the particle size of the fly ash is very important. Especially fine particles are needed because they improve the packing of the combination of gravel, sand and cement particles. This results in a denser concrete, which is consequently stronger and more durable.
value valuable
resource resource
__ upgrading
w QC,
(upgrading)
by-product 0~__ waste
Figure 1
/
~.~ ///
/
/
utilization
_ ~t~:~c~li~on
storage
--D
niches
D
time
.......
Trends in by-products marketing
HIGH PERFORMANCE CONCRETE The use of high performance concretes (high strength and high durability) is becoming more and more common. At the same time there is a tendency towards concrete mixtures which need only minor compaction energy for optimal filling of the formwork. These concretes have a high slump and must therefore be very stable in order to assure homogeneity (see figure 2). For the proportioning of these concrete mixtures, special additives and ultra fine fillers like silica fume (particle size 0.1-1.0 microns), are needed. However, the availability of these ultra fine fillers is very limited and as a consequence the price is high. Because of the developments indicated, it is expected that the need for ultra fine fillers will strongly increase.
291 relative compressive strength 1.1
"
mortar f/c = 0.33
1.00.9 0.8
""--------1-- 41---- 9 0 d
0.7
_
A
"
-
28 d
~
0.6
2d 0.5 10
20 D-50 (10 -3 mm)
30
40
Figure 2
Concrete strength versus fly ash fineness (D-50) (6)
3
FLY ASH FINENESS
It is well known that fly ash fineness is a major parameter for its effect in concrete. The fine particles in the grain size distribution have a relatively high contribution to the strength development of concrete. This was shown in various research projects (2, 3, 4). Also by KEMA this effect was demonstrated (5). A typical result is shown in figure 3. Fly ashes were classified and mixed to given D-50 values (being the mean particle size of the size distribution). It can be seen that both mortal and concrete strengths increase, if finer fly ashes are added. The effect is stronger for concrete and is more pronounced for the finer particle size range. The usual fly ash particle size range is up to 200 micrometer. For the improvement of concrete properties, however, fine fly ash particles are preferable in the range up to 10 micrometer. These fine fraction can be separated from the bulk amount fly ash, or the coarser fly ash can be processed to finer particle sizes. Separation is cheaper but less effective because the amount of fine particles in the original fly ash is very limited.
292 strength (N/mm~) I
veryilstrong
very~durable I I I I I
60
normal 0
Figure 3
very liquid and stable
workability
Trends in concrete properties CLASSIFICATION OF FLY ASH
Two low-NO x fly ashes with lower and higher carbon content (indicated as B and M) produced by Dutch power stations were selected for the tests. The ashes were processed in an air classifier and, if necessary, mixed to obtain certain gradings. The particle size distribution was measured by Malvern 2600 C analyzer and the gradings were qualified in terms of characteristics of the size distribution, i.e. D10, D50 and D90, but also by the grading modulus G (4). G = (6/((1/dl) -
(lld2))lln(d21dl)
In the formula dl and d2 represent the diameters of the smallest and the largest size particles of a group between two successive sieves. Between these sieves the size distribution is assumed to vary linearly to a log scale. Table 1 gives the corresponing values. Concrete cubes (150x150x150 mm) were cast and stored in the fog room (20 ~ and RH > 95%). The cement content of the reference mixes was 320 kg/m 3, which was substituted with 20% (by weight) fly ash in the case of the fly ash mixtures.
293 For all mixtures normal hardening Portland cement was used. The maximum aggregate size was 31.5 mm, while the slump of the fresh concretes was adjusted at 70 mm (plus or minus 10 mm). The amount of water needed for a given workability (the slump of the concrete cone = 70 mm) was measured for all concrete mixtures. Then cube compressive strength was determined at ages of 7 and 28 days. The results are presented in table 2. The results prove that fly ash fineness is a major factor for concrete properties. MICRONIZATION OF FLY ASH 5.1
Processing
It was decided to process fly ash to particles in the one micrometer range, because this will result in an important market for fly ash as a high valuable resource. Besides by air classification fly ash fineness was increased by grinding to particle sizes down to between about 5 and 10 micron (6). The ultra-fine range was also reached in Japan, by vaporization at about 2400 ~
and condensation of fly ash (7). The present
project, investigates whether fly ash can be micronized economically to less than 5 pm and to see if the performance of this product in mortar and concrete is satisfactory.
5.2
Materials
The effects of the various types of fillers were tested with normal hardening Portland cement (PC-A). A typical fly ash was used (57.0% SiO2, 26.4% AI203, 4.4% Fe203, 4.2% C and 1.8% CaO). The fine fraction was collected in the bag-filter of an air classifier, whereas the ultra-fine fraction was obtained by grinding. It was found that the mean particle size being 21.6 l~m for the input fly ash was reduced to 9.9 l~m for the air classified fly ash and to 1.6 l~m for the ground one. A picture of micronized fly ash is given in figure 4. Particle size distribution curves are given in figure 5.
294
Figure 4
SEM-picture of micronized fly ash
100
undersize (%)
/
/
4
5O
J
/
/
/
/ /'
r
/
/
I
3 /
/
/
// //
/
/ / ,/
fly-ash 1 !"] Input
/
2 r-I classified 3 r-] ground 4 I-'1 silica fume
o. _ _ _ _ ~ . . . ~ 0.1
Figure 5
.~
1
10
100
Jill
1000
Particle size distributions of the classified and ground fly ashes in comparison with the input fly ash (percentage passing versus sieve opening in micrometers)
295 The silica fume involved, had an SiO2 content of 92% (m/m) and a mean particle size of 0.12 ~m. In the tests, a 50% to 50% (m/m) combination of silica fume and micronized fly ash was also applied. 5.3
Laboratory scale tests
In the concrete compositions, 360 kg/m 3 PC-A cement was used; the maximum grain size of the river gravel was 31.5 mm. Because a major objective was to realise highly fluid mixtures, the fresh concrete slumps were 230 plus-minus 40 mm. This resulted in a water to cement ratio of 0.32 for the mixtures containing fillers and 0.35 for the reference mixture (no filler added). An overview is given in table 3. If needed, chemicals called super plasticizers were added to improve the workability of the fresh concrete. The amounts of filler added were 5%, 10% and 15% (m/m cement). After 3, 7, 28 and 91 days curing, the compressive strength was determined (see table 4). The results indicate that the strength values are significantly higher for the mixtures with higher filler content, especially in the case of filler types micronized fly ash and silica fume. Note: In table 3 and 4 the reference mixtures are indicated as CREF 1, 2 and 3. The various types of fillers are AC (= air classified PFA), AG (= ground PFA), SF (= silica fume) and SA (= 50% AG plus 50% SF). The numbers 5, 10 and 15 indicate the amount of addition of filler (m/m cement). 5.4
Full-scale test
Based on the results of the laboratory tests two mixes were designed for the full-scale tests, being precast L-shaped elements (100 x 150 x 300 cm3). Special attention was given to the workability in order to realise a mixture which can be placed with minor compaction energy. So the PC-A cement content was raised to 410 kg/m 3 and about 2% (m/m cement) melamine-sulphonate super plasticizer was added. The maximum grain size in the mixes was 16 mm. In mixture 1 the water to cement ratio was 0.36 and in mixture 2, 0.38. The addition of micronized fly ash was 12% (m/m cement) in both cases (see figure 6).
296 The moulds were filled in one batch of 1.2 m 3 fresh concrete. After that compaction was needed for 2 minutes. One day later the elements were demoulded. Mixture 1 showed airencapsulations at the surface, while the element made from mixture 2 showed an excellent surface texture. At an age of 28 days, concrete compressive strength was determined from drilled cores. It was found that after one day hardening the cube strength was already about 55 N/mm 2. The 91 days strength values proved to be about 95 N/mm 2 for both mixes. .:
...
.....:.
9
..
.
.
.:... .
.
.
.
.
.
9
ill
Figure 6
High performance concrete element with micronized fly ash
6
CONCLUSIONS
It is a challenge to use the potentials of the by-products of coal combustion. Because of environmental, economical and technical reasons the present worldwide utilization rate of 35% will strongly be stimulated. In many countries like the Netherlands these by-products are fully accepted, which results in a 100% utilization, mainly in the building materials industry. By adequate measures like quality control and up-grading, fly ashes prove to be excellent raw materials for the building industry. Fly ash fineness proves to be an important parameter for the quality of the products.
297
By appropriate grinding it is possible to micronize fly ash to sizes under 5 l~m. By micronizing all fly ash can be processed, while in the case of air-classification the output of fine material is very limited (< 10%). The effect of addition of these ground fly ashes on properties of concrete was determined and compared to the effects of air classified fly ash, silica fume and combinations of these two types of fillers. It was found that in concrete, the fluidity was positively effected by these types of fillers. So, the irregular shape has no significant effect. Also high strength values were reached. Mixes with micronized fly ash behaved well during full-scale tests. It can be concluded that micronized fly ashes are excellent high performance fillers for concretes.
REFERENCES "Innovation for a sustainable future". Proceedings of the 12th International symposium on Management & Use of Coal Combustion Byproducts (CCB's). ACAA, January 1997. G. Wooley. "Effects of fineness and loss on ignition on concrete performance". Report of the Association of Quality PFA suppliers, UK, 1989. P. Schiessl and R. H~rdtl. "The change of mortar properties as result of fly ash processing." IBAC Mitteilungen, pp. 247-294, 1989. B.P. Hughes and AI-Ani. "PFA fineness and its use in concrete". Magazine of concrete research, no. 147, pp. 99-106, 1989. H.A.W. Cornelissen, C.H. Gast. "Upgrading of fly ash for Utilization in Concrete". Fourth Canmet/ACI Conference on Fly Ash, Silica Fume, Slag and Natural Pozzolans in Concrete, Istanbul, 1993. R. H~irdtl, 1991. Effectiveness of Fly Ash Processing Methods in Improving Concrete Quality. In: Waste Materials in Construction, ISBN 0-444-89089-0, pp. 399-406. Y. Matsufuji, et al., 1993. Study on Properties of Concrete with Ultra fine Particles Produced from Fly Ash. Fly Ash, Silica Fume, Slag and Natural Pozzolans in Concrete, Proceedings International Conference, ACI SP 132, vol. 1, Istanbul pp. 351-365.
298 Table 1
sample code
Size distribution characteristics and LOI
LOI*
D10
D50
D80
G
(~.m)
(~m)
(~.m)
(l/mm)
(%)
B0
7.4
39.2
133.5
225
4.56
B1
4.5
23.0
75.3
456
4.50
B2
6.6
17.4
46.1
462
4.14
B3
3.2
9.5
32.2
850
5.70
M0
6.5
24.2
96.8
299
8.31
M1
4.4
17.7
53.5
563
8.04
M2
4.2
15.7
45.3
600
7.46
M3
2.9
8.7
36.3
919
7.94
loss on ignition at 815 ~ for 10 minutes Table 2
sample code
Concrete test results
water content
(dm31m 3)
compressive strength (N/mm 2) 7 days
28 days
reference
163
30.3
38.3
B0
163
23.0
31.4
B1
160
25.2
33.3
B2
161
23.9
32.9
B3
156
28.2
38.7
M0
162
24.5
33.3
M1
160
27.0
36.8
M2
158
27.3
36.0
M3
157
27.9
38.7
299 Concrete compositions and workability data
Table 3
Sample
fly ash
SF
code"
(kg/m 3)
(kg/m 3)
w/c
SP
Slump
Spread
Spread
(%)**
(mm)
(static)
(jolting)
(mm)
(mm)
CREF1
0
0.32
2.75
170
320
420
CREF2
0
0.35
2.5
220
470
540
CAC05
18
0
0.32
2.5
190
330
420
CAC10
36
0
0.32
2.5
230
430
520
CAC15
54
0
0.32
2.5
250
500
570
CAG05
18
0
0.32
2.5
230
420
520
CAG10
36
0
0.32
2.5
240
540
600
CAG15
54
0
0.32
2.5
270
610
>700
CSF05
18
0.32
2.5
220
400
500
CSF10
36
0.32
2.5
220
350
470
CSF15
54
0.32
2.5
190
320
440
C = concrete; AC = type of filler; 0.5 = % filler (m/m cement) superplasticizer as weight percentage of cement plus 0.2 filler
300 Table 4
Properties of hardened concrete containing the indicated types and amounts of fillers
Sample code
filler type
filler
(%)
Compressive strength (MPa) 3 days
7 days
28 days
91 days
CREF1
no
56.0
68.7
81.8
90.6
CREF2
no
38.8
49.8
64.2
69.3
CAC05
AC
5
56.3
68.2
83.6
92.3
CAC10
AC
10
52.3
64.9
81.6
93.7
CAC15
AC
15
49.7
62.3
79.8
92.4
CAG05
AG
5
53.0
65.3
80.0
85.9
CAG10
AG
10
52.6
66.2
86.2
96.0
CAG15
AG
15
55.0
69.6
95.8
105.9
CSF05
SF
5
55.1
69.4
94.0
95.3
CSF10
SF
10
58.3
70.6
96.6
103.9
CSF15
SF
15
57.9
74.3
99.3
104.8
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
THE EFFECT OF THE DUTCH BUILDING MATERIALS DECREE ON THE BY-PRODUCTS FROM COAL-FIRED POWER STATIONS
M.P. van der Poel N.V. Sep- Dutch Electricity Generating Board Arnhem, the Netherlands
Summary The Dutch government published the Building Materials Decree in November 199~ This new legislation sets the boundaries for imissions of organic and inorganic components from building products into the soil. The immision of building material,, according to the Building Materials Decree is calculated from the results of regulatq leaching test on these building materials. This paper will highlight the effect of the Building Materials Decree on the by-products from coal-fired power stations.
1.
The Building Materials Decree
The Building Materials Decree, referred to as the Building Materials (Soil and surf~ Waters Protection) Decree is published in November 1995 and applies to the use q building materials in a work on or in the soil or in surface water. Building materials in the meaning of this decree are characterised as granular (stonelike) building materials used outside. If the total contents of silicon, calcium c aluminium (with the exeption of metallic aluminium) in a building material together amount to more than 10% m/m of the total building material, it is a building materi~ in the meaning of this decree. The aim of the decree is to set the environmental conditions from the point of view soil and surface waters protection for the use of primary and secondary materials, or in the terrestrial soil, in the surface water and on or in the soil beneath surface water. Although the decree primarily lays down rules with a view to protecting the quality soil and surface waters, it intends also to contribute towards achieving other
302
environmental policy objectives, such as reuse. Because of the continuity of reuse for some building materials special categories are incorporated. For the unmoulded use of fly ash from power plants no special category is being incorporated because the continuity of the present applications is not in issue.
2.1
By-products from coal-fired power stations
Three energy sources are mainly used for the large-scale generating of electricity in the Netherlands. In 1995 these are coal (45%), gas (47%) and nuclear fuels (8%). Most of the carbon from coal is converted to heat during combustion between 1300 and 1700 ~
depending on the type of burner. The inorganic fraction in the coal is
left as ash, produced by the fusion of clay minerals. Water loss combined with rapid cooling of the hot flue gas in the power station exhaust system prevents the elements forming crystal structures, and the minerals are turned into a glassstructure. The process is similar to that which takes place during a volcanic eruption. Fly ash therefore is about 70% glass. Fly ash and bottom ash from coal-fired power stations and products derived from these, are stonelike secondary building materials in the meaning of the Building Materials Decree. For the power industry fly ash and bottom ash from pulverised-coal power stations are the most relevant by-products in relation to this decree.
2.2
Fly ash
The hot fluegas is fed through electrostatic filters which remove remove 99% of the fly ash. In 1995 the Dutch power stations produced 878,000 tons of pulverised coal fly ash. The produced fly ash can all be utilised. The greatest part goes to the cement industry and concrete industry to be used as: -
partial substitute for cement
-
filler used in the manufacture of Portland fly ash cement raw material for clinker production filler and binding agent in concrete
-
substitute for cement in cellular concrete.
303
Fly ash is also utilised as: -
filler in asphalt mixes
-
raw material for the production of artificial gravel.
All the applications of fly ash in the Netherlands are so called moulded applications which means that leaching is determined by diffusion. 2.3 B o t t o m a s h
Bottom ash is produced by the melting and sintering particles of ash in the boiler. Because of the gravity, bottom ash falls down into collecting tanks with water underneath the boiler. Bottom ash is also a product that can be utilised very well. Bottom ash is mainly utilised in the application of: -
road construction
-
concrete block manufacture.
3.
Immission standards
The starting-point for determining a maximum permissible load on the soil is that the multifunctionality of the soil must be guaranteed. As it is not realistic to employ building materials without some burden on the soil, a marginal soil-load is accepted on the assumption that this does justice to the starting-point of maintaining multifunctionality. Marginal soil-load has numerically interpreted as: a load on the soil resulting from leaching from the building material which mathematically results in an increase in the solid phase of the soil of no more than 1% of the contents of pollutants in relation to the long-term targets for soil in 100 years, averaged over one metre of standard soil deemed to be homogenous.
It is assumed that in general this marginal soil-load will also protect the groundwater in a sufficient way. On the basis of the marginal soil-load concept and the average background ieveU in the Netherlands, the maximum permissible immission standards for inorganic
304
substances were determined. When determining these immission standards, it was necessary for a number of substances to rise the immission level to allow the continued use of building materials. That means that in general the immission standards are low, especially for mobile elements. For this mason it is not allowed to use unmoulded fly ash, for instance as a layer in road constructions, any longer. This particularly because of the initial leaching of Molybdeen. In appendix A and B leaching characteristics of fly ash and bottom ash are given. 4.
Classification in categories
Building materials are divided into category 1 and category 2 building materials. - category 1: the building materials referred to in this decree as category 1 may be used without taking isolation measures. However, control measures should be taken to prevent the building materials mingling with the soil and to allow removal. - category 2: building materials with a greater immission into the soil than the
marginal load of 1% may only be used as category 2 building material if the immision can be reduced to below the maximum immision standards belonging to this category. Permanently controlled isolation measures are necessary for materials belonging to category 2. Isolation measures are understood to be measures which virtually rule out any contact of that building material with rainwater or groundwater when that material is used. Regulations for these isolation measures and the control and monitoring measures appurtenant to these measures are stated in the decree and elaborated in a ministerial decision. 5.
Determining of standards
The composition standards for inorganic substances and the immission standards for inorganic substances in category 1 and in category 2 building materials must be determined by one of the laboratories designated by the Environment Minister and the Minister for Transport, Public Works an Water Management. With the aid of difficulty functions and correction factors immission into the soil as a result of emission from a building material will be calculated - after leachability from that building material has been determined in the laboratory in accordance with the
305
draft Dutch standard NEN 7340. Defining the functions and correction factors has been a very difficult process. This because the long-term behaviour over a period of 100 years by extrapolation from the results of laboratory tests is a rather theoretical approach. In this framework it can be told that studies by the Dutch Electricity Generating Companies have stated that the behaviour of fly ash in the laboratory is different from the behaviour of fly ash under field conditions. 6.
Directions for use
One of the departure points of the decree is that it is the task of the party using building materials in a work to ensure that the building materials in question do not become mixed with the soil or aquatic sediment during the life of the work, and that they are indeed removed when the work in which the building materials have been used is removed. These obligations apply in principle to all building materials - with the exeption of clean earth - whose use is permitted under the decree, irrespective of where they are used. In addition to the duty of removal a number of other directions for use formulated as general rules have been incorporated in the decree. In figure 1 the directions for use for the various categories of building materials when used on or in the soil are given.
306 Figure 1 Soil Protection Act (Landfill) Decree
Immission1~
Special category Slag's from waste incineration plants
Category 2 building materials
U2 >
-isolation measures: * 0,5 m above mean highest groundwater level * isolation (capped from above) - control measures in connection with duty of removal: * minimum quantity 10,000 tonnes (1,000 tonnes for foundation layers in road construction) - other control and monitoring measures (management and maintenance)
Category 1 building materials
Special
Building
- control measures in connection with duty of removal
category
Materials
(generally normal management and maintenance)
tar-holding
Ul >
Decree not
asphalt
enforced
aggregate s org
G
Composition G
:Composition value for standard soil
Sorg
:Limit-value for organic components
Ul
:Leaching limit-value for category 1 building materials
U2
:Leaching limit-value for category 2 building materials
7.
Approvals and other forms of proof
On 1 J a n u a r y 1988 the decree will c o m e into full force for all building materials. The user of a building material is than required to have available information on the
307
composition of that building material and the expected immisions as a result of that material being used. For nearly every building material whose application requires a report in advance the information on composition and immission must accompany that report. These materials are category 1 earth, all category 2 building materials and building materials containing tar-holding asphalt aggregate. For the remaining category 1 building materials the competent authority may request the mentioned information from the user, up to five years after the building material has been used. In effect, this duty means that the user should demonstrate or be able to demonstrate that he is employing a category 1 or a category 2 building material. The user can provide the proof concerning the quality of the building material by handing over an approval recognised by the Minister of Housing, Spatial Planning and the Environment and the Minister of Transport, Public Works and Water Management on the basis of certification of the product. As a recognised approval the product certificate and the attestation can be distinguished. The product certificate means a certificate stating a product conforms to certain product specifications. However this certificate does not mean that the building material will satisfy the requirements of the Building Materials Decree in every instance and every application, as this also requires a test of the specific application to the immission requirements. An attestation, in relation to the Building Materials Decree, indicates that a building material, which has a certain composition and leaching properties as described in the product certificate, if employed as stated in the attestation, satisfies the composition and immission requirements of the decree. The attestation will therefore in effect make a statement about a specific building material in a certain use in relation to the category in which that building material and use fall. The user may also attempt by other means - such as a manufacturer's declaration that the building material satisfies the composition and immission requirements of the decree. To do so, the decree imposes on the user the duty to engage an accredited
308
of equivalent foreign laboratory to access composition and immission. These assessment must be carried out in accordance with the rules set in or by virtue of this decree. A building Materials Decree Approvals Manual, in which guidelines and test criteria are laid down, is currently in preparation.
8.
Consequences for fly ash and bottom ash
With the publication of the Building Materials Decree for the power industry development toward achieving environmental certification of construction materials in relation to fly ash and bottom ash has reached an important phase. Because of the Building materials Decree and the leachability of fly ash, this building product has to be used in moulded applications. Leaching of moulded applications is determined by diffusion. For bottom ash selenium (Se) is a critical element. In thin layers however the application of not certificated bottom ash as a category 1 building material is almost 100%. Certificated bottom ash can be used in its entirety as a category 1 building material. In 1996 the first bottom ash under certificate (attestation) is supplied.
References [1]
Dutch decree of 23 November 1995, containing regulations on the use of
building materials {Building Materials (Soil and Surface Waters Protection) Decree on or in the soil or in surface waters}. [2]
Milieugienische kwaliteit van primaire en secondaire bouwmaterialen in relatie
tot hergebruik en bodem- en oppervlaktewaterenbescherming, RIVM-rapp.no. 771402006, RIZA-rapp. no. 93.042 [dec.1993]. [3]
Annual report of 1995 from NV GKENliegasunie.
APPENDIX A
man and ann'ronmsnt
EC-vliegas
Building material:
NV8052 wkl leachlng characteristics
~dentificat~on number: 17 D r
m
composition
US=10 columntest in rngkg
adjusted value granular materials .bmd
A.
U1
U2
S1
088
700
37500
N
mun
n
o m
.d(n I ) m n m m m u m m n,Ul o
s
~ o m
osa7 3200
h
550
5
21€6
1278
Om2
cd
o m
007
1000
17
o m
o m
o m
0011
C4
042
2W
25000
2
0018
0011
0010
0025
Ct
130
I200
129300
17
1728
1393
0160
4673
CU
072
350
37500
7
0080
0101
0007
0-5
He
o m
008
500
I
o
o m
o m
MO
028
091
12500
16
5810
5143
0350
15280
NI
110
3 m
~ ~ 0 0 0 10
0064
0089
0010
0878
58m75LQC.3
NA: No illormalion.rsllabb. ERR: ~land.rddsvt.lon
2.m
m
n>U2 bp(-1 I
9
kpbdfn 1))
0948
061s
2881
0803
0056
0483
wl*yn d d h
D D I6
I1
0450
0624
1357
0608
'
D
aqua regia in mgkg
APPENDIX B
m n and nrimnmnt
EC-bodemas
Building material: identfiicationnumber:
NV8053.wkl leaching characteristics
170~93
U S 4 0 columntest in mgkg
agustd v a l u ~
composition
granular materials .b""n(
As
Ul
In
31
066
700
37500
~1
s m
mm
7 m m
w
o m
007
iom
042
260
25003
Q
133
I200
Cu
072
350
N 61
mm
8dln.l)
0.108
0.m
54
2.857
41
0 UX
57
12Y)OO 37500
n*mum -hum
m u 1 mln bglmem) -1.5
lop(*qn.l)) 0478
2.398
0.24)
14.800
8
0.334
0.378
0.W4
0.-
0.050
3
118%
0.548
0.W
0.a
0.001
0.700
1
-1.782
0.524
'
81
0.024
0.030
0.001
0 240
.1.830
0.448
'
D
81
0.053
0.099
0.010
0.740
1
-1.550
0.482
'
D
-2.553
0.151
o m
ow
sao
54
0.W3
0.001
0.w
0005
Mo
028
091
12500
58
0.144
0.135
0.010
OUr,
7
-1 078
0.19
NI
110
370
25000
01
0.097
0.131
0.005
1.m
1
-1.M
0.581
Pb
1 m
em
1m00
9 ,
005
043
5000
s.
OM
0.10
Sn
027
240
V
1 e4
9200
125000
Zn
3 w
1d.m
12wm
Ma, 2YJ00
290
4.10
Cl
mOO
660000
CN mmp
007
0 1
CN rdl
001
o m
2500
13W
10000
4WW
F Iol
dd.lhr
0.520
Y
81
w I h
0.002
500 00
ma, 12500
NA: No inlormn(nn ndlmbb. ERR: s1.darddsvlat~n zem
'
' ,'
D
D
aqua regia in mg/kg
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
311
PREDICTION OF ENVIRONMENTAL QUALITY OF BY-PRODUCTS OF COAL-FIRED POWER PLANTS
ELEMENTAL COMPOSITION AND LEACHING Ruud Meij KEMA P.O. BOX 9035 6800 ET Arnhem, the Netherlands
ABSTRACT
In the Netherlands the elemental compositions of the various streams of coal-fired power plants are well recorded. The information of these studies is used to calculate enrichment factors for the trace elements in ash, the vaporization percentage of minor and trace elements in flue gases, the degree of removal of gaseous minor and trace elements from flue gases in flue-gas desulphurization installations and the leaching percentage of the elements in ash. These relative parameters combined with trace-element analyses of the coals are used to predict the concentrations of trace elements in the ash, in the leachate and in the flue gases in the gaseous phase. The model is also valid for co-firing with secondary fuels. 1
INTRODUCTION
In 1995 29% of electricity in the Netherlands was generated using coal. Only imported bituminous coal is fired. Coal is imported from all over the world. Major suppliers are Australia and the USA. Other suppliers are Colombia, South Africa, Indonesia, Poland and China. Today mostly blends are fired. In the Netherlands the only boilers installed are pulverized coal-fired dry bottom types. The flue gases are cleaned by high-efficiency coldside electrostatic precipitators (ESPs) and by flue-gas desulphurization (FGD) installations of the lime(stone)/gypsum process. Table 1 shows some typical values for a 600 MWe coal-fired power plant in the Netherlands. The by-products are bottom ash, collected ash, gypsum and sludge of the waste-water treatment plant. The collected ash from the electrostatic precipitators (ESPs) is called pulverized fuel ash (PFA) in the UK and fly ash in the USA. In this paper it will be called PFA.
312 The policy in the Netherlands is in principle not to produce waste, but to produce usable residues (for the environmental legislation in the Netherlands see the paper by Van der Poel (Van der Poel, 1997). The electricity generating companies in the Netherlands founded a special firm for the marketing of the coal-firing residues: "de Vliegasunie" (Dutch Fly Ash Corporation). This firm also stimulates research of and experiments with applications. Long-term disposal of coal-firing residues is impossible at present. So far the Dutch Fly Ash Corporation has realized almost 100% utilization of all by-products. For more information on this subject see the paper by Van den Berg (Van den Berg, 1997). Table 1
Averaged mass flows at a modern coal-fired power plant of 600 MWe
description
unit
net capacity
MWe
600
full load hours
h.a-1
6,000
,,
description
%
thermal efficiency
40.5
unit
ratio bottom ash
12/88
bottom ash
ton.a1
15,600
PFA
ton.a1
115,000
energy demand
Mj.a-1
3.2-101
gypsum
ton.a1
41,000
coal demand
ton.a-1
1.2-106
sludge
ton-a1
600
coal, ash content
% (w/w)
coal, caloric value
MJ.kg1
27
coal, sulphur content
% (w/w)
fly dust emission
desulphurization efficiency % collection efficiency ESP
%
MEASUREMENTS
11
ton.a1
60
92
process water FGD m3.h"1
100
99.75
limestonedemand
ton.a1
24,000
0.7
P E R F O R M E D AT C O A L - F I R E D P O W E R P L A N T S IN
THE NETHERLANDS
2.1
Introduction
In this chapter an overview is given of the research in the field of (trace) elements performed at coal-fired power plants in the Netherlands. It concerns complete mass balance studies, studies limited to some streams and leaching studies.
313
2.2
Mass balance studies at power plants in 1980-1992
Following the reintroduction of coal as a fuel for power plants, the environmental consequences for electricity generation have been thoroughly studied; for instance in the Dutch National Coal Research Programme (NOK). A fairly important environmental aspect is trace elements. The concentrations and distributions of trace elements in coal, ash, and in flue-gas ( in the vapour phase) were determined in sixteen mass balance studies in coalfired power plants. The first flue-gas desulphurization (FGD) system was installed in the Netherlands in unit 13 (CG-13) of the Gelderland power plant in 1985. Extensive testing was performed at this unit in the following year (1986) in order to study the fate of (trace) elements in a coal-fired power plant equipped with a wet flue-gas desulphurization facility of the limestone/gypsum type. This aspect was researched in detail (Meij, 1989). An important aspect in these mass balance studies are the relations between the various streams, from which relative parameters could be deduced. 2.3
Studies at some related streams at power plants in 1988 and 1993-1995
Trace elements are emitted into air in solid (fly ash) and gaseous states. The emissions in the solid state are low due to their high degree of removal in ESPs. The emissions in the gaseous phase are relatively more important. In a plant with FGD equipment both emissions are further diminished. Hence, from an environmental point of view, the gaseous emissions require further research. Consequently, in 1988 the removal of gaseous minor and trace elements in FGD plants was studied at all Dutch units equipped with FGD systems. The concentrations of the gaseous trace elements were measured in the flue gases both upstream and downstream of FGD installations together with the concentrations in the feed coal. The relative parameters which could be deduced were the vaporization percentage and the removal in the FGD installation.
314 In the years 1993-1995 26 samples of feed coal with the corresponding pulverized fuel ash were analyzed for their elemental composition. It concerns mostly blends and represents the recent Dutch policy of coal purchase. 2.4
Studies of the leaching behaviour of bottom ash and fly ash in 1991-1995
Leaching behaviour of by-products has been studied at KEMA since 1980. However old data are not useful, because different leaching tests were used. It was KEMA who took the initiative to standardize leaching tests. Nowadays the column test and the availability test are mostly used in the Netherlands. In 1991 and 1995 45 bottom ash samples were studied: elemental composition and leaching behaviour established by the column test. In 1993-1995 26 fly ash samples were studied: elemental composition and leaching behaviour. The elemental composition was fixed after an aqua regia digestion and after a total digestion or INAA (instrumental neutron activation analyses; method without digestion). The leaching behaviour was established by the column test and the availability test. 2.5
Studies of the leaching behaviour of fly ash under field conditions
The leaching behaviour of pulverized fuel ash under field conditions has been studied in large lysimeters (height 3m80 and 0m95) at the KEMA premises since 1993. This project should lead to a better understanding of the leaching process in field condition and therefore a lot of parameters are monitored: water balance, composition of pore water and leachate, including pH, Eh and speciation of As, Cr and Se (Meij, et al., 1994; Van der Hoek et al. 1995).
315
Studies at power plants during co-firing in 1993-1996
2.6
In the years 1993-1996 11 test series were performed at coal-fired power plants (7x) and at a test facility at KEMA (4x) during co-firing secondary fuels such as sewage sludge, paper sludge, wood and petroleum-coke. In these test-series all the relevant streams were monitored and compared with the situation without co-firing including leaching behaviour.
PARAMETERS DEDUCED FROM STUDIES AT COAL-FIRED POWER PLANTS 3.1
Introduction
The studies mentioned in chapter 2 yield typical parameters that provide the relations between the streams concerned. These parameters are independent of the situation at that particular moment and can be used in a general way in models for predicting the composition of the streams and the leaching behaviour of the by-products (Meij, 1994).
Relation between elemental coal composition and elemental
3.2
ash-composition After combustion of the coal, ash remains. In general the ash contains the same elements as were present in the coal, but enriched in the ash by a factor equal to 100/(ash content in %). Three types of ashes are to be considered: -
ash collected on the down side of the boiler and called bottom ash or slag
-
ash collected in flue gases by flue gas control devices, such as electrostatic precipitator (ESP), this type of ash is named pulverized fuel ash (PFA) in the UK and fly ash in the USA
-
ash that escapes the flue-gas control devices and will be emitted through the stack, called fly ash.
In this paper the three types of ash are called bottom ash, PFA and fly ash, respectively.
316 The enrichment in the ash depends on the type of ash and the particular element. The term "relative enrichment" was introduced to properly describe the behaviour observed (Meij et al., 1983). The relative enrichment factor (RE) is defined as:
RE =
(element concentration in ash) 9(% ash content in coal) (element concentration in coal) 100 Classification of elements based on their behaviour during combustion in
Table 2
boiler and ducts with their Relative Enrichment factor (RE) class
bottom ash
PFA
fly ash 1)
behaviour in installation
I
= 1
=1
= 1
not volatile
IIc lib Ila
<0.7 <0.7 <0.7
=1 =1 =1
1.3<..<2 2<..<4 >4
III
<<1
<1
i
I
volatile, but condensation within the installation on the particles very volatile, hardly any condensation
1) emitted fly ash and PFA from last hopper of ESP (finest fraction) Based on the RE factor, elements can be grouped into three classes. The background of the classification is the behaviour of the elements during combustion in the boiler and further behaviour in the ducts, air preheater and ESP. The three classes are given in table 2. Class II is further divided into three subclasses. These subclasses refer to the degree of volatility. For the three types of ash of each element studied the RE factors have been established in studies mentioned in chapter 2. All these RE factors are combined in a database. In table 3 the elements are classified into the three classes as mentioned in table 2.
317
Table 3
Classification of elements based on research performed in the Netherlands
class
I
AI, Ca, Ce, Cs, Eu, Fe, Hf, K, La, Mg, Sc, Sm, Si, Sr, Th and 13
class
IIc lib Ila
Ba, Cr, Mn, Na and Rb Be, Co, Cu, Ni, P, U, V and W As, Cd, Ge, Mo, Pb, Sb, TI and Zn
class
III
B, Br, C, CI, F, Hg, I, N, S and Se
Class I elements are defined as elements that do not vaporize during combustion. Their
concentration in all ash types is the same (see table 2). The RE factor is about one. However, for some elements there is a redistribution among the various ash types, i.e. bottom ash, PFA (collected) and fly ash (in the flue gases downstream of the ESP). Those elements are vaporized in the boiler. Concomitantly with the route of the flue gases through the boiler, ducts, air preheater and ESPs, the temperature decreases from about 1600 ~ to about 120 ~
Depending on the chemical compound, the dew point will be
passed somewhere on this route and condensation will start. Condensation occurs on the surface of the fly ash particles. Also, particles can form through nucleation of vaporized material and growth through coagulation and heterogeneous condensation. The smallest particles have the largest specific areas. Therefore, on a weight basis, the condensing elements are found in greatest concentrations on the smallest particles. All elements that condense within the installation are grouped in class I1. The RE factor of the bottom ash is less than 0.7 because elements originally present in the vapour phase have no chance of condensing on the bottom ash particles. The RE factor of the PFA from the collection tank is about one for elements of class I1" the factor for the smaller particles exceeds 1.3. The smaller particles are found in the last two hoppers of the ESP and in the flue gases downstream of the ESP. Elements that occur in compounds with a low dew point condense only partly within the installation and, in the absence of an FGD plant, they are totally or partly emitted in the vapour phase.
318 They are grouped as class III. Their RE factor is very small (<< 1), especially in the bottom ash and to a lesser extent in the PFA of the collection tank. The RE factor of the smallest fly ash particles, as found in flue gas downstream of the ESP, can be high (see table 2). 33
Relation between elemental coal composition and concentrations of gaseous elements in the flue gases
In chapter 3.2 the relation between the elemental concentrations in coal and the various ashes are discussed. Class III elements are generally for a small part present in ash. The largest part is present in the vapour phase. Just as RE factors, a parameter is introduced for these elements: the vaporization percentage (Meij et al., 1983). Based on research as mentioned in chapter 2, the class III elements can roughly be divided into three groups (Meij, 1994): -
displaying almost complete vaporization (CI, F, I)
-
with a typical vaporization of about 50% (B, Br and Hg)
-
with low vaporization (Se and As = class IIc).
However, since the introduction of FGD, the major part of these elements is removed in the FGD. Another parameter, also based on research as mentioned in chapter 2, is introduced: the degree of removal. The results are (Meij, 1994): -
B, Br, CI and l are removed for >80-90%,
-
F, Hg and Se are removed for >50%.
34
Relation between ash-composition and leaching
The leaching of elements, as mentioned in chapter 2.3, is studied together with the elemental composition of the ash. The parameter, which could be deduced from these studies, is the leaching percentage relative to the composition. This is done for bottom ash with respect to the column test and for the pulverized fuel ash with respect to the column test and the availability test. The results are given in figure 1. The Dutch leaching test and other test are discussed in detail by Van der Sloot (Van der Sloot, 1991) and also by Meij et al. (Meij et al., 1994).
319 Thus the leaching percentage of a great number of ashes are obtained. The mean value can be used to predict the leaching behaviour of future ashes of that order. However the range of the mean leaching percentage is fairly large. It is not possible to predict leaching behaviour in detail with a straightforward parameter. There are more parameters concerned. In the lysimeter project (Meij et al., 1994) more information is obtained about the leaching mechanism. However, profound knowledge of a particular pulverized fuel ash is not sufficient to predict leaching of other pulverized fuel ashes. Nevertheless a prediction of leaching behaviour of an arbitrary pulverized fuel ash could be possible in combination with an evaluation of the data in the file "Database Leaching". It has to be said that an accurate prediction is not always needed. Only when it is really necessary, for instance when the concentration of a particular element comes close to a standard. In that case an advanced model for predicting the concentrations of that particular element is needed.
320
)
Group 3 10<.< 100%
o,ou0~/
, ~~ ,~, ~
/)
o, ~
~
V
~:~.co
c,
0.1<..<10%
Increasing leaching,
Group 1 <0.1%
pulverized fuel ash
pulverizef fuel ash
Columntest L/S-IO
Columntest L/S-IO
Availability test L/S-IO0
KEMA database leaching
KEMA database leaching
KEMA database leaching
bottom ash
Figure 1
Leaching behaviour of elements in bottom ash and pulverized fuel ash as produced in Dutch coal-fired power plants in different leaching tests. The relative leaching with respect to the composiotion of the ash grouped into three classes. The figures in brackets are derived from detection limits
4
DATABASES
4.1
Introduction
The measurements, as mentioned in chapter 2, and the deduced relative parameters, as mentioned in chapter 2, are recorded in databases. The databases are (see also figure 2: a database Coal: elemental composition, ash content and other quantities b database Bottom ash: elemental composition and RE factors c database Pulverized Fuel Ash (PFA): elemental composition and RE factors d database fly ash (emitted into air): elemental composition and RE factors
321 e database Gaseous Elements (emitted into air): concentrations in the flue gases, vaporization percentages and degree of removal in FGDs f
database leaching bottom ash: composition of the leachate, pH and leaching percentages
g database leaching pulverized fuel ash: composition of the leachate, pH and leaching percentages h database Co-firing: all the data and relative parameters as mentioned in the databases a up to and including g, but classified according to the secondary fuels. The foremost database is the one concerning the coal composition. Therefore the next section more detail will provide on this database. 4.2
Database coal
It appears that the differences in elemental concentration between lots from the same geographical region is relative small. Therefore all the samples from one region are combined. For one country of origin the mean value together with the standard deviation will used henceforth. The weighted averaged coal composition, as fired in coal-fired power plants in the Netherlands in a particular year, can be calculated, based on the origin of the coal for that year, together with the standard deviation. PREDICTION OF ENVIRONMENTAL QUALITY OF BY-PRODUCTS The databases as mentioned in chapter 4 contain data on the composition and leaching of various streams of the Dutch coal-fired power plants. Because of this one can get information of the composition (mean and standard deviation) in the past. One can extrapolate into the future. However, besides composition, relative parameters are always as well recorded. These relative parameters can be used to predict the composition of the various streams. A general scheme is presented in figure 2. The basis is the coal composition. If one knows the coal composition, the composition of the various ashes and flue gases can be calculated.
322 If the coal composition is not known but the origin of the coal is known, the coal composition can be taken from the database. This latter approach will yield a fairly good prediction. This approach is also valid for blends and was tested in practice with 21 samples from 6 different Dutch power stations in 1993 and 1994. A comparison between the measured and the predicted values of the coal composition (42 elements) yields a mean R 2 value of 93+7%. The same procedure was followed for the composition of the pulverized fuel ash: it yields a mean R2 value of 93+8% The prediction for some individual elements can be less accurate, such as (in decreasing order) Hg>Br>Be,Se>Cd,Zn. The origin of the coal in a particular year is well recorded, so for that year the averaged composition together with the standard deviation of the various streams can be calculated. Hence, the representative figures for that particulate year are obtained. It turns out that all the relative parameters also apply to co-firing, at least for the study cases, in which at the most 10% of the coal was replaced by a secondary fuel. So the model is also useful in predicting the consequences of co-combustion. The prediction of leaching is only indicative, as discussed in section 3.4. However, in many cases this is sufficient.
323 COAL DATABASE elemental composition per
+
;
RE-factor bottom ash
composition bottom ash
I
RE-factor PFA
RE-factor fly a s h
composition P F A (esp)
+
comp. fly a s h
vaporization
cone. in flue g a s e s
. . . . .
........
leaching parameters . . . . . .
cocombustion composition
country
..............
........
.+
i.p. a,ea+c r a m e t,,-+ e r s +-[I
relnoval FGD
l re|nol val FGD
emission (leachate)
emissions into air
emissions into air
I. . . . .
einission (leachate)
Figure 2
Scheme for prediction of the composition of the various streams a coalfired power plants
6 -
CONCLUSION the elemental composition of coal and the outgoing streams, such as bottom ash, pulverized fuel ash (ESP-ash), emitted fly ash, flue gases and leachate, is recorded in databases
-
besides elemental composition of the outgoing streams, relative parameters are also deduced. These parameters are: relative enrichment factors, vaporization percentages, removal percentages and leaching percentages
-
these relative parameters also apply to co-firing
-
the elemental compositions of coal blends are on average easy to predict
324 -
the elemental compositions of the outgoing streams are on average easy to predict, assuming a known coal composition or a predicted coal composition. Some elements are less well to predict in ash, such as mercury and bromine
-
the prediction of the elemental composition of leachate is only indicative. This is due to the fact that leaching depends on several parameters. However, for most cases it can be sufficient. If necessary, e.g. when the leaching is close to a standard, an advanced model for predicting the concentrations of that particulate element is needed. This model can be made based on information derived from the database and from the lysimeter experiment. ACKNOWLEDGMENTS
This research has been funded by the Dutch Electricity Production Sector and by the 'Vliegasunie' (the Dutch Fly Ash Association).
REFERENCES
BERGH, J.W. VAN DEN, 1997. By-products management and quality control. These proceedings. HOEK, E.E. VAN DER, MEIJ, R., 1995. Speciation of arsenic, selenium and chromium on pulverized fuel ash in laboratory and practice. Third international conference on the biochemistry of trace elements. May 15-19, 1995 Paris. MEIJ, R., 1989. Tracking trace elements at a coal-fired power plant equipped with a wet flue-gas desulphurization facility. In: KEMA Scientific and Technical Reports, 7 (5), pp. 267-355.
325 MEIJ, R., SCHAFTENAAR, H.P.C., 1994. Hydrology and chemistry of pulverized fuel ash in a lysimeter or the translation of the results of the Dutch column leaching test into field conditions. In: Proceedings of the international conference WASCON '94, held in Maastricht, The Netherlands, June 1-3, 1994. The proceedings are published in "Studies in Environmental Science 60, Environmental Aspects of Construction with Waste Materials", edited by J.J.J.M. Goumans, H.A. van der Sloot and Th. G. Aalbers, published by Elsevier, Amsterdam, ISBN 0-444-81853-7. MEIJ, R., 1994. Trace elements behavior in coal-fired power plants. Fuel Processing Technology 39 pp. 199-217. MEIJ, R., KOOIJ, J. VAN DER, SLUYS, J.L.G. VAN DER, SIEPMAN, EG.C., and SLOOT, H.A. VAN DER, 1993. The emission of fly ash and trace species from pulverized coal-fired utility boilers. In: Proceedings of the Vlth World Congress on Air Quality, held in Paris, May 1983, part IV, pp. 317-324. POEL, M. VAN DER, 1997. Environmental legislation. These proceedings. SLOOT, H.A. VAN DER, 1991. Systematic leaching behaviour of trace elements from construction materials and waste materials. In: Waste Materials in Construction, J.J.J.R. Goumans, H.A. van der Sloot and Th. G. Aalbers (editors), Studies in Environmental Science 48, Elseviers Science Publishers B.V., ISBN 0-444-89089-0 pp. 19-36.
This Page Intentionally Left Blank
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
327
SHORT LEACHING TEST COMPARED TO A COLUMN LEACHING TEST FOR
INTERNAL QUALITY CONTROL OF COAL BOTTOM ASH Eline E. van der Hoek and Frans J.M. Lamers KEMA P.O. Box 9035 6800 ET Arnhem, The Netherlands ABSTRACT In an internal quality control system, necessary to obtain environmental certification, it is possible to use a short leaching test. It is desirable to use short leaching test procedures instead of the Dutch regulatory column test (NEN 7343) which lasts over 21 days. Therefore, a study is performed comparing the leaching results of coal bottom ash from the column test by two short leaching test (first step cascade test, NEN 7349, and the CEN 292 WG 2 compliance test). The CEN test method is a possible short test to replace the column test. 1
INTRODUCTION
Coal bottom ash is mainly used as a granular building material in road constructions. To guarantee its use as construction material it is necessary to comply to the immision demands of the Dutch Building Materials Decree. For granular materials the immision demands must be tested by measuring the emission from a column test (NEN 7343). The column test procedure lasts over 21 days. This long test procedure gives problems because big lots of bottom ash have to be stored before it can be used. There is not always enough room at the power station to store these big lots. In the Building Materials Decree it is possible to use environmental certificates. To obtain this certificate an internal quality control procedure must be used. It is possible to use short test procedures in this internal quality procedure as long as the test results can be translated to the test results of the column test (Lamers, 1997). The Dutch fly ash corporation has started a study to the possibility of using a short leaching test for examining the emissions from coal bottom ash.
328 First, the first step of the cascade test (NEN 7349) was studied. The cascade test is a single stage batch test at liquid to solid (L/S) ratio of 20 I/kg; it lasts for 23 hours. For many years the cascade leaching test (NEN 7349) has been used for the quality control of bottom ash and many leaching data from this test are available. There is a great deal of experience with this test procedure in the laboratories of the power stations. Therefore, it seemed attractive to use the first step of the cascade test as short test. Results of the cascade test study indicate that statistical correlations can not be found because of external factors influencing the test results. This is further described under "cascade test". It was decided to study the possibility of another test, the CEN 292 WG 2 test (CEN test). The CEN test is a two stage compliance test which is completed within 24 hours. In theory the CEN test is more comparable with column test. The CEN test procedure that was used consist of a two stage test and the final L/S solid ratio (L/S = 10) is the same as the column test. Furthermore, by using the CEN test for quality control of bottom ashes the power plants will follow European developments. For a good statistical comparison, it is important to know the reproducibility of the different test procedures. Therefore, first the reproducibility of both the column test (NEN 7343) and the CEN test (CEN 292 WG 2 compliance test) procedures was studied. The results of this study are described under "CEN test". 2
CASCADE TEST
2.1
Method
The first step of the cascade test (NEN 7349) was studied as a possible short test procedure. The first step of the cascade test is a single stage batch procedure at L/S 20 for 23 hours. The leaching results of 15 column tests where compared with results of the first step of the cascade test (KEMA, 1996a). The column test and chemical analysis were performed at the laboratory of KEMA, while the cascade test and chemical analysis were performed at laboratories of the power stations.
329
The leaching These
of potential
elements
critical elements
are potential
S e , M o , B a , V, S O 4 a n d S b w a s m e a s u r e d .
critical as was found
in e a r l i e r c o l u m n
test results (Lamers,
1997).
2.2
Results
It w a s n o t p o s s i b l e
to relate leaching
of the Se leaching
data from coal bottom
the cascade
test. Sometimes
S e is l e a c h e d
in t h e c o l u m n
r e s u l t s o f b o t h t e s t s . In f i g u r e 1 a n e x a m p l e ash percolated
m o r e S e is l e a c h e d
from the column
in t h e c a s c a d e
is g i v e n
test and from
test and sometimes
more
test.
Selenium 0"3
i
,
~
.................. 9 I.................. i................................................. .~................. ~................. i...................................... i.................t.................. ;.................. ~.................. i.................. i J.........i................................... i.................. ~.................. i................... i................................... i.................. i.................. ;.................. ;.................................... i...
E
0.2
~
0.1
o l---
.
.
.
.
.
'
.
.
.
.
.
.
0.05
0
30559 30769 31450 31860 32195 32896 32917 33308 30723 31063 31859 31873 32895 32916 33278
Sample number column test
Figure
1
Leaching cascade
of Se from test
~
c a s c a d e test
15 coal bottom
ashes
in t h e c o l u m n
t e s t a n d in t h e
330 Both the difference between the tests and external factors can cause differences between the test results. Differences in test procedures which can cause differences in the amount of element leached are: leaching time percolation versus shaking pH differences differences in L/S ratio and reproducibility of test procedure. External factors which can cause differences in the amount of element leached are aging processes (the test were not always performed at the same time) pH differences (caused by aging or by heterogeneity in the sample) test performances at different laboratories different methods for analysing the percolate and sampling and sample handling. In the cascade test results many external factors could have influenced the results. There are not enough test results to perform a good statistical analysis. For a good statistical evaluation it is important to decrease the number of external factors which can influence the results. It is also important to know the reproducibility of the test. 3
CEN TEST
3.1
Method
The two stage batch test of the CEN 292 WG 2 was studied as a possible short test (KEMA, 1996b). The test procedure consists of a two stage test at L/S 0-2 for 6 hours and at L/S 2-10 for 18 hours. To test the reproducibility 10 column tests and 10 CEN test were performed at one bottom ash sample. 20 Kg bottom ash sampled according to NVN 7302 was dried at 40 ~ for 24 hours. The sample was divided in 10 subsamples and these samples were crushed and sieved. All particles were smaller then 4 mm. Each subsamples was split into two parts.
331 One part was submitted to a column test and the other part to a CEN test. For comparison the fractions of the column test were combined to two fractions: L/S 0-2 and L/S 2-10. In the percolates the pH and electrical conductivity were measured. The percolates were analyzed for the critical elements Se, Mo, Ba, SO4 and Sb. 3.2
Results
The leaching of Sb was in all samples below detection limit and was not further considered. The deviation in the chemical analysis, based on the accuracy of the analysis, was compared to the deviation in the whole test result. The deviation in the analysis was smaller than the deviation in test, which indicates that the deviation in the analysis can be neglected. As an example the Se test results with their analytical deviation (triangles) are shown in figure 2. In figure 2 also the emission demands of the Building Materials Decree are given for an application height of 1,5 m, which is considered the maximum height of utilization of coal bottom ash. The average leaching concentration of Se is 0.026 mg/kg which is below the emission demand of 0.036 mg/kg. One column test result gave a Se leaching concentration which was above the emission demand. Se was leached from the CEN test at higher concentrations than from the column test. Heterogeneity of the subsamples can also change the test results, however no extremes were measured among the 10 different tests.
332
Se, L/S 10
0.06
"'T ................. ." ................. .T................. T ................. T ................. T ................................... ,~.................. r ........................................................................
i
~-j
_]
i
.
.~
.
. i .
~
J
0.04!
+
i
] +
~
~
~
i
"i"
+
i
"~
i
!,
t~:l
~
,,,i
i
i.
+
i ................i............:+;<........ i.............i;"i ............it ,..-
9~_
+
0,03
+
+
+..+................. +........... + ........
+
..+...........
~
,.
: 1
i
~
L
+
i
i
i
i
I
i
,-j,
i
+;~
~, I
+
............. +........................... +..........................................
................+............
+
+
+
+
+
!
~
~
/
~
t
I..........T.............t ..............." ..............; ...............t .............":~- ............t ..............r~ ...............!........-~:".'~
"
~.............+.................+ .............i ............++" ..............i............+
80.~
.- ................. ; ................. :
~
+
+
"" ..........+:............+. ................................................... + ............+......... +...............+...........~,
+
!
~
i
cat
+i
i
] ! ~....................................................................... i
,+
+
i /~ i ~., +................. ~.................. +................. +............ :
+
!
...............................................................
! .,.+ .............................. + + .......++...........++ .............+i..............p..............................
................................................................................................................................... +................i ..............--. +
+
t
+
|
+
+
+
A
............................................................................. ................+i...............'~................................ +i .............+!...............+i ..............+I ..............t,................. + ................................................... ! I I i+ , .................. ~+ ~ i i o.0~+.+ 1
"
................ +............. +............. +......... +........... +................ i............ ~ ............... +................. +............... i............... +................. i................ 1................ +.................+..................i .................i .................;
2
3
column test
Figure 2
4
5
6
sample number
CEN 292-test
7
8
9
10
analytical deviation
The concentration and analytical deviation of Se leached from the column test and the CEN test. The calculated emission demand for utilization as cat 1 from the Building Materials Decree is indicated as a straight line
The repeatability of a leaching test can be expressed by the variation coefficient. In the testing protocol for the Building Materials Decree it is assumed that the total variation coefficient of the measurement including sampling errors, sample preparation, analysis and the performance of the leaching test is 25% (RIVM, 1995). For most of the elements leached from the bottom ash we measured a variation coefficient < 20% in both tests. Only the variation coefficient for Se in the column test was 28% and for Mo was 27%. For regulatory purposes, it is possible that the variation coefficient becomes even higher, because in that case the variation of sampling and sample preparation will also be included. Se is the most critical parameter for coal bottom ash. This implies that a higher amount of bottom ash is falsely rejected than was expected in the testing protocol.
333 Relation of the column test and the CEN test All experiments were performed on the same bottom ash sample. Therefore, no statistical correlation between the results of column test and the CEN test can be found (there is only one point). The test results at different fractions (L/S ratio 0-2 and 2-10) are plotted in figure 3, to obtain an idea of the correlation of the two tests. Except for sulphate, the reliability (r~) of the lines is above the 50%. The slopes indicate that, except for Ba, the concentration leached from CEN test are higher than leached from the column test. The
correlation lines start almost from the origin, that indicates that the leaching is linear at increasing L/S ratio. t"
.o "6
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. , i
.o
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iii
~r~ ~
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.
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334 3.3
Discussion on leaching mechanism
In general, to be able to apply short tests it is important to understand the leaching mechanism of the construction material. Leaching is dependent on many factors. The most common leaching mechanisms is leaching controlled by the availability of the elements and leaching controlled by solubility or sorption processes (van der Sloot, 1996). The leaching mechanism will determine the test results. Leaching controlled by the availability is mainly dependent on L/S ratio, whereas leaching controlled by solubility or sorption is mainly dependent on pH. Factors such as aging and sample handling are important when the leaching concentration is controlled by the pH. The L/S ratio in a test controls the amount of leaching at an availability controlled mechanism. Field leaching studies are helpful in revealing the underlying leaching mechanism. In another study the leaching from coal fly ash in the field is studied by using lysimeters. In this study both the hydrology and the transport of elements have been monitored starting from 1993 (Meij and Schaftenaar, 1994). The collected data is used to model the leaching process in which both the hydrology and chemical reactions are considered. The results indicate that the leaching of Se from coal fly ash is availability controlled and can be modelled by a linear sorption model, whereas the leaching of Ba is solubility controlled. Figure 4 shows a comparison between the barium solubility and the barium concentration (transposed to the activity of Ba 2§ in the pore water over 44 months of natural weathering. The resemblance between the Ba activity and the solubility line of barite indicates that barium leaching is solubility controlled.
335
Bariet BaS04
,41
i
"~ '~F~'_ "x ~
a
-cT-
--
IB
In o .J
0
. . . . . .
!
i
4
Log[S04]
Banet |
L
z
t=26 . i
Figure 4
e
t=9
m
t,,15
~
t,,19
A
t,,20
x
t,,23
,
11,28
c~
111.30
~
ts34
~
t=39
m
t--41
so
t,,24
i
Activity of barium and sulphate calculated with MINTEQA2 in the pore water of a lysimeter filled with coal fly ash and compared with the solubility of Barite
Leaching models can be helpful in translating short batch test results to results into the results of characterization tests such as the column test. 3.4
Conclusion and further research
The CEN test method seems to be a possible short test to replace the column test for internal quality control. In order to perform a good statistical analysis, it is necessary to obtain a representative data set of both column test and CEN test from coal bottom ash. Special care has to be taken for sampling, sample handling and test performances. Knowledge of the underlying leaching mechanism can be used in translating the test results from short batch tests to results from characterization tests.
336 At this moment a study has started to obtain a representative data set. Figure 5 shows the first results of this study. In figure 5 the results of Se leaching from different bottom ashes from 5 different power stations as measured with the CEN test are compared with the results of the column test. The results indicate a good correlation between the two tests. Selenium L/S 10 14 y = 0.7416x + 0.5512 R2 = 0.935 9
v r
i lo
,=,= O tj
9
8
9 Series1 i linear correlation line
U
]
6
~
4
~
2
o 0
0
2
4
6
8
10
12
14
16
Concentration leachate CEN test (1~1/I)
Figure 5
Correlation between the leaching of Se from the CEN test and from the column test from different coal bottom ashes
REFERENCES LAMERS, F.J.M., VAN DEN BERG, J.W. and BORN, J.G.P. 1997 Environmental certification of bottom ashes from coal fired power plants and of bottom ashes from municipal waste incineration, WASCON 1997 MEIJ, R. and SCHAFTENAAR, H.P.C. 1994 Hydrology and chemistry of pulverized fuel ash in a lysimeter or the translation of the results of dutch column leaching test into filed conditions, in Environmental aspects of construction with waste materials, proceedings of the international conference on environmental implications of construction materials and technology developments, Maastricht, the Netherlands. Eds. Goumans, J.J.J.M., van der Sloot, H.A., Aalbers, Th. G., 491-507.
337
KEMA, 1996a (van der Hoek, E.E., Avis L.H. en Venhuis L..P.) De cascadetest als mogelijke verkorte test voor de beoordeling van E-bodemas. 53627-KES/VVBR 96-3122. KEMA, 1996b (van der Hoek, E.E. en Avis L.H) Onderzoek naar de mogelijkheid van de cen-292 test als verkorte test voor de certificering van e-bodemas. 54320.300-KET/PTE 96-3024. RIVM, 1995 (Aalbers, Th.G. en Derksen, G.B.) Toesten van bouwkwaliteit aan normen en eisen. Rapportnr 771402010. VAN DER SLOOT, H.A. 1996 Developments in evaluating environmental impact from utilization of bulk inert wastes using laboratory leaching test and field verification, Waste Management, Vol 16, 65-81.
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Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997Elsevier Science B.V. All rights reserved.
339
Retention mechanisms in mortars of the trace metals contained in Portland cement clinkers I. SERCLERAT, P. MOSZKOWICZ Laboratoire d'Analyse Environnementale des eroc~d~s et des Syst~mes Industriels LAEPSI INSA-Lyon, Bat 404, 20 Avenue Albert Einstein, 69100 Villeurbanne France
Portland cement clinker is made of limestone and clays heated up to 1450~
These raw materials, as
well as the fossils fuels fired in the kiln, contain trace amounts of heavy metals, just like any natural resource. Furthermore, flammable residues such as low-grade coals, tyres, solvents, are commonly used in substitution of coals or fuels. The co-firing of waste derived fuels WDF in cement kiln must comply with the European regulation 94-67 (December 16th 1994). Some of these WDF may present significant metal contents. However it must be highlighted that the wastes fed in the kiln never exceed a few percent of the mass inlet, so in most cases, the total metal content is not significantly affected by the co-firing of WDF 1, excepted if a waste exhibits a level of one metal specially higher than the average content of the raw material. It is mainly the case for lead fairly abundant in some used oils, and zinc brought by old tyres ; that's why these metals have been chosen for our study, along with chromium which is known for producing soluble and hazardous chromates in cement. The first part of the experiments carried out in this study concerned the leaching behaviour of industrial mortar bars contacted with deionised water. We showed that the metal concentrations in the leachates are very low and often non detectable. These results confirmed previous works related in the literature 25. Furthermore, various measurement campaigns also proved that the mortars coming from kiln burning WDF do not exhibit metal release higher than the samples produced with classic fuels4"5. Yet the leaching behaviour of the trace metals is not fully explain in the literature. In order to explain the effectiveness of their retention in the mortar bars, the second part of our study was dedicated to the understanding of the containment mechanisms of the trace metals, and of the experimental parameters controlling their release. We have therefore designed more specific experiments : 9 Laboratory samples have been enriched in metals during clinkerisation, up to ten times the usual metal content in industrial samples. They should ease the observation of the metal release, and therefore provide a better understanding. Also, complementary leaching tests in chemical or physical contexts more or less aggressive for the cement matrix appeared very fruitful : 9 The influence of the pH on the metal release have been studied, using leaching tests of monoliths in various leachants, more or less aggressive for the cement material ;
340 9 Extraction experiments on crushed material have been undertaken to assess the fixation of the metals in the matrix itself by purely chemical mechanisms, after elimination of all the transport phenomenon. These leaching parameters have obviously nothing to do with the conditions of use of cement based material, and are merely experimental tools to assert the release mechanisms. This paper is essentially focused on the second part of our experiments, concerning the retention
mechanisms of the trace metals.
Experimentais Four industrial clinkers have been chosen amongst French production for their levels of chromium, zinc or lead higher than the average. The corresponding raw materials had also been provided, from which tt replicates >> of the industrial clinkers were synthetised in an electrical laboratory furnace. Corrective additions of metals were made, so as to obtain identical metal contents in the industrial samples and in the laboratory replicates (table 1). The release behaviour of the mortars made from these first two sets of samples were compared, using the tests described below.
II ][ il
mg]k~ 'I Industrial '"
I Replicate II Industrial
'
Cr
Pb
ZH
101 98 58 62
32
2oo
6 2 13 24
192 224 228 246
H Replicate H I Industrial I 18 ....m Replicate 155 14 248 table I : Heavy metal contents of the industrial samples and their laboratory replicates
A third set of clinkers was made in laboratory furnace, from an industrial raw meal that have been enriched in chromium, zinc and lead before the clinkerisation. Three metal levels were chosen, up to ten times the maximum concentrations usually encountered in industrial samples (table 2). The levels of hexavalent chromium have also been checked afterwards.
mg/kg
II HI|
Crtotal
Pb
"
Zn
B ...... i'80 .... 150 230 M 1005 680 1090 H 1810 1805 1920 table 2 : Heavy metal contents of the enriched laboratory cements
Cr(W) io0 610 1120
These three sets of samples have been crushed after addition of gypsum, so as their final composition and their hydraulic properties are similar to industrial Portland cement. Mortars have been prepared with the usual Cement/Sand/Water ratio = 1/3/0.5, and were mould in cylindrical bars ; after 28 days of maturation at 20~
and 98% of relative humidity, they were cut
into disks of 1,4 cm high and 8 cm in diameter (volume 70,4 cm 3, surface 132 cm 2, weight 152+_2g).
341
These test samples were submitted to sequential leaching tests in deionised water during 100 days overall, according to the following: 9 Static, batch leaching tests in sealed polyethylene bottles 9 Ratio : volume of leachant / sample surface = 5cm (i.e. 660 dm 3 of leachant ; liquid to solid mass ratio = 4,33) 9 Immersions:
- either 10 sequential contacts of 1d - 1d - 1d - 4d - 7d - 7d - 7d - 14d - 28d - 30d (total 100d), - or 1 continuous contact, during which aliquots are withdrawn at the times above mentioned. 9 Leachant : deionised water excepted otherwise stated.
Complementa~ experiments 9 Sequential leaching tests had been carried out as described here above, though using an alkaline leachant (pH 12.7) which is typically non aggressive for the cement matrix, or in pH conditions regulated at 7 by nitric acid additions. 9 Extraction tests: Mortar samples crushed to 100~m were contacted till equilibrium with solutions of various pH (liquid to solid mass ratio - 10). The influence of the pH o f the leachant upon the effectiveness of the fixation has been established, by using contact solutions maintained at various pH values between 6 and 13.
Results The leachates have been analysed by ICP AES or graphite furnace SAA. The detection limits are 4lag/l for chromium, 10lag/1 for lead and 3 lag/] for zinc.
l-Comparision of industrial samples and laboratory replicates The table 3 presents the total release of metals by the mortar bars contacted with deionised water. Most o f the leachates exhibit metal levels below the detection limits.
cumulative amount leached' in 100d (lag) Zn Pb Cr I Industrial .....! Replicate II Industrial II Replicate HI Industrial HI Replicate
NS NS NS NS NS NS NS NS NS NS NS NS NS 29 NS 33 NS NS table 3 : Metal release from the industrial samples and the laboratory replicates (10 immersions, total duration 100d) NS 9non significant because too many leachates concentrations < detection limit The metal release being non measurable in deionised water, these two sets of samples have been tested in conditions more aggressive for the cement matrix, to make sure that the clinkerisation in
342 laboratory furnace provides samples whose leaching behaviour is representative of industrially-made clinkers. Lead and zinc
For the industrial samples and their laboratory replicates, the lead and zinc concentration in the leachates are consistently under or close to the detection limits, whatever the conditions applied: sequential leaching of monoliths in deionised water (table 3), in alkali, at pH 7 regulated, or even during the tests of extraction from crushed material at various pH values. Hence there is no measurable difference between the two sets of samples, considering the release of lead and zinc. Chromium
The chromium release is also quite low, and sometimes under the detection limits. The cumulative chromium extracted over the test duration can though be worked out. In the different chemical conditions tested, the metal release from the monoliths appeared to be directly linked to the chromate content of the solid. The clinkering conditions of the sample (industrial kiln or laboratory furnace) has no influence. This result has been confirmed by the extraction tests on crushed material, as shown in table 4 : IIdeionised water 0tga)
pH 7 (ptga)
791 I industrial 50 789 I replicate 31 355 H industrial 15 489 II replicate 22 1005 HI industrial 66 1036 .... !II replicate 84 table 4 : Solubilisationof chromium from crushed material
% Cr 6+ extracted at pH7
110 106 105 108 110
Concerning chromium, the samples made from laboratory clinkers present a leaching behaviour very similar to the corresponding industrial samples. Concerning lead and zinc, and within experimental accuracy, the laboratory samples do not exhibit any obvious discrepancy from the industrial ones. The study has therefore been pursued using solely the enriched laboratory clinkers.
2- Release mechanisms studied on enriched laboratory samples
The mortars made from enriched clinkers exhibit a measurable metal release (table 5), the first extracts being the more concentrated. The amount of lead and zinc leached are still very low ; the corresponding concentrations in the leachates are in the lag/l range. The chromium levels are somewhat higher. For the sample H (enriched to ten time the usual content in industrial samples), the concentrations in the leachates reach 150~tg/1, in the experimental conditions here applied.
343
Cumulative amount leached (ttg)
Enriched
Pb
Cr
samples
Zn
34 NS NS 245 34 26 M 459 101 23 H table 5 : Metal release from the laboratory enriched samples (10 immersions total duration 100 d) NS ' non significant because too many leachate concentrations < detection limit B
~nc
The figure 1 shows the extraction of zinc from crushed mortar at various pH values. Zinc appears to be insoluble for the pH higher than 8. Its concentrations in the leachates are lower by of orders of magnitude than the solubility of the common zinc compounds such as hydroxides or carbonates, thus indicating a chemical bounding in the solid phase. 25C
pg/I Zn F"
20C 15C 10C 50 0
I
7
8
9
10
11
pH
12
figure 1 : Extraction of zinc from crushed material - sample H Such a low solubility strongly limits the release of zinc by the mortar monoliths : The level of zinc in the leachates of mortars blocs are consistently in the ~tg/l range. The cumulated amount leached in 10 immersions (total length 100 days) are between 15 and 25 ~tg; there is no significant influence, neither of the chemical conditions of leaching, nor of the metal content in the solid. Lead The solubilisation curve of lead from crushed mortar is presented figure 2. Lead is partly extractable from the mortar in strongly alkaline conditions, but it is bound in the solid for pH values under 12.5. Just like for zinc, the equilibrium concentrations against pH are far lower than the expected solubility of lead
compounds 6,7.
344
5C) pg/I Pb
)C)
5C)
0
7
8
9
10
12
11
13 pH 14
figure 2 : Extraction of lead from crushed material at various pH - Sample H The various experiments undertaken have pointed out the proportionality between lead release and its level in the mortars. The results of lead released from monoliths are therefore expressed as percentages of the metal content of the solid (figure 3) 9 1,4%
% lead leached in 42 days
1,2% 1,0%
0,8% 0,6% 0,4% 0,2% 0,0%
!
pH 7
!
D-water
pH 12,7
figure 3 : Percentage of lead leached from mortar bars, in various leaching conditions The mortar bars exhibit a good retention of lead when they are contacted with neutral or moderately alkaline solutions (the leachant pH rose up to about 11 during the tests in deionised water); but when the leachant pH is very high, lead is partly released by the mortars. The comparison of the two previous curves shows that lead release is strongly influenced by a solubilisation process controlled by the pH value.
Chromium Whatever the leaching conditions, the release of chromium is directly proportional to the chromate content of the sample. The extraction test at various pH (figure 4) provided important results : - As foreseen with the non-enriched samples (w 2.), the chromium in the leachates is solely in its hexavalent form. Hence the trivalent chromium is never solubilised. - The total amount of chromate of the sample is extracted for pH values under 10. - Chromate, though usually soluble, is bound in the solid in the pH range 1 l-13 ; we point out the fact that this range corresponds to the pH domain in which the ettringite phase is stable according to Damidot and Glasser 8.
345
Cr extrait (rng/I) 30252015-
1050
6
I
I
I
I
7
8
9
10
11
12
13 pH 14
figure 4 : Extraction of chromium from crushed material at various pH - Sample H The release of chromium by the mortar bars is shown figure 5. Just like in extraction tests, only hexavalent chromium is leached. The mortars contacted with deionised water release less metal than in aggressive conditions such as pH 7 regulated. Surprisingly, the release is quite high in an alkaline solution of pH 12.7, although the extraction test proved that the solubilisation of chromium is minimal for a pH value of 12.5. Hence the release of chromium cannot be explain taking into account only the influence of the pH.
2,0%
chromium extracted in 42 days
1,5% 1,0% 0,5% 0,0%
. . . . . I
pH 7
l
D-water
I
pH 12,7
figure 5 : Percentage of chromium leached from mortar bars, in various leaching conditions
Discussion The release of zinc by the monoliths contacted with deionised water is very low (In our tests, less than 25~tg of zinc is extracted in 100 days of leaching, whatever the metal content of the mortars). It is due to the fact that this metal is bound in the solid as a compound which is insoluble in water. This result can be extended to the various chemical conditions applied in our tests, as zinc is chemically retained in the matrix in the pH range 8-13. It must be highlighted that the elevated alkalinity of cement ensure an elevated pH in the mortar bars even if the surrounding solution is fairly
346 aggressive : this remark explains why the zinc release by the monoliths is still low when the leachant is maintained at pH 7. The release of lead from the monoliths is controlled by a solubilisation process strongly dependant of the pH conditions applied to the material. This metal being bound in the matrix for the pH lower than 12.5, its release is limited as long as the alkalinity of the leachant is not too elevated, which is the case of leaching in deionised water or at pH 7 regulated. Very high values of pH would be reached only during a prolonged contact between the cement and the liquid, or when the mortar is submitted to an extremely alkaline solution. Such conditions of pH does not occur during the real utilisation of cement based materials. An other part of this s t u d y I has proved that the chromate is chemically bound in the ettringite structure, in substitution for sulfate. The interpretation of the leaching results requires to consider the specific chemical properties of such chromate-ettringite, and especially its solubilisation mechanisms. As an example, the figure 6 presents the evolution of chromium concentration during an unique continuous contact. The metal level rise up and becomes stable after a few days. This stationary value is directly proportional to the chromate content of the solid. The release of chromium is due to the partial dissolution of ettringite in which it is contained as an impurity. The stationary level of chromium is due to the quick saturation of the bulk with respect to ettringite. Depending upon the chromate level in the solid, the bulk is in equilibrium with ettringites containing different amounts of metal impurity. The partial dissolution of these phases liberates a corresponding amount of chromate into solution. 200 pg Cr 150
XX
X
X
l
x
100 ~ , 50
X
eB
*M
"
xH
poe
0 0
50
75 j o u r s 2 5 0
figure 6 : Amount of the chromium released in the leachant during one prolonged contact
Conclusions The first part of our study has shown that the traces metals occurring in Portland cement clinker are retained in the relevant mortars bars when they are submitted to deionised water. The metal concentration in the leachates are consistantly under or close to the detection limits. Furthermore, samples enriched in metals up to ten times the levels usually encountered in industrial samples also exhibit very low metal release.
347 To explain these results, the work presented in this paper focused on the understanding of the retention mechanisms of the trace metals, and on the identification of the parameters controlling their release: -
Zinc is bound in the solid and is nearly insoluble in the chemical conditions applied ; therefore its release by mortar bars contacted with deionised water is very low. This result has be extended to leachants in the pH range 7-13.
-
Lead is nearly not released in deionised water: we showed that this metal is insolubilised by the cement matrix, provided the leachant pH is under 12.5. It must be pointed out that in real conditions of use of cement materials, the contact water never reaches such elevated pH.
- The trivalent chromium is bound in the mortars. The soluble chromate ions are partly retained in the matrix owing to their fixation in the ettringite phase. Their release is linked to the dissolution mechanisms of this phase, and the quick saturation of the leachant with respect to ettringite appears as the limiting factor.
Acknowledgement
The authors would like to acknowledge the Association Technique de l'Industrie des Liants Hydraulique (,4 TILH) and the Agence De l'Environnement et de la Maitrise de l'Energie (ADEME) for supporting these studies.
References
1. Sercl6rat, I. Les m~taux traces clans le clinker de ciment Portland- R~tention dans les mortiers et integration dans les hydrates de ciment. Thesis ISAL 960140, INSA Lyon, France (1996). 2. Germaneau, B., Bollote, B. and Defosse, C. Leaching of heavy metals from mortar bars in contact with drinking and deionised water. Emerging technologies symposium on cement and concrete in the global environment, Chicago (1993). 3. Sprung, S. and Rechenberg, W. Einbindung von schwermetallen in Sekundarstoffen durch Verfestigen mit Zement. Betonteschnische Berichte 1986-88, Beton Verlag Ed., Dusseldorf (1989). 4. An analysis of selected trace metals in cement and kiln dust. R&D serial N~
Portland
Cement Association, Skokie, ILL (1992). 5. Kanare, H.M. and West, P.B. Leachability of selected chemical elements from concrete. Emerging technologies symposium on cement and concrete in the global environment, Chicago (1993). 6. Pourbaix, M. Atlas des ~quilibres ~lectrochimiques ~t 25~
Gauthier-Villar, Paris (1963).
7. Sanchez, F. E,tude de la lixiviation de milieux poreux contenant des esp~ces solubles. Application au cas de d~chets solidifids par liants hydraulique. Thesis ISAL 960118, INSA Lyon, France (1996). 8. Damidot, D. and Glasser, F.P. Thermodynamic investigation of the CaO-A1203-CaSOa-H20 system at 25~ and the influence ofNa20. Cem. Concr. Res. 23 : 221 (1993).
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Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997Elsevier Science B.V. All rights reserved. STUDY OF CEMENT-BASED SEWAGE SLUDGE ASH
MORTARS
349 CONTAINING
SPANISH GROUND
Monzo J., PayS, J., Borrachero M.V., Bellver A. and Peris-Mora E. Grupo de Investigaci6n en Quimica de los Materiales (GIQUIMA) Departamento de Ingenieria de la Construcci6n Universidad Polit6cnica de Valencia Camino de Vera s/n 46071-Valencia (Spain)
Abstract
A study of cement based mortars containing spanish ground sewage sludge ash is presented. The influence of original and ground sewage sludge ash on mortars workability and compressive strength has been studied. An initial decrease of workability is observed when 30 % of Portland cement is replaced by original ash. When ash grinding time increases a little increased of workability is observed. Mortars containing a 15 % of ash cured at 40~ for 14 and 28 days showed equal or higher compressive strength than control mortar. No significative differences sere observed among mortars containing ash with different grinding times.
Introduction
As a consequence of water treatment processes, a large amount of sewage sludge is obtained. A part of this sewage sludge is used in agriculture as organic fertilizer and soil amendment. But, depending on the origin of water treated (municipal, industrial, ...) chemical parameters change, and in some cases accumulation of heavy metals and other toxic compounds can be occur, producing adverse impacts on human health and the environment (1). The incineration of sewage sludge is one alternative to manage the excess of sewage sludge production, that some cities are using. This method permits to reduce the volume until 90%, and the sewage sludge ashes obtained can be deposit in controlled landfills. However, landfill sites space limitations and environmental problems have guided the investigations of altemative uses in construction. Sewage sludge ash (SSA) has been used to manufacture bricks (2), to incorporate into concrete mixtures (3,4),in asphaltic paving mixes (5) and mortars (6). In a previous research (7) some properties of cement-based mortars containing SSA were studied. The objective of the present work is to study the influence of grinding of SSA on workability and strength of cement-based mortars. SSA were obtained from sewage treatment plant of Pinedo (Valencia, Spain), that produces about 2,000 tons/year.
350 EXPERIMENTAL Materials. Portland cement used for mortar preparation was conforming to the specifications of ASTM type I. Fine aggregate was natural sand with 2.94 fineness modulus. SSA were obtained from sewage treatment plant of Pinedo (Valencia, Spain). Sikanol-M was used as plasticizer.
Apparatus and procedures. Samples of original SSA were ground using a laboratory ball-mill (Gabrielli Mill-2). SSA samples were introduced into the bottle-mill containing 98 balls of alumina (2 cm diameter) and were ground during 2.5, 5 and 10 minutes. Mortar specimens cast in square prismatic mortar molds with internal dimensions of (40x40x160) mm were used. Preparation of mortars was carried out according to ASTM C-305 test (8), mixing 450 g. of Portland cement, 1350 g. of natural sand and 225 mL of water for control mortar and the rest of mortars replacing by mass a 15% of Portland cement by original or ground SSA. Mortars were put in a mold for obtaining specimens, which were stored in a moisture room (20-x1~ for 24 hours. Afterwards the specimens were demoulded and cured by immersion in 40-x1~ water in order to activate the hydration process until testing at 3, 7, 14 and 28 days. Mortars for workability studies were prepared according to ASTM C-305 (8), mixing 450 g. of Portland cement, 1350 g. of sand and varying water volumes between 200-225 mL for control mortar. The rest of mortars were prepared replacing growing percentages of Portland cement by ground SSA and workability test were developed following ASTM C-109 (9) test. Some tests were developed using a mortar plasticizer (Sikanol-M) in a 0.1% in weight respecting SSA + cement. Freshly prepared mortars were placed into a conic mold which is centered on the flow table. Mortar was put on two layers and compacted with a wooden tamper (10 times). Afterwards, the mold was removed and the table was dropped 15 times (one per second). Flow table spread (FTS) was given as a mean of maximum and minimum diameters of the spread cone.
Results and discussion SSA obtained from water treatment plant was analyzed and the results obtained are presented in Table 1. From among these data can be emphasized the high concentration of sulfate in SSA (12.4 % expressed in SO3content ). High concentration of sulfate are due, chemical reagent used in water treatment. Workability (FTS). The influence of original and ground SSA on mortar workability has been studied. In Figure 1. Flow Table Spread (FTS) versus SSA grinding time is represented for mortars containing a 30% of SSA and 0.5 water cement ratio. In this figure is compared the platicizer influence on FTS. An initial decrease of workability is observed when a 30% of control mortar cement is replaced by original SSA (SSA 0), a more marked decrease is observed in mortar containing plasticizer. A different behavior is observed, for mortars with or without plasticizer, when SSA grinding time increases. Mortars containing plasticizer increase FTS when grinding time do. The most important increases is observed between SSA 0 and SSA 2.5. The absence of plasticizer shows a decrease of FTS when grinding time increase. In all cases workability of mortars containing plasticizer was higher than mortars without it.
351 Table 1. Chemical Composition of Original Sewage Sludge Ash and Portland Cement ~,~
,
~6~
~
.
~.~ . . . . .
,,
Moisture
0.5
....
Loss on ignition
5.1
3.02
Insoluble Residue
16.1
0.95
SO3
12.4
3.54
7.4
2.85
SiO2
20.8
21.00
CaO
31.3
62.87
MgO
2.6
1.05
AI20 3
14.9
4.94
P205
6.7
0.1
Fe20 3
200 175 A
E E
O0
0.1% sikanol 150 -
i-
U.
125
100 t Control
0% s i k a n o l i SSA 0
i S S A 2.5
i SSA 5
1 S S A 10
S S A grinding t i m e (min)
Figure 1. FTS values of mortars containing 30% of SSA versus SSA grinding time
In Figure 2. FTS versus volume of water for mortars containing a growing replacement of cement by ten minutes ground SSA and 0.1% (in weight) of plasticizer is represented. As could be expected, a increase of FTS is observed when water volume do, but this behavior is more pronounced when SSA percentage is slow (15 and 30%). Probably, the important adsorption of water on SSA particles surface determines the short increase of FTS when high SSA percentages (45 and 60%) are used. Compressive Strength (Re). Preliminary studies make clear, in first place, that SSA did not present autocementicious hardening, whereas, secondly, mixtures of Ca(OH)2 -SSA hardened in few days. This behavior indicated that SSA could present pozzolanic activity. The influence of original and ground SSA on mortars compressive strength has been studied. Mortars containing a 15% of ash and 0.5 water / (cement + SSA) ratio were cured at 40~ and tested at 3,7,14 and 28 days (Figure 3.). No plasticizer was used. The results obtained showed higher Rc in short
352 curing time (3 and 7 days) for control mortar (without ash). When curing time increases (14 to 28 days) mortars containing ash showed equal or higher Rc than control mortar. This fact confirm pozzolanic behaviour of SSA. No significatives differences were observed among mortars containing SSA with different grinding times. 175 -
~.
p.
"
15%
150
125 ,,
,,,
100 200
. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
205
210
215
220
225
Volumeof water(mL) Figure 2. FTS values of mortars containing from 15% to 60% of SSA versus volume of water
50SSA2.5 45-
.,~
SSA10
control
c o n t r ~
7
40
A
A
n =E
/
1
v
0
35
i
30
25
0
. . . . . .
!
7
. . . . .
I
. . . . . .
14
Age (days)
v
21
. . . . . .
28
Figure 3. Compressive strength of mortars containing 15% of SSA with different grinding times
4
. . . . . . .
0
i
7
. . . . . .
T
. . . . .
14
i
21
. . . . . .
28
Age (days) Figure 4. Flexural strength of mortars containing 15% of SSA with different grinding times
353 Flexural Strength (Rf). The influence of original and ground SSA on flexural strength of mortars has been studied (Figure 4). The results obtained showed higher Rf for control mortar than ash mortars except for 28 days curing time that mortar containing 2.5 minutes ground SSA that gave same Rf than control mortar. No significative tendency between SSA grinding time and Rf is observed.
Conclusions 1. High concentration of sulfate are present in SSA, due to chemical reagents used in water treatment 2. A cement replacement by SSA in mortars produces a decrease of workability, being a more marked decrease when mortar contains plasticizer 3. A increase of workability in mortars containing plasticizer is observed when SSA grinding time do 4. Mortars cured f4 to 28 days at 40~ containing 15 % of Portland cement replaced by ground SSA gave equal or higher compressive strength than control mortars. No significative differences were observed among mortars containing SSA with different grinding times 5. Flexural strength for mortar was higher than 15 % SSA replaced mortars except for 28 days curing times
Acknowledgment We would like to express our gratitude to Mr GermAn Rodriguez and Mr Alejandro Mulet from Consell Metropolit/~ de l'Horta for providing us the samples of SSA, SIKA S.A. (Valencia office) and Cementos Asland (Puerto de Sagunto plant) for their support for this research projet.
References 1. Dean, R.B. and Suess, M.J. "The risk to health of chemicals in sewage sludge applied to land" Waste Manage. Res. 1985, 3, 251-278 2. Allenman, J.E. and Berman, N.A. "Constructive sludge management: biobrick" J.Environ. Eng. Div., ASCE, 1984, 110, 301-311 3. Tay, J.H. "Sludge ash as filler for Portland cement concrete" J. Environ. Eng. Div. ASCE 1987, 113,345-351 4. Tay, J.H. and Show, K.Y. "Clay blended sludge as lightweight aggregate concrete material" J. Environ. Eng. Div., ASCE 1991, 117, 834-844
354 5. A1 Sayed M.H., Madany I.M. and Buali A.R.M. "Use of sewage sludge ash in asphaltic paving mixes in hot regions" Constr. Build. Mater. 1995, 9, 1, 19-23 6. Bhatty, J.I. and Reid, J.K. "Compressive strength of municipal sludge ash mortars" ACI Mater. 1989, 86, 394-400 7. J.Monz6, J.Payfi, M.V.Borrachero and A.C6rcoles, Cement and Concrete Research, 1996, 26, 9, 1389-1398 8. ASTM C-305-80. "Mechanical Mixing of Hydraulic Cement Pastes and Mortars of Plastic Consistency". 9. ASTM C-109-80. "Standard Test Method for Compressive Strength of Hydraulic cement Mortars (Using 2-in. OR 50-mm Cube Specimens)".
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
355
FLY ASH - USEFUL MATERIAL FOR PREVENTING CONCRETE CORROSION
S. Mileti61, M. I!i61, J. Ranogajec 2 and M. Djuri6 2 IIMS Institute for Materials Testing, Beograd, Yugoslavia, 2Faculty of Technology, Novi Sad, Yugoslavia
Abstract Large quantities of fly ash is produced in our country every year. Most of fly ash is got from lignite that means that this material is not so useful for concrete production. This paper presents results of investigations of sulphate corrosion of concrete made from portland cements and portland cements with the fly ash addition. The addition of fly ash appears to be very useful for preventing sulphate corrosion of concrete even in the case of very strong ammonium-sulphate corrosion according to our results. The effect of ammonium sulphate solution on the durability of Portland cements (various C3A content) with partial replacement of 30% mass percent'fly ash was investigated. Results show that fly ash addition to Portland cement can improve resistance to ammonium sulphate attack. Key words- Portland cement, fly ash, corrosion, ammonium-sulphate
1. Introduction Concrete has been widely used as the most important constructional building materials in the world. More and more attention has been paid to the mechanical properties and durability of cement concrete. Generaly, the durability and the degradation coefficient of the concrete has been considered as a dominant factor in addition to the fact that the mechanical properties could satisfy the demand of the construction design. In practise, concrete buildings suffer simultaneously mechanical, chemical and physical attacks. Therefore, the effect of mechanical stresses must be taken into consideration when durability and corrosion resistance of a concrete are estimated, i.e., the study of stress corrosion of concrete is necessary and very important for durability. 1-4 Chemical degradation of concrete is the consequence of reactions between the constituents of cement stone, i.e. calcium silicates, calcium aluminates and above all calcium hydroxide etc., with certain substances from water, solutions of soil, gases, vapours, etc.5-8 The most important aggressive ions are: SO4 2", Mg 2+, NH4 § CI-, H § HCO3-. Primarily, the types of chemical corrosion of concrete can be devided into two groups, i.e., expansive corrosion and dissolving corrosion, with respect to the cause of failure of concrete. The attack of sulphate ions on cement stone can cause expansion, in general due to the formation of ettringite C3A.3CaSO4.32H20, in the shape of prismatic crystals. 9'1~ The consequences are damages to the concrete and destruction at worst. The concrete corrosion by ammonium sulphate, for example, covers the most aggressive corrosion on concrete, neither balancing nor creation of protective gel takes place. In this case concrete is damaged not only by expansion, but also by dissolving the cement stone.
356 In this investigation the method of Koch and Steinegger 8 'xs used to test the sulphate resistance of the cements. According to the authors criterion of the sulphate resistance was the quotient:
Rc=Flexural
Strength of the Sample Stored in the Sulphate Solution Flexural Strength of the Sample Stored in Water
/1/
The results show that there is a considerable influence of the mineral composition of Portland cement clinker and cement on the behaviour of concrete in the presence of aggressive sulphate and ammonium ions. For the manufacture of concrete resistant to the attack of aggressive ions special attention should be paid to the selection of cement. ~0
2. Experimental To investigate the resistance of cement to sulphate attack Portland cement and Portland fly ash cement manufactured in Yugoslavia were used" 9 Portland cement B (PCB)- according to the European cement standard EN 197-1: CEM-I 9 Portland fly ash cement B (cement clinker B) with 30% fly ash (PCBP)- according to the European cement standard EN 197-1: CEM II/B-V 9 Portland cement K (PCK)- according to the European cement standard EN 197-1: CEM-I 9 Portland fly ash cement K (cement clinker K) with 30% fly ash (PCKP)- according to the European cement standard EN 197-1: CEM II/B-V The potential phase analysis, chemical contents, physico-chemical and mechanical properties were determined for all starting materials used. Cement pastes were prepared by Koch-Steinegger method. 8 Specimens of 1x 1x6 cm were molded and compacted by vibration. After one day at 100% relative humidity the specimens were demolded and kept immersed in water for 21 days. After that, samples were immersed in the aggressive solutions of different concentrations for different periods of time. Control samples were prepared and stored in distilled water under the same conditions as reference. As aggressive solution, ammonium-sulphate concentrations 2.5%, 5%, 7.5% and 10% was used, but, results only for 10% ammonium-sulphate solution are presented. The mass change of samples, SO42- content change in solution and flexural strength were measured after 7, 14, 28, 56, 90, 180 and 270 days of storage in the aggressive solution. Other testing methods used in this work are: 1. Determination of standard strength (EN 196-1) 2. Chemical analysis (EN 196-2) 3. Determination of setting time (EN 196-3) 4. Determination of the sieve residue (EN 196-6) 5. Determination of specific surface (EN 196-6) 6. Calculating the potential phase analysis (ASTM C 150)
357 3. Results and discussion
The selected aggressive environment represents very strong aggressiveness to ensure fast results for the real conditions which can be present in underground waters in Yugoslavia. TABLE 1 Potential phase composition of Portland cement clinker Potential phase composition, %mass CaS C28 C~A C4AF
Portland cement clinker KB KK 57.5 67.0 13.5 12.7 13.3 6.6 8.7 9.1
The potential phase analysis of the Portland cement clinkers is given in Table 1. It can be seen that the cements have low and high C3A content in clinkers of 6.6% and 13.3% influencing the sulphate resistance. The ordinary Portland cement is not resistant to the attack of sulphates because it has a considerable content of tricalcium aluminate - C3A, whose hydrates react with sulphate ions, giving expansive compounds. Portland cement with increased resistance to sulphates must have a low content of C3A. According to the literature the difference in the C3S content could be significant regarding sulphate resistance too. TABLE 2 Fly ash chemical composition Chemical composition, %mass LOI SiOz A120~ Fe20~ CaO MgO SO~ N%0
K20 Hydrated water Insoluble residue
Fly ash 5.7 50.9 21.7 11.6 6.5 2.7 0.05 0.3 0.7 34.6 76.6
Table 2. presents the chemical composition of fly ash. According to the high content of SiO2, A1203 and Fe203 and the low content of CaO the fly ash is suitable for cement production though loss on ignition was relatively high.
358
TABLE 3 Chemical composition of cements Chemical composition, %mass
Cement PCB 19.7 7.0 2.7 62.0 0.1 0.8 0.1 2.0 2.2 0.4 0.4 0.07
SiO 2 A1203 Fe2Oa CaO Insoluble residue LOI CaO free SOa in CaSO 4 MgO Alkalies as NaaO K20 MnO
PCK 21.0 5.3 2.9 63.8 0.1 0.7 0.4 1.7 1.4 0.3 0.3 0.07
PCBP 14.0 6.2 2.7 44.7 20.2 3.0 0.0 2.0 2.4 0.4 0.2 0.04
PCKP 15.6 4.9 2.9 47.5 18.7 2.9 0.0 1.5 1.2 0.4 0.3 0.05
The chemical composition of the cements is presented in Table 3. All cements meet Yugoslav standard JUS B.C1.011. Portland fly ash cements have a higher loss on ignition and contain less free CaO than Portland cements. TABLE 4 Ph~,sico-chemical properties of cements Physico-chemical properties Sieve residue at 0.09 mm sieve, %mass Density, g/cm ~ Specific surface, cmZ/g Setting -standard consistence, %mass -initial time, min -final time, min Volume stability -Le Chatelier test, mm
PCB 1.8 3.1 3320
Cement PCK PCBP 2.6 5.2 3.2 2.9 3100 3720
25.8 165 225
23.8 165 225
28.0 240 330
27.5 255 360
1.0
1.5
1.0
1.0
PCKP 6.0 2.9 3710
Table 4. presents figures characterizing fineness, density, standard consistency, setting time and volume stability of the test cements. Obviously the addition of fly ash raises the water demand for standard consistency and sieve residue and extends setting time, but has no significant influence on other characteristics. All characteristics are in compliance with Yugoslav standard JUS B.C1.011.
359 TABLE 5 Standard strength of cements Strengths, MPa PCB Flexural: -2 days -3 days -7 days -28 days Compressive: -2 days -3 days -7 days -28 da~cs
]
PCK
Cement [ PCBP
PCKP
4.4 5.3 7.2 8.0
3.7 4.4 7.4 8.9
2.5 3.6 6.2 8.3
2.1 2.9 4.7 8.4
15.7 19.8 30.2 40.3
13.2 16.0 32.8 50.9
8.8 14.9 24.2 39.5
7.4 10.4 19.4 44.9
Table 5. gives values for flexural and compressive strengths of cements after 2, 3, 7 and 28 days. Due to the clinker phase composition, Portland cement PCK had lower initial strength but higher later strengths after 7 days. The Portland fly ash cements had lower compressive strength even after 28 days than the corresponding Portland cements. In this way, complete characterization was implemented regarding all the cements used in this investigation. Figs. 1. and 2. presented mass change of the samples immersed in mentioned aggressive solution. Generally, it is obvious that Portland ash cements had much lower mass change than Portland cements. Test samples from cement PCB lasted only 56 days due to expansion components formation. Capability of mass change for cements PCK and PCKP compared to cements PCB and PCBP was much higher. Those cements also lasted longer. The reason for this must be in clinker composition diferencies.
25
20
/
0 15 o10 o o~
~
_.____-Q
Q
j.. l j
I
l0
-//
--I--PCB
--e--
,
0 0
I
50
,
I
100
,
I
,
150
i
200
,
PCBP
i
250
,
300
Time (days) Fig 1. Mass change for Portland cement PCB and Portland fly ash cement PCBP
360
25
20
./ I f
o 15 Im
vl
- - 9 PCK 9
--V-- PCKP
I
50
,
I
100
,
I
,
150
I
200
,
I
250
300
Time (days) Fig. 2. Mass change for Portland cement PCK and Portland fly ash cement PCKP Figs. 3. and 4. presented SO4 2- content change of the aggressive solution where test samples were immersed. Generally, it is obvious that for Portland ash cements had much higher SO42content change than for Portland cements. Test samples from cement PCB lasted only 90 days due to expansion components formation. It is asumed that all SO42 content changes in solution was directly connected with SO42 bonding in test samples with aluminate components. Capability of SO42 content change for cements PCK and PCKP compared to cements PCB and PCBP was much higher. Those cements also lasted longer. The reason for this must be also in clinker composition diferencies.
361 1300 1200 II00
1000 900 --D--PCB l --o--PCBP
800 0 o
700
r.~
600
e-'
500
,
0
I
,
50
I
,
100
I
,
I
150
,
200
I
,
250
300
Time (days) Fig.3. SO42- c o n t e n t change for Portland cement PCB and Portland fly ash cement PCBP
1300 1200 I100 1000 [ --o--PCK I - - A - - PCKP
900
~
o 800
~
0
700 -
o* r.t3
6~176 f
500
0
I
50
,
I
100
,
I
,
150
I
200
,
I
250
,
300
Time (days) Fig.4. SO42" c o n t e n t change for Portland cement PCK and Portland fly ash cement PCKP
362
10
0.8
0.6
0.4
~O~o~
0.2
0
--i--PCB --o--PCBP
50
100
150
200
i 250
m u I 300
Tmae (days) Fig. 5. Sulphate resistance coefficients for Portland cement PCB and Portland fly ash cement PCBP, according to Koch-Steinegger
tO
0.8
0.6
--m-- PCK I --o-- PCKP
0.2
0.0
0
,
I
50
,
I00
,
150
-
I
200
I 250
w 300
Ttme (days) Fig. 6. Sulphate resistance coefficients for Portland cement PCK and Portland fly ash cement PCKP, according to Koch-Steinegger Koch-Steinegger figures with sulphate resistance coefficients are presented in Figs. 5 and 6. It can be seen from Koch-Steinegger method that the Portland cement with fly ash has better resistance to sulphate aggression for the both kind of Portland cements. Hence, no one
363 of the tested cements shows satisfactory resistance, what is understandable because 10% (NH4)2SO 4 solution was used instead of 4.4% Na2SO4 solution as aggressive medium. Used Portland cements with low content of C3A with and without 30% fly ash (PCK and PCKP) had better sulphate resistance than Portland cements with high content of C3A (PCB and PCBP). The results of sulphate susceptibility tests according to Koch-Steinegger characterized by degradation coefficients are presented in Figs. 5 and 6. From the diagrams, it can be clearly seen that cements with the addition of 30% of fly ash showed distinct higher resistance against the ammonium sulphate solution. The increase of corrosion in the very begining for all cements is a normal phenomenon, because the creation of expansive compounds closes the pores and makes cement paste impervious to aggressive ions. However, further increase in the volume within the paste very quickly results in cracking. For the Portland cement PCB with the high content of C3A this occured after 28 days only. Samples from Portland cement PCK and cements with fly ash addition have endured 90 days. This can be explained by the fact that fly ash in cements has formed a protective layer thus retarding corrosion process and increasing durability. Portland cements, on the other hand, showed, depending on C3A content, either linear or exponential type of degradation after initial period of forming the protective layer. This layer obviously became negligible due to the action of NH4+ ions thus opening new pores and accelerating corrosion process again. The investigations are evidently encouraging, because the addition of fly ash has pointed to realistic prospects for its positive effect. Therefore, cement PCBP, with the addition of fly ash, shows good resistance to the aggressive attack by sulphate solution, although this cement is, due to its phase composition, very unsuitable in that sense. It is evident that resistance of Portland cement to sulphate attack is directly related to its content of C3A. This was confirmed even in the case of complete elimination of physicochemical influence of fly ash on the properties of cement (bonding Ca(OH)2, filling pores, etc.).
4. Conclusion
The results of testing the attack by aggressive sulphate solutions allow the conclusions: 1. The resistance of cements to sulphate attack is higher with a lower content of aluminate in clinker PCK and especially with addition of fly ash to the cement. 2. The addition of 30 % of fly ash to Portland cement as a replacement improves the durability of Portland cement to a considerable degree. 3. Koch-Steinegger method shows that both Portland cements did not resist the strong attack of 10 % (NH4)2SO 4 solution.
following tricalcium of clinker extremely
Reference
Biczok,I.: Concrete Corrosion, Concrete Protection. 8th Edition. Budapest, 1972. Mehta,P.K.: Mechanism of sulfate attack on Portland cement concrete - Another Look. Cem.Concr.Res. 13 (1983) pp. 39-51. Mitrovid, N. and Du~id, V." Sulphate corrosion of concrete. Proc. of Yugoslav. Symposium, Split, 1985, pp. 59-75. Moskvin,V.M., Ivanov,F.M., Alekseev,S.V. and Guzeev,E.A.: Korozija betona i 2elezobetona. Moskva, 1980.
364
10. 11.
12.
Regourd,M.: Structure and behaviour of slag Portland cement hydrates. 7th Intern. Congr.Chem.Cem., Paris, 1980, Vol. I, pp 278-291 Taylor,H.F.W.: Crystal structure of some double hydroxide minerals. Mineral.Mag. 39 (1973) pp. 247-256. Miletid, S. and Ilid,M." Sulphate corrosion of portland cement with various mineral compositions. Proc. 15. Symposi. Corrosion and Protection of Materials, Beograd, 1995, pp. 255-262. Koch,A., Steinneger,H.: An rapid method for testing the resistance of cements to sulphate attack. Zement-Kalk-Gyps 13 (1960), No.7, pp. 317-324. Matsufuji,Y., Koyama,T. and Harada,S.: Service life predictive method of building materials. Proc. Durability of Building Materials and Components 7, Vol. 1, Published by E&FN Spon, London, 1996, pp. 45-53. Miletid, S. and Ilid,M." Cement stone corrosion in sulphate environment. NTP VJ, 46, No. 3 (1996) pp. 23-31. Miletid, S., Ilid,M., Ranogajec,J. and Djurid,M." Sulphate corrosion of Portland cement and Portland cement mixed with fly ash and slag as a function of its composition. Proc. XVI Symposi. on Nordic Concrete Research, Helsinki, 1996, pp. 339-340. Ranogajec,J., Ilid,M., Miletid, S., Lazar,S. and Milinkovid,Lj." Effect of sulphate and ammonium ions on the cement stone corrosion. Proc. XX Congr. JUDIMK, Cetinje, 1996, pp. 97-102.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
FLY ASH AS THE BASIC MATERIAL BINDERS
365
FOR INORGANIC
PRODUCTION
M. Ili6, S. Mileti6, R. Djuri~,i6
IMS Institute for Materials Testing, Bul.vojv.MiYiCa 43, Belgrade, Yugoslavia
Abstract Among several other possibilities for the fly ash utilization in building materials industry and civil engineering (cement production, concrete production, blocks production etc.) another way to utilize this material is to make inorganic binder of it. This paper presents results of our investigation of lignite fly ash from thermopower plant for the cement and the inorganic binder production. Results are satisfying regarding quality requirements for Portland and masonry cement. This kind of materials could be produced with the fly ash as main or replacing raw material. Key words- Fly ash, deposit, building materials
1. Introduction For several years, organizations connected with the building industry have been involved in research on energy conservation in Portland cement concretes. These organizations have been encouraging the use of less energy-intensive materials, specifically pozzolans such as fly ash as admixture or partial replacement for the relatively more expensive Portland cement. In addition to volume/mass replacement, these pozzolans can react chemically with the calcium hydroxide of the Portland cement to produce CSH gel, which is cementitious and contributes to the strength increase of the Portland cement TM. Today, all of the world have one problem yet. This is a problem about solid waste, great number of landfill with solid waste and problems about minimization of waste and recycling and utilization, too. Thermo-power stations, which are using coal as the fuel, mostly have installed very efficient equipment for preventing the emission of solid particles (fly ash) to the atmosphere. Electrostatic precipitators with 3 to 5 fields collect about 99% of solid particles. For our country it is about 10,000,000 t per year of this material. Wetted material is then transfered to deposit places which presents very big ecological problem 5-7. By definition, fly ash is a fine powder of mainly spherical, glassy particles having pozzolanic properties and consisting essentially of SiO2 and A1203. Fly ash is obtained by electrostatic precipitation of dust-like particles from the flue gases of furnaces fired with pulverized coal. Deposit places are mostly very profesionally arranged and protected, but there are still deposit places which are unconvenient for such aim. The aim of our investigations was to establish procedure for deposits of waste fly and bottom ash investigation for their safe utilization in building materials production. Reaching this aim means reducing enormous ecology problem so as the getting equal quality products (cement, mortar, concrete, bricks, etc.) 89 ' . Our basic9 aim was
366 to establish procedure for safe utilization of deposited waste fly and bottom ash due to their chemical composition, radioactivity, unburnt carbon, demands. If it is about the utilization in cement production e.g., it should be noted that the activity of fly ash depends not only on its own properties, but on the physical and chemical properties of the cement employed, even within the same cement type. The fly ash should therefore be tested with the cement intended to be used in practice in mortar and concrete. Unless otherwise specified, an ordinary Portland cement should be used to test basic activity. For years the IMS Institute has been investigating possible reuse of wastes in the building materials production. Waste is unavoidable associate of power processes, chemical processes and industrial and mining operations. The biggest waste generators in our country are the thermo-power plants. Hundred of square kilometers of good soil are replaced with deposit places for fly and bottom ash. Performed investigations were based on geological testing, chemical testing, radioactivity, heavy metals content, content of impurities, physical characteristics and pozzolanic activity of mixture of deposited fly and bottom ash. Those investigations were pert'ormed on large number of deposits.
2. Materials and methods The following materials were used in our investigation: 1. Cement Cement used in this research conformed with the European specification EN 197-1 for common cements. 2. Gypsum Powdered gypsum used in this research was waste phosphogypsum conformed with the Yugoslav standard JUS B.C1.032, as the dihydrate gypsum. Gypsum was added in cement for regulation of setting time in amount of 3 % mass. 3. Mineral admixtures The mixture of fly and bottom ash, produced as the waste material and conformed with the Yugoslav standard JUS B.C1.018. was used as mineral admixture. The unit sample should be representative for the test purpose. The taking of each sample of at least 4 kg for complete testing is recommended. From this sample a laboratory sample of at least 1 kg is obtained by subdividing, such as quartering. Performed investigations were based on the following methods: 9 9 9 9 9 9 9
geological survey chemical testing radioactivity heavy metal content content of impurities physical characteristics pozzolanic activity All the testing were performed according to the following testing methods:
367 1. Determination of strength (EN 196-1) 2. Chemical analysis (EN 196-2) 3. Determination of setting time and soundness (EN 196-3) 4. Determination of fineness (EN 196-6)
3. Results and discussion
Results of geological investigations of the deposit are presented in Table 1.
Number of coreholes 17
TABLE 1 Geological investigations of the deposit Total core Average drilling Maximum drilling length, m length, m length, m 149 9.22 12.0
Minimum drilling length, m 6.5
It could be concluded that the deposit is covered with sufficient number of coreholes and enables the right evaluation of deposit in such manner to establish the opencast mining in proper way. Our investigations were based on chemical testing, radioactivity, heavy metals content, content of impurities, physical characteristics, and pozzolanic activity of fly ash. Those investigations were performed within very long period so the variation could be estimated of some characteristics in the function of time and place of origin. We performed laboratory, semi-industrial and industrial tests with this material as the mineral admixture for cement production too. Composition of fly ash from deposit is given in Table 2. TABLE 2 F1~r ash composition Composition Silicon dioxide (SiO 2) Aluminium oxide (A120~) Ferric oxide (FelOn) Calcium oxide (CaO) Magnesium oxide (M~O) Sulphur trioxide (SO~) Sodium oxide (Na20) Potassium oxide (K20) Loss on i~nition Insoluble residue
Mass percent 28.44-37.93 8.25-12.37 7.11-9.20 20.42-24.48 2.01-3.40 1.29-3.56 0.20-0.35 0.30-0.50 13.68-24.44 26.84-52.36
The results summarized in Table 2. indicate that the fly ash constituents up to 100 % varies but the main parameters mostly satisfy the requirements of JUS B.C1.018. Loss on ignition is out of limits given in mentioned standard but it is due to hydrated water bonded to ash particles and not to unburnt carbon, and so it could be used. It is evident from the chemical analysis of the fly ash that the variations of some constituents (A1203, CaO, LOI) are large, but it could be said that besides those variations, chemical composition satisfies the requirements of JUS B.C1.018 for the most important constituents. No contamination was evident.
368 Radioactivity of fly ash was, also, measured because it is very important by point of view of ecology and health, and this value is 0.398-0.520 Bq/kg. It satisfies the national limits. The study therefore suggests that partial replacement of cement by fly ash enables utilization of deposits as waste materials. Pozzolanic activity is determined and presented in Table 3. TABLE 3 Pozzolanic activity of fly ash Pozzolanic activity, Flexural strength Compressive strength
MPa 0.4-2.4 1.1-9.2
It is generally known that fly ash addition in cement results in lower initial strengths and with constant increasing in time. Activity of this kind of fly ash is rather low but it doesnt affect the basic characteristics of cement produced with the addition of 30 % fly ash. Chemical composition of cement produced with 30 % of fly ash are given in Table 4. TABLE 4 Chemical composition of cement with 30 .% of fly ash Composition , Mass percent SiQ 15.98 5.15 A1203 2.05 Fe203 CaO 43.11 Insoluble residue 23.69 1.25 Moisture at 105 ~ C Loss on ignition 3.31 0.74 CO 2 in CaCOz CaO free 0.12 CaO in CaCO~ 0.94 1.07 CaO in CaSO 4 1.53 SO~ in CaSO 4 CaS 0.00 MgO 1.01 Alkalies as Na20 0.28 K20 0.27 MnO 0.05 FeO 0.71 0.06 P205 C1_
The results indicated that the composition of cement satisfies the requirement of JUS EN 197-1 for common cements. The content of the minor oxides such as P205, MnO, MgO, alkalies is so small that they do not affect the cement properties. Physical-mechanical properties of same cement are given in Table 5.
369 TABLE 5 Physical-mechanical properties of cement with 30 % of fly ash Properties Unit Value % Sieve residue at 0.09 mm 4.40 Specific surface (by Blaine) 3790 Density g/cm 3.00 % Standard consistence 28.70 Setting time min -initial 165 -final 330 Soundness (Le Chatelier) mm 1.0
cm~/w
Properties Flexural strengths Compressive strengths Shrinkage Heat of hydration
[
Unit MPa MPa mm/m J/g
[
3 days 3.5 12.5 -0.019 166.0
7 days 5.1 20.9 -0.125 250.0
TABLE 6 Chemical composition of masonry cement Composition, % mass BM SiO 2 + Insoluble residue 19.70 A120~ 3.22 Fe~O~ 2.97 CaO 53.67 P20~ 0.06 Moisture at 105 ~ C 0.82 Loss on ignition 0.00 15.84 CO2 CaO flee 0.53 SO~ 1.77 MgO 1.51 K20 0.17 MnO 0.04 0.14 Na20 S 0.00 Alkalies as Na20 0.25 Water soluble sulphates SO~ 1.43 1.72 SO 4 Water soluble alkalies Na~O 0.14 K20 0.17 Alkalies as Na20 0.25
28 days 7.1 35.0 -0.737 -
90 days 8.1 41.1 -0.812
NM 21.97 2.98 1.52 48.87 0.05 0.89 0.00 19.96 0.86 2.14 1.61 0.07 0.04 0.10 0.00 0.15 1.20 1.44
0.04 0.06 0.08
370 The properties of cement, which are presented in Table 5. indicate that this cement exibits lower density, small increase in water demands, good strengths and low heat of hydration due to fly ash addition. In other words, this cement satisfies the requirements of JUS B.C1.013 for the cement with the low heat of hydration. Flexural and compressive strengths were measured at 3 days, 7 days, 28 days and 90 days according to the procedure laid down in JUS. Three specimens were tested at the end of each curing period.The results obtained for compressive and flexural strength make it possible to formulate a hypothesis as to the role played by the different percent of fly ash. Masonry cements (signed as B M and NM) produced with the fly ash as main constituent are presented by characteristics presented in Tables 6 and 7. Chemical characteristics shown in Table 6. are usual for such kind of material. Soluble constituents are rather low. TABLE 7 Physical-mechanical and mechanical properties of masonr~r cement Properties Unit ] BM NM % Sieve residue at 0.09 mm 0.50 1.64 Specific surface (by Blaine) 5511 6060 Density g/cm 2.92 2.81 % Standard consistence 26.3 24.7 Setting time min -initial 210 105 -final 315 225 Soundness (Le Chatelier) mm 0.0 0.0 % Entrained air content 14.0 17.5 % Water retention 79.5 89.6
cm~/w
Water / binder 0.50 0.533
Strength, MPa 7 days 28 days flexural compressive flexural compressive
Flow, mm
BM
NM
BM
NM
BM
NM
BM
NM
BM
NM
164 180
158
3.9 3.9
2.0 -
15.7 14.0
6.6 -
5.5 5.1
3.2
21.4 19,9
10.9 -
Physico-chemical and mechanical characteristics of both masonry cements presented in Table 7. are usual for such kind of material, although fly ash is used as main constituent. Both samples of masonry cements produced with fly ash as main constituent satisfies the requirements of Yugoslav standard JUS B.C1.010 for such kind of material. Based on presented results, the maps of quality of some characteristics and the project for opencast mining were made. Environment protection project for opencast mining and dispatch stations were also made, concerning measures for air and noise pollution. The production of above mentioned cements started few years ago very sucessfully.
371 4. Conclusion
Results of large scale investigations on various deposit places of waste mixtures of fly and bottom ash led us to following conclusions and procedure: 1. 2. 3. 4.
Classification of wastes geared to their possible reuses as input material. Basic geological examinations. Detailed geological examinations. Chemical, physical, mechanical testings and examination of possible harmful ingredients (radioactivity, heavy metals, etc.). 5. Monitoring of some properties in function of time or space namely origin. 6. Laboratory, semi-industrial and full scale tests in users plants. 7. Design of opencast mining. 8. Design of environment protection during opencast mining and dispatch. 9. Design of production divisions for preparing waste material to be used in some production. 10.Mining and dispatch. 11.Quality assurance and constant technical surveillance. In this way all possible negative consequences of inadequate use of waste material are eliminated. It must be said that each deposit have to be examined by different procedure due to differencies between deposit places and origin of fly and bottom ash. That was our conclusion based on very large scale of examination on deposit places. Based on very large number of investigations from which only small part are shown in this paper, it could be concluded that the variations in quality of fly ash are large enough, so the safe utilization of this material could be done with permanent testing and separation.
References
1. Mileti6, S., Stefanovir, M., Djuri~ir, R., Proc. WASCON'91, Confon Environmental Implications of Construction with Waste Materials, NOVEM, Maastricht, Netherlands, (1991), p.4. 2. Davis, R., ASTM Spec. Techn. Publ. No.99, p. 3. 3. Zivanovi6, B., Djoki6, S., Mileti6, S., Ka~arevir, Z., Tehnika, 5, Beograd, (1984), p.6. 4. Ramakrishnan, V., Coyle, W.V., Brown, J., Tlustus, P.A., Venkataramananujam, P., Effects of Fly Ash Incorporation in Cement and Concrete, Materials Research Society, Pittsburgh, (1981), pp. 233-245. 5. Popovi6, K., Dimic D., Kameni6, N., Krstulovi6, P., Miletid, S. Se~un, Lj., Proc.RILEM Intern.Symp.Test Quality and Quality Assurance in Testing Laboratories for Construction Materials and Structures, Paris, France, (1989), p.5. 6. Dragirevir, Lj., Zivanovi6, B., Miletir, S., Zbornik radova, I Jugoslovenski simpozijum o keramici, II, SHD, JUDIMK, Beograd, (1981), p.5. 7. Berry, E.E., Malhotra, V.M., ACI Journal, Proceedings Vol.77 (2), (1980), p. 59 8. Pearson-Gallovay, "Civil Engineering", N.Y.7/8, 1950. 9. Sersale, R., Proc. of 7th Inter.Congress on Chemistry of Cement, Paris, Vol. I, IV-l/3, (1980)
This Page Intentionally Left Blank
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
373
A study of the potential of utilising electric arc furnace slag as filling material in concrete Catharina B~iverman and Francisca Aran Aran 1 Department of Chemical Engineering and Technology Royal Institute of Technology Stockholm, Sweden Abstract Sand in concrete has been substituted by steel slag from an electric arc furnace. The physical properties investigated were break load and compressive strength. The two materials showed similar results. The leaching properties of the materials were also investigated to study whether the steel slag concrete could be acceptable from an environmental point of view. The leaching properties of both materials showed similar results, except for chromium leaching. The release rate of chromium has been calculated, and the results show that the chromium leaching should be no problem.
Introduction and background Large quantities of waste material are produced annually in the steel industry. The use of these secondary materials in civil engineering applications is of great interest nowadays, both to decrease the amounts landfilled and to replace natural aggregates like sand and gravel. The utilisation of these materials is only interesting if there are no environmental effects caused by the substitution. Recycled scrap iron generates a slag with a high metal content. The slag used in this study comes from a scrapmetal-based steel industry, and is an electric arc furnace (EAF) slag. EAF slag is presumed not to have any binding properties useful in concrete production, and can, therefore, be used to replace natural aggregate in concrete. The aim of the study was to investigate whether steel slag can be used in concrete as filling material to substitute sand. The influence of steel slag in concrete was investigated and compared to normal concrete. The steel slag concrete was also compared to steel slag to study the influence on leaching of cement in contact with the slag.
Material and Methods The slag used in this study was an EAF slag from Fundia Steel AB, Sweden. The slag was crushed and sieved to give a particle size distribution similar to that of natural sand.Two types of concrete were prepared: a normal concrete with natural sand as aggregate, and a steel slag concrete with steel slag as aggregate. The preparation technique was the standard technique for preparing concrete specimens. The specimens were kept in a moisture chamber at 100% humidity for 7 days, the first day in the mould, and at 50% humidity for 21 days.
Physical properties The physical properties investigated were compressive strength and breaking load in order to investigate whether steel slag concrete could make a good product. The steel slag concrete had a higher density than the normal concrete. The compressive strengths and breaking loads of the two different types of concrete were similar, and they can both be considered as medium strength concrete. The steel slag concrete was, however, more brittle than the normal concrete.
Leaching tests The leaching properties of the steel slag concrete were studied to see whether the material is acceptable from an environmental point of view. Batch leaching tests were performed on crushed material with a particle size less than 0.16 mm, at constant pH, 9.5, 10.5, 11.5, 12.5 and 13.5, at liquid-to-solid ratio of 5 and at pH 9.5, 11.5 and 13.5 at a L/S ratio of 100. Samples were taken after one and seven days. A second series of experiments performed on the steel slag concrete and on the normal concrete involved batch leaching tests using slabs. This test can be used to study the diffusion, from deeper portions of the sample. Similar tests have earlier been used for diffusion measurements in granite (~. Concrete slabs, 6*40*40 mm, were made. Ten slabs were put in a Teflon holder and placed in a vessel, see figure 1. A volume of 300 ml of water was added to the vessel. The leachate pH was kept at 9.5, 11.5 or 13.5 by Resident in Barcelona, Spain.
374 the addition of nitric acid or sodium hydroxide. The vessels were closed and samples were taken after 1, 2, 4, 8, 16 and 32 days.
Figure 1.
Ten slabs of concrete were put in a Teflon holder and placed in a vessel.
Analyses The samples were analysed with ion chromatography. The ion chromatography system consists of a Dionex model DX-300, with both suppressed conductivity detection and post-column reaction with UV/VIS-detection. The content of Ca, K, Mg, Mn, S, Co, Cr, Cu, Zn, C1, Pb, Ni were determined.
Results and discussion
Crushed sample The natural pH of the leach water in contact with the samples was the same for the two concrete samples, whereas it was one pH unit lower for pure steel slag. This shows that the final pH of the leachate is controlled by the cement and that the substitution did not considerably affect the final pH. As the pH in our experiments was manually adjusted by the addition of acid, we observed that it was difficult to keep a constant pH, due to the high pH buffering capacity of the materials. This problem did not occur when the pH was kept at a higher pH than the initial pH. The buffering reactions are probably the dissolution of oxides, mainly lime (CaO). This effect was the same for both steel slag concrete and normal concrete. Chromium was the only element of those analysed for which the leach pattern differed significantly between steel slag concrete and normal concrete. The chromium concentration in the batch experiments is shown versus pH in figure 2. The results show that the chromium leaching from the steel slag concrete follows the leach pattern for steel slag at high pH, above 12, but that at a pH below 12 the leaching from the steel slag concrete is higher than that from both the steel slag and the normal concrete. The minimum chromium leaching from the steel slag concrete was observed at the natural pH of the material. The concentration at p is in the same range as that of the maximum level for Swedish drinking water, which is 0.9 ~M <2~. 8-
~, . . . . . . . . . . . . . . . . . . . . . . . . ~ ~ m i ......
6-
|
m
Normal Concrete
9 ....
~ i | Steel Slag C o n c r e t e ' ~: w: m: m."
|
i
q ' .....
I
i
"m- -
Steel Slag
|
......\..
j
S
~: ,:
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9
10
11
!
!
12
13
14
pH
Figure 2.
Chromium concentration in the leachate for the different batch experiments versus pH of the experiment. The samples were taken after 1 day.
375
Slabs In the experiments with concrete slabs, the concentrations in the leachate from the normal concrete were similar to the concentrations in the experiments with crushed material at the corresponding pH. There were no significant differences between the samples taken at various times. In the leachate from the steel slag concrete, the chromium concentration for the experiments at pH 9.5 and 11.5 was constant at about 0.5 laM, independent of the time at which the samples were taken, whereas at pH 13.5 the chromium concentration showed a strong time-dependence, shown in figure 3. The release rate was 1" 10 -6 mol/(m2*day). It would take about 50 years to dissolve all the chromium in this material if the release rate is constant. The solubility of chromium at this pH is of the order of 10 .2 mol/dm 3(3), and it is thus far from saturation.
4
g "~
3
6
8
2
f
1
0
1~0
2~0
3~0
4~0
Time (days)
Figure 3.
The concentration in the leach water in the leaching test at a pH of 13.5. The release of chromium was time dependent at this pH.
Leaching mechanisms The time dependence of the chromium release is shown in figure 3. After the initial quick rise, the leaching rate is constant. The pore diffusion does not limit the leaching in this case. The leaching rate found experimentally is typical of a case where the dissolution or reaction rate is the controlling factor. A comparison of the surface area exposed to leaching in the two experiments showed that the finegrained particles had a specific surface area at least 86 times larger z than the slabs (outer surface). The leached amount was, however, only 14 times larger (in g/kg concrete) from the ground material than from the slabs, after 7 days of leaching. Assuming that the surface area of the particles is 86 times larger than that of the slabs and that the dissolution rate is the same for slabs and fine particles, when the solution is far from equilibrium, these results show that the outer exposed surface of the slabs contributes 16% of the leaching and that the exposed inner surface must be 5 times larger than the outer surface. This is the same as though the outer 0.8 m m depth of the concrete slabs contributes to the leaching as if it is fully available for leaching.
General discussion The main variable controlling the leaching of heavy metals is the final pH of the solution (4). This was not affected by the addition of steel slag instead of sand to the concrete. The leaching of most elements was not affected, with the exception of chromium, as the leaching was similar for both concrete types. The greater leaching of chromium from steel slag concrete compared to that from normal concrete may not be an environmental problem as the leaching at the normal pH (about 12.5) is very close to the drinking water limit and cannot, therefore, be considered to be a problem. The release rate, even under extreme conditions (pH 13.5 shown in figure 3) is very low.
2 The surface area, in this discussion, is calculated assuming that all particles are cubes with a 0.16 mm side. The smaller particles contributes with a larger surface area. There is, therefore, reason to believe that the specific surface area is larger than has been assumed.
376 If the steel slag concrete were used to make a swimming pool (25m* 12m*2m) and the walls were not coated, it would take more than three years with no change of water, and no cleaning either, to exceed the drinking water limit for chromium, assuming a constant release rate and that the release rate is similar in neutral water as at pH 13.5. This is, of course, not true, but it is improbable that the release rate is larger at neutral pH, or that the rate increases with time. It is, therefore, probable that it would take much longer than three years in a real situation to exceed the drinking water limit.
Conclusions In this study, steel slag was used to replace sand as an aggregate in concrete. The compressive strength and break load were the same for both materials but the steel slag concrete broke into much smaller pieces. The leaching from the steel slag concrete and from regular concrete are similar for all metals except chromium. The release rate is, however, low for chromium so it should not be an environmental problem. The release is lowest at the pH that is natural for the concrete, i.e. about 12.5. The results of both the physical study and the leaching tests show that this steel slag can well be used as a substitution for sand in medium strength concrete. There are, as far as we can see, no environmental risks associated with the utilisation. The benefits are that virgin material is saved and that the steel slag does not have to be deposited in landfills.
Acknowledgements The authors thanks Gunnar Klingstedt and Lennart Magnusson at the Cement and Concrete Institute in Stockholm for help in the preparation of the concrete samples and in the measurement of physical properties of the specimens. They have also contributed to the work in valuable discussions. We also thank Dr Luis Moreno and Professor Ivars Neretnieks at the department of Chemical Engineering and Technology at the Royal Institute of Technology for their contributions to this work. This work was financed by the Swedish Environmental Protection Agency/AFR.
References 1. Skagius, A-C.K. Diffusion of dissolved species in the matrix of some Swedish crystalline rocks, PhD thesis, Department of Chemical Engineering, Royal Institute of Technology, Stockholm, Sweden, 1986. 2. SLV FS 1989:30, Statens livsmedelsverks kung6relse om dricksvatten, ISSN 0346-119X, 1989. 3. Stumm, W. and J.J. Morgan, Aquatic Chemistry, Chemical Equilibria and Rates in Natural Waters, Third Edition, John Wiley & Sons, ISBN 0-471-51185-4, 1996. 4. Chandler, A.J., T.T. Eighmy, J. HartlEn, O. Hjelmar, D.S. Kosson, S.E. Sawell, H.A. van der Sloot, J. Vehlow, International ash working group, International perspective on characterisation and management of residues from municipal solid waste incineration, summary report, December 1994.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997Elsevier Science B.V. All rights reserved.
377
PROPERTIES OF P O R T L A N D C E M E N T MORTARS INCORPORATING HIGH AMOUNTS OF OIL-FUEL ASHES.
J. Payfi, M.V. Borrachero, J. Monz6, M.J. Blanquer and E.Gonzfilez-L6pez. Grupo de Investigaci6n en Quimica de los Materiales de Construcci6n (GIQUIMA) Departamento de Ingeniefia de la Construcci6n Universidad Polit6cnica de Valencia c/Camino de Vera s/n 46071 Valencia (Spain)
Abstract
The residue of oil-fuel burned at the electrical power plant of Grao de Castell6n (Spain) has been incorporated in Portland cement mortar and concrete. The used oilfuel ash presented a high percentage of magnesium compounds because of magnesium oxide addition for removing slag and ashes into boiler and pipes. Several researches had been carried out on stabilization of toxic metals also occurring in oil-fuel ashes (specially vanadium and nickel), by mixture with coal fly ashes and cement. In our case, the presence of magnesium compounds in the composition of the studied oil-fuel ashes could alter the mechanical and chemical properties of cement matrix in fresh and hardened mortar and concrete. We present here the chemical, physical and mineralogical characterisation of oilfuel ashes and the behaviour of Portland cement mortars incorporating high amounts of these oil-fuel ashes: workability, water demand, setting time, expansion and and compressive strength developments. Preliminary results demonstrated a high absorption of water on oil-fuel ash particles, which promotes an increasing of water/cement ratio for a given workability; a setting acceleration of Portland cement/oil-fuel ash pastes was observed, due to the presence of carbonates. On the other hand, no significant expansions in specimens due to the presence of magnesium compounds were detected and, consequently, mechanical properties of hardened mortars containing oil-fuel ashes did not drop with curing time, but compressive strength for mortars containing OFA were very lower than control mortar ones.
Introduction
The combustion of oil-fuel at electric power plants yields oil ashes as a solid residue. Oil-fuel ash (OFA) composition depends on, logically, the nature of the oil fuel burned and the additives used for removing them from the boiler and pipes. Oil fuel ashes produced by combustion of oil-fuel are enriched in heavy metals, specially in vanadium, which would be recovered by the steel industry ~. However, in many cases, the residue is simply stabilised for avoiding leaching processes which could release toxic elements to the environment. Stabilization of oil ash wastes has been carried out within fly ash/cement/lime matrix; the stability of the solid blocks prepared also has been investigated from structural and chemical aspects 2, as soon as fouling community development when materials were used as concrete reefs 3.
378 Composition of OFA will be altered by the use of additives for removing ashes and slag from the boiler and pipes, as for example, using magnesium compounds (magnesium oxide and magnesium hydroxide); in addition to changes in chemical composition, new compounds will be synthesised by reactions between additives and oil fuel ash residue. These compounds could play an important role on the final properties of the stabilising matrix in mixtures within cement and other materials (for fresh and hardened materials). Finally, if OFA has been exposed to atmospheric agents, several reactions (carbonation, hydration, dissolution, selective dissolution, ion exchange,...) would be promoted and the nature of OFA compounds substantially altered.
Experimental section
Oil fuel ashes were obtained from the electric power plant of Iberdrola in Grao de Castell6n (Spain). A representative sample of OFA was chosen and dried at 105 ~ for 24 hours. Part of the sample was ground using a laboratory ball-mill within alumina balls 4. Cement was an ASTM type I Portland cement; a natural sand was used for preparing mortars (3.56 modulus fineness). Sikatard was used as setting retarder (phosphate-based additive) and Sikanol-M as plasticizer. OFA samples were studied by several techniques: Thermogravimetric analysis (TGA850 Themogravimetric Measuring Module, Mettler-Toledo), Scanning electron microscopy (JEOL JSM-6300, equipped with microanalysis based on energy dispersive X-ray), X-ray diffraction (PW1710 Based Diffractometer) . Preparation of mortars, setting determination, and mechanical strength measurements were carried out according to the corresponding ASTM procedures. Workability of mortars was determined using a flow table, measuring the spreading of mortar cones 5 (FTS values).
Results and discussion
Morphological chemical and mineralogical characterization. OFA sample before grinding was made up of particles between 10 lam and 3 cm in diameter. After grinding, fineness of the OFA sample was similar to ordinary Portland cement one; electron microscopy studies on ground OFA samples showed that crushed particles were very irregular in shape, with diameter from 1 to 40 ~tm and particles showed rough surfaces (see Figures 1 and 2), and surface chemical composition was determined by means X-ray dispersive energy: sample showed a high content in magnesium; nickel, vanadium, iron, silicon and calcium were also present; and sulphur also was detected. Figure 3 shows X-ray dispersive energy surface elemental analysis pattem of ground OFA. X-ray diffraction analysis of OFA demonstrated that the sample was a complex mixture of compounds; several crystalline substances were identified6: olivine-like compounds (forsterite Mg2SiO4, MgxFe2.xSiO4, fayalite Fe2SiO4), periclase-like compounds (periclase MgO, MgxNi~.xO), bunsenite NiO), magnetite-like compounds (magnetite Fe304, magnesioferrite MgFe204, trevonite NiFeO4), vanadates (magnesium vanadium oxide Mg3(VO4)2, nickel vanadium oxide Ni3(VO4)2), hydrated halloisite (A12Si2Os(OH)4nH20) and magnesium hidroxide sulphate hydrate (Mg4(OH)6SO48H20). X- ray diffractogramm for OFA was showed in Figure 4. Thermal analysis of OFA permitted to determine the presence of other important substances; Figure 5 shows the thermogravimetric TG and DTG curves for OFA. Weight loss near 110 ~ was attributable to moisture and loss between 100 and 250 ~ was due to hydration
379 water; weight loss in 250-400 ~ range was attributed to water loss due to hydroxide groups, and weight loss in 400-600 ~ range was attributed to CO2 evolution due to carbonate anion decomposition. Probably, as thermogravimetric analysis suggests, OFA contains several substances, which were not detected by X-ray diffraction analysis, as magnesium carbonatehydroxide hydrates (MgCO3)x(Mg(OH)2)ynH20. The study of thermogravimetric analysis of substances as (MgCO3)4(Mg(OH)2)5H20 permitted to assign the corresponding losses.
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380
2000
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M: M a g n e t i t e - l i k e c o m p . O: Olivine-like e o m p . P: Perielase-like c o m p . V: V a n a d a t e c o m p . H: Halloisite S: M a g . H y d r o x i d e - s u l p h a t e
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5 10 15 20 25 30 35 40 45 50.55 60 65 70 75 80 2O(Degrees) Figure 4. X-ray diffractogramm of OFA sample
DTG
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200
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300 20
500
400
30
40
600 50
700 60
800 70
Figure 5. TG and DTG curves of OFA sample
900 ": 80
90
*C m.tn
381
Constancy of volume for Portland cement/OFA pastes The presence in OFA samples of magnesium oxide and sulphate compounds could produce dangerous expansion processes of cement paste due to hydration of MgO to form brucite:
MgO + 1-I20~ Mg( OH) 2 or due to reaction of sulphates with calcium aluminate hydrates to yield ettringite 7. Expansive processes in cement/OFA pastes were monitored using Le Chatelier needles, measuring the distance between the indicator needles after a thermal treatment in water. Expansion values found for pastes with different cement/OFA ratios are summarized in Table 1. In all cases, the distance between indicator needles did not exceed 10 mm, suggesting that expansive reactions did not affect appreciably the volume of the pastes. Consequently, cracks due to expansion processes will not occur when using OFA in mortar and concrete, and structural stability is guaranteed in relation to internal attack processes.
Table 1. Distances between indicator needles (Le Chatelier needles) for different Portland cement/OFA pastes. Portland cement (%)
OFA (%)
cement/OFA ratio
Distance between needles (mm)
85
15
5.67
2.3
70
30
2.33
3.5
55
45
1.22
1.6
40
60
0.67
2.1
Setting time of cement/OFA pastes Usually, ordinary Portland cement contains gypsum as retarder; so, setting time is controlled by chemical interaction between calcium aluminate hydrates and calcium sulphate dihydrate. The presence of other compounds in cement could affect its setting, since these substances can act as retarders or accelerators. In the setting process of cement, two periods are distinguished, the initial set and the final set 8, and the Vicat needle is used almost universally. When OFA is added to Portland cement, an acceleration of setting was observed; so, if the percentage of OFA is equal or greater than 30 %, a flash set is caused. Setting times for cement/OFA pastes are summarized in Table 2. Setting acceleration observed when OFA is mixed with Portland cement could be attributed to the presence in OFA of magnesium hydroxide carbonates (MgCO3)x(Mg(OH)z)ynH20; these compounds would interact with tricalcium aluminates forming carboaluminate compounds 9. So, an intense crystal formation accompanied by crystal bonding, yielding an strong three-dimensional network is achieved.
382
Table2. Setting time for cement/OFA pastes OFA (%)
Water for defined consistence (%)
Initial set (minutes)
Final set (minutes)
Setting time (minutes)
0
28
82
113
31
15
31
30
122
88
30
33
10
39
29
45
38
12
25
13
60
44
13
17
The addition of a phosphate-based retarder (Sikatard) in cement/OFA pastes (0.1% of cement+OFA weight) increased notably initial and final sets. Probably, the precipitation of calcium-phosphate salts ~0hinder carboaluminate network crystallization. Table 3 summarizes setting time values for cement/OFA pastes using this retarder.
Table3. Setting time for cement/OFA pastes using a setting retarder OFA (%)
Wmerfordefined consistence(%)
Initial set (minutes)
Final set (minutes)
Se~ingtime (minutes)
15
31
165
265
100
30
33
120
165
45
45
38
60
125
65
60
44
60
120
60
Workability of cement/OFA mortars Control mortar was prepared mixing 450 g of Portland cement, 225 mL of water and 1350 g of natural sand, whereas OFA replacing mortars were prepared replacing part of the cement by ground OFA, with variable water/(cement+OFA) ratios. For studying the influence of OFA on workability of cement mortars, flow table spread (FTS) of mortar cones was measured according to the already reported method 5. OFA replacing mortars (15-60 % replacement of cement by OFA) were prepared with 0.5 w/(c+OFA) ratio and FTS values are depicted in Figure 6. FTS values for mortars prepared by the same way and using plasticizer (Sikanol-M, 0.1% in weight of the total amount c+OFA) also are represented in Figure 6. It can be noticed that the presence of OFA decreases strongly the workability of mortars. When plasticizer was not used, replacing percentages greater than 15 % producen poorly workable mixtures; the addition of plasticizer enhances workability only for 15 and 30 % OFA replacing mortars. For enhancing workability of mortars containing high volumes of OFA, greater w/(c+OFA) ratios were necessary. So, in Figure 7, FTS values for mortars containing 45 and 60 % of OFA are represented (using plasticizer and increasing w/(c+OFA) ratio). Morphology and roughness of ground OFA particles and, probably, adsorption of water on particle's surface, are responsible for decreasing workability of mortars.
383
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160
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140
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120 100 0
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30
45
60
OFA replacing percentage Figure 6. FTS values for mortars containing OFA (w/(c+OFA)=0.5)
180 -
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160 140 120
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loo 7" 0.5
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0.7
w/(c+OFA) ratio Figure 7. FTS values for mortars containing 45 and 60 % replacing percentages
Strength development of cement/OFA mortars Prismatic mortar specimens (16x4x4 cm) were prepared and, after 24 hours at 20 ~ in a moisture room, were cured at 40 ~ until test age. Mortars were prepared using retarder and plasticizer. Figure 8 shows compressive strength development of mortars containing ground OFA; w/(c+OFA) ratios are indicated for each replacing percentage. It can be noticed that values for mortars containing OFA showed very lower compressive strengths than control mortar one. This fact could be attributed to three factors: firstly, the increase on w/(c+OFA) ratio specially for mortars containing high amount of OFA; secondly, the presence of some compounds found in OFA may hinder the adequate development of the cementitious matrix; and, thirdly, the use of plasticizer could increase the air content with an irregular distribution. However, no decreasing on compressive strength was observed in any case with curing time, suggesting that dangerous expansive reactions in the cementitious matrix were not occur.
384
5o t
control
40- - C o n t r o l (0.5)
30-. 20
o~ r/3 r~ d~
I
--15 %(0.5)
15%
r~
- - 30 % (0.56) --45%(0.61)
101
30 %
45 %
o
0
14
7
2'1
-.-60%(0.67)
2'8
Curing time (days) Figure 8. Compressive strength development for cement/OFA mortars. For each mixture w/(c+OFA) ratios are given in parentheses
Better mechanical properties were obtained when OFA was used without grinding and after sieving ( for obtaining a material with particles less than 5 mm in diameter, coarse particles were crushed and sieved; retained material on sieve was finally discarded). This sieved OFA was used for replacing part of the sand (10-30 % in weight). So, mortars were prepared as follows: 450 g of cement, 0.5 water/cement ratio and 1350 g of sand+OFA. Specimens were molded and cured in the same conditions than above, and compressive strength measured at 28 days curing time. Figure 9 shows the relationship between compressive strength and the replacing percentage of sand by sieved OFA. Despite w/c ratio and the total amount of cement are maintained constant for all prepared mortars, also compressive strength value decreased with increasing replacing percentage; this fact suggest that OFA altered the cementitious matrix, and, consequently, mechanical properties of cured mortar worsen.
50
~
40 -
9-~
30
o
20 . 0
.
.
. . 5
.
10
15
20
25
30
OFA replacing percentage Figure 9. Compressive strength values for mortars containing sieved OFA replacing sand (40 ~ curing temperature, 28 days).
385
Conclusions Based on the experimental reported in this paper about the incorporation of oil-fuel ashes (OFA) in cement mixtures, the following conclusions are made: 1- Ground OFA showed irregular particles with rough surfaces; magnesium is the main metallic element, and nickel, vanadium, iron, silicon, calcium and sulphur also were detected. Different compounds were identified by X-ray diffraction and thermogravimetric analysis: olivine-like compounds, periclase-like compounds, vanadates, magnetite-like compounds and magnesium hydroxide-carbonate compounds.
2.- No expansive processes were detected in OFA/cement pastes, and structural stability due to intemal attack is guaranteed. 3.- An acceleration of setting is produced due to the presence of OFA in cement pastes, and when high percentages of OFA are used a flash set is caused; the use of setting retarder permitted increasing initial and final sets of OFA/cement pastes. 4.- Morphology and surface roughness of OFA particles are responsible to decreasing workability of mortars containing OFA. Workability was enhanced using a plasticizer, but an increasing of w/(c+OFA) ratio is necessary for mortars containing high OFA replacing percentages. 5.- Mortars containing ground OFA showed very lower compressive strengths than control mortar ones, although no drop in compressive strength was observed with curing time. 6.- The substitution of sand by sieved OFA permitted to obtaining mortars with better mechanical properties, but compressive strength decreased with the increase of replacing percentage of sand by sieved OFA.
Acknowledgement We would like to express our gratitude to Iberdrola (Grao de Castell6n power plant), SIKA S.A. (Valencia office) and Cementos Asland (Puerto de Sagunto plant) for their support for this research project.
References 1.- S. Akita, T. Maeda, H. Takeuchi, Recovery of Vanadium and Nickel in Fly Ash from Heavy Oil. J. Chem. Tech. Biotechnol., 1995, 62, 345-350 2.- C.S. Shieh, I.W.Duedall, Chemical Behaviour of Stabilized Oil Ash Artificial Reef at Sea, Bull. Mar. Sci., 1994, 55, 1295-1302 3.- W.G. Nelson, D.M.Savercool, T.E.Neth, J.R.Rodda, A Comparison of the Fouling Community Development on Stabilized Oil-Ash and Concrete Reefs, Bull. Mar. Sci., 1994, 55, 1303-1315
386 4.- J. Pay~i, J. Monz6, M.V. Borrachero, E. Peris-Mora. Mechanical Treatment of Fly Ashes. Part I: Physico-Chemical Characterization of Ground Fly Ashes. Cern. Con. Res., 1995, 25, 1469-1479. 5.- E. Peris-Mora, J. Pay~i, J. Monz6. Influence of Different Sized Fractions of a Fly Ash on Workability of Mortars. Cern. Con. Res., 1993, 23, 917-924. 6.- V. Primo. DRXWIN v1.3, A Graphical and Analytical Tool for Powder XRD Patterns. 7.- I. Odler, I. Jawed. Expansive reactions in Concrete. Materials Science of Concrete, vol II. J. Skalny and S. Mindess Ed., The American Ceramic Society, 1991, pp. 221-247. 8.- F.M. Lea. The Chemistry of Cement and Concrete. Chemical Publishing Company, New York. 3rd Edition, 1971, pp. 362-364. 9.- J.I. Bhatty. A Review of the Applications of Thermal Analysis to Cement-Admixture Systems. Therrnochimica Acta, 1991, 189, 313-350. 10.- V.S. Ramachandran. Admixture and Addition Interactions in the Cement-Water System. Il Cemento, 1986, 83, 13-38.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
387
THE USE OF F L Y A S H TO I M P R O V E T H E C H L O R I D E R E S I S T A N C E OF C E M E N T M O R T A R S
Joseph G Cabrera, G R Woolley and K E Hassan ~
~ ~, .
Civil Engineering Materials Unit (CEMU) Department of Civil Engineering University of Leeds, Leeds, .LS2 9JT, UK
ABSTRACT This paper presents a laboratory study on the strength and chloride penetration resistance of fly ash and ordinary portland cement concretes and mortars. The concrete mixes were 30/20 structural concrete with different workabilities with target slump values of 50, 75 and 125 mm. The mortar specimens were sieved from the site concrete mixes and cured for various ages before exposure to a sodium chloride water solution. Two laboratory techniques of chloride exposure were used to assess the behaviour of opc and opc/fa mortars. These were the intermittent test and the continuous capillary absorption test. The exposure to chloride salt in the laboratory was an attempt to simulate site exposure conditions, i.e. splash zone in road bridges and direct contact with wet contaminated soil. In each case the chloride concentration profile was obtained at different exposure periods and used for the calculation of chloride diffusion coefficients. The results of the study confirm the unique advantage of fly ash as a workability enhancer, i.e. reduction of the water content without impairing the concrete workability. The results show that the concrete with and without fly ash achieves approximately the same strength at the age of 28 days. It is found that the coefficient of diffusion is not dependent on the method to salt exposure and most importantly that even at short ages the use of fly ash enhances the chloride penetration resistance of concrete. This technological advantage is enhanced by the environmental impact caused by using fly ash which, on the one hand reduces the energy required to produce concrete, and on the other, avoids the disposal of fly ash as inert waste.
388 Introduction
In many civil engineering materials applications the use of industrial waste is not only environmentally desirable, but it is beneficial in technological terms, since wastes used in appropriate circumstances can and do improve the properties of materials for construction. Durable concrete demands resistance to aggressive environments. Reinforced concrete not designed tbr the appropriate environmental conditions may sustain chemical or/and physical damage, which in the event leaves the reinforcing steel unprotected and susceptible to corrosion. The resistance to fluid and ion penetration is directly related to the permeability and diffusion properties of concrete. In turn these properties are strongly affected by the pore structure characteristics, the degree of hydration and the nature of the cement hydrates. A durable concrete is defined as that which exhibits adequate and expected performance during its design life. This definition is not related to the environmental conditions and therefore, it is implicit within it that the performance level required increases as the environment becomes more aggressive. Designing for performance not only requires designing for strength, but also designing for resistance to fluid and ion penetration. The measurable parameters are then intrinsic permeability and diffusion coefficient. Permeability is measured by monitoring the rate of fluid flow which occurs due to a pressure gradient (1). It is calculated by assuming nonturbulent flow and using the Darcy's numerical model (2). Because ions move within concrete in a solvated state, permeability gives a good indication of the resistance of concrete to ion penetration (3). Diffusion is defined as the process of ion movement through a solid which is caused by a concentration gradient. The rate of ion movement is used to calculate the diffusion coefficient by using the first (steady state) or second (non-steady state) Fick's Law. In practice the movement of ions occurs through a combination of pressure and concentration gradients and therefore it is important to be clear on what it is that a test indicates with reference to the potential durability of concrete. Chloride ions in solution penetrating concrete may reach the reinforcing steel and initiates a process of corrosion which evantually leads to the loss of steel section and reduction of serviceability of the reinforced concrete (4, 5). Exposure to chlorides occurs by: a)
b) c)
continuous exposure of the concrete surface to salt laden water, by intermittent splashing of salt water, by chlorides carried by wind in a solvated state, particularly close to sea shores.
These different exposure modes may result in different chloride concentrations at the concrete surface and depending on the moisture content of the concrete, might penetrate the concrete by diffusion and/or permeation with the additional complication that if the concrete is dry, capillary water absorption will occur.
389 To reduce the rate of chloride penetration, concrete should be designed for minimum total porosity particularly by reducing the capillary voids (6, 7) and by providing a "chemical trap" so that chlorides can be immoblised by reacting with the hydrated phases forming the walls of the concrete voids. It is now accepted that the best method to reduce the rate of ion penetration into concrete is to use water reducing agents and/or pozzolanic or hydraulic additions (8, 9). This paper evaluates the effect of fly ash on the chloride resistance of concrete exposed to salt water and presents diffusion measurement data under two simulated exposure conditions: a) intermittent splashing, b) capillary-osmosis absorption The paper explores the effect of fly ash replacement on the water demand of concrete designed to different workability levels. Finally. it discusses the environmental and durability consequences of using fly ash to produce concrete.
Materials and Sample Preparation Specimens for the study were prepared from in-situ structural concrete mixes used in the construction of Drax power station in the North of England (10). The mixes contained 33% (by weight~ fly ash as cement replacement. Three specific workabilities of 50, 75 and 125 mm, as measured by the slump test, were used in this study. For control purposes, opc concrete mixes of similar workabilities were also made and tested. Details of the different mix constituents are given in Table 1. The effect of fly- ash inclusion on reducing the water demand without impairing the workability of concrete has been acknowledged for a long time. In fact the mix design concept proposed by Cabrera in 1985 (11) and used by Hassan et al (12) in a recent design procedure is based on the effectiveness of fly- ash to reduce water demand without impairing workability. Hassan et al showed that a 10% replacement of opc by fly ash will reduce the water demand by 3-4%. The mixes designed for this study show that the reduction of water to maintain constant workability- (see Table 1) increases when the target workability increases, i.e. for the mix with 50 mm slump 10% opc substitution gave a 4% water reduction, while for the mix with 125 mm slump, the water reduction per each 10% substitution was 6.4%. Mortars were sieved from the different concrete mixes and cast into 100 mm cubes. The mortar cubes were cured for a period of either 3 or 28 days according to BS 1881 (13). The cubes were then stored in an environmental chamber maintained at 20~ and 65% relative humidity until the age of 77 days prior to chloride exposure.
390 P R O G R A M M E OF TESTING
Compressive Strength Mortar and concrete cubes (100 mm side) were used to measure compressive strength for the different mixes studied in this programme. The test was carried out in accordance with BS 1881: Part 116 (14).
Intermittent Splashing An apparatus was designed to discharge by fine overhead spray a small amount of 3.2% sodium chloride and water solution, this is shown in Figure 1. The mortar cubes were placed on an open grid set in a shallow bath. Mounted overhead were 12 No. spray heads connected to a header tank and compressed air supply. A bulk storage tank with internal agitator contained a 3.2% sodium chloride solution which continuously filled the smaller header tank. An electrically time control system caused the overhead spray to operate at 15 minute intervals, dousing the cubes arranged below. The solution then ran to waste. The splashing test was completed in a stable 20~ environment. Each series of tests, which ran through 84 days, contained an equal number of fly ash and ordinary Portland cement cubes. The chloride concentration profiles were measured at exposure ages of 14, 28, 56, 70 and 84 days.
Capillary Osmosis Absorption Test For the continuous exposure condition a simple metal bath was prepared and fitted with an internal plastic support. The mortar cubes were placed upon the support and all adjacent faces of all cubes and the cube/tank sides were sealed with a polymer sealing compound to exclude air. A clear plastic tube was fitted to the bath to allow filling to the underside of all cubes with a 3.2% solution of sodium chloride. This arrangement is shown in Figure 2. At intervals of 14.28.56.70 and 84 days the cubes were tested tbr chloride content at different depths so that a concentration profile could be obtained.
Chloride Concentration Profile After the desired exposure period, the cubes were removed from the chloride environment and carethlly wiped dry with clean tissue, before sample drilling. Powder samples were collected from the surface down to depths of 0-10 mm, 10-20 mm and 20-30 mm using a Hillti hammer with 12 mm diameter drill. The procedure was repeated for cubes prepared with 50, 75 and 125 mm slumps. The powder samples obtained from each experimental set up were mixed with boiling water and tested for chloride content using the Quantab titration methods (15).
391 The Quantab titration method used in this work was checked against X-ray spectrometry on samples taken from the mortar cubes. Figure 3 plots results of the two methods giving a correlation of 0.714. When major outliers were excluded a correlation of 0.894 was obtained. This confirmed that the Quantab Titration method is an acceptable method for determining chloride concentration in this experimental series. RESULTS
Compressive Strength The compressive strength results, moisture contents and densities of the concretes and mortars are given in Table 2, the compressive strengths are also presented graphically in Figure 4. Typically the lower water/cementitious ratio mix gave higher compressive strength and the opc mixes gave high compressive strength when compared with the pfa mixes at 7 days and comparable strengths at 28 days. It is clear that the difference in strength is narrowed down with time from 7 to 28 days and from other research (11, 17) it is expected that at a later age the pfa mixes will eventually give higher strength than the opc mixes.
Chloride Penetration The chloride concentration profiles for the different opc and opc/fa mortars were obtained after 84 days of exposure to chloride environments either by splashing or capillary absorption. Typical examples for the chloride concentration versus depth are shown in Figure 5 and 6. The results show that the chloride concentration decreases across the mortar depth. The chloride distribution with depth of the specimen is a decay function, which can be statistically represented by the following expression: C1 = ae ba
(1)
Where: C1 d a,b e
= = = =
chloride concentration mortar depth constants natural logarithm base
The chloride diffusion coefficients are then calculated using the following expression: ~ - -
Co
e
x
392 Where: = Co = Cx D = t = erf
the chloride concentration at the surface of the specimen the chloride concentrations at depth x of the exposed surface of the specimen diffusion coefficient time the error function (16)
The diffusion coefficients of the different opc and opc/fa mortars cured for 3 days or 28 days are listed in Table 3. DISCUSSION W a t e r D e m a n d and Strength
Replacement of a proportion of cement with fly ash modifies the rheology of plastic concrete allowing a reduction in water demand for the same workability (17). Based on the results given in Table 1, a statistical equation with r 2 = 0.99 was obtained. This equation can be used to determine the approximate water reduction for any target slump when designing trial mixes. The equation is: Rw = 9.7 + 0.08 (S)
(3)
Where: Rw S
= water reduction (%) = slump (mm)
This equation is presented graphically in Figure 7. Since for any concrete the water/cementitious ratio limits the maximum strength achievable (11) it should be expected that in the short term (28 days) the compressive strength of the fly ash mixes should equal the strength of the equivalent opc mixes, since in reality, there is a trade off between reduction of opc and reduction in w/c + fa ratio, i.e. less opc lower strength, but lower w/c+fa ratio higher strength. In the long term (beyond 28 days) the reaction between fly ash and the lime generated by the hydration of cement reduces the porosity of the concrete and most importantly reduces the average pore size, leading to higher strength and improvements in performance related parameters (11, 18). Table 2 shows that the strength of the opc and opc+fa mixes at 28 days is statistically approximately the same. It also shows the effect of increasing the water cementitious ratio. An increase from 50 to 125 mm slump (2 89times) reduces the strength by one third on the opc concrete, but only by a quarter in the opc-fa concrete. This is the result of the lower water cementitious ratio of the fly ash concretes and mortars.
393 Chloride Diffusion
The chloride concentration on the surface of the specimens tested was found to be different in the two tests carried out. The surface chloride concentration on the specimens subjected to capillary absorption were higher than the chloride concentrations of specimens subjected to the intermittent splash experiment. However, the most important finding was that the diffusion coefficient calculated from any of the tests carried out was approximately the same and was not affected by the estimated surface concentration. Figure 8 shows that statistically the diffusion coefficients are the same for both modes of exposure. Thus the idea that the exposure mode affects the value of the chloride coefficient of concrete is not supported by the findings of this study. The results presented in Table 3 show that even at short age, before the beneficial effect of the fly ash lime reaction has taken place, the chloride diffusion coefficient of the opc concrete is reduced by at least 25% for low workability (50 mm slump) to 42% for the high workability (125 mm slump) when fly ash is used. This reduction in terms of enhanced service life is considerable, to say the least. Apart from this obvious technological advantage, fly ash that is not utilised is disposed of in landfill sites for which there is a cost, this becomes unnecessary if all available fly ash is incorporated into concrete. The energy reduction arising from the reduction of cement content is also considerable as has been shown in reference (19). CONCLUSIONS From the results of the short term laboratory experiments the following conclusions are offered: Fly ash replacing part of the ordinary portland cement in concrete reduces the amount of water required to maintain a target workability value. This reduction increases as the workability increases. .
.
.
Concrete mixes designed to a target workability show that at 28 days of age, 33% replacement of opc by fly ash gives approximately the same compressive strength as opc mixes. The chloride diffusion coefficient of opc concrete is reduced from between 25% to 42% by the substitution of opc by fly ash. The beneficial effect of fly ash is enhanced for mixes with high water-cementitious ratio. The incorporation of fly ash in concrete results not only in a better durable material but reduces the energy required to produce concrete, thereby reducing pollution levels arising from the high temperatures required to produce clinker. Furthermore, the problems associated with the disposal of fly ash in landfill sites are potentially eliminated.
394 REFERENCES
THE C O N C R E T E SOCIETY. Permeability testing of site concrete - A review of methods and experience. Report of Concrete Society Working Party, on Permeability of Concrete and its control. Technical Report 31, The Concrete Society, London, 1987. CABRERA J G AND LYNSDALE C J. A new gas permeameter for measuring the permeability of mortar and concrete. Magazine of Concrete Research, Vol 40, No 144, pp 177-182, 1988.
M I D G L E Y H G AND ILLSTON J M. The penetration of chlorides into hardened cement pastes. Cement and Concrete Research, Vol. 14, 1984, pp. 546-558. .
BUENFELD N R. Chlorides in concrete. Construction Repair and Maintenance, January 1986, pp.7-10. CABRERA J G AND GHOUDDOUSSI P. The effect of reinforcement corrosion on the strength of the steel/concrete "bond". Bond in Concrete RILEM International Conference. Topic 10; Proc. Environmental Influence on Bond, pp. 10-11. Riga, Latvia. October 15-17, 1992. CABRERA J G, DODD T A H AND NWAUBANI S O. The effect of curing temperature on the chloride ion diffusion of superplasticised cement and fly ash cement pastes. ACI, SP-139 Durable Concrete in Hot Climates, pp 61-76, 1993. CABRERA J G. The measurement of concrete porosity. Concrete Research Seminar, University of Leeds, 1984. BAMFORTH, P B. Concrete classification for R C Concrete structures exposed to marine and other salt-laden environments. In Structural Faults and Repair- 93. Edinburgh 1993. BUENFELD N R AND NEWMAN J B. The permeability of concrete in a marine environment. Magazine of Concrete Research. Vol. 36, No. 127, pp. 67-80, June 1984.
10.
W O O L L E Y G R. Construction of structures at Drax power station using pfa concrete. International Symposium on the Use of pfa in Concrete. Editors J G Cabrera and A R Cusens. Vol. 1, pp 313-321, Leeds 1982.
395 11.
CABRERA J G. The use of pulverised fuel ash to produce durable concrete. How to make today's concrete durable for tomorrow. ICE, UK, Thomas Telford, pp 141150, London, 1985.
12.
HASSAN K E, CABRERA J G AND BAJRACHARYA, Y M. The influence of fly ash content and curing temperature on the properties of high performance concrete. To be published in proceedings of 5th International Conference on Deterioration and Repair of Reinforced Concrete in the Arabian Gulf. Bahrain, October 1997.
13.
BRITISH STANDARDS INSTITUTION. Testing Concrete - Part 111, Method of normal curing of test specimens (20~ method), London, BS 1881" Part 111" 1983
14.
BRITISH STANDARDS INSTITUTION. BS 1881" Part 116, Methods for determination of compressive strength of concrete cubes. BSI, London, 1983.
15.
LEES T P. Field tests for chlorides in concreting aggregates. Construction Guide. Cement and Concrete Association, Wexham, 1982.
16.
CRANK, J. The mathematics of diffusion. Second Edition. Clarendon Press, Oxford, 1975.
17.
CABRERA J G AND HOPKINS C J. The effects of pfa on the rheology of cement pastes. International Symposium on the Use of pfa in Concrete. Editors" J G Cabrera and A R Cusens. Vol. 1, 1982, pp 141-150.
18.
FRAAY A L A, BIJEN J M AND MDETTAAN Y. The reaction of fly ash in concrete: a critical examination. Cement and Concrete Research,. Vol. 19, 1989, pp 235-246.
19.
CABRERA J G AND W O O L L E Y G R. Life cycle benefits of calcium silicate replacements. Proceeding of the Intemational Symposium on Inert Waste" An Opportunity for Use, Editors: J G Cabrera and G R Woolley, Leeds, 1995, pp 215220.
396 Table 1. Mix constituents of opc and OPC Concrete Slump w/c opc sand gravel (mm) ratio 1220 50 0.52 300 700 1220 75 0.56 300 700 1220 125 0.62 300 700
opc/fa concrete. OPC/FA concrete w/c opc fa sand ratio 0.45 200 100 700 0.47 200 100 700 0.49 200 100 700
gravel 1220 1220 1220
reduction in water content (%) 13.5 16.1 21.0
Table 2. Compressive strength of opc and opc/fa mortars and concretes. At 7 days of age 50 mm slump
opc mortar opc/fa mortar opc concrete opc/fa concrete 75 mm slump opc mortar opc/fa mortar opc concrete opc/fa concrete 125 mm slump opc mortar opc/fa mortar opc concrete opc/fa concrete
I
Compressive strength (MPa) 49.1 37.7 47.5 40.3
Moisture Content (%)
Wet Density (kg/m3)
6.35 6.70 5.00 4.50
43.8 36.7 39.6 34.2 33.0 23.8 32.0 25.6
At 28 days of age Moisture Content (%)
Wet Density (kg/ms)
2260 2249 2440 2434
Compressive strength (MPa) 60.8 55.0 57.6 49.6
5.85 6.20 5.00 4.50
2250 2249 2443 2420
7.10 7.40 4.60 4.80
2219 2230 2432 2422
50.9 48.2 52.3 48.1
6.70 7.15 4.00 4.50
2235 2217 2448 2417
7.40 7.45
2203 2228 2468 2423
47.8 43.6 42.6 37.5
6.90 6.65 4.50 5.5
2213 2208 2415 2400
.
Table 3. Non-steady state chloride diffusion coefficients. OPC/FA mortar Slump OPC mortar x loll(m2/sec) (mm) x 10"ll(m2/sec) 28 days 3 days 28 days 3 days curing curing curing curing SPLASH TEST 50 2.31 2.59 1.82 2.29 75 1.79 1.40 1.66 1.57 1.81 125 4.05 3.50 2.77 1.80 CAPILLARY 50 2.00 2.38 2.34 1.52 ABSORPTION 75 1.47 2.07 1.28 2.11 TEST 125 5.01 3.42 2.85
397
, j ~ ' "
m
,9
:'
~a
Figure
1.
Intermittent
~.. . . . .
...
. .,
~
,~
: ',; ~
.,~..:.
. . . . "
, ~,.,"
splashing
apparatus.
;.
.~;~!. .
.
,~.
.~.~...
, : e , . ~ . :,.~. ?.,~.:...,. ..,~,~
' ~,
,
.,
...':.~
~"9
". ;'.'
..
2~;
,~-..
....." ": ~"~ ~':~ :~.~!i:.. . . . . .
.:-
~:
.
r
~t~;~::~,~'.: ....
~ ..~.~fi~:?
Figure
,
'
2.
Capillary-osmosis
absorption
set-up.
,
~ ....
.
~.~
r ....
.,.
..,. 9 il
398
/ / / I 600
E
o_ o_
500
~
>..
-J/
. ~D"
[D
E 8
4OO
t
*6
Q
el.. u~ 3OO
I
"
c o t)
Points marked omitted
(~
200
@ o or..)
100 I
|
@
I
I 200
I00
Chloride
I 300
contenf
%.J
-
I 400
I 500
Qucnfab
fifration
I 600
I 700
(ppm)
Figure 3. Relation between chloride results from spetrometry and quantab titration.
60
f~ o
c-
~ {/3 ~n
slump
75ram
slump
50
40
125ram
Gb
-,-v)
50ram
30 ff
iI
iI
iI
tt
II
II
f
20
Q_
E
o
10
I
7
I
Age (days)
28
Figure 4. Compressive strength of opc and opc-pfa concrete mixes.
slump
399
600
500
F
o._ o._ 400
\
\
\
300
200
100
I
1
5
I
10
1
15
I
20
25
Devth
,
!
I
.50
50
(-"qm)
Figure 5. Typical chloride profile of specimens exposed to the splashing environment. 600
..-..
500 \
E f_l.. 400 c rD
u
300
o c~
200
\\
\\\\
100
,'0
1'~
2'0 DepTh
2'~
3'0
(mm)
Figure 6 Typical chloride profile from specimens exposed to the capillary osmosis environment.
s'o
400 20
18
16
iz. r 0
() -C (]) ,r'Y
i2
10
I
30
F
.50
I
70
I
90
!
1 10
130
Slump value ,,mm) /
Figure 7. Relation between reduction in water content (arising from substitution of opc by fa) and slump value of concrete.
6
_
Y.
~n
4
X c"
9 ,~ " ~ ~ L ~ m '
3
of
equality
O
f'"
s
r 0)
2
._
E L_ if) -,-r
1
0
I
1 Capillary
I
2 osmosis
1
I
3 absorpfic",
I
4 (X
5
10 -1'
'm2/sec)
I
6
F igure 8. Relationship between chloride diffusion coefficients obtained from splash and capillary absorption tests
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
LOW
LIME
BINDERS
MORAVIA
BASED
ON
401
FLUIDIZED
JAN DROTTNER* - SILESIAN POWER PLANT O S T R A V A , 7 0 9 74
JAROMIR HAVLICA TECHNICAL UNIVERSITY OF B R N O , 637 00
BED
ASH
CO.
BRNO
ABSTRACT
Cement, s t e e l m i l l and e n e r g y i n d u s t r y are a f t e r o r g a n i c mass decomposition the s e c o n d major sources of CO 2 emission which create, w i t h o t h e r gases, g r e e n h o u s e effect, m o r e o v e r , the w a s t e d i s c a r d e d f r o m the last ones is still not u t i l i z e d to the extent the e n v i r o n m e n t a l p r o b l e m s of t o d a y call for. The r e s e a r c h team tested several s o u r c e s of f l u i d i z e d bed combustion (FBC) ashes with a goal to i n c o r p o r a t e t h e m into manufacture of b u i l d i n g elements. Some of the FBC ash samples showed distinctly different both chemical and mineralogical composition from the prevailing majority indicating good potentials even as a c o n s t i t u e n t of a l t e r n a t i v e s i l i c a binder. This fact m i g h t l e a d t o w a r d s c o n s i d e r a b l e s a v i n g s of t r a d i t i o n a l cement or lime which production exploits dwindling natural deposits and t h r o u g h thermal treatment of l i m e s t o n e (together with a fuel) r e l e a s e s large a m o u n t s of CO 2 to the a t m o s p h e r e . In q u e s t i o n are a s h e s f r o m i n s t a l l a t i o n s w h e r e low g r a d e coals or coal w a s t e are b e i n g c o m b u s t e d . INTRODUCTI ON :
Since the nineties, our research t e a m has been exerting efforts to solve the energy and building industry problems of h a n d l i n g and p r o c e s s i n g s o l i d r e s i d u e s f r o m the f l u i d i z e d bed c o m b u s t i o n (FBC) t e c h n o l o g i e s . In m a n y a r e a s w o r l d w i d e one can m e e t w i t h d e p o s i t i o n s of low g r a d e coal or coal w a s t e (slurry) w h i c h w e r e e i t h e r n e g l e c t e d due
Senior Materials products handling
Engineer, program and u t i l i z a t i o n
Associated Professor, of c e m e n t i o u s p h a s e s
field '
of
leader
in
thermodynamics
FA and
and FGD kinetics
402 to the s t a t e - o f - t h e - a r t of energy generation technique by 'its shipment costs or simply for e n v i r o n m e n t a l aspects. This is the case of the L i m b u r g Area, the Lorraine Area, the Upper Silesian Basin to name only some. Fresh wind of the world energy crisis in the seventies a t t r a c t e d the a t t e n t i o n to the f o r g o t t e n energy d e p o s i t i o n s with a goal to d e v e l o p t e c h n o l o g i e s to make the most from the least. The FBC installations meet the goal, moreover, the reduction of SO x and NO x is a n o t h e r evident feature. The a d d i t i o n of SO 2 sorbent at the c o m b u s t i o n process gives rise to larger amount of solid residues to be handled, d e p o s i t e d or b e t t e r utilized than it is in case of c o n v e n t i o n a l coal boilers. FLUIDIZED
BED
ASH
- MINERALS
Fuel feed consists of a p o w d e r e d coal and sorbent calculated a c c o r d i n g to the sulphur content in the coal to create conditions for the r e a c t i o n s taking place at t e m p e r a t u r e around 850 ~ when decomposition of sorbent and oxidation of the sulpherous c o n s t i t u e n t s from coal are c o m p l e t e d CaO + SO + / 0 ~CaSO 4 C a l c i u m sulfate in t~ e form2 of 2 anhydrite II remains fixed in the ash cutting down the SO x emissions by 96 %. As the combustion temperature is kept well below the d e c o m p o s i t i o n temperature of anhydrite II also a glass phase present in c o n v e n t i o n a l high t e m p e r a t u r e ashes doesn't orginate. Clay m i n e r a l s in the coal are t h e r m a l l y d e c o m p o s e d during the c o m b u s t i o n p r o c e s s to yield the ash h y d r a u l i c and pozzolanic reactivity. The SO 2 sorbent is dolomite or ~ r e f e r e n t i a l l y limestone w h i c h u n d e r g o e s d e c o m p o s i t i o n w i t h i n 650 ~C to 850 ~ PHASE
COMPOSITION
OF
ASHES
Rough sand like bed ash creating the fluidized bed is c o l l e c t e d s e p a r a t e l y from fly ash. Due to their chemical and mineralogical differences (9), the applications may differ s u b s t a n t i a l l y (i0). In the ash collected from AFBC and CFBC installations chemical analyses confirm presence of free lime, in PFBC i n s t a l l a t i o n the free lime is replaced by c a l c i u m carbonate. The main component of AFBC and PFBC ashes is anhydrite, the content of which exeeds sometimes 20 per cent. Further v a r i a b l e amount of quartz, hematite, magnetite and unburnt carbon are being detected. High content of AI20 originates from thermal d e c o m p o s i t i o n of clays obtained in ~he coal. With the circulated and e s p e c i a l l y with the p r e s s u r i z e d t e c h n o l o g i e s contrary to the a t m o s p h e r i c ones increase of a residential time of a combustible p a r t i c l e in the fluid bed p r o d u c e an ash of almost total carbon burnout (Table i). These set facts make the FBC technology a d o p t a b l e for the c o m b u s t i o n of low grade fuels as coal waste, coal slurry, oil shales, wooden chips, peat, etc.
403 In The Nederlands, the fluidized bed ashes are employed in p r o d u c t i o n of low grade b i n d e r s - calcined p r o d u c t s (4) which comply the building industry demands put on the end product. The morphology character of fluidized bed ashes is different from conventional fly ashes due to the c o m b u s t i o n temperature (5, I0, ii). FBC ash p a r t i c l e s are of irregular shape, g l a s s y phases and m u l l i t e are m i s s i n g alltogether, there is no e v i d e n c e of cenospheres. Specific surface area is higher d e p e n d i n g on the type of s e p a r a t o r installed. It is also w o r t h to mention that d i f f e r e n t fractions differ not only by fineness but also in its chemical and m i n e r a l o g i c a l composition. REACTIVITY
OF
THE
FBC
ASHES
These ashes exhibit cementitious properties. During hydration processes, lime and anhydrite react with AI203 y i e l d i n g e t t r i n g i t e 3CaO . AI203 . 3CaSO 4 . 31H20 d e c r e a s i n g the initial h i g h e r pH value (6, 12). This reaction d o e s n ' t take place with the PFBC ash or with ashes where the free CaO content is negligible. Second reaction w h i c h c o n t r i b u t e s to the h a r d e n i n g process is the f o r m a t i o n of an a m o r p h o u s C S ( A ) H gel. Third reaction which may be observed with mixes reach of c a l c i u m sulphate leads to h y d r a t i o n of a n h y d r i t e II to gypsum. Fourth reaction taking place in systems containing periclase, magnesium oxide (from dolomite sorbent) is slow hydration to brucite. W o r t h note is that hydraulic properties of ashes w i t h low content of either reactive CaO or CaSO 4 may be modified (PFBC ash, low grade coal feed installations) by adding the m i s s i n g reagent and thus stimulate the ettringite or CS(A)H gel formation. From this statement one may conclude a wider scope of utilization possibilities w i t h systems w h e r e the e t t r i n g i t e complex is e i t h e r m i s s i n g or is c o n t r o l l e d to the acceptable extent. Systems c o n t a i n i n g e t t r i n g i t e phase alone as a h y d r a u l i c binder are not known. RELEVANT
STANDARDS
The FBC ashes exhibit h y d r a u l i c and p o z z o l a n i c p r o p e r t i e s producing hydration products close to other c o n s t i t u e n t s added to p o r t l a n d clinker listed in ENV 197-1. These c o n s t i t u e n t s are clearly defined, results from long term d u r a b i l i t y tests and field a p p l i c a t i o n are available. Moreover, fly ash in concrete constructions is governed by EN 450 supervised by EN 206 standard. These standards, its s t i p u l a t i o n s and l i m i t a t i o n s were combined at our r e s e a r c h work. Conclusions from research affect the standard tests required, m a n y p r o c e d u r e s have to be modified. As these FBC ashes are s t e p i n g out of the line of materials tested in b u i l d i n g industry thus, the existing standards ~ s t i p u l a t i o n s create an obstacle for their utilization. This goes for L.O.I. - the p r o c e d u r e puts f o r w a r d false data due to d i s s o c i a t i o n of c a l c i u m carbonate and m a g n e s i u m carbonate.
4 0 4
I n s t e a d of d e t e r m i n a t i o n of u n b u r n e d c a r b o n s h o u l d be set next to carbon dioxide in case of presence of c a l c i u m c a r b o n a t e or m a g n e s i u m c a r b o n a t e . If h i g h level of iron o x i d e is detected, then d e t e r m i n a t i o n of b i v a l e n t iron is r e c o m m e n d e d . As quality parameters of ash from p l a n t to p l a n t differ, it seems to be p r u d e n t the s t a n d a r d l i m i t a t i o n s s h o u l d be laid d o w n on the final c o n s t r u c t i o n or c o n s t r u c t i o n e l e m e n t in phase of its p r o d u c t i o n . P o z z o l a n i c a c t i v i t y p l a y s an i m p o r t a n t role in a s s e s s m e n t of materials e m p l o y e d . The determination by EN 450 fulfills its purpose, rapid thermochemical test largerly adopted may considerably s p e e d up the q u a l i t y assessment phase of the ash s o u r c e (7,8). E
X
P
E
R
I
M
E
N
T
A
L
Based on p r e l i m i n a r y tests the recent r e s e a r c h a c t i v i t i e s w e r e f o c u s e d on the e v a l u a t i o n of a b i n d e r c o n t a i n i n g s i g n i f i c a n t a m o u n t of FBC ashes t o g e t h e r with other industrial by-products originating in the industrial part of Czech R e p u b l i c - the O s t r a v a r e g i o n in a civil e n g i n e e r i n g as a r o a d b a s e c o n s t r u c t i o n material. The following by-products and a c t i v a t o r s w e r e e m p l o y e d as basic constituents: Ordinary Portland cement CEM I 42,5 with specific area of 364 m 2 / k g (Blaine). Ground granulated blast furnace slag (GGBFS) w i t h specific area of 320 m ~ / k g (Blaine). FBC fly ash w i t h s p e c i f i c area of 415 m 2 / k g (Blaine) c o l l e c t e d by electrostatic percipitator from the p o w e r plant Tginec, w h e r e coal m i n e w a s t e s are c o m b u s t e d (Figure la). G r o u n d n ~ t u r a l g y p s u m (60 % of C a S O 4 . 2H20) w i t h s p e c i f i c area of 310 m /kg (Blaine). Crystalline blast furnace slag (CBFS) 0 - 8 mm. The o x i d i c c o m p o s i t i o n of the s t a r t i n g m a t e r i a l s is g i v e n in the T a b l e 2. From blends of these constituents, mortars of identical workability (slump) w e r e p r e p a r e d . The w o r k a b i l i t y was set by mortar A with w a t e r to b i n d e r r a t i o of 0,50, no p l a s t i c i s e r was a d d e d in any of the blends. The initial setting time of the respective b i n d e r was determined according to the procedure described in EN 196-3 standard. The s o u n d n e s s Le C h a t e l i e r test a c c o r d i n g to the EN 196-3 on e a c h b i n d e r was c a r r i e d out. The SO 3 c o n t e n t in each of the b i n d e r was d e t e r m i n e d by EN 196-2 standard. Test mortar specimens 40 x 40 x 160 m m were prepared, demolded after 24 h o u r s and w a t e r cured. A f t e r d i f f e r e n t curing time the f l e x u r a l and c o m p r e s s i v e s t r e n g t h t o g e t h e r w i t h the bulk density were determined. In m i x D (Table 3a) a part of FBC m i n e r a l s was r e p l a c e d by e q u i v a l e n t part of gypsum. A p r e f e r e n c e of the same w o r k a b i l i t y w i t h all t e s t e d blends over the same water-to-binder ratio (as the EN 196-1 s t a n d a r d requires) was preferred due to the selection of p r o c e s s i n g "Roller Compacting" technology. -
-
-
-
-
405 Addition of g y p s u m in m i x D should s i m u l a t e the increase of the SO 3 content in FBC m i n e r a l s and its effect on the b i n d e r quality, i n c o r p o r a t i o n of i n d u s t r i a l p o z z o l a n a like GGBFS should prove its impact on compressive strength at later ages. C o m p o s i t i o n of the blends with all s u m m a r i z e d data are p r e s e n t e d in Table 3a. Compressive strength development is shown in Figure 2. The 90d c o m p r e s s i v e s t r e n g t h s led for the s e l e c t i o n of the blend marked "C" to its further modification for the given p u r p o s e - R o l l e r C o m p a c t i n g S t a b i l i z e d Road C o n s t r u c t i o n . As this c o n s t r u c t i o n is not i n t e n d e d to be a top finish layer there is no necessity to carry out the d e i c i n g salt resistance and f r e e z e - t h a w r e s i s t a n c e tests admitting bringing those interesting data with similarities as associated with c e m e n t - f l y ash composites. A gravel u s e d in the laboratory prism tests was a CFBS 0-8 mm a g g r e g a t e w i t h a poor d i s t r i b u t i o n curve in the area of the fine particles. A m i x t u r e m a r k e d H as shown in Table 3 was c o m p a c t e d by v i b r a t i o n and cured for three days in m o i s t room, then d e m o u l d e d and kept untill t e s t e d in wet b u r l a p in p l a s t i c bags. To improve che d e f i c i e n c y of fine p a r t i c l e s in the CFBS aggregate, i n c r e a s e d p o r t i o n of GGBFS was i n c o r p o r a t e d into the mix. C o m p r e s s i v e and flexural s t r e n g t h s are l i s t e d in T a b l e 3 b. Simultaneously the binder itself was tested from the m i n e r a l o g i c a l point of view. HYDRATATION,
MINERALOGICAL
TRANSITIONS
The scope of this part of the r e s e a r c h was to e n l i g h t e n on the m i n e r a l o g i c a l p r o c e s s e s t a k i n g place in dry and low c a l c i u m content h a r d e n i n g composite. Components focused on were the portlandite and C S ( A ) H phases, the former to be the most susceptible phase to carbonation creating the pH barrier of h y d r a t e d p r o d u c t s against corrosion and carbonation respectively, the latter b u i l d i n g the s u p p o r t i n g s k e l e t o n of the h a r d e n e d paste. This r e s e a r c h part has b e e n split into test of standard curing c o n d i t i o n s and hydrothermal curing c o n d i t i o n s to bring forth data on the h y d r a t e d p r o d u c t s and p h a s e s c o n t r i b u t i n g to the s t r e n g t h s of the c o m p o s i t e (13). M e t h o d s u s e d were the XRD combined with DTA. Finally, from the a u t o c l a v e d test it was expected to support the evidence once d e t e r m i n e d by p r e v i o u s Le C h a t e l i e r test - the s o u n d n e s s of the composite. PREPARATION
OF T H E
PASTE
SAMPLES
The H p a s t e b l e n d 36,5 gr CEM I 42,5; 233,6 gr FBC; 206,5 gr GGBFS m i x e d w i t h 13 % of w a t e r was c o m p a c t e d into b r i q u e t e s w h i c h were m o i s t cured in the d e s i c a t o r untill tested. In case of the a u t o c l a v e test the b r i q u e t e s a f t e r 1 day in d e s i c a t o r u n d e r w e n t the 3-6-3 hours regime at 190 ~ and 12 atm.
406 COMMENTS
OF
RESULTS
Initial setting time R e p l a c e m e n t of cement by low free CaO content FBC ashes will p r o l o n g the initial setting time c o n s i d e r a b l y as a consequence of C3A and C3S r e d u c t i o n a c c o m p a n i e d by a low heat of hydration. W a t e r - t o - b i n d e r ratio Higher water demand is influenced by p a r t i c l e c h a r a c t e r i s t i c s of the FBC material.
fineness
A c t i v i t y index The value linked with the b l e n d A suggests 28d and 90d p o z z o l a n a a c t i v i t y index.
and the
excellent both
C o n s u m p t i o n of cement It ~ s evident that the use of FBC m a t e r i a l plus other p o z z o l a n a c o n s t i t u e n t s as GGBFS enhances the r e d u c t i o n of cement c o n s u m p t i o n drastically. Soundness of the c o m p o s i t e Le Chatelier tests confirmed conformity of all binders tested with the standard condition. Also the autoclave test p e r f o r m e d on the paste b l e n d H p r o v e d it's soundness. However, this test is highly recommended to be carried out with blends whenever the utilization of FBC ash is considered as the behaviour of the thermally activated clay m i n e r a l s and it's impact on d u r a b i l i t y has not yet been elucidated. Detected phases exhibiting binding p r o p e r t i e s where CSH (II) and e t t r i n g i t e in case of standard curing and CSH (I) in hydrothermal process conditions, p r e s e n c e of p o r t l a n d i t e has not been detected (Figure ib, c, d, e). Blend s e n s i t i v i t y to the temperature during h a r d e n i n g A d d i t i o n a l test focused on the H m o r t a r s e n s i t i v i t y towards the lower t e m p e r a t u r e d u r i n g curing period brought in data that limit the a d o p t a b i l i t y of the set "Roller Compacting" processing t e c h n o l o g y as a result of low cement inclusion c o n n e c t e d with low e v o l u t i o n of heat of hydration. CONCLUSIONS
:
T e s t e d b l e n d H with the FBC ash employment p r o v e d compliance with the set standards ~ stipulations for the c o n s i d e r e d processing technology. Adoptability of this technology is, however, profoundly t e m p e r a t u r e dependant. Cheap lightweight building elements manufactured at elevated t e m p e r a t u r e a c c o m p a n i e d with p r o d u c t i o n of dry m o r t a r s may be the viable p r o c e s s i n g s o l u t i o n to the e x a m i n e d FBC ash.
407 REFERENCES i. EN 450
Fly ash for control, 1994 2. E N V
Cement,
concrete,
197-i
composition,
3. E N V 206
Concrete, criteria,
specifications
performance, 1989
4. N O K
The Netherlandts 1992 5. Odler,
definitions,
and quality
and c o n f o r m i t y criteria,
production,
National
requirements,
Clean Coal
placing
and
Programme,
1992
compliance
Reports
1991,
I.
Composition, p r o p e r t i e s and possible u t i l i z a t i o n of high sulphur ashes p r o d u c e d in flue gas desulphurization, Paper 63, Tenth International Ash Use Symposium, January 1993, Vol. 2 6. M u l d e r ,
E.
Formation and d i s i n t e g r a t i o n of ettringite in a mixture of fly ashes as base road c o n s t r u c t i o n material, S u p p l e m e n t a r y Papers of the Tenth International Ash Use Symposium, January 1993 7. H a s s e t t ,
J.
e.a.
C h a r a c t e r i z a t i o n of p o z z o l a n i c cementitions m a t e r i a l s by a novel heat of hydration technique, Paper 63, llth International Symposium on Use and Management of Coal C o m b u s t i o n By-products, January 1995, Vol. 2 8. B r a n d ~ t e t r ,
J.
9. B r a n d ~ t e t r ,
J.;
Determination of the pozzolanic materials, Czech Pat No 231365/1987 Drottner,
activity
of alumino
silicate
J.
Composites based on solid residues of fluidized bed coal combustion and other by-products, Fifth CANMET/ACI International Conference on Fly Ash, Silica Fumes, Slag and Natural Pozzolans in Concrete, S u p p l e m e n t a r y Papers, p.p. 389-411, June 1995 i0.
W o m Berg,
W.;
Puch,
K. H.
Utilization of residues from fluidized VGB Kraftwerks technik 73 (4), 1993
bed combustion
plants,
ii. Odler, I.; Skalny, J. Potential for the use of fossil fuel c o m b u s t i o n wastes by the construction industry, Materials Science of Concrete, Vol III, Amer. Ceram. Soc., W e s t e r v i l l e 1992 12.
Havlica,
13.
Sauman,
J.;
Sahu,
S.
Mechanism of ettringite and m o n o s u l f a t e Res. 22, p.p. 661-667 (1992) v
Z.
formation,
Cem.
Concr.
Study of reactions between CaO or 3CaO.SiO 2 a n d ~ - 2 C a O . S i O 2 and power station fly ashes under hydrothermal conditions. Fifth Int. Symp. Chem. Cem. IV, p.p. 122-134 (Tokyo 1968) .
408
Table 1. Combustion technology and phase composition of solid residue (in mass %) 9.
.:.
.... :.
'
II
I
bed.combustion
rbe.t
9A t m o s p h e r i c
High Temperature combustion > 1200
0/C
- " Fluid~d ............
"
..
Pressurized
I:~BC_,CFBC_:.
p~C
9 860 II
Illl
I
I
5 0 - 90
Glass phase Mullite
3 -20
X - Ray amorphous
30 - 50
AS
Clay, shale Feldspar
0-1
Dead burnt lime
0-3
1 -3
1-2
1-2
0
Periclase
0-1 2 -20 2-
Magnetite Anhydrite
30 - 50
1 -3
0
5 - 22
Free reactive lime Hematite II
10
0-2
0-2
3 - 10
3 -15
4-
0-2
0-2
,,.
15
2-6
10-25
1 0 - 25
Calcite
0
0-1
I0-
Quartz
2-8
3-10
3 - 10
15
Table 2. Oxidic composition of the starting materials Method used: XRF, gravimetric analyses
SiOz,.
9.~:
":. .:i l..:.. .":.;:.::.::::i.:.[i:ji
22.46
|
41,8
37,3
.'::E!;I:-
5.13
"
21.4
6,6
7.2
~F.e,~i" ........ ......... 9 ................... 9'."".. : " i.."............. :,, :_;__.,",,....: :,?.i1
3.24
5.7
1,6
0.7
-i.At~.o.;:.....
:...: .... .7..:.:.:::
37,6
...... 9 .:.!:;
61,42
10.7
37.8
40.8
! M-~..:.: .;:..:. ~....:. : ... :..!:.,!.::/-
1.29
6.2
14,4
1{),6
I:K~O:.":.:.::.: .;...i.:........... : . ::.i:.;:.. i:::.l
0.90
2.3
0.4
0,5
_ Na,,O ~.:.!i:" :.i;. " ' .i ':"..:i: "!:i~:L
0,52
0,9
0.3
0,2
~TiOr
::!: .: .:.....)~..:.:
0,35
1,2
0,3
0,1
!MaO
"
n
0.1
0.6
0, 5
iCaO.
" :......
"
........ '.
!pLo..~..:
free C a O .
.
.
L,O,I.
!C
.
.
.
.
.
.
': .
" ....
9 ..
9 9 .
$0~... :
;:~.:::;L
.::.::."::.
.
..:......:'
.
.
.
n
0.8
0.1
0.1
2.55
3.75
1.71
1,65
0.2()
! ,48
.
.::
.. :_
9.
.
1.92
99
:
n ........ not d e t e r m i n e d
--
.............
4,9
409
Fig. I a
x ~ pattern of FBC
ash
Q
Q..Quarz A..Anhydrite II C..Calcite
A
Fig. 1b
~
pattern of H paste after 60 days of standard curing
I-I
E
CSH (II) t Q
c
Fig. 1c
E..Ettringite
T.
fl
x a D patttern of H paste after hydrothermal process
Q CSH (I) A
40
35
C
30
A
25
20
15
10
5 2 theta
Fig.
1d
Fig. 1e
DTA-TG curves of H paste after 60 days of standard curing
hydrothennal process 1000 "'C 0
" C mg lO0O 0
mg
800
800
2o 600
600 400
DTA-TG curves of H paste after
• i
400
4o
2O0
200
0 0
40
80
min
i
0
40
80
rain
410
Table 3 a) Composition of the mortars in grams, the same workability
....................................................... Rret 11~ i i i 111 Ill Illlll I I
CEM 1 42,5
n
,,,, ,A
45 "
.
C
225 225
18f)' 135 135
1350 0,5 210 5,0 2.91 2227
1350 0,59 220 5,0 3,15 2240
1350 0.51 220 5.0 2.66 2198
~c GGBFS Nat. gypsum..
H
D
.
315 135
.....
18o
CBFS O-8
1350 0,38 120 0,0 2,55 2246
Quartz sand W Initial setting time (min) Le Chatelier (mm) Binder SO3 cont (%) Unit weight, 90 d (k~, (k~/m 3)
36.5 233.6 206.5
4O 254 225
1324.4
1441
217
234
n n 2194
n n 2194
130.5 135 4.5 1350 0.53 290 0,0 2,91 2236
Table 3 b) Strengths, flexural/compressive, MPa .
R 3.26/15,7 4,12/22,2 5,94/35.2 6,66/43,7 13,5/51,6 14,1/60,3 100,0
Days
7 28 90 180 90 d. per cent
A
.
.
.
.
.
B
.
.
.
.
.
C
.
.
.
.
.
.
u 11...
D
1.29/5.1
0,60/2,2
0.40/2.0
3,36/ 15.3 6.09/28.9 9,29/49,4 12,8/53.9 12,8/59,2 104,5
1,30/5,2 3,95/17,0 7,58/34,5 11.5/45,9 12,2/51,9 88,9
1,40/5,8 4,4/19,2 7,37/36,8 11,7/47,6 12,5/49,7 92,2
. . . . . . . . . . . . . . . . . . .
.
- .....
0
H ,4 3/ i 1 9
1,28/5.6 4,47/18.5 7.05/33.3 14,5/50.1 15,3/52,0 97,1
3,14/10.4 5,16/14,2 5,9/17.0 6.4/19,0 ,-
Fig. 2 Compressive strength developement
I " l "
a~B~
"~ omo..-'"
I ,->KIB
fro ~"
.,..,
'-'-I II.-I
,~
2'0
.~
,'o
s'o
Time, days
~,
7'0
8~
I
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
STRUCTURAL PERFORMANCE OF REINFORCED CONCRETE MADE WITH SINTERED
ASH AGGREGATE
P. J. WAINWRIGHT 1 P. ROBERY 2
1. SENIOR LECTURER, DEPARTMENT OF CIVIL ENGINEERING, UNIVERSITY OF LEEDS, ENGLAND.
2. DIRECTOR, G. MAUNSELL & PARTNERS, CONSULTING ENGINEERS, BIRMINGHAM, ENGLAND
Summary Previously published work has described a process in which the bottom ash from MSW incinerators has been treated to produce a coarse aggregate suitable for use in concrete. Data has been given on the production parameters, the physical properties of the aggregate and their influence on some of the material properties of the concrete, the data presented here is an extension to that work. The structural performance of concretes made from aggregates derived from MSW incinerator residues is compared with that of concretes made from a natural aggregate and from a commercially available lightweight aggregate (Lytag). Tests were performed to measure strength, deformation, bond, load deflection and cracking characteristics of reinforced concrete beams. The results showed that overall the performance of the incinerator aggregate was similar to that of Lytag and compared favourably with that of the natural aggregate. The structural behaviour of all beams was predictable using the design standard that was applicable at the time the tests were carried out.
411
412
Introduction
Previously published work by the authors (1- 7) described a process where by an artificial aggregate was produced from the ashes derived from municipal solid waste ( MSW ) incinerators. The ashes were treated to remove all ferrous and non ferrous metals and were further crushed, blended with clay, pelletised and then fired in a kiln. The properties of the aggregates made with ashes from two different sources were compared with those of a natural gravel aggregate and a number of tests carried out to measure the performance of the material in concrete. Tests were performed over a period of over 4 years measuring such properties as :- compressive and tensile strength development, elastic modulus, shrinkage and creep. A study was also undertaken into the durability of the concretes made with the material looking in particular at the corrosion rates of reinforcement and the susceptibility to attack by the alkali silica reaction. As expected the results showed that the material did not perform as well as the natural aggregate but never the less, considering the nature of the material, its performance was considered satisfactory and more importantly there were no indications of any long term detrimental reactions taking place. In the majority of the tests undertaken in the work referred to above comparisons have been made between the ash aggregate and a natural gravel aggregate. However by its nature the ash aggregate is more like a lightweight aggregate than a natural aggregate and it was therefore decided to carry out a study comparing its properties with that of a commercially available lightweight material. Emphasis in this study was placed on the structural performance in reinforced concrete beams as well as looking at the more fundamental properties of small scale concrete specimens.
Experimental work Materials and their properties The aggregate used in this study was that manufactur~ from the ashes of the Dutch incinerator in Rotterdam as described previously ( 7 ). Comparisons were made between this material and :a). A 10 mm. single size lightweight aggregate made from sintered fly ash and sold under the trade nmne of Lytag. b). A 10mm single size, irregular, quartzitic gravel aggregate.
The relevant properties of the aggregates are shown in Table 1 Mix Proportions All concrete mixes were designed to have nominally the same strength at 28 days and similar aggregate volume proportions. The sand / cement ratio was kept constant at 2 / 1 ( by volume ) and the coarse aggregate volume concentration was 0.48 for the gravel aggregate concrete and 0.4 for the other two. The concrete made from the natural aggregate had a water / cement ratio of 0.59 and a nominal cement content of 315 kg / m 3 ; whereas for the concretes made from the ash and Lytag aggregates the figures were 0.48 and 385 kg / m 3 respectively.
413
Both the lightweight aggregates were pre-soaked for 30 minutes before casting and all concretes were made to nominally the same workability of 100mm slump. Following casting all specimens were kept under water at 20~ until testing.
Tests Performed The following tests were performed at the appropriate ages in accordance with the relevant British Standard ( 8 ) :9
Compressive strength on 100mm. cubes
9
Flexural strength, static and dynamic modulus of elasticity on 100 x 100 x 500mm. prisms.
The following non standard tests were also performed to help assess structural performance :9
Bond Strength Bond strength was determined using two methods (9) namely the Pull out test to measure anchorage bond and the Transfer test to measure adhesion bond.
9
Reinforced concrete beam tests.
The load deflection, ultimate load carrying capacity and cracking characteristics were assessed on a number of reinforced concrete beams details of which are given in Table 2. Three different reinforcement ratios were chosen to cover the range from under to over reinforced. Beams were tested over a span of 2.4m using a two point loading system. All structural design calculations on the beams were carried out in accordance with the Code of Practice that was appropriate at the time the tests were being conducted. The beams were loaded to their design load then cycled 5 times from 5kN to the design load before being loaded to failure. Measurements were taken on mid span deflection, strain distribution and crack width.
Results
Compressive / Flexural Strength Results for compressive strength are shown in brief in Table 2 and in some more detail in Fig 1. All mixes show similar strength development characteristics and were relatively close to their design strength of 40.0 N/mm 2 at 28 days. Comparing the Lytag concretes with the Ash concretes ( made with the same w/c and cement content ) it can be seen that the former was approximately 30 % stronger at 28 days which is probably indicative of a better quality material. Flexural strength results are summarised in Table 2 and it is interesting to note that, as with compressive strength the results for Lytag and gravel are similar yet unlike compressive strength the ash concrete achieves a higher strength albeit only marginally ( approx. 7% ).
414
Elastic Modulus The relationship between elastic modulus (dynamic) and strength is shown in Fig. 2; the ash and Lytag concretes show a similar relationship, but as expected, for a given compressive strength the natural gravel concrete has a modulus which is approximately 1.6 times larger than the other two. Bond Characteristics Results from a typical pull out test showing bar slip against average bond stress are shown in Fig. 3. Somewhat surprisingly the ash concrete shows a significantly lower slip for a given bond stress than the other two. The stress required to produce a given slip is believed to increase with an increase in both the mortar strength and ire elastic modulus of the concrete. Direct comparisons between the natural aggregate concretes and the other two are therefore ditticult because, on the one hand the mortar strength of the former is likely to be lower because of the higher water / cement ratio yet, on the other hand, the modulus of elasticity is higher. Comparisons between the two lightweight concretes should though be valid but no reason can be given at present to explain the significantly lower slip values of the ash aggregate concrete
The results of average bar stress against average bond stress as measured by the bond transfer test are shown in Fig. 4 and show a more logical trend than the pull out test. For a given bar stress the bond stress for the gravel concrete is approximately 40% lower than for the two lightweight concretes, this is difference is probably due to the weaker mortar strength of the former as reported earlier.
Structural Behaviour Some of the results of tests carried out to study the structural behaviour of reinforced concrete beams when loaded in the manner described previously are summarised in Table 3.
It is to be expected that the lightweight aggregate concrete beams will deflect more at ultimate design load than those made with natural aggregate concrete due to the lower modulus of elasticity. The results however show no obvious trend and in fact in only one of the series ( series 1 over reinforced ) does the natural aggregate concrete show a lower deflection than either of the two lightweight concretes. Load deflection curves for series 3 ( reinforced with 2 Y -16 bars) are shown in Fig. 5, in this particular set of tests the ash concrete beams show the highest deflection ( which was to be expected ) but the Lytag beams deflected the least. The differences though are only small, for example at the design load of approximately 30.0 kN the differences between all three tests is only about 12 % and from Table 3 the maximum difference at the design moment for all tests was no more than 15%. The structural performance of the beams was also assessed using the UK design code which was applicable at the time, the characteristic curve for the flexural strength of a member was obtained from a graph of :Mu fcu bd 2
against
As. bd
fcu
415
The relationship between these two terms is shown in Fig. 6 together with the results of all the twelve beams tested. It can be seen that the results for all beams lie outside the ultimate limit state envelope used by the code and that they all follow a similar trend regardless of the type of concrete used. The figures for both average crack width and spacing at the ultimate design moment are given in Table 3; both can be seen to be greater for the lightweight concretes than for the natural aggregate which is what was expected. However more importantly none of the values recorded were outside the limits recommended by the code and in general the values for the ash concrete are lower than those of the Lytag concrete. The strain distribution about the centre of each beam was measured for each increment of load and in every case the stains were seen to be distributed approximately linearly above the neutral axis. In addition the position of the neutral axis was in general lower for the lightweight concrete beams due to their lower elastic modulus.
Conclusions The following conclusions can be drawn from the work reported here. 9 The deflections of reinforced concrete beams at the ultimate design moment made with the ash aggregate compare favourably with those of beams made from the natural aggregates and were less than those beams made with Lytag concrete. 9 The average crack width and spacing was similar for lightweight aggregate concretes; they were both higher than the natural aggregate concrete but still within the limits of the structural code. 9 The shape of the stress strain distribution curve was similar for all beams but the depth of the neutral axis in the lightweight concrete beams was lower. The anchorage bond strength was higher for the natural aggregate concrete than for the lightweight concretes of similar mortar strengths. The cohesion bond was however lower when comparisons were made on the basis of equal concrete compressive strength. 9 Overall the structural performance of the incinerator ash concrete was similar to that of the Lytag concrete and compared favourably with the natural aggregate concrete.
416
References
Wainwright, P.J and Boni ,S.P.K. Artificial aggregates from domestic refuse. Concrete Vol. 15. No. 5 May 1981, pp 25-29. Boni, S.P.K. The use o f sintered pelletised domestic refuse as an aggregate in concrete. Ph.D thesis, Dept.Civil Engineering, Leeds University, England. 1980, pp.285. Wainwright, P.J. and Boni, S.P.K. Some properties o f concrete containing sintered domestic refuse as a course aggregate. Magazine of Concrete Research Vol. 35, No. 123. June 1983, pp. 75- 85. Hadzinakos, I. The chemical and other properties o f sintered refuse slags as an aggregate in concrete. Ph.D thesis, Dept. Civil Engineering, Leeds University, England. 1980, pp. 243. Robery, P.C. The use o f domestic incinerator ash as an aggregate in concrete. Ph.D thesis. Dept. Civil Engineering, Leeds University, England. 1982, pp.393. Wainwright, P.J. et al. A review o f the methods o f utilisation o f incinerator residues as a construction material. Proc. Int. Conf. Low - cost and energy saving construction materials. Rio de Janeiro, Brazil, July 1984. Wainwright, P.J. and Robery,P.C. Production and properties o f sintered incinerator residues as aggregate for concrete. Proc. Int. Conf. Environmental Implications of Construction with Waste Materials, Maastricht. Netherlands, November 1991. ( Studies in Environmental Science No. 48 published by Elsevier ) pp. 425 - 432. British Standards Institution. Methods for Testing Concrete. BS 1881. 1983.
Snowdon. l.C. Classifying reinforcing bars for bond strength. BRE current paper CP36/70, Nov. 1970. Building Research Establishment, Watford, England
417
Table 1
Aggregate Properties
Source
Bulk Density ( kg/m3) Loose 1525 1059 896
10mm gravel Ash Lytag
Relative Density
Water Absorption %
2.61 2.29 1.75
0.6 10.11 12.20
Rodded 1626 1121 962
Table 2 Reinforced concrete beam details Beam No.
H1 H2 H3 H4 G1 G2 G3 G4 L1 L2 L3 L4
Agg. Type
Comp. Str. N/mm 2
Ash
28d. 37.5
Gravel
37.5 37.5 45.0
Lyta9
45.0 45.0 50.0 50.0 50.0
Flex. Str. N/mm 2
6m.
28d. 4.6
44.6
4.6 4.6 4.3
52.3
4.3 4.3 4.35
57.8
4.35 4.35
r
Elast. Mod. KN/mm 2 28d. 20.2
6m. 4.68
20.2 20.2 32.7
4.44
32.7 32.7 23.3
5.18
23.3 23.3
bxd mm.
Reinf. Type
Steel ratio
6m. 22.4
35.8
25.5
120xl 72
120xl 64
120xl 72
3-Y16 3-Y16 2-Y16 2-Y12 3-Y16 3-Y16 2-Y16 2-Y12 3-Y16 3-"(16 2-Y16 2-Y12
2.92 2.92 1.95 1.08 3.03 3.03 2.09 1.12 2.92 2.92 1.95 1.08
Note :- Overall dimensions of all beams 2600 x 120 x 200mm.
Table 3 Summary of reinforced concrete beam results Beam No.
Cracking moment kNm. Mcr
UIt. des. moment kNm. Mud
H1 H2 H3 H4 G1 G2 G3 G4 L1 L2 L3 L4
4.59 5.99 3.86 3.78 4.68 5.29 3.85 3.49 3.99 5.82 3.74 3.48
24.9 26.2 19.3 12.2 25.0 26.5 19.0 12.0 27.6 28.4 20.5 12.6
,
Ult. des. failure moment kNm. Mu 35.7 36.8 25.3 15.2 34.9 35.7 24.5 14.6 37.4 38.1 26.1 15.4
Exp. ult. failure moment kNm. Mue 35.5 36.8 24.4 17.4 38.1 42.0 27.6 18.8 37.1 39.9 28.9 18.7 ,
Deflection at Mud mm. 7.10 6.45 7.05 5.6 7.00 7.00 6.95 5.65 7.8 7.45 6.3 5.1
Av. crack width at Mud mm. x 10-2 9 4 6
6 6 5 10 6 5 7
Av. crack spacing at Mud mm. 78 88 74 78 75 70 70 78 88 84 70 78
418
Fig. 1 Compressive strength development 7o
4b L y t a g 9 g ravel
30
0
Aash
0
50
100 Age
20O
150
days
Fig. 2 Relationship between compressive strength and elastic modulus 7O
~
~~
f
5o
jr
40 f/)
A ash 9Lytag
30 -I
~o
.,, m
ip
ii!, gravel
,m
w
...
0
20
25
30
Dynamic
35
mod.
40
elasticity
45
kNImm 2
Fig. 3 Relationship between average bond stress and bar slip 15
~ 10
I
. --.-I
o
9gravel AAsh 5 I
0
0.2
0.4
0.6 Bar slip mm.
0.8
1.2
419
Fig. 4 Average bond stress against bar stress 4.5 4 ~E 3.5
"o
J c
9Lytag 9gravel
2.5
w
2
9ash
1.5 ,.,,'.....
0.5 0
-
~,r
,t
0
50
100
150
200
250
,
300
350
400
Av. bar str. Nlmm z
Fig. 5 Load deflection relationship for beams with 2 Y- 16 reinforcement _.._,=.=
i.,' /
9Lytag 9gravel
L
9
J
0
5
10
15
20
25
deflection m m .
I'asI
Fig. 6 Characteristic curve for flexural strength 0.3
9lytag
0.25
des. ult. state
0.2 o.15
=E
=,.
F~
0.1 0.05 0
J
0.05
0.1
0.15
0.2
As.fylbd.fcu
0.25
m
les. limit state
0.3
0.35
9gravel
This Page Intentionally Left Blank
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
Investigating
Waste/Binder
Interactions
421
By Neural
Network
Analysis
C. D. Hills, J. A. Stegemann, N. R. Buenfeld Imperial College of Science, Technology and Medicme, London, SW7 2BU, UK
ABSTRACT Sohdifymg cement/waste mixtures comprise highly complex chemical environments. The difficulty in predicting either positive or detrimental interactions between different components hampers the design of sohdification treatment processes and can result m either environmentally unacceptable products which require expensive remediation, or unnecessary overdosing of additives. Development of a rehable diagnostic and predictive system for assessing treatability of wastes by sohdification would therefore be a great benefit to users. No efforts m this regard have yet been undertaken because of the enormous number of variables involved. However, a preliminary exploration of the concept of using neural networks for this purpose has shown that it is possible to tram a network to predict solidified waste properties to an acceptable degree of accuracy. The present work reviews factors that influence the properties of stabilised/sohdified wastes, presents research objectives designed to improve waste sohdification technology and preliminary results from neural network modehng studies.
1. I N T R O D U C T I O N Solidification systems employing cementitious and pozzolanic binders rely upon hydration reactions to physically and chemically brad waste species to form products that are more stable in the environment. The fundamental principle behind this technology appears relatively simple and involves the controlled application of hydraulic binders, such as ordinary Portland cement. Correctly applied, these binders minimize the solubility of contaminants and entrap them in the structure of a sohdified product by sorption, lattice inclusion I and enclosure in the matrix porosity. Sohdification is used to significantly improve the handling and facilitate the safe disposal of a wide range of hazardous waste materials and contaminated land 2. In practice, the wide choice of sohdification systems coupled with the fact that wastes from different processes, and with differing characteristics, can not all be treated alike means that choosing an appropriate binder for a given waste stream is a complex undertaking. Inappropriate solidification formulations can result in handling difficulties during processing and placement, failure of the sohdified material to set and harden into a durable monohth, physical degradation of the matrix with time, and/or provision of a chemical environment m which contaminants are not fully immobilized. All of these results indicate an inadequate protection of the environment. Reports of environmental problems associated with considerable quantities of commercially sohdified wastes 3,4,5,6, ~ despite the application of testing regimes required by UK regulatory authorities, suggest that there are limitations on the use of hydraulic cements in this application, which are not yet defined. This work explores some of the factors responsible for effective waste solidification. Its purpose is to address the potential of neural network analysis for contributing to a better understanding of the relationship between waste and binder and variables of importance to solidified product properties in both the short and longer term.
422
2. I N T E R F E R E N C E S WITH C E M E N T H Y D R A T I O N In preparing concrete, l0 to 20% of hydraulic binder is blended with inert aggregate to yield a high-strength building material. Small amounts of chemical admixtures (<2% by weight of cement) can have a significant impact upon the properties of concrete 7. Durmg sohdification, small quantities of cementitious binder can be exposed to large quantities of chemicals. For example, a waste contaming 45% by wet weight of sohds might be treated with as little as 10% by weight of hydrauhc cement. A considerable number of compounds are now known to affect hydration processes s,9,1o,11, causing effects such as: acceleration or retardation of set (which, in the extreme, can result in flash set or complete inhibition of hydration), false set, increased water consumption, and matrix disruption by expansive reactions 12,1~,10
Figure 1 shows an example electron photomicrograph of a Portland cement bound solidified product about 1 year of age. The solidified product contained 40% blended neutrahsed heavy metal waste and 15% binder, on a wet weight basis. The microstructure is devoid of the crystalline phases which are common in conventional cement and concrete and is typical of the evidence of persistent interference with hydration.
Figure 1. Micrograph showing typical solidified product microstructure (1 year old sohdified inorganic waste).
Both organic and inorganic waste species are capable of interference with hydration mechanisms by hindering transport processes which occur into and out of cement particles during normal hydration 14,15, and consequently impair microstructural development ~ Nonpolar organic compounds, such as polychlorinated biphenyls and many polynuclear aromatic compounds, tend to be water insoluble and will tend to partition with any solid phase in preference to water 17. Although this may have the effect of reducing their leachability, sorption of even small quantities of organics onto the solid binder can interfere with hydration. Many soluble polar organics do not tend to react with cement, and their leachability is therefore not reduced by solidification; organic compounds which do produce salts, complexes or precipitates may also remain mobile by compromising hydration reactions. Inorganic species are generally considered most amenable to cement-based treatment, because they tend to be reactive with hydraulic binders. However, it is this very reactivity which also
423
affects cement hydration reactions. The severity of retardation effects for a variety of anions and cations has been r a n k e d by several authors e.g. lS.9, but to a large extent this is an oversimplified approach, and only possible when workmg with pure compounds, r a t h e r than blended species as they are found in real wastes. In real wastes there are complicating factors which p r e v e n t simple ranking, for instance there may be confounding effects resulting from a particular combination of cation and anion, the concentration of each, the presence of other compounds in the waste, and curing conditions. By way of a simple example, less than approximately 1% of CaC12 retards the set of Portland cement, greater amounts cause acceleration 19, and more than approximately 4% causes matrix disruption 20. Similar effects have been observed for actual sohdified products; use of Portland cement to solidify an electric arc furnace dust resulted in initial brief retardation followed by rapid acceleration of hydration at 30% dry weight electric arc furnace dust, but only a strong inhibition of hydration at 60% electric arc furnace dust 21 Figure 2 illustrates the synergistic effect of the addition of interfering agents on the heat of hydration of OPC. Compounds found in waste materials were added to OPC paste as metal hydroxides and sodium salts in single and multiple additions 12. Whereas most additions caused either retardation or acceleration, a combmation of metal hydroxides was found to cause indefinite retardation of hydration reactions.
3. I N F L U E N C E OF B I N D E R C H E M I S T R Y
The chemistry of a hydraulic binder system can be tailored by the inclusion of other cementitious materials, such as blast furnace slag, and/or pozzolanic mineral admixtures, such as coal fly ash. Alternative binders to Portland cement can sometimes help overcome deleterious effects induced by a waste 22. For instance, in the example of the sohdified electric arc furnace discussed earlier, use of an activated blast furnace slag formulation in place of Portland cement resulted in a consistent mild retardation, which simplified control of the process 21 Figure 3 shows the effect of binder type on the a m o u n t of chromium leached from solidified waste forms containing a zinc and chromium rich plating waste extracted with distilled w a t e r for 24h, by end over end tumbling accordmg to the modified DIN 38414 test 2~. Particularly at high waste additions, the concentration of chromium in the leachates was far higher for blended ordinary Portland cement (OPC) t h a n for calcium a l u m m a t e cements (CACs) with the individual differences between pozzolan additions bemg of less importance. The relatively high chromium concentration leached from some samples indicated that chromium was speciated as Cr 6§ in this waste. As the solubility of Cr 6§ is not pH sensitive, the observed differences in leachability imply a change in speciation or uptake of the ion into the cement hydration products.
4. O T H E R F A C T O R S In general, wastes presented for sohdification have undergone p r e t r e a t m e n t in some way. Appropriate p r e t r e a t m e n t can be important to maximize efficient use of binder. For example, preneutrahzation saves using valuable cement just to neutralize acidity and can help reduce mterferences by taking interfering agents out of solution 24. Precipitated species are incorporated in the sohdified waste matrix as sludge particles, and species which were not precipitated r e m a m available for incorporation in cement hydration products. Curing conditions are another factor to which lamentably httle attention is paid. The effect of some interferences is t e m p e r a t u r e dependent 25, which suggests t h a t use of an appropriate curing
424
temperature is important in obtaining the desired final properties in a waste form. For solidified wastes, as for concrete, freezing, drying and physical disturbance of a setting mixture have detrimental effects. Therefore, protection from freezing conditions, high humidity and placement prior to set are all essential for obtaining a high quality solidified product.
Figure 2. Effect of additions on the heat of hydration of OPC (after Hills et al. 12) RETARD
ACCELERATE OH. SO4
4
Pba Hg
Cd,'
Mg
Cr 9
9
Fe
~o3
~3
Ni Combined a n k h s
OPCCONTROL
5 E2 Cu Zn
1
9
9 Combined cations and anions CN
HPO4,
k " Combined cations 0 ii , 0.00e+0 1.00e-5
l 2.00e-5
, 3.00e-5
4.00e-5
Time to maximum heat (reciprocal seconds)
Figure 3. Chromium leached from blended OPC and CAC binders with increasing waste contents 30 ~. "~
9 9 9
,- 20 o
C A C / f u e l ash C A C / m e t 9 kaolin CAC/silica fume
9 OPC/fuel ash I'-I OPC/meta kaolin 9 OPC/silica fume
o .xz
E .5
10
10
20
30
Waste addition rate (% w/w)
425
Finally, it is worth noting here that other factors such as the method and efficiency of blending waste and binder and other aspects of processing such as mix design will all exert some influence on solidified product properties.
5. A P P L I C A T I O N OF NEURAL NETWORKS The many and complex potential interactions between the variables in a solidified product are not possible to evaluate by simple means. The aforementioned discussion of the deleterious mteractions between waste and binder has concluded that a considerable number of materials are involved and that synergistic relationships are implicated. With this in mind, alternative means of identifying the relationships of importance are required and could incorporate an examination of existing data. To-date little work has been carried out on evaluating existing data in order to aid future solidification practices. The reason for this may lie in the difficulty of approaching such a complex and clearly nonlinear problem with conventional data analysis techniques. However, it may be possible to address the problem using neural network analysis. Neural network analysis is capable of processing of a variety of inputs, and therefore has advantages over conventional techniques for finding relationships between variables in complex systems. A neural network consists of a number of interconnected processors which are arranged in layers: are capable of receiving incoming information from several sources, weighting the information, and transmitting it in a single output to the next layer of the network. Learning by a neural network consists of changing the weights on the information transferred through the processors in a systematic fashion based on training. When enough training data is available neural networks can generalise to predict combinations of inputs not previously encountered, enabling them to be used to identify patterns in large data sets of many variables 26 An inevitable consequence of neural network analysis is that any relationships identified between variables in a dataset are non-mechanistic. While this might appear to be an obvious drawback, the advantage is that relationships can be identified without a preliminary mechanistic hypothesis. Thus, it may be possible to develop new insight into mechanisms, based on identification of new relationships through neural network analysis. The application of neural network analysis to cementitious systems has been demonstrated elsewhere e.g. 26,27,2s. Glass et al. 29 have quantified a wide range of factors influencing the binding of chlorides in concrete from published data. Preliminary experiments concerning the application of neural networks to analysis of laboratory solidification data, have been positive and encouraging. When applied to the data from a laboratory program involving 150 mixes produced from different waste/binder combinations to predict setting times, strength development and leachate metals composition an average accuracy of greater than 7 5% was achieved 30 Table 1 summarises the networks constructed and the results obtained. Figure 4, shows the predicted vs measured leaching of molybdenum and zinc. The close correspondence between measured and predicted concentrations clearly illustrates the potential of this approach. Additional preliminary neural networks were constructed using specific information concerning the metals of interest, such as ionic radius, atomic weight, and valency and this has enabled other relationships to be generalised, however, further work is required before these results can be verified.
426
Figure 4. Predicted vs m e a s u r e d leaching values of molybdenum and zinc from laboratory solidified wastes (after Hills et al., 1997)
A
8
Mo measured
~.
Mo predicted
....... 41~......
Zn measured
....... ~ ......
Zn predicted
g
=o
~ 6 o r,J
~4
J
I J
J J
J ......... _.~ . . . . . .
o
0
10
~--"~
~ .
........
20
30
.
. . . -~......~.~.~
40
Waste addition (% w/w)
There is an urgent need to improve our knowledge concerning waste fixation through both short and long term research. Three main areas of research and development are required to be addressed as shown in Table 2.
6. O B J E C T I V E S FOR F U T U R E D E V E L O P M E N T
Neural network analysis is a potential tool in several of these areas and could lead to a knowledge-based diagnostic and predictive system for predicting interactions between components of solidified waste, and final product properties. This would be a great benefit to users of the technology by: providing insight into the most important parameters in design of durable formulations of low leachability, eliminating or reducing the time and expense of treatability studies, streamlining performance testing enabling process refinement during full-scale solidification based on feedback of quality analysis and control data, and reducing the requirement for and subjectivity associated with h u m a n expertise. Table 3 lists parameters which can be found in the existing data and could be used to train and test neural networks to examine the relationships between solidified waste properties.
The relationships expected to be of particular interest are those between: waste/binder dosages, the presence of agents which normally interfere with cement hydration, and engineering properties. 9 different measures of the transport characteristics of the solidified waste matrix (e.g., water absorption, and leachability in ANSI/ANS 16.1, and 9 waste and binder composition and pH and acid neutralization capacity of the solidified product. 9
427
Table 1 Summary of data from neural network analysis of laboratory solidified wastes (after Hills et al. 3o) INPUTS
NN 1
NN2
NN3
Cement type (5 types)
YES
Pozzolan (3 types)
YES
Waste Type Content
YES
NN4
NN5
Measured and predicted from NN3 Measured and predicted from NN4
Setting time (mins)
NO
If measured
NO
YES
Strength (kPa)
NO
If measured
YES
NO
Leachate metal (5 metals) Hidden neurons OUTPUT
Average accuracy
YES 12
NO 22
Leachate concentration (mg/1) 63.2%
56.8%
YES
14
15
14
Setting time (mins)
Strength (kPa)
Leachate concentration (mg/1)
76.8%
74.2%
75.3%
Table 2 Research and development effort for improving solidification technology RESEARCH AREA OBJECTIVE Basic Science elucidating waste binder interactions improving/optimising binders developing novel systems Applied Science improving materials processing characterising solidified product performance Regulation developing long-term monitoring framework specifying characterisation procedures developing guidelines for safe containment structures
7. CONCLUSIONS When Portland cement is used to solidify waste materials the complex interactions that result cannot be easily characterised. Consequently, design of solidification formulations is complicated, and entails a high risk of failure. Neural network analysis can be used to identify complex relationships between large numbers of variables and it may be possible to elucidate factors of importance to solidification from existing data.
428
Preliminary studies have been carried out by neural network analysis with the result that contaminant leachate concentrations of laboratory-prepared solidified wastes were predicted to an acceptable degree of accuracy. In addition, a number of other relationships were generalised indicating that this method may have considerable potential to aid our understanding of the important variables involved.
Table 3 Example of variables of potential importance for solidification COMPONENT VARIABLE Binder type composition mineralogy acid neutralisation capacity Waste type composition, mineralogy acid neutralisation capacity Combined setting time (Solidified Product) strength development moisture content specific gravity age water adsorption hydraulic conductivity acid neutralisation capacity pore water composition monolithic leachability (e.g. ANSI/ANS 16.1) leachability (distilled water/acid batch extraction etc.) freeze thaw resistance wet/dry weathering resistance
8.
REFERENCES
1Glasser F.P. (1992) Chemistry of Cement Solidified Waste Forms In: Chemistry and Microstructure of Solidified Waste Forms. (Ed. R.D.Spence). Lewis Publishers, London, pp 1-35. 2 USEPA, (1994) Innovative Treatment Technologies: Annual Status Report, (6th. Edition). USEPA Report EPA-542-R-94-005, September 1994. Environmental Data Services Ltd. (1988) ENDS Report 158, p. 8. 4 Environmental Data Services Ltd. (1989) ENDS Report 173, pp. 8 -9. 5 Environmental Data Services Ltd. (1990) ENDS Report 186, p. 10. 6 Environmental Data Services Ltd. (1992) ENDS Report 205, p. 11, 33. 7 Rixom, M.R. and Mailvaganam, N.P., (1986) Chemical Admixtures for Concrete, (2nd. Edition). London. 8 Jones, L.W., (1988) Interference Mechanisms in Waste Stabilisation/Solidification Processes. Literature Review. tlazardous Waste Engineering Research Laboratory, Office of Research and Development. US EPA, Cincinatti, OH. USA. 9 Conner, J.R., (1990) Chemical Fixation and Solidification of Hazardous Wastes. Van Nostrand Reinhold. New York. ~oMattus, C.H. and A.J. (1996) Literature Review of the Interaction of Select Inorganic Species on the set and Properties of Cement and Methods of Abatement Through Waste Pretreatment.
429
Stabilization and Solidification of Hazardous, Radioactive and Mixed Wastes, 3rd Volume, ASTM STP 1240. (Eds.) Gilliam, T.M and Wiles, C.C. ASTM. 609-633. 11 Hills, C.D. and Pollard, S.J.T. (1997) The Influence of Interference Effects on the Mechanical, Microstructural and Fixation Chracateristics of Cement Solidified Hazardous Waste Forms. J. Haz. Mat. (in press). 12 Hills, C.D., Sollars, C.J. and Perry, R., (1993) Ordinary Portland Cement-Based Solidification of Toxic Wastes: The Role of OPC Reviewed. Cem. Concr. Res., 23, 196- 212. 13 Wastewater Technology Centre, (1992b), Engineering Properties Testing of Solidified Residues, Ontario Waste Management Corporation Working Paper, Zenon Environmental Inc., February, 1992. 14 Forsen, L., (1933a) Die Chemie des Portlandzments in Komplexchemischer Darstellung, Zement, 6, 73- 78. 15 Forsen, L., (1933b) Die Chemie des Portlandzments in Komplexchemischer Darstellung, Zement, 7,87-91. 16 Hills, C.D., Sollars, C.J and Perry, R., (1994) Calorimetric and Microstructural Study of Solidified Toxic Wastes: Part 2 -a Model for Poisoning of Hydration. Waste Managm., 14, 601 -612. 17 Caldwell, R.J., and Cote, P.L., (1990). Investigation of Sohdification for the Immobilization of Trace Organic Contaminants, Haz. Waste Haz. Mat., 7, 3, 273 -282. 18 Wilding, C.R.,Walter, A. and Double, D.D., (1984) A Classification of Inorganic and Organic Admixtures by Conduction Calorimetry. Cem. Concr. Res., 14, 185 -194. 19 Lea F.M. (1970) The Chemistry of Cement and Concrete. (3rd. Edition) Edward Arnold, London, UK 20 Wiles, C.C. and Barth, E., (1992) "Solidification/Stabilisation: is it Always Appropriate"? Stabilization and Solidification of Hazardous, Radioactive and Mixed Wastes, 2nd Volume, ASTM STP 1123. (Eds.) Gilham, T.M and Wiles, C.C. ASTM. 18 -32. 21 Shi, C., Stegemann, J.A., and Caldwell, R., Use of Heat Signature in Solidification Treatability Studies, accepted for presentation at the 10th International Congress on the Chemistry of Cement, Goteborg, Sweden, June 2-6, 1997. 22 Koe, L.C., Hills, C.D, Sollars, C and Perry, R. (1992) Hydration Reactions During the Sohdification/Stabilisation of Toxic Wastes. Proc. Pacific Basin Conf. on Hazardous Waste. Pacific Basra Consortium for Hazardous Waste Research, Bangkok, Thailand 23 Council of the European Communities, (1993) Draft Directive on the Landfilling of Waste. Report SYN 335, Brussels, June 1993. 24 Stegemann, J.A., Shi, C., and Caldwell, R.J., (1994). Response of Various Sohdification Systems to Acid Addition, presented at WASCON'94- The International Conference on Environmental Implications of Construction Materials and Technology Developments, Maastricht, The Netherlands, June 1-3, 1994. 25 Wastewater Technology Centre, (1992a), Accelerated Testing of Solidified Residues. Working Paper for Ontario Waste Management Corporation, Toronto, Canada. 26 Buenfeld, N.R. and. Hassanem, N.M, (1995) Neural Networks for Predicting the Deterioration of Concrete Structures, Nato/Rflem Workshop on the Modelling of Microstructure and its Potential for Studying Transport Properties and Durability, St Remy-Les-Chevreuse, 10-13 July, 1995, pp. 413-430, Kluwer Academic Publishers, Dordrecht. 27 Buenfeld, N.R. and. Hassanein, N.M, (1997a) An Artificial Neural Network for Predicting Carbonation Depth in Concrete Structures, Chapter 4 of 2nd ASCE Monograph on Artificial Neural networks in Civil Engineering, American Society of Civil Engineers. 2s Buenfeld, N.R. and. Hassanem, N.M, (1997b) Neural Networks for Modelling the Influence of Cement Chemistry on Concrete Durability, 10th International Congress on the Chemistry of Cement, Goteborg, Sweden, 2-6 June, 1997. 29 Glass, G.K., Hassanein, N.M and Buenfeld, N.R. (1997). Neural Network Modelling of Chloride Binding. Mag. Concr. Res. (m press).
430
30 Hills, C.D., Hassanein, N.M. and Buenfeld, N.R. (1997) An Examination of Factors Affecting Solidified Hazardous Waste Forms by Neural Network Analysis. Proc. Int. Congr. Waste Solidification -Stabilisation Processes, Nancy, France 28 Nov. -1 Dec. 1995
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice © 1997Elsevier Science B.V. All rights reserved. THE USE OF MATERIALS
MSWI
BOTTOM
ASH
431 IN
HOLLOW
CONSTRUCTION
E. JANSEGERS
Gewestelijk Agentschap voor Netheid, Net Brussel, de Broquevillelaan 12, Brussel, Belgium Abstract Municipal Solid Waste Incineration (MSWI) bottom ash has been used to produce hollow building stones under industrial conditions. Different concrete mixes in which a part of the natural aggregates has been replaced by pretreated aged bottom ash have been tested. The stones were tested on their stability, durability and environmental aspects. The results are very promising. A patent in the name of <> has been granted.
Unlike full concrete building
stones, these building stones have thin walls from which components formed during curing can migrate out without causing cracking or any other destruction of the concrete stone.
1.
INTRODUCTION
The use of bottom ash in road applications has been one of the solutions to avoid dumping it on landfills. <>, the company that collects and treats household waste and similar industrial waste in the Region of the Capital of Brussels (1 million inhabitants),, and owner of the incineration plant (500 000 T/year), has tried to elaborate another way of recycling bottom ash and more precisely by using it for the fabrication of hollow concrete building stones. The research from others has proved that concrete mixtures using bottom ash were of bad quality. Cracks in the concrete were detected and this is probably the reason why research has stopped examining this solution of recycling. Although the production of thin-walled hollow concrete elements requires in general a better concrete mortar condition than the production of full concrete elements, it has been discovered that thin-walled hollow elements do not present cracks. Any componentwhich may form during the curing step of the manufacturing process can migrate out of the thin walls of the concrete stone without this resulting in cracking or other destruction or alteration of the concrete element.
432 The aim is to produce building stones using pre-treated aged bottom ash as a substitute for gravel.
By encapsulating the bottom ash in a cement matrix of sufficient compression
strength and of sufficient dimensions, an immobilisation of the pollutants is aimed. This has been checked by studies on technical behaviour (compression strength and secondary sideeffects) and on environmental behaviour (increase in diffusion resistance).
2.
P R E T R E A T M E N T AND M A T U R A T I O N C O N D I T I O N S
At the incineration plant of Brussels (grate furnace), the bottom ash is cooled in a waterextractor. The large elements (> 120 mm) are separated from the ash. The remaining ash is then elementary de-ironed by means of a magnetic drum. Two tests were performed on bottom ash that previously had leaked out for about 2 .weeks. This allows a good air exposure and a decrease of humidity. The carbonation process is stimulated. In the first test, the ash was sieved and de-ironed a second time in another installation. An intensive iron separation is important to avoid the formation of rusty spots.
A second test was performed with, beside a de-ironing, a complementary removal of non ferrous materials. This was done to compare the technical and environmental values with the first experiment.
An air separation was used to remove the fine non-burned fraction.
The bottom ash was broken in a percussion breaker.
This was done to compare the
differences in behaviour of the stones, made with the two types of bottom ash. In the two tests, the bottom ash was subjected to a maturation process of at least 3 months. Therefore the organic compounds can putrefy, metals can be convened to less soluble metalhydroxydes and salts can be mineralised to less soluble forms. The heat released by the exothermic reactions results in a decrease of humidity.
3.
STUDIES ON T E C H N I C A L B E H A V I O U R
Stones were fabricated using bottom ash as a substitute for the gravel used in classical stones (1/3 of the weight).
During laboratory tests the compatibility between bottom ash and different types of cement was tested.
The pre-treated and aged bottom ash was fed with water and cement in a concrete mixer at a factory where buildin~ stones were made. After .qevo.r~l minl~t~ nF rniYino
th~
nr~, ....
14
433 concrete was poured into moulds and these moulds were vibrated for a few seconds. The resulting elements were then removed form the moulds and cured a first time for several hours in a substantially closed area, in which temperature and humidity rose as a result of the heat and moisture released due to the reaction formed by mixing the different components. The so formed elements are packed and stored and are submitted to a second curing for 7 days under atmospheric conditions. Several thousands of blocks were made with the bottom ash from the two tests (one without and one with separation of the non ferrous metals). The building stones, which are now ready for use, belong to the 6/1.6 category (compressive strength superior t o 6 MPa and apparent specific gravity inferior to 1 600 kg/m~). Their dimensions are 29 by 19 by 19 mm.
Figure 1 : presentation of the building stones Their dimensions, their mechanical and their physical characteristics are conform to the Belgian norm BENOR for the production of building stones typically used for indoor applications (NBN B21-001). No swelling or shrinkage effects occurred. The shrinkage class is less than respectively 0,6 mm/m and 0,4 mm/m (comparable with the values of classical stones) for the first and the second test (with separation of non ferrous materials). Frost resistance was good. The behaviour of alternating wet and dry cycles showed no effect on the stones. Water absorption was a little high in the second test (> 8%), but since no problems with frost •resistance or wet & dry cycles has been observed, this does not seem to have consequences.
434 The compatibility of the stones with interior wall renderings (paints, ceiling coatings and waterproof coatings) was examined during a period of 3 months after application of the coatings. No effects (even with artificial rehumidification) different from those occurring with conventional building stones were noticed. No spots, salt-formations, efflorescence, flaking or blisters appeared. In the second test, building stones were made with the sieved bottom ash, but also with the broken fraction. The humidity of the broken bottom ash was lower than in the non broken ash. The non broken fraction contained a little more fine material than the broken bottom ash.
The compressive strength was a little higher for the stones made with the broken
bottom ash than the stones with the non broken bottom ash. 4.
STUDIES ON E N V I R O N M E N T A L B E H A V I O U R
In formed materials, diffusion processes are responsible for Iixiviation. Therefore, a test for measuring the diffusion was carried out in order to determine the amount of components released.
Diffusion tests were performed according to the NEN 7345 method, which measures the cumulative diffusion during 64 days.
Test cubes were sawn from the building stones. As the minimum dimensions of 40 mm could not be reached due to the hollowness of the stones, the dimensions of the test cubes were in the first test (on the stones made by bottom ash where the non ferrous metals were not separated) 80 x 80 x 35 mm. The evolution of pH was similar for all the stones, so differences in pH cannot explain possible differences in lixiviation.
All parameters were
significantly lower than the Flemish project values (in general more severe than the values from the <>from the Netherlands), except for copper where the values were exceeded. (mg/rr~) As
27
Sb
Ba
100
Se
Cd Cr Co Cu Hg Mo Ni
Pb
1,1 55 8,5 25 0,8 11 15
60
Sn V Zn Br CICN F-
SO4 2-
6 0,7 20 86 90 30 36 000 7 360 27 000
Table 1 : lixiviation limits in formed materials according to the Flemish project
435 For lead and zinc the best immobilisation was reached. sulphate were diffusion-controlled.
The lixiviation of chrome and
The high values found for the lixiviation of copper
indicate that part of it is present in a mobile chemical form.
Wash-up and exhaustion
effects are the limiting factors for this lixiviation behaviour. A second test was performed (on stones made using bottom ash from which non ferrous metals and the non-burned fraction have been separated) on cylindrical test samples with a diameter of about 10,5 cm where only the upper circle was not imbedded in an acryl resin. This was done in order to avoid exhausting effects detected in the first test. For none of the metals lixiviation values could be measured. The calculated values were made with the detection limits as values, and have to be considered as ceiling values.
The measured
values were lower than the calculated values, but both were beneath the values fixed for the Flemish Region. There were no differences observed between the building stones using broken or using non broken bottom ash. In parallel with the second diffusion test, test samples made with 8% cement and the same pre-treated, aged bottom ash were also subjected to a diffusion test. In order to simulate the lixiviation behaviour of bottom ash used in bound applications as foundations from road constructions.
The critical parameter was copper that exceeded the Flemish lixiviation
limits for bound applications.
5.
CONCLUSIONS
The recycling of bottom ash in hollow concrete building stones is technical and environmental viable. The thin walls of the stones allowthe components formed during curing to migrate, out without causing cracking or any other destruction of the stones. The bottom ash used to make these stones has to be de-ironed to avoid rusty spots and has to undergo a separation of the non ferrous metals to avoid exceeding values on environmental limiting values. Before the treatment the bottom ash is preferably leaked out for two weeks to allow a good aeration. After the treatment the bottom ash has to be aged, for instance for 3 months. The stones are fabricated using this pre-treated and aged bottom ash as a substitute for the gravel used in classical building stones. The compressive strength is high enough for indoor building stones, and unwanted sideeffects such as lack of frost resistance, swelling or shrinkage effects, bad behaviour after wet and dry cycles and non compatibility with interior wall renderings are absent.
436 No technical nor environmental problems were observed for the building stones in which broken bottom ash instead of non broken bottom ash was used. The use of bottom ash to fabricate hollow concrete building stones is a good technical and environmental solution to avoid dumping it on landfills and to recycle materials.
The
diffusion tests have shown that the only possible critical parameter in building stones is the lixiviation of copper, but that this is, after a good separation of non ferrous metals, probably completely resolved.
5.
REFERENCES
1. C.R.O.W. (1988) : Resten zijn geen afval (meer) - Afvalverbrandingsslakken. Publication 15, C.R.O.W., Ede, The Netherlands •2. W.T.C.B. (1992): La r6cup6ration, dans le secteur de la construction, des mfichefers d'incineration des dechets m6nagers. Unpublished internal report, Belgium 3. W.T.C.B. (1994): R6cuperation des mfichefers.d'incin6ration pour la fabrication de blocs de construction. Unpublished internal report, Belgium 4. W.T.C.B. (1996): Utilisation des mfichefers d'incin6ration dans le secteur de la construction. Unpublished internal report, Belgium 5. V.I.T.O. (1995): Milieuhygienische beoordeling van het gebruik van AVI-bodemassen in snelbouwstenen. Unpublished internal report, Belgium 6. V.I.T.O. (1997): Milieuhygi6nische evaluatie van AVI-bodemassen als secondaire grondstof. Unpublished internal report, Belgium
Goumans/Senden/vander Sloot, Editors Waste Materials in Construction:Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
437
Using C H E M F R O N T S , a geochemical transport program, to simulate leaching from waste materials Catharina B~iverman, Luis Moreno and Ivars Neretnieks D e p a r t m e n t of Chemical Engineering and Technology Royal Institute of Technology Stockholm, Sweden Abstract A coupled geochemical and transport program, CHEMFRONTS, has been developed for calculating water transport in porous media. The program was developed to handle sharp reaction fronts such as redox and pH fronts. CHEMFRONTS is based on reaction kinetics and can therefore handle non-equilibrium systems such as glass phase dissolution. After modification in the program, we have successfully simulated partially saturated systems with indiffusing carbon dioxide and oxygen. Unsaturated systems are, however, very complex. All the information required to make reliable simulations are not available. Simulations of these systems can, therefore, only be used to show trends and the influence of different variables. In spite of these uncertainties, we find these simulations meaningful to perform.
Introduction Ashes and slags have physical properties that make these secondary material suitable as gravel substitutes in civil engineering applications. Difficulties in predicting the future leaching from these materials have often led to limitations in utilisation. To assess what will happen with a secondary material in the future, knowledge about the processes controlling the long-term leaching is required. Laboratory experiments can give good information about what will happen in the short-term, but as an experiment rarely lasts for more than a few years (often only for hours) the results cannot be extrapolated to make reliable predictions of long-term behaviour. Solid materials in a porous medium that comes in contact with mobile water react unless the water already is in equilibrium with the solids. Solid phases dissolve and new phases may form. Dissolved species are transported by advective flow and diffusion. These processes are a natural part of the evolution leading to, for example, weathering of rocks. With computer programs the leaching processes can be simulated for much longer times than is practical to observe experimentally. Programs can be used to simulate the geochemical evolution in time and space. They are often based on assumptions that there is local thermodynamic equilibrium at every point. Some models include dissolution and precipitation reactions. Examples of equilibrium programs are: HALTAFALL (Ingri et al., 1967), MINEQL (Westall et al., 1976), MICROQL (Westall, 1979), EQ3/6 (Wolery, 1992; Wolery and Daveler, 1992), PHREEQE (Parkhurst et al., 1980), WATEQ (Truesdell and Jones, 1974), and MINTEQA2 (Allison et al., 1991). Transport models, in addition, account for the mass balance. Many computer programs have been developed by combining an equilibrium model and a transport model. Examples hereof are: CHEQMATE (Harworth et al., 1988), TRANQL (Cederberg et al., 1985), PHASEQL/FLOW (Walsh et el., 1984), DYNAMIX (Liu and Narasimhan, 1989), and HYDROGEOCHEM (Yeh and Tripathi, 1991). If the reactions are not fast enough for equilibrium to be reached, dissolution and precipitation kinetics can be used. This approach is used in programs such as CHEMTRNS (Noorishad and Carnahan, 1987), PRECIP (Noy, 1990), MPATH (Lichtner, 1990), and CHEMFRONTS (B~iverman, 1993 and B~iverman et al., 1996).
438 Simulation of evolution of sharp fronts such as redox fronts and other simultaneously moving fronts has shown to be difficult with equilibrium based transport programs. CHEMFRONTS was originally developed to handle these type of fronts. Such fronts occur, for example, when oxygenated water infiltrates reduced rock. At redox and pH fronts that develop, some minor species can accumulate. These phenomena have been observed in nature, for example in the uranium mine in Poqos de Caldas in Brazil, where the uranium ore is located at the redox front (Cross et al., 1991). We have also observed accumulation at a pH front in laboratory experiment where iron has accumulated (Bfiverman, 1997a). As CHEMFRONTS is based on reaction kinetics it can handle non-equilibrium systems, such as glass phase dissolution and reactions with gas phases. We have found it necessary to modify CHEMFRONTS to be able to handle partially saturated systems with indiffusing carbon dioxide and oxygen. These systems are very complicated though and therefore no reliable predictions can be performed at this stage. More knowledge about the processes involved is required for this. Even though these simulations have large uncertainties, they are meaningful to perform. They can give information about what can happen and how variables, such as water infiltration rate, oxygen access etc. influence the leaching and therefore be of use when a construction incorporating waste material is planned. MODEL In a system where water flows through a solid material, chemical reactions between the solid and the liquid phase occur. The whole system can be defined as consisting of components, at most one for each element. The mass balance of the system is shown in equation 1
~) ((~Yj) + VWj at
M o~Xm - Z Vmj
m=l
at
( j = 1. . . . . N) (1)
where the first term, the amount of component j that has accumulated in the system, plus the transport of component j by fluid flow and diffusion, VWj, is equal to the changes in the mineral phase. ~ is the porosity, Yj is the total aqueous concentration of component j, including that present in the various complexes, M is the number of minerals, Vmj is the stoichiometric coefficient for the mineral m and the component j, Xm is the concentration of mineral m in the solid phase, and t is the time. The quasi-stationary state approximation (Lichtner, 1988), which this model is based on, describes the evolution of geochemical processes as a sequence of stationary states. A single volume of water that flows through a column reacts with the solid phase, dissolves parts of the mineral until the water volume is saturated with the existing minerals, and new minerals form if supersaturation is reached. The reactions of a subsequent volume of water are similar to the previous ones, as the changes in the mineral phase are small. When the mass of the components in the solution is very small compared to that in the mineral phase, the volume of water that must flow through the system to dissolve a substantial amount of a mineral is very large compared to the total volume of the column and its minerals. The accumulation of species in the water can then be ignored and the first term of equation 1 be neglected without any substantial loss in accuracy. When there is a large advective flux, the diffusive flux can be small by comparison. In this model the transport by diffusion is not accounted for, which limits the use of the program to systems where the flux is controlled by the advection. Equation 1 then becomes
439
M V-d-~z -~z_, Vmj ()Xm(z) m=l Ot
dYj=
(2)
for a one dimensional flow. v is the water flux in the z direction. Equation 2 states that the change in concentration of a component in the water is equal to the negative change of OXm components in the mineral phase. The mineral precipitation or dissolution rate, ~ , is given by OXm(r,t) = ~m(r,t) Im(r,t)
~)t
(3)
where ~m is a logical factor, which is unity both if the solution is supersaturated so that precipitation is possible, and if minerals are present so that dissolution is possible. Otherwise, ~m is zero. The rate of dissolution or precipitation, Ira, is Im(r,t) = CZm(r,t)kfm(Qm(r,t)-K~)
(4) the product of the specific surface of the mineral, am, the mineral reaction rate, k f , and the driving force of the system (Qm(r,t)-K~). The driving force is the difference between the ion activity product of the water solution, Qm, and the ion activity product at saturation (the inverse of the equilibrium constant of the formation for the mineral m)
J Qm(r,t) = I-I (aj(r,t)) vmj
j=l
(5)
Expressing the total concentrations in terms of component concentration we finally arrive at equation 6
M dCdz-- vl m~=l Vmj c)Xmot(z) = a function of (6) where A is the matrix which transforms component concentration C to total concentration Y for any component. The system of equations (6) are solved by a standard solver for systems of stiff differential equations. This gives the concentrations in water along the whole column. From these, the dissolution and precipitation at every point is obtained from equations 3-5. The depletion or accumulation of the minerals at every point is obtained by integrating equation 3 over the time step chosen. In practice, a small time step is taken and bookkeeping of the mineral changes is made. It is then found that in some points a mineral previously present has been depleted. The location where the mineral now starts to appear is the new location of that front. If some mineral has formed which was not there previously, a new front location is formed for that mineral. A new time step is started by solving (6) at the new front locations. This is repeated for as long as needed. It may be noted that the various fronts separate eventually, and the rate of the front movement becomes independent of the mineral reaction rate in the special case when the product of the specific surface and the reaction rate o%kfm is constant (B~iverman, 1993, B~iverman et al., 1996). Figure 1 shows an example where the pyrite and uraninite fronts
440
are coupled and move together. These fronts will not separate. The other fronts in the system are well separated from each other. Then there is no need for further calculations by time stepping. Extrapolation of the front locations in time can then be made.
5e-3
1~
4 e -3
~
9 ~
[] E"
3e-3
~r
2e-3
9
[] 9
1e-3 " J ~ t 0e+0 ...........
0
i~
1000 time
Figure 1.
Pyrite
o
Hematite
~
Chalcedony Uraninite
~:~ /t;~ :~
500
Kaolinite
9 K-Feldspar
~
1500
2000
2500
(years)
The location of the different fronts versus time. The pyrite and uraninite fronts are coupled, as uraninite is accumulated at the redox front.
The calculation typically starts with a homogeneous one dimensional column where water is flowing through with a constant rate. The calculations are made within a continuous column. The column is sub divided into several regions in each of which the same minerals exist. The regions are separated by "fronts". The positions of the boundaries are moved in every calculation step as the fronts (boundaries) move. As new minerals are formed or old minerals are exhausted, new regions are added or old ones removed. There is no minimum size of the regions, and their sizes can differ considerably. When the water from one region moves into the next it is generally not in equilibrium and the solid phases react. Some solid phases dissolve and the dissolved species form complexes. Other and new minerals may precipitate. The component concentration profile along the column is calculated by integration of Equations (6). In this process, the mineral dissolution and precipitation rates are also obtained. The process is illustrated by the following example where K-feldspar (KAISi3Os) is dissolved forming a silica mineral (here chalcedony (SiO2) and kaolinite (AlzSi2Os(OH)4), a clay mineral): KAISi308 (s) + H+ (aq) + 0.5 H20 K+ (aq) + 2 SiO 2 (s, 1) + 0.5 A12Si205 (OH)4 (s)
(7)
The dissolution and precipitation rate profile is shown in figure 2 when a front already has propagated a distance downstream in the column. At a front at the distance of 0.1510 m, in the figure, the water first comes in contact with the K-feldspar. The mineral starts to dissolve. This results in supersaturation of both kaolinite and chalcedony that precipitate. Protons are consumed, potassium and some silica are released into the solution. The process stops when the aqueous solution is saturated with respect to K-feldspar, kaolinite and chalcedony. This happens at a distance of about 0.1514 m, in the figure.
441 100
-
............~ a l c e d o n y
A
L
r
~
E
Kaolinite
o
E
~K-feldspar x -100
-"
,
0.1505
,
0.1510
Distance
Figure 2.
,
0.1515
0.1520
(m)
The dissolution (negative) and precipitation (positive) rate of the minerals involved when K-feldspar is dissolved.
The example used when CHEMFRONTS was developed was based on the uranium ore development in Polos de Caldas (Cross et al., 1991). The project was an international study of analogue processes. This study concerns redox and hydrolysis fronts and uranium mineralisation propagation at the Osamu Utsumi mine, an open pit uranium mine in Brazil, see figure 3. 9
600 m
. . . . . . . . . . . .
:.i!i:iii" -
,
,,
-
~.:::.:::
.......................
i ~ g / ~ ' . ~ . ! : ! ~ ! : i ~ ! : i ~ ! ~ : ~
iiilE4;iiiii!iii!iiilHi!ili]ililili!!iii!iiii;i!iii
ii
......................................................................................
iiiiiiliiiiiii.............. iiiiii
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~
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~
~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~~~:i~:N~:~~.......i~ ~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~:~
:-:-:':-:-:-:':--:-:-:-:-:-:':-:-:-:-:
:!.!:i.i:i.i:i.i:i.i:i.hi.i:~.!:!.hb!:~.!:~.h!.h~.h~.i:i.i:i.i:i.i:i.i:i.h~.i:i.i:i.h!.h!.!:~.h~.!:~.h~.h~.h~.!:~.hi.~.h~.hi.hi.id.i:i.hi.i •:•:::•:::•:::•:::•:::•::••:::•:::.:::•:::•:•:•:::.:::•:::••::•:::•:::•:::•:::.:::•:::•::••:::•:::•:::•:::.:::•:::•:::•:::•:::•:::•:::•:::•:•••:::••::•:::•:::•: •:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:.:•:•:•:•:•:•:•:.:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•••:•:•:•:•:•:•:•:•:•:•:•:•••:.:•:•:•: •:•:•:.:.:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:.:•:•:•:•:•:•:•:•:•:•:•••:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:•:
R e d u c e d rock
Figure 3.
W
O x i d i s e d rock
i
U r a n i u m mineral ( 2 0 0 - 2 5 0 0 p p m )
[~
M a i n fractures
A schematic picture of the uranium mine in Polos de Caldas, Brazil.
The upper part of the pit is oxidised and separated from the deeper-lying reduced rock by a redox front. Uraninite nodules are found in many places just below the redox front in the reduced rock.
442 The mineral distribution after 38 000 years of water infiltration is shown in figure 4. The front rates in the simulation agree well with the field results (B~iverman, 1993).
8000Chalcedony 6000 -O
t .............
K-feldspar
.o 4000-
i
...............................................................................................................
I Kaolinite O r t~
2000-
i
| | |
I Uraninite*1000 ......
| t
0 0.0 Figure 4.
Hematite 012
014
016
0 I.8
11.0 meter
The mineral distribution after 38 000 years of water infiltration.
This is an example where the flow rates are slow and front movements several orders of magnitude slower. The separated fronts can be extrapolated with no loss of accuracy in the calculations, as equilibrium is reached between the fronts. The kinetic approach gives an explanation for the uranium ore formation, just below the redox front, that could be difficult to explain by equilibrium calculations.
Examples of simulations performed with C H E M F R O N T S
Weathering of mine waste In the former case the time scale was hundreds of thousands of years. The front movement were only dependent of the mineral content and the front position could thereby be extrapolated. CHEMFRONTS can also be used to predict geochemical transformations in unsaturated porous media near the atmosphere-geosphere interface. In this case, the time scale is a few years. Extrapolation of front movements is not possible because reaction with the gas phase will take place in all the column. In this example, we simulate the breakthrough of copper in drainage water from mining waste, using data from the Aitik site in northern Sweden. Sulphide weathering, pH-buffering and copper (im)mobilisation are included. The principal processes in this case are as follows. The waste rock dump is partially saturated and the air in the pore-space contains oxygen (3-21%; Bennett et al., 1994). Sulphide minerals oxidise in this environment, which produces acidity, sulphate and dissolved metal species. The release of dissolved copper is of environmental interest and as its mobility is influenced by pH, it is necessary to include the processes affecting pH conditions. The acidity reacts with pH-buffering minerals, for example, calcite. The simulated breakthrough curves for the waste rock heap are shown in figure 5. During the first 13 years, the effluent pH is near neutral and aqueous concentrations are close to the measured concentrations of effluents from large experimental columns filled with fresh waste
443 rock (Str6mberg et al., 1994). The effluent pH then declines as a consequence of the calcite depletion, pH is for a short period buffered near 5.5 as a result of copper mobilisation. After this phase the pH stabilises near 3.5. The peak concentration of copper is obtained between 14 and 15 years, which is the result of the release of previously accumulated copper. -i
10
-
--%
-2 10
|
--
H
|
|
|~
,,m -..
w,
,,m m
~
m
m
~
,,m m
~
i
1 0 -3 __
|
Total Cu
1 0 -4 __
[ ........................................
'
o 1 0 -5 __ .l-a 9
1 0 -6 _
o
1 0 -7 _
(D
9
,_.+ ............................................ Ill-il -i-2
-8
10
-
Total Fe
.~
HgsIB~m,.ll~HBau~m . . . . . . . . . . . . . . . . . . . . al,*lCJ..~,~Smml.UUWll*U~DS~ll~' -9
10
-
-10
10
-
I 6
I 8
I 10
I 12
I 14
I 16
I 18
I 20
Time, years Figure 5.
The leachate concentration in the effluent from the waste rock heap.
Simulation of a leaching experiment CHEMFRONTS has been used to simulate a serial batch experiment where leach water has been flowing through a series of bottles containing waste material. The flow-through rate has been L/S= 1 in 20 days. The simulation and experiment has been carried out with electric arc furnace steel slag, blast furnace slag and municipal waste combustion bottom ash (B~iverman, 1997 a, b). The input data to the simulation has been chosen based on total composition of the materials, minerals identified with X-ray diffraction analysis, Scanning electron microscopy has been included in the simulations, both primary minerals and secondary minerals formed in leaching experiments. Trace minerals have been chosen to be the most probable, using an equilibrium program EQ3/6 and discussion with a geologist. The aqueous complexes included in the simulation are chosen from the results of the EQ3/6 simulation. In the simulation only the most important elements are included. The basic components, original minerals, and secondary minerals for the simulation with EAF steel slag are shown in table 1. In addition, 38 different aqueous species were included in the simulation.
444 Table 1.
The components and minerals used in the simulation with MSWIBA.
Components: calcium, hydrogen, iron, sulphur, copper, silicon, chlorine, sodium, lead, chromium, oxygen and carbon dioxide Original minerals"
larnite (~-Ca2SiO4), ferrous oxide (FeO), halite (NaC1), cuprous oxide (Cu20), Ca(OH)2, anglesite (PbSO4), chromic oxide (Cr203) and anhydrite (CaSO4) S e c o n d a r y minerals" calcite (CaCO3), gypsum (CaSO4.2H20), cupric oxide (CuO), goethite (FeOOH), lead chromate (PbCrO4), Ca-ferrite and andradite (Ca3Fe2(SiO4)3) Figure 6 shows the simulated results of the EAF slag experiment, for the iron, copper, chromium and silicon concentrations in the first leach water. The strange profile of the silicon concentration is caused by dissolution of one mineral and precipitation of another. In the whole column larnite (13-Ca2SiO4) dissolves. When the concentrations of iron and calcium, together with silicon, have reached saturation level, andradite (Ca3Fe2(SiO4)3) starts to precipitate. As the concentration of calcium continues to increase, the precipitation of andradite continues. This causes the decrease in iron concentration. As the silicon concentration in the steel slag leach water is below the detection limit, it is impossible to compare the results from the simulation with the experimental results.
Molar
1E-04 O....~---O ..-0
1E-05 -
,r
....
C r ' " O ' " O ' ' " ~
.O"
1E-06 d 1 E-07 _
~,.,,,
,~.....,Lk_._
..4~....
A ....
,dk-----d~---'A
"'"0"'-
Fe
...... @ .....
C u
---O--
Cr
""1[~..---~"'"
,q ~....... ~
...... "0" ....... O ....... "0" ...... " , 0 ....... 0 ....... -@" ....... 0 ....... 0
1E-08-
---A'-- SiO2
IE-09 -
1E-10s 3
Figure 6.
l
I
4
5
6
I
l
I
7
8
9
1'0
Cell
The concentration of iron, copper, chromium and silicon in the simulated column with steel slag.
The concentration at the end of the simulated column was compared with the experimental column. For most of the components the agreement was satisfying, but for copper and sodium the results disagree. There may be many reasons for this mismatch. The simulation was performed with only 12 elements whereas there were many more in the experiment. Some important elements may have been left out of the simulation that would considerably affect the result of the simulation. There may also be some solid phase or aqueous complex of importance for the solubility, of the elements included in the simulation, that were left out of the simulation. Some elements are very sensitive to the redox potential. As it is difficult to measure reducing Eh
445 in an oxic environment, the redox potential in the simulation may be wrong compared to the real value in the experiment, as it was based on the measured value. The simulations using experimental data for validation can be continued for longer time to simulate the long-term leaching from a waste material. This has been done and is reported elsewhere (Baverman et al., 1997a). The simulation conditions, flow rate, access to oxygen, concentration of inflowing water etc., can be varied.
Discussion There are several difficulties encountered when making computer simulations of long-term processes in waste materials. The materials are heterogeneous and formed under extreme conditions. Large parts of the waste materials are glasses. Geochemical computer models are developed for calculations of problems in geological systems. They often require well defined systems with thermodynamic data to make reliable simulations. These data are often not available for waste materials, especially not for the glass phases. It is therefore difficult to make reliable simulations in space and time for systems containing waste materials. CHEMFRONTS has been found to be a reliable program for studying reactions along advective water flow paths through porous bedrock systems. We have also been able to use CHEMFRONTS to simulate unsaturated flow in ashes, slags and mine waste. The program cannot, at this stage, be used to make good predictions of what will happen in an unsaturated system, as more knowledge about the processes is required. It can, however, be used to give an indication of potential risks etc. Even though these simulations have large uncertainties, we find them meaningful to perform.
Acknowledgments Financial support has been provided by AFR/Swedish Environmental Protection Agency.
References Allison, J.D., D.S. Brown and K.J. Novo-Gradac (1991) MINTEQA2/PRODEFA2, A geochemical assessment model for environmental systems, EPA/600/3-91/021, March. Bennett J. W., Gibson D. K., Ritchie A. I. M., Tan Y., Broman P. G. and J6nsson H. (1994) Oxidation rates and pollution loads in drainage, Correlation of measurements in a pyritic waste rock dump, Bureau of Mines Special Publications SP06A-94, 400-409. B~iverman, C. (1993) Development of "CHEMFRONTS", a coupled transport and geochemical program to handle reaction fronts, SKB Tech. Report 93-21, Swedish Nuclear Fuel and Waste Management CO, Box 5864, S-102 48 Stockholm, Sweden. B~iverman, C. (1997a) The importance of the pH buffering capacity - Comparison of various methods to estimate the pH properties of a waste material, Fifth Annual North American Wasteto-Energy Conference, Research Triangle Park, North Carolina, USA, April 22-25, 1997. B~iverman, C. (1997b) Long-term leaching mechanisms of ashes and slags; Combining laboratory experiments with computer simulations, Ph.D. thesis, TRITA-KET R65, ISSN 1104-3466. B~iverman, C., A. Sapiej, L. Moreno and I. Neretnieks (1997) Serial Batch Tests Performed on Municipal Solid Waste Incineration Bottom Ash and Electric Arc Furnace Slag, in Combination With Computer Modelling, Waste Management and Research, in press. Cederberg, G.A., R.L. Street, and J.O. Leckie (1985) A groundwater mass transport and equilibrium chemistry model for multicomponent systems, Water Resour. Res. 21 (8), 10951104.
446 B~iverman, C., Str6mberg, B., Moreno, L. and Neretnieks, I. (1996) CHEMFRONTS: a coupled geochemical and transport simulation tool, KAT 96/36, submitted to Journal of Contaminant Hydrology. Cross, J.E., A. Harworth, I. Neretnieks, S.M. Sharland, and C.J. Tweed (1991) Modeling of redox front and uranium movement in a uranium mine at Polos de Caldas, Radiocim. Acta, 52/53,445-451. Harworth, A., S.M. Sharland, P.W. Tasker and C.J. Tweed (1988) A guide to the coupled chemical equilibria and migration code CHEQMATE, Harwell Laboratory Report, NSS R113. Ingri, N., W. Kakolowicz, L.G. Sillen, and B. Warnquist (1967) High speed computers as a supplement of graphical methods - V. HALTAFALL: A general program for calculating the composition of equilibrium mixtures, Talanta, 14, p 1261. Lichtner, P. (1988) The quasi-stationary state approximation to coupled mass transport and fluid-rock interactions in a porous medium, Geochim. Cosmochim. Acta 52, 143-165. Lichtner, P. (1990) Redox front geochemistry and weathering: theory with application to the Osamu Utsumi uranium mine, Polos de Caldas, Brazil, Submitted to Chemical Geology, Special issue on the Polos de Calds, December 3. Liu, C.W., and T.N. Narasimhan (1989) Redox-controlled multiple reactive chemical transport, 1. Model development, Water Resour. Res., 25(5), 869-882. Noorishad, J., and C.L. Carnahan (1987) Development of the non-equilibrium reactive chemical transport code CHMTRNS, DE-AC03-76SF00098; LBL-22361. Noy, D.J. (1990) PRECIP: A program for coupled groundwater flow and precipitation/dissolution reactions, National Environment Research Council British Geological Survey, Technical Report WE/90/38C. Parkhurst, D.L., D.C. Thorstenson, and L.N. Plummer (1980) PHREEQE - A computer program for geochemical calculations, Report USGS/WRI 80-96, NTIS Tech. Rep. PB81167801. Str6mberg B., S. Banwart, J.W. Bennett and A.I.M. Ritchie (1994) Mass balance assessment of initial weathering processes derived from oxygen consumption rates in waste sulfide ore. Bureau of Mines Special Publication SP 06B-94, 363-370. Truesdell, A.H., and B.F. Jones (1974) WATEQ, a computer program for calculating chemical equilibria of natural waters, J. Res. U.S. Geol. Surv. 2(2), 233-248. Walsh M.P., S.L. Bryant, R.S. Schechter, and L.W. Lake (1984) Precipitation and dissolution of solids attending flow through porous media, AIChE J., 30(2), 317-328. Westall, J.C., J.L. Zachary, and F.M.M. Morel (1976) MINEQL: A computer program for the calculation of chemical equilibrium composition of aqueous system, Tech. Note 18, 91 pp., Dep. of Civ. Eng., MIT, Cambridge, Mass. Westall, J. (1979) MICROQL: 1 A chemical equilibrium program in BASIC, EAWAG, Swiss Fed. Inst. of Technol., Duebendorf, Switzerland. Wolery, T.J. (1992) EQ3NR, A computer program for geochemical aqueous speciationsolubility calculations: Theoretical manual, user's guide and related documentation (Version 7.0), Lawrence Livermore National Laboratory, September 14. Wolery, T.J., and S.A. Daveler (1992) EQ6, A computer program for reaction path modeling of aqueous geochemical systems: Theoretical manual, user's guide and related documentation (Version 7.0), Lawrence Livermore National Laboratory, October 9. Yeh, G.T., and V.S. Tripathi (1991) A model for simulating transport of reactive multispecies components: model development and demonstration, Water Resour. Res., 27(12), 3075-3094.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
447
OVERVIEW OF G E O C H E M I C A L PROCESSES C O N T R O L L I N G L E A C H I N G CHARACTERISTICS OF MSWI BOTTOM ASH Jeannet A. Meima and Rob N.J. Comans*
ABSTRACT In the past few years important progress has been made in understanding the geochemical processes that control leaching of various elements from MSWI bottom ash. An overview of that progress is presented here. Bottom ash leaching has been observed to depend strongly on the ageing/weathering of the solid material. Three major stages in weathering have been identified, each stage having a characteristic pH which is controlled largely by Ca-minerals and pCO2, but also by soluble-A1 and -SO4. In the first two stages, which are characterised by relatively unweathered bottom ash of pH> 12 or pHI0-10.5 respectively, the general processes are thought to be precipitation/dissolution of relatively soluble minerals. In the third stage, the leaching of several elements has decreased, including the potential contaminants Cd, Pb, Cu, Zn, and Mo. This reduction in leaching is due to (a) the neutralisation of bottom ash pH, (b) sorption processes or formation of more stable mineral species, and (c) reduced leaching of dissolved organic carbon (DOC). The latter process is important for Cu in particular, since >90% of the dissolved copper may be associated with DOC. It will be discussed how the obtained knowledge of these geochemical processes can be applied to make reliable predictions of the long-term behaviour of bottom ash in the environment.
1. INTRODUCTION Combustion residues, such as Municipal Solid Waste Incinerator (MSWI) bottom ash and alkaline coal fly ash, are produced world-wide in ever-increasing quantities. Combustion residues, however, may pollute the environment because they are enriched in potentially toxic elements relative to soils and sediments [1,2]. For a proper assessment of the environmental impact of the utilisation and disposal of these ashes it is, therefore, necessary to understand both the short-term and the long-term processes that affect the mobilization of potentially hazardous elements from ash residues. Recently, considerable progress has been made in understanding the behaviour of alkaline (waste) materials in the environment. By considering waste materials as mineral assemblages, similar to rocks and soils, it has been shown (a) that upon weathering the high-temperature solids that form during combustion will transform into naturally occurring secondary minerals [e.g. 3-8], (b) that element leaching can be modelled/predicted by common geochemical processes such as dissolution/precipitation, sorption, redox, and complexation processes [e.g. 1,3,7,9-21], and (c) that element leaching depends strongly on the ageing/weathering of the solid material [5,7,8,19]. This paper gives an overview of geochemical processes that control element leaching from MSWI bottom ash at different stages of weathering. Although we focus on MSWI bottom ash only, the geochemical processes discussed here will contribute to a more general understanding of the behaviour of combustion residues or alkaline (waste) materials in the environment.
*Author to whom correspondence should be addressed; telephone: +31 224 564218; fax: +31 224 563163; email address: [email protected]
448 2. MSWI BOTTOM ASH
Incineration is a viable management strategy for treating combustible municipal solid waste that cannot be recycled. The waste volume is strongly reduced while exothermic energy is recovered. Incinerators usually operate at temperatures averaging from 850-1000~ depending on fumace-design and on the caloric value of the waste [22]. The residence time of the waste in the incinerator varies from 45 to 90 minutes. The heavier ash residue that is collected Table 1 from the combustion chamber is called Element concentrations generally found bottom ash. Other waste streams are in MSWI bottom ash world-wide* electrostatic precipitator (ESP) ash, air pollution control (APC) residues, and grate concentration (ppm) elements siftings. The hot bottom ash is quenched in <500000 Ca, Fe, O, S i a water tank immediately after < 100000 A1, C, Mg, Na incineration. The grate siftings are usually mixed with the bottom ash, whereas in <20000 K, Pb, S some incinerators ESP ash is also mixed < 10000 Cu, P, Ti, Zn with the bottom ash. <5000 Ba, C1, Cr, Ni Given the variation in waste <2500 F, Mn, N, Sr composition, furnace configuration, <500 As, B, Br, Co, combustion temperature, retention time, Mo, Sb, Sn, V and quenching process, the elemental < 100 Ag, Au, Cd, Cs, composition of MSWI bottom ash from Ga, Hg, I, La, different origin is remarkably similar: bulk Rb, Sc, Se, Y chemical analyses of ashes from different facilities usually fall within an order of * data from Ref 18 magnitude [2]. Table 1 gives the order of various element concentrations found in MSWI bottom ash world wide [18]. MSWI bottom ash is a highly reactive material because it consists predominantly (>70%) of X-ray amorphous, glassy constituents [2,14,23] and because it has a relatively high surface area due to internal porosity [11,18,22]. The high glass-content results from the rapid cooling (quenching) of the hot, partly molten material. Petrographic analysis has led to the following classification scheme of intact bottom ash particles [ 14,24]:
1. non-combusted materials (15-45 %) 9 e.g. waste glass, soil minerals (pyroxenes, quartz, and feldspars), metals, metal alloys, and organics 2. melt products (55-85%) 9 glasses (isotropic silicate glass, schlieren, and opaque metallic glass) 9 crystalline complex silicate minerals e.g. melilite group minerals which are rich in Fe and Ca and depleted in A1, and scapolite-like minerals which are rich in Ca and Na and depleted in Fe 9 crystalline complex oxide minerals e.g. lime, iron oxides, and spinel-group minerals
Table 2 Overview of geochemical processes that have been reported to control the leachingof maior and trace elements from st different stages of weathering. and minerals/species identified in the matrix by spectroscopic techniques.
MSWl bottom ash
Element
unweathered (A-type) bonom ash, pH>12 contmlllng mechanism identltled specles
quenchsdlnorrcarbonated (&type) bottom ash, pHlblO.5 controlling mechanlam identltled species
carbonated (Glype) bottom ash, pHgS.5 controlling mechanism Idemifled species
Ca. S04, C03
solubllltycontrol by portlandlte'and gypsum'
salub~l~tycontral by e~ringlte'.'.~' I and gypsum5'~"~'2~'3~25
enr~ngne5.'.'3,anhydrite2' CalCile25.1.?4.21.27
solubilityconlrol by calcite5' and gypsum'
anhydme7.
gypsum2u4223.27
CaO"'" whihockde"; CaHP04", camplex-(Ca)s~licates"~~' gibbsite", c o r ~ n d u m ' ~ ~ AIO'~, ', elementaValloy'4~2'.complex~ l i c a t e s ' ~ . her~ynite'~. ~', M~ACO?
solubilitycontrol by glbbsite' or amorphous AI(OH)~"
amorphous AI(0H);
solubll~lycontrolby ferrihydr~te'
magnetde5.'. hematites7,wusiiie'*. maghem~te'~, Iron oxnde"
A1
sol~billty~ontrol by hydrous Al-sllicate7
Fe
anhydrile'.2', calcite'.". partland~te",eomplex(Ca1slllcate"
anhydrie",salubll~tycontral by calclte"
campiex-s~licates~'
solub~l~tycantrol by g i b b s ~ l e ' ~ " " ' ~or amorphous AI(oH)I'.B'~
magneli~e'.~',hematvie', pseudobraakite2'
solubil~tycontrolby ferrihydr~te"~
hematite2' " 232', w ~ s l i t e ~ ~ ' ~ ~ ~ ' , maghem~te". goeth~te~', pyre'"', pseudobrookde2', magnetite'"'232' ulvospinel" hereynile", jacobsite2'. chr~mlte'~."
''.
'',
magne~ite'~, dolomde", M ~ s o , . ~ H ~ o ' ~ .solubil~tycontrolby M g ~ calcite', ssplolde'. or sepiolite2', complexsilicater2'. dolomlte5 MgA120,'4 q u a d s ~ " 4 2 3 2glasses2' 7. ", ~ ~ I ~ b i l i t y c o n by t r o~llite' l
ca~cde~'.~
Mg
solubililycontrol by brucae'
complex-silicates2'
ralubil~tycanlrolby b r u ~ i l e "or~ ~ ~ magnesite"
SI
Solubllltycontrol by hydrous ~ksilicate'
quartz',27, ~om~lex-s~l~cates'~'
ralub~l~tycantral by wairakite" or amorphous S10,'
melall~dallay"
organic c~mplexatlon'~ 'P.2'.2' and 'g.2'28 ~ o l ~ b ~ l l t y e by ~ n~enarite'~ tr~l or CU(OH)~"
metalliciall~y'~
solubii~lycontrolby c e r r ~ s ~ t e " . " ~ ~ complexs!l~cate'~'~ ". pbOi4 or ~ b ( 0 H h ""or s a r p t t o n c ~ n f r o l ~ ~
m e t a l l i ~ a l l ~ ycomplex-~ilicate'~. '~, sorpt~oncontrol'e'Oor PblO-. PblCalO-. PblFeiO-. and PblAIIO solubililycontrol by ~ h l o r ~ ~ r o m o r p h ~ t e ' ~phases"; sorbed to neolormed clays2*
metalldalloy"
~ ~ l ~ b i l i t y ~ o by n t rzinclte".". ol Zn(OH1,". or ZnSi03" 14, sorpt~onconlr01'~
sorptionzontro12"or s o l ~ b ~ l ~ t y ~ o by ntrol zinc~te"
Na, K. CI MO
~~lubtlilycontrol by p~welliie'~
Cu
Pb
Zn
salubllltycantrol by ~incile'~
Cd
solubil~lyanlrolby otavile" or Cd(0Hl~"
Dlher
"
hydrous Al-silacate5, complex-silicates2"." 2327 complex-(~d~)rilicates~' haltte2,sylvde2, complei-(~al~)sibcates'~~' ~ o l ~ b ~ l t t y e ~by n tp r o~lw e l l i t e ' ~ . " ~ ~ Ca~d)," metalliciall~~" ID". cucl". N ~ C U P O ~ " ,
carnpiex-silicate"
rmc~te".ZnSO.". ZnClF, elemental Zn". Irankl~n!te'*
~ ~ l ~ b i l i t y ~ o by ntrol otavite'", t ' b ' 9 2 8 or sorption to caklle"" possible solublldycontrolillng mlnerals lor Mn, Ba, and V: MnO(0H)": barlte". Pb,V2OIt'
or amorphous SiO,'
hydrous ~ l - s ~ l ~ c a tquartl',', e ~ ~ ~ ~ illtte8. ', complex-sklicatei'
sorption l o
CaMoO.'P. PblMo-rlch particle2' ferrihydrlte'"" CulOphases 1e.g. cuprlte. teno~ite)'~. organlc complexahonz' and sorption to metalliciall~y'~, ~omplexsilicate'~. amorphous A I ( O H ) ~ ~ ~CulCalO-, ~~' CuIFdO-. CUIAIIO-. CUIAIIOISIC~-.CulS-phases", sorbed to neoformedclayszD
ZnlOphases (e.g. ~ i n c i t e l ' ~ . metall~clallo~'~, complex-s~licate'~. ZnlCalO-. ZnlFelO-, ZnlAllO-, ZnlAVOlS phases", sorbed to neolormed claysz9
sorpt~oncontrol'~'~
rutlle2.". bar~te"". graph~ticcarbon". taenle". (spemes containing Sb, Sr. NI. Cr. As. Ag, Rb, Nd". Cr+xldes2')
NI sorbed to neoformedclaysZg
Mineral formulas' anhydnte. CaSOI, barite, BaS04; bruclle. Mg(0HL; calcde. CaCO,; cerruslte. PbCO.; chloropyromorphlfe. Pb5(POJ3CI;chrom~te,FeCr~0.; corundum. A,O i .: cuprite. Cu20; dolomite. CaMg(C03)7,enringite. C&I~(SO&OH)I~.~~H,O; lerrihydrite. Fe(OH)3, Iranklinite. ZnFe,O,, glbbsite. AI(OH),; goethlte, FeOOH: gypsum. CaS042H20, halite. NaCI, hematite, Fe?03,hercynde. FeAi20.; illlte. )6,Mg..Al~sSbsOlo(OH)2; Iron oxlde, Fe20.; iacobsde. MnFezO,. pyrite. FeS; quartl, SO2; rutile. TlO?, seplollte. Mg2S~.0610H)..1.5H~O: syklte. KCI; maghemite. Fe203;magnesde, MgCO,, magnetole. Fe,O,; otavtte. CdCO.: portlandtte. Ca(OH),; paweilde. CaMa04, pseudobrooklte.F~PTIOS; taenite. Fe.Ni; tenonte. CuO: ulvosptnel. Fe,Ti04; walraktte. CaAhSi,O,,.ZH,O; whiilock~te.CadP04)2;wustite. FeO: zincite. ZnO.
450 The sequence of reactions during incineration has been compared to a melt of melilite-bearing igneous rock which can be described by a CaO-MgO-A1203-SiO2-Na20-FeO system [14]. Equilibrium, however, is not obtained, which is illustrated by residual organics in the bottom ash (Table 1) and by the occurrence of thermodynamically incompatible phases in the bottom ash (e.g. quartz and mellilite, ref. 14,24). A detailed overview of minerals and species identified in MSWI bottom ash is given in Table 2. Potentially hazardous elements in MSWI bottom ash include heavy metals (e.g. copper), oxyanions (e.g. molybdenum and antimony), and soluble salts (e.g. sulphate and chloride) (see Table 1).
3. W E A T H E R I N G As shown above, MSWI bottom ash consists primarily of metastable solids. Upon weathering these solids transform into naturally occurring secondary minerals. Weathering has been shown to strongly affect the leaching of major and trace elements from MSWI bottom ash [5,7,8,19]. In general, weathering reactions in MSWI bottom ash have been shown to be similar to those observed in alkaline soils and volcanic ashes [4,5] and basalts [4]. Three major stages in weathering have been identified, each stage having a characteristic pH that is controlled largely by Ca minerals and pCO2 [7]. Figure 1 illustrates the changes Ca-leaching as weathering continues. Important characteristics of the three weathering stages are discussed below.
10000 _1 E
t~
o
~_ gypsum
1000100
_~,
-
t
\
10-
calcite~
1
2
,
,
,
," ~ ~ ,
4
6
8
10
12
14
pH Figure 1. Total dissolved Ca in type A (O), type B ( g ) , and type C (A) bottom ash leachates at L/S=5 as a function of pH, and MINTEQA2 predictions assuming equilibrium with different mineral phases. The style of the lines indicates the category of bottom ash on which the modelling was performed: ...... A-type; ~ B-type; - - C-type bottom ash. (Modified after Ref. 7.)
451 (A) unweathered bottom ash, with pH > 12 Stage-A represents the initial alteration processes which take place when the dry bottom ash first contacts water, which is in the quench tank. Reactions include the hydrolysis of the oxides of Ca, A1, Na, and K, and the dissolution/reprecipitation of hydroxides and salts of these main cations [7,25,30]. The resulting bottom ash pH is strongly alkaline (12.4) and controlled by the solubility of portlandite (Ca(OH)2) [7]. (B) quenched/non-carbonated bottom ash, with pill 0-10.5 In stage B bottom ash pH has been decreased to 10-10.5 by the formation of ettringite (Ca6A12(SOa)3(OH)12.26H20), gibbsite (AI(OH)3), and gypsum (CaSOn.2H20) [7,25]. When the three minerals coexist, no degrees of freedom are left and pH is fixed at pH 10 [7,25]. Due to continuing hydrolysis secondary minerals such as amorphous Fe/Al-(hydr)oxides, hydrous aluminosilicates, and possibly zeolites begin to precipitate [4,5,7]. Soluble salts will be leached rapidly with percolating water [e.g. 4,5,7,11,30]. Biodegradation of residual organic matter and dissolution of reduced mineral phases may create a reducing environment [5,18]. (C) carbonated bottom ash, with pH8-8.5 In stage C bottom ash pH has further decreased to equilibrium values of 8-8.5 by absorption of CO2 and subsequent precipitation of calcite (CaCO3) [e.g. 5,7,8]. The CO2 required for this carbonation may infiltrate from the atmosphere or come from biodegradation of organic residues [5,26,30]. The neoformation of Fe/Al-(hydr)oxides and hydrous aluminosilicates continues. Similar to the weathering of volcanic ashes, these hydrous aluminosilicates are an intermediate reaction product in the transformation of glasses to clayminerals [6]. The 2:1 clay mineral illite seems to be the final product of glass weathering in MSWI bottom ash [6]. Weathering has been shown to have a significant effect on the leaching of trace elements from MSWI bottom ash [8,19]. The leaching of Cd, Pb, Cu, Zn, and Mo from C-type bottom ash, for example, is generally significantly lower than from more fresh bottom ash [8,19]. A potentially important mechanism is the sorption of trace elements to neoformed (amorphous) Fe/Al-minerals [8,19,20,29]. Furthermore, the neutralisation of bottom ash pH from > 10 to 8-8.5 and the formation of less soluble secondary minerals of trace-elements also contribute to reduce leaching [ 19]. Lower trace-element leaching from weathered bottom ash does not seem to be caused primarily by a prior release of these elements from the residues during storage [ 19,31 ].
4. LEACHING The rate at which an element is leached from the bottom ash is dependent on its abundance in the bottom ash, its availability to the solution, the dissolution kinetics of the primary solids containing the element, whether or not the element will reprecipitate as a secondary solid or will sorb to solid substrates, and the kinetics of these precipitation/sorption reactions [13]. Kirby and Rimstidt [13] have identified 3 basic types of solution behaviour duringbatch leaching of MSWI bottom ash: 1. availability, which means that there is a lack of concentration-change due to exhaustion of a phase. This type of behaviour is usually observed for soluble salts, such as Na, K, and C1 [5,7,13,18]. Furthermore, molybdenum may show this type of behaviour at strongly alkaline pH [7,18]. In general, the higher the Liquid to Solid (L/S) ratio, the more elements will show this type of behaviour.
452 2. kinetic, which means that the rate of mass transfer from the solid to the liquid phase or v.v. is the concentration-limiting step. The contact time between the solid and the liquid phase usually determines whether kinetics are important or not. In general, two steps can be observed in element leaching from bottom ash: a fast release of the element, which is generally completed within 24 h, followed by a slow release or re-binding which may continue for more than 1 week [5,8,13]. The leaching of silicon, for example, is strongly influenced by slow dissolution/precipitation kinetics of silicate-minerals [5,8,13]. Furthermore, the slow transformation of the primary high-temperature solids into stable secondary solids has been shown to affect the leaching of several other elements as well [7,19]. Little is known, however, about the kinetics of these weathering reactions. Alternatively, a slow release may also be the result of diffusion processes, which are believed to become important when the residues are monolithic in form (e.g. incorporated into asphalt pavement), when they are compacted to low permeability, or when they are overlain by an impermeable barrier [ 18,32]. 3. equilibrium, which means that the concentration of an element is controlled by a dissolution/precipitation equilibrium or by a sorption equilibrium. Various elements experience retention in the bottom ash matrix by these processes: Table 2 gives an overview of the proposed controlling-mechanisms for MSWI bottom ash at different stages of weathering. Below, we review underlying geochemical processes, such as complex formation, dissolution/precipitation, sorption and redox reactions, which control element leaching from MSWI bottom ash.
5. G E O C H E M I C A L PROCESSES C O N T R O L L I N G LEACHING
complexation processes Hydrolysis and complexation with carbonate are the dominant inorganic complexation reactions in bottom ash leachates. These reactions cause, for example, the solubility-curves of amphoteric elements such as Fe, A1, Zn, Cu, and Pb to follow V-shaped patterns as a function of pH [33]. Figure 2 illustrates the effect of hydrolysis on the solubility of Zn. As a result, pH is a dominant controlling parameter in element leaching from (waste) materials, which is in correspondence with experimental data [e.g. 7,12,16,19,30,34,35]. Other potentially important inorganic complexes include Cd-C1 complexes, which may become significant in leachates from fresh bottom ash [16]. MSWI bottom ash releases substantial amounts of dissolved organic carbon (DOC) originating from incomplete burning of the original waste and/or subsequent biodegradation processes [7]. Copper, which is known to have a very high affinity for organic material [36], has been shown to be bound for >90% to DOC in leachates of both fresh and 1.5-year old MSWI bottom ash [21 ]. The conditional stability constants of these Cu-DOC complexes have been determined using a competitive ligand exchange / solvent extraction technique [21]. Figure 3 illustrates the importance of this organic complexation on the leaching of Cu from fresh MSWI bottom ash, and shows that Cu-leaching under environmental conditions (pH>7) is dominated by this process.
453
1E-02 _1
g
c N
1E-03
-
1E-04
-
1E-05
-
1E-06
-
O
~
o
. o
....
Zn(;~,:
1E-07
I "
"
o.: "
I
5
'
-
7
"
- 1
"
I
".o ~i
,"
8' I
I
9
"z"(OH)/I
11
13
pH Figure 2. The effect of inorganic complexation on the solubility of zinc. The solid line represents the predicted total concentration of zinc in equilibrium with the mineral zincite (ZnO). Dashed lines represent concentrations of corresponding Zn-species. Symbols represent total dissolved Zn in type B bottom ash leachates at L/S=5. Data were taken from Ref. 19.
1 E-04
O
'
~Total
E-O5
r~ 1 E - 0 6
Cu
I1~,,
-
c u'§
1E-07
"
Cu-0rg
-
"
-
- ~ ----
" ~ . Cu-inorg ,
6
-
0
7
~
,"
,
,
8
9
10
11
pH Figure 3. The effect of organic complexation on the leaching of Cu (O) from fresh MSWI bottom ash at L/S=5. The solid line represents the predicted total concentration of Cu in equilibrium with the mineral tenorite (CuO). Dashed lines represent corresponding concentrations of Cu 2§ the sum of the inorganic Cu-complexes, and the sum of the organic Cu-complexes. (Modified after Ref. 21.)
precipitation/dissolution processes
Precipitation/dissolution processes control bottom ash pH (see above) and the leaching of in particular major elements from MSWI bottom ash (Table 2, Figure 1). In the case of major elements, solubility-controlling minerals indicated by geochemical modelling generally correspond to minerals detected by spectroscopic analysis of the bottom ash (Table 2). Precipitation/dissolution processes may also control the leaching of trace elements from Aand B-type bottom ash. Proposed controlling processes for trace-element leaching, however, are often indicated by geochemical modelling only (Table 2) because low bulk concentrations hamper the detection of trace-element species by means of spectroscopic techniques [2,19]. A
454 step-wise approach for the geochemical modelling of element-concentrations in equilibrium with potential solubility-controlling minerals is given in Meima and Comans [7].
sorption processes Sorption is a general term which refers to all processes, except the precipitation/dissolution of pure mineral phases, which remove a chemical species from the aqueous solution to a solid phase. Sorption processes are expected to be important when suspensions at equilibrium are undersaturated with respect to known solubility-controlling minerals. Potential sorbent minerals in MSWI bottom ash are amorphous or crystalline Feand Al-(hydr)oxides, hydrous aluminosilicates, and calcite [8,20]. Recent studies have shown that surface complexation reactions can successfully describe the leaching of trace-elements from combustion residues, such as MSWI bottom ash and coal fly ash [9,20,37-39]. In addition, trace-elements have been found to be associated with secondary and potential sorbent minerals in weathered MSWI bottom ash (Table 2). A step-wise approach for the modelling of surface complexation or surface precipitation processes in heterogeneous systems such as MSWI bottom ash is described by Meima and Comans [20]. This approach is based on (1) the database of surface complexation and surface precipitation reactions and associated equilibrium constants for sorption of ions on Hydrous Ferric Oxide [40], (2) 'selective' chemical extractions to obtain the available sorbent mineral concentrations, and (3) leaching of the bottom ash at pH-values unfavourable for sorption to obtain the available trace-element concentrations. The identification and modelling of sorption processes in heterogeneous solid systems such as MSWI bottom ash is, however, at its beginning. Because of their potential importance, these processes deserve considerably more attention in future research.
redox processes In fresh MSWI bottom ash the prevailing redox conditions are oxidizing [7,11]. During disposal or utilization of the bottom ash, however, the redox potential may decrease strongly by biodegradation of residual organic matter and/or by the presence of reduced mineral phases [5,18]. Relatively low redox potentials were recorded, for example, in percolate from landfilled combined MSWI bottom and fly ash [41] and in a 6-week old storage of fresh MSWI bottom ash [7]. Variations in bottom ash EH may affect metal mobilities by: 9 directly changing the oxidation states of redox sensitive elements to more soluble/insoluble species. The leaching of Cu [18,35,42,43], Cr [18,43], As [43], and V [43], for example, has been shown to increase toward more oxidizing conditions, whereas the leaching of Fe was decreased [7,35,43]. 9 changing the amount of redox sensitive metal surfaces (Fe/Mn-(hydr)oxides) available for sorption [35]. 9 changing the degree of (co)-precipitation or complexation with other redox sensitive cations and anions, e.g. the precipitation of heavy-metal sulphides [ 18,35,41,43]. The cited studies show that the influence of EH on metal solubilities in MSWI bottom can be significant and that further research on this topic is required.
455
CONCLUSIONS AND RECOMMENDATIONS
F O R F U T U R E RESEARCH
The leaching of major and trace elements from M S W I bottom ash has successfully been described on the basis of geochemical processes such as complexation, precipitation/dissolution, and sorption processes. For the prediction of the long-term behaviour of M S W I bottom ash in the environment the results imply that: 9 materials should be tested at pH 10 and/or pH 8.3, depending on whether or not the materials are used in contact with air and may become carbonated; 9 the concentrations of toxic elements in leachates are likely to be greatest in the earliest stages of disposal: the most soluble phases dissolve rapidly, while the capacity of secondary minerals to bind trace elements may not be large enough. Furthermore, m o l y b d e n u m is very mobile at alkaline pH. These problems may be overcome by (1) neutralising the pH of the bottom ash, and (2) by adding sorbent minerals to the bottom ash [44]. 9 on the long-term, the leaching of toxic elements is likely to be reduced by the neutralisation of bottom ash pH and by sorption to neoformed minerals. Future research should concentrate (a) on a further identification/modelling of these sorption processes and (b) on the kinetics of these weathering/sorption reactions, i.e. the period of time that is required to obtain a sufficient reduction in trace-element leaching. 9 Little is known about the processes that control element leaching from M S W I bottom ash in reducing environments, which, therefore, also requires further research.
REFERENCES 1. 2. 3. 4. 5.
6. 7. 8. 9. 10. 11.
Eary L.E., Rai D., Mattigod S.V., and Ainsworth C.C. (1990) Geochemical factors controlling the mobilization of inorganic constituents from fossil fuel combustion residues: II. Review of the minor elements. J. Environ. Qual. 19, 202-214. Kirby C.S. and Rimstidt J.D. (1993) Mineralogy and surface properties of municipal solid waste ash. Environ. Sci. Technol. 27, 652-660. Mattigod S.V., Rai D., Eary L.E., and Ainsworth, C.C. (1990) Geochemical factors controlling the mobilization of inorganic constituents from fossil fuel combustion residues: I. Review of the major elements. J. Environ. Qual. 19, 188-201. Kirby C.S. A Geochemical analysis of municipal solid waste ash, Ph.D. Dissertation, Virginia Polytechnic Institute and State University, 1993. Zevenbergen C. and Comans R.N.J. (1994) Geochemical factors controlling the mobilization of major elements during weathering of MSWI bottom ash. In Environmental Aspects of Construction with Waste Materials (Eds J.J.J.M. Goumans, H.A. van der Sloot, and Th. G. Aalbers), pp. 179-194. Elsevier Science B.V., Amsterdam. Zevenbergen C., van Reeuwijk L.P., Bradley J.P., Bloemen P., and Comans R.N.J. (1996) Mechanism and conditions of clay formation during natural weathering of MSWI bottom ash. Clays and Clay Minerals 44, 546-552. Meima J.A. and Comans R.N.J. (1997) Geochemical modelling of weathering reactions in municipal solid waste incinerator bottom ash. Environ. Sci. Technol. 31, 1269-1276. Meima J.A., van der Weijden R.D., Eighmy T.T., and Comans R.N.J. The effect of carbonation on traceelement leaching from municipal solid waste incinerator bottom ash. Submitted for publication. Theis T.L. and Richter R.O. (1979) Chemical speciation of heavy metals in power plant ash pond leachate. Environ. Sci. Technol. 13, 219-224. Fruchter J.S., Rai D., and Zachara J.M. (1990) Identification of solubility-controlling solid phases in a large fly ash field lysimeter. Environ. Sci. Technol. 24, 1173-1179. Theis T.L. and Gardner K.H. (1992) Dynamic evaluation of municipal waste combustion ash leachate. In: 5th International Conference on Ash Management and Utilization. pp27-65. Arlington, VA.
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Comans R.N.J., van der Sloot H.A., and Bonouvrie P.A. (1993) Geochemical reactions controlling the solubility of major and trace elements during leaching of municipal solid waste incinerator residues. In Municipal Waste Combustion Conference., pp. 667-679. Air and Waste Management Association, Williamsburg, VA. Kirby C.S. and Rimstidt J.D. (1994) Interaction of municipal solid waste ash with water. Environ. Sci. Technol. 28, 443-451. Eighmy T.T., Eusden jr. J.D., Marsella K., Hogan J., Domingo D., Krzanowski J.E., and St~npfli. D. (1994) Particle petrogenesis and speciation of elements in MSW incineration bottom ashes. In Environmental Aspects of Construction with Waste Materials (Eds J.J.J.M. Goumans, H.A. van der Sloot, and Th. G. Aalbers), pp. 111-136. Elsevier Science B.V., Amsterdam. Eighmy T.T., Eusden Jr J.D., Krzanowski J.E., Domingo D.S., St/impfli D., Martin J.R., and Erickson P.M. (1995) Comprehensive approach toward understanding element speciation and leaching behavior in municipal solid waste incineration electrostatic precipitator ash. Environ. Sci. Technol. 29, 629-646. van der Sloot H.A., Comans R.N.J., and Hjelmar O. (1996) Similarities in the leaching behaviour of trace contaminants from waste, stabilized waste, construction materials and soils. The Science of the Total Environment 178, 111-126. Van der Sloot H.A. (1996) Developments in evaluating environmental impact from utilization of bulk inert wastes using laboratory leaching tests and field verification. Waste Management, 16, 65-81. Chandler A.J., Eighmy T.T., Hartl6n J., Hjelmar O., Kosson D.S., Sawell S.E., van der Sloot H.A., and Vehlow J. (1997) Municipal solid waste incinerator residues. In: Studies in Environmental Science, 67, Elsevier: Amsterdam, The Netherlands. Meima J.A. and Comans R.N.J. The leaching of trace-elements from municipal solid waste bottom ash at different stages of weathering. Submitted for publication. Meima J.A. and Comans R.N.J. Application of surface complexation/precipitation modelling to contaminant leaching from weathered MSWI bottom ash. Submitted for publication. Meima J.A., van Zomeren A., and Comans R.N.J. The complexation of Cu with dissolved organic carbon in leachates from municipal solid waste incinerator bottom ash; determination of conditional stability constants. (manuscript in preparation) Theis T.L. and Gardner K.H. (1990) Environmental assessment of ash disposal. Crit. Rev. Environ. Control. 20, 21-42. Zevenbergen C., Vander Wood T., Bradley J.P., Van der Broekck P.F.C.W., Orbons A.J. and Van Reeuwijk L.P. (1994) Morphological and chemical properties of MSWI bottom ash with respect to the glassy constituents. Hazard. Waste Hazard. Mater. 11,371-382. Eusden Jr. J.D., Holland E.A., and Eighmy T.T. (1994) Petrology, bulk mineralogy, and melt structure of MSW bottom ash from the WASTE program. In Proceedings of the 16th annual Canadian Waste Management Conference, Calgary. Comans R.N.J. and Meima J.A. (1994) Modelling Ca-solubility in MSWI bottom ash leachates. In Environmental Aspects of Construction with Waste Materials (Eds J.J.J.M. Goumans, H.A. van der Sloot, and Th. G. Aalbers), pp. 103-110. Elsevier Science B.V., Amsterdam. Johnson C.A., Brandenberger S., and Baccini P. (1995) Acid neutralizing capacity of municipal waste incinerator bottom ash. Environ. Sci. Technol. 29, 142-147. Pfrang-Stotz G. and Schneider J. (1995) Comparative studies of waste incineration bottom ashes from various grate and firing systems, conducted with respect to mineralogical and geochemical methods of examination. Waste Management & Research 13,273-292. Johnson C.A., Kersten M., Ziegler F., and Moor H.C. (1996) Leaching behaviour and solubilitycontrolling solid phases of heavy metals in municipal solid waste incinerator ash. Waste Management 16, 129-134. Zevenbergen C., Bradley J.P., Van der Wood T., Brown R.S., Van Reeuwijk L.P., and Schuiling R.D. (1994) Natural weathering of MSWI bottom ash in a disposal environment. Microbeam analysis 3, 125135. Belevi H., St~impfli D.M., and Baccini P. (1992) Chemical behaviour of municipal solid waste incinerator bottom ash in monofills. Waste Management & Research 10, 153-167. Stegemann J.A., Schneider J., Baetz B.W., and Murphy K.L. (1995) Lysimeter washing of MSW incinerator bottom ash. Waste Management & Research 13, 149-165. Kosson D.S., van der Sloot H.A., and Eighmy T.T. (1996) An approach for estimation of contaminant release during utilization and disposal of municipal waste combustion residues. J. of Hazard. Mater. 47, 43-75. Stumm W. and Morgan J.J. (1981) Aquatic Chemistry (2nd edn). John Wiley, New York. Theis T.L. and Wirth J.L. (1977) Sorptive behavior of trace metals on fly ash in aqueous systems. Environ. Sci. Technol. 11, 1096-1100.
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DiPietro J.V., Collins M.R., Guay M., and Eighmy T.T. (1989) Evaluation of pH and oxidation-reduction potential on leachability of municipal solid waste incinerator residues. In International Conference on Municipal Waste Combustion, pp. 2B.21-43. U.S. EPA and Environment Canada, Hollywood. FL. Buffle J. (1988) Complexation reactions in aquatic systems: an analitical approach; Ellis Horwood Series in Analitical Chemistry; Ellis Horwood: Chichester. Dzombak D. and Morel F. (1992) Modeling the leaching of metals from hazardous waste incineration ash. Proceedings of the Incineration conference; Albuquerque, New Mexico. Kersten M., Moor C., and Johnson C.A. (1995) Emissionspotential einer milllverbrennungsschlackenmonodeponie ftir schwermetalle. Mtill and Abfall, 11,748-758. Van der Hoek E.E. and Comans R.N.J. (1996) Modeling arsenic and selenium leaching from acidic fly ash by sorption on iron (hydr)oxide in the fly ash matrix. Environ. Sci. Technol. 30, 517-523. Dzombak D.A. and Morel F.M.M. (1990) Surface Complexation Modeling: Hydrous Ferric Oxide. John Wiley & Sons, New York. Hjelmar O. (1989) Characterization of leachate from landfilled MSWI ash. In International Conference on Municipal Waste Combustion; U.S. EPA and Environment Canada: Hollywood, FL, pp 3B. 1-19. Van der Sloot H.A., Hoede D., and Comans R.N.J. (1994) The influence of reducing properties on leaching of elements from waste materials and construction materials. In Environmental Aspects of Construction with Waste Materials (Eds J.J.J.M. Goumans, H.A. van der Sloot, and Th. G. Aalbers), pp. 483-490. Elsevier Science B.V., Amsterdam. F/~llman A-M and Hartl6n J. (1994) Leaching of slags and ashes - controlling factors in field experiments versus in laboratory tests. In Environmental Aspects of Construction with Waste Materials (Eds J.J.J.M. Goumans, H.A. van der Sloot, and Th. G. Aalbers), pp. 39-54. Elsevier Science B.V., Amsterdam. Comans R.N.J., Meima J.A., and Geelhoed P.A. (1997) Development of a technology to reduce the leaching of contaminants from MSWI bottom ash by the addition of sorbing components. Presentation at WASCON 1997 and submitted for publication.
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Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
459
HEAVY METAL BINDING MECHANISMS IN CEMENT-BASED WASTE MATERIALS
Christian Ludwig, Felix Ziegler and C. Annette Johnson Swiss FederalInstituteof EnvironmentalScience and Technology(EAWAG), CH-8600 DObendorf,Switzerland ABSTRACT: Field and laboratory experiments were carried out to elucidate the geochemical and hydrological mechanisms that are important to understand the binding mechanisms of heavy metals in landfills with cement based waste materials. The focus of the work was on Zn(II), firstly in sorption experiments with calcium-silicate-hydrate, and secondly as a component in the leachate from a field lysimeter experiment. The leachate of the lysimeter containing cemented electrofilter ashes was sampled during rain events in order to determine the leaching processes. It was found that most of the rainwater was in intimate contact with the waste material in this field system and that while hydrological factors caused changes in concentrations of up to 100% (8-161xM), the concentration range was controlled by geochemical factors. The residence time of the water was sufficiently long to be able to describe Zn concentrations by thermodynamic calculations. The pH values in the leachate ranged between 12.5 and 13.1 where Zn2SiO4(s) appeared to be the most stable phase. Comparison with laboratory experiments suggested that alternative mechanisms could be important. In the laboratory experiments Zn appeared to be incorporated into the calcium-silicatehydrate frame forming solid solutions that have varying solubilities depending on the Ca/Zn ratio in the Cal.x-Znx-silicatehydrates. The field data agreed well with this alternative model. Thermodynamic and kinetic factors are discussed and compared with respect to the geochemical and hydrological contributions.
INTRODUCTION Today, over 80% of Switzerland's municipal waste is incinerated. The heavy metal-rich flue gases produced during incineration are scrubbed with electrostatic precipitators followed by aqueous washing treatments to remove acidic gases and potentially harmful heavy metals including mercury. The resulting electroprecipitator ash and the solid residues of the aqueous treatment are mixed. This filter ash (FA) is predominantly inorganic and is rich in heavy metals ~, and therefore, has to be handled as hazardous waste 2. Though altemative processes designed to quantitatively separate heavy metals waste residues are in development 3'4, landfilling is at present a common practice. In Switzerland, FA is mixed with cement before it is disposed of in mono landfills. Laboratory tests have shown a reduction in the heavy metal concentration in the leachate of cement-stabilized FA by factors of 5 to 205. The mechanisms which reduce the mobility are still a matter of debate. Thus, the long-term leachability of the cemented FA cannot be predicted. In order to make such predictions, it is necessary to understand the geochemistry and hydrology of these systems. The hydrology of a landfill determines whether and how long water can come into contact with the solid material. The geochemical processes determine the reactions between the solid phase and the leachate. Depending on the composition of the landfill and the residence time of the water in contact with the solid phases, concentration of elements are kinetically or thermodynamically controlled. In recent work, Johnson et al. 6 and Kersten 7 have shown the usefulness of thermodynamic calculations for the interpretation of the heavy metal concentrations in the leachates of a landfill with municipal solid waste incinerator ash. However, kinetic factors play an important role for slow geochemical processes that cannot be estimated by thermodynamic calculations. For kinetically-controlled geochemical processes, the product concentrations may scale with the reactivity of the different reactants. Here, the known ligand exchange rates around hydrated metals can be used to estimate the reactivities, e.g. for dissolution 8'9 or adsorptionl~ processes. For our field studies, we have chosen a pilot landfill that was built for scientific research purposes 5'11. The advantage of this site is that the landfill is completely filled and that CO2 contamination of the leachate can be avoided. Of special interest was the investigation of the effect of preferential flow during rain events upon the concentrations of the dissolved cations and anions in the leachates. Our field studies were accompanied by laboratory experiments and theoretical investigations about the geochemical reactions of importance. In this paper we have chosen Zn(II) as an example for a trace
460 element. Cycling of Zn is of major interest due to its high concentration in the FA. Zn is also a suitable element for laboratory experiments because it is highly soluble and allows experiments above the detection limits of common methods of analysis. EXPERIMENTAL Materials. CaCO3 (p.a.), ZnC12 (p.a.), KC1 (p.a.), CaC12 (s.p.), NaC1 (s.p.), NaOH and HC1 titrisol at various concentrations, Si standard solution (1000ppm SiCI4 in 5M NaOH) and concentrated HNO3 (s.p.) were obtained from Merck. SiO2 (Aerosil 300) was purchased from Degussa. All solutions were prepared from 17Mg2 ultrapure water (Barnstead Nanopur) which was filtered through a 0.2 ~m in-line filter. For the laboratory experiments the ultrapure water was boiled under Ar. HDPE-flasks for the field sampling and for the sorption experiments were leached with acid (--0.6M diluted from concentrated HNO3). The pilot landfill. The lysimeter is located next to the old landfill "Teuftal" in Mfihleberg (Kanton Bern) and was constructed TM for scientific research. The landfill contains cemented FA in form of cubic blocks that have an edge length of 0.5m. The plant is approximately 1.5m deep, has a surface area of 16m 2, and is covered with clay-silt (0.2m), gravel (0.8m), and humus (0.3m) layers. Additional installations at the existing sampling station were made to prevent the samples from CO2 contamination and to prevent the drainage solutions from blocking the tubing. Sampling and field measurements. Only under wet conditions did we find enough discharge for sampling. A measuring cell with a rotating stirrer and a cell volume of 0.04dm 3 was connected to the drainage outlet to measure temperature, conductivity and pH with a testo 252 field equipment. It was not possible to perform on-line pH measurements because pH-electrodes become unstable in basic solutions. The pH-electrodes were calibrated using Merck titrisol buffer solutions (7, 10, and 13). The conductivity cell was checked with KC1 solutions. The conductivities and pH values were corrected for the temperature at 25~ The temperature corrections for the conductivity measurements were based on leachate samples. Concentrated samples were diluted and the conductivities were measured at different temperatures between 6 and 27~ The established framework was used to interpolate. HDPE-flasks (0.25dm 3) were filled with sample, sealed thightly, and stored for further investigations at about 1012~ which is close to the sampling temperature, to reduce the possibility of precipitation. Sample analysis. A1, K, Na, Si, and Zn were measured with ICP-OES (Spectro, Spectroflame). The samples were diluted with acid by a factor of 5 to give a pH value of about 2. Diluted samples were analyzed for SO42- and C1 using an IC (Sykam) equipment with a Sykam (A04) column. CO32- was measured without pretreating the sample solutions using a TOC (Shimazu 5050) analyzer. The samples were sealed before the measurements to minimize CO2 contamination. Laboratory experiments. The experiments were carried out in a glove box under Ar. Calciumsilicate-hydrates (C-S-H) with a Ca/Si ratio of about 1 was synthesized after Atkins et. al ~2 by mixing 12.11 g of CaO (prepared by heating CaCO3 at 900~ for 24 hours) with 12.89g SiO2 and suspended in 0.5dm 3 water in a l dm 3 HDPE bottle. This suspension was shaken for 7 days on a rotary shaker (Btihler, Swip SM 25) at 150rpm. The suspension was then centrifuged (10 minutes, 6000rpm). The separated solid was vacuum-dried. X-ray powder diffraction spectras of the product were in agreement with the spectra of C-S-H as obtained by Taylor 13. Sorption experiments were performed in presaturated solutions (S1) with respect to C-S-H with a composition of [Si(IV)]=0.1mM, [Ca(II)]=3mM, [OH-]=8.3mM, [C1-]=0.1M, [Na+]=0.1M resulting in a pH value of 11.7. For the Zn experiments a stock suspension ($2) was prepared by adding 1g C-S-H to 0.5dm 3 of the presaturated solution and equilibrated for 7 days on the rotary shaker at 150rpm. Then, 1cm 3 of $2 was added to 50cm 3 S 1 and was equilibrated, again for 7 days. Afterwards aliquots of a Zn stock solution was added to obtain final Zn concentrations of 0.96, 0.48, 0.19, 0.096, 0.048, 0.019, and 0.0048mM. The suspensions were equilibrated for 4, 28, 53 or 87 days. In a withdrawn sample the pH value was measured using a combined glass electrode (Metrohm 6.0202.100). The remaining sample
461
was filtered (0.451am nylon, Whatman) and acidified with 0.3cm 3 of concentrated HNO3. Zn concentrations below 0.1ppm were measured by anodic stripping voltammetry (DP-ASV, Metrohm VA-Stand 694, VA-Processor 693) and above this value with AAS (Perkin-Elmer 5000).
THEORETICAL ASPECTS Heavy metal solubility is controlled by a spectrum of very slow to fast geochemical processes. Most important are sorption, dissolution, and precipitation reactions at mineral surfaces and diffusion and transformation reactions in the solid phases. Reactions in solution are generally fast. The different processes can be classified in order of their rates as follows: SOLID transformation ~ diffusion
<
SOLID SURFACE desorption ~ dissolution < adsorption ~ precipitation < diffusion
<
SOLUTION reaction in solution
The importance of understanding whether a particular reaction is slow or fast is two-fold. For relatively slow reactions, hierarchy of importance is established depending upon the rates. For relatively fast reactions, equilibrium is attained. Thus, geochemical reactions are assigned to two different categories of processes: 1) thermodynamically and 2) kinetically controlled reactions. Kinetic Control of Geochemical Processes. Investigations of kinetic processes have concentrated on the solid/water interface because of their importance in environmental systems. Mineral dissolution, sorption and complexation processes in solution have been systematically investigated. Although the many factors involved make predictions difficult, rates of adsorption l~ of cations onto mineral surfaces and the rates of dissolution 8'9 of these surfaces can be correlated with rates of ligand exchange in homogeneous solutions using linear free energy relationship (LFER). This has been discussed by Casey and Westrich 8, Ludwig and Casey TM, Ludwig et al. 9'15, and Casey and Ludwig 16 and is illustrated in Fig. 1. This illustration shows a cation (M) sorbed at the mineral/water interface which is dissolving. The surface is acting as a surface ligand that is replaced by water molecules. The rate-controlling reactions in desorption may be analogous to those of dissolution, and crystal growth to those of adsorption. The first order rate coefficients 17 for water exchange around a dissolved cation ranges from ~ 10-8 s-1 for Rh3+(aq) to ~ 101~ s-1 for Cs+(aq). The reactivity of a metal centre changes by the nature of the complexing ligands 9 (or by hydrolysis) TM (Fig. 1B) or by the oxidation state ~7. The determined relationships can be used to rank the reactivities among different species. The reactivities for the different cations of the same oxidation state increase as Pt << Ru << V <
(2)
for the bivalent cations and as Rh << Ru < Cr << A1 << V < Ga < Fe << In << U << La
(3)
for trivalent cations. The alkali metals are known to be highly reactive. Thermodynamic Control of Geochemical Processes. At thermodynamic equilibrium, steady-state conditions are assumed; i.e., the rate of formation and decomposition of a particular species are equal. Model concepts, relationships between species, and the corresponding equilibrium constants are used to describe a system. In Fig. 5 the sorption isotherms for a metal ion at a mineral surface for some different cases after Stumm TM are illustrated.
462
A
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Fig. 1 A) Simplified reaction scheme for one possible rate controlling process during dissolution or adsorption. The process consists of more than one reaction step. M indicates a cation bound to surface hydroxyl-groups. Incoming waters replace the surface ligand at the metal center. B) Analogue scheme showing a complexing ligand that influences the reactivity of the metal center.
Heavy metals that are highly reactive according to the series shown in relationship (2) and (3) (e.g., Zn(II), Cd(II), Cu(II), Pb(II)) dissolve or desorb quickly into an undersaturated solution. These species can thus be potentially found in high concentrations in leachates. The amount present in the solid phase must be taken into account. Depending on the residence time of the water in the cement pores, the highly reactive elements reach solubility limits that are controlled by different mechanisms (Fig. 5). In ideal cases, minerals such as simple mineral oxides, hydroxides and the orthosilicates dissolve congruently. For most silicates, this is not the case 19'2~ as the reactive species leave the mineral structure only in the top layers. Further dissolution is then controlled by diffusion through the inert part of the mineral structure. If this process is very slow or not possible, the reaction is controlled by the reactivity of the silica frame work and not by the solubilities. Further restrictions to a simple model are posed by solid solution phases (e.g., Cal_xZnxSiO4) that have different solubilities depending on the value x. The difficulty with solid solution interpretations in complex systems, is that it is not always known which species are involved, and also the behavior is very often not ideal. The predictions are therefore difficult. Hydrological Control of Geochemical Processes. Leaching processes involving geochemical reactions depend on the solid/solution interactions. It is therefore of prime importance to understand the hydrological processes within a landfill. There are a number of indicators for preferential flow in landfill systems and because this water plays a minor role in leaching processes, it is important to assess the
463
amount of water in intimate contact with the waste material, and the average residence time of water in the landfill. In a simple model, we assume that there are preferential flow-paths. The water that has relatively short contact with the solid material is referred to as "fresh" rain water. In contrast the "old" landfill water preequilibrates in the cement pores. The partial discharge of "old" landfill water (Qold) and of "fresh" rain water (Qfresh) is then given by the equations21: Qold - Qtotal-C/Cstart - Qstart
(4)
Qfersh
(5)
and -
-
Qtotal.(Cstan-C)/Cstart
where c, Cstan, Qtotal, and Qstart are respectively, the electric conductivity, the electric conductivity before the rain event, the total discharge, and the total discharge before the rain event. The sum of Qstart, Qold, and Qfres, gives the total discharge Qtotal. RESULTS AND D I S C U S S I O N S The Zn concentrations in the leachate were found to be between 8 and 161LtM after the rain event from June 21-23, 1996 (Fig. 2). The values are comparable to rain events sampled in the same season, irrespective of the discharge rates that range between 0.55 and 30.2dm3/h. The dilution during a rain event appears to reduce concentrations by only 40 to 60% in all of the measured cases. In order to understand the effect of hydrology, it is necessary to study some additional parameters. The electric conductivity measured as a sum parameter shows reduction in salinity as the discharge increases. However, the change is not as large as one expects from the dilution. It becomes obvious when the conductivity starts to increase parallel to the discharge, before conductivity and discharge finally return to their starting values (Fig. 3B). Preferential flow calculations as defined by changes in conductivity (4, 5) indicate that only a small amount of water passes through the landfill without reacting with the cemented FA (Fig. 3C). It must be noted that the conductivities of the preferential flow water are assumed to be similar to the water from the reference field that is negligible in comparison to the landfill water. There is a direct correlation between Zn concentration and conductivity (Fig. 2). F r o m this, we 0018-
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Fig. 2 The variation of Zn concentration in the leachate after a rain event. The symbols indicate the measurements before (o) and after (o) reaching the discharge maximum. The conductivity was used as a sum parameter that correlateswell with the Zn, and most of the other cation and anion concentrations.
464
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Fig. 3 A) rain and rain sum B) electric conductivity corrected for 25~ and C) discharge separated into "fresh" rain water and "old" landfill water using the conductivity as a tracer assuming equilibrium between the solution in the pores and the solid phases before the rain. The dashed line indicates the discharge (Qstart) before and after the rain event.
465
can obtain a value of concentration of Zn which is most probably in quasi equilibrium with our waste material. The processes that lead to this concentration must be related to the residence time of the water in the landfill. The residence time of the landfill water estimated from the total landfill volume (24m3), the total pore volume (55%, 13.2m3) 5, the amount rainwater estimated to pass through the landfill (25%) 5 and an annual average precipitation of 1050mm is estimated to be on average 3 years. However, since there are large variations in flow, the residence time during strong rain events can be as low as a few weeks. According to the relationship (2), Zn(II) is very reactive. In agreement with the long residence time for pore water, we expect that Zn in the leachates is controlled by the solubility of solid Zn phases or by the adsorption/desorption equilibria at the surfaces. The Zn content in FA ranges between 12 and 52g/kg 22. From dissolution experiments 9 carried out at low pH-values, Ca2SiO4 was found to be more reactive than Zn2SiO4, but compared with other orthosilicates show high and similar reactivity. Westrich et al. 23 have studied the reactivities for Cal_xMgxSiO4 solid solutions for x-0.5 and found that they were between those of the endmembers Ca2SiOn(s) and Mg2SiOn(s). From this point of view, we may expect that Cal_xZnxSiO4 solid solutions may equilibrate comparably fast as their endmembers. The Zn concentrations in the field leachate are presented in Fig. 4 as a function of pH. The lines represent the solubility limits for ZnO(s) and Zn2SiOn(s). On the logarithmic scale, the secondary role of the hydrology become obvious. The changes of the Zn concentration during a rain event are much smaller than those expected from a change in the pH value. ZnO(s) and Zn2SiOn(s) have been chosen because they are known to be the most insoluble phases for the systems under investigation.
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7
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Fig. 4 The solubility curves of ZnO(s) and Zn2SiO4(s) were calculated as a function of the pH values and compared with the field (O) and laboratory (r-l) experiments. The calculations considered all important species using Zn(II), Si(IV), OH-, CI, 8042- and CO32- as components. The deviations of the dashed from the solid lines indicate the influence of the present CI-, SO42 and CO32- ions in solution. The ZnzSiO4(s) curve was calculated taking a solution in equilibrium with 2.8mM Si(IV) into account, which was the typical concentration before and after a rain event. The arrow on the right shows the variation of the Zn concentration during a rain event. The arrow on the left shows the direction towards higher Zn concentrations for the isotherm shown in Fig. 5.
The measured Zn concentrations in the field leachates are more than one order of magnitude below the solubilities of Zn(OH)z(S) or the even more stable ZnO(s), but are well represented by the calculated Zn2SiO4(s) solubility curve. This indicates possible thermodynamic equilibrium with this solid phase,
466 although other mechanisms may not be excluded. The laboratory data do not appear to be controlled by a single solid phase and are probably controlled by other mechanisms. The isotherm in Fig. 5 shows that relatively high Zn concentrations adsorb on C-S-H even after 4 days. A surface site density of ---10-32mol/g was estimated from the first three adsorption measurements (o) which is in good agreement with the surface site densities measured for other silicates 24'25 and cement 26 minerals. The points shown as (,,) are in the supersaturated region with respect to ZnO(s). Nevertheless, the concentrations in solution still increase with increasing total amount of Zn in the system. The fact implies control due to solid solution formation of Ca]_x-Znx-S-H phases or surface precipitation at the C-S-H phase. Formation of Zn(OH)2(s) as a precursor of the more stable ZnO(s) may not be excluded, but this seems unlikely because the solubility of the Zn(OH)2(s) is higher than the Zn concentrations in solution. Additional experiments in the range below the saturation of ZnO(s) show that the Zn sorption increases with time (Table 1). We speculate that the effect is due to the slow diffusion of Zn into the C-S-H phase. These experiments are in good agreement with those of Kersten and J o h n s o n 27 who conducted preliminary experiments with synthesized solid solutions that have solubilities well below that of ZnO(s) and depend on the Ca/Zn ratio in the solid phases. For comparison, the amount of Zn ]~ in the cement phase of the landfill (0.23mmol/g, corrected for pore size) along with the corresponding Zn concentrations in the leachate are plotted in Fig. 5. Although, the laboratory experiments were carried out at a single pH value about one pH unit lower than the field values, a rough comparison shows that the concentrations in the leachate are in good agreement with the measured isotherm. The comparison of laboratory and field experiments shows that Zn concentrations in the leachate can be explained by the solubility of Zn2SiOn(s), or alternatively by surface nucleation or solid solution. Z n 2 S i 0 4 (s))
1 --
2 :)(s) 7tnH~_n,....,2(s)
Ji.
__
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\
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i
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-2-3-
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-4-5-
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-7
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-7
i
i
f
-6
-5
i -4
i -3
i -2
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Fig. 5 Sorption isotherm of Zn(II) to C-S-H at a pH value of 11.7. The measurements indicated as (O) were undersaturated and those indicated as ( i ) were supersaturated with respect to ZnO(s). The different lines represent schematic isotherms that can be addressed to a) adsorption only, b) adsorption and heterogeneous nucleation for low activation energies, the line indicates the solubility for ZnO(s), c) the same as b) where the line indicates the solubility for Zn(OH)2 (s), d) adsorption and surface precipitation via ideal solid solution, e) adsorption and heterogeneous nucleation of a metastable precursor, and f) the same as e) but with transformation of the precursor into the stable phase. The shown solubilities of the solid phases are only valid for the laboratory experiments. (The schematic isotherms were adapted from Stumm TM)The hatched area represents the concentrations in the lysimeter.
467
Table 1: Decrease of aeqous Zn(II) with time in two laboratol), experiments. time Zn(II)(aq) [~_M] Zn(II)(aq) [laM] [d] (Zntot= 48gM) (Zntot= 4.8gM) 4 8.7 0.39 28 6.1 0.24 53 4.1 0.19 87 3.5 0.17
CONCLUSIONS The field data show that most of the rainwater is in intimate contact with the cemented FA leading to high salt concentrations in the leachate. The residence time of the water is sufficiently long to allow us to describe Zn concentrations by thermodynamic equilibrium calculations. The field Zn concentrations can be represented by solubility control with Zn2SiOn(s). However, a comparison between field and laboratory data indicates that other mechanisms such as solid solution or surface nucleation could explain the field observations equally well. The Zn concentrations predicted by the sorption experiments at C-S-H are in good agreement with the concentrations measured in the field leachate and suggest that Zn mobility is indeed influenced by interactions with cement minerals. It appears that the field system is sufficiently reactive that sorption and dissolution processes play an important role in the control of trace element concentrations. There is, however, no p r o o f that solid diffusion and slow transformation processes do not play a role. It will be the task of future work to investigate binding and transformation mechanisms in the solid phase as well as to fred methods of discerning thermodynamic and kinetic processes.
Acknowledgments. We thank Mr. M. K~ippelifor assistance with the field experiments, Dr. A. Amirbahman for reviewing the manuscript, Dr. B. St~iubli (Amt ftir Gew~isserschutz und Wasserbau, Kanton Ztirich) for giving us the opportunity to investigate the pilot landfill, Dr. M. Ochs (BMG Umwelttechnik AG, Ztirich) who coordinated the different research projects at the landfill for helpfull discussions, Mr. H. Naef (Sieber Cassina und Partner AG, Bern) for introducing CL at the landfill, Mr. W. Imboden (EMPA, Dtibendorf) for helping to construct the sample station, Mr. W. Messerli and Mr. Schtitz (both Deponie Teuftal AG, Mtihleberg, Kanton Bern) for installing HDPE tubing and Mr. F. Nadenbousch (Werkhof, Mtihleberg, Kanton Bern) for his hospitality during the field work. CL especially acknowledges W.H. Casey for many helpful discussions about dissolution kinetics.
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7.
Baccini, P. and Brunner, P. H. Behandlung und Endlagerung von Reststoffen aus Kehrichtverbrennungsanlagen. Gas - Wasser- Abwasser 65:403-409 (1985). Swiss Federal Government. Technical Ordinance on Waste Management (=Technische Verordnung fiber Abf~ille, 1994). St/iubli, B. Rtickst~inde aus der Kehrichtverbrennung: Wie den Fluss der Stoffe lenken? Abfall-Spektrum 2:21-27 (1994). Jakob, A., Stucki, S., and Kuhn, P. Evaporation of Heavy Metals during the Heat Treatment of Municipal Solid Waste Incinerator Fly Ash. Environ. Sci. Technol. 29:2429-2436 (1995). Immobilisierung yon Rauchgasreinigungsrtickst~inden aus Kehrichtverbrennungsanlagen, Schlussbericht zum Projekt IMRA April 1991. Projektleitung Amt ftir Gew~isserschutz und Wasserbau, Kanton Ztirich (1991 ). Johnson, C.A, Kersten, M., Ziegler, F., and Moor, H.C. Leaching Behaviour and Solubility-Controlling Solid Phases of Heavy Metals in Municipal Solid Waste Incinerator Ash. Waste Management 16:129-134 (1996). Kersten, M. Aqueous Solubility Diagrams for Cementitious Waste Stablilization Systems. 1. The C-S-H Solid-Solution System. Environ. Sci. Technol. 30:2286-2293 (1996).
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22. 23. 24. 25. 26. 27.
Casey, W.H., and Westrich, H.R. Control of dissolution rates of orthosilicate minerals by divalent metaloxygen bonds. Nature 355:157-159 (1992). Ludwig, Chr., Casey, W.H., and Rock, P.A. Prediction of ligand-promoted dissolution rates from the reactivities of aqueous complexes. Nature 375:44-47 (1995). Hachiya, K., Sasaki, M., Ikeda, T., Mikami, N., and Yasunaga, T. Static and Kinetic Studies of AdsorptionDesorption of Metal Ions on a 7-A1203 Surface. 2. Kinetic Study by Means of Pressure-Jump Technique. J. Phys. Chem. 88:27-31 (1984). Ochs, M., St~iubli, B., and Wanner, H. Eine Versuchsdeponie ftir verfestigte RtickstS.nde aus der Rauchgasreinigung von Kehrichtverbrennungsanlagen. (to be submitted to MOll und Abfall). Atkins, M., Glasser, F.P., and Kindness, A. Cement Hydrate Phases: Solubility at 25~ Cem. Concr. Res. 22:241-246 (1992). Taylor, H.F.W. Cement Chemistry. Academic Press, London (1990). Ludwig, Chr., and Casey, W.H. On the mechanisms of Dissolution of Bunsenite [NiO(s)] and Other Simple Oxide Minerals. J. Colloid Interface Sci. 178:176-185 (1996). Ludwig, Chr., Devidal, J-L., and Casey, W.H. The effect of different functional groups on the ligandpromoted dissolution of NiO and other oxide minerals. Geochim. Cosmochim. Acta 60:213-224 (1996). Casey, W.H., and Ludwig, Chr. The dissolution of oxide minerals. Nature 381:506-509 (1996). Burgess, J. Metal Ions in Solution: Basic Principles of Chemical Interaction. Ellis-Norwood Limited, Chichester UK (1988). Stumm, W. The Chemistry of the Solid-Water Interface. Wiley-Interscience, New York (1992). Brantley, S.L., and Chen. Y. Chemical Weathering Rates of Pyroxenes and Amphiboles, in: Chemical Weathering Rates of Silicates Minerals, Eds. White, A.F., and Brantley, S.L. Mineralogical Society of America, Washington (1995). Nagy, K.L. Dissolution and Precipitation Kinetics of Sheet Silicates, in: Chemical Weathering Rates of Silicates Minerals, Eds. White, A.F., and Brantley, S.L. Mineralogical Society of America, Washington (1995). Johnson, C.A., Richner, G., Vitvar, Th., and Schittli, N. Hydrological and geochemical factors affecting composition in municipal solid waste incinerator bottom ash, Part I: The hydrology of Landfill Lostdorf, Switzerland. (to be submitted to J. Contam. Hydrol.). Brunner, P.H., and M6nch, H. The Flux of Metals Through Municipal Solid Waste Incinerators. Waste Management and Research 4:105-119 (1986). Westrich, H.R., Cygan, R.T., Casey, W.H., Zemitis, C., and Arnold, G.W. The Dissolution Kinetics of mixed-Cation Orthosilicate Minerals. Am. J. Sci. 293:869-893 (1993). Wogelius, R.A. and Walther, J.V. Olivine dissolution at 25~ Effects of pH, CO2, and organic acids. Geochim. Cosmochim. Acta 55:943-954 (1991 ). Zemitis, C.R. Dissolution kinetics of three beryllate minerals. MS thesis, University of California Davis, U.S.A. (1993). Schweizer, Chr.R. (unpublished data, Ph.D. thesis in preparation). Kersten, M., and C.A. Johnson. (unpublished data).
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
469
I C P - M S , H y d r i d e G e n e r a t i o n - I C P - M S , a n d C Z E for the S t u d y (Analysis and Speciation) of Solidification/Stabilisation of Industrial W a s t e c o n t a i n i n g A r s e n i c .
C. Vandecasteele, K. Van den Broeck and V. Dutr6 Department of Chemical Engineering, University of Leuven, de Croylaan 46, B-3001 Heverlee, Belgium.
ABSTRACT
Different analytical methods were applied to analyse the leachates obtained from leach tests on S/S industrial waste containing arsenic and the treated percolate water from a landfill. ICP-MS was used to determine total As and other elements, hydride generation-ICP-MS for the determination of As (III) and capillary zone electrophoresis (CZE) for As (III) and As (V). Interferences and matrix effects were studied in detail and corrected for. For the leachates, no corrections were required, for treated percolate water with lower As concentrations, in the case of ICP-MS important corrections were necessary. Results for different types of leach tests (extraction, static and semi-dynamic) and for treated percolate waters are presented.
470 I. INTRODUCTION
In this laboratory a detailed study is made of solidification/stabilisation of industrial waste containing arsenic (1) (2) (3). This includes: optimisation of the solidification/stabilisation method using extraction tests; study of the leaching of contaminants from monolithic samples using static and semidynamic leach tests; - characterisation and treatment of percolate waters from a landfill where such solidified waste is landfilled. -
-
For such a study analytical methods are required to determine accurately the total As concentration of the leachates and simultaneously the concentration of important elements such as Ca, Sb and Pb. ICP-MS (inductively coupled plasma-mass spectrometry) was used for this purpose. As in the case of As, interferences may occur due to formation of polyatomic ArCl-ions, this interference was studied in detail. Moreover, matrix effects that may be an important source of systematic errors in ICP-MS were considered. Since the leaching behaviour and the toxity of As depend strongly on the oxidation state of the As, speciation of As in solution is required. Hydride-generation ICP-MS was used for the speciation of As, as it allows, under well chosen conditions, to determine only As (HI) and no As (V). A more rapid method for speciation of As (HI) and As (V) was found in CZE (capillary zone electrophoresis). In this paper the analytical methods and the interferences will be described and results obtained will be intercompared and discussed.
2. EXPERIMENTAL
Waste material and percolate water The waste material originates from a metallurgical solidification/stabilisation according to different recipes.
process
and
was
subjected
to
The percolate water originates from a landfill where the solidified/stabilised waste is landfilled together with other waste material. This percolate water is treated in a.sewage treatment plant with FeCi3 and Ca(OH)_, and filtered to lower the metal concentration. The filtrate was also analysed. ICP-MS The ICP-MS used is a PLASMAQUAD PQ-2 Plus from VG-Elemental. It is used in standard operating conditions with a Meinhard nebuliser for sample introduction.
471
Hydride generation-ICP-MS A continuous flow hydride generator was used with a membrane gas/liquid separator to separate ASH3. The set-up is shown schematically in Fig. 1. Arsine diffuses through the silicone rubber membrane and is carried with the argon flow to the plasma. Sample uptake rate is 6 ml/min; uptake rate of the NaBH4/NaOH (15 g/l-i/0.1 M) solution is 3 ml/min. The sample solution is buffered with acetic acid/sodium acetate (0.1 M) to pH=5. Under these conditions no As (V) is reduced to As (III). CZE A Waters Quanta 4000 capillary ion analyser with on-line UV-detection at 185 nm was used for this work. A 60 cm long by 75 [.tm i.d. fused silica capillary coated with a polyimide layer on the outside was used. Before each use the capillary was conditioned for 30 min with buffer solution which containes 2 ml/100 ml of OFM (osmotic flow medifier solution) to reverse the electro-osmotic flow. The instrument was used with a negative potential applied to the injector side of the capillary, so that the anions migrate towards the anode because of their own mobility. Moreover the coating of the capillary with OFM reverses the electro-osmotic flow and thus increases the overall migration speed of the anions towards the anode. Table 1 summarises the experimental parameters. For sample injection both hydrostatic injection and electromigration were used (4). Extraction tests The DIN 38414 $4 test was used, to examine the effectiveness of the immobilization of As and other elements in S/S samples. 1 1 of distilled water is added to 100 g of dried substance (<10 mm), so that the liquid-to-solid ratio, L/S, is 10 and shaken for 24 h in a closed vessel. Static and semi-dynamic tests Static (non-agitated) leach tests were carried out on monolithic solidified waste samples. The samples (ca. 35 g) were suspended in the centre of a vessel containing 350 ml of distilled water (L/S = 10). During the time of leaching, sample aliquots are periodicially taken from the homogenised solution and analysed. Semi-dynamic leach tests, whereby the leachant is periodically replaced after intervals of static leaching, were also carried out. The samples (ca. 35 g) are suspended in 600 ml of leachant (distilled water) in a closed vessel, so that the liquid-to-surface area ratio (L/SA) is 10.
3. RESULTS AND DISCUSSION
Initial characterisation of the waste material The waste material was analysed by ICP-MS after dissolution in a boiling mixture of conc. HNO3 and HCI, filtration and dilution. Different batches were obtained and the concentrations ranged from 23 to 47 % (wt) for As, from 16 to 35 % for Sb and from 8 to 23 % for Pb.
472
A leachate of the raw waste material contained ca. 95 % of As (III), as shown by titration with iodine and with bromate. X-ray diffraction of the waste material also indicated that most of the As occured as As (ill) (As203). ICP-MS Table 2 gives experimental results for the apparent As concentration found in HCI solutions as a function of the chloride-concentration. As soon as the Cl-interference becomes significant, it must be corrected for. As the leachates contain only low concentrations of chloride, no correction is necessary. The situation is, however, different for treated percolate from the landfill, that contains much higher chloride concentrations (ca. 16 g/l giving an apparent As concentration around 800 ~g/l), so that correction (see below) is necessary. Fig. 2 shows the signal as a function of the Na2SO4 concentration for As, along with that for In. Important signal suppression starts to occur above about 200 mg/l of Na2SO4. For the leachates the total salt concentration was below 2 g/l. Therefore the samples were diluted by a factor of 10. For the percolate and filtrate samples (salt concentration 30-40 g/l) higher dilution factors were required. In all these cases using In as an internal standard after appropriate dilution was sufficient to correct for matrix effects and instrumental drift. Hydride generation ICP-MS Fig. 3 gives the % of As (III) found by HG-ICP-MS as described in solutions containing a total As concentration of 100 ~tg/l, but with varying amounts of As (ill) and As (V) as a function of the % As (V). The agreement between % As (HI) added and found is excellent, proving that only As (III) and no As (V) is measured. Fig. 4 shows that in the presence of NaC! an important signal enhancement occurs for As (ill). As verified from the mass spectrum this was not due to mass spectral interference. The reason for this enhancement is not yet clear. For a similar experiment with Na2SO4, no enhancement or suppression was noticed below I g !-t of Na2SO4. To correct for the NaCI matrix effect, especially for the treated percolate, standard addition was used. CZE The effect of NaCI, KNO3 and NaNO3 on the As (Ill) and As (V) were studied and it was shown that for concentrations below 1 g/I no significant interference effects occured. Leaching tests on S/S samples For the leaching test all As concentrations were higher than 5 mg/l, so the interference of CI was negligible. Moreover, as the salt concentrations were low, 10- fold dilution combined with the use of In as an internal standard was sufficient to correct for matrix effects.
473
Extraction tests Fig. 5 gives the As concentration in the leachate obtained for extraction tests on solidified samples prepared using addition of cement (1.1 weight unit per unit of waste), water and varying amounts of lime to the waste sample. Addition of lime lowered the As concentration in the leachate to around 5 mg/l. Static tests Fig. 6 shows the results for a static test carried out on a sample solidified according to the same recipe, but with 1.0 weight unit of lime per unit of waste. The concentrations of As and Ca in the leachate are shown as a function of the leach time, along with the pH-values of the leachate. The plot for As shows an initial increase of the concentration followed by a decrease to approximately 5 mg/1. The results can be explained by the formation of a slightly soluble CaHAsO3 precipitate. At the end of the static leach test, a sample was taken from the leachate and analysed for As (HI) by HG-ICP-MS (100-fold dilution) and for total As by ICP-MS. Ca. 87 % of the As was in the As (III) state. Semi-dynamic leach tests Fig. 7 shows the results for Ca and As for a semi-dynamic leach test, for a sample solidified according to the same recipe. The results are presented as the cumulative fractions released (CFR) over the total leach period as a function of the square root of the leach time. After an initial leaching period of ca. 12 h, the CFR plots can be fitted by a straight line, indicating that diffusion is the release mechanism. Percolate water The percolate water after water treatment (addition of FeC13 and Ca(OH)2 and filtration was analysed. Because of the low As-concentration, the interference of CI in ICP-MS can no 9 35 ~7 longer be neglected. The correction for Ar35C1 on As was based on the ratios A r - - C l / A r C1, ~5 35 35 37 Ar-C1/-C10 and Ar- CI/- C10. The latter two ratios were determined experimentally by adding chloride to the sample matrix, the former was deduced from the 35C1/37C1 ratio. Moreover standard addition was applied to correct for matrix effects. The results are shown in Table 3, indicating that correction for the Cl-interference is necessary and that the 3 correction methods give results that are in good agreement. Results for As (Hi) obtained by HG-ICP-MS are also given in Table 3. Calibration with aqueous standards gave somewhat higher results than standard addition, probably because of the previously described signal enhancement due to NaCI. Ca. 10 % of the total As is present in the As (III) form. A study of the treatment of the percolate water was also carried out. Increasing amounts of CaO were added to precipitate As. The percolate was diluted (20 ml in 50 ml) and 0, 0.15, 0.2 and 0.3 g of CaO were added. After filtration the samples were analysed using CZE, HG-ICPMS and ICP-MS. Table 4 summarises the results. The results for As(V) obtained by CZE are in good agreement with those for total As obtained by ICP-MS. The As (HI) concentrations were below the detection limit of CZE, but could be detected by HG-ICP-MS. For an amount of CaO of 0.3 g/50 ml almost all As precipitates.
474 Table 1 : Experimental parameters for CZE.
pH of buffer solution concentration of sodium phosphate buffer voltage applied injection time for hydrostatic injection electromigration injection time applied voltage
10 2.5 mM 10kV 90 s 60 s 20 kV
Table 2 9Apparent As concentrations as a function of increasing C! concentration.
CI concentration (g !-l) 0.0035 0.007 0.0175 0.035 0.07 0.175 0.35 0.7 1.75 3.5 7.0 17.5
Apparent As concentration (lag !l ) 0.29 0.33 0.995 2.61 5.66 13.9 26.1 53.9 122 231 386 836
475
Table 3 : Results for As in the treated percolate water.
ICP-MS (100-fold diluted samples) * Correction method No correction Standard addition 4~ + std. addition 35C1160 + std. addition 37C1160 + std. addition
As-concentration (Bg/l) 1050 + 315 828 213 + 10 214 + 10 224 + 10
HG-ICP-MS (10 fold diluted samples)
As (HI) concentration (~tg/l) 24.1 +_ 3.1 20.9
No correction ** Standard addition ***
* ** ***
n = 6, the results are the mean and standard deviation for 6 independent measurements (6 samples each with 3 repeats). n - 3, the results are the mean and standard deviation for 3 independent measurements. Deduced from 4 samples with additions of 0, 5. 10 and 20 Bg/l As (III), each measured three times.
Table 4" Results for As (III), As (V) and total As in percolate waters treated by adding CaO. Results in mg/1.
CaO added (g/50 ml)
0 0.15 0.2 0.3
* ** ***
CZE* CZE* hydrostatic electromigration As (V) As (III) As (V) As (III)
ICP-MS**
Hg-ICP-MS***
As total
As (III)
55.5 + 1.8 44.3 _+ 0.4 40.7 + 3.5 <7.5
55.9 42.1 38.9 0.65
<5 <5 <5 <5
on 20-fold diluted samples on 200-fold diluted samples on 100-fold diluted samples
59.6 42.4 37.8 <0.75
<0.5 <0.5 <0.5 <0.5
+ 0.7 _+0.1 _+ 0.5 + 0.27
0.440 _+ 0.067 0.249 + 0.025 0.285 + 0.003 0.0096 + 0.0012
476
NaBH 4
Argon in
\.
/
,~, PTFE tube
To ICP-MS torch Silicone rubber tube . I. Waste
rn
.....
,
Acidified sample/blank
Fig. 1
Set-up for hydride generation.
120 115 110 105
I --O-- In J
mX~As J
100
~
95
~9
90
~'
85
"~
80
70 65
6ot o
Fig. 2
200
400 600 Na2SO4 concentration (mg I1)
800
Signal suppression in ICP-MS for As and In as a function of the Na2SO4 concentration.
1000
477
USE OF HAc/Na-acetate BUFFER TO DETERMINE THE
Aslll
-"'~
% AS III added =
I
'I
I
1
!
I
I
I
I
10
20
30
40
50
60
70
80
90
~'-
100
% AsV
Fig. 3
Percentage As (III) found by HG-ICP-MS in solutions containing I00 l,tg/1 As, as a function of the percentage As (V), compared to the percentage As (III) added.
350 T
300 ~ --~ 250+ =O}
t
200+ ~
j~e~ ,0 Z
100
~
rr
I
0
0
1000
2000
3000
4000
5000
NaCl concentration (mg I1)
Fig. 4
As (III) signal (relative units) in a solution containing varying concentrations of NaC1.
% AS III found
478
10000 A
g
lOOO
i o
u
'1o
Q
0
t
t
I
I
i
t
I
t
2
4
6
8
10
12
14
16
Ca a d d e d a s C a O ( g l l O g w a s t e )
Fig. 5
As concentration in the leachate for extraction tests on S/S samples prepared with addition of different amounts of CaO.
200
1000
15
0
v
0
150
0
0
0 O0
0
CT~
8
CD
0
8
,f
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-
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-
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-10
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9
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~7
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101 time
-
%
'
leoch
Fig. 0
-11
600
V
o9 <1: O c-r O (D
12
- 8O0 V
0
~
1
102 (hours)
Rc,ult.,, of a ,ratio leach te,,,,t on an S/S sample.
o
0 105
-8
9
479
3.5
9
9
9
9
w
3+
2
0
9
2.5 Jr
9
F
0
1.5
0.5 0
0 0
0 1o0
0 200
300
400
500
600
t
i
i
700
800
900
sqrt leach time (s 1/2)
Fig. 7
Results of a semi-dynamic leach test on an S/S sample.
REFERENCES
1. Dutr6 V. and Vandecasteele C." Solidification/stabilisation of hazardous arsenic containing waste from a copper refining process, J. of Hazardous Materials, 4___00(1995) 55. 2. Dutr6 V. and Vandecasteele C." Solidification/stabilisation of arsenic-containing waste 9 leach tests and behaviour of arsenic in the leachate, Waste Management, 1__55(1995) 55. 3. Dutr6 V. and Vandecasteele C.; An evaluation of the solidification/stabilisation of industrial arsenic-containing waste using extraction and semi-dynamic leach tests, Waste Management, in press. 4. Van den Broeck K. and Vandecasteele C." Capillary electrophoresis for the speciation of arsenic, Accepted by Microchim. Acta.
This Page Intentionally Left Blank
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice
9 1997 Elsevier Science B.V. All rights reserved.
481
Application of computer modelling to predict the leaching behaviour of heavy metals from MSWI fly ash and comparison with a sequential extraction method P. Van Herck, B. Van der Bruggen, G. Vogels, C. Vandecasteele Department of Chemical Engineering, University of Leuven de Croylaan 46, B-3001 HEVERLEE, BELGIUM e-mail: Peter. VanHerck@cit kuleuven.ac.be
Abstract: Combustion residues m general and MSWI fly ashes m particular form a major environmental problem as they are polluted with heavy metals. The heavy metals can be removed from the fly ash by leaching with e.g. the acid waste water obtained m the wet scrubber of the air pollution cono'ol system. The remaining fly ash can be landfilled or valorised e.g a. a construction material. An understanding of the leaching reactions and of the factors that injTuence leaching is very important for the treatment of fly ash. Therefore the compositionof the fly ash was determined and the it~uence of pH was examined. Also a sequential extraction procedure was performed on the fly ash. hi this paper, the results of experimental leaching tests of fly ash are compared with computer calculations of the thermodynamic equifibrium of the leaching solution - f l y ash system. The computer program MINTEQA2, usedJbr this purpose, allows to predict the metal concentrations in the leaching sohaion, the minerals that precipitate, and the pH of the leaching sohttion at equifibrium. ComparL~on of the experimental and calculated leaching data with the results from the sequential extraction procedure allowed to verify the accuracy of the sequential extraction procedure.
1. I n t r o d u c t i o n Yearly around 2.8 million tonnes of municipal waste is generated in Flanders, Belgium. From this total amount of MSWI 30% is selectively collected; 43% of the rest fraction is processed in incinerators and 57% is landfilled (1). Incineration of one tonne of municipal waste leads to the formation of 10 to 50 kg of fly ash depending on the type of incinerator. The Combustion residues in general, and fly ash in particular, form a major environmental problem. Due to the fact that the legal standards for the emission of contaminants are getting more stringent the air pollution control system of incinerators must be improved, resulting in an increase of the amount of residues. The fly ash is contaminated with heavy metals and PCDD/PCDF. It must be considered as hazardous according to the Flemish environmental legislation (2). Generally, dust particles are removed from the incinerator flue gas by means of an electrostatic precipitator. Wet scrubbers remove in a first stage HCI and HF from the flue gas and in a second stage SO2. The first stage produces an acid waste water mainly containing HC1 and to a lesser extent HF. This acid solution can be used to leach the fly ash, in order to remove part of the heavy metals, as in the 3R process developed in the Karlsmhe Nuclear Research Centre (3). Part of the research in this paper is along similar lines as the work of Comans et al. (6) and Eighmy et al. (7). In the latter work, the emphasis was, however, on the spectroscopic characterisation of fly ashes, whereas one of the main subjects of the present paper concerns computer predictions and
482 calculations that will be compared with experimental results in order to gain a better understanding of the leaching process. Theoretical calculations were performed using the simulation program MINTEQA2. These results were also compared with the results of a sequential extraction method in order to test the accuracy of this method.
2. M a t e r i a l s and m e t h o d s 2.1 Definitions The leachability of the fly ash can be influenced by several parameters. Two of these parameters are acid dose (mol/kg) and L/S-ratio (l/kg). The parameter AD, the acid dose, is defined as: AD =
amount o f H + - ions added mass q / fly ash
mol (--~g )
The L/S-ratio is defined as: L /S =
volume o f leaching solution mass o f fly ash
l (~) kg
2.2 Material and characterisation The fly ash was obtained from the 'Houthalen Waste Incineration Facility" (Houthalen, Belgium), a municipal solid waste facility with an annual capacity of 98 000 ton. The fly ash was collected by a classical electrofilter. The fly ash was totally dissolved and the concentrations were measured by ICP-MS (Inductively Coupled Plasma Mass Spectrometry). This method gives the total composition of the fly ash. To obtain an approximation of the leachable fraction of the fly ash, a leaching test was conducted in a highly acid environment (18 mole HCI per kg of fly ash), and the metal concentrations in the leachate were measured. The total amount of CO32 bound in the fly ash was determined by leaching the fly ash in an acid environment (1M HCl-solution), whereby the carbonate is converted into CO2. Nitrogen was blown through the leaching solution to remove the COz, which was trapped and neutralised in a 1 M NaOH solution. The carbonate concentration was then determined by titration. The phosphate concentration in the leaching solution was determined using the M o l y b d e n e blue spectrofotometric method (8). The total amount of chloride (the sum of the amount of chloride ions from the HCI leaching solution, and the amount that is leached from the fly ash) was determined by titration with AgNO3. The particle size distribution was determined with a system of sieves with variable sieve size (0, 45, 63, 100, 200, 300, 400 and 500 ~tm diameter). With a sequential extraction procedure the total amount of a metal in the fly ash is divided in different fractions. To this end the fly ash is sequentially leached with different solutions, each selective for a given fraction. The procedure is based upon that of Kirby and Rimstidt (11). Table 1 gives the different fractions with their specific leaching solutions. In each fraction a L/S-ratio of 200 l/kg is used.
483
1 2 3 4 5 6
Fraction water soluble acid soluble oxidizable easily reducible moderatelyreducible residue
Leaching solution distilled water 0.5M CH3COOH 0.1M Na4P207 0.175M (NH3)2C204 + 0.1M H2C204 0.1M Na2EDTA + 0.3M NH2OH.HCI disolution with HCIO4 and HF
Table 1: Sequential extraction procedure (11) The oxidizable fraction is the fraction of metals that is organically bound or occurs as sulfide salt. The reducible fractions consist mainly of Fe- and Mn-oxides which may contain other metals. The easily reducible fraction consists of amorphous oxides and the moderately reducible fraction mainly of crystalline oxides. The metals in the residue are encapsulated in the silicate matrix of the fly ash.
2.3 Leaching experiments The acid waste water from the wet scrubber may in practice be used as leaching solution. The solution contains mainly HC1 and to a lesser extent HF and has a pH around 0.5. During the research pure HCI solutions with different concentrations were used (4,5). The fly ash and the leaching solution were brought in contact by mechanical shaking for a period of 3 hours. Then, the final pH was measured and the solution was filtered over a glass microfibre filter. Metal concentrations in solution were measured by ICP-MS. The influence of AD was investigated by using different acid concentrations, resulting in an AD ranging from -2 mol/kg to 18 mol/kg, giving a good view of the evolution of pH and metal concentrations as a function of AD. Negative values of AD were obtained by replacing the HCl-solution with a NaOH-solution. The leaching experiments were performed with a L/S of 10 1/kg.
2.4 The computer program M I N T E Q A 2 MINTEQA2 version 3.11 (9, 10) is a geochemical model capable of calculating equilibrium aqueous speciation, adsorption, gas phase partitioning, solid phase saturation states, and precipitation-dissolution of metals. MINTEQA2 was used to determine the equilibrium concentrations in the leaching solution as a function of pH. In a second stage, the influence of certain components of the fly ash on the equilibrium pH of the solution was examined. Input concentrations for each component were based on determined values for the mass fraction of every component in the leachable fraction of the fly ash, as presented in the Results section (with L/S = 10 1/kg). Contrary to the work of Eighmy et al. (7), components were not entered as finite solids but as components in solution.
3. R e s u l t s 3.1 Characterisation of the fly ash The total composition and the composition of the leachable fraction of the fly is presented in table 2. Major constituents of the fly ash are A1, Ca, Na and K. The other metals have lower concentrations. The rest fraction (about 30 %) consists mainly of silicate compounds (7, 11,
12).
484
Component
Ag AI Ca Cd CI Co CO3 Cr Cu F Fe
g/kg fly ash leachable fraction <0.1 73.0 204.2 0.3
g/kg fly ash total dissolution <0.1 107.0 229.0 0.3 54.0
Component
Hg K Mg Mn Na Ni Pb PO4 Sn SO4 Zn
<0.1 22.1 0.6 0.8 0.6 16.9
0.1 0.5 6.3
g/kg fly ash leachable fraction <0.1 38.7 7.9 0.7 46.4 0.1 2.9
g/kg fly ash total dissolution 0.3 65.0 16.7 1.2 54.0 0.2 3.4 27.0 0.9 58.0 12.0
0.1 9.4
Table 2" Composition of the fly ash The particle size distribution indicates that the particles are mostly smaller than 200 lam in diameter. Only 7 % is larger than 200 ~m. The largest fraction is retained on the sieve corresponding to the 45-63 ~am fraction. This particle size distribution is similar to that of other fly ashes (13).
3.2 Sequential extraction Table 3 gives the results for the sequential extraction procedure for the metals Al, Ca, Cu, Zn, Cd and Pb. The results are given as the fraction of the total amount of the metal (%) found in each step. Fraction water soluble acid soluble oxidizablc easily reducible moderately reducible residue
AI 5.5 48.5 3.4 18.2 (1.6 23.9
Ca 21.4 61.7 2.2 1.3 7.8 5.6
Cd 7.2 90.2 0 0 0.7 1.8
Cu 4.7 80.2 1.7 5.4 1.0 7.1
Pb 6.3 78.8 7.2 1.9 3.2 2.7
Zn 3.7 83.2 1.3 2.9 0.2 8.6
Table 3 Results of the sequential extraction procedure: fraction (%) of tile metals in each step The leaching solution in the 'Water soluble" step reaches a pH of 9.9. Only Ca exhibits a high solubility in water. Cd, Cu, Pb and Zn have a high acid solubility. The pH in this step reaches 3.4 after 3hrs. The other steps show a low leachability for all the metals exept for AI. The fraction of AI in the '~asily reducible" step is large and a considerable amount stays in the residue. This points out that the residue consists mainly of alumino-silicates.
3.3 pH of the leaching solution The final pH of the leaching solution is higher than the original pH because of the alkalinity of the fly ash. Figure 1 shows the evolution of the final pH after extraction as a function of AD. Between acid dose 0 and 4 mol/kg, the pH decreases rapidly. Here, only small amounts of basic metal salts dissolve, so that the added acid is not neutralised. Between acid dose 4 and 12 mol/kg, the pH decreases more slowly and becomes nearly constant. The added acid is
485 consumed to neutralise the dissolving basic metal salts (buffering capacity of the fly ash). When using an AD > 12 mol/kg, the amount of basic salts is insufficient to neutralise the added acid giving a logarithmic decrease in the pH.
L 8 -
6-4-
,o
e
a 9176 ** ~
2-
.
00,
L 2
L 4
~ 6
t
t
~
,
~
8
10
12
14
16
*
*t
18
J
20
acid d o s e (mol/kg)
Figure 1" Final pH as a function of the acid dose (L/S=10 1/kg, extraction time=3hrs)
3.4 Simulation 3.4.1 Theoretical solubility
Infigure 2, the leaching diagT'am for zinc is presented. In the figure the experimental data are presented along with the calculated concentrations (mg/g fly ash) as a function of the pH of the solution a~er leaching. The leaching diagram represents the quantitative partitioning between the different phases (15). The indicated minerals are the controlling solids at the specific pH value. At pH 6, for example, about one third of the total zinc concentration remains in solution (figttre 2), whereas one third is precipitated as smithsonite and one third as ZnO.SiO2. ~
-•
.A
10 ~'
8
8
6
Smithsonite (ZnCO3) / Zn,(O/H)sCl2z
~ ~
2 0
t
----+--------~-----I--
1
2
t
3
4
5
-
t
6
7
8
9
10
11
12
13
14
,n Figure 2" Leaching diagram of zinc (,%Experimental values; m calculated) Other solubility controlling solids are Zns(OH)sCI2 (pH 6 to 8) and ZnO (pH 7 to 13). At low pH the equilibrium concentration is restricted only by the available amount of Zn in the fly ash. At higher pH values, the concentration at equilibrium is determined by solubility
486 restrictions. At pH 8, different zinc minerals occur and no more zinc remains in solution. At pH > 13, zinc solubility increases again, due to the formation of hydroxide complexes. The theoretical predictions are in excellent agreement with the experimental data. When using the sequential extraction method the results of the first two steps can be compared with these calculated results. In the water soluble step zinc shows a very low leachability at a pH of 9.9 which is reflected by the low values in the leaching diagram. At a pH of 3.4 of the acid soluble step in the sequential extraction method the leaching of Zn is already completed. Thus the water and acid soluble steps give accurately water and acid soluble fraction. The simulation results for aluminium are given in figure 3 (15). Between a pH of 4 and 13, aluminium is precipitated as diaspore (AIOOH). The leaching diagram is in very good agreement with the experimental values. However, considering the high concentration of aluminium on the fly ashes, aluminium might also exist as an alumino-silicate compound. This was proven by the sequential extraction procedure where a high residue fraction for A1 was found. The small water soluble fraction is in good agreement with the low solubility at the pH of 9.9. At pH 3.4 the leaching for AI is not yet completed. Thus the fraction of AI and Ca in the acid soluble step is too low. It is also possible that if in the sequential procedure the pH slightly shifts, different results can be obtained. A final pH in this step of 1.5 would be more appropriate. 80
70 60 *~
~A A
50
~ 3o
Dklspore (AIOOH)
~ 20 ~ 1o 0 0
1
2
3
4
5
6
7 pll
8
9
10
11
12
,
I
13
14
Figure 3" Leaching diagram of aluminium
The leaching diagram for calcium is given in./igure 4 (15).
A
2O0 ~ ' ~ ,
>,~ E=
o~ 15o
~
t
'~~
"8 lOO .~
,3
Ca-silicates
2H ,
~
Hydroxyapatite (Ca3(PO,,,)2)
1~\ A,
50
.....
0
1
-
+
I ~__~
. . . . . .
2
Calcite (CaCO,j) ---~----- .... ' ~ .
.
3
4
5
.
.
.
Dolomite(CaCO3.MgCO3)
,
6
7 pH
8
9
10
11
Figure 4: Leaching diagram of calcium
12
13
14
487 Gypsum was found to be a controlling solid over the whole pH range; at high pH values hydroxyapatite, calcite and dolomite appear as controlling solids. The higher soluble fraction found for Ca compared to other metals can be explained by the observed leachability at pH 9.9. At pH 3.4 the leaching for Ca is not yet completed. Thus the fraction Ca in the acid soluble step is too low. And again a final pH of 1.5 in this step would be more appropriate. The leaching diagram for lead is given in figure 5 (15). Only one controlling solid occurs: chloropyromorphite at a pH lower than 9, and Pb(OH)2 at high pH values. When the fly ash is leached with water the fly ash shows a low leachability for lead as can be seen in the leaching diagram. The water soluble step in the sequential extraction method gives also a low leachability. At pH 3.4 the leachability according to this diagram is 0 %. According to the acid soluble fraction in the sequential extraction method it should be almost 100%. This difference can be explained by the absence of a high Cl-concentration in the acid soluble step where CH3COOH is used. Thus chloropyromorphite can not be formed and the leachability is higher. 3T.
A
2,5 =
2
1,5
Chloropyromorph (P~(P04)3Cl)
~'~-" 0,51 "~ a. 0
A
0
~ , 2
1
b(OH)2I
3
~A& I &.~l$ 4 5 6
t 7 pH
i 8
9
10
, 12
11
13
14
Figure 5 Leaching diagram of lead Theoretical calculations and experimental determination of the cadmium equilibrium are in good agreement Oqgure 6) (15). Under very basic conditions, Cd(OH)2 occurs. At a pH of 3.4 of the acid soluble step in the sequential extraction method the leaching of Cd is already completed. Thus the acid soluble step gives an exact view of the acid soluble fraction. The water soluble step gives a fraction of 7.2%. This step has a pH of 9.9 and at that moment the leaching of Cd already starts. 0,3
~"
_ r
0,25 0,2
/~A ,~
o,15
~
0,1
~
,if A
A
avite (CdCOj)
Cd(OHh
\
0,05 0 0
1
2
3
4
5
6
7
8
9
10
I1
Figure 6: Leaching diagram of cadmium
12
13
488
3.4.2 Equilibrium pH It is clear that the final pH of the leaching solution is very important for the leaching of the metals and is therefor important for the sequential extraction procedure. The pH is determined
by the acid dose and the dissolution of matrix elements. Therefore modelling was also applied to predict the equilibrium pH of fly ash leached with an acid solution of given AD. When the fly ash is brought in contact with the acid solution, basic oxides will dissolve causing an increase of the pH, giving the fly ash a certain buffering capacity. The equilibrium pH aider dissolution of basic oxides was calculated by MINTEQA2 for various values of AD. It was found that the variation of pH is mainly related to the AD and the dissolution of aluminium and calcium (CaCO3, CaO and A 1 2 0 3 ). 200
14 .! ~
180
~
,9 9 9 9
-~
160
exp pH
- ' - - - O - - - calculated pH
140
- ....
Ca in solution
120
~"
~
AI in solution
100
~=
80
~r
f
9
60
~
40
-
r 20
0
io
m -2
0
1
2
3
4
5
6
7 A D
8
9
10
11
12
13
14
15
16
17
(mol/k g)
Figure 7: Experimental and calculated pH, calcium and aluminium concentration as a function of the AD
Figure 7 compares experimental and calculated pH values that are in good agreement. The experimentally determined concentrations of calcium and aluminium are also shown. Up to a AD of 4 mol/kg, aluminium oxides do not dissolve and CaO is the main neutralising basic oxide. At a AD higher than 10 mol/kg, calcium is nearly completely dissolved and A1203 is the main neutralising basic oxide. Once the soluble amount of aluminium is in solution, the buffering capacity of the fly ash is consumed and increasing the AD leads to a logarithmic decrease of the pH. When the sequential extraction procedure is used for a fly ash with a different A1 or Ca composition the pH of the acid soluble step changes, leading to a different result. So the composition of the fly ash has an influence on the results of the sequential extraction procedure.
4. C o n c l u s i o n The main elements of the leaching process are solubility and precipitation of heavy metals. Leaching diagrams obtained by simulation calculations are a useful means to obtain a clear picture of the precipitation equilibria for different metals. At a given equilibrium pH of the leaching solution, the leaching diagrams indicate which metals are in solution, and which minerals are precipitated. By comparison with experimental data, the simulation program has proven to give reliable results. Because the results obtained by calculation are in good agreement with experimental results, it may be assumed that the simulation of the leaching behaviour can be extended to fly ashes with different compositions, or to other leaching conditions.
489 The leaching diagrams can be used to verify the accuracy of the sequential extraction procedure. By comparing the results of the water and acid soluble step of the sequential extraction procedure with the leaching diagrams, it is clear that the acid soluble step should be performed with a more concentrated acid in order to obtain a pH of 1.5. At this pH the leaching process of all of the investigated metals is finished. Otherwise it would be possible to obtain different resutls when the pH of the leaching solution in the acid soluble step shifts slightly. Also the choice of acid can be important for evaluating the leachability of the fly ash. Lead precipitates when HCI is used, but with acetic acid lead has a high leachability. The other metals show no influence of the choice of acid. The results of the water soluble step gives a good agreement with the leaching diagrams. During the leaching process, the pH of the leaching solution increases as basic metal salts dissolve. The equilibrium pH is determined by the AD, and the amount of CaCO3, CaO and A1203 on the fly ash. Thus the composition of the fly ash can influence the results of the acid soluble step because the pH changes.
References 1. Wille D., De Boeck G., hn,entarisatie Huishoudelijke AJi,alstoffen in Vlaanderen in 1994, Productie, hlzame#ng en Vetwerking, Openbare Afvalstoffenmaatschappij voor het Vlaamse Gewest, Publicatie nr. D/1996/5024/4, April 1996. 2. Senelle R., Dujardin J., van Damme M., VLAREM II, Die Keure La Charte, Brugge, 1995. 3. Vehlow J., Brown H., Horch K., Merz A., Schneider J., Stieglitz L., Vogg H., SemiTechnical Demons#'ation of the 3R-Process, Waste Management and Research, 1990, vol. 8, 461-672. 4. Gong Y., Kirk D.W., Behaviour of mm#cipal so#d waste incinerator fly ash, Journal of Hazardous Materials, 1994, vol. 36, 249-264. 5. Van Herck P., Vandecasteele C., Wilms D., Characterisation of fly ash from municipal waste incineration and study of the leaching #1 view of metal removal, Proceedings of Solid Waste Management: Thermal Treatment and Waste-to-energy Technologies, Washington, 1995, Air & Waste Management Association, Pittsburgh, 723-728. 6. Comans, R.N.J., Meima, J.A., Model#ng Ca- sohtbi#ty in MSW1 bottom ashes leachates, Elsevier, Proceeding of the international conference on environmental implications of construction materials and technology developments, Maastricht, The Netherlands, 1-3 June 1994, 103-110. 7. Eighmy T.T., Dykstra J., Krzanowski J.E., Domingo D.S., St/~mpfli D., Martin J.R., Erickson P.M., Comprehensive approach towards understanding element speciation and leaching behaviour in municipal so#d waste incineration electrostatic precipitator ash, Environmental Science & Technology, 1995, vol. 29, 629-646. 8. Clesceri L.S., Greenberg A.E., Trussell R.R., Standard methods for the examination of water and wastewater, American Public Health Association, Washington DC, 17th Edition, 1989, part 4, 177-178. 9. Allison J.D., Brown D.S., Novo-Gradac K.J., MINTEQA2/PRODEFA2, A geochemical assessment mode/for em,ironmental ~systems: Uset"s man,al, Environmental Research Laboratory U.S. EPA, Athens, GA, 1991. 10.Felmy A.R., Girvin D.C., Jenne E.A., MINT'EQ- A computer progT"am for calculating aqueous geochemical equilibria, U.S. EPA Project 600/3-84-032, US EPA, Washington D.C., 1994.
490 11.Kirby C.S., Rimstidt J.D., Mineralogy and Surface Properties of Municipal Solid Waste Ash, Env. Sc. Technol., 1993, vol. 27, 652-660. 12.Derie R., Les cendres volantes des incinOrateurs d'ordures mdnag~res. StructureReactivitY, Chimie Nouvelle, 1993, vol. 10, nr 37, 1091-1097. 13.International Ash Working Group, An international perspective on characterisation and management of" residues from municipal solid waste incineration, Summary Report, International Ash Working Group, 1994. 14.van der Hoek, E.E., Comans, R.N.J., Speciation of As and Se during leaching of fly ash, Environmental Aspects of Construction with waste materials, Elsevier, Proceeding of the international conference on environmental implications of construction materials and technology developments, Maastricht, The Netherlands, 1-3 June 1994, 467-476. 15.Van der Bruggen, B., Vogels, G., Van Herck, P., Vandecasteele, C., Simulation of the leaching behaviour of heavy metals from municipal waste incineration fly ash, submitted for publication in Journal of Hazardous Materials.
Goumans/Senden/van der S l o o t , E d i t o r s Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
MODELS
FOR LEACHING
OF POROUS
491 MATERIALS
Pierre M O S Z K O W I C Z (1), Radu BARNA 0)(2), Florence S A N C H E Z 0), Hae Ryong BAE 0), Jacques M E H U (2) (1) LAEPSI, INSA Lyon, 20 av A Einstein, 69621 Villeurbanne, France (2) P O L D E N , INSAVALOR, BP 2132, 69603 Villeurbanne, France
Abstract The release of soluble species contained in solidified/stabilized wastes are assessed by leaching tests. Interpretation of experimental results must be supported by precise modeling of the different phenomena involved hydrodynamics, dissolution, chemical interaction, diffusive transport. The models are presented, which can apply according to the leaching scenario (with or without advection). -
solubilization shrinking core model, diffusionnal model, coupled dissolution/diffusion model.
1. Introduction The Laboratory of Environmental Analysis of Industrial Systems and Processes (LAEPSI) of INSA, Lyon and the division P O L D E N INSAVALOR have consecrated a significant part of their research over the past few years to the study of stabilized waste leaching tests. Interpretation of the results leads to the evaluation of the environmental quality of the obtained materials. Solidification using hydraulic binders gives rise to porous monolithic materials. The pollutants initially contained in the wastes are confined and may even be stabilized within the solid matrix. Certain mechanisms of solidification/stabilization are well known, whereas others are still the object of research. There are numerous factors which govern waste solidification/stabilization. and their choice can be optimized in order to obtain materials with characteristics meeting the technical and environmental specifications of the considered scenarios : good mechanical strength, good leaching behaviour, etc... The release of soluble species contained in a porous monolithic cement-based block in contact with water is the result of complex and coupled phenomena (at the block surface and within the block: -
-
water transfer in the porous medium up to saturation, dissolution of the species in the porewater according to the local chemical context,
492
-
transport of species in solution due to the effect of concentration gradients, change of species solubility in the porewater (including possible reprecipitation) if the context of the latter has undergone certain modifications, due to the pH profile for example, following the release of pH controlling species (portlandite...).
The leaching studies of porous structures in different scenarios of liquid/solid contact lead to the use of several models to describe pollutant release. Two cases can be distinguished: 1- Leaching of monoliths without advection : Several models can describe mass transfer in the porous solid: solubilization shrinking front model, diffusional model, coupled dissolution/transport model. ; 2- Leaching of granular beds with advection : in this case the hydrodynamics of the system must be taken into account by the percolation-leaching model. If the leaching imposes particular conditions of leachant flow around the solid, the mass transfer mechanisms (solubilization, diffusional transport etc...) must be integrated in the hydrodynamic model..
2. L e a c h i n g m o d e l s w i t h o u t a d v e c t i o n
2. 1. Solubilization shrinking core m o d e l This model can describe the case of the main elements of the matrix without major chemical interactions (example: Ca, Na, K,...). It considers the coupling of the two phenomena: instantaneous solubilization (up to the saturation limit of the pore water solution) of the species present in the solid phase of the porous matrix saturated with water and its diffusional transport in the pore solution (characterized by the diffusion coefficient D), without chemical coupling with another species (the common ion effect is not taken into account). The main constitutive hypotheses of the model are : 1- Initially, at t = O, the solute has a uniform concentration in the solid equal to SO(kg/m3). 2- The solution is saturated by the solute with a constant concentration Csat, as long as the aqueous phase in the porous matrix is in equilibrium with the solid phase still containing the solute. d C d2C 0 < x < X(t)
d t
3x 2
d X
x=xm = So d t i
x > X(t)
C = Csat,
S = SO
It is therefore a problem of a "shrinking front": the mobility of the dissolution front is governed by the mass balance at the front position X(t).
493
The rate at which the dissolution front shrinks within the solid is proportional to the square root of time X ( t ) = Ks/i-. Between this front and the liquid interface, transport of pollutants in the pore water takes place by diffusion. O n the other side of the front towards the core there is no mass transfer. The parameters of the model are 9Csat, So and D. The concentration at the liquid/solid interface varies according to the leaching scenario. If renewal is sufficient, this concentration can be considered as zero. .Front
C=O Liq
I
x=O
x
I
x=X(t)
Fig. 1 9The shrinking front solubilization model (one species)
2.2. Diffusional m o d e l The diffusional model is widely used to interpret leaching tests (tank leaching test N E N 7345). Strictly speaking, this model describes the transport of one species initially present, completely dissolved in the pore water. This description can be extended to the case of very soluble elements whose solubility does not change according to the variable physico-chemical leaching context and which are instantaneously and quantitatively solubilized in the pore water: Na, K, C1...) Numerous experiments, in different scenarios, have shown that a diffusional model (based on Fick's law) can correctly describe the released flux J (kgs -1 m -2) of the very soluble elements which are not constitutive elements of the solid matrix : --+
f "
=-Dax
0c dn
where D a= apparent diffusion coefficient (m 2 s-1) C = volumetric concentration, (kg m -3) Generally used assumptions for applying this equation are : isotropic porous media, in which the structure is constant with time. The solubilization is considered to be instantaneous and not mass limited. The concentration at the solid/liquid interface is zero (which is the case if water renewal is sufficient). The apparent diffusion coefficient D a (in m2/s) is assumed to remain constant in time and space, which implies, in particular, that the solid is saturated with water from the beginning of the process and that no physical or chemical alteration disturbs the diffusion phenomenon.
494 The fundamental diffusion equation becomes: Of
(~2C
~2C
0 2 C "]
d t - Da x ~ , d x 2 + d y 2 + d z 2 ) For a solid of infinite length (x e [0, oo D in contact with the liquid via a normal plane surface of direction x, the flux of the leached material can be written :
J(t) =-Da d~xl x=0 =
Co x ~ a
if C o is the initial leachable concentration. The two parameters Co and Da can be identified by two distinct experimental tests (availability test and tank leaching test). We proposed an approach based on the simultaneous identification of the two diffusional model parameters, from only one tank leaching test if the leaching time is sufficient to reach depletion of the released species in the solid core of the leached sample. In figure 2 the values of the standard deviation z of the simulated and experimental sodium concentrations in the leachate are represented, according to the varying values of D a and C 0. The optimal values of the parameters C O and D a are used in figure 3 (continuous curve) to simulate the released sodium flux J (mgs -1 m -z) in comparison with the average experimental flux (pOints).
__/-
.....
---
z
3
4
5
6
7
8
~t Fig. 2 : Optimal values of C Oand D a
Fig. 3 : Simulation of Na release
Long term simulation for any scenario involving solid/leachate contact is then theoretically possible. The application limits of the model are reached when the physical characteristics of the material itself are modified (increased porosity, destruction of the porous structure...). We have also observed a 100 to 1000 times lower "residual flux": the release continues although the defined (theoretical) diffusional flux has stopped. A possible explanation of this phenomenon could be the continuation of release of a less soluble phase after depletion of the more soluble phase.
495
For the same experiment, the comparison between the solution of the diffusional model (parameters" C o and Da), and the shrinking core model (parameters Csat, So and D) shows that"
Co=A
IoD a =SO a 2
e
' Csat
and
'I
Da =1
-
t
D
=J
Co erf 2 , f ~
Based on the available experimental results (solute mass released in the leachate), it is not possible to distinguish between the two models : as long as the solid "remains" semi-infinite, the mass released is proportional to the square root of time. In conclusion, with respect to the initial hypothesis, the two models are equivalent for the modelling of the release of specific species. However, an important difference is that the concentration profile of the species in the pore water can be calculated from the shrinking core model.
2. 3. C o u p l e d d i s s o l u t i o n / d i f f u s i o n
model
This model must be considered in the case of elements whose solubility depends on the variation of the physico-chemical context, pH in particular (example" amphoteric metals). Experiments have shown that amphoteric metals release is controlled by solubility in the pore water context. The pH evolution within the solid and especially at the solid/liquid interface is a significant parameter. Lead is a typical case of such behaviour. Pb additive release in a mortar elaborated from Portland cement (sequential leaching of identical samples of the same monolithic material containing PbO) in contact with different chemical contexts : demineralized water (W), controlled pH 5 and pH 10, alkaline water at pH 12.5 (AW) is presented in the figures below. 6000
[ ........
300 ""
5000 1W 4000
9 pH5 ApHIO
2000
:
1ooo
~jm
0
-" ~
150
i
100 _A 9 F-
50
i ~
0
m
250
200
OAW
r
3000
s
1
~ 2000
t
9
~ 4000
9
OpH5
9
1
0 o
2000 time
tim e (hours)
4000 (hours)
Fig. 4" Leaching of Pb in different leachants The release of Pb is sensitive to the chemical context of the leachant and cannot be described by the simple diffusional or shrinking core model. A coupled dissolution/diffusion model can describe the release of chemically more complex species contained in a stable porous matrix in contact with water. In the case of a porous matrix containing two leachable components: calcium hydroxide and lead hydroxide, the modelling of release can be decomposed into several stages :
- release of portlandite, described by a shrinking front model ;
496
-
calculation of the induced pH profile, assuming that local chemical equilibrium takes place in the porewater ; - determination of local lead solubility (by calculation or from specific experimental determination); - description and calculation of lead transport by diffusion in the porewater and/or at the solid/liquid interface.
Different simulations were carried out and compared to the results obtained from leaching of cement-based matrixes containing lead. The coupled dissolution/diffusion model allows representation of the leaching tests results using demineralized water and confirms the interracial character of lead release. As long as lead in solid form is present in the matrix zone near the leaching surface of the matrix, its release is controlled by a solubilization phenomenon at the solid/liquid interface. In this case, the leaching model can be simplified: diffusional transport of lead within the matrix can be neglected. A model based on the shrinking core model to describe the calcium release from which the pH evolution at the solid/liquid interface is assessed (Figure 5) and variable lead solubility according to this pH (assessed in an Acid Neutralization Procedure) allows a good representation of the phenomena (Figure 6). The case of lead is specific:the release seems to be governed (on the time scale of our laboratory leaching tests) by a lead solubilization phenomenon at the solid/liquid interface, itself governed by the pH as the relevant parameter. 12,2 12 11,8 _ 11,6 11,4 ~ 11,2 ~" 11 10,8 10,6 i 10,4
0
i l l
f// ~[[ J ! ! ~!! I I
0
!
i
r
m
/
I I
I
f
_-1
/
r
#
=
~'-3
|
...............
1000 2000 3000 4000 5000 t (h)
Figure 5" Simulation of pH evolution near the solid/liquid interface.
-5 I . -6
0
.
.
2
.
. . . . . ...................
4
6
8
.
.
10
12
14
pH
Figure 6" Experimental solubility curve according to pH
The observed discontinuities on the simulation of pH evolution at the matrix cement/solution interface result from the sequential character of the leaching test carried out (periodical renewal of the leachant). Knowledge of the pH near the matrix cement/solution interface allows, from the experimental solubility curve, evaluation of the lead concentration at saturation in this zone. The quantity of lead released can be estimated via an interfacial transfer coefficient (figure 7).
497 2000
1500 l 9 Poin~exp. '~-'Simulation
1000
500 0 0
1000
2000
3000
4000
5000
t (h)
Figure 7 : Simulation of lead release. Leaching with demineralized water
3. Leaching models with advection Leaching of granulate materials is usually carried out by percolation tests. Transport by advection must then be taken into account. The 6rst level of modelling concerns characterization of the hydrodynamic regime of liquid through the column. Dispersion must be taken into account. The general equation governing solute transport in the mobile liquid phase is as follows (one dimensional model) :
c dt
c - D
/l ~ , d2xJ _ v d x
+ R
where v is the velocity of the liquid, D the hydraulic dispersion coefficient and R a source term (corresponding to dissolution flux). The phenomenon of advection-dispersion can classically be translated by a model consisting of n identical contactors of continuous stirred open reactor type each containing a mobile phase in contact with the solid phase. The presence of stagnant zones can also be taken into account. The porous solid is therefore in contact, in each reactor, with the mobile liquid as well as the stagnant liquid. The solute is exchanged between the solid phase and the liquid phase as well as between the mobile and immobile liquid phases. In the following figure, the hydrodynamic model of percolation-leaching is presented.
Q
Cm,j-1
Vm Cmj
V.
.... d,r; o
Cmj
Cim.i .........
Figure 8 : Hydrodynamic model of percolation-leaching of porous granular material
498 The experimental use of a tracer allows identification of model parameters (n number of reactors, mobile fraction fm of the volume and transfer coefficient K (between the two liquid zones). But problems of interaction between the porous medium and the tracer (diffusive transport in the porous system, surface sorption phenomena, etc...) may complicate interpretation of experimental results. The balance of tracer in the cell j for time i'At, while flowing and in the stagnant zone (Coats and Smith model), is : i i d V i W dCi,j QCm,j_I-QCm,j = ~ - ( Cj)=fm--~ n n dt
W dCiim,j
+
f
- - ~ dt
lm n
i dCim,j
dt
- K(Ci
i ) m,j -- Cim,j
where :
Cm,i: the volumetric concentration of tracer in the mobile fraction and in the cell j, kg/m 3, Cim,j : the volumetric concentration of tracer in the stagnant zone and in the cell j, kg/m 3,
V : the volume accessible to the fluid, m 3, fm : the mobile fraction in the liquid phase, fxm : the immobile fraction in the liquid phase, I<2: the mass transfer coefficient, h -1. In the following graph the experimental and simulated results are presented for the behaviour of the lithium ion used as hydrodynamic tracer during the leaching of slags. The model takes into account physico-chemical interferences between the tracer and slags: the hypothesis used neglects diffusive transport in the porous system, only taking into account sorption phenomena. The parameters used for the simulation are : n=15, fm =0.8 and K=4h -1.
0,8
0,6
.
E 0,4
I- Ei-Li module
0,2 0
0
500
1000 1500 2000 2500 3000 min
Fig. 9 : Identification of hydraulic regime using a tracer in a column containing leached slags The model of leaching with advection must take into account the release of the solute by transfer from the porous solid matrix to the flowing liquid (source term). The models describing the solute transfer in the porous material are either the dissolution shrinking front or purely diffusional. The corresponding models have been presented above. The global model is therefore applicable to elements whose solubility does not vary according to modifications in the local physico-chemical context in the column.
499
The results for calcium release contained in the slags in a percolation column are presented :
1000q-
-I12 ~\
model [] exp
100 l
'~t 1
0,01
I
0,1
I
1
I
10 h
I
100
I
r
1000 10000
Figure 10 : Calcium release from slags by percolation leaching Three periods can be demonstrated: during about one hundred hours, the leachate contains a practically constant concentration and equal to the initial concentration corresponding to the equilibrium of the aqueous solution in prolonged contact with the slags ;
-
- in a second period, calcium concentration decreases in the leachate according to a law inversely proportional to the square root of time: calcium release is controlled by diffusional transport in the granular solid ; -
after 3000 hours, calcium concentration falls, which corresponds to depletion in the porous solid.
4. Conclusion The different models presented here can be used to describe the release of soluble species contained in the porous monolithic or granular materials during leaching. The hydrodynamic scenario must be considered to determine the limiting stage of the mass transfer process, taking into account possible evolution of the solubility of the released species. The development of leaching tests aiming to characterize pollutant retention in solidified wastes must be carried out with great care and supported by precise modelling of the different phenomena involved.
500
References A D E N O T , F., BUIL, M. Modeling of the corrosion of the cement paste by deionized water. Cement and Concrete research, 1992, vol. 22, p. 489-496 BARNA, R Etude de la diffusion des polluants dans les d~chets solidifi& par liants hydrauliques. Th~se doctorat, INSA Lyon, 1994, 210 p. BARNA, R., MOSZKOWICZ, P., VERON, J., TIRNOVEANU, M. Solubility model for the pore solution of leached concretes containing solidified wastes. Journal of Hazardous Materials,1994, vol. 37, p. 33-39. COTE, P. Contaminant leching from cement-based waste forms under acidic conditions. Ph. D. thesis, McMaster University of Cincinnati, 1991, 191 p. CRANK, J. Free and moving boundary problems. Oxford: Clarendon Press, 1988, 425 p. CRANK, J. The mathematics of diffusion. Second edition, New York: Oxford University Press, 1990, 414 p. HINSENVELD, M. A shrinking core model as a fundamental representation of leaching mechanisms in cement stabilized waste. Ph.D. thesis, University of Cincinnati, 1992. MOSZKOWICZ, P., POUSIN, J., SANCHEZ, F. Diffusion and dissolution in a reactive porous medium: Mathematical modelling and numerical simulations. Journal of Computational and Applied Mathematics, 1996, Vol. 66, p. 377-389. ROSENBROCK, H.H. An Automatic Method for finding the Greatest or Least Value of a Function. Comput. Journal, 1960, vol 3, p 175-184. SARDIN, M., SCHWEICH, D., LEU, F.J., G E N U C H T E N , M.TH.VAN. Modeling the nonequilibrium transport of linearly interacting solutes in porous media: a review. Water resources reserch, 1991, vol. 27, p. 2287-2307. SANCHEZ, F. Etude de la lixiviation de milieux poreux contenant des esp~ces solubles: Application au cas de la lixiviation des ddchets solidifi& par liants hydrauliques. Th~se doctorat, INSA Lyon, 1996, 245 p.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
501
A G E N E R A L I S E D M O D E L FOR THE A S S E S S M E N T OF L O N G - T E R M L E A C H I N G IN C O M B U S T I O N R E S I D U E LANDFILLS James N. Crawford, Ivars Neretnieks, and Luis Moreno Department of Chemical Engineering and Technology, Division of Chemical Engineering, Royal Institute of Technology, S-1 O0 44 Stockholm, Sweden Summary A generalised solid waste leaching model (GSWLM) has been developed which may be used to simulate the changes in leachate chemistry which occur in a waste landfill over long periods of time without considering detailed weathering processes. The model simulates the reactive properties of the waste and how these change in time as the pH- and redox-buffering capacities of the waste are exhausted. Reactions that control the solubility of heavy metals may be included in the reaction and transport model to ascertain the mobility of heavy metals and thus leachate pollution loads as a function of time. The model is intended to be simple and flexible enough that it can be adapted to different waste types with relative ease. A particular advantage of the model is that it can use data from rather simple experiments such as pH and redox titrations.
1.0
Introduction and Background
Pollution of groundwater by contaminants released from hazardous wastes is a serious and growing problem. The prediction of the long-term leaching behaviour of combustion residues is thus becoming an increasingly important issue as awareness grows of the future pollution risks associated with landfills of such wastes. It is not entirely clear if landfill effluent concentrations of heavy metals decrease over time or whether they may be very low for a long time and then increase dramatically when the pH- and redox-buffering capacities of the waste are exhausted. Such a sudden increase in leachate concentration is the scenario for what is commonly referred to as a chemical time bomb. Hydrogeochemical modelling may be used to provide an answer to this question as phenomena that can be observed in the laboratory or in the field are, for all practical purposes, instantaneous from a geological time perspective. This paper deals with a methodology that may be used to gain valuable insights into long-term leaching processes and ascertain whether chemical time bombs are likely to occur.
502 A waste landfill containing ash from municipal solid waste incineration (MSWI) may be thought of as a large chemical reactor. Carbon dioxide, oxygen, and acidic rainwater slowly infiltrate the landfill and react with the waste. Soluble products from these reactions are transported downstream with the water. Further down in the landfill, these soluble species may precipitate as new, secondary minerals, or they may take part in additional reactions. After a long time, the waste near the surface of the landfill will be exhausted with respect to its pH- and redox-buffering capacity, and be depleted of its readily soluble constituents. Downstream along the fluid flowpath there may still be an unreacted buffering capacity. When the leachant passes from a depleted zone to an unreacted zone, the chemistry of the leachant may alter considerably. The region separating these zones is often referred to as a reaction front. These reaction fronts move slowly through the waste as the buffering capacity is used up in an on-going process. At the reaction front, the composition of the leachant may change very suddenly. At a pH-front, for example, the pH of the leachant can change from acidic to alkaline over a short distance. Upstream of the pH-front, the buffering capacity of the waste is exhausted, whilst downstream there remains unreacted buffering capacity. This process is illustrated schematically in figure 1, below: Material with exhausted buffering capacity
/ /:
Unreacted buffering capacity
/
Inflowing water
[H'], [0,],.,) & [C03 a] //
/
//
Reaction front
Figure 1.
Schematic diagram of a reaction front
The width of the reaction front depends upon reaction rates, dispersion, and the effects of diffusion. If we assume that chemical reactions are fast and advective flow dominates, the reaction fronts may appear razor sharp. If, on the other hand, there is a lot of dispersion, the reaction fronts may be very diffuse and the chemistry of the leachant will vary more gradually from zone to zone. There may be a number of different constituents in the waste which buffer the leachant at different pH levels. Minerals that buffer at high pH values (e.g., CaO) are generally exhausted before those that buffer at lower pH values (e.g., Kaolinite). There may be a number of different pH reaction fronts along a flowpath where
503 different minerals control pH-buffering. If the leachant flow is dominated by advective transport and there is little dispersion, the pH-profile will vary in a stepwise fashion along the flowpath. Similar reaction fronts may also be seen for redox processes. The concentration of dissolved oxidant in the leachant (generally oxygen in natural systems) may thus be seen to decrease significantly downstream from a redox front. Many oxidation reactions (e.g. oxidation of residual organic or sulphide materials) generate acidity. Depending upon the amount and type of reducing material initially present, aciditygenerating redox reactions may consume the pH-buffering capacity of the waste rapidly if gaseous diffusion of oxygen is a dominating transport mechanism. If the landfill is saturated and there is very little gaseous diffusion, the solubility of oxygen in the infiltrating leachant and the flowrate of the leachant will limit the amount of acidity which is generated. The mobility of heavy metals is dependent upon a number of controlling variables that determine their solubility. The pH and the redox state of the leachant, as well as the concentration of complexing agents (both organic and inorganic) are arguably the most important of these. Sorption of metal ions to oxides and humic material is important, but this is a process that tends to retard leaching. A model that does not include sorption processes will thus give a worst case scenario for leaching (i.e. where there are no retardation processes). It has been observed that binding of metal ions to colloidal material (having essentially the same binding properties as the solid phase) may be a very important mechanism for transport of contaminants (McCarthy et al., 1989). Colloidal materials are typically suspended particles of less than 2~tm diameter. Colloids may be inorganic substances (e.g. clay particles), organic humic substances, or even bacteria. Techniques developed for geological systems can be used to simulate the geochemical evolution of a waste repository, provided that the mechanisms of leaching are understood in sufficient detail. This type of modelling is valuable as it allows us to evaluate scenarios and anticipate leaching processes that may be impossible to observe in the laboratory or in the field due to the long time-scales involved.
2.0
Methodology
Complex geochemical models require very detailed knowledge about the mineralogy and reactive properties of the waste material. Combustion residue-type wastes are notoriously difficult to characterise owing to the different origins of the waste and
504 variations in the combustion processes used to make them. Moreover, a large proportion of the waste (Yan, 1995) may consist of glass phases that have poorly defined compositional and reactive properties. Dissolution of these glass phases over time gives rise to secondary minerals in a process, which is similar to clay formation in nature (Zevenbergen, 1995). Because of the chemical complexity of combustion residues, the emphasis is upon simple methods that rely on relatively uncomplicated laboratory experiments to obtain input parameters. By examining the fluxes of only the most important reactants and buffering constituents, it is possible to construct a simple mechanistic model for the leaching system. The generalised solid waste leaching model (GSWLM) uses hypothetical minerals to describe the reactive properties of the waste (Crawford, 1996). A simple version of this model that has been used to simulate the reactive properties of MSWI reasonably closely is one that contains three pH-buffering minerals and one redox-buffering mineral. The generalised pH- and redox-buffering reactions are:
AOH (~) + H + <::> A + + H20
(1)
BOH (s) + H+ r
B+ + H20
(2)
COH (s) + H+ r
C+ + H20
(3)
CO3 -2 + 2H +
(4)
R (~) + 0 2 r
In addition to the mineral reactions, there are also three aqueous reactions accounted for in this simple model"
CO3 -2 %- H + r -2
CO3 + 2 H H20 r
+
HCO3
(5)
<=>H2CO~
(6)
H + + OH
(7)
The quantities and thermodynamic constants for the pH-buffering reactions may be obtained by fitting a numerical titration of the generalised model to an experimental pH titration curve. Figure 2 shows the experimental pH titration characteristic and the generalised pH-buffering model (optimised in the least-squares sense) for a typical bottom ash from municipal solid waste incineration (MSWI BA):
505 p H - t i t r a t i o n C h a r a c t e r i s t i c f o r MSWI B A 10.0 9.5 9.0 8.5 8.0 pH
7.5 7.0 6.5 6.0 5.5 5.0 0.001
0.000
0.002
0.003
0.004
0.005
0.006
tool H+ a d d e d / g waste
Figure 2.
Acid Neutralising Capacity of a typical MSWI BA
In figure 2, experimentally determined data (from pH-static experiments) are given as scattered points and the solid line corresponds to the simulated generalised model. The MSWI BA used in this investigation comes from a waste incineration plant located in Link6ping, Sweden. This bottom ash has been used as a "benchmark" waste for a number of projects in a collaborative effort between the Swedish Geotechnical Institute in Link6ping, and the Royal Institute of Technology in Stockholm, Sweden. As there are few experimental points and some spread in the data for this titration characteristic, there may be a number of different parameter combinations that fit the batch data almost equally well. It has been found that these sub-optimal parameter combinations exhibit roughly the same characteristics when a flow-through system is simulated. This uncertainty seems to only have a minor influence on the observed leaching behaviour. The partial pressure of oxygen in reducing wastes, sediments, or groundwater is generally very low if there is little diffusive transport (gaseous diffusion) of oxygen into the system. Under reducing conditions, the partial pressure of oxygen (Pro) may be as low as 10 .3o - 10 .7o atm. For a redox buffer mass balance, however, it makes little difference whether the equilibrium partial pressure is 10-a~ atm
or
10 -70 a t m .
The
equilibrium partial pressure of oxygen in the pore water may influence the solubility of heavy metals present in trace quantities in the waste, but it will not influence the rate of redox buffer exhaustion.
506 For all practical purposes, the oxygen concentration in the pore water may be considered as zero and the rate of redox buffer exhaustion is equal to the rate of oxygen transport into the waste. For this reason, the equilibrium constant for the redox-buffering reaction is set to give an arbitrarily low equilibrium
Po2, and we
consider only the total reducing capacity of the waste in the mass balance. It is difficult to directly measure the reducing capacity of a combustion residue waste material as would exist in a landfill-leaching environment. Oxygen is an impractical reagent to work with as it has relatively slow reaction kinetics and is a major atmospheric constituent. In a method originally proposed by Van der Sloot (1993), and further developed by Dziwniel et al. (1996), Ce(IV) is used as an oxidant. In this method, an excess of acidified Ce(IV) is added to a slurry of the waste material. After the waste has been allowed to equilibrate with the solution, the supernatant fluid is filtered and the amount of remaining Ce(IV) oxidant is determined by potentiometric titration with an Fe(II) reductant. Simple arithmetic then gives the reducing capacity of the waste as the difference between the amount of oxidant initially added to the waste, and the amount remaining after reaction (determined by titration). Ce(IV) was chosen as its oxidisting potential lies closer to that for oxygen than stronger oxidising agents such as permanganate ion (MnO4), for example. The reducing capacity of the MSWI BA was estimated to be 1.22 equivalent moles OJkg of waste when Ce(IV) was used as an oxidant. When MnO4-was used as an oxidant, the measured reducing capacity was 2.75 equivalent moles OJkg of waste. The stronger MnO4 oxidant thus gives an estimate which is 2.26 times that estimated by oxidation with Ce(IV). This result shows clearly, the importance of choosing an appropriate oxidising reagent when measuring the redox-buffering capacity of the waste material. The loss on ignition (LOI) for this waste material at 550~ is roughly 4.3% by mass. We may write the empirical formula of the organic content as n-CH20, if we assume the average composition of possible organic phases in the waste. If we also assume that the loss on ignition is due solely to pyrolysis of organics, there is a reducing capacity of approximately 1.43 equivalent moles OJkg of waste. Considering that a proportion of the organic material may be refractory (non-oxidisable) under landfilling conditions, the assumed organic content correlates well with the reducing capacity as measured by oxidation with Ce(IV). It is actually uncertain, how much of the measured reducing capacity is organic, and how much is due to oxidation of metallic Fe, or Fe(II) compounds present initially in the waste. Organic material, however, has a much greater capacity for oxidative
507 generation of acidity than the Fe compounds which are likely to exist in the waste. For this reason, assuming that the reducing capacity is largely organic gives a worst case scenario for acidity generation in the model. 3.0
Coupled Transport and Reaction Simulation Results
The generalised model describing the reactive properties of the waste was simulated using the CHEMFRONTS advective transport and reaction simulation program (B~iverman, 1993). This program was designed to simulate flow-through systems where there is negligible dispersion or diffusion. The program has the ability to simulate the kinetic dissolution of minerals, but owing to the absence of appropriate kinetic data for the waste reactions, local equilibrium has been assumed for the simulations. It was considered important that the influence of a gaseous phase in the waste be investigated as this may have a significant influence upon the leaching characteristics of the waste when diffusion is the dominating mechanism for transport of O2 and CO2. As the program cannot simulate diffusional processes directly, a method first adopted by Str6mberg et al. (1995) was employed in order to simulate a constant gaseous partial pressure in the landfill. The equilibrium constants for the reactions were chosen so that the Pco: and Po2 would correspond to the normal atmospheric partial pressures of these gases. Two cases have been investigated: a saturated landfill scenario, and a partially saturated scenario (Po2= 0.21 atm, and Pco: = 10.3.5 atm). As a case study, certain assumptions were made concerning the pH of leachant entering the system, as well as its infiltration rate. Because the local equilibrium assumption has been used in these simulations, the results can be scaled to arbitrary landfill depth and leachant flux. The leaching parameters assumed for the coupled reaction and transport simulation are summarised in table 1: Table 1.
Leaching parameters assumed for coupled reaction and transport simulation Parameter
Value
Incoming leachant pH
5.5
Leachant dissolved O2 concentration
10 mg/1
Waste depth
1m
Leachant flux
0.1 m3/m2year
In addition to the generalised mineral reactions, a number of reactions describing the solubility of lead (Pb) were included in the simulation model. The Pb solubility model
508 was kept relatively simple and the influence of organic complexation, or colloid binding has not been considered. The following aqueous species were assumed to exist for dissolved Pb: PbCO3, Pb(CO3)2-2, PbHCO3 +, PbOH +, Pb(OH) 2, Pb(OH)3-, Pb2OH +3, Pb3(OH)4+2, Pb(OH)4 2, PbSO4, Pb(SO4)2-2. The possible
solubility-
controlling minerals for Pb were presumed to most likely be: PbCO3 (Cerrusite), PbO (Litharge), PbO2 (Plattnerite), Pb(OH)2, PbSO4 (Anglesite), and PbS (Galena). The thermodynamic constants for these reactions were obtained from the MINTEQA2 database (Allison et al., 1991). Figure 3 shows the effluent pH, and the concentration of Pb and dissolved oxygen as a function of time for a saturated landfill as predicted by the 1-dimensional coupled reaction and transport model: Landfill Effluent Concentrations (Saturated
Conditions)
1.0E-03 ;, . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.0E-04
'
~
F
-7
T-Pb+2 I
10.5
I. . . . o2(aq) i
I---p.
"6 1.0E-05 E
! 100
~~ 1 0 E - 0 6
9.5 pH 1.0E-07 C O O
~"
Redox
front breakthrough
9.0
1.0E-08 1 85
1.0E-09 ,
10E-10
' 0
Figure 3.
'
,
'..
.
.
.
100000
,
200000
300000
Time (yearn)
400000
500000
9
9
8.0
600000
Leaching of Pb in a flow-through system under saturated conditions
Figure 4 shows the effluent pH, and the concentration of Pb and dissolved oxygen as a function of time for a partially saturated landfill as predicted by the 1-dimensional coupled reaction and transport model:
509 Landfill Effluent Concentrations (Partially Saturated Conditions)
1.0E-03
11.0
1.0E-04
10.5
O 1.0E-05 E .,,,,,
10.0
t,,,-
.9.,., 1.OE-O6 9.5
A
t~ 1.0E-07 [ ~ Oe"
pH
~ l ,,Q 1.0E-08
1.0E-09
I. . . .
02
1.0E-IO 0
Figure 4. Because
5000
10000
15000
20000
25000
8.0 30000
35000
Time (years)
Leachingof Pb in a flow-through system under partially saturated conditions we
have assumed local equilibrium and neglected
dispersion, the
breakthrough characteristics shown in figures 3 and 4 appear rectangular. In reality, the concentration fronts can appear much more gradually in the leachate than what these diagrams may suggest. In spite of this, however, the results should be roughly the same for the depletion time of the pH and redox-buffering capacity in the waste material. 4.0
Discussion
As may be seen from the results, the pH-buffering capacity in the partially saturated case is depleted roughly 20 times faster than in the saturated case. The equilibrium pH levels are also noticeably lower under partially saturated conditions. This is due to aqueous dissolution of atmospheric CO2, which is a source of protons in the pH-buffer mass balance. When there is a constant partial pressure of CO2 and 02 in the waste (as in the partially saturated case), the organic redox buffer has no influence on the proton mass balance as the dissolved concentration of carbonate is determined by the Pco2. When the landfill is water saturated, however, acidity generated by the redox reaction influences both the effluent pH and the rate of pH buffer depletion. This explains why the effluent pH increases slightly after the redox-buffering capacity is exhausted in figure 3. Another feature of interest, which is revealed in the coupled reaction and transport simulations, is that the pH-buffering minerals take proportionately longer to deplete at
510 lower pH levels. As may be seen from the discontinuities in the simulated pH-titration curve (figure 3), the three generalised pH-buffering minerals are present in roughly equal quantities. The breakthrough times for the pH-reaction fronts, however, differ significantly when the waste is leached in a flow-through system. The rate of pH-buffer depletion is related to the difference in the total proton concentration entering and leaving the system. At high pH levels, the dissociation of water to form hydroxyl ion (OH) is a source term in the proton mass balance and the pH-buffer is thus exhausted at a faster rate than would occur at a lower pH-buffering level. This also means that a pH-buffering mineral, which buffers at the same pH as the infiltrating leachant, will (in principle) never be exhausted. In both leaching scenarios, the concentration of dissolved Pb generally decreases over time. In the saturated case, the solubility controlling mineral for Pb was found to be Pb(OH)2 under reducing conditions and PbO2 (Plattnerite) when the redox-buffering capacity was depleted. As may be seen from figure 3, the Pb solubility is actually higher under reducing conditions when Pb(OH)2 is controlling. In the partially saturated case, the solubility-controlling mineral for Pb was found to be PbO2 over the entire leaching time. At no point was PbS2 found to be solubility controlling, even though the equilibrium P02 was roughly
10 6~
atm under reducing conditions.
The predicted Pb concentrations are higher under partially saturated conditions than for saturated conditions. This is largely due to the increased carbonate (CO3-2) complexation that occurs under partially saturated conditions. Neither chloride (C1-), nor sulphate (SO42) complexation was found to influence the Pb solubility appreciably over the pH ranges encountered in the simulations. 5.0
Conclusions
The results obtained from the coupled reaction and transport modelling seem to indicate that a chemical time bomb for Pb does not exist. Outside of the range of pH buffering, which was examined in the model (i.e. beyond the experimental pHtitration end point), it is not possible to say with absolute certainty how the Pb leaching may change. As the depletion time for pH-buffering minerals increases with decreasing pH, however, it is likely that any eventual increase in leachate Pb concentration will be very gradual. The solubility model, which was used in the simulations, may over predict the aqueous concentrations of Pb as precipitation with phosphate (PO42) has been neglected. On the other hand, transport of colloidal Pb may be a very important
511 mechanism and the total Pb pollutant loads may actually be higher than those predicted by the model, which gives only the dissolved Pb concentrations. The amount of colloidal Pb in the leachate will depend upon filtering processes in the landfill, as well as in the underlying aquifer. Notwithstanding this, however, the amount of Pb transported as colloidal material is not likely to increase dramatically after the depletion of the pH- and redox-buffering capacity of the waste. This means that the leachate Pb concentrations, which are to be observed soon after the commencement of leaching (i.e. after breakthrough of the first leachant pore volume), should decrease over time and will not increase catastrophically as would be the case in a chemical time bomb scenario. As may be appreciated from this simple example for Pb leaching, coupled reaction and transport modelling can be a useful tool for predicting changes in landfill leachate composition over geologically significant periods of time. The generalised model, although very simple, can also give insights into how the leaching of heavy metals is related to the chemistry of the leachant, as well as the eventual depletion of the pHand redox-buffering capacity of the waste material.
Acknowledgement This paper contains results from a research project aimed at the development of methods to predict the long term changes, release, and transport of heavy metals from landfills of solid wastes. The financial support and encouragement of the Swedish Waste Research Council (AFR) is gratefully acknowledged.
6.0
References
Allison, J.D., Brown, D.S., and Novo-Gradac, K.J., (1991) MINTEQA2/PRODEFA2, A geochemical assessment model for environmental systems 9Version 3.0 User's Manual. Environmental Research Laboratory, U.S. EPA, Athens, Georgia
Bfiverman, C., (1993) Development of "CHEMFRONTS", a coupled transport and geochemical program to handle reaction fronts. SKB Technical Report 93-21. Swedish Nuclear Fuel and Waste Management Company, Stockholm, Sweden. Crawford, J.N., (1996) The long term release of heavy metals from combustion residues and slags. Licentiate Treatise, Department of Chemical Engineering, Royal
Institute of Technology, Stockholm, Sweden
512 Dziwniel, T., Crawford, J.N., and Neretnieks, I., (1996) Redox properties from waste incineration. Undergraduate thesis, Department of Chemical Engineering, Royal Institute of Technology, Stockholm, Sweden. McCarthy, J.F., and Zachara, J.M., (1989) Subsurface transport of contaminants. Environmental science and Technology, Vol. 23, No. 5, pp. 496-502 Str6mberg, B., Moreno, L., and Crawford, J., (1995) Modelling of transport and initial weathering processes in a sulphidic mining waste rock heap. In Proceedings of Groundwater Quality, Remediation and Protection (GQ '95). Vol. 1 pp. 171-180, IAHS Pub. No. 225 Van der Sloot, H.A., (1993) Determination of the reducing properties and the reducing capacity of construction and waste materials, Draft version of a standard for assessing the in waste materials, Soil and Waste research Department, Netherlands Energy Research Foundation (ECN), Petten Yan, J., (1995) On leaching characteristics and dissolution kinetics of combustion residues. Licentiate Treatise, Department of Chemical Engineering, Royal Institute of Technology, Stockholm, Sweden Zevenbergen, C., Van Reeuwijk, L.P., Bradley, J.P., Keijzer, J., and Kroes, R., (1995) Leaching of heavy metals from MSW incineration bottom ash in a disposal environment. In proceedings of Sardinia '95, Fifth International Landfill Symposium. Vol. III, pp. 369-377, October 1995
Goumans/Senderdvan der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
513
INFLUENCE OF 1HE 1YPEOF CEMENT USED ON 1HE LEACHING OF CONTAMINANTS FROM SOLIDIFIEDWASTE CONTAINING ARSENIC V. DutT6 and C. Vandecasteele Katholieke Universiteit Leuven Department of Chemical Engineering W. De Croylaan 46 - 3001 Heverlee - Belgium Inlroduclion The waste material studied originates from a metallurgical process and contains large amounts of arsenic (32 wt%), as As203. Besides arsenic, the waste also contains Sb (17%) and Pb (14%). The waste was solidified with cement and pozzolanic materials to reduce the leachability of the contaminants from the waste. The solidification procedure was optimised and the concentration of arsenic in the leachate of the extraction test DIN 38 414 was lowered from ca. 5 g/I to ca. 5 mg/I. Addition of lime played a major factor in this reduction [11. The low concentration of arsenic (5 mg/I) was reached due to the formation of both CaHAsO3 and Ca(OH)2 in the leachate of the S/S waste. Formation of CaHAsO3 alone cannot lower the arsenic concentration beneath ca. 55 mg/I. It can be said that the concentration of arsenic in the leachate of the solidified waste material is highly dependent on the concentration of calcium in the leachate, and on the formation of Ca(OH)2. The concentration of calcium in the leachate is, of course, dependent on the amount of calcium a d d e d to the waste during the solidification process. In this paper, the effect of cement alone, without lime addition is discussed.
Increasing the
percentage of CaO in the amount of cement a d d e d is also likely to reduce the leachate arsenic concentration.
To investigate this, different types of cement were used in the solidification recipe.
Cement types can differ in their composition, water demand, setting time, price, etc...
In Table 1,
the composition of the cement types used is given. Experimental The waste material was solidified according to the recipes in Table 2. The amount of cement for each S/S sample is given as the ratio of binder mass to waste mass, together with the water to cement ratio [W/C). Sample 'eco' has a higher water-to-cement ratio than the other samples, because Ecobind 50 has a higher water demand. ca. 1.35 cm.
The solid waste specimens have a diameter of ca. 4.5 c m and a height of
The volume to surface area ratio {V/S) ranges from 0.41 to 0.44.
Before the
solidification product was subjected to any leach test, it was allowed to harden for 2 weeks. The DIN 38414 extraction test [2], and the N2 semi-dynamic leach test [3] were applied to the samples. In the semi-dynamic leach test, a leaching volume of 600 ml was used. Over the 3 week duration of the experiment, 21 renewals were performed. F__.xlracJiontest The results of the extraction test are presented in Table 3, where the concentration (mg/I) of Ca, As, Sb and Pb is given, together with the leachate pH. From these tests it can clearly be seen that the concentration of arsenic in the leachate of the S/S samples decreases as the pH value of the leachate and the calcium concentration increase. The concentration
of antimony hardly changes
between the different recipes.
The low lead
514 concentrations are a consequence
of the rather low pH values of the leachates.
Lead leaching
increases only very rapidly above pH 12 [4]. The relationship between As and Ca leaching, the pH value and the type of cement used can better be understood from Figure 1. The amount of Ca leached from the S/S sample, and the leachate pH can be related to the amount of CaO present in the cement used for solidification. As more Ca is present in the leachate, the pH of the leachate rises, and accordingly the concentration of As decreases.
Thus, the cement type used is more effective in the order: Cem III, Cem II, Cem I,
Ecobind 50. Nevertheless, the concentration of arsenic in the leachate, in this series of experiments, does not reach the minimum value of ca. 5 mg/I, which is obtained when lime is added to the S/S sample.
The still high values for the As concentration are a consequence of the low Ca
concentrations.
Calculating the saturation indices (Sl} for the compounds CaHAsO3 and Ca(OH}2
shows that the first compound is formed, whereas the second is in a state of undersaturation. Semi-dynamic leaching The 4 different S/S samples [Table 2} were subjected to the semi-dynamic leach test, whereby after each interval of static leaching, the leachate pH was measured. The pH increased during the first leaching intervals, but rapidly a constant value of ca. 11.5 was reached, for all the samples.
No
significant difference in leachate pH between the different S/S samples, as occured for the leachate pH values of the extraction test, was observed.
The leachates collected over the entire leach test
were analysed for Ca, As, Sb and Pb. The fractions released over the different leach periods were summed to calculate the CFR value and plotted versus the square root of the leach time. In Figure 2, the CFR plots for As are presented. From these plots it can be seen that the fraction of arsenic leached, at the end of the test, from the S/S samples increases in the following order: Cem I [3.3%} < Eco (7.0%} < Cem III [8.2%} < Cem II (10.7%}. Also the CFR values for calcium increase in the same order: Cem I [5.5%} < Eco (6.6%} < Cem III [7.6%} < Cem II {8.4%}. Also the CFR plots for Sb and Pb follow the same trend. The cumulative fractions of Sb released for the 4 S/S samples vary between 3.7 % and 5.0%; those for Pb vary between 0.05% and 0.07%. For the leachate of each static leach period, the saturation index Sl for the compound CaHAsO3 is calculated using the measured concentrations of Ca, As and the leachate pH.
During the total
duration of the semi-dynamic leach test (21 periods of static leaching} the Sl values for all of the 4 S/S samples remained negative, indicating that the leachate is in a state of undersaturation in regard to the compound CaHAsO3 and that no precipitate is formed. Thus, the elements Ca and As coming out of the S/S sample and into the solution do not affect one another. Their concentration in the leachate is the result of the ease with which they are released from the S/S sample, and thus of the integrity of the monolithic S/S sample. This explains the order of the CFR plots for the 4 samples: it is the same for all the elements (Ca, As, Sb and Pb}. The sample that releases the highest fraction of the element As (sample Cem II}, also releases the highest fraction of Ca, Sb and Pb. The same is true for the sample that releases the second, third and fourth highest fraction of the respective elements. The CFR plots for the elements Ca, Sb and Pb can all be fitted by a straight regression line. The plots for arsenic initially show a different behaviour, but after 6 intervals of static leaching, the plots can also be fitted with a straight regression.
From the slope of these straight lines, the effective diffusion
coefficients can be calculated. These are given for Ca and As in Figure 3.
515 Conclusions Samples were prepared with different types of cement. At the end of the extraction test performed on the samples, chemical equilibrium was reached in the leachate, with formation of the compound CaHAsO3. Ca(OH)2 however was not formed. With only the formation of CaHAsO3 in the leachate, the arsenic concentration cannot be lowered to the value of 5 mg/I that is reached with samples prepared with lime addition (where the Ca concentration reaches values of ca. 900 mg/I).
The
concentration of Ca in the leachate and the leachate pH can be related to the amount of CaO present in the cement type. During the semi-dynamic leach test, performed on the same set of samples, the compound CaHAsO3 is not formed in the leachate. The release of the contaminants is a consequence of the physical structure of the monolithic S/S sample. No clear relation could be found between the composition of the cement and the concentration of the contaminants in the leachate. Leachability indices can be calculated and can be compared with those from S/S samples prepared with lime. This indicates that As has a smaller leachability index (is more released} from the S/S samples prepared with only cement compared to the S/S samples prepared with lime addition, whereas Ca has a higher leachability index (is more retained]. References [1]
Dutre V. and Vandecasteele C., An evaluation of the solidification/stabilisation of industrial arsenic containing waste using extraction and semi-dynamoic leach tests, Waste Manag., in press.
[2]
DIN Deutsches Institut f~r Normung, DIN 38 414 S4, Oktober 1984.
[3]
C6te P.L., Constable T.W. and Moreira A., An evaluation of cement-based waste forms using the results of approximately two years of dynamic leaching, Nucl. Chem. Waste Manag. 7 (1987) 129-139.
[4]
Conner J.R., Chemical fixation and solidification of hazardous wastes, Van Nostrand Reinhold, New York, 1990.
Table ] Composition of cement types, major compounds (wt%) Cement type AI203 CaO Fe203 Cem III/B 42.5 HSR L 8.5 46.5 2 Cem II/A-M 32.5 R 8.6 53.3 3.6 Cem I 52.5 5 63.3 3.1 Ecobind 50 4 68 3
SiO2 28.5 24.5 19.8 20
516 Table 2 Solidification recipes for the 4 S/S samples Sample
C e m III/B 42.5 HSR L
C e m II/A-M 32.5 R
C e m I 52.5
Ecobind 50
W/C
c e m III
2.2
0
0
0
0.55
c e m II
0
2.2
0
0
0.55
ceml
0
0
2.2
0
0.55
eco
0
0
0
2.2
0.64
Table 3 Leachate pH and leachate concentration (mg/I) Sample
pH
Ca
As
Sb
Pb
c e m III
11.41
222
295
24.6
nd
c e m II
11.63
255
134
15.4
0.01
cem I
11.81
204
94.6
14.7
0.02
eco
12.03
485
18.3
14.8
0.31
500 ~
E
v
400
u m
12.2
12.0
~" 60
12.0
11.8 =
11.8 =Q,,
100
11.2
o
11.0
0
11.6 ~ .= U 11.4
0 m .C
70 }'" 50 = 40 30 .= 20 0 10
300
uGt
12.2
200
Sampie
l
I
---4--- pH.I
11.6 "~ 11.4 11.2
, Cem III
, Cem II
Cem
Cement type
11.0 Eco li
I-'-
t CaO (%) pH
Figure 1 Leachate concentration of As, Ca and leachate pH of the 4 samples (left); a m o u n t of C a O (wt%) in the c e m e n t Iypes (right)
U
517 12 10
n
,,,,,
D
I:z: i,i, (J
n
[]
o o X o x []
0 0
[]
~
0
D 0 X0 X0 X0 0X 0X 0X X0 X0
n
o
O
0
0
O
~
0
0121 ~
o o o 0 ~ 0 X
x
X
x
x
loi Cem III Cem II
x
Ceml Eco
X
x
A AA
A AA
A A
i
i
i
t
I
i
200
400
600
800
1000
1200
sqrt leach time
I
1400
(s 112)
Figure 2 CFR plot for arsenic for the 4 S/S samples
Arsenic
3.0E-10
Calcium
2.5E-10 ~"
2.0E-10
E
1.5E-10
r
1.0E-IO
~" ~
~
5.0E-11 O.OE+O0
I~ Cem II
Cem III
I
Eco
I
Ceml
8.E-10 7.E-10 6.E-10 5.E-10 3.E-10 2.E-10
1.E-IO i
O.E+O0
Cem II
t
Cem III
Figure 3 Effective diffusion coefficients De {cm2/s} for As and Ca
Eco
,/
Ceml
This Page Intentionally Left Blank
Goumans/Senderdvan der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
519
V e r i f i c a t i o n of l a b o r a t o r y - field l e a c h i n g b e h a v i o r o f coal fly a s h a n d M S W l b o t t o m a s h as a r o a d b a s e m a t e r i a l
Dr. Ir. J.P.G.M. Schreurs Intron, Institute for materials and environmental research B.V., P.O. Box 5187, 6130 PD Sittard, The Netherlands Dr. H.A. van der Sloot ECN, P.O. Box 1, 1755 PG Petten, The Netherlands Prof. Dr. Ch. Hendriks TU Delft, P.O. Box 5048, 2600 GA Delft, The Netherlands
ABSTRACT For an environmentally acceptable utilization of primary and secondary construction materials regulations have been developed in the framework of the Soil Protection Act (WBB) and the Clean Water Act (WVO). This has resulted in the Building Materials Decree. An extensive field study has been carried out to determine the relation between predictions based on laboratory leaching tests and the release as established by measurements at applications of secondary materials in roadbase constructions. The study consisted of an extensive sampling program based on previous release estimates, an estimate of actual water management, leaching of the underlying sand layer as well as calculation and modelling of release based on these profiles. For many relevant elements a quantifiable release has been established after about 10 years of field exposure and for most elements the predictions based on laboratory tests agree with the field data fairly well.
Introduction For an environmentally acceptable utilization of primary and secondary construction materials regulations have been developed in the framework of the Soil Protection Act (WBB) and the Clean Water Act (WVO). This has resulted in the Building Materials Decree [1]. Based on leaching tests in the laboratory and their translation to practical situations, a theoretical relation has been established between the desired level of protection of soil and groundwater and the results of laboratory leaching tests. Since results of laboratory leaching tests cannot be translated directly to field conditions, RIVM developed translation factors and formulas, which are described in the Guidance to the Building Materials Decree. This concerns corrections for the temperature difference between lab and field, the degree of contact with water under field conditions, and the extrapolation to longer time scales. For instance to the reference point of 100 year as used in the concept of marginal burdening of soil. For road construction applications the question is here to
520 what extent laboratory leaching data can be correlated with field observations taking into account the correction factors specified above. The CROW Coordination Committee Construction Materials and Environment, originally the Steering Committee Environmental Hygiene and Road Construction, supported in 1993 a research plan prepared by INTRON and ECN to provide a relation between predictions of release based on laboratory test data using the calculations specified in the Building Materials Decree and measured impact in applications in practice. Four characteristic situations in road construction have been selected to verify the degree of agreement between predicted emissions using the prescribed calculations and the field situation after about 10 years of field exposure. In these road base applications coal fly ash and MSWl bottom ash have been used as stabilization layer. The aim of the study was the verification of the predicted release from secondary materials with measured immission in practice in the four selected road construction applications using results from laboratory leaching tests. The following approach was followed to study the relation between the predicted immission on the basis of laboratory data and formula's specified in the Building Materials Decree with the measured impact under field conditions: - extensive sampling at the four selected road construction applications, determination of release under field conditions, -prediction of release according to the Regulation of Construction Materials, -comparison of predicted release with measured release under field conditions, -evaluation of observations. -
The determination of release in practice has been based on: I. The concentration decrease in the construction material relative to the starting situation. II. The net concentration increase in the underlying soil and collected leachate. II1. The difference in leaching between fresh and ,,field aged" construction material. The release in practice as derived from these different complementary methods is compared with the calculated release based on predominantly percolation controlled system or a predominantly diffusion controlled system. To provide a sufficient statistical basis for the comparison 10 cores were sampled for each of the four road base constructions in the study. The prediction of release was carried out in accordance with the formulas specified in the Building Materials Decree. Geochemical reaction / transport modelling has been carried out to support the interpretation of the field measurements. Figure 1 gaves a view of the approach described above. Selection of sites Four conditions in road base application were selected. In all cases, secondary materials have been applied as stabilization layer on a drainage layer of sand. As cover materials, asphalt concrete, sand and road construction bricks have been utilized. The selected applications (code) are: CA, coal fly ash cement stabilization under asphalt. CZ, coal fly ash cement stabilization under sand. at the Coloradoweg (Province of South Holland, Municipality of Rotterdam, year of construction: fall 1983). VF, Feniks under road construction bricks. MA, MSWl Bottom ash under road construction bricks.
521 and the Vondelingenweg (Municipality of Rotterdam) (year of construction: mid 1986) The coal ash cement stabilization at the Coloradoweg is a stabilization of pulverized coal ash with blast-furnace slag cement (9 %). Feniks is a stabilization of MSWl bottom ash (from Roteb, Rotterdam), with sand and blast-furnace slag cement in the following proportions, 60 %% MSWI bottom ash, 40% sand and 3 % cement The MSWI bottom ash consists of sieved, aged MSWI bottom ash from ROTEB and applied as granular material. Sampling The release of constituents from the construction materials containing secondary materials has been established by measuring the decrease in concentration in the construction material itself as well as from the concentration increase in the underlying sand layer. By taking cores through the stabilization layer well into the underlying soil using a core-catcher (PVC inter tube in the stainless steel drilling core) an undisturbed profile of construction material and adjacent soil can be obtained. The PVC cores were capped, transported and stored. In the laboratory, the cores have been sliced in the full length and subsequently sliced in segments for chemical analysis. For each core (10 per application) this has resulted in concentration profiles in the construction material and the adjacent soil. These data have been used to determine the release from the secondary construction material in the exposure period of about 10 years. For each application samples have been taken from the field exposed construction material to assess it residual leaching behavior in a column test (NEN 7343) [2] after aging/weathering. For the prediction of release, the predominant transport mechanism of constituents is a crucial property that needs to be addressed to allow a comparison of lab test data with field measurements. The leaching mechanism - percolation or diffusion - can be identified for practical applications based on water management, verification of water content in the core profiles and on transport modelling. With the exception of application CA - coal fly ash under an asphalt top cover, the leaching behavior is controlled by percolation. Under asphalt, the release is largely diffusion controlled. A study of the water management of the four applications has provided information on the liquid to solid ratio (L/S) under field conditions and information on the occurrence of interferences such as temporary contact of the stabilization layer with ground water as a result of a high water table. The calculated liquid/solid ratio based on formulas in the Building Materials Decree in agreement with the study of water management and projections from (collected) leachate production. Determination of release from field data
For the determination of release, three methods have been applied: I. I1. III.
Concentration decrease in the secondary construction material. Concentration increase in the underlying soil and/or the collected percolate. The difference in leaching between ,,fresh" and ,,aged" material.
In the figure 2 the determination of E~ and E, is shown.
522 Release E~ The determination of release (E~), which is based on the concentration decrease of the constructionmaterial, can be carried out when a proper reference can be found or when a concentration decrease is observed in a part of the construction material. For instance, at the interface of construction material and soil. In the latter case, the part of the construction material that is not or not appreciably leached can be used as reference. Based on a statistical evaluation of all cores sampled, it is possible to identify whether a statistically significant decrease relative to the reference level has occurred. Release E, The determination of release E,, which is based on the concentration increase in the underlying soil directly in contact with the construction material, can be carried out when a proper reference for the original background concentration can be found or when a concentration increase is observed in a part of the sampled profile. In the latter case, the concentration in the section further away from the construction material/sand interface can be used as reference. Based on the results of all the cores, the statistical significance of a measured increase in the underlying sand layer can be identified. The concentration measurement in samples from the sand layer are based on a mild acid extraction as the elements causing the increase in soil are surface bound. Elements that are not sorbed in the soil are transported to deeper soil layers or to ground water. From a concentration increase in collected percolate in combination with the fraction that is retained in the sand flayer, a better estimate of release for mobile elements (e.g. anions) can be given. This implies that only in case of leachate collection as applied in demonstration projects this aspect can be covered adequately. For the road base applications at the ,,Coloradoweg" and the ,,Vondelingenweg" a leachate collection system has been placed for monitoring purposes thus allowing to close the mass balance for mobile species. Release E,~ The release of constituents from the construction material can also be assessed from the difference in leaching of the fresh material and the leaching of field-exposed material. The underlying assumption is that no appreciable change in leaching behavior has occurred. For aged and weathered materials this cannot be fully excluded. This manner of determining release is further identified as E,~.
Release calculation according to the Building Materials Decree. In the Building Materials Decree a distinction is made between granular materials (N) and monolithic construction materials (V). In general, the leaching of the former will be governed by percolation, while leaching of the latter is dominated by diffusion. For both types of materials leaching test protocols and models are available (Guidance document Building Materials Decree) that allow a prediction of the release under field conditions. For the four applications predictions of
523 release have been made for the respective exposure time to determine whether measurable release can be expected in practice. For the calculation of the immission (Ep) in case of a percolating system the following formula is used [3]:
Ep = d b * h * [ E(10)-Eg ] * [ 1 - e "x'Lspractice] / [1-e -".1~
with:
db
dry density of soil kg/m 3 height of the applied secondary construction material (m) leached quantity at LS=10 in the standardized column test (mg/kg). E(10) average leached quantity for clean soil at LS=10 (mg/kg) E0 constant related to the degree of interaction of the constituent with matrix K LS liquid/solid ratio (l/kg) LSpracticeliquid/solid ratio reached in practice (l/kg) period of field exposure (y) J N net infiltration rate (mm/y) h
The correction for the leaching of soil (E0) implemented by the regulator to account for the difference in leaching between lab and field is not applied in this study. It would lead to negative release predictions. In view of the fact that leaching data for the precise materials applied in the four available, data from other sources obtained or very similar materials have been used For the studied applications only in case CA diffusion is relevant mechanism of release for diffusion controlled systems as specified in the Building Materials Decree
works are not instead. transport. The is given by:
Ed = E(64d) * fte= * fext,v with: fter, correction for temperature difference between lab and field (0.7 here) fext.v correction for the degree of water contact and the correction for a different exposure time (64 days to 11 years, here 7.81). The coal ash stabilization under asphalt concrete is assumed to be permanently wet, which implies that the water contact correction factor becomes unity. From earlier studies leaching information on similar material as applied in this work has been used as the relevant data for the actual material applied in these is not available.
Comparison of predicted and measured release in practice From the field samples and the various calculations many release data with their uncertainties have been obtained. In view of the relatively large variations in measured and predicted release values a qualitative assessment of the comparison has been adopted. A difference between predicted and measured release at a time scale of about 10 years of less than a factor 2 is considered good to
524 very good. A difference of a factor between 2 and 5 is considered of comparable order. In case of differences by more than a factor of 5 the agreement is poor or not existing. In table 1 and 2 the results of the different methods to determine the leaching in the field are presented, for detailed information see [4]. From a comparison of predicted and measured release a large part of the measured constituent (65%) shows a good to very good agreement between field measurements (E~, E, or E,~) and the calculated release (E d or Ep) based on laboratory test data. It should be realized that prediction of n_oorelease and observing no measurable release in 10 years is also a good prediction. The level of agreement is even better, when in case of mobile elements (e.g. B, Cr, Mo, V and SO4) the release E, is corrected for the release to ground water. For these mobile elements, the agreement between the concentration increase in the underlying soil and the predicted release is often poor. The high mobility of these elements is related to the fact that all of these elements occur as oxyanionic species (e.g. borate, molybdate), and as such feature limited interaction with soil. For these elements a good agreement can be obtained, when the mass balance is closed by calculating release from concentration and volume measurements of collected percolate. This has been possible for Mo and Cr in coal ash application CZ and for Cu in MSWl bottom ash application VF and VA. The agreement between E,a and Ep is generally good. The application CA, however, the discrepancy between E, and Ed can be explained by the occurrence of a diffusion resistance at the interface between construction material and unsaturated sand layer. Based on geochemical modelling, the water management and the measured concentration profiles in the construction material as well as in the underlaying sand layer it can be concluded that transport of mobile species to ground water is highly unlikely. This implies that the E, measurement in practice is in this case the most reliable determination of actual release. The poor agreement between prediction and measurement necessitates a adjustment of the prediction to account for this diffusion resistance. In all cases, the release determination based on a concentration decrease in the construction material is the least accurate.
525 Conclusions
For the first time an extensive field study has been carried out to determine the relation between predictions based on laboratory leaching tests and the release as established by measurements at applications of secondary materials as road base materials. The study consisted of an extensive sampling program based on previous release estimates, an estimate of actual water management, leaching of the field-exposed material, concentration profile analysis of the construction material and the underlying sand layer as well as calculation and modelling of release based on these profiles. The main goal of the study was to verify the predicted release based on laboratory leaching tests with measured release under field conditions. For many relevant elements a quantifiable release has been established after about 10 years of field exposure. As expected, the variations in the measured concentrations in the soil samples were large, for some elements up to 200 %. This can be due to the small increases relatively to the background level or variability in composition between the different cores. Based on the results of validation studies of leaching tests the calculated release from laboratory data also have relatively large uncertainties (20 - 50 %). In spite of the sometimes large uncertainties in measured release most elements show a reasonable to good agreement between predicted release and measured release in the field. From this first large scale lab-field verification conclusions can be drawn:
of leaching
behavior, the following specific
General The method of sampling an analysis applied in this study provides the possibility to assess actual release after about 10 years of field exposure of construction materials. Future demonstration projects can make valuable use of these observations. Recommendations have been made in the Handbook for leaching characteristics [5].
Methodology for release measurement in practice Based on the water management in practice and geochemical modelling it proved possible to distinguish between predominantly diffusion-controlled release (coal ash under asphalt cover, CA) and percolation dominated release (other applications, coal ash and MSWI bottom ash under respectively sand and bricks). Under a closed top cover such as asphalt concrete diffusion proves to be the controlling release mechanism. The choice of the most useful methodology to determine the release (E~, E, or E,~) in practice depends on the specific situation under consideration. Determination of release from the concentration increase in the underlying soil (E,) is most appropriate for elements that sorb strongly to soil (e.g. Co, Cu, Pb, Zn). In view of the relatively low background concentrations in soil (after mild acid extraction) this approach is usually the most sensitive. In case of release by diffusion release from soil profile analysis (E,) is appropriate for all constituents. Elements which do not or weakly interact with soil (e.g. B, Cr, Mo and SO4) will be transported to deeper layers and ultimately to ground water. In this case the release based on the release between fresh and aged material can give a good estimate. Another option is the combination of a soil increase and measurements of collected percolate (only relevant for demonstration
526 projects). The release from concentration decrease of the construction material (E~) is only applicable to homogeneous materials. Due to the subtraction of two large numbers the uncertainty is generally high in this method.
Comparison of measured release in practice and predicted release according to Building Materials Decree. For several elements (CA: Cd, Cu, Ni, V and SO4; CZ: Cr, Cu, Mo, V, Zn and SO4; VANF: Cu, Pb, Ni) the predicted release according to the formulas specified in the Building Materials Decree agree well with measured release under field conditions after about 10 years exposure. In case of diffusion controlled release, the occurrence of a diffusion resistance as a result of a large difference in the level of water saturation between the construction material and the underlying soil needs to be taken into account. In all applications the release of Zn measured in practice is significantly higher than predicted based on release calculations of the Building Materials Decree. A possible explanation can be the initially higher pH in the construction material shortly after placement compared to the pH in the leaching tests that form the basis for the release prediction. -
-
Formulas in the Building Materials Decree For the evaluation of field data through a comparison of measured release in practice with predictions based on formulas in the Regulations of Construction Materials, the correction factor for soil (Eo) is omitted as it leads to negative release predictions. Based on the release predictions some elements (Cd, Co, Ni, and Cu) are hardly leached from coal fly ash stabilization leaching to a hardly measurable release in practice. These predictions are in agreement with the different field release measurements. Modelling of transport using the code ECOSAT on measured profiles in the construction material and the adjacent soil has provided indications that a diffusion is substantially less than predicted on the basis of saturated conditions. In practice the underlying sand layer is unsaturated. This leads to a diffusion resistance at the stabilization/soil interface. By introducing a correction factor in the diffusion coefficient, the results may be translated to varying degrees of saturation. For the coal ash stabilization under sand and the MSWl bottom ash under bricks, the pH of the construction material after about 10 years exposure proves to be significantly less than shortly after placement. This is caused by release of mobile pH controlling species and/or carbonation. This change in pH can lead to a significantly different leaching behavior in subsequent time intervals. The prediction of release over long periods of time (e.g. 100 year) on the basis of laboratory experiments (higher pH) can deviate significantly from the true behavior in practice. Anions such as sulphate, but also oxyanions such as molybdate, borate and chromate are poorly retained in the underlying soil. After a relatively short time (<10 y), the leachable fraction of these components can be washed out completely to the ground water in a percolation dominated scenario. In the calculations according to the Building Materials Decree this aspect is addressed for sulphate and CI, but not for mobile oxyanions. -
-
527 Literature
1. ,,Building Materials Decree. Specifications to the regulation", Staatsblad van het Koninkrijk der Nederlanden, 20 December 1995, 247. 2. ,,Determination of leaching from granular construction materials and wastes by means of a column test", NEN 7343, June 1994. 3. Th.G. Aalbers et. al .... Milieuhygienische kwaliteit van primaire en secundaire bouwmaterialen in relatie tot hergebruik en bodem- en oppervlaktewateren bescherming", RIVM rapport 7771402006/RIZA rapport 93.042, 1993. 4. J.P.G.M. Schreurs en H.A. van der Sloot, ,,CROW project: Relatie uitlooggedrag laboratorium praktijk bij wegenbouwkundige projecten", Intron rapport nr. 96082, 1995. 5. ,,Manual Characterization of Leaching Behavior", Handboek Uitloogkarakterisering (in Dutch), CROW, 1994.
528
RELD STUDIES
LABORATORY WORK
I Concentration profile in road base
I
! concentration profile in sea below applk:atk3n
L ...........
!
J
~
Concentrations in percolate
Column experiment on ~esh and aged rnatedal
1
I c==,=.~ o,.~=
]
!.....
Figure
1. Approach
of the vertification
Model calculabons
l
; Calculation of release [ k................
!
]
! CalculatJo~of release
i
Standmdleachingtest da~ I
1
i c==.~ of.,,.= ! j
I
I
Calculation of lm~sk~n, i
~
of laboratory-field
leaching
behavior
!
529
I
/segment
g
/
Co.~,-,~,n---. t
in~d layer (ReleasecalculalJonEll)
/
'
7J_J~
I
"~ I
I|
8 )
I in cons~cl~on material
I (ReleasecalculationEl])]
8
i --T--
-
Figure 2. View of the segmentation of the cores and the determination of E~ and E~ from concentration decrease and increase in the construction respectively the underlying soil.
530 Table 1. Comparison of predicted (Ep) and measured (E,, E,,, E,,,) release (mg/m 2) for the percolation scenario (shoulder of the road construction), location Colorado road Component
I[
E,
(sd)
E,,
(sd)
E,,
(sd)
(280) (6O0) (300) (140)
2049
(1945)
-
-
924.500
(528.000)
-
-
Mo V Zn
954 8130 -
(1980) (6715) -
300 1580 300 135
SO4
857.700
(353.000)
-
Or
1134
II
Ep 290 1470
(848)
440 4 2
830.000
Table 2. Comparison of predicted (Ed) and measured (E,, E,, E,,) release (mg/m 2) for the diffusion scenario (central part of the road) Project CA, location Colorado road I C~176
II
E,
(sd)
E,
(sd)
E,~
-
-
Mo V Zn
91 548 338
(130) (593) (502)
31 60 183 130
(35) (28) (270) (270)
1767 24
SO4
46.740
(25.422)
15.400
(10.000
132.150
Or
-
(sd) -
(718) (48) -
(34.200)
II
Ed 317 443 208 2.7 58.300
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
531
Leaching of chromium from steel slag in laboratory and field tests - a solubility controlled process? A M . Fallman Swedish Geotechnical Institute, S-581 93 Linkrping, Sweden e-mail" annfal@geotek, se
Abstract Differences in concentrations in leachates and leaching conditions were found in laboratory and field tests on electric arc furnace steel slag. Solid phases were assumed to control the leachate concentrations of chromium and barium. Geochemical modelling was used to indicate solubility controlling solid phases. The mineral BaSO4(c ) and solid solutions of Ba(S,Cr)O 4 were predicted to control the concentrations of barium and chromium in the leachates, respectively. These solids were not expected to be present as primary minerals in the steel slag and were rapidly formed, within 6 hours.
1.
INTRODUCTION
Steel slags are utilised in roads, as construction and paving materials, since its physical properties are similar to natural materials like gravel. Around 77%, equal to 19 Mtonnes, of the electric arc furnace (EAF) steel slag produced world wide was utilised in civil engineering during 1990.1 Environmental concerns regarding utilisation and landfilling of steel slag focus on the content of heavy metals and specially on the leachable quantities, where chromium have gained special attention due to its toxicity in the hexavalent state. 2 The conditions under which the leachates are produced varies, depending on purpose, design and local conditions. Differences such as rate of infiltration, penetration of air through the material, and particle sizes of the material may influence the resulting concentrations of e.g. metals in the leachates. 3 Laboratory leaching tests represent conditions that may be different from the field situation. 4 Variations in concentration of more than one order of magnitude were found for chromium, vanadium and lead in leachates from EAF slag under different leaching conditions in field and laboratory tests (dilution was excluded). 3, 4 Interdependencies between the concentrations of chromium, barium, sulphur and calcium in leachates from steel slag were seen in an earlier study. 5 In addition, Fruchter et al. 6 found that solid solutions of barium, chromate and sulphate controlled the concentration of chromate in the leachates from fly ash (laboratory and field tests). The objective of this study was to test if the concentrations of chromium and barium in leachates from EAF steel slag in different leaching systems could be solubility controlled.
532 2.
M A T E R I A L S AND M E T H O D S
2.1
Material and sample preparation.
The material used was slag from scrap based production of low alloyed steel in electric arc furnace. The slag was emptied below the furnace, excavated, and transported while still hot to an intermediate storage for cooling (water sprinkling). The non-magnetic, screened 0-300 mm fraction (sampled at one occasion) was used in all tests reported in this paper. The total composition of the slag (average o f two test samples) is presented in Table 1.4 (Digestion in LiBO 2 melt as well as HNO 3 in a Teflon bomb; 7, 8 analyses by Inductively Coupled Plasma Emission Spectrometry (ICP-AES) for major elements and Inductively Coupled Plasma Mass Spectrometry (ICP-MS) for As, Cd, Co and Pb).
Table 1 Total composition (mg/kg) of the steel slag. 4 Element
Element
Fe
242000
Zn t
244
Ca
221000
Nb
198
AI
219000
Sr
178
Si
57700
Cu t
166
Mg
45200
Zr
Mn
39300
Ni t
45
Cr
7760
Pb t
21.5
55.2
P
4600
Mo
20.6
Na
3700
Sn
17.5
Ti
2800
Co t
5.8
S
1260
As t
5.3
V
1210
Cd I
0.45
Ba
728
Be
<1.2
K
460
Hg t
<0.41
W 410 1 Digestion in HNO 3 and Teflon bomb.
2.2
Laboratory Leaching Tests.
Leaching tests were carried out as column tests, pH static tests and single batch tests. 4, 5, 9 The laboratory samples were air dried at 50 ~ and stored in closed containers until use. Oversized material, according to the specifications, was size reduced in a jaw crusher. Ultra pure water ( <0.2 mS/m) and 0.1 mM HNO 3 (column test) were used as leachant. Bottles and beakers of polypropylene (PP) or polyethylene (PE) were used in the leaching tests and bottles of PE were used for storage of leachate samples. The column was made of polyacrylate.
533 Leachates were pressure filtered (N2) in a polytetrafluoroethylene (PTFE) coated stainless steel filtering device through a 0.45 lam cellulose nitrate filter (D 100 mm). Column tests were performed on test samples of grain size of 0-20 mm (columns with the diameter 0.1 m and height 0.9 m) with inflow in the bottom at a flow rate corresponding to a leachate to solid (L/S) ratio of 0.1 per day (maintained by a peristaltic pump, Gilson, Minipuls 3). 4 A 1.2 lam glass fibre filter was placed in the column at the outlet as a pre-filter. The outflow leachate was further passing through a 0.45 lam filter and accumulated in a closed bottle under N2-atmosphere. Accumulated samples at L/S-ratios of approximately 0.1, 0.3, 0.7, 1.0, 2 and 4 were taken out. The pH static tests were carried out at L/S=5 for 24 hours on test samples of grain size 0-4 mm (75-150 g)under continuous measurement and adjustment of pH by a titrator (Radiometer TIM 90) to +0.05 units, with either 0.1-0.5 M HNO 3 or 0.5 M NaOH. 5 The leachates were mixed with a PTFE coated propeller. The beakers were covered but not air tight during the test. Single step batch tests were carded out at L/S ratios 1.7 and 4.25 for 3, 6, 24, 48, 168 or 336 hours with end over end rotation (10 rpm) on grain sizes 0-1 mm and 0-8 mm. 9 Care was taken to minimise the head space in the bottles (100 or 250 ml). The tests were carried out in a two level factorial design 1~ with the parameters L/S ratio, grain size and time. The factorial design was evaluated according to an additive model based on the saturation indices, defined below, and based on 2 standard deviations (SD).
2.3
Lysimeter. A lysimeter of the size 3.0 x 3.0 x 1.2 m 3 made of plywood and covered with pre-formed HDPE liner was filled with approximately 10 m 3 slag corresponding to 21 tonnes of material. 4 The percolate was collected at the centre of the bottom. A geotextile attached to a synthetic draining layer on top of the bottom liner prevented particles from entering the leachate collecting system. The percolates were continuously flow measured and proportionally sampled by a tipping bucket system placed indoors in the basement of a nearby building. The sampled leachate was stored in a bottle under argon gas to prevent CO 2 uptake and oxidation. 2.4
Test Road. The slag was used as road base material in a 50 m long section of an unpaved road (width 5 m and average height 2 m). 3 The test part was divided into two sections: one with the 0-300 mm size fraction and the other with the fraction 11-300 mm where material of grain size below 11 mm had been removed. Two leachate collection units (lx4 m 2 at 3% slope) were placed in each section at depths of 0.5 and 1.5 m below the road surface. The leachates were collected in a similar system as in the lysimeter but the leachates were stored open to the atmosphere. Samples for pH measurements were taken directly from the tipping bucket. The experimental program in this study was started 1.5 years after the test road was built. 2.5 Leachate Analysis. Leachates were analysed with respect to pH, redox potential (Pt and calomel electrodes) and concentration of elements (ICP-AES, ICP-MS). Chromium was in some cases analysed by Atomic Absorption Spectrophotometer with Graphite Furnace (AAS-HGA). Sulphate was analysed by ion chromatography in the leachates from lysimeter, test road and batch tests. In the other leachates it was assumed that all sulphur were present as sulphate. Alkalinity was measured by titration with 0.1 M HCI (not in the batch test leachates).
534 2.6
Geochemical model.
The geochemical code MINTEQA2 version 3.11 was used for speciation calculations. 11 Saturation indices (SI) were derived from the calculated ion activity products (IAP) and solubility products (Ks) for the relevant minerals (SI=log IAP-log Ks). Input data were measured pH, Eh and concentrations of inorganic substances in the leachates (AI, As, Ba, Ca, Cd, Cr, Cu, Fe, K, Mg, Mn, Mo, Na, Ni, Pb, SO42-, Si, V and Zn). The pH measurements in the batch test leachates were not reliable due to disturbances of Na. The pH values for these leachates were instead obtained from the calculations (ion balance assumed). The redox potential and the pH were not allowed to change in the simulations and solids were not allowed to precipitate. Redox reactions were specified for Fe(II)/Fe(III), Mn(II)/Mn(III) and Cr(III)/Cr(VI). The state of oxidation was fixed for V(V), As(V), Cu(II) and S(VI). The temperature used was 25 ~ The thermodynamic data base provided by the code was used with amendments as shown in Table 2. Further changes in the data base were addition of Mo and revision of the constants for hydroxides of Cr(III) and Cu(II). 12-14
Table 2 Selected stability constants amended or chan~;ed in the data base. 15 Mineral
Ks
Ba(S0.77Cr0.23)O4
- 10.13+0.07
Ba(S0.96 Cr0.04)O4 BaSO 4
-9.79+0.06 -9.78+0.06
3.
RESULTS AND DISCUSSION
The differences in leaching conditions between the tests are illustrated by the differences in pH, Figure 1. Leachates from the laboratory systems were alkaline, while neutral from the lysimeter. The test road showed pH values comparable to those from the laboratory tests in the 0-300 mm material while the 11-300 material gave pH values between the laboratory tests and the lysimeter. The leachate compositions were also differing between the tests, see Table 3. Chromium concentrations were highest in the lysimeter and test road with 11-300 mm material, while barium concentrations were highest in the laboratory tests and the test road with 0-300 mm material. The conclusion was that different chemical conditions were dictating the leaching systems and that geochemical modelling was necessary to further understand the mechanisms controlling the leachate concentrations.
535 14 13
[]
Batch tests < 1 mm
12
9
Batch tests <8 mm
11
---o---- Column
10
•
-
"-
Lysimeter
+
Test road 0-300 mm
•
Test road 11-300 mm
9 8
7 6 0.0001
I
t
t
t
I
0.001
0.01
0.1
1
10
L/S
Figure 1
pH in the leachates from different tests on steel slag.
Table 3 Concentrations in leachates (mg/l) from the different laboratory and field tests. Parenthesis mark maximum values measured previously but not used in this study. 3 Element
Ca
S(tot)
SO42--S
Ba
Cr
Batch test <1 mm
255-910
<0.16-2.62
<0.1-0.8
0.11-51
<0.0005-0.030
Batch test <8 mm
130-280
1.1-5.0
0.3-3.2
0.41-3.8
<0.0005-0.075
pH static
94.9-1900
3.08-10.5
NA
0.26-3.61
0.0031-0.060
Column
108-236
1.38-18.3
NA
0.32-1.21
0.013-0.13
Lysimeter
35.4-54.8
5.4-49
5-97
0.072-0.12
0.086-0.73
Test road 0-300 mm
350-900
NA(6.9)
0.2-13
7.2-15.6
0.024-0.048 (0.10)
Test road 11-300 mm
9.1-170
NA(30)
0.9-11
0.20-0.51
0.032-0.081 (0.66)
NA= not analysed
Results from the speciation calculations expressed as SI versus pH are given in Figure 2 for BaSO4(c), gypsum, witherite, and in Figure 3 for BaCrO 4 and barium sulphate/chromate. The barium concentrations in the leachates appeared to be controlled by barium sulphate for all leachates except those from the batch test with grain sizes <1 mm. Gypsum did not limit sulphate. The steel slag is produced in a reducing system where sulphates are not likely formed. 16 Less than half of the sulphur content was analysed as sulphate in most (75%) of the leachates from batch tests. Sulphate minerals could be formed aider oxidation and were likely formed as secondary minerals. The high SI values for BaSO 4 in some column and test road
536 leachates could possibly reflect an overestimation of the SO42--Sfraction. However, this was not confirmed. Co-precipitation with Sr was proposed by Fruchter et al. 6 to explain over saturation of BaSO 4. This could not be confirmed in this study since Sr was not analysed. Barium carbonate (witherite) could be an additional candidate controlling barium. The Sis for witherite, see Figure 2, were pH dependent, and all the calculated systems were falling into the same pH dependency pattern. The high Sis for witherite in the test road with 0-300 mm material could be due to in-growth of carbonate during the exposure to the open air. The Sis for witherite in the column leachates were lower than for BaSO 4 and thus less likely to control the barium concentrations. Barium chromate was the chromium mineral in the MINTEQA2 database showing Sis closest to zero. However, these were too low to indicate solubility control of the chromium concentrations (see Figure 3). Rai et al. 15 found that solid solutions of barium, sulphate and chromate were obtained from over saturated solutions of these components. They also determined the stability constants and ratios between sulphate and chromate in the solid solutions. The introduction of these stability constants into the MINTEQA2 database resulted in calculated Sis close to 0 for all leachates except from the batch tests on <1 mm material, see Figure 3. The Sis of the Ba(S,Cr)O 4 solid solutions were close to those for barium sulphate, but different from the Sis for barium chromate. Neither of the solid solutions with 4% and 23% chromium, respectively, could be regarded as dominant in the control of chromium concentrations. Fruchter et al. 6 also found that these solid phases appeared to control the leachate concentrations of chromium in their studies on coal fly ash. Chromium is likely found in Cr20 3 or Cr-(Fe,Mg)-oxides in the slag. 2 Chromates were likely formed after oxidation. The measured redox potential was used to assess the partitioning between Cr(III) and Cr(VI) in the leachates. The Cr(VI) activity at pH 4-6 was 23, 9 and 11 magnitudes lower, respectively, than the Cr(III) activity. In spite of these low modelled activities the calculated Sis for the Ba(S,Cr)O 4 solution with 4% chromium were close to saturation. The factorial design evaluation of the Sis from the batch tests leachates gave as result that BaSO4(c ) and the Ba(S,Cr)O 4 phases might be solubility limiting (i.e. SI-0 within the confidence interval of 3 SD) for the tests on <8 mm samples (see Table 4). For the small grain size <1 mm samples only BaSO 4 (c) (only L/S=I.7) appeared to represent saturation. There was no time effect on SI for these solids.
Table 4 Predicted intervals for SI (3 SD) based on factorial design interpretation of the batch tests. Grain size 0-1 mm No L/S effect BaCrO 4
L/S 1.7
L/S 4.25
[-3.40, -1.83]
Grain size 0-8 mm No L/S effect
L/S 4.25
[-2.79, -1.22]
[-1.81, 0.93] [-2.93,-0.191
BaSO4 (c)
L/S 1.7
[-1.72, 1.021 [-1.81, 0.931
Ba(S0.96,Cr0.04)
[-2.22,-0.211
[-1.46, 0.551
Ba(So 77,Cro 9"~)
[-2.13, -0.25]
[-1.30, 0.58]
537
1.5
BaS04(c)
+ o o
0.5
5
-0.5
6
7
8 I~ x 9
•
11
~ + 13
>(
~o rn
-1.5
pH
0 -0.5
r~
I
I
I
I
I
I
I
6
7
8
9
10
11
12
CaS(M.2H20 (Gypsum)
~ -1.5
I
5
/
/
~
~
~
*
~
_
0 +
9
0
-2 -2.5
•
o X
-3
+
IITO
cl rm
x
-3.5 -4
o
I
I
I
I
I
i
J
i +.
5
6
7
8
9
10
11
1 2 4+
1-
-0.5 -1 -1.5 r~
-2
I
13
t~
BaCO3 (witherite)
Batch tests <1 mm Batch tests <8 mm
S
Lysimeter O
Column
+
Test road 0-300 mm
•
Test road 11-300 mm
-2.5 -3 -3.5 -4
- pH-static tests pH
F i g u r e 2.
S a t u r a t i o n i n d e x ( S I ) v e r s u s p H f o r B a S O 4 ( c ), C a S O 4 - 2 H 2 0 ( g y p s u m ) , B a C O 3 ( w i t h e r i t e ) in t h e l e a c h a t e s f r o m d i f f e r e n t tests.
538 I
I
I
I
I
I
I
I
I
5
6
7
8
9
10
11
12
13
BaCr(N
+
9
++
UfIN
4-+
C,I
r,~-3
go o
pH 1.5
Ba(S0.96, CrO.O4)04
+ o O
0.5
I
9
~,
__5--"o~
w,.,w
T,#}
-0.5
o
-1.5
pH 1.5
Ba(S0.77, o0.23)04
[]
Batch tests <8 mm
0.5
o
I
-0.5
Batch tests <1 mm
/
6
I
I
~
9
IX
t+
~,/,~,0o
+
i~
Lysimeter I
13
-1.5
/
2
o
Column
+
Test road 0-300 mm
•
Test road 11-300 mm pH-static tests
pH Figure 3.
Saturation index (SI) versus pH for BaCrO4, Ba(S0.96Cr0.04)O 4 Ba(S0.77Cr0.23)O 4 and in the leachates from different tests.
539 4.
CONCLUSIONS
It is indicated from this study that concentrations of barium and chromium may be controlled by the solubility of well defined solid phases. The mineral BaSO 4 and solid solutions of Ba(S,Cr)O 4 were predicted to control the concentrations of barium and chromium in the leachates, respectively. Solubility controlled conditions were rapidly achieved, within 6 hours, in the batch test. None of these phases were expected to exist as primary minerals in the steel slag. However, the presence of these predicted solids associated with leached slags need to be verified.
ACKNOWLEDGEMENTS This work was funded by the Swedish Waste Research Council under grants AFR Dnr 324/92 and 126/93 and in addition based on work funded by Fundia Special Bar AB.
5.
10
REFERENCES
The Management of Steel Plant Ferruginous By-Products, International Iron and Steel Institute, Committee on the environmental Affairs and Committee on Technology, Brussels (1994). Ye, G., Burstr6m, E. and Fallman, A-M. Utilisation and Stabilisation of Steelmaking Slags. Swedish Waste Research Council, AFR-Report 57, Stockholm (1995). Fallman, A-M. and Hartl6n, J. Utilisation of electric arc furnace slag in road construction. In Environmental Geotechnics, M. Kamon (ed.), pp 703-708. Balkema, Rotterdam (1996). Fallman, A-M. and Hartl6n, J. Leaching of Slags and Ashes - Controlling Factors in Field Experiments versus in Laboratory Tests. In Environmental Aspects of Construction with Waste Materials. Goumans, J.J.J.M., van der Sloot, H.A., Aalbers Th.G. (eds.), pp 3954. Studies in Environmental Science 60, Elsevier, Amsterdam (1994). Fallman, A-M. and Aurell, B. Leaching tests for environmental assessment of inorganic substances in wastes, Sweden. Sci. Total Environ. 178:71 (1996). Fruchter, J.S., Rai, D. and Zachara, J.M. Identification of solubility-controlling solid phases in a large fly ash field lysimeter. Environ. Sci. Techn. 24:1173 (1990). ASTM D 3682-91 Standard Test Method for Major and Minor Elements in Coal and Coke Ash by Atomic Adsorption. SS 02 81 83 Metal content of water, sludge and sediments - Determined by flameless atomic absorption spectrometry - Electrothermal atomization in a graphite furnace General principles and guidelines; Swedish Standards: Stockholm, (1986). Fallman, A-M. End point criterion in batch leaching tests for granular solid wastes. Environ. Sci. Technol. (submitted) Box, G.E.P., Hunter, W.G. and Hunter, J.S. Statistics for Experimenters; John Wiley & Sons: New York, (1978).
540 11 12
13 14
15
16
Allison,J.D., Brown, D.S. and Novo-Gradic, K.J. MINTEQA2/PRODEFA2, A Geochemical Assessment Model for Environmental Systems: Version 3.0 User's Manual. Environmental Research Laboratory; U.S. EPA: Athens, GA (1990). Rai, D., Zachara, J.M., Schwab, A.P., Schmidt, R.L., Girvin, D.C. and Rogers, J.E. Chemical Attenuation Rates, Coefficients and Constants in Leachate Migration. Volume 1" A critical Review. EPRI EA-3356, Electrical Power Research Institute. Palo Alto (1986). Rai, D., Zachara, J.M., Eary, L.E., Girvin, D.C., Moore, D.A., Resch, C.T., Sass, B.M. and Schmidt, R.L. Geochemical behaviour of Chromium Species. EPRI EA-4544, Electrical Power Research Institute. Palo Alto (1986). Comans, R.N.J., van der Sloot, H.A. and Bonouvrie, P.A. Geochemical reactions controlling the solubility of major trace elements during leaching of municipal solid waste incinerator residues. In Municipal Waste Combustion Conference. Kilgroe, J. (ed.), pp 667-679. Air and Waste Management Association, Pittsburg, PA (1993). Rai, D., Zachara, J.M., Eary, L.E., Ainsworth, C.C., Amonette, J.E., Cowan, C.E., Szelmeczka, R.W., Resch, C.T., Schmidt, R.L., Girvin, D.C. and Smith, S.C. Chromium Reactions in Geological Materials. EPRI EA-5741, Electrical Power Research Institute, Palo Alto (1988). Rosenqvist, T. Principles of extractive metallurgy. McGraw-Hill, New York (1974).
542
Abstract Today Municipal Solid Waste Incinerator Bottom Ash is to a large extent used in road construction on Denmark. It is primarily used as fill in embankments and as granular subbase course. This is the result of development during the last 20 years; at the same time incineration of waste has become an substantial element in the handling of waste in Denmark. The use of bottom ash as subbase in light and medium trafficked roads, paths and parking areas has proved to be successful. In 1993, a more heavily trafficked test road was constructed. Layers of 20 cm bottom ash from various plants was used as subbase and compared to the standard material which is virgin, relatively uniform sand. After 3 years of medium to heavy traffic, the test road shows no signs of rutting or progressive damage or other signs of rapid deterioration of the road construction. However, a period of 3 years has been estimated to be too short to recommend general use of bottom ash in heavily trafficked roads. This paper treats the results from the test road and the Danish experience gained from the use of incinerator bottom ash in road construction.
543
1.
Introduction Municipal Solid Waste Incinerator Bottom Ash (MSWI bottom ash or bottom ash in
following) is the solid residual product, which is found at the bottom of the incinerator in plants used for burning municipal solid waste. Incineration of municipal solid waste is an integrated part of the Danish way of handling waste. The general priorities are: 1. Minimise waste production, 2. Recycle, 3. Incinerate, 4. Dispose in landfill. In the 1970'ies and 1980'ies several examinations were carried out in Denmark to describe the properties of MSWI bottom ash for construction purposes. The main result of the tests was that sorted bottom ash is well suited as subbase for lightly trafficked roads, parking areas, paths and other trafficked areas. Tests also showed that bottom ash can be used as fill material for embankments. The amount of waste which is delivered to incinerator plants has increased in recent years. Based on information for 1994, the annual amount of waste delivered to the 31 incinerator plants in Denmark is 2.1 mill. metric tons. The incineration process results in a weight reduction of some 75%. After a cooling and storage period, magnetic metals and larger particles are removed, so that the remaining ash, sorted bottom ash, can be used for building and construction works. Since the road sector is one of the largest consumers of sand, gravel and stone materials, it has been this sector which has used the major part of the bottom ash produced in Denmark. Ten years ago, the State has introduced taxes to stimulate recycling: Today, ECU 28 per ton is charged for incineration in plants which produce both electricity and heating whereas a tax of ECU 35 per ton is charged for plants that only produce heating. For depositing waste, the tax is ECU 45 per ton, and from 1997, it is not allowed at all to dispose waste which can be burnt.
544
0
Recycling as subbase material The results of the tests and investigations which were carried out in the early 1980'ies by
Schmith (ref. 1 and 2) showed that it is reasonable to use bottom ash as subbase for lightly trafficked roads and areas. Based on this information counties and local authorities have used considerable amounts of bottom ash as general fill and as subbase layer. The use is typically concentrated in construction projects which take place in the same county as the incinerator plant. The experiences gained from the construction phase are positive. In general, the sorted bottom ash can easily be handled and laid, on condition that the moisture content in the bottom ash is close (+ 3%) to the optimum moisture content, and measurements have documented that strengths can be obtained which are fully satisfactory for the construction job. Since 1989, the construction projects for light traffic, where bottom ash has been used, have been carried out based on Material Specifications and Working Specifications (Pihl et al. Ref. 3). Regulations from 1983 based on public law issued by the Ministry of the Environment restricted the permits for the use of bottom ash seen from an environmental point of view.
/
,
~penhagen
9 Incinerator Plants
Fig. 1
Map of Zealand, showing actual incinerator plants and the test road.
545
3.
Test road 1993 To investigate a possible use of sorted bottom ash for moderate to heavy traffic, a full
scale test was carried out in 1993. The test sections are situated on a newly constructed ring road passing the town of Sk~elskor on Zealand, see Fig. 1. The starting point was a traditionally constructed road on a subgrade of normal Danish moraine clay. The test road was constructed in 1993, a total of 2.3 km with a cross section as shown in Fig. 2. The traffic load was predicted to be moderate to high with many relatively heavy vehicles.
o , - A s p h a l t Wearing Course 30 mm ,. . . . 20,/oo 0 ~ ..~ ~- /1__Asphalt Concrete 90 mm ~--. ;;.....~,-~-.:.L:.-:~.-..~::.~.~ ;,:-..;~::..------ Gravel Base 200 mm
....... ----,,---..~ ~ ' . . - i ~ . . : ~ - : ~ ; .
~ 9
~
. . . . . - . - . . _ - . .
~/~:~:~,~.../,:
:
.
.
.
.
.
.
..
9
:?;i-::~::i!).?.i~:~:(".~~::.:~): ::!i~):i.)::.~i":~:/)"i!-- Bottom Ash/Sand ~.~/~.~~ f,~..~;:,~!$~,~/r 7~ !/~r~<~ Moraine Clay/Fill 9
.
.
.
-
.
.
.
.
.
-. . . . . . .
9
.
.
300 mm
Fig. 2 Cross section of the test road Four sections of 200 m each were given a different subbase material; three different bottom ashes, from the Amager Plant, Vestforbraending Plant and KAVO Plant - and a traditional material - relatively uniform sand - as a reference material all used as subbase material. A private enterprise AFATEK was responsible for running the project; furthermore the Danish VKI (Water Quality Institute) and the Danish Road Institute participated in the project. The project period was from Spring 1993 to Autumn 1995 and later the Danish Road Institute has kept the test sections under observation and carried out tests on the road.
4.
Laboratory testing The aim of the tests in the laboratory was to evaluate the properties of the bottom ash
for construction purposes compared to the properties of the reference material. The following tests and material examinations were carried out: Grain size distribution (Fig. 3), grain density,
546
SILT
100
GRAVEL
SAND
/ - Sub ase
S
I
5o
/
o
.06
.02
Fig. 3
STONE
J
.2
I~ncinerator
/
.6
2
6
Bottom Ash.
20
60
200
Grain size distribution for sorted bottom ash and sand (average value)
compaction test and CBR-tests, Los Angeles tests, loss of ignition, etc. Furthermore, freeze-thaw properties, permeability, capillarity were also examined. Some of the material tests were carried out at the time of delivery before laying on the road site, some after laying and others after 1 and 2 years in operation; materials were taken from all test sections. An important result from the analyses of the bottom ash was that crushing during construction work was apparently the strongest influence on the material, stronger than the influence of traffic in the following two years on the completed road. Some vibration tests in the A
Test Grain density fraction 0 - 16 mm
Ps
(t/m3)
2.51-2.52
2.65-2.67
2.57-2.58
Loss of ignition 1000 oC fraction 0-1 mm
GI
(%)
14.8-14.9
11.3-11.8
15.4-16.7
Los Angeles according to ASTM C131
LA (B) (%) LA (C) (%)
53 46
53 49
45 48
Vibration dry density and moisture content according to ASTM 2049-69
Pd, m~
(t/ms)
1.54-1.55
1.65-1.67
1.54-1.54
(%)
22.7-23.0
20.3-20.5
23.3-23.3
Wv
A: Bottom ash, Amager Plant B Bottom ash, Vestforbr~ending Plant C Bottom ash, KAVO Plant
Table 1." Laboratory tests of sorted bottom ash
547 laboratory showed that crushing was particularly big when the material was dry, reduced when it was wet, and least when the material was almost water saturated. The permeability of the bottom ash was measured in the laboratory according to the principle of measurement on water saturated materials with decreasing water pressure. The permeability of the bottom ash was five to ten times lower than that of the reference material (sand material). Bottom ash / Sand Amager Plant Vestrorbramding Plant KAVO Plant Sand 1)
Permeability rn/s 0.5-0-7 0.9-1.4 0.2-0.4 7-8
1)
x 1 0 -6 x 1 0 -6 x 1 0 -6 x 1 0 -6
Measured on water saturated material with decreasing water pressure
Table 2: Permeabifity from laboratory tests
5.
Tests on the road The aim of the tests on the road was to evaluated the functional properties of the bottom
ash material. The purpose was also to observe any special conditions which appear when handling bottom ash including spreading in layers, regulating and compacting. In the construction period the subgrade, subbase and the basecourse was controlled according to current specification, in regard to material control and compaction. Isotope measurements were carried out to control the compaction of traditional materials, while it was necessary to use the sand replacement method in order to determine the degree of compaction of bottom ash. For the measurement of the traffic, equipment was installed to make it possible to measure vehicles and axles continuously, and during the first two years all vehicles were weighed on two different days. The traffic volume on the two-lane road was converted to an ADT of
548 approx. 5.5 104 equivalent 10 ton axles in the one direction and approx. 3.0 104 equivalent 10 ton axles in the other direction, which is considered to be moderate to heavy traffic according to the terminology of the Road Standards. The difference in the traffic volume in the two directions is due to the fact that there is a brewery at the end of the road, so that heavily loaded vehicles leave the factory. They return to the factory empty. Bearing capacity was measured by means of static plate bearing tests on the subgrade, on the layer of bottom ash and on the reference layer as well as on top of the unbound gravel basecourse. Four times during the period the completed road was measured by means of the Falling Weight Deflectometer. The following conclusions can be drawn from the static plate bearing tests: The average E-value for the bottom ash is 81 MPa and the bearing capacity of the material is therefore at the same level as that of the reference material consisting of sand, with an average E-value of 79 MPa. Bottom ash from Vestforbr~ending Plant has on average the highest E-value. The same pattern can be seen for the dynamic E-values (from the Falling Weight Deflectometer measurements): Amager bottom ash and KAVO bottom ash have significantly lower E-values than Vestforbr~ending bottom ash and the reference material (sand). In order to prove the form stability and structural quality of the bottom ash and therefore its suitability as an element in road construction, measurements of evenness and rutting on the road surface were carried out (longitudinal and transversal evenness). Comparisons between the sections with bottom ash and the reference section should prove its suitability. The following conclusions can be drawn regarding evenness and rutting: There was no difference in evenness between the four test sections, three with bottom ash and one with sand as subbase material. There was no significant change in rutting in the project period for the four test sections. The measured mean values were small and in practice within the uncertainty of measurements. According to these facts there was no visible damage, unevenness or rutting on the road sections.
549
The conclusion showed that there was no difference between the sections with bottom ash and the reference section.
6.
Conclusions After three years of relatively heavy traffic on the test road no rutting, development of
damage or other signs of rapid deterioration of the road construction has been observed. Recycling of MSWI bottom ash as subbase material for lightly trafficked roads and other areas can therefore continue, seen from a road construction point of view. However, the most recent examinations have resulted in consequences for continued use. The relatively low permeability which has been measured on the bottom ash materials has resulted in a recommendation: no use must take place in connection with difficult moisture and subgrade conditions, where the bearing capacity of a road construction is dependent on quick drainage. Since loss of ignition is not thought to be suitable as evaluation for the properties of bottom ash as subbase material, this has been replaced by other limits for unburnt matter in bottom ash. It is also necessary to change the present general construction specifications, so that excessive compaction and unnecessary traffic directly on top of the bottom ash is kept to a minimum, in order to avoid that the bottom ash is crushed too much during laying, compaction and later in the construction process. Three years is considered to be too short a period to give a general recommendation regarding the use of MSWI bottom ash on moderate to heavily trafficked roads. The relatively low measured E-values (bearing capacity) and the lower permeability in relation to traditional road materials are negative conditions which must be considered. It has, however, been decided to continue observations of the test road.
550
7.
The future Since there are good results of the use of Danish MSWI bottom ash as fill in the con-
struction sector, recycling of the material in this way will continue in the future. The evaluation of the latest laboratory and test road results lead us to the conclusion that Danish MSWI bottom ash under certain conditions can also continue to be used as subbase in roads, paths and other trafficked areas. New Material Specifications and General Construction Specifications were approved in 1996 (Pihl and Milvang, Ref 4). The present environmental demands are expected to be changed in the near future. Possibly more stringent rules will be seen for use of bottom ash in areas of great interest for drinking water, but it is not expected that this will have any influence on the continued use of bottom ash as fill and subbase material in the Danish road sector.
References Schmith, N.B. (1983) Incinerator Residue as a Pavement Material III, Laboratory tests, Note 176, Institut for Veje, Trafik og Byplan, Danmarks teknisk Hojskole & Vejdirektoratet, Statens Vejlaboratorium 2.
Schmith, N.B. (1983) Incinerator Residue as a Pavement Material IV, Recommendations and economic evaluation
Note 177, Institut for Veje, Trafik og Byplan,
Danmarks teknisk Hojskole & Vejdirektoratet, Statens Vejlaboratorium 3.
Pihl,
K.A.,
Ahrentzen,
forbrcendingsovnsslagge,
P.,
Kalsmose,
Laboratorierapport
K. 66,
(1989)
Bundsikringslag
Miljoministeriet,
Skov-
af
og
Naturstyrelsen & Vejdirektoratet, Statens Vejlaboratorium Pihl, K.A., Milvang-Jensen, O. (1996) Bundsikringslag a f forbr~endingsslagge, Vejteknisk Institute, Rapport 78, Vejdirektoratet
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
Acid Resistance of Different and Solidified Wastes
Monolithic
551
Binders
J.A. Stegemann a and C. Shi Water Technology International Corporation operators of the Wastewater Technology Centre and the Canadian Clean Technology Centre Burlington, Ontario, Canada L7R 4L7 aPresent address: Imperial College of Science, Technology and Medicine, London SW7 2BU
Abstract Laboratory tests which add acid to a ground sohdified product are useful for examinmg acid neutralizing capacity and dissolution of metals as a function of pH, but do not examine the effect of acid on the monolithic structure of the cement-based matrix, which is also important for reducmg leachability of contaminants. In this study, specimens of five different cementing systems (portland cement, portland cement with silica fume, alkali-activated ground blast furnace slag, coal fly ash with lime, and high alumina cement with lime and gypsum), with and without waste addition, were immersed in three types of acid (pH 3 nitric acid, pH 3 acetic acid and pH 5 acetic acid) to mvestigate the effects of acid attack on a monolithic matrix. It was found that calcium silicate hydrate-based formulations with a low Ca/Si ratio have the greatest acid resistance. Acid resistance of sulphated high alumina cement specimens was poor. 1. I N T R O D U C T I O N Recent results from a field study show that diffusion-limited release of alkalinity is an important control on leachability of metals from wastes treated by solidification with cement binders (Stegemann et al., 1996). The initial porewater pH of a cement-based sohdified waste is typically higher than 12, resulting in high solubility and leachability of amphoteric metal contammants. In the disposal environment, acidic or neutral conditions in contact with the solidified waste will deplete the excess lime and soluble alkalis which cause this high alkalinity and high leachability. Depending on the nature of the hydration products composmg the sohdified waste matrix, incongruent dissolution can lead to formation of a less alkaline to acidic outer "leached layer" on the surface of a monolithic solidified waste, enclosmg an inner zone of higher alkalinity. The formation of this leached layer in portland cement-based matrices has been demonstrated by several authors (as summarized by Hmsenveld and Bishop, 1996). The pH of the leached layer will range from that of the mner zone to that of the surrounding environment, and will necessarily include a zone with pH 10 to 12, where the solubility of amphoteric metal contaminants will be low, resultmg in secondary precipitation of metals diffusing outward from the highly alkaline m n e r zone (as well as inward from more acidic zones). Alkalinity from the interior of the monolith will replenish the leached layer by diffusion so that this mechanism for capture of amphoteric heavy metals remams active. In addition to ensuring the controlled release of alkalinity important for immobilization of amphoteric heavy metals, the monolithic structure of a solidified waste limits the surface area of contact between groundwater and all contaminants, mcludmg organic compounds and anions, so that leachmg of all contaminants is controlled by diffusion rather than advection.
552
By contrast, in a non-monohthic sohdified waste, uncontrolled contact between the waste and acidic or neutral groundwater will result in an extended period of high alkalinity and high leachability, without the benefit of the leached layer contaminant capture mechanism. Although, depending on the sohdified waste composition, an optimal pH environment for precipitation of amphoteric metals may be established over time, the extension of the period of high leachability will by then have resulted in transport of a much higher fraction of contaminants mto the environment. Also, because leachant transport will be by more rapid advection rather than by diffusion, there will be increased contact with acidifying or neutralizmg influences in a nonmonohthic waste and the pH will drop below the optimal range for metals precipitation more rapidly than in a monolithic waste. Thus, the monohthic structure of a sohdified waste has important benefits for mmimizmg release of contaminants into the environment, which an equally alkalme non-monohthic material can not provide. Typical properties which are measured to assess the durabihty of the monohthic structure of a sohdified waste are: unconfined compressive strength, and freeze/thaw, wet/dry and biological weathering resistance (e.g., Wastewater Technology Centre, 1991, United States Environmental Protection Agency, 1993). Maintenance of an appropriate pH has long been recognized as being critical for precipitation of metals, but acid resistance is also an important aspect of the matrix durability of a monohthic sohdified waste. Resistance of a cement-based matrix to acid attack will depend not only on the ability of the matrix components to neutrahze acid, but also on the matrix microstructure, which determmes the surface area in contact with acid, and the characteristics of the degradation products from acid attack, which may form a protective surface layer (Pavlik, 1994). The response of different solidified waste matrix components to acid addition was the subject of previous work (Stegemann et al., 1994), which included a review of the chemical stability of cement hydration products over a range of pHs. This aspect of acid resistance was explored experimentally by using different binders to create a range of solidified products with different hydration products, and measurmg their acid neutralization capacity and dissolution in a series of batch extractions of ground samples (<0.2 mm) with mcreasmg amounts of nitric acid (Stegemann and Cote, 1991). The present work undertakes to examine and compare the effect of acid on the physical structure of different matrices by m e a s u r m g the rate of corrosion of monolithic specimens of the same sohdified products upon immersion in acid.
2. M E T H O D S AND M A T E R I A L S
2.1. Preparation of Monolithic Specimens Five different sohdification bmders, representmg different hydration products, were chosen for this work. The binders were: portland c e m e n t , portland cement with silica fume, alkah-activated ground blast furnace slag, coal fly ash with lime, and high alumina cement with lime and gypsum. The formulations were identical to those used in the previous study of acid neutralization capacity of ground sohdified products, and are summarized in Table 1. The different bmder systems have been numbered from 1 through 5, and a "W" mdicates the batches contammg waste. Binders 1 to 4 were chosen to allow comparison between the corrosion resistance of portland cement, and blended cements contammg industrial by-products. These four systems also have different Ca/Si ratios (shown in the last row of Table 1). Bmders 1, 2 and 4 were used to solidify a platmg sludge containing heavy metals; the set of binder 3 (activated blast furnace slag) was mhibited by the plating sludge in previous trials, so it was not suitable for testmg as a monolith.
553
Bmder 5 was selected to produce a high proportion of ettringite (3CaO.A12Oa.3CaSO4.32H20) upon hydration, and was used to solidify a hazardous waste incinerator ash c o n t a m m g high levels of chloride and sulphate. The proportion of h y d r a t e d lime to high alumina cement was increased and gypsum was omitted in solidification of the ash, to allow the sulphate and chloride from the ash to form ettrmgite and calcium chloroalummate (3CaO.A12Oa. CaCl~. 10H20). Table 1. Formulations for monolithic s Component , Portland cement Silica fume Blast furnace slag Sodium metasilicate Class F coal fly ash High calcium h y d r a t e d lime High alumina cement Gypsum Metal plating sludge Hazardous waste ash Water Ca/Si mole ratio
Percentage 1 lW 2 100 40 80 20 . . . . . . . . . . . . . . . . . . . . 60 . . . . 40 50 40 3 3 1.4
ected to acid attack in batch (based on dry weight) 2w 3 4 4W 5 32 . . . . . 8 . . . . . 92.5 7.5 . . . . . 80 48 . 20 12 10 . . . 60 . . . 30 60 40 . . . . 50 40 40 50 40 1.4 0.5 0.5 0.5 i
i
,
5W
10 30 60 50 -
All batches without waste were prepared at a water to cement ratio of 0.4, whereas a water to solid ratio of 0.5 was required in solidification of the waste. Monolithic specimens were cast in clear polystyrene 4.5 cm diameter x 7.4 cm high right circular cylmder moulds. Samples were moist cured at 22~ for one year prior to testing.
2.2. Corrosion E x p e r i m e n t s Prior to exposure to an acidic environment, each monolithic specimen was s a t u r a t e d and cemented mto its polystyrene mould using epoxy glue such that only one end of the cylinder was exposed. A thus partially enclosed monolithic specimen of each formulation was suspended with the exposed surface facing downwards in 1 L of three types of acid solution: 1) Nitric acid, continuously adjusted to a target pH of 3. Nitric acid, HNO3, was chosen because it is a strong mineral acid which dissociates completely, whose salts are soluble. It has been used in previous work for m e a s u r e m e n t of acid neutralization capacity of ground samples, 2) Acetic acid, continuously adjusted to a target pH of 3. Acetic acid, CH3COOH, is a weak organic acid (K~ at 25oC = 1.8 x 10 -5) which is often used in laboratory tests such as the USEPA Toxicity Characteristic Leaching Procedure to represent organic acids produced by decay of organic m a t t e r in landfills, and 3) Acetic acid, contmuously adjusted to a target pH of 5, to provide a less aggressive comparison to acetic acid at pH 3. The nitric acid solution h a d an initial pH of 3, i.e., a concentration of 1 mmol/L; the acetic acid solutions had initial pHs of 3 and 5, corresponding to concentrations of approximately 56 and 0.0056 mmol/L, respectively. The pH of all three series of experiments was manually monitored and corrected to pH 3 or 5, by addition of more concentrated nitric or acetic acid, as appropriate. This method of pH adjustment had the effect of substantially increasing ionic strength, and particularly nitrate or acetate concentrations, over the course of the experiment. For most specimens, the acid leachant was changed once or twice because of volume mcrease and algal growth, which r e t u r n e d the nitrate or acetate concentrations to their initial values.
554
The depth of complete disintegration of the matrix from the end of each cylinder and acid consumption were measured and recorded over immersion periods of up to 19 months. The precision and accuracy of the measurements was about 0.5 mm.
3. R E S U L T S AND D I S C U S S I O N In practice, the high alkalinity of the cement-based matrices made the target pH of 3 difficult to mamtam, and the average pH measured for both the "pH 3" nitric and acetic acid solutions was actually 3.7, ranging from as high as 10.8 at the begmnmg of the experiment, to lower than 3 immediately after addition of acid. A target pH of 5 was maintained easily, although measured pH values did range between 11.5 at the beginning of the experiment and 3.4, immediately after acid addition. For the sake of simplicity, the acid solutions will be referred to by their target pHs in the following discussion. The results from the acid attack by pH 3 nitric acid, pH 3 acetic acid, and pH 5 acetic acid are shown in Figures l, 2 and 3, respectively, as plots of corroded depth (primary vertical axis) as a function of time (horizontal axis). Each figure shows five separate charts, one for each bmder type, with results plotted for specimens with and without waste. The amount of acid added in order to m a m t a m the target pH has also been plotted as a function of time, with reference to a secondary vertical axis on the right side of each plot. In comparing the plots, it should be noted that the time scale in Figures 1 and 3 ranges from 0 to 600 days, but in Figure 2 it ranges from 0 to 100 days. On the "corroded depth" axes, scales of 0 to 12 ram, or 0 to 120 mm were used, depending on the requirements; the scale for mmoles of acid addition was adjusted as needed to display the data.
3.1. Corrosion by pH 3 Nitric Acid Figure 1 shows that 19 months of immersion in pH 3 nitric acid resulted in corroded depths of less than 2 mm for all the bmders tested, with the exception of the high a l u m m a cement/gypsum system assumed to have formed an ettringite matrix. Addition of plating sludge did not appear to negatively affect the acid corrosion resistance of portland cement products, with or without silica fume. These results are consistent with an observed decrease in solubility of the portland cement and portland cement/silica fume formulations containmg plating sludge in previous acid neutralization capacity experiments on ground products (Stegemann et al., 1994). However, addition of platmg sludge resulted in greater corrosion of the monolithic fly ash/lime product, which also crumbled at the cylinder edges, as well as showing increased corrosion over the exposed face. While strength development in this system was indicative of successful hydration, the waste appears to have changed the hydration products, as was also suggested by the absence of a calcium silicate hydrate (CSH) plateau in the acid neutralization capacity curve found in previous work. After an initial lag period of low corrosion, the high ettrmgite bmder was rapidly destroyed in less than 6 months; replacement of gypsum with mcinerator ash resulted in even more rapid deterioration in only 4 months.
3.2. Corrosion by pH 3 Acetic Acid Because acetic acid is a weak acid, a much higher concentration was required to achieve pH 3, than for the nitric acid. Consequently, immersion in pH 3 acetic acid (Figure 2) was much more aggressive than pH 3 nitric acid to all formulations. The pure portland cement formulation was mitially corroded by the acid, and then turned salmon-coloured and expanded dramatically, crackmg its mould. It is thought that the expansion was caused by reaction of the acetic acid with the free lime from the portland cement to form calcium diacetate. Calcium diacetate, Ca(CH3COO)2.2H20, has a high solubility (I~p ~ 6), but the high concentration of acetate (up to l0 moles/L, because of the continual pH correction of the leachate) was sufficient to force it out of
555 12
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558
solution. An orange colour was also observed in the outer layer of the portland cement formulation suspended in nitric acid. Other authors attributed an orange colour change observed for portland cement-based solidified wastes immersed in acetic acid to oxidized iron (Cheng and Bishop, 1996a). This explanation is consistent with the absence of a colour change in the specimens not contammg portland cement. Addition of silica fume consumed the excess calcium hydroxide generated by the portland cement, but the portland cement/silica fume specimen crumbled suddenly and completely, after exhibiting no corrosion for 7 weeks, suggesting that expansion reactions as well as acid attack may have affected its structural integrity. Calcium depleted from the calcium-rich CSH of this formulation may have reacted with the acetate to cause this effect. The portland cement and portland cement/silica fume specimens containing waste did not suffer any noticeable expansion, and were gradually but quickly corroded. The activated blast furnace slag and fly ash/lime binders had not yet disintegrated when ~he experiment was discontinued after 80 days, but exhibited corrosion depths of 6 and 2 mm, respectively. Expansion did not appear to occur in these binders, and may have been averted by the lack of easily soluble calcium, as the Ca/Si ratio of these systems was 0.5. Addition of waste to the fly ash/lime system resulted in its complete disintegration within a week, showing again that the modified hydration products in this formulation had a poor acid resistance. The high alumina cement formulations also disintegrated within a week, due to the high solubility of ettrmgite and calcium chloroaluminate at low pH. 3.3 Corrosion by pH 5 Acetic Acid Comparison of Figures 1 and 3 shows that pH 5 acetic acid was also more aggressive than pH 3 nitric acid, to all but the fly ash/lime/waste formulation and the high alumina cement formulations, with corroded depths between 6 and 12 mm measured after 16 months of immersion. It is possible that the deterioration of specimens in this experiment was due to a combination of acid attack and expansion. Because of the lower concentration of acetic acid, a different mechanism occurred than in the pH 3 acetic acid experiments; rather than causmg massive expansion and failure, formation of smaller amounts of calcium diacetate lead to cracking at a smaller scale which facilitated acid attack. Specimens with less soluble calcium, i.e., with a lower Ca/Si ratio binder (activated blast furnace slag and fly ash/lime) and/or contaming waste, were not as subject to expansion and therefore deteriorated less. The high alumma cement specimens containing incinerator ash immersed in pH 5 acetic acid were the only ones of this bmder type which had not completely corroded by the end of the test. A sharp discontmuity is apparent in the corrosion and acid addition curves after 22 weeks, which may be indicative of formation of a protective layer of aluminum triacetate and/or calcium diacetate by reaction of the acetic acid with alummum and calcium liberated by ettrmgite and calcium chloroaluminate dissolution. This protective layer may not have had an opportunity to form at pH 3, because of the extreme rapidity of dissolution of the specimens. The discontmuity in the corrosion and acid addition curves for the r e m a m m g specimens after 25 weeks corresponds to a replacement of the leachant, and is unlikely to be related to calcium diacetate precipitation, as the amount of acetate added was insufficient to exceed solubility limits. For the same reason, formation of a calcium diacetate precipitate under natural conditions is unlikely. 3.4. G e n e r a l i z e d Factors in Acid Attack In general, the corrosion plots exhibited three stages: (1) a lag period before (2) acceleration of deterioration, followed by (3) a decrease in rate of deterioration. It is postulated that the lag period is attributable to depletion of easily soluble alkalinity from the surface layer of the specimens, which resulted in consumption of acid, but left a structurally stable
559
matrix. In the case of the portland cement formulations the soluble alkalinity was initially likely to have been mainly sodium, potassium and free lime; later on, once the pH of the corroding surface layer dropped below 12.5, and for the lower Ca/Si ratio products and those containing waste, it is expected that decalcification of calcium silicate hydrate (CSH) took place. For the CSH-based matrices, visible deterioration of the matrix structure would not be expected until the pH of the corroding layer decreased below 9.9, where CSH coexists with more soluble silica gel (see review in Stegemann et al., 1994). The lag period was longer for the lower Ca/Si ratio solidified products because the higher Ca/Si ratio products contained free lime. Free lime is more soluble than CSH, and leaves a higher porosity matrix as it dissolves, increasing exposure to acid attack. Cheng and Bishop (1996b) found a porosity of 0.8 in the decalcified layer of portland cement-based sohdified wastes. For the high alumina cement formulations, the lag period may have been caused by dissolution of alkalis and free lfine; deterioriation of the matrix structure would be anticipated to start when the pH of the surface layer dropped below approximately 11 (see review in Stegemann et al., 1994). The period of accelerated corrosion was linear as a function of time. In the case of the CSH-based matrices, it resulted from increased dissolution of the silica-rich CSH, alummosilicates and silica gel r e m a m m g after decalcification of the CSH, as the pH of the corroded layer dropped from 9 to below 5. Agam, a higher Ca/Si ratio increased the vulnerability to acid attack in all three series of experiments. The corrosion rate was highest for the high alumina cement matrices, in which ettrmgite or calcium chloroaluminate dissolved rapidly. In the portland cement and portland cement/silica fume systems, with and without waste, and in the activated blast furnace slag and fly ash/lime systems, it is postulated that a protective surface layer consistmg mamly of silica gel, and also containmg a l u m m a and iron compounds, gradually developed over time. F u r t h e r leaching of alkahs and calcium and inward movement of acid to the corrosion front then became controlled by diffusion through this layer (Pavlik, 1994). The benefit of the silica gel protective layer was most evident in the relatively low corrosion of the activated blast furnace slag and fly ash/lime specimens upon immersion in pH 3 acetic acid. The low free lime content of these matrices resulted in a particularly dense silica gel protective layer. This layer has an additional benefit in sohdified wastes, in that the silica gel can adsorb heavy metal contammants at pH values as low as 5 (Schmdler et al., 1976). A high proportion of free lime, or other soluble calcium, such as ettringite, can result in secondary precipitation of calcium salts, which may contribute to formation of a protective layer but can also lead to expansion, as was seen in the acetic acid experiments. The corrosion rate did not decrease over time for any of the high alumina cement specimens, except the one made with incinerator ash and immersed in pH 5 acetic acid discussed m 3.3, as dissolution of the ettrmgite and/or calcium chloroaluminate matrices in nitric acid was complete, leaving no residue nor depositing a protective layer on the surface of the monohth. Other authors have found that, at the same concentration, mineral acids are more corrosive to cements than weak acids (Pavlik, 1994 and Bayoux et al., 1990). Such was not the case in this series of experiments, because continued addition of acetate caused precipitation and expansion reactions which lead to increased matrix deterioration in the weak acid.
3.5. Acid C o n s u m p t i o n of Different Matrices A straight line relationship between depth of corrosion and acid addition would be expected for matrix deterioration by chemical dissolution. Plots of the depth of corrosion as a function of the amount of acid added are not shown here, but a straight line passing through the origin was fitted to the data for each specimen by the method of least squares, to determine slopes of mm of corroded depth per mmole of acid added. For the pH 3 nitric acid and pH 5 acetic acid experiments, most
560
correlation coefficients (R) were found to be between 0.8 and 1, indicating that the corrosion depth and acid addition data were highly correlated, except when the corroded depth was too small to be accurately measured (e.g., activated blast furnace slag corroded by nitric acid), or when deterioration mechanisms other than dissolution were a factor (e.g., fly ash/lime/waste formulation in nitric acid). For the pH 3 acetic acid data, correlation coefficients higher than 0.8 were determined only for the blast furnace slag and fly/ash lime systems. These formulations were the only ones which maintained their structural integrity in the pH 3 acetic acid; under aggressive attack by acetic acid other deterioration mechanisms, e.g., expansion, cracking and crumblmg, came mto play for the other formulations. For comparison with acid neutralization capacities measured for ground samples (Stegemann et al., 1994), the mm of corroded depth per mmole of acid added were converted to mmoles of acid added per gram of dry cement using the surface area exposed to the acid, and the solidified product densities and cement contents. Table 2 summarizes the resulting values for each specimen. The approximate amount of acid per gram of dry cement in the formulation, which was required to achieve complete matrix destruction, was read from the acid neutralization capacity (ANC) curves generated previously. It was assumed that the all matrices were completely destroyed at pH 5. Table 2. Amount of acid required for complete matrix destruction Batch mmol of acid/ lW 2 2w 3 4 4w 5 5w of dry cement 1 62 16 47 25 17 7.0 7.0 8.9 pH 3 HNO3 19 160 ** 55 360 800 41 39 20 pH 3 CH3COOH ...... 1.6.0 64 9.0 22 12 12 43 21 I1 pH 5 CH3COOH 21 23 16 20 13 5.0 ll 8.0 6.4 .~NC to pH 5* > 17 * based on Stegemann et al., 1994 ** slope not calculable because sudden crumbling followed a period of no corrosion Values for which the correlation between corrosion depth and acid addition was poor (i.e., R<0.8) are shaded. .
_
IIIII
While the precision and accuracy of the slopes is not high, because of low precision of the depth measurement relative to the depth of corrosion, and poor correlation between the data for some specimens, it is possible to make some general observations" Most obviously, the amount of acetic acid consumed at pH 3 was far greater than the acid neutralization capacity to pH 5 found previously. This is a consequence of the method of acid replenishment in this experiment, which resulted in buildup of large acetate concentrations, creatmg a buffer system which also consumed acid. This effect is not noticeable for the pH 3 acetic acid experiments, which had lower acetate concentrations, except for the high ettrmgite batches. For the portland cement/silica fume formulation in pH 5 acetic acid it appears that significantly less acid was required for monolithic matrix dissolution than predicted by the acid neutralization capacity to pH 5. This may be explained by matrix deterioration through mechanisms other than dissolution, as discussed earlier. For most specimens, however, the amount of acid required for dissolution of the monolithic matrix was similar to or greater than that predicted by the acid neutralization capacity of ground solidified products to pH 5. Lower Ca/Si ratio monolithic products seem to have a higher acid resistance than might be expected solely based on the ANC to pH 5, which may be because of formation of a denser, less soluble silica gel at low pH. Unfortunately, there is insufficient data to statistically confirm this fmdmg. It also appears that the portland cement and portland cement/silica fume batches with waste consumed more acid before dissolving than those without waste, and that this effect was greater for the monolithic specimens than it was for the ground solidified product. It seems that the waste
561
acted as an aggregate in the CSH matrix which decreased physical deterioration. phenomenon was also observed for the fly ash/lime formulation in acetic acid.
This
4. CONCLUSIONS The stages leading to deterioration of monolithic cement-based binders and sohdified products under acidic conditions can be conceptuahzed as follows: 1) neutralization of soluble alkahs such as sodium and potassium from the surface, 2) dissolution of soluble alkalinity, such as free lime and ettrmgite from the surface, 3) formation of a protective surface layer of decalcified CSH and silica gel containing iron and aluminum silicates, and 4) diffusion control of contmued neutrahzation of the alkalinity from the interior of the monolith through the surface layer. The decalcified CSH and silica gel surface layer is capable of adsorbmg metals at pH values down to 5. Because stages 1 and 2 occur relatively rapidly, monohthic products which do not contain CSH (e.g., high ettrmgite formulations) are at a serious disadvantage because they simply contmue to dissolve rapidly, do not form a protective surface layer of silica gel and therefore also have no capacity to retain metals at lower pH values. Formation of secondary precipitates by reaction of matrix dissolution products (e.g. calcium) with the the acid may be a comphcating factor which can result in enhancement of the protective layer, or accelerated deterioriation. Although information about acid neutrahzation capacity is a great help in interpreting data from acid attack experiments on monohthic binders and solidified products, it is not by itself a good predictor of the acid resistance of monohthic products. Whereas formulations contammg portland cement generally have the highest acid neutralization capacity, CSH-based monohthic products with a lower Ca/Si ratio (e.g., activated blast furnace slag and fly ash/lime) exhibit greater acid resistance over time. The free lime in portland cement provides a high buffering capacity but is rapidly dissolved in acidic solutions, leaving a vulnerable porous silica-rich layer, whereas the CSH of lower Ca/Si ratio rapidly forms a dense silica-rich protective layer. The addition of plating sludge to the binder systems was shown to change acid resistance. The acid resistance of the monolithic portland cement formulations containing waste (with and without silica fume) increased, compared to specimens not containing waste, but also compared to the ground material. The plating sludge appears to have acted as an aggregate which improved the durability of the matrix structure. The plating sludge interfered with formation of CSH in the activated blast furnace slag formulation, to the extent of inhibiting strength development, so that the product could not be tested as a monohth.
5. R E C O M M E N D A T I O N S The focus on acid neutralization capacity of ground sohdified products fails to recognize the contribution of the monolithic structure of a solidified product to immobilization of contammants. For maximum acid resistance, design of sohdification formulations should ensure formation of low Ca/Si ratio CSH. Development of a standard test for measurmg acid resistance of monohthic specimens should focus on minimizing potential comphcating factors, such as secondary precipitation of reaction products of the acid and matrix. Ideally, automatic rather than manual pH adjustment to pH 5 with nitric acid is recommended for assessment of acid resistance of monohthic specimens. A less expensive alternative is to use pH 5 acetic acid with periodic leachant renewal, rather than pH adjustment.
562
6. R E F E R E N C E S
Bayoux, J.P., Letourneux, J.P., Marcdargent, S., Verschaeve, M. (1990), "Acidic Corrosion of High Alumma Cement", Proceedings of the International Symposium on Calcium Aluminate Cements, Queen Mary and Westfield College, London. Cheng, K . Y . and Bishop, P.L. (1996a), "Morphology and pH Changes in Leached Sohdified/Stabilized Waste Forms", Stabilization and Solidification of Hazardous, Radioactive, and Mixed Wastes: 3~d Volume, ASTM STP 1240, T. Michael Gilliam and Carlton C. Wiles, Eds., American Society for Testing and Materials. Cheng, K.Y. and Bishop, P.L. (1996b), "Property Changes of Cement-based Waste Forms Durmg Leaching", Stabilization and Solidification of Hazardous, Radioactive, and Mixed Wastes: 3~d Volume, ASTM STP 1240, T. Michael Gilliam and Carlton C. Wiles, Eds., American Society for Testing and Materials. Hmsenveld, M. and Bishop, P.L. (1996), "Use of the Shrinking Core/Exposure Model to Describe the Leachability from Cement Stabilized Wastes", Stabilization and Solidification of Hazardous, Radioactive, and Mixed Wastes: 3~d Volume, ASTM STP 1240, T. Michael Gilliam and Carlton C. Wiles, Eds., American Society for Testing and Materials. Pavlik, V. (1994), "Corrosion of Hardened Cement Paste by Acetic and Nitric Acids - Part I: Calculation of Corrosion Depth", Cement and Concrete Research, Vol. 24, No.3, pp. 551-562. Schmdler, P.W., Fuerst, B., Dick, R., and Wolf, P. (1976), "Ligand Properties of Surface Silanol Groups; I. Surface Complex Formation with Fe a§ Cu2§ Cd 2§ and Pb 2§ Journal of Colloid and Interface Science, Vol. 55, No. 2. Stegemann, J.A., and Cote, P.L. (1991), Investigation of Test Methods for Solidified Waste Evaluation,- Appendix B: Test Methods for Solidified Waste Evaluation, Environment Canada Unpublished Manuscript Series, Document TS- 15. Stegemann, J.A., Caldwell, R.J., and Shi, C. (1996), "Laboratory, Regulatory, and Field Leaching of Sohdified Wastes", Proceedings of the International Conference on Incineration and Thermal Technologies, Savannah, Georgia. Stegemann, J.A., Shi, C. and Caldwell, R.J. (1994), "Response of Various Sohdification Systems to Acid Addition", Presented at WASCON'94, Maastricht, The Netherlands. United States Environmental Protection Agency (1993), Technical Resource Document: Stabilization~Solidification and its Application to Waste Materials, EPA/530/R-93/012, Office of Research and Development, Washington DC 20460. Wastewater Technology Centre (1991), Proposed Evaluation Protocol for Cement-Based Solidified Wastes, Environment Canada Pubhcation EPS 3/HA/9, Ottawa.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
563
R e s e a r c h and S t a n d a r d i z a t i o n P r o g r a m m e for D e t e r m i n a t i o n of L e a c h i n g B e h a v i o u r of C o n s t r u c t i o n Materials and W a s t e s in the Netherlands. R.J.J. van H e i j n i n g e n I and H.A. van der Sloot 2
1. B a c k g r o u n d . In 1980 a study group ('SOSUV') was formed in the Netherlands with the aim to develop a set of standard laboratory tests for the determination of the leaching behaviour of solid combustion residues. The SOSUV group drafted test procedures to simulate leaching under different field conditions for both granular waste materials and monolithic waste matter. After the early work of the SOSUV group a substantial research programme supported by the government was launched to test a great variety of other bulk waste materials with potential for utilization ('Mammoth-project'). As a reference also natural construction materials (e.g. cement, sand, gravel, concrete) were included within the project. During the course of the Mammoth-project two national standards were published: - NVN 2508 'Determination of leaching characteristics of combustion wastes' - Draft-NVN 5432 'Determination of the maximum leachable quantity and the emission of potentially hazardous components from products in which combustion wastes are incorporated'. In recent years a many other granular waste materials and waste-derived products have been studied. It was observed that the leaching behaviour of a great many substances could be interpreted on the same basis as used for the characterization of combustion residues. The standard methods NVN 2508 and NVN 5432 for combustion residues apparantly also apply for many other waste materials. In 1990 a special task group was asked to extend the scope and applicability of NVN 2508 and NVN 5432 to comparable materials and to develop additional standards necessary for detailed characterization of other construction materials and wastes. With support of the Netherlands Standardization Institute (NNI) and the Netherlands Agency for Energy and the Environment (NOVEM) an ambitious multi-year project programme was started [TSP 90]. The leading role in the execution of this programme is performed by a newly established standardization committee (NNI-Committee 390 011).
1 Eindhoven University of Technology 2
Netherlands Energy Research Foundation ECN
564 Apart from the national research and standardization programme a variety of field s t u d i e s was launched by several interested parties in order to obtain direct leaching information from specific road construction projects, ground works, and waste deposits. In part these field experiments were also ment to link the measured contaminants diffused in the soils around the constructions with predictions based on the results of standardized laboratory experiments with the applied materials.
2.
Scope
and contents
of the Research
and Standardization
Programme.
For the environmental characterization of construction materials and wastes several tools are needed: - a classification system based on the most important physical and chemical characteristics; standardized test methods for the determination of the chemical composition and the leaching properties; guidelines for the interpretation of test results in relation to the proposed end use; legal acceptability criteria and procedures 9 In order to make such tools available for all relevant materials a step by step standardization programme was prepared along the following lines: the programme must result in a complete set of standards and guidelines based on a sound scientific understanding of potential environmental impacts; - the programme must be supported by all relevant parties from government, building industry, and advisory bodies; for a priority list of constructiom materials and waste products the most important standards need to be available within a period of 2-3 years (see appendix); all standard procedures have to be screened with respect to practical consequences and costs. With the above mentioned guidelines in mind a Standardization Programme was started with the intention to prepare as a first step all readily obtainable standards on the basis of existing knowledge. The standardization programme consists of 15 projects on the subjects: Sampling protocols and techniques 9 Protocols for sample storage and pre-treatment 9 Destruction and chemical analysis of the solid phase Leach tests for inorganic and organic compounds Analysis of the leachate Influence of reducing conditions 9 Influence of particle size distribution 9 Geometry aspects in leaching of monolithic matter
565 Reference materials Interpretation of test results The interrelations between the projects (indicated as 'V-projects') are shown in Fig. 1" Fig.l"
OVERVIEW
OF PROJECTS
FOR
DEVELOPMENT
OF STANDARD
TESTS
FOR
WASTE
CHARACTERIZATION
Vl - NEN 7310
V!
SAMPLING
PROTOCOLS
V2 - NEN 7320 PROTOCOLS
FOR SAMPLE
AND SAMPLE ANALYSIS OF I N O R O A N I C S V3 o NEN
I
7320
STORAGE
PREPARATION ANALYSIS OF ORGANICS V I 0 - NEN 7330
Destruction & Analysis
V16
Destruction & Analysis VI6
Reference Materials
Reference Materials v LEACHING TESTS FOR ORGANIC COMPOUNDS
LEACHING TESTS FOR INORGANIC COMPOUNDS
VII - NEN 7342
V7 - NEN 7341 v4/5
Influence o f particle-size V6 - NEN 7348
Influence o f Redox conditions
Influence of
for leaching
| Leach test for [Granular Materials [
76-N--~ 7-~
Influence of Redox conditions
Quick
testa
Surface Area for Monolith v15
Quick
tests
VI2 - NEN 7346
v8 - NEN 7345 V13
v14
_partt___5~le'sl_____'e ~12- NEN7344
V8 - NEN 7343
Leachtest for Granular Materials
Availability
V4/5
Availability for leaching
V13 Surface Area for Monolith
Leachtest for Monolithic Materials
R e f e r e n c e Materials for L e a c h i n g
Leach test for Monolithic MateriaLs
R e f e r e n c e Materials for L e a c h i n g ,4/
~/
>1 OVERALLASSESSME~L l-
[ RES~RCUPROJECTS [
L a b - Field Relations
In the same time a coherent Research Programme was started in order to solve practical and scientific problems with respect to possible future standards on material characterization. This part of the programme consists of 9 projects on the subjects: 9 Translation of laboratory results to practice 9 Durability of products under field conditions 9 Modelling of long-term leaching behaviour 9 Interaction of leached contaminants with soils Enhanced leaching of non-mobile components by dissolved organic matter and particulates ('facilitated transport') 9 Selection of extracts for bioassessment
566
The coherence of the research projects (indicated as 'O-projects') is shown in Fig. 2. Fig.2: O V E R V I E W OF R E S E A R C H PROJECTS IN RELATION TO RELEASE FROM W A S T E AND W A S T E - D E R I V E D PRODUCTS
O1 Translation factors lab to practice (Temperat., rainfall)
03
07
O2 Durability of monolith under field conditions 08 " Facilitated transport "
Modelling soil interaction and transport
Verification of monolith leach models- 3D 09
Selection of extracts for bioassays 04105 Integrated model for assessment of release with time
06 Verification of models in historical sites
RELATION B E T W E E N LAB TESTS AND PRACTICE OF UTILIZATION AND DISPOSAL
The main results of the entire Research and Standardization Programme are summarized in the Handbook on Leaching [HBU 94][HBU 97] prepared under supervision of the Netherlands Foundation for Legislatation and Research in Ground Work, Water, and Road Constructions and Traffic Control Techniques (C.R.O.W). In part I of this handbook general guidelines for the characterization of leaching behaviour are presented, together with a survey of the developed standard test procedures. In part II of the handbook for some 50 building materials and waste materials the physical properties, leaching characteristics, and most relevant aspects from an environmental point of view are highlighted.
3. Scope and contents of the individual field experiments. The third series of activities on leaching concerns individual field studies launched by several interested parties in order to obtain direct leaching information from specific road and building construction projects, ground works, or waste deposits. The field experiments
567 were primarily ment to check wether or not contamination would occur due to leaching of the used materials. Some field tests were focussed on this primary goal but other experiments were more comprehensive and focussed also on the understanding of the leaching mechanisms and the chemical interactions with the surrounding soils. From the results of the latter field studies the relation between the observed leaching in the real world and the results of standardized laboratory tests on the materials used may be derived. The results of the 6 most interesting field studies are summarized in part III of the Handbook on Leaching [HBU 96] prepared under supervision of C.R.O.W. It concerns the following cases: a Use of granular material (MSWI bottom ash) in an embankment to support a motorway b Road foundation with compacted granular materials (sand, pulverized coal bottom ash, concrete granulate) c Road foundation with cement-bound granular materials (pulverized coal fly ash) d Concrete sluice construction in and above surface water e Riverbank protection with granular material (steel slag) f Road surface with paving-stones. Two of these cases are presented in more detail elsewhere in this Proceedings (case c: by Schreurs and Van der Sloot; case e: by Sonneveldt e.a).
4. Results obtained to date.
To date the Research and Standardization Programme has resulted in a series of new national standards for leaching tests with respect to inorganic compounds (NEN 7340, 7341, 7343, 7345, and 7349) as well as organic compounds (NVN 7344 and 7350). Also protocols for sampling (NEN 7300 and NVN 7301, 7302, and 7303), sample storage and pre-treatment (NEN 7310 and NVN 7311, 7312, and 7313), and standards for the destruction and analysis of both inorganic compounds (NEN 7320 and NVN 7321, 7322, 7323, and 7324) and of organic compounds (NVN 7330) were developed. Additional tests for the assessment of the influence of specific material properties or conditions such as pH, redox, and dissolved organic matter are well underway [NOR 95]. For validation of the new leaching standards a comprehensive round robin test programme has been executed. Also the standards for the destruction and analysis of both inorganic and organic compounds have been validated in a round robin test for a variety of important building materials. For the sampling and pre-treatment protocols a validation programme is in preparation.
568 The research programme has given first results with respect to the derivation of factors for the 'translation of laboratory test results to practice' 9 Also basic insight has been obtained in the factors that complicate the material characterization such as durability, facilitated transport in the surrounding soil, and biodegradation. Together with the information obtained from the individual field experiments (see section 3) a fair degree of quantitative understanding of the leaching and diffusion mechanisms has been reached for most currently used building materials in ground works and road or embankment constructions. The analysis of the 6 cases mentioned in part III of the Handbook on Leaching has learned that to date it is possible: 9 to generate basic leaching insight in the majority of chemically stable materials which is suitable to assess (at least qualitatively) the environmetal consequences of the commonly used applications with these materials; 9 to assess quantitatively the environmental consequences in a number of specific field situations by applying semi-empirical 'translation factors' to the results of standard laboratory tests; For materials applied under conditions where the infiltration of water is more or less inhibited (highly unsaturated conditions) and for materials that change strongly under the infuence of air and water (e.g. materials with reducing properties or biologically active materials) it is not possible to predict the leaching behaviour on the basis of the present standard tests. For these circumstances only with the help of additional characterization tests and the application of suitable calculation models a more quantitative assessment of the environmetal consequences is possible. In both areas progress is being made through projects presently underway. A recent overview of the results obtained to date and the progress made in Dutch legislation with respect to leaching was presented during the symposium 'Secundaire materialen primair toepassen'. The proceedings of this seminar also contains a comprehensive literature survey [SYM 96].
5. References
[TSP 90]
Taakstellend Plan ter ondersteuning van normcommissie 390 011 'Uitloogkarakterisering van bouwmaterialen en afvalstoffen', NOVEM B.V., Utrecht (1990).
[NOR 95] Normcommissie 390 011 'Uitloogkarakterisering van bouwmaterialen en afvalstoffen', Jaarverslag 1994, doc. 390 011/95-08, NNI, Delft (1995).
569 [HBU 94] Uitlogen op karakter, Handboek Uitloogkarakterisering, I Testmethoden en II Materialen, C.R.O.W, Ede (November 1994). [HBU 96] Uitlogen op karakter, Handboek Uitloogkarakterisering, III Praktijk, C.R.O.W, Ede (May 1996). [HBU 97] Uitlogen op karakter, Handboek Uitloogkarakterisering, Aanvulling op delen I en II, C.R.O.W, Ede (January 1997). [SYM 96] Symposium 'Secundaire materialen primair toepassen', gebruik van uitloogproeven in de praktijk, C.R.O.W, Ede (8th May 1996).
570
Appendix List of materials treated within the Netherlands Standardization Programme: Classification (grain size)
Species
Priority list
Other materials
Powders and sludges Category a
Conventional materials
Cement Lime
Category b
Fly ashes
Pulverized coal fly ash MSWI fly ash
Category c
Gypsum
Phospho gypsum FGD-gypsum
Category d
Sludges
Sewage treatment sludge Waste water treatment sludge Drinking water purification sludge River sediment Dredge material
FBC fly ash
Fine-granular materials (0-4 mnl Category a
Soils
Clay Sand Peat Decontaminated soils
Category b
Fine-granular
Construction debris sand Pulverized coal bottom ash Blast furnace slag Coal gasification slag
Course-granular materials (0-60
FBC bottom ash Sand blasting waste
mm)
Category a
Natural materials
Gravel Lava
Crushed rock
Category b
Coursegranular
PC fly ash aggregates Concrete granulate Coal mine waste Asphalt concrete granulate
Masonry granulate Construction debris
Category c
Slags
Blast furnace slag Phospho slag LD steel slag MSWI bottom ash
Metallurgical slags
Stablized wastes Bituminous mixes
Monolithic matter and products Category a
Hydraulic products
Concrete Concrete products Asphalt concrete
Category b
Autoclaph products
Calcium silicate bricks Light-weight concrete
Category c
Baked products
Masonry bricks Paving-stones
Waste derived masonry bricks Waste derived ceramic products
Category d
Monolithic slags
Monolithic blast furnace slag phospho slag LD steel slag
Monolithic metallurgical slags
Goumans/Senderffvander Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
UTILIZATION
571
OF FLUE GAS DESULPHURIZATION
IN T H E C E L L U L A R C O N C R E T E
BY-PRODUCTS
TECHNOLOGY
Brylicki W., Lagosz A. University of Mining and Metallurgy, Department of Material Science and Ceramics Krak6w, Poland
ABSTRACT In the fly gas desulfurization processes by use of other methods than the wet lime one, or during the fluidized bed combustion, the waste materials, being the mixture of fly ash, desulfurization products and residual sorbent are usually produced. Their complex and very otten variable chemical and mineralogical composition is a factor substantially limiting the possibility of their reuse. In this work the chemical and mineral composition of wastes produced in different desulfurization processes has been determined. These wastes were used as a natural gypsum replacement in the so-called PGS cellular concrete technology (,,foamed-gaseous-silicate concrete"). The growth of fresh mixture, phase composition of fresh and hardened concrete, morphology of hydration products and practical properties of hardened cellular concrete were examined. As it results from the studies, the waste material from desulfurization process can be used as a component in cellular concrete technology. The fly ash can play a role of finegrained aggregate or cementitious material.
572 1. INTRODUCTION The reduction of sulfur and nitrogen oxides emission to atmosphere is nowadays one of the most important tasks in the natural environment protection. The desulfurization and denitrification programs can be realised effectively when the efforts in the field of coal desulfurization, flue gas desulfurization and denitrification and waste disposal are co-ordinated. Among the over 100 methods of flue gas desulfurization invented until now, only some of them have been put into practice. The wet lime method should be pointed out as the first one. In some countries, where the desulfurization operation has been advanced early (USA, Japan, Germany), more than 80% of desulfurization installations work following this wet lime method, in which the de-dusted gases are subjected to the purification. The pure gypsum dihydrate, produced as a desulfurization product, can subsequently substitute the natural gypsum raw material in building materials production and in other technologies. When the desulfurization waste material is a multicomponent mixture if fly ash, desulfurization products, unreacted sorbent and other compounds formed on fuel combustion, the other technologies should be taken into account. The authors are of the opinion that the Polish cellular concrete technologies PGS or UNIPOL (fly ash or combined variant) can be useful for this purpose. The experiments aimed with the determination of different flue gas desulfurization byproducts usability in cellular concrete technology by PGS method. The calcium sulfates and sulfites from waste materiel controlled the growth and setting of concrete mixture, and consequently the strength development of cellular concrete. The gypsum component, which is commonly used in the first step of cellular concrete technology, acts as a calcium aluminate and aluminoferrite hydration retarder. It reacts quickly with these phases, originating from fly ashes and cement, as well as with the hydrated calcium aluminates produced in reaction between AI powder and Ca 2§ ions from the solution. On the surface of AI powder particles the ettringite and hydrated calcium monosulfate are thus formed and hinder the hydrogen release. The AFt and AFm phases have no special influence on the properties of hardened concrete, because they fill the pores on growth and early maturing. The proper growth of concrete mixture is disturbed at low gypsum component content. The quick setting is thus observed and the mixture does not attain the fight structure [ 1]. At the presence of calcium sulfites C3A-CaSO3-12H20 is formed instead of ettringite [2]. This compound oxidates to ettringite [2,3]. 2. EXPERIMENTAL 2.1. Materials
2.1.1. Fly ash ,,Skawina" The fly ash was produced from the black coal in the ,,Skawina" power station. It complies with the standard requirements for the raw materials used in cellular concrete production. The chemical composition of fly ash is given in Table 1. 9 specific gravity 2.12 g/cm 3 9 specific surface 3050 cm2/g (Blaine) 9 water demand 27%
573
2.1.2. Ground burnt lime ,,Tarnrw Opolski". The ground burnt lime complying with the standard requirements for the raw materials used in cellular concrete production was used. The properties of lime are as follows: 9 active CaO + MgO content 87.1% 9 time of slaking 28 min 9 temperature of slaking 68~ 9 total CaO content 93.5% 9 total MgO content 3.4% 9 loss on ignition 3.0% 2.1.3. Flue gas desulfurization by-products. The following FGD by-products were used: 9 from wet bialkaline method (,,Chrzanrw" thermal power station) 9 from semi-dry ,,dry scrubbing" process (,,Sosnowiec" thermal power station) 9 from fluidized bed combustion with desulfurization, both pressure installation (PFBC Canada) and atmospheric air circulation installation (FBC Canada). The natural Polish gypsum raw material was also used as reference. In Table 2 the chemical composition of materials is shown (including free CaO). The chemical composition data for particular by-products show the significant differences between the sulfate components and free CaO contents. One should also notice the differences in sulfite contents, determined as SO2. The phase composition of FGD by-products was studied by XRD and DTA-TG methods. The results are shown in Table 3. In some cases the products of desulfurization reaction was not identified because of poor crystalline form or low content. Therefore the presence of particular phases was deduced rather from chemical analysis data and DTA-TG curves (they are market by ,,*" in Table 3). The sulfate and sulfite contents were calculated from the chemical composition. The results of this evaluation are as follows: 9 35.9% CaSOa-2H20 and 15.1% CaSO3-0.5H20 for bialkaline process, 9 15.5% CaSO4-2H20 and 14.1% CaSO3-0.5H/O for dry scrubbing process, 9 26.8% CaSO4 and 0.8% CaSO3 for pressure FBC, 9 30.1% CaSO4and 0.9% CaSO3 for FBC, 9 24.0% CaSO4 for dust from FBC installation.
2.2. Examination of pastes prepared with flue gas desulfurization products. The FGD by-products were used to produce the cementitious materials used in cellular concrete technology by PGS method. The composition of particular binders mixtures is given in Table 4. The calculations based on assumed SOs content, the same as in reference batch with natural gypsum component. The sulfites were calculated as equivalent to sulfates. The dry components were ground to the Blaine specific surface of ca. 4100 cm2/g. The samples were mixed with water at water to solid ratio 0.35 and atter preliminary curing at 65~ for 1 hour autoclaved at 190~ atm for 12 hours cycle. The autoclaved samples were examined by XRD and SEM. The chemical composition of these samples was very similar and the two main mineral components - hydration products were detected by XRD: tobermorite Cas[Si6018H2].4H20and hydrogarnet CasA12[SiOa](OH)8. Some amount of calcite and 13-quartz, being the unreacted residue, is also visible. The XRD patterns exhibit
574 some differences of peak intensities which can reflect the different contents and degree of crystallinity for particular phases. This was confirmed by SEM observations. The samples, despite of the similarity of phase composition, reveal the presence of different tobermorite and hydrogranet forms. The degree of crystallinity may be affected by the presence of sulfites, which promote the poorly crystallized, defected structures of hydration products. One can conclude that the microstructure of autoclaved pastes prepared with desulfurization byproducts depends on the composition of these materials. This fact may be of significant importance in concrete technology. The average microstructures of autoclaved pastes are shown in Fot. 1-6. 2.3. Evaluation of concrete mixture growth in the presence of desulfurization products.
2.3.1. Examination of hydrogen release. This experiment was carried out basing on the test of aluminium powder activity. The effect of different desulfurization by-products on the rate of hydrogen release the fresh concrete mixture was measured and the results are given in Table 5 and 6. As it can be seen in Table 5 and 6 the hydrogen release depends on the composition of concrete mixture, particularly on the content of sulfites. At high sulfite content, as in the sample with desulfurization by-product II from semi-dry process, the reaction of AI powder with calcium hydroxide is strongly hindered. The experiment has been carried out for a long time in this case, with and without agitation of mixture. This is of interest that the hydrogen release is not stopped and agitation improves it significantly. 2.3.2. Examination of fresh concrete mixtures. The concrete mixtures were produced following the proportions taken from PREVAR (concrete prefabrication plant in Skawina near Krakrw in Poland): 9 dry components for 1 m s of concrete - 620 kg, including 312 kg of cementitious material and 308 kg of fly ash ,,Skawina", 9 water- 310 1 ( w / s - 0.5), 9 A1 powder- 380 g, 9 surfactant as 38% solution of Sulfapol - 620 g. The procedure of concrete mix production was as follows: Firstly the surfactant was poured to the water and, after agitation, the A1 powder. Subsequently, the other components (cementitious material and fly ash) were added to this suspension and homogenized (3 min and 15 min in the case of the batch with by-product II, because of retarded hydrogen release). The mixture was poured into the moulds (~ 80 mm, h-80 mm). The concrete mixtures were cured at 65~ for 1 hour and subsequently autoclaved. The process was not disturbed at all. The visual examination of samples gave positive results (only the samples with SII showed slow growth). The concretes were enough strong to remove the spews and to keep cutting. The spews were examined by XRD, after rinsing with acetone. The similar phase composition was found with such hydrates as calcium hydroxide Ca(OH)z, ettringite 3CaO.AIzOs.3CaSO4.32HzO, in some cases traces of calcium carboaluminate hydrate 3CaO-AIzO3.CaCOs-121-120 and hydrated calcium aluminate C4AH13. The cellular concrete samples were autoclaved in PREVAIL and the elements thus produced did not exhibit any destructive effects. Because of small amount of materials in this
575 tests, the full examination of concretes was not possible. However, one could find a higher apparent density (5-20%) then for control sample as well as similar compressive strength and capillary suction. Simultaneously, in the opinion of authors, the concrete mixture design procedure was in this case too simplified. For industrial implementation the concrete mixture composition should be further optimized. In Table 7 the results of standard tests, according to Polish standard PN-89/B-06258, for the cellular concrete with by-product ,,I" from bialkaline process are shown. 3. CONCLUSIONS. 9 The waste materials from different desulfurization installations show the presence of different products formed on desulfurization, being the compounds of SO2 and SO3. Depending on the process these products are: gypsum dihydrate, gypsum hemihydrate or anhydrite with more or less calcium sulfite or calcium sulfite hemihydrate. 9 At high amount of sulfites in waste materials used to produce the binder (even more then 50% S in sulfite) the retardation of H2 release takes place. 9 The desulfurization products affect the microstructure of autoclaved pastes and concrete mixtures. The degree of crystallinity and crystal size is strongly influenced. 9 The cellular concrete samples produced with different desulfurization wastes exhibit similar or slightly less advantageous physical properties than the reference sample. This, in the opinion of authors, may be altered by better optimization.
Fot.1. Autoclaved paste prepared from reference sample SO with well developed plate-like tobermorite crystals.
.%,, . .... j! %iiii!ili! ~
~:~:~:~:~::::.::~. ...
~!!!ii!~ iiii!~~,~,~,~!,,~'~!
~
i i i i ij:i!i:i!iliiii!!!iii!iiiiiii:!!il ii~iI....
576
References: 1. Jatymowicz H., Siejko J., Zapotoczna-Sytek G. Autoclaved ceihdar concrete technology. Ed. Arkady, Warszawa 1980 (in Polish). 2. Shiino H., Yasue T., Arai Y. Effect of Calcium Sulfite on Setting of Portland Cement. (;ypsum & Lime 188 17 (1984). 3. Yasue T., Mihara M., Arai Y. Synthesis and Characteristics of New Compound in the System CaO - A1203 - CaSO.~.nH20. (;)psunt & l, ime 173:5 (1981). Fot.2. Autoclaved paste prepared from sample SI with defected plate-like tobermorite crystals.
Fot.3. Autoclaved paste prepared from sample SII. The spherical,porous fly ash partcles and poorly crystallized, defected tobermorite crystals are visible.
577 Table 1. Chemical composition of fly ash.
AI.203
L.O.I.
SiO2
Fe203
[%]
[%]
[%]
[%]
1,23
53,78
8,46
26,66
MgO
CaO
_
[%]
.
[%]
.
1,57
6,42
Na20
SO3
[%]
K20
[%]
[%]
1,72
3,04
0,52
Table 2. Chemical composition of FGD by-products. Component in weight %
Sample Number L.O.I. I II III IV V
26,04 2t,77 14,24 5,08 24,21
S i O 2 A1203 Fe203
CaO c
5,94 8,43 11,16 19,87 12,00
41,94 40,05 43,54 44,23 30,66
2,86 7,54 5,51 5,75 6,00
1,05 1,86 8,42 4,28 11,96
MgO 0,95 2,10 0,34 0,59 0,42
SO3 16,72 8,54 15,75 17,71 14,11
SO2
CaO w 2,01 ,,, 7,39 17,32 15,34 9,93
7,52 7,00 0,88 0,94 -
material from bialkaline process (,,Chrzan6w") II - material from dry.scrubbingprocess (,,Sosnowiec") III - material from pressure FBC installation (,,Canada") IV - material from FBC installation with atmospheric air circulation (,,Canada") V- material from dedustingof FBC instanation gases I -
Table 3. Phase composition of FGD by-products. Flue gas desulfiarization by-product (as in Table 2)
Phase I
CaSO4-2H20, CaSO4, CaSO4-0.5H20 CaSOa-0.5H20,
Ca(O~
+ t*
+*
t
+
+
+
t
+
+
t +
+ +
+
F
+
+
+
+
+ +
t
Fe203
(mullite), 3 CaO. Al203-3 CaS O4.32H20 CaSiOa.CaCO3.CaSO4-151-120 unbumed carbon
V
+
CaO, ~]SiO2
IV
+
CaSO3 CaCO3,
III
+
3A1203-2SIO2
+ +
+
578 Table 4. The compositions of binder mixtures. Cementitious material components SI 14,1 37,7 48,2
FGD by-product ground lime fly ash natural gypsum
Mixture number* [% by weisht ] Sill SIV 19,9 17,7 34,6 35,3 45,5 47,0
SII 19,4 36,6 44,0
SV 23,7 35,6 40,7 -
SO ** 38,0 54,0 8,0
* - S(I-V) - binder with FGD product as in Table 2. ** - SO - referencebinder with natural gypsum
Table 5. The effect of binders on the rate of hydrogen release. Time [min] SI 0,5 8
9
16 20
44 58
Volume of H2 [cm3] at t=20~ (composition of binders as in Table 4) Sill SIV SV 075 0 0 10 28 8 46 63 32
SO 0 6 61
Table 6. The effect of binder SII on the rate of hydrogen release. Time [min]
Volume of H2 cm 3] at t=20~ Test following the procedure Test during continuous as in Table 4 a~itation of mixture 0 0
8
0
0
16 20 40 56
1 4 35 46
3 7 46 66
579 Table 7. The results of standard tests for the cellular concrete with by-product I. Test property
Result
Volume density at dry state Compressive strenght Freeze and thaw resistance
av. 646 kg/m 3 av. 4.30 MPa no fissures, scalls, cracks, weight loss; compressive strength decrease of 14.3% av.4.7 cm 0.6 mm/m.
Capillary suction Shrinkage Heat conductivity coefficient at dry state - 0.146 W/m-K Bock's Method
Fot.4. Autoclaved paste prepared from sample SIII. The well developed hydrogranets are visible.
580 Fot.5. Autoclaved paste prepared from sample SIV with perfect tobermorite twinned forms.
Fot.6. Autoclaved paste prepared from sample SV with tobermorite crystals intergrowths.
Goumans/Sendergvan der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
STATE
581
OF THE ART OF G_yPSUM RECOVERY FOR A SPANISH POWER PLANT
E.Peris Mora; J. Monz6; J. Paya; M.V. Borrachero Dept. Ingenieria de la Construcci6n Universidad Polit~cnica Valencia, Spain
1.-
PREAMBLE
The problems of pollution of water, soil and atmosphere are increasingly considered, by the planning and development policies of most of the states of the world. Nowadays, the technology and management of the environment present a great growth, and receive important amounts of public fund investments from the European Union. In the past, most environmental policies have had a corrective character (who contaminates must pay). Afterwards, the trend was modified and legislation aimed to preventing and avoiding the damage; was the Environmental Impact Assessment system. The different laws limiting atmospheric emissions and the EIA Directive are respectively under these two points of view. The "Fifth Action Program in Environment Matter of the European Union" actually is based on the correction, prevention and on the better participation and planing. Atmospherical pollution was one of the first environmental problems approached by the European Community, when in the 1970s Council Directive (70/220/EEC) on "Approximation between the State members related to the measures to take anti-pollution of the air by gases originating from vehicle motors" was approved followed bY that regulating the vehicle emissions (72/306; 74/290/EEC). Fuel power plants are probably the most thermo-efficient system to convert combustion energy. The size and design of a combustion power plant are created to obtain the maximum rentability of the chemical reactions. Other thermic machines like motor cars, etc., are designed to produce a maximum rate of H.P./Engine-weight. An special case occurs when the objectives are the utilization of a specific fuel: industrial or agricultural wastes, or a particular fuel with low calorific power such as lignites. After the petroleum crisis in the 70's (between 1970s/74 the price of the oil barrel multiplied by six) many coal power plants were constructed. Thus, many power plants to use low power and high sulphured coals were built. Environmental;protection was not then considered as a priority criteria; only several decades later the consequences offered new criteria. Nowadays technologies to correct pollution are available, and the social body has accepted the relevance of some new environmental protection values. Thus, "clean production" should be presumed more expensive than the "dirtyone". In any situation, the decision of the "cleanliness" degree that will be demanded to any technology, will always be a political decision, in response to some social values.
2. - O B J E C T I V E
OF THE
PRESENT
STUDY
This study developed after the requirement from Technical Management of " C E N T R A L T E R M I C A DE ANDORRA", belonging to ENDESA (Spanish Public Enterprise). The program titled "Project Teruel" represents important investments in the factory, with the purpose of reducing the emissions, especially of sulphur oxides, coming from the combustion of coals with high sulfur contents.
582
The construction of the new desulphuration plant has already begun. The project anticipates a reduction of current emissions by 90%. The technology consists in the utilization of 700.000 tons of limestone (CaCO3) per year with the production of approximately 1.000.000 tons of gypsum (CaSO4.2H20). As it happens with other residues produced in large bulks, the plasters generated in the desulphuration process cause a management problem; however, they also possess a potential profitability if an adequate market for them is found.
3.-
ANTECEDENTS
AND HISTORY
Enormous investments are required to reduce the sulfur emissions produced from coal power plants. In the United Kingdom, only one company has done the necessary investments to obtain a satisfactory reduction in the emissions of sulphur dioxide; this was accomplished when the factory was a public enterprise. In Spain only a public company ENDESA has decided for cleaner combustion after a hard public opinion struggle against the consequences of air pollution. All this suggests that funds are more easily obtained when political factors are presents. In the United States, the protection of the environment has been promoted exclusively through the pressure of the social movements in the sixties. The market and the cost of energy has not stopped in the United States the enaction of laws for reducing the pollutant emissions. Title IV of the "Amendment to the Clean Air Act" of 1990 has imposed new limitations to the emission of SO 2 and NOx from fuel power plants, which must be adapted to new fuels or procedures (Locker, 93; South & Bailey, 95; Barsotti & Kalyonku, 95). In Japan, desulphuration gypsum has been employed as an alternative material to the traditional chemical plaster since the 70's (Nagata, 1995). The Japanese anti-pollution laws and measures against the emission of sulphur oxides date of 1968; the first desulphuration plants are running since 1972. Other desulphuration facilities were installed between 1972-1977 in 18 central thermal and in 18 metallurgical industries, in addition to in other 13 industries.
In Canada (Luckevich, 1993) desulfogypsum are an alternative to mineral plaster began to be applied in the 80's. The cited author thinks that "... in North America .. there is much to learn from the European and Asian experience insofar as the b y - p r o d u c t s utilization". In Europe, during the last years, there has been an important increase in the non-polluting legislation, within in the European Union and in the OECD (Franz Wirsching, Rolf Hiiller & Rainer OlejnikJ, 1994). The Residues Directive (91/156/EEC, of 18 March 1991) contains a catalogue with by-products that are defined as: "substances or objects described in the Annex I and that are referred to waste materials". In the OECD Directive (Control of Transborder Residue Movements intended for Recovery: C(92)39) desulphogypsum is considered as a material included in the " green list" (non-dangerous). German legislation, which included initially plasters Fuel Gas Desulphuration (FGD) as residues, now considers it as a "product". In the Netherlands, lignite power plants equipped with decontamination systems have introduced in the market important quantities of gypsum plaster (400.000 tons, Moonen, 1991). As consequence of mineral resources shortage, the Dutch administration favors the research and development of technologies that provide construction materials as recovery products (Winden, Zwan, Zeilmaker 1991). The experience with desulphogypsum plasters is not extended to all the countries of the European Union. Experimental work with ashes and plasters in Italy, (Gera, Mancini, Mecchia, Sarrocco & Scheneider 1991) appeals to the use of imported desulphogypsum plasters because there are not power plants in this country with systems of desulphuration.
583
In Sweden (Timm, 1991) 52 million tons of wastes are produced annually. 20 million are recovered,either as energetic material or as by-products. Within those, 500.000 tons are combustion wastes, including plasters. Environmental protection laws do not consider including the products of decontamination of gaseous emissions as dangerous materials, but the "Swedish Environmental Protection Agency" has developed environmental protection procedures, setting very strict limits, mainly to reduce the emissions that had caused the destruction of thousands of small lakes.
4. - Q U A L I T Y
OF THE
MATERIAL
Wirsching (93, 95) defines the plaster FGD as:
"The plaster produced in plants of desulphuration (gypsum FGD or desulphogypsum) is calcium sulfate dihydrate very pure, wet, crystalline, finely granulated. It is produced specifically through processes of desulphuration in those which, through injection of calcium carbonate to the scrubbers, is therein after submitted to oxidation processes and dehydrated. The plaster FGD is produced in the fuel power plant under criteria of quality for that material". 4.1. G e n e r a t i o n
of the g y p s u m F G D .
The process of desulphuration is based on the following reactions:
. c,,co, . 2u ~ - . c,,so, 89
§
Of this reaction results that each 64 grams of SO 2 will consume 100 grams of calcium carbonate CaCO3, but will issue to the atmosphere 44 grams of carbon dioxide CO 2. The calcium sulfite hemihydrated is submitted to forced oxidation in all the modern systems to produce calcium sulfate dihydrate:
CaSO,. 2 HzO + S02 + HzO ~
Ca(HS03) 2 + 0 z + 2H20 ~
Ca (HS03) z + 11_120
C a S O , . 2 H 2 0 + HzSO 4
CaCO 3 + HgSO 4 + 1120 ---,, CaSO42HzO + CO z
That globally: 2CaSO,.2HzO
+ 0 z + 3HzO ~
2CaSO4.2H20
In the first step, the insoluble calcium sulfite reacts with more sulphur dioxide producing calcium bisulfite (soluble); this reacts spontaneously with the atmospherical oxygen, producing calcium sulfate dihydrate and also is produced more sulfate, by direct reaction of calcium carbonate with the sulfuric acid, that is generated in the oxidation reaction.
584
4.2. P r o p e r t i e s The "Decision of the Council of the OECD over the T r a n s b o r d e r Control of Residues intended for Recovery Operations" (C(92)39) establishes some lists (green, amber and red). In principle, the gypsum FGD were included in the green list, of where it were eliminated to instances of "EUROGYPSUM", that requested that it would be recognized its character of "product". This request was based on the following considerations: The gypsum FGD, as a matter outweighs secondary, it is used by European industry of the plaster. Its economic value is comparable to the natural plaster and its price is conditioned by transportation freight The gypsum FGD is transported by highway, railway or load ships without need of special cautions. In the German legislation (Federal German Emission Protection Law) only are considered as residues, the materials that they can not be reutilized and they must be rejected. In the United Kingdom, France, Belgium, Holland, Denmark and Austria the gypsum FGD has the character of "product".
4.3. G r a n u l o m e t r i c
distribution
The grain of the natural or artificial gypsum present important differences. The figures 1 and 2 (Glasscock, 93) allow to compare the different size distributions of the natural and artificial plasters. The first is generated by grinding of mineral chunks (3" to 6") and the second by the directly formed crystals during the desulphuration process into the scrubber. These last have been formed in an stirred environment, where the presence of numerous condensation nucleus generate grains with very uniform shape and size. O~ O3
~
100
80 60
U L
~
40
= E o
20 0 0
1
10 100 Particle Size (microns)
1000 100
. c_
--
m
80
60 >= 40
._
NIPS
4----
Natural (AJabaster) Landplaster
m
E
20
300
212
150
106
75
53
38
27
19
13
Particle Size (microns) Natural
"0-
Synthetic
9,4
6.6
4.7
3.3
2.4
585 4.4. Normative The parameters that indicate the quality of the desulphogypsum, because its utilization as by-product in European industry of the plaster are withdrawals in following Table I (Wirsching, 1993). The "United States Gypsum Company" (USG) has a long experience in the characterization and the marketing of the american plasters; It establishes a series of characteristics for the artificial plasters,(Henkels, Gaynor, 1995). Ojanpera and Cabbage (1993) (see Table II) offer a comparison between natural plasters and desulphogypsum in the one which are emphasized important advantages of these last with respect to several of the characteristic parameters.
Table IL ~ U t y of Gypsum FGD and natural
Table I. Quality criteria for Gypsum FGD Parameter
% weight
Water content
< 10
CaSO4.2H20
> 95
Soluble MgO
< 0.1
Propriety
Gypsum FGD
Free water %
6-10 !
Combinated water %
CaMg
,
Natural Gypsum
~
0.2-3
!
19-21
19-21
0--5
0.15-1.6
CI-
< 0.01
Carbonates %
Na20
< 0.06
caso,.~o %
91-99.8
97-98.5
SO2
< 0.25
SiO, ppm
1700-7000
5000--6500
5-9
AI20s ppm
1000-5000
<5000
white
F%O3 ppm
140-1300
600-4200
neutral
MgO ppm
100-800
5000-14000
Na + K ppm
300-1400
200-500
SO2 ppm
10-700
0
Chloride ppm
10-100
70-80
pH Color Smell Toxicity
no-toxic
5. WORLD WIDE PRODUCTION (Unpredictable trends) Different evaluations have been done to estimate the plasters total of desulphogypsum currently produced and reutilized in each one of the different countries, and the world over. Any estimation would be a provisional figure, as consequence of the figures of consumption is conditioned by the following circumstances: 1. The petroleum crisis and its spectacular rise in costs in the seventies was the cause of the implementation of the utilization of sources of fossil fuels, that they had been rejected by its poor energetic properties (low power) and environmental (high contents in sulphur). 2. The prices readjustment of the liquid fuel has not made to reduce energy demands originating from the coal. Nuclear energy has been submitted to critical on the part of the public opinion, in such a way that the construction projects of new nuclear plants has been reduced in the last years.
586
3. The Rio de Janeiro United Nations Conference of 1992 established a series of commitments, about the limitations in the pollutant emissions to the atmosphere, above all centered in the emissions of C02, hydrocarbons and substances related with "green house effect" and in the depletion of the ozone layer. 4. Without may have been object of international procedures of that range, many countries in those which acts the free market have developed a vast application of desulphuration systems that in most cases produce a momentary "excess" of gypsum of good quality in the by-products market. 5. The technology of clean combustion is an elaborated system that, as of high initial investments allows to reach materials of uniform quality.
6. R E U S E O F G Y P S U M F R O M 6.1. G y p s u m c o n s u m p t i o n
DESULPHURATION
PROCESS
For better understand the tendencies on the production and reutilization of desulphogypsum we can refer the situation in EEUU and Japan. The table III shows the production, imports and consume of gypsum in EEUU. Barsoty & Kalyoncu (1995) estimate proximately 20.000.000 tons the production of FGD in EEUU in 1993, before the enaction of IV amendment in the Clean Air Act. Table III: Production and reuse of Combustion Products in EEUU Fly Ash
Bottom Ash
Slag Furnace
GypsumF GD
Production ton
43,4
12,8
5,6
18,4
Use ton
9,5
3,8
3,1
1,0
22
30
55
In Japan desulphogypsum is the most important secondary material used in construction (Makansi & Ellison, 1993) (Fig. II). Because they are not mineral exploitation and the absence of land to be used like damping area, the policy of wastes is in this country especially strict. The consume gypsum demand has been estimated in 9,3 millions of tons (Nagata, 1995). A total 25 % are gypsum from power plants desulphuration what is recovered in 100%. The 43 % is imported from Tahiland and Mexico.
6.2. A p p l i c a t i o n s Very diverse applications are mentioned in the technical literature, three different type of refereed applications can be distinguished: a) Applications with higher added value. In this reutilization the gypsum is transformed in building materials (pre-fabrication) to be applied like a finished product Board walls, Blocks for construction; Shape-molded elements; Artificial reefs.
587
b) Use as prime material. The gypsum can be used, directly or after a semifinished transformation in a material to be applied in building or civil works Mortars; Artificial gravel; Additive or cement filler for concrete; Special cement; Self-leveling floors material; Concrete for pipelines; Mortar for mining; c) Use in a big amount of quantities, to be applied in civil engineering and environmental purposes. These are the applications were an high volume of secondary materials are used. Material of mines landfill; Residues stabilization ; Material for highway bases; Soil stabilization
7. - C O M M E R C I A L
ASPECTS
In other estimation Makansi & Ellison (1993) evaluate in 27 million of tons of plaster used per year in EEUU. A part is imported from Canada and Mexico (between 30% and 40%). In Canada, the annual consumption is about 3.5 million of per year tons. In 1986 the FOB price of the natural mine gypsum was of approximately 8 S/ton (according to the Mineral Industry Survey of the Bureau of Mines, EEUU 1977). In consequence of this low price it is difficult the introduction of artificial gypsum produced in very expensive installations. Until the enaction of the "IV amendment to the Clean Air Act" in EEUU, the preoccupation for obtaining desulphogypsum in a commercial quality was constituting only a marginal phenomenon (Locker, 1993). Now, the things have changed in this country; the companies consuming the by-product demand specifications of quality very rigorous, the impurities presence, color, etc. will have to be absolutely guaranteed in the product. The price and the difficulty of the transportation is the factor more unfavorable. In Europe the principal consumers countries of synthetic gypsum are Germany, Holland and Austria (Makansi & Ellison, 1990). The clean combustion technology, originated from Germany and Japan has been developed, until obtaining plasters that fulfill the specifications demanded by the consumers. Nevertheless, the adjustment of the European markets, to the substitution of the mineral plasters by the by-products, it is very difficult and it has obligated to improve the quality. In Spain, the natural gypsum market consumes more than 2.5 millions tons every year in a disperse little manufacture industries. The future introduction of more than one millions tons of gypsum FGD will represent a problematic management.
8. F I N A L
CONSIDERATIONS
We can analyze, some of the common characteristics defining this situation. a) The industries consuming gypsum are not new installations; they are decades on a traditional activity, consuming a cheap prime material. Nevertheless, in several countries the activity is based on imported minerals (EEUU, Japan, Holland), although the plaster constitutes a relatively small part of the production costs (Makansi & Ellison, 1990); (Henkels, Gaynor & Garceau, 1993); (Kuntze, 1993); (Nagata, 1995). b) Frequently, the location of transformation gypsum industry are where the gypsum mines are abundant and inexpensive of extracting. The products that are manufactured in that industry are not expensive, the most of them devoted to construction. The plaster products can deal affected for uncontinuous demands (campaigns) (Locker, 95). c) The reason for assuming the clean combustion is to protect the environment, taken as a political decision. The initiative has begun in the countries, in those which the free competence is an economic and ideological
588
paradigm. The market occurred for those by-products has been appeared, after to have generated it in a non return way. We can interpret the situation as a stages succession: 1) To clean the atmosphere produces pollution by solid residue; management of these is viable and known, before of the beginning of its appearance 2) To sell the by-product can produce benefits 3) To put in the market the by-product could have indirect environmental advantages, as the avoiding of the opening of new mines, energetic saving, etc. (Wirsching, 1993); (Wirsching, 1995);
(South & Bailey, 1995); (Armour, 1995); (Jensen, Freimut & Wetegrove, 1995) d) Industry consuming gypsum does not accept unconditionally, the presence of a new product, not even to "zero cost". The improvement of quality of the produced plasters has gone increasing, throughout the years in which it has been developed the technique, and yet today it continues improving. The exigency of quality for the product, composition and regularity as supply are an exigible condition by all the gypsum consumers. The acceptance of the alternative materials, as matters outweigh secondary by industry is not only an economic problem, but cultural (Harness & Shang, 1993); (Barsotti & Kalyoncu, 1995); (Takshina,
Ukawa & Kimura, 1995). e) On important problem for the management of synthetic gypsum is the "diversity of scale" between producer and consumers. For the gypsum producing companies, the adaptation to some requirements of a "small consumer" constitute an important restriction. As consequence of this diversity of productions and requirements, they have come out on the market intermediary companies (Feeney & Novak, 1993); (Makansi
& Ellison, 1990); (Jurkowitsch & Hiiller, 1995); (Lani, College & Babu, 1995); (Welliver, Roth & Colijin, 1995); (Larrimore, 1995). f) The production of the by-product in the power plant is uniform and will be produced every day, with or without market. The various markets existence in volume, pace of consumption and geographical dispersion would compel, in any event, to have a system stocks storage, that deadens the fluctuations of demand. The capacity of the deposits of those stocks will have to anticipate, the situation of "not reuse" for unlimited periods of time (Armour, 1995).
9. C O N C L U S I O N S As a conclusion: The gypsum FGD production generates a big amount of the by-product. The reuse of this material does not present technical problems in the most of his applications. Nevertheless the activity consuming the greatest volume of materials, the civil engineering, in public works, appears like the most shy consumer. A short number of the papers consulted describe experiences consuming a big volume of material in banks, highway bases and works of this character. In this case it is tended furthermore to propose, the mixtures employment of other by-products, also generated in the power plant, as bottom and fly ashes. In most of these experiences are presented promising results, but always is necessary the technological research at appropriate scale (Golden & Saylak, 1993); (Clarke
& Smith 1991 and Clarke 1992, 1993); (Bhat & Rogers, 1995).
589
10. R E F E R E N C E S Armour D.W., FGD gypsum management phylosophy at two midwestern utilities. Proceedings of 4th International Conference on FGD and other Synthetic Gypsum, May 17-18, 1995, Toronto, Canada. Barsotti A.F. and Kalyoncu R., Implications of glue gas desulfurization on the mineral industries. Proceedings of 4 th International Conference on FGD and other Synthetic Gypsum, May 17-18, 1995, Toronto, Canada. Bhat P.A. and Rogers K.J., Geotechnical engineering study of FGD gypsum for landfill operation. Proceedings of 4 th International Conference on FGD and other Synthetic Gypsum, May 17-18, 1995, Toronto, Canada. Clarke L B . and Smith I.M., Management of residues from coal utilisation: An overview of FBC and IGCC byproducts. Proceedings of the International Conference on Environmental Implications of Construction with Waste Materials (WASCON '91), p81, November 10-14, 1991, Maastricht, The Netherlands. Clarke LB., Applications for coal-use residues. IEA Coal Reseach, 1992 Clarke L.B., Utilisation options for coal-use residues: An international overview. Proceedings of American Coal Ash Association 10th International Coal Ash Utilisation Symposium, Jam 1993, Orlando, FL, USA. Clarke L.B., Applications for Coal use residues: an internacional overview. Proceedings of the International Conference on Environmental Implications of Construction materials and technology developments (WASCON94), June 1-3, 1994, Maastricht, The Netherlands.
Feeney S. and Novak J., Case study in the use of synthetic gypsum generated from a wet FGD system. Proceedings of 3~ International Conference on FGD and Chemical Gypsum, September 26-28, 1993, Toronto, Canada. Gera F., Mancini O., Mecchia M., Sarroco S. and Schneider A., Utilization of ash and gypsum produced by coal burning power plants. Proceedigns of the International Conference on Environmental Implications of Construction with Waste Materials (WASCON '91), p733, November 10-14, 1991, Maastricht, The Netherlands.
Glasscock J., Qualification testing and results analysis for natural and synthetic gypsum sources. Proceedings of 3 ~d International Conference on FGD and Chemical Gypsum, September 26-28, 1993, Toronto, Canada. Golden D.M. and Saylak D., High-volume utilization of FGD gypsum in civil engineering projects. Proceedings of 3~d International Conference on FGD and Chemical Gypsum, September 26-28, 1993, Toronto, Canada.
Harness J. and Shang J., What are the barriers for waste utilization ?. Proceedings of 3~ International Conference on FGD and Chemical Gypsum, September 26-28, 1993, Toronto, Canada. Henkels P.J. and Gaynor J.C., Characterization of synthetic gypsum. Proceedings of 4th International Conference on FGD and other Synthetic Gypsum, May 17-18, 1995, Toronto, Canada. Henkels P.J. Gaynor J.C. and Garceau N.F., The conversion of United States gypsum company's east Chicago, indiana plant to nipsco FGD gypsum. Proceedings of 3~ International Conference on FGD and Chemical Gypsum, September 26-28, 1993, Toronto, Canada. Jensen K., Freinmth C. and Wetegrove, Energy saving FGD gypsum calcination methods. Proceedings of 4th International Conference on FGD and other Synthetic Gypsum, May 17-18, 1995, Toronto, Canada. Jurkowitsch H. and Hiiller R., Handling and processing FGD gypsum for application in building materials in europe. Proceedings of 4 th International Conference on FGD and other Synthetic Gypsum, May 17-18, 1995, Toronto, Canada. Kuntze R.A., Alpha -plaster from FGD gypsum and potential markets. Proceedings of 3~ International Conference on FGD and Chemical Gypsum, September 26-28, 1993, Toronto, Canada. Lani B., College J. and Babu M., Results of Thioclear testing: magnesium-lime FGD with high SO2 removals and salable byproducts. Proceedings of 4th International Conference on FGD and other Synthetic Gypsum, May 17-18, 1995, Toronto, Canada. Larrimore L., Evaluation of uses for byproducts gypsum from the Chyoda FGD process. Proceedings of 4 th International Conference on FGD and other Synthetic Gypsum, May 17-18, 1995, Toronto, Canada.
590
Locker D., Byproduct gypsum: White water or the alchemist's dream ?. Proceedings of 3 ~dInternational Conference on FGD and Chemical Gypsum, September 26-28, 1993, Toronto, Canada.
Luckevich L.M., The tecnology of gypsum utilization in the 1990"s. Proceedings of 3~ International Conference on FGD and Chemical Gypsum, September 26-28, 1993, Toronto, Canada 9 Ludwig U., Khan N.Y. and Hiibner, High performance anhydrite and hemihydrate binders hrom flue gas desulphurization and chemical gypsum. Proceedings of 4th International Conference on FGD and other Synthetic Gypsum, May 17-18, 1995, Toronto, Canada. Makansi J. and Ellison W., Utilization of gypsum from flue gas desulphurization. Indian J. Environmental Protection, vol 10, N-~2, Feb 90, 81-93. Moonen L., FDG gypsum and self-levelling floor screeds. Proceedigns of the International Conference on Environmental Implications of Construction with Waste Materials (WASCON91), p507, November 10-14, 1991, Maastricht, The Netherlands.
Nagata N., Gypsum utilization in Japan. Proceedings of 4th International Conference on FGD and other Synthetic Gypsum, May 17-18, 1995, Toronto, Canada. Ojampera R., Raman K. and Hale R., FGD gypsum production and utilization in Ontario, Canada. Proceedings of 3~ International Conference on FGD and Chemical Gypsum, September 26-28, 1993, Toronto, Canada. South D.W. and Bailey K.A., FGD systems: What utilities chose in phase I and what they mightchoose in phase H. Proceedings of 4 th International Conference on FGD and other Synthetic Gypsum, May 17-18, 1995, Toronto, Canada.
Takashina T., Ukawa N. and Kimura K, Improving FGD gypsum calcination methods. Proceedings of 4th International Conference on FGD and other Synthetic Gypsum, May 17-18, 1995, Toronto, Canada. Timm B., Swedish environmental policy on utilization of by-products from fuel combustion. Proceedings of the International Conference on Environmental Implications of Construction materials and technology developments (WASCON94), June 1-3, 1994, Maastricht, The Netherlands. Weiliver W.R., Roth T.J., Brown J.R., Hudson M.S. and Colijn H., Powerchip gypsum: A new choice in material handling for industrial gypsum. Proceedings of 4 Internahonal Conference on F(~I5 and other Synthetic Gypsum, May 17-18, 1995, Toronto, Canada. 9
th
9
Wirsching F., FGD gypsum definitions and Legislations in the European Community, in the OEDC and in Germany. Proceedings of 3 International Conference on FGD and Chemical Gypsum, September 26-28, 1993, Toronto, Canada. Wirsching F., Hiiller R. and Oiejnikj R., FGD gypsum definitions and legislation in the european communities, in the OECD and in Germany. Proceedings of the International Conference on Environmental Implications of Construction materials and technology developments (WASCON94), June 1-3, 1994, Maastricht, The Netherlands. Wirsching F. and Kirchen G., Fillers and coating pigments from FGD gypsum. Proceedings of 4 th International Conference on FGD and other Synthetic Gypsum, May 17-18, 1995, Toronto, Canada.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
FINE
GRINDING
OF
HARD
Sidor Jan a, W 6 j c i k
CERAMIC
WASTES
591
IN
A. Mariusz b and K o r d e k
ROTARY-VIBRATION
MILL
Jacek c
A c a d e m y of M i n i n g and M e t a l l u r g y - T e c h n i c a l University, A v . M i c k i e w i c z a 30, A-3, 30-059 Cracow, Poland aFaculty of M e c h a n i c a l E n g i n e e r i n g and Robotics, D e p a r t m e n t of Technological Equipment andEnvironment Protection, bFaculty of M a t e r i a l s Science and Ceramics, Department of B u i l d i n g Materials, CFaculty of Mining, Department of Mineral Processing, E n v i r o n m e n t a l P r o t e c t i o n and U t i l i z a t i o n of Wastes,
ABSTRACT
Some results of g r i n d i n g of two high p u r i t y and hard ceramic wastes as c o r u n d u m insulator of s p a r k i n g plugs (CSP), filaments of sodium lamp tubes (CLT) and s e m i c o n d u c t i v e silicon (Si) were p r e s e n t e d in this paper. E x p e r i m e n t s consists of p r e l i m i n a r y crushing and g r i n d i n g in v i b r a t o r y c r u s h i n g (VC) and final v e r y fine g r i n d i n g in rotary v i b r a t i o n mill (RVM) . Results r e v e a l e d that ground m a t e r i a l c h a r a c t e r i z e w i t h a high p u r i t y and a v e r y fine p a r t i c l e size d i s t r i b u t i o n b e l o w 5+10~m. Ground wasted corundum p o w d e r can be again used in ceramic technological p r o c e s s e s or can be used as abrasive m a t e r i a l while silicon p o w d e r can be u s e d in the p r o d u c t i o n of high p u r i t y silicon carbide. 1
9I N T R O D U C T I O N
T e c h n o l o g i c a l ceramic wastes of high p u r i t y and hardness are g e n e r a t e d in any of special ceramic technologies. High p u r i t y means that ceramic wastes does not c o n t a i n more than 0.1% of total harmful impurities while h a r d n e s s is close to 6+10 in Mohse's h a r d n e s s scale. Ceramic wastes are u s u a l l y c r e a t e d in such unit o p e r a t i o n s as: forming, sintering, m e c h a n i c a l treatment and a s s e m b l i n g of ready goods and increase the p r o d u c t i o n costs when they are in m a n y tons. Ceramic wastes c r e a t e d in forming o p e r a t i o n does not make problems b e c a u s e they can be again used in t e c h n o l o g i c a l p r o c e s s e s but the ones such as- d e f e c t i v e byproducts, cut t e c h n o l o g i c a l surplus or d e f e c t i v e final p r o d u c t s created in the r e m a i n i n g ones are inconvenient. This t r o u b l e s o m e are made by n e c e s s i t y of g r i n d i n g of such hard m a t e r i a l s to particle size d i s t r i b u t i o n close to raw materials, i.e. b e l o w 5+10~m k e e p i n g high p u r i t y of powders d u r i n g grinding. The p o s s i b i l i t i e s of u t i l i z a t i o n of ceramic wastes are c o n d i t i o n e d by the following considerations: - high p u r i t y and price a p p e a r i n g from high q u a l i t y of raw materials,
592 - r e n e w a l a p p l i c a t i o n of g r o u n d c e r a m i c w a s t e in the same or other technological processes influences on lowering of p r o d u c t i o n costs, - p r o t e c t i o n of n a t u r a l e n v i r o n m e n t . The o r i g i n a l l a b o r a t o r y v i b r a t i o n m a c h i n e s as V C and RVZ~ c h a r a c t e r i z e d w i t h low a c t i v i t y on n a t u r a l e n v i r o n m e n t were used to p r o v i d e the e f f e c t i v e n e s s of g r i n d i n g p r o c e s s e s and to high p u r i t y of c o r u n d u m and s i l i c o n c e r a m i c wastes. C h o i c e of these m a c h i n e s w e r e done on the b a s i s of t h e i r h i g h t e c h n i c a l and technological advantages confirmed in previous preliminary g r i n d i n g of h a r d c e r a m i c m a t e r i a l s and c e r a m i c wastes. The very h i g h d e g r e e of c o m m i n u t i o n and h i g h p u r i t y of p r o d u c t s were obtained in VC during crushing of the titanium silicon c a r b i d e (I), a l u m i n i u m nitride, f e r r o s i l i c o n (2'3) P o s i t i v e results were o b t a i n e d in v e r y fine g r i n d i n g and c o l l o i d a l g r i n d i n g in RVZ{. The p r e l i m i n a r y e x p e r i m e n t s of g r i n d i n g of h a r d m a t e r i a l s such as t i t a n i u m s i l i c o n carbide, s i l i c o n c a r b i d e (4), c o r u n d u m w a s t e s (s), a l u m i n a 16'7), fly ashes (8), h y d r o x y a p a t i t e (9) and other m a t e r i a l s cI~ w e r e c a r r i e d out in RVZ~. 2. E X P E R I M E N T A L 2. l . M a t e r i a l s .
The f o l l o w i n g c e r a m i c w a s t e s were u s e d in c r u s h i n g and grinding experiments: - corundum insulators of sparking plugs (CSP) of maximum d i m e n s i o n s ~ 1 4 / ~ 4 x 6 5 (tube), - c o r u n d u m f i l a m e n t s of s o d i u m lamp tubes (CLT) of m a x i m u m d i m e n s i o n s ~ 9 . 6 / ~ 8 x 9 0 (tube), - s e m i c o n d u c t i v e s i l i c o n (Si) of 2 0 x 2 5 x 4 0 (irregular grains). 2.2.
Vibration
crusher.
KW 40/1 V C (Fig.l) was u s e d to p r e l i m i n a r y c r u s h i n g of CSP. C r u s h e r ' s jaws set in s w i n g i n g m o t i o n will a l w a y s m o v e in the opposite directione i t h e r i n w a r d or outward. C r u s h i n g takes p l a c e w h e n jaws m o v e i n w a r d w h i l e d u r i n g t h e i r o u t w a r d m o v e m e n t the c r u s h e d m a t e r i a l is p o u r e d down. B a s i c p a r a m e t e r s of KW 40/1 are as follows- inlet d i m e n s i o n s - 100xl60mm, - m a x i m a l feed d i m e n s i o n : 40mm, - slit a d j u s t m e n t range- 1-10mm, - f r e q u e n c y of jaw v i b r a t i o n s : 8+25Hz, - m o t o r power: 2.2kW, - output: 5 0 + 5 0 0 k g / h , - dimensions: 840x420x920, - mass: 145kg. 2.3.
Rotary
vibration
mill.
L A M O W - C - 5 x 2 R V M (Fig.2) was u s e d to p r e l i m i n a r y c r u s h i n g of C L T and fine g r i n d i n g of CSP, C L T and Si. G r i n d i n g is done by s e t t i n g the m i l l c h a m b e r w i t h the g r i n d i n g m e d i a and g r o u n d material in the motion combine vibration motion with the v i b r a t i o n a m p l i t u d e p e r p e n d i c u l a r to the axis of c y l i n d r i c a l c h a m b e r and the r o t a r y motion. B a s i c p a r a m e t e s of L A M O W - C - 5 x 2 are as follows-
593
fen
~----~
3
L[~
Figure 1 S c h e m e
of KW 4 0 / 1 V C l-jaw, 2 - k i n e m a t i c v i b r a t o r , 3v i b r a t o r d r i v e , 4 - e l a s t i c unit, 5-slit adjustment mechanism -
v i b r a t i o n s f r e q u e n c y : 14Hz, v i b r a t i o n a m p l i t u d e : 2+8mm, c a p a c i t y of the c h a m b e r : 5 a n d n u m b e r of c h a m b e r : 2, m o t o r p o w e r : 1.5kW, dimensions: 1270x800x960, - mass: 240kg.
2.4. Experimental
procedure
i '
t_W---
Figure 2 S c h e m e of L A M O W - C - 5 x 2
RVM:l-chamber unit,2-chamber, 3-vibrator,4-vibrating unit frame, 5-mill drive,6-shield, 7-base plate,8-vibro-insulator i0
liter,
of preliminary
crushing.
The m a i n o b j e c t i v e of p r e l i m i n a r y c r u s h i n g w a s to o b t a i n d i s i n t e g r a t e d m a t e r i a l of s u c h p a r t i c l e s i z e d i s t r i b u t i o n that it c o u l d be i n t r o d u c e d i n t o the RVM. C e r a m i c w a s t e s of C S P was f e e d into V C in n a t u r a l s h a p e w h i l e Si w a s f i r s t h a n d c r u s h e d till the dimensions were smaller than 40mm. The CTL was p r e l i m i n a r y c r u s h e d in L A M O W - C - 5 x 2 b e c a u s e of t h i n t u b e w a l l s a n d a v o i d a n c e of m e t a l l i c p o l l u t i o n f r o m jaws. C r u s h i n g w a s c a r r i e d out on d r y in c h a m b e r with polyamide lining using corundum grinding media. In p r e l i m i n a r y c r u s h i n g the slit was i m m a n d the v i b r a t i o n f r e q u e n c y w a s 18Hz.
2.5. Results
of p r e l i m i n a r y
crushing.
The p a r t i c l e s i z e d i s t r i b u t i o n b e f o r e a n d a f t e r c r u s h i n g of s h o w n on Fig. 3 a n d 4. C o r u n d u m w a s p o l l u t e d w i t h metallic contaminations d u r i n g c r u s h i n g a n d w a s c l e a n e d b y the use of m a g n e t i c s e p a r a t o r . C S P a n d Si w e r e
2.6. Experimental
procedure
of fine grinding.
The p r e l i m i n a r y c r u s h e d f r a c t i o n c l a s s of c o r u n d u m C S P b e l o w - 0 . 4 0 m m w a s u s e d in g r i n d i n g w h i l e c e r a m i c w a s t e s of C T L w a s u s e d on the w h o l e . The p a r t i c l e size d i s t r i b u t i o n of p r e l i m i n a r y c r u s h e d c e r a m i c w a s t e s d e t e r m i n e d on wet F r i t s c h m i c r o s i e v e s was s h o w n in T a b l e i. B o t h c o r u n d u m w a s t e s C S P a n d C T L w e r e g r o u n d on wet in d i s t i l l a t e w a t e r in L A M O W - C - 5 x 2 d u r i n g 1 , 2 , 4 a n d 7h. The c h a m b e r has p o l y a m i d e l i n i n g a n d it c a p a c i t y w a s 5 d c m 3. The f e e d c o n s i s t s of 8 0 0 g of m a t e r i a l , 6 0 0 0 g of c o r u n d u m g r i n d i n g
594
2
2
loo
,o
8o 70 60 5O ,o
IIIIII
I
II 1111 I~1111111
I I / /
::tt '
1
I ~1111
I1-1111 1/11111111 IIIII11 111 I I IIIIII / I I IIIIIII
lllllfl l lllll
I IIIIIIII/ I Ill II I illli~ llt IIII
II I IIIIII I II I III IIIIII I II
:!I I lIU~ll l'lllilll l~,l!!
I9
!!!!!!!!
I~[T[R, (ram)
I0
19
Figure 3 Particle size d i s t r i b u t i o n of C S P b e f o r e a n d after crushing: 1-feed, 2product Table
1 Particle
Class,#m CSP,
1 1 I
IIIIII1 /I II III
size
%
CTL , %
4,3
22.8
-
-
DIAMETER. (ram)
Figure 4 Particle size d i s t r i b u t i o n of S i b e f o r e and after crushing: 1-feed, 2product
distribution 250
9 ........3 1 5 - 2 5 0
.......:...... 9
l] 1 U 1] 11
18.4
of
CSP
125
24.7 18.2
and
CTL
wastes.
i 5-63 I, :,01 i,0 1 14.6
2.4
12.8
17.3
10.3
53.2
media
a n d 4 0 0 c m 3 of water. G r i n d i n g of S i was c a r r i e d out on d r y in air in this same R V M d u r i n g 2 , 4 , 6 a n d 10h. The c h a m b e r has a l s o p o l y a m i d e l i n i n g a n d it c a p a c i t y was 5 d c m 3. The f e e d c o n s i s t s of 1 0 0 0 g of m a t e r i a l and 4 0 0 0 g of s i l i c o n g r i n d i n g media. Particles size d i s t r i b u t i o n w e r e d e t e r m i n e d u s i n g laser g r a n u l o m e t e r F r i t s c h A n a l y z e t t e 22. S a m p l e s of all fine g r o u n d p o w d e r s w e r e a l s o o b s e r v e d u n d e r s c a n n i n g m i c r o s c o p e J E O L 5400. 2.7.
Results
of fine g r i n d i n g of C S P .
R e s u l t s of fine g r i n d i n g of C S P w e r e s h o w n in T a b l e 2. SEM of c e r a m i c w a s t e s b e f o r e a n d a f t e r g r i n d i n g (sample C S P - 4 ) were s h o w n on F i g u r e 5 a n d 6. Table
2 Results
Sample G r i n d i n g time,
of
CSP.
. . . . .
CSP-4
CSP-2 2
CSP-3 4
5.2
4.5
3.3
2.6
16.8
15.0
12.5
I0.2
-10~m
71.6
79.2
83.7
89.7
-4~m
41.7
49.2
56.1
63.3
h
size
ds0, #m
Particle
size
dg0, ~m
Fraction,
fine g r i n d i n g
CSP- 1 1
Particle
Fraction,
i
from
i
7 ,
,
,
595
F i g u r e 5 S E M of c e r a m i c w a s t e CSP before grinding, 50x. 2.8.
Results
of
fine
grinding
Figure CSP-4
of
6 S E M of c e r a m i c waste a f t e r g r i n d i n g , 2000x.
CTL.
R e s u l t s of fine g r i n d i n g of C T L w e r e s h o w n in T a b l e 3. S E M of c e r a m i c w a s t e s b e f o r e a n d a f t e r g r i n d i n g (sample C T L - 4 ) w e r e s h o w n on F i g u r e 7 a n d 8. Table
3 Results
from
fine
grinding
of
CTL.
li ................................I .................................. :
......... ~i:.:.',::.::::~'.~:~~,.::.i.:. i:.':: ...:i:: .;':. :.,,i :::, :.. :.~<.:~::::i::i:l:..:,:'..:.c~L~'2.. .iil':'.:.",:ic~.~,~:..:..~ii'.."::'~.: 4.::I
Particle
size
ds0, #m
5.2
4.5
3.3
2.6
Particle
size
dg0, #m
16.8
15.0
12.5
i0.2
-10~m
71.6
79.2
83.7
89.7
-4#m
41.7
49.2
56.1
63.3
Fraction, Fraction,
A l s o o t h e r 10h g r i n d i n g of C T L w i t h h i g h e r m a s s of f e e d (1250g) w i t h a n o t h e r set of g r i n d i n g m e d i a w e r e done. R e s u l t s s h o w n that ds0 was 1 . 5 ~ m a n d dg0 was 7 . 9 # m w h i l e f r a c t i o n - 1 0 # m was 92.9% a n d - 4 # m was 78.9%.
596
F i g u r e 7 S E M of c e r a m i c w a s t e CTL b e f o r e g r i n d i n g , 50x.
F i g u r e 8 S E M of c e r a m i c w a s t e CTL-4 a f t e r g r i n d i n g , 2000x.
2.9. Results of fine grinding of Si. R e s u l t s of fine g r i n d i n g of Si w e r e s h o w n in T a b l e 4. SEM of w a s t e s t u f f s b e f o r e a n d a f t e r g r i n d i n g (sample Si-4 w e r e shown on F i g u r e 9 a n d i0. Table
4 Results
Sample
Grinding
time,
h
from
fine
grinding
of
Si.
2
4
Si -3 6
Si -4 I0
Particle
size
ds0, #m
16.7
ii.4
9.5
5.6
Particle
size
dg0, #m
53.6
41.8
39.0
21.0
-10#m
55.2
66.2
70.8
85.3
-4#m
31.9
41.1
45.8
59.4
Fraction, Fraction,
A l s o o t h e r 4h g r i n d i n g of Si w i t h an a d d i t i o n of 0.2% of a c t i v a t o r was c a r r i e d out. R e s u l t s s h o w n that ds0 was 7 . 9 # m and dg0 was 2 8 . 3 # m w h i l e f r a c t i o n - 2 0 # m was 8 0 . 1 % a n d -8#m w a s 50.4%.
597
Figure 9 SEM of ceramic waste S i - 4 after grinding, 2000x.
Figure 10 SEM of ceramic waste S i - 4 after grinding, 10000x.
3 .CONCLUSIONS
Results of presented experiments allows to draw the following conclusions1 A p p l i c a t i o n of V C and R V M c r e a t e d the p r a c t i c a l p o s s i b i l i t y of g r i n d i n g p r o c e s s e s - fine c r u s h i n g i fine g r i n d i n g of c o r u n d u m and silicon wastes. 2 P r e l i m i n a r y g r i n d i n g can be c a r r i e d out in V C when grains are b i g g e r than 3+10mm or in R V M when grains are b e l o w l+3mm. 3 A p p l i c a t i o n of chamber with p o l y a m i d e abrasive r e s i s t a n c e lining and g r i n d i n g media made from the same m a t e r i a l as g r o u n d ones make the autogenic g r i n d i n g and p r o v i d e high p u r i t y of g r o u n d materials. 4 Corundum wastes CSP despite of higher particle size d i s t r i b u t i o n before g r i n d i n g were g r o u n d faster in the first stage of process (l+3h) than C T L and they were only slightly finer but close to C T L after next 4h of grinding. 5 The c r u s h i n g process of ceramic wastes of Si runs more e f f e c t i v e l y in V C but g r i n d i n g in R V M much slowly. Process was two times a c c e l e r a t e d when g r i n d i n g a c t i v a t o r was used. 6 Applicated in presented experiments relatively big l a b o r a t o r y devices make the p o s s i b i l i t y for their usage in the i n d u s t r y conditions for g r i n d i n g of the high q u a l i t y ceramic wastes with 9 Mohse's hardness. The output of V C KW 40/1 amounts 40+60 and 7 5 + 8 5 k g / h for c r u s h i n g of c o r u n d u m and silicon wastes r e s p e c t i v e l y
598 while for L A M O W - C - 5 x 2 is ranged from 0.5+2 and from 0.2§ for g r i n d i n g of c o r u n d u m and silicon wastes r e s p e c t i v e l y d e p e n d i n g on particle size distribution. RVZ{ MOW-G-50 and MOW-H-200 with 50dm 3 and 200dm ~ c a p a c i t y of chamber can be used when the need of higher output will be required. 4. R E F E R E N C E S
i.
Sidor J.,"Fine g r i n d i n g of Ti3SiC 2 powders u s i n g rotary v i b r a t i o n mill", FOURTH EURO CERAMICS - Vol.l- pp. 121-128. D e v e l o p m e n t s in p r o c e s s i n g of a d v a n c e d ceramics - Part i. Edited by C.Galassi, C.N.R. - IRTEC, Faenza Italy (1995).
2.
Sidor J . , " S o m e results of e x a m i n a t i o n of superfine crushing in the v i b r a t o r y crusher", VIII Conf. of Probl. in Constr. and Expl. of Met. and Cer. Machines, Zakopane 22-26 Jan. 1996 Vol. III, Ed. by Polish Acad. of Sc., W a r s z a w a
3.
Sidor J . , " V i b r a t i o n machines in fine g r i n d i n g processes", I Conference-Seminar "Kierunki R o z w o j u M a s z y n B u d o w l a n y c h i PMB", Kielce 9-10 Oct. 1996, pp.15-18, Edited by Instytut M e c h a n i z a c j i B u d o w n i c t w a i GS, W a r s z a w a (1996)
4.
J.Sidor, E . E r m e r - K o w a l c z e w s k a , "Preliminary investigation of the c o l l o i d a l g r i n d i n g of SiC in the l a b o r a t o r y rotaryv i b r a t i o n mill", Material Science Bulletin, Nr4, (1989)
5.
Sidor J . , " P r e l i m i n a r y tests of g r i n d i n g of high quality corundum scrape in r o t a r y - v i b r a t i o n mill ", III Inter. S y m p o s i u m II M i l l i n g and Air Classif. in Industry II , Rudy 13 15 Oct. 1993, Ed. by Inst. Chemii Nieorg., Gliwice
6.
Sidor J., "Preliminary tests of a prototype of the industrial rotary-vibration mill for fine milling of aluminum oxide", Physicochemical Problems of Mineral Processing, 26, p.57-64, W r o c l a w (1992)
7.
W6jcik M . A . , S i d o r J.,Kabala J.,Palyz J.,Purska E.,"Effect of the m o r p h o l o g y and grains size d i s t r i b u t i o n of alumina ceramic mass on the green and fire d e n s i t y and shrinkage of moulders", LIGHT METALS 1996, AIME 1996
8.
Sidor J . , W 6 j c i k M . A . , " D i s i n t e g r a t i o n of the ashes in the r o t a r y - v i b r a t i o n mill", WASCON'94 Conference, Maastricht, E n v i r o n m e n t a l Aspects of C o n s t r u c t i o n with Waste Materials, E l s e r v i e r Science B.V., Amsterdam, H o l l a n d (1994)
9.
Sidor J., "Some results of the i n v e s t i g a t i o n of superfine and colloidal grinding of h y d r o x y a p a t i t e in a rotaryv i b r a t i o n mill", CERAMICS 46, 1994, Polish Ceramic Bulletin 8, Polish A c a d e m y of Science, pp. I17-126, (1994)
i0.
Sidor J. , "Some results of investigation of superfine g r i n d i n g of selected waste m a t e r i a l s in a r o t a r y - v i b r a t i o n mill", XXVI Cracow Conference "Eng. and Prod. Techn. in Benef. Raw and Waste Materials", Ustro~ 7-8 Sept. 1994, Edited by AGH Krak6w (1994)
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
599
Influence of the calcium content on the coal fly ash features in s o m e innovative applications.
Paola Catalfamo, Sebastiana Di Pasquale, Francesco Corigliano Dipartimento di Chimica Industriale, UniversitY. di Messina, Salita Sperone 31 (Casella Postale29), 1-98166 Sant'Agata di Messina Italy.
Letterio Mavilia Istituto di Chimica, Facolt~. di Ingegneria, UniversitY. di Reggio Calabria, via E. Cuzzocrea 48, 89100 Reggio Calabria, Italy.
Abstract Despite a low chemical reactivity, recent trends in the innovative uses of coal fly ash based on the chemical properties have been successful. Lots of fly ash showing sharply alkaline reaction (water pH rising to 11-12 upon contact) usually are high-calcium (> 3-4%), most of which is present as CaO. These lots are suitable to be used as remotion agents of heavy metals in wastewater or retentive agents of them in polluted solids. Lots giving substantially neutral reaction are usually low-calcium (0-3%). They are suitable to be converted by hydrothermal treatments into zeolitic products, where higher calcium contents interfere. A reverse destination is destitute of good results. Causes are discussed.
1. INTRODUCTION. The production of electric energy from coal combustion produces far more fly ash than is currently reused. Consequently, new uses have to be promoted in order to contain the need of disposal. The conventional uses of coal fly ash (CFA) are mainly based on their physical properties. Despite a low chemical reactivity, recent trends addressed towards uses based on the chemical properties have been successful. These involve the conversion by hydrothermal treatments into partially zeolitised products, the use of CFA as an adsorber of heavy metal ions from wastewater, compost and soils or (less recently) as silicate supplier in lime and cement. In these uses of chemical type it is possible to observe different performances according to whether high calcium oxide or low calcium oxide CFA (i.e., basic or neutral CFA) are employed [ 1-5]. The presence of an excess of the free basic oxide differently influences the chemical behaviour of ashes: on one hand. it stops their conversion into zeolites [1,2]; on the other, it gives the ashes absorbtion properties which permit the removal of heavy metal from wastewater and a stronger retention of them when present in polluted soils and compost [3-5].
600 Calcium contents ranging from 1% to 10% have been found in coal fly ash produced in the Italian Electricity Board (ENEL) power stations [ 1,6,7]. In this communication we have tried to suggest, on the basis of previous work. the most convenient "chemical" utilization of coal fly ashes according to their calcium content.
2. BASIC OR NEUTRAL COAL FLY ASHES. From the point of view that mainly concerns the present discussion, it is important to observe that there are lots of fly ash showing sharply alkaline reaction (water pH rising to I 112 upon contact) and lots giving substantially neutral reaction. This different feature must depend on whether strong basic oxides (such as MgO, CaO, Na20, K20) exceed the acid ones (in particular SO2 and SO3) or not in the coal combustion products and therefore it is connected with the coal composition. Among strong basic oxides in fly ashes CaO, undergoes to the largest variation [6,7]. Thus, it is not surprising that the CaO content is the best indicator of the alkaline or neutral reaction of fly ash.
3. COAL FLY ASHES CONVERSION INTO ZEOLITIZED PRODUCTS. It has been shown that, under hydrothermal conditions, it is possible to achieve the crystallization of zeolitic phases from the amorphous fraction of coal fly ash. For example, this fraction is capable of evolving into zeolite A after 4 hrs of hydrothermal treatment; this zeolite developes into zeolite P if the treatment time goes on for 6-8 hrs or if the Si/AI ratio grows up to 3.0 [8-10]. It has, however, been observed that materials with calcium content exceeding certain limits cannot be converted into zeolites [1,2]. The authors have investigated the effective interference of the concentration of calcium ions on the nucleation and/or crystallisation of zeolite A or P. It was found that in a broad range of concentrations (0.01-0.07 tool dm -3) the calcium ion interferes with the crystallization without completely suppressing it, but rather causing it to decrease in proportion to its presence. These results have been explained in terms of a specific interaction between calcium and silicate ions, which removes the latter from the tbrmation of the aluminosilicate gel capable of evolving into zeolites with the exception of those hydroxysodalite-like [ 1,2]. According to the above reasoning, it is possible to discuss some results recently reported in the literature [11-14] just concerning coal fly ash hydrothermal treatments to obtain zeolitised products. In work reported by Singer and Berkgaut [11] the treatment has been applied to a fly ash sample containing 3,2% CaO (2,3 as Ca) and to one more containing 9,4% CaO (6,7 as Ca). The results significantly point out that zeolite P formation takes place with the first sample and hydroxysodalite with the second. An almost similar treatment reported by Lin and Hsi [ 12] has been applied to a single sample containing 5,2% CaO (3,7 as Ca). Again the zeolite P formed is consistent with our proposed rule. In this case HS has been obtained as well, but its formation was correctly explained by the authors with the higher NaOH concentration (4-1 ON) used in an alternative treatment of the same sample. In the study pertbrmed by Amrhein et al. [13] the treatment has again been applied to a single sample containing 5,67% Ca (7,94 as CaO). In this case HS forms, that is recognizable from the reported diffraction patterns although the authors quote it as "unnamed sodium aluminum silicate hydrate". Again the kind of zeolite forming is in agreement with our
601 proposed rule. As regards less recent literature [ 14], the treatment was been applied to 17 coal ash samples. For all low- or lowest-calcium content, except one named DK. Well, the results show that, in the optimum range of alkali concentration, all samples give rise to zeolite P except the latter, where HS forms. In none of the cited literature [ 11-14] were the authors able to explain the different feature, but we remark that those results should have been expected if our proposed rule had been employed by them. Thus, if it is desired to change coal fly ash into true zeolite (HS has neither channels nor exchange properties), it is necessary to use Ca-poor samples. The interference of calcium can be suppressed by complexation with EDTA, which has been shown to take place selectively in the conditions of zeolite synthesis [ 1-2]. By adding an amount of EDTA equivalent to that of calcium in the hydrothermal treatment, Ca-rich samples may zeolitise, as EDTA removes calcium combined with silicate, and this can then hydrolyse to hydrated silica. Nevertheless. in the conversion of coal fly ash into a partially zeotitised product the use of EDTA has a hardly bearable cost. Thus. it is reasonable to reserve to this use low-calcium grades.
4. COAL FLY ASHES AS "HEAVY METAL ABSORBERS". On the contrary, Ca-rich samples are suitable as heavy metals absorbers. It has been noted that Ca-rich samples exhibit good absorbtion properties in removing from wastewater moderate amounts of metal ions as copper, zinc, cadmium and chromium [3-5,t5]. Heavy metal absorbtion is maximized, pH rises from initial values of 5-6 to final values of 10-! t. In the absence of them, a pH of 10-11 is quickly established. Ca-poor samples show constant pH values (-_- 6) and do not remove any amount of heavy metals from aqueous solutions. Consequently the use of coal fly ash for the removal of metals in wastewater streams, is strictly related to their chemical composition. The absence of absorbtion properties seems evident in a material such as CFA, where unreactive phases are largely prevalent. Nevertheless, alkaline Ca-rich CFA containing reactive sub-micron sized CaO particles can explain the alkaline pH values (10-11) transmitted to the aqueous solutions and the heavy metal removal capacity, according to the slight solubility of the heavy metal hydroxides. Heavy metals removal is therefore strictly connected with the initial chemical ash composition and particularly with the presence of calcium as oxide. Indeed. it must give an alkalinity to the solutions in order to precipitate heavy metals hydroxides (pH> 8-9).
5. CONCLUSIONS. When fly ashes are destined to uses based on their chemical properties, in order to obtain good results it is useful to separate them into low-calcium, neutral, and high-calcium, basic, samples. The first are appropriate to be converted into zeolitic products by hydrothermal treatments, the latter are suitable as remotion agents of heavy metals in wastewater or retentive agents of them in polluted solids, but not vice versa. Additional examination of this subject could however provide further insights.
602
Acknowledgements We gratefully acknowledge the Italian MURST 40% fund for its contribution to this study.
REFERENCES [1 ] Catalfamo, P., Patan6, G., Primerano. P., Di Pasquale, S., Corigliano, F., 1994. The presence of calcium in the hydrothermal conversion of amorphous aluminosilicates into zeolite: Interference and removal. Mat. Eng., 5(2): 159-173. [2] Catalfamo, P., Patan6, G., Primerano. P., Di Pasquale, S., Corigliano, F., Ital. Patent Application MI 93A 002119, 5 Oct. 93. [3] Patan6, G., Di Pasquale, S., Corigliano, F., Mavilia, L., 1996. Use of zeolitised waste materials in the removal of copper (II) and zinc (II) from wastewater. Ann. Chim., 86: 87-98. [4] Mavilia, L., Postorino, G., Patan6, G., Di Pasquale, S., Produzione di zeoliti da ceneri di carbone ed esempi di applicazioni. XVI Cong. Naz. Merc., 1-3 September 1994, Pavia, pp. 110-118. Idem, Ibidem, Incontro Scient. Cost. Divisione Chimica-Ambiente, 10-12 October 1994, Roma, p. P 14. [5] Patan6, G., Mavilia, L., Corigliano, F., 1996. Chromium removal from wastewater by zeolitised waste materials. Mat. Eng., 7(4): 509-519. Patan6, G., Di Pasquale, S., Catalfamo, P., Corigliano, F., Rimozione di metalli pesanti da acque reflue. I Congr. Naz. Chim. Ambientale, 20-21 November 1995, Roma, p. 106. [6] Puccio, M., 1983. Le ceneri di carbone. ITEC Press, Milano, Italia. [7] ENEL-DSR, Internal report n. 187 AB/sr, October 1987. [8] Corigliano, F., Mavilia, L., Primerano, P., Zappia, V., ltal. Patent. n. 1217447 (Appl. 20395 A/88, 29 Apr. 88). [9] Mavilia, L., Zappia, V., Catalfamo, P., Primerano, P., Corigliano, F. Utilizzazione industriale di materiali a base di silice amorfa: III Ceneri di carbone, Cong. Naz. Merceol., 10-13 October 1988, Messina-Taormina, pp. 1829-1838. [10] Catalfamo, P., Corigliano, F., Primerano, P., Di Pasquale, S., 1993. Study of the precrystallization stage of hydrothermally treated amorphous aluminosilicates through the composition of the aqueous phase. J. Chem. Soc., Faraday Trans., 89( 1): 171-175. [11] Singer, A., Berkgaut, V., 1995. Cation exchange properties of hydrothermally treated coal fly ash. Environ. Sci. Technol., 29(7): 1748-1753. [12] Lin. C.F., Hsi, H.C., 1995. Resource recovery of waste fly ash: synthesis of zeolite-like materials. Environ. Sci. Technol., 29(4): 1109-1117. 113] Amrhein, C., Haghnia, G.H., Soon Kim. T., Mosher, P.A., Gagajena, R.C., Amanios, T., De La Torre, L., 1996. Synthesis and properties of zeolites from coal fly ash. Environ. Sci. Technol., 30(3): 735-742. [14] Kato. Y., Kakimoto, K., Tomari, M.. 1984. Analysis and utilization of coal ash. Report No. Spey 16 of the Ministry of Education, Science and Culture, Tokyo, p. 311. I 15] Ferrero, F., Gaglia Prati, M.P., 1996. Coal fly ash and alginate for the removal of heavy metals from aqueous solutions. Ann. Chim., 86: 125-132.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
P R O C E S S I N G A N D A P P L I C A T I O N OF P H O S P H O R I C G Y P S U M
R. van Selst
Head of the materials department of Intron, Institute for materials and environmental research B.V., P.O. Box 5187, 6130 PD Sittard, The Netherlands
L. Penders
Processengineer of Kemira Agro Pernis, The Netherlands
W. Bos
Processengineer of Hydro Agri Rotterdam, The Netherlands
603
604
ABSTRACT
In the Rotterdam Port Area in the Netherlands two phosphoric acid producers are situated. Together they produce approximately 1,8 million tons per year of phosphoric gypsum which is up till now discharged in the Nieuwe Waterweg and the North Sea. This phosphoric gypsum is to a certain extent contaminated with the impurities of P205, F, organic matter, Radium, ect. Research has been established into different applications of this gypsum. Phosphoric rock selection, the BAT process and purifying the gypsum result in a gypsum which could in the future replace natural gypsum in several applications. The findings of the purification trials and proposed pilot plant along with a specification of the purified material are presented in this paper.
Keywords: phosphogypsum, applications
605
Introduction At the production of phosphoric acid the residue gypsum is generated, called phosphoric acid gypsum, or also phosphoric gypsum for brevity's sake. Kemira Agro Pernis B.V. and Hydro Agri Rotterdam B.V. in Rotterdam and in Vlaardingen have a yearly production of 1.05 and 0,8 million tons of phosphoric gypsum respectively. This gypsum has up till now been discharged into the Nieuwe Waterweg and the North Sea.
Within the scope of the expiring of the licence for the discharge of phosphoric gypsum by these companies, research has been established into different applications of this gypsum.
Processes Both phosphoric acid producers are situated in the Rotterdam Port Area and have so called wet processes. Hydro Agri Rotterdam has a hemidihydrate Double Filter process which is shown in figure 1. Kemira uses the Nissan-H Repulp process shown in figure 2. In the wet phosphoric acid proces the phosphate rock is digested with sulfuric acid. The resulting gypsum is then filtrated. Apart from the phosphoric acid also a 28 percent Silicon Fluorine Solution is extracted. Both processes have a very high efficiency (98-99%) and belong to the B.A.T. processes.
Experimental Characterization of phosphogypsum The two phosphogypsums were characterized for various chemical constituents. See table 1.
Purification of phosphogypsum
radioactivity aspects. In order to purify the gypsum a fundamental research 1 was conducted into the appearance of the Radium in the rock and the gypsum. The phosphate rock was digested with nitric acid. The residu and the gypsum then produced was analyzed using a
606 scanning electron microscope.
hydrocyclone tests. Phosphogypsum and water, in the proportion 1:3 by volume was thoroughly mixed in a mixer slurry tank. After attaining uniformity in the gypsum slurry the fine fraction (< 15 i~m) was removed using a one-stage hydrocyclone. Under and overflow samples were collected. The water was removed by decantation and filtration. The gypsum samples were analyzed for various constituents.
Results and discussion radioactivity aspects. From a literature survey it was found that there are very little publications on the appearance of Radium in both phosphate rock and gypsum 2. In the phosphate rock few relatively big (50 i~m) particles containing Strontium and Barium were found (photo 1). In the gypsum the Sr/Ba containing particles were different in shape and smaller (photo 2). On the basis of a existing relation 2 between Sr/Ba and Ra it is proved that radium is already present in solid particles in the phosphate rock. Further research is focussed on the different appearance of the Radium in the phosphate rock and the verification of the relationship between Sr/Ba and Radium containing particles.
hydrocyclone tests. The analyses of the samples, purified by hydrocyclone treatment are reported in table 3. From this result it can be seen that the samples collected at the bottom, i.e. underflow, had a comparitively lower level of Radium. The underflow samples were also more white which is probably caused by the removal of the organic matter by the hydrocycloning 3. In the overflow less than 5% by weight of the total feed solids were found. The results are in line with the literature 2.
607
proposed pilot plant for the purificatin of phosphogypsum Based on the results obtained a pilot plant for the purification of phosphogypsum of the capacity of 5 ton/h has been proposed (see figure 3). In this pilot plant the impure phosphogypsum shall be mixed with water. In the mixing tank the solid content is reduced to _ 10% m/m. This slurry is pumped to a one stage hydrocyclone set. In the hydrocyclone the fine fraction (< 15 i~m) of the gypsum is preferentially removed to lower the radioactivity concentration. The underflow will contain _ 55% m/m solids. The pH of the underflow in the mixing tank will be increased to 5 using lime. On the vacuum belt filter the gypsum is washed with water to remove soluble salts. The upgraded, washed and dewatered gypsum is than ready for calcination.
specifications of the purified phosphogypsum On the basis of the research specifications for the purified phosphogypsum have been worked out. They are however based on the following conditions: - use of selected phosphate rock types; - best available technique for phosphoric acid production; - purification and dewatering of the gypsum.
In table 3 the specification for the purified phosphogypsum is given.
608
Conclusions 1. Phosphogypsum can be purified using hydrocyclones. The research showed that the organic matter and Radium can be reduced considerably by this treatment. 2. Based on phosphate rock selection, a BAT phosphoric acid process and a purification treatment a gypsum can be produced which can be compared with natural gypsum. 3. A pilot plant of the capacity of 5 ton/hour is likely to be fabricated and installed by the end of this year.
References 1. Schreurs J .... Naar een Raium-arm fosfogips", Intron report 96271, 1996. 2. Moisset J . . . . Complete
removal
of Radium from
phosphogypsum",
third symposium
on
phosphogypsum, Orlando, Florida, 1990, vol. 1, 181-196. 3. A Singh et.al .... An improved process for the purification of phosphogypsum", Construction and Building materials, 1996, vol.10, no. 8, 597-600.
609
,,,e r
~ X,~,Crystallisati~ l , I
l Ph~176 ~c'~ i
_o
on I
"'er I
t
~,~sum
-- Dihydrate,.,,..._.jjl ~ ,a ,
Figure1. HemidihydrateDoublefilterProcessfromHydroAgriRotterdam
a
610
Phosphate rock~
- "
.....
/-----\
Re-Crystallisation
-~Hemihydr~e~~.
Dihydrate
i
Filter
,, Washwater ,
, Repulp
t Filter
j
Phosphoric acid I i Gypsum I Figure 2. Nissan-H Repulp Process
Flowsheet Gypsum Upgrading Unit
Gv~sumslurry
Maaswater
Hot coolinqwater To river
Was h-water
60°C
ri
+1
Vacuum filter
I
+
612 Table 1. Chemical analysis of the two phosphQgypsums Constituents
CaO
%
Kemira Agro
Hydro Agri
dihydrate gypsum
dihydrate gypsum
32
32
(97-98% CaSO42aq)
(97-98% CaSO42aq)
P20~
%
0,45
0,1
SiO 2
%
1,2
0,8
AI203
%
< 0,1
< 0,1
Fe203
%
< 0,01
< 0,01
F (tot)
%
0,5
1,0
Cd
mg/kg
< 0,5
< 0,5
Hg
mg/kg
0,08
0,09
As
mg/kg
< 0,5
< 0,1
Pb
mg/kg
1,6
1,5
mg/kg
0,5
< 0,5
Zn
mg/kg
2,5
< 0,5
Cr
mg/kg
1,2
1
Cu
mg/kg
5
< 0,5
Ra-226
Bq/kg
500
450
Th-232
Bq/kg
25
< 8
Table 2. Analyses of phosphogypsum before and after hydrocyclone purification Sample
Before
After (underflow)
Radium
Bq/kg
490
284
Thorium
Bq/kg
15
13
613
Table 3. Upgraded synthetic gypsum Parameter
Formula
Unit
Value
H20
Wt%
<10
CaSO4*2H20
Wt%
> 97
1.
Free Moisture
2.
Calcium Sulphate
3.
pH value
4.
Colour
White
5.
Odour
neutral
6.
Magnesium oxide, water solube
MgO
Wt%
< 0,1
7.
Sodium oxide, water soluble
Na20
Wt%
<0,1
8.
Potassium oxide, water soluble
K~O
Wt%
< 0,02
9.
Chloride
CI
Wt%
< 0,01
pH
>5
10. Toxicity
non-toxic
11. Organic substances
2)
12. Phosphorus pentoxide, total
P20~
Wt%
< 0,3
13. Phosphorus pentoxide, water soluble
P20~
Wt%
< 0,01
14. Fluoride, total
Wt%
< 0,5
15. Fluoride, water soluble
Wt%
< 0,03
i~m
> 60
mg/kg
<1
18. Heavy Metals (E Pb, Cr, Ni, As, Cu, Zn, Hg)
mg/kg
<10
19. Radium
Bq/kg
< 120
20. Thorium
Bq/kg
< 20
21. Kalium
Bq/kg
<10
16. Dso 17. Cd
2)
Cd
actual value to be specified but will have no negative impact on setting time and/or colour
614
Photo 1. Particles in the phosphate rock containing Strontium and Barium
615
Photo 2. Particles in the phosphoric gypsum containing Sr/Ba
This Page Intentionally Left Blank
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
617
VALORIZATION OF LEAD-ZINC PRIMARY SMELTERS SLAGS D. MANDIN*, H.A.VAN DER SLOOT**, C. GERVAIS***, R. BARNA***, J. MEHU*** *METALEUROP RECHERCHE (TRAPPES, FRANCE) **NETHERLANDS ENERGY RESEARCH FOUNDATION (PETTEN, THE NETHERLANDS) ***POLDEN & L A E P S I - I N S A (LYON, FRANCE)
1. INTRODUCTION Lead and zinc primary smelters in Europe produce and dump every year about 1 Mt of slag. The industrial objectives of the Brite Euram project on which we will report here are to valorize the primary lead and zinc slags, and to avoid their dumping through their use in safe construction and civil engineering materials representing high tonnages. To maintain its competitiveness regarding overseas competitors, the European non-ferrous metallurgy industry has to avoid the burden of dumping cost, and must hence develop beneficial uses for its slags. While in recent years, significant improvements in both "conventional" (lead blast furnace or ISF) or "new" (Kivcet - QSL) processes have been achieved, the solid slags generated by these processes are still largely dumped. The five smelters operate the processes mentioned here under" .......................................................................................................................................................................................... .........................................................................................................................................................................................
SMELTER
i! .. ..
COUNTRY
FURNACE TYPE
.......................................................................................................................................................................................
BRITANNIA ZINC METALEUROP
:::: .. ii :.
United Kingdom
IS F
i!
France
ISF
i::
Germany
ISF
Italy
ISF
:: i:
MIM Htittenwerke ENIRISORSE (Temav)
. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
BERZELIUS St61berg METALEUROP ENIRISORSE (Temav)
Germany
QSL
France
LBF
Italy
Kivcet
. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .........................................................................................................................................................................................
ii !~ = .. .. ..
~ ii :~ :.: :: ..
:: ..
i,!
The main goal of the work is to work towards a common approach and understanding of the leaching properties of "metallurgical slag". This implies a better understanding of the phenomena involved in the leaching process in order to be able to adapt, if necessary, the metallurgical process to produce a slag which chemical composition, particle size and morphology, will result in reduced leachability. And further to develop cost effective processes and techniques to use the non-ferrous slags in environmentally compatible construction and civil engineering materials, as a partial or total substitute to quarry raw materials. Slags from other industrial processes such as blast furnace slag, steelslag and phosphate slag are used in coastal protection and in roadbase applications [1,2]. The environmental consequences of these applications have been shown to meet strict utilisation criteria.
2. E X P E R I M E N T A L 2.1 Characterisation and leaching behaviour To reach the objectives, slags from the seven plants under consideration were sampled in such a way as to obtain three samples representing different operating conditions for each furnace. The sampling of the slag was carried out over one month to obtain homogeneous lots both for studying the leaching behaviour and for studying the material treated for beneficial application. All 21 samples were
618
subjected to full chemical characterisation, determination of relevant physical properties, determination of fine mineralogical composition and leaching according to the most commonly applied leaching procedures in Europe. The leaching tests applied are: the French X31-210 [3], the DEV $4 [4], an acetic acid extraction similar to the EP tox [5], a modified version of the Swiss TVA [6] and the British repetitive shaker test [7]. To gain information on the factors controlling release from Pb/Zn slag tests addressing specific aspects of leaching have been applied such as a test to assess the pH dependence of leaching [8] and experiments to evaluate the redox properties of the slags [9]. Specific aspects of leaching addressed are the particle size distribution, the mode of stirring and method of filtration, the sensitivity to the liquid/solid ratio as obtained from a column test, test performance in closed bottles and exposed to the atmosphere and finally special exposure conditions such as sea water and acid rain water. The role of the chemical speciation of elements in the slag matrix and its influence on the leaching properties of slag is addressed in two ways: by analysing the leaching behaviour of pure phases assumed to be present in the slag matrix and by geochemical modelling [ 10] using the data from the pH static leaching test as input. The ageing of slags covers changes due to carbonation, oxidation and slow changes in mineral phases. These aspects are covered by laboratory testing and field exposure measurements. Slag has been exposed to the atmosphere for more than a year to observe changes in leaching controlling properties. 2.2 Improvement of the slags by metallurgical parameters
A better understanding of the factors controlling the slag leaching behaviour can lead to measures to improve the environmental properties of slag by controlling relevant parameters in the metallurgical process. For this purpose, the leaching data generated in the first part of the work are evaluated against the operating conditions of the different furnaces, as they do not run under the same operating conditions. Aspects addressed in this part are related to the basicity of slag, the occurrence of metallic Pb and to the level of crystallinity of the slags. In full scale operating facilities, tests have been performed in order to decrease the metallic lead content in the slags. In addition, experiments on slag cooling and granulation have been carried out, as these bear on the level of slag cry stall in ity . 2.3 Development of beneficial applications of slags
The following options for beneficial use of slags as substitute for natural raw materials have been evaluated: Aggregates for concrete (partial sand replacement) Structural fill in embankments and use as road construction materials Base materials for brick and tile manufacture For different applications different specifications may apply. First the technical suitability of the slags for a given application needs to be determined. When this proves to be the case, the material can be prepared and its actual properties tested. Then the environmental aspects of use of the material in construction will be addressed. For the identification of technical properties end users are involved in the programme. - Concrete
In the case of application in concrete, the concrete properties will be evaluated at sand replacement levels of 0 (reference), 10, 50 and 100 %. The durability of the concrete is evaluated through measurements of freeze-thaw resistance, alkali-aggregate reaction, long term and autoclave expansion. The leaching behaviour of slag-based mortars and concretes is evaluated using leaching tests focusing on the properties of the intact product rather than studying size reduced material. The major pollutants will be monitored (e.g. heavy metals, Ba, sulphate)as well as major elements or parameters relevant for the leaching behaviour of the material (A1, Ca, Na, K, pH). - Application on embankment
and road construction
In road construction secondary materials can be applied in different layers of the construction. The investigations have been carried out to compare the properties of the following mixes containing slag : Road cover ~ surface bitumen concrete with slag addition as gravel Road base ~ slag treated with cement and lime, ~ slag concrete with cement and gravel, ~ asphalt concrete with lime gravel.
619
Following the verification of the technical performance of the mix designs, field demonstration has been carried out at a level of 600 m 2 in 5 sections of different composition. - Application in bricks and tiles The possible application of slag as sand replacement in a special bricks and tiles manufacturing process is based upon transformation of a lime-sand mixture in an autoclave process. The performance has been studied at laboratory scale with various quantifies of slag as a partial substitution of sand. After demonstration of satisfactory performance, an experiment on a larger scale has been performed at a sand lime bricks plant. - Study of environmental performance of slag containing construction products The environmental performance of the slag containing construction products is verified in long term laboratory leaching tests (3 months, 6 months, 1 year) under a number of exposure conditions, such as acid rains, alkaline waters, oxidizing and reducing conditions, salted and fresh water, water saturated testing and discontinuous exposure to water. Pilot scale tests have been designed and manufactured in close collaboration with the final product producers and users.
3. R E S U L T S 3.1. Slag characterisation and leaching behaviour 3.1.1 Physical and mineralogical properties of slag Through proper collection of 21 lead and zinc primary smelter slag samples from smelting installations in 4 European countries, it has been possible to cover a broad spectrum of slag qualities from both a chemical and mineralogical point of view. Zn assays ranging from 5 to almost 12 % and Pb assays ranging from 0.7 to 5 % have been measured with one extreme sample of 14% Zn and 13% Pb. All slags exhibit a same general size distribution trend, covering the 0,2 - 4 mm size range, with very few materials finer than 100 mm. This renders them similar to natural washed sands obtained in quarry operations. Mineralogical investigations have shown that granulated slag consist all largely of an Fe-Ca bearing aluminosilicate glass, as reflected by the presence of moderate amounts of S iO2, FeO and CaO, together with minor amounts of A1203. Both in lead and zinc slags, metallic lead droplets are embedded in this aluminosilicate glass phase and are largely inaccessible. Accounting for less than 1 % of the slag material, their diameter varies from one slag to the other, probably depending on the viscosity of the slag and on the performance of the forehearth.
3.1.2 Comparison of national leaching tests for lead and zinc slags The national leaching test procedures for all slag from the different facilities have been placed in perspective by plotting them in relation to the pH dependent leaching data as obtained from the pH static leach test. In the case of non-controlled pH conditions, the pH of solution is dictated by the slag itself and evolves in the neutral to alkaline range (pH 7 to 10). The higher values reflect the behaviour of slags with an elevated lime content. In this pH domain the leachability of Pb and Zn is generally moderate to low 1 - 10 mg/kg. In the case of acidic pH, either under CO2 flushing or under acetic acid leaching, all analysed elements exhibit higher releases. In figure 1 all test data are given in relation to the final pH in the leachate. The measurement of pH in leachates is rather critical due to the low buffer capacity of Pb and Zn slag. Exposure to CO2 in the atmosphere may result in pH shifts that could lead to erroneous conclusions and less consistent behaviour than actually occurring. More stringent recommendations for leach test protocols would be required to obtain more reliable results. 3.1.3 Leaching behaviour of Pb/Zn slag - Acid Neutralisation Capacity. The availability for leaching according tot NEN 7341 [11] is carried out to assess the distribution between potentially assessable chemical phases and mineral incorporation of elements. For Pb and Zn, in particular, the difference between total composition and availability illustrates the degree to which these elements are incorporated in the matrix. An additional property derived from this test is the the Acid or Base Neutralisation Capacity (ANC/BNC) that can be derived from the acid or base consumption to reach the preset pH value of 7. For Pb/Zn slags this in the order of 0.0015 - 0.007 Mol/kg.The relatively low Acid Neutralisation Capacity compared to other types of industrial slags imply that Pb/Zn slags do not significantly affect their surroundings in terms of pH control. On the contrary, the pH of the surrounding medium is important in assessing slag
620 utilisation/disposal scenarios. Application in acidic environments should be avoided. 1000
100
~
pHstat Pb Pb slag O Zn slag .... 9 pHstat Zn - - ~
k
A
~,XN~~ JI_ON_~" ~
I ~ "6
8
0., T
0.01
1 z1
-
5
6
7
. ~ k o ~ _ ~162
8
pH
9
10
1000 -rIf, 100+ ~
~o
8
o ~
,
~
&O "
\\\k
o.1
0.01
5
6
.
7
O0
,
~176
~
4
12
13
...m.... pHstat Pb tx Pb slag o Zn slag +pHstatZn
"C ~4-
11
~
8
9
~
/?
/'7
~
10
Pb
11
12
13
pH
Figure 1. Results of pH dependent leaching behaviour for Pb and Zn from Pb and Zn slags in comparison with regulatory test data (X31-210, DEV $4, EP tox and TVA).
621
- Reducing properties of Pb/Zn slag. Using a recently modified procedure [ 12] the reducing properties of Pb and Zn slags has been assessed. At low pH (pH < 5) the slag material exhibits clearly reducing properties, as the redox potential is more than 50 mV below the pH- Eh curve for normal oxidised water. At pH 7 - 11, however, the reducing properties are not apparent. This is in strong contrast with types of reducing slags such as steel blast-fumace slag. This type of slag has been shown to exhibit reducing properties over the entire pH range, which was attributed to the leaching of reduced S - species as main carriers of reducing capacity (S partly presents as the relatively soluble CaS)[13]. In the case of Pb and Zn slag the leachability of sulphur species is very limited. S is apparently tied up in insoluble mineral phases (e.g. ZnS, PbS, FeS). In the case of Pb and Zn slag, the reducing properties are governed by reduced Fe and Mn species. Since these constituents show a very strong pH dependence in their leaching behaviour[13], this is reflected in the measured reducing properties of the slag as a function of pH. pH dependence of leaching. To assess the behaviour of slags as a function of pH, pH static tests are carried out using a liquid to solid ratio of LS = 10 and pH control in the pH range from pH - 4 - 12. A pH controller is steered by computer dosing HNO3 or NaOH depending on whether acid or base is required. The amount of acid or base is kept small to avoid significant change in the initial liquid to solid ratio. In figure 2 the pH dependent leaching behaviour of A1, Cu, Cr, Mn, Pb and Zn is given. The general trend in the leaching behaviour has been found to be very similar for a given element almost irrespective of its origin. It reflects the degree of solubility control by mineral phases at the surface of the slag. The difference in leaching levels between slags is in part due to the alkalinity of slags. This consistent behaviour can be exploited further to optimize slag properties for use or disposal, if necessary. - Leaching as a function of liquid to solid ratio (L/S). Leaching in a column reflects the behaviour of a material under percolation conditions. For this purpose the column test according to NEN 7343 [ 14] has been carried out covering the liquid to solid (LS, 1/kg) range 0.1 - 10. This tests gives an indication of leaching behaviour on the long term. In figure 3 the column test results for Pb are graphically presented as a function of LS. -
1000 . - - ~ . : - - - : . . : . - - - . . . . : - - . - . - : . - . - . - : - - . . . . - : . - - . - . - - : - . - . . ~ . . . - . - ~ - - . . . . - - ~ - . - . - - - : . . . - - ~ - . -
....
[5,'
Pb slag 1
C.
Zn slag 1
~,
Pb slag 2
~"
Zn slag 2
O
Zn slag 3
0.0
~.
100
E ~ r~
Pb
lO
~
1
~
0.1
D
D D
V
',:7 O
V O
O ~7
L-1
O
0.01 0.001 I 0.03
I
I
I II
I
0.1
I
I
I
I III1
|
1
1
i
i
i
i i
10
Liquid/solid ratio (LS) Figure 3. Column test data of Pb for four Pb and Zn slags showing cumulative release expressed in mg/kg covering the liquid to solid ratio 0.1 to 10. The availability test data are included as horizontal lines.
622
10
1000
AI
Gibbsite
g
1012
rite
g g
lC
g
o
1
o
~
8
0.1
8
C
o
~
0.(
0.01
o
0.001
I
~
5
7
pH
I
I
9
11
0.0( 5
13
7
pH
9
11
13
1000
0.01
MnHPO4
100
Mn
10
0
Cr(OH)
O
"~,
~ 0.001
i
0
o
8
O
3
lo
I
I
I
5
7
9
11
pH
o
0.0001 13
t
i
t
5
7
9
pH
o
I
r
11
Cerrusite ."
I OOC
100
10C
O
lC
10 ""
. -~
o
0.1 8
0.01
1000
I O00C
~
- ..~
0.001
o
0.0001
~,,,, ....
0.1
o
0.01
8 0.1
0.001
Zn
Pb
0.0001 3
I
I
I
I
5
7
9
11
pH
13
0.01 3
I
t
I o
5
7
9
pH
t
11
13
Figure 2. pH dependence of leaching and geochemical modelling of leaching behaviour from Pb/Zn slag (using MINTEQA2). The solubility control of the metals implies that the release rate will be constant over time. In column tests Pb shows a continuous increase of the cumulative release with time, which is consistent with the identified release mechanism. For comparison the availability for leaching as obtained by NEN 7341 is indicated for each of the slags. The leaching behaviour differs significantly for the slags studied. The pH of the slags with the highest leachability for Pb is either rather low (pH 5) or high (pH 10).
623
- Leaching behaviour of pure phases. The concentrations dictated by solubility pure Pb phases has been modelled. In figure 4 the modelling data are presented. 10000
.........
'z.-. . . . . . . . . . . . . . . . . . . . .
~
~
\ 1000 0 ~.~ ~
100
\
~" \,
\,, \ \
\
"'. ',
\.
\k
10
/
-,
\\o
~ii-.. H , ,
2.
d, o
.... 7 ' /
Pb(OH)2 .....
0,1 O r~
. . . . . . . PbSiO3 .... A
0,01 0,001 0,0001
PbO
Pure Pb phases ..
~
|
5
7
-
, 9
pH
--
~___ 11
,
13
PbCO3 sPbSiO3
[]
sPbO
~::
Pbrmtal
~i
Aged Pbmetal
o
sPbCO3
Figure 4. Modelling and measurements of Pb leaching from pure Pb phases. At Metaleurop leaching of pure Pb phases has been carried out using the French 3-step leaching test [3]. These data have been put into the graph for comparison. The leaching data for PbO correspond well with the leachability predicted for PbO, allthough the first two steps of the leaching test are undersaturated with respect to PbO solubility (slow kinetics of dissolution?). The leaching behaviour of Pb metal corresponds to that of PbO, which is not very surprising as Pb metal oxidizes relatively fast. Aged Pb metal matches in the first step with PbO, but in later steps the solubility moves in the direction of PbCO3 solubility. The solid PbSiO3 leachability does not correspond well with the PbSiO3 solubility curve. It matches better with either Pb(OH)2 (or PbCO3). The solid PbCO3 solubility as measured in the leaching test corresponds reasonably well to the modelled PbCO3 solubility (slightly supersaturated), except for the first leaching cycle that has a rather high concentration. The pure phases seldom remain pure as external influences may convert the surface of pure Pb phases to other forms, which may then become solubility controlling. The fact that pure PbSiO3 corresponds closer to PbCO3 than to the silicate, can be caused by a small contribution of leadcarbonate in the silicate. Since it is the least soluble phase, it may become the controlling phase. Based on these modelling and laboratory observations, the following mechanism can be derived. When metallic Pb droplets occur embeded in slag or as individual particles among slag, the leaching behaviour of Pb will initially reflect the behaviour of PbO as Pb oxidizes rapidly. This stage does not last in contact with the atmosphere, since due to interaction with CO2 from the air PbO will be react to PbCO3. This is the more stable phase upon aging of slag. If this mechanism proves to be correct, the question can be posed whether testing the fresh reactive slag is representative for slag leaching behaviour in environmental exposure conditions. Due to the unknown progression of the oxidation/carbonation, the Pb leaching results on fresh slag must be quite variable. This is in agreement with the observations. It might be enlighting to verify for slag with a high Pb leaching level in relatively fresh slag the change in Pb leachability with time. - Role of ageing in leaching from slag. To study the consequences of ageing of slags on the leaching behaviour through exposure to the atmosphere and moisture, three containers have been set up in the laboratory in which at three levels porous cups have been installed to be able to analyze porewater pH and EH with time. One container (h -- 0.7m, O - 0.4 m) was filled with slag and kept fully saturated, the second was intermittently wet and dry and the third was kept wet under a fine sand cover. After one year the porewater in the bottom of the container was analyzed for a range of major and trace constituents to be compared with the pH-stat leaching information. In the permanently wet container the uptake of 02 and CO2 is slow and consequently the oxidation/carbonation proceeds relatively slow
624
compared to the other configurations. In the intermittently wet/dry cycling container, the changes are most pronounced as the interaction with 02 and CO2 are largest (gas phase diffusion 104 times faster than in water). In the container with a fine sand cover the change in pH and redox is least affected as the uptake of CO2 and 02 is further restricted. For the most part the conditions are oxidized. Mildly reducing conditions can develop in systems that are relatively closed from the atmosphere. The pH may remain around pH - 10 for a long time depending on the degree of exposure to the atmosphere. The site specific conditions determine which condition of the one tested here will prevail. In general, the consistency of the leaching data on aged slag with the pH-stat data is very promising. - G e o c h e m i c a l m o d e l l i n g o f slag leaching behaviour. Using the geochemical speciation model MinteqA2[ 10], the leaching data obtained from the pH stat experiments on Pb and Zn slagshave been modelled to identify the potential solubility controlling phases. This information is relevant for decisions on possibilities for modification to improve slag leaching characteristics. It also helps to identify the influence of the degree of amorphous glass versus crystaline phases. The leaching data obtained for major, minor an trace elements have been transformed to mol/1 concentrations and used to calculate saturation indices for all possible mineral phases. Based on the evaluation of the saturation indices, those phases that are likely controlling phases due to a good match between measured concentrations corrected for activity (Debye-Huckel) and solubility of specific phases over all or a part of the pH range studied have been selected. In figure 2 potential solubility controlling phases have been inserted. L e a d - T h e most relevant solubility controlling phase for Pb solubility from Pb and Zn slag appears to be cerrusite - PbCO3. This is consistent with the observations on pure Pb phases. Only in case of ISF Gran 3 solubility control of Pb by PbO and PbSiO3 matches better with the observed solubility curve than cerrusite. XRD measurements have revealed the presence of metallic Pb droplets and PbO in this sample. Z i n c - T h e solubility of Zn is largely matched by the solubility dictated by zincite - ZnO. - M i n e r a l o g i c a l aspects. T h e basicity of slags, which is reflected in the leachable lime content, dictates the final pH in the leachate to a large extent. As has been shown before pH is an important controlling factor off slag leachability. The mineral phases identified as solubility controlling phases are not the same as those identified as bulk mineral phases by XRD of the total slag samples. This illustrates the point that leachability is controlled by mineral phases at the surface of slag particles, which is not necessarily the same as the mineral composition within the slag panicles. In the case of the Pb phases this is already clear, but for other element the same applies (Table la and b). Table la. MINERAL PHASES IDENTIFIED IN THE SLAG BULK MATRIX SATURATION INDICES pH 12.0 10.5 9.5 8.5 7.5 6.5 4.14 6.77 8.10 14.72 10.60 7.72 FRANKLINITE -0.44 -2.31 -3.70 -5.20 -6.66 -8.22 P-WOLLSTANIT -11.32 -12.89 -14.75 -18.30 -20.74 -25.30 GEHLENITE WUSTITE 1.84 2.44 1.27 -0.50 -2.47 -4.47 WURTZITE 2.55 3.64 3.77 2.63 2.18 1.77 -6.51 -4.25 -3.77 -4.90 -5.10 -7.38 ANORTHITE HERCYNITE -2.13 1.83 2.11 0.26 -0.84 -3.92 FECR204 9.29 12.81 12.26 9.49 8.29 5.02
5.5 10.72 -9.21 -29.21 -6.51 1.14 -9.13 -7.32 0.30
4 9.28 -10.55 -33.18 -9.71 0.19 -9.87 -10.86 -9.59
Table 1b. SOLUBILITY CONTROLLING MINERAL PHASES BASED ON GEOCHEMICAL MODELLING USING LEACHATE ANALYSIS DATA SATURATION INDICES pH 12.0 10.5 9.5 8.5 7.5 6.5 5.5 4 Fe(OH)2 0.04 0.84 -0.28 -2.04 -4.01 -6.01 -8.05 -11.26 NORSTRANDITE -2.08 -0.51 0.20 0.15 0.58 0.05 -0.63 -0.80 QUARTZ -2.58 -1.60 -1.13 -0.67 -0.28 0.08 0.67 1.61 PB(OH)2 (C) 3.49 2.51 0.58 0.05 0.13 0.21 - 1.23 -3.50 CERRUSITE -2.70 -0.56 -0.95 -0.36 0.22 0.78 -0.58 -2.98 ZN(OH)2 (E) -0.506 -0.927 - 1.576 -0.916 -0.916 -2.086 -3.727 -5.97 Note: Values close to zero indicate solubility control; indices are given in log-scale units.
625
- C o m p l i a n c e t e s t a n d c o n c i s e test. For quality control purposes, it is important to have short procedures that allow a quick verification to determine that the material is consistent with a previous characterisation. In CEN TC 292 a compliance test for granular materials has been developed[ 15]. It consists of three extraction options. Here the two step option LS - 2 and LS - 2 - 10 has been applied to be able to relate the test results to the column test data. In addition, two pH controlled experiments were carried out as described in the proposed concise test for granular materials[ 16]. The pH conditions applied are pH 4 and pH 12 (LS= 10). In figure 5, the data are plotted as a function of pH and compared with the pH static information for the five slags tested. The data show a good correlation between pH dependent leaching behaviour and the individual measurements in the CEN test and the added pH controlled tests for the concise test. The markedly different behaviour of GRAN 3 is also recognised in the concise test. For quality control purposes, the relevant condition for a given application can be selected.
1000C
1000 %
Zn
X
100(
%
100
6'
%
& o'J
Q
%
10C
I=
& & O
%
tO
~ lC
,.. ,..I
0 [] o ,&&
c-
o
e-
% A
E 10
r
_o
A
!._.
%
%
|A
~
o
e-
t@
~
o
o
o
0.1
0.01
I 3
5
[]
~ 7
I
pH
9
0.1
Pb
n
% 0.01
I 11
13
3
o A
I
I
I
5
7
9
pH
O
&
I 11
13
Figure 5. Correlation between CEN compliance test, concise test and pH static leach test data for five different slags.
3.2 Improvement of the slags by metallurgical parameters The main goal of the work is to utilize the slags in civil engineering applications. This goal should preferably be reached without alteration of the process. However, if the leachability would prove to be too high for utilization of slag as a substitute to raw materials, it could be necessary to modify some aspects of the process. The leachability of Pb is considered most critical in this respect as metallic Pb will have an influence on leachability. The following measures have been tested to reduce the metallic lead content: - selective settling of lead droplets - transforming of lead in a more stable phase e.g. silicate - diminishing the accessibility The two experiments involving lead blast furnace slag must be considered as preliminary. The first focused on slag tapping at higher temperature and the second on re-melting and cooling at different speeds.
626
3.2.1 Slag tapping at higher temperature Two forehearth operations have been conducted, each of them being achieved at full scale over a three shift period. The following results were obtained: Total lead release NFX31-210 (mg/kg) Normal forehearth operation - slag temperature 1185 ~ 5.4 Overheated forehearth operation - slag tapping temp. 1210 ~ 1.3 The decrease in lead leachability is not directly linked to an improvement of lead droplet settling. It rather appears to be related to a modification in slag mineralogy.
3.2.2 Re-melting and cooling The cooling mode has been studied at the laboratory. After re-melting the slag was submitted to three cooling modes: rapid cooling through water quenching, moderate cooling through tapping into ingot and slow cooling by controlling the temperature decrease at 2 ~ The following results were obtained: Treatment Lead leachability (NF X 31-21 O)
Reference slag 4.0
Rapid cooling 0.7
Moderate cooling Slow cooling 0.4 0.2
To explain the observed leaching behaviour three mechanisms can be identified: settling of metallic lead, transformation of metallic lead in lead silicate and modification of the crystallinity of the slag. Experiments are underway to identify which mechanism is prevails.
3.3 Environmental assessment of utilization scenarios of slags 3.3.1 Materials and recipes On the basis of technical and environmental criteria, a preliminary study allowed the choice of a substitution level of aggregates by slags for each type of material manufactured by the related industrial partners. The recipes retained for environmental evaluation are presented in the following table. materials matenals
recipes reference
[ 1
concrete kg/m3
J
~
, sand gravel cement water
792 1010 320 200
slags 11% ISF slags M6taleurop
slag sand gravel cement water
81.4 732 1010 320 200
breeze blocks kg/m3
brick
sand-cement
(%)
(%)
sand -bitumen
(%)
L sand gravel cement water
1197 803 200 130
sand lime
90 10
slags 15.2 % slag sand gravel cement water
granulate road binder
94.5
granulate bitumen
5.5
slags 47 %
164 1075 803 200 130
100 4.5
slag granulate road binder
47 46.5
slags ] granulate[ road binder
47 46.5
5.5
slags 50 % slag granulate filler bitumen
50 47 3 4.5
~
%
slags 8-15 % ISF slags Britannia Zinc LBF slags
M6taleurop
~
slags sand lime 11% 81.7 736 1010 320 200
sla_9_g~16 % slags I "]72 sand [ 1075 gravel ] 803 cement[ 200 water I 130
I 8-15 I 75-82 I 10
5.5
granulate 147 filler [3 bitumen 14.5
3.3.2 General presentation of the methodology The environmental validation of the obtained materials was carried out by studying leaching behaviour of the materials under specified conditions. The methodology described in the French experimental standard X30-407 and the methodological guide of the WG6 of the CEN/TC 292 were used : The
627
valorization scenarios were described in terms of influence factors. The parametric tests were used to measure the respective influence of these factors (pH) on the leaching behaviour of the materials (solubility, dynamics and release intensity). The limits of our impact study are : The materials are observed in only one stage of their life cycle (that for which the materials were designed) migration in the immediate environment (soils, surface waters, underground water) of the leached constituants from the site of use is not followed up toxicity for man or ecological impact on the flora and fauna as a secondary effect is not evaluated The aim is essentially to measure the flux of pollutants in the considered scenarios
3.3.3 Presentation of the study The following table presents the scenarios and most probable use criteria for each material.
Materials
concrete
scenarios
wall
breeze blocks
brick
wall
wall
v'
v'
foundation
in contact with the air
sand cement
sand bitumen
base layer
flooded not in contact with the air normal road
v' v'
road in contact with underground water
3.3.4 Presentation of the experimental programme Analysis of the scenarios allowed us to retain influence criteria whose effects were evaluated using the parametric tests presented in the following table"
~ e n a r parametric tests
leaching on crushed samples
i
o
s
in contact with air
flooded not in contact with air
normal road
road in contact
with underground water
acid neutralizing capacity solubility according to
pH
test at equilibrium carbonation with demineralised water leaching on with a synthetic monolithic solution of porewater material with acid solution coupled with a wetting/drying cycle
*
*
The programme consists of several levels of environmental evaluation. Level 1 9 Parametric tests in the laboratory to evaluate release while varying one influence parameter (e.g. solubility tests measure variation according to chemical context).
628
Level 2 :
Pilot tests in real environmental exposure conditions integrating the variation of climatic factors. There are two sizes : Level 2A :small size pilots ( l m 2) for concrete, breeze blocks, bricks, Level 2B: large size pilots (150 m 2) for road construction materials. The levels of exposure tested for the different materials are summarized in the following table : Materials
concrete
"s~narios experimental ie3~al~ level 1
~
level 2A
wall
foundation
breeze blocks wall
bricks
ISF BZ
wall
ISF MER
ISF MER
ISF MER
LBF MER
LBF MER
LBF MER
ISF MER
ISF MER
level 2B
sand cement normal road
sand bitumen
flooded road
normal road
flooded road
ISF MER
ISF MER
ISF MER
ISF MER
LBF MER
LBF MER
LBF MER
LBF MER
ISF BZ ISF MER
ISF MER
3.3.5 Presentation of the pilots Level 2A : It concerns concrete, breeze blocks and bricks. For each material there is a reference recipe and a recipe with the slags ISF provided by Metaleurop (concretes, breeze blocks) or Britannia Zinc (bricks). The pilots have been installed since October 1996.The materials provided in monolithic form were assembled in a wall with a surface of lm 2 and arranged to slope at 30 ~ in relation to the vertical. The lower surface of the wall is covered with a waterproof membrane. A system collects the run-off and percolation waters. Each sample, which is collected according to the arbitrary climatic conditions, is analysed (lead, zinc, calcium, sodium, potassium, filtered and the volume and pH are measured. The follow-up is in progress. Level 2B : It concerns road construction material. For each material there is a reference recipe with slag ISF from M6taleurop. The four experimental surfaces were made in B6thunes (Pas de Calais) by Colas and have been in operation since May 1996. The tested materials form the base layer of the road structure (just beneath the road surface layer). Each experimental surface is of 158 m 2. The following diagramme shows the structure as a whole :
surface n ~ 1 sand bitumen reference
surface n~ sand bitumen and slags
surface n~ sand cement reference
ISF MER
surface n~ sand cement and slags ISF M E R
surface layer draining bituminous concrete 0,04 m base
sand bitumen 0.24 m
sand bitumen with slags 0.24 m
layer sand cement 0.38 m
sand cement with slags 0.38 m
drainage layer in 10/20 limestone 0.05 m
foundation layer 9red schist 0.50 m
The following figure shows the collecting device of run-off and percolation waters.
629 slope 2 % bitumen border
i
4 cm
BITUMINOUS CONCRETE run-off water
I
SB: 24 cm or
SC:38 cm
5 Clff'~
BASE LAYER -SAND CEMENT - BITUMINOUS SAND
~
waterproof membrane
DRAINING LAYER percolation water
m
50 cm
L
FOUNDATION LAYER
GROUND
The level 1 laboratory studies are also in progress. The results of these trials and those concerning the pilots will be available in autumn 1997.
4. CONCLUSIONS The regulatory test methods can be placed in perspective by comparing the test results in relation to the pH static leaching test. The measurement of pH in leachates is rather critical due to the low buffer capacity of Pb and Zn slag. More strict control of the final pH in leachate analysis is needed for materials with very low buffer capacities as exposure to the atmosphere for even a relatively short time may bias the results in an unacceptable manner. The leaching behaviour of elements from Pb and Zn slags is generally very consistent. Exceptions can be related to significant differences in slag properties, such as degree of ageing and degree of crystallinity. If metalic Pb is present in slag or mixed with slag, fresh slag will show high Pb leachability due to high PbO solubility. Upon aging PbO converts to PbCO3, which in the geochemical modelling has been shown to be the most important solubility controlling phase in most cases. These observations lead to the question whether testing of fresh slag is the appropriate way of judging slag leaching behaviour for Pb for environmental assessment purposes. The geochemical modelling has revealed that solubility controlling phases are not the same as those observed in mineral analysis of bulk slag. The often new phases formed at the surface of a materials exposed to moist, atmospheric conditions dictate the leaching behaviour of the slag material in practice. An important difference has been observed for the controlling phase of Pb in ISF GRAN 3 where PbO and PbSiO3 are the most likely solubility controlling phases in contrast to cerrusite (PbCO3) in the other slags. Based on the present understanding of the leaching behaviour of slags, as produced, as a secondary raw
630 material in construction, conditions for beneficial application can be idemified. The conditions that need to avoided can be clearly identified and codes of practice for reuse can be defined.
Acknowledgement This work has been carried out in the framework of Brite Euram project BRE2-CT94-0585.
5. REFERENCES [1] D. Mandin. EC/CANMET Workshop Waste Minimization and recycling, Lisbon, 16-19 September 1996. [2] WASCON 1994: Environmental aspects of construction with waste materials Eds. J.J.J.M. Goumans, H.A. van der Sloot and Th. G Aalbers, Elsevier, Amsterdam, 1994. [3] D6chets: Essai de Lixiviation X 31-210, 1988. Association FranCaise de Normalisation (AFNOR), Paris. [4] DEV $4: German standard procedure for water, waste water and sediment testing - group S (sludge and sediment); determination of leachability ($4). Institiit fOr Normung, Berlin, 1984. [5] EPA Toxicity Test Procedure (EP-tox), Appendix II, Federal register, Vo145(98), 1980, 33127 - 33128. Govemment Printing Office, Washington D.C. [6] Bericht zum Entwurf fOr eine technische Verordenung fiber Abf~ille (TVA), 1988. D6partement F6d6ml de llnt6rieur. Switzerland. [7] AEAR R 10737 Equilibrium shaker test. [8] R.N.J. Comans, H.A. van der Sloot, P. Bonouwie. Geochemical Reactions Controlling the Solubility of Major and Trace Elements During Leaching of Municipal Solid Waste Incinerator Residues. Proceedings Municipal Waste Combustion. VIP 32. Air & Waste Management Association Pittsburg, Pennsylvania. 1993. 667-679. [9] H.A van der Sloot, D. Hoede and R.N.J. Comans. The influence of reducing properties on leaching of elements from waste materials and construction materials. In: Environmental aspects of Construction with waste materials. Eds. J.J.J.M. Goumans, H.A. van der Sloot, Th.G. Aalbers, Elsevier Science Publishers, Amsterdam, 1994, 483 - 490. [10] Felmy, A.R., D.C. Girvin, and E.A. Jenne, MINTEQ--A computer program for calculating aqueous geochemical equilibria, EPA-600/3-84-032, U.S. Environmental Protection Agency, Athens, (1984). [ 11] NEN 7341 Determination of the availability for leaching from granular and monolithic contruction materials and waste materials. Sept 1993. [ 12] Draft NVN 7348 Determination of reducing properties and reducing capacity of waste materials and construction materials. 1994. [ 13] H.A. van der Sloot. Developments in evaluating environmental impact from utilization of bulk inert wastes using laboratory leaching tests and field verification. Paper presented at "Bulk Inert Wastes: an opportunity for use", Leeds, September, 1995. Special issue Waste Management, 16 (1-3), 1996, 65-81. [ 14] NEN 7343 Determination of leaching from granular contruction materials and wastes by means of a column test. June 1994. [ 15] Compliance test for leaching of granular materials, CEN TC 292 Characterization of Waste, Working Group 2 Draft European Standard. June 1994. [16] H.A. van der Sloot, D.S. Kosson, T.T. Eighmy, R.N.J. Comans and O. Hjelmar. An approach towards international standardization: a concise scheme for testing of granular waste leachability. In: Environmental aspects of Construction with waste materials. Eds. J.J.J.M. Goumans, H.A. van der Sloot, Th.G. Aalbers, Elsevier Science Publishers, Amsterdam, 1994, 453-466. [17]CEN TC 292 Characterization of waste - Methodology for the determination of the leaching behaviour of waste under specified conditions ENV, March 1997.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
631
THE L O N G - T E R M ACID N E U T R A L I Z I N G CAPACITY OF STEEL SLAG Jinying Yan, Catharina Biiverman, Luis Moreno and Ivars Neretnieks Department of Chemical Engineering and Technology Royal Institute of Technology, S-IO0 44 Stockholm, Sweden A B S T R A C T . The long-term acid neutralizing capacity (ANC) of the electric arc furnace slag was investigated by batch pH titration, and the neutralizing processes of the slag were evaluated by reaction path modeling. Reaction time plays an important role for the determination of the ANC for the steel slag. The relatively slow reactions may give large contributions to the ANC for a long-term leaching process. pH-dependent reactivities of the steel slag were found in the high pH range. The neutralizing rates at high pH were much slower than that at relatively low pH. Below pH 9, the reaction rates became less pHdependent and usually fast. The features of neutralizing reactions of the slag may be explained by its mineralogical composition and dissolution kinetics in the neutralizing processes.
1
INTRODUCTION
Steel slag is one of the important sources of industrial wastes. There is a long tradition in using the slag as potential secondary raw material in constructions and other fields. Recently more attention has been given the waste materials of environmental concern. 1 The possible environmental impact of the slag has to be investigated in order to assess whether the slag satisfies the requirements of a respective application, especially on the release of toxic elements in the leaching processes of the waste materials. The pH environment of a given leaching system usually controls the dissolution of the waste matrix, and determines the release behaviors of many toxic elements. 2 The acidbase properties of solid wastes have been found to considerably influence the leaching of solid wastes by changing the pH environment. If a given aqueous system is well understood, the pH environment of the system can be defined by any two of these three parameters: (1) total dissolved acidic or alkaline substances, (2) the pH value of the system, and (3) the neutralizing capacity (acid neutralizing capacity (ANC) or base neutralizing capacity (BNC)). 3-4 Neutralizing capacity is, therefore, considered as one of the important acid-base properties of solid wastes. This capacity factor is significant for determining the shift of acid-base behavior of solid waste in leaching processes, and very useful for predicting the potentially reactive amount of solid waste in neutralizing reactions. Many attempts have been made to assess acid-base properties of solid wastes including the neutralizing capacities. 5-8 However, how to define this factor is still an open question because of the complicated neutralizing reactions that take place during leaching. The central problem is that traditional test methods may only account for relatively fast reactions that take place in the initial leaching processes in the relatively short term. The long time reactivity of a solid waste must be considered. For long-term considerations, we must investigate the influence of reaction time on the determination of the neutralizing capacity, and define the importance of kinetic factors for the neutralizing reactions in leaching processes. Many elements, minerals and even gas phases, as well as a wide pH range are involved in a heterogeneous waste system. Because of the complexity of the waste system, computer modeling should be one of the most useful approaches to give a good insight into the neutralizing processes. However, as current experimental methods do not take account of the long-term effect in the leaching test as mentioned above, most modeling work uses some simple geochemical models to find key factors that control the leaching
632 behaviors of solid wastes. 9-15 In these models, thermodynamic equilibrium was assumed by which release of a species was determined by the solubility of a controlling mineral. The stability of the mineral was then connected with the pH environment. The knowledge of the key minerals that controlled the leaching environment became essential to understand the leaching processes. These models can not be used in dealing with reaction paths and kinetics. In some circumstances, the reaction path may become very important and kinetic factors may have a large influence on the leaching processes. In addition, little work has so far been done to address the neutralizing capacity even for short-time leaching processes of steel slag. In this paper, we combine long-time experimental investigations with reaction path modeling to evaluate the neutralizing processes of the steel slag. Long-time experiments were used to investigate the time dependent neutralizing reactions, both short- and long-time neutralizing capacities and some kinetic information. The reaction path modeling was used to simulate the neutralizing processes. The simulations are performed over a pH range that may be interesting in the leaching of slag. The changes of solid phases and the chemical speciation in the aqueous phase were calculated in these simulations. The simulation results then were compared with experimental observations whereby a better understanding could be established for the complex neutralizing processes. 2
METHODOLOGY
2.1
Experiments
2 . 1 . 1 Material
The steel slag used in this study is a scrap metal based steel slag produced from the electric arc furnace process. The main chemical composition of the slag is shown in Table 1. The samples of slag were oven dried at 105 + 1 ~ to constant weight. In order to minimize the effects of diffusion in the particles, the slag was finely ground to particle sizes less than 0.160 mm. TABLE 1. Major elements in the steel slag Element
Ca
Mg
Na
K
AI
Fe
Mn
Ti
P
(mmol/g)
5.51
1.86 0 . 0 2 0.01 0.81 4 . 3 3 0 . 7 2 0.06 0.15
S
Si
0.04
2.05
(% by weight) 22.2 4 . 5 0 0 . 0 4 0.05 2.20 24.3 3 . 9 0 0.28 0.46 1300" 5.80 * The unit is ppm.
2.1.2 pH titration procedure and analysis of leachate Long-term batch pH titration experiments were carried out for the steel slag by using the automatic titrator (Metrohm 719S Titrino). The slag sample was mixed with pure water in a liquid to solid ratio 5:1 (40 g ground slag and 200 ml of water) in a plastic bottle. HNO3 solution (usually 1 M) was used as titrant. The acid was automatically added to keep a constant pH value (a given pH + 0.02). The experiments were performed at different pH levels, and were run from 4000 to 6000 hours. The acid neutralizing capacities (ANC) of the steel slag were determined for different titration times and for various pH levels. The leachates were analyzed after one week. The samples of leachate were taken from the batch experiments. The main cations, calcium (Ca2+), magnesium (Mg2+), sodium (Na § and potassium (K § were determined by using the DIONEX DX-300 Series Ion Chromatography System with suppressed conductivity detection.
633
2.2
Modeling of the neutralizing processes
2 . 2 . 1 Modeling concept The conceptual description and approach to conceive the solid waste-water reaction processes can be found in our previous work. 16-17 A reaction path model was used to simulate the pH titration process of the steel slag. The original theoretical approach was proposed by Helgeson et al. 18-20 to calculate a reaction path that involves one or more irreversible processes in which a series of successive partial equilibrium states may result in a state of local equilibrium for a reaction system. The basic idea used in the modeling of neutralizing reactions is to simulate the pH titration process as it is performed in the laboratory. A very small amount of slag is added to the water with known initial concentration. By using a geochemical reaction path model the rate of change of the water composition and the dissolution of the slag is calculated. The slag is modeled as consisting of its individual elements as found in the total analysis. The potential formation of secondary minerals is checked for from a list of many possible minerals the most supersaturated is precipitated first, then the next in line and so on. Some (more) acid and some more slag are then added and a new reaction path is calculated superimposed on the first. The procedure is repeated until all the interesting pH range has been covered. The whole titration process is considered as an irreversible process, but partial equilibria between solid phases and aqueous phase are assumed to exist.
2.2.2 Computer code A geochemical computer code--EQ3/6 was employed to calculate the reaction path of the titration process. 21 In these calculations, solid waste and acid solution were treated as two special reactants. Relative reaction rates between solid waste and acid titrant were handled by a relative rate law, 22 and by defining the rate constants for these two special reactants respectively. An internal model is included for complex reactions and calculation of activity coefficients for the species in the aqueous phase. The main results of modeling are given in the distributions of species (including proton) in both solid and aqueous phases during the simulations, and equilibrium states of the reaction system along a certain reaction path.
2 . 2 . 3 Modeling system For the modeling of the neutralizing reaction, the steel slag was represented by a fourcomponent CaO-MgO-SiO2-A1203 system. We also include oxides of sodium and potassium, as well as carbonate and sulfate, so that the original chemical composition of the slag could be represented. This is also reasonable for the mineralogical composition of steel slag. 23-24 In the reaction path modeling, these four components and relative reaction rates of steel slag were taken as main modeling parameters. Other minor constituents of the slag were usually considered to be less important for the neutralizing reactions. The system was assumed to be closed so that the interactions of the system with the atmosphere were ignored. In addition, the system was in an oxidative environment as the titrant was HNO3. According to the experimental leaching characteristics of the steel slag, only a part of the steel slag reacts in the neutralizing processes. The initial guess for the reactive fraction can be taken from special availability experiments of various components. The final reactive fraction of the slag will be obtained from the modeling by curve fitting in order to get the best agreement with the experimental results. In the simulations, equilibrium was maintained among aqueous species and potential minerals. Although the computer code can automatically make optimum-choice from the database based on the phase rule, some adjustment was always required. The problem is related to the formation of the secondary minerals. Those that would form from purely thermodynamic consideration are often not minerals that are observed to form under similar conditions. Rarely are the secondary minerals determined or sought for in leached slags. We have therefore tested if the choice of secondary minerals is crucial. It has been found that an individual mineral is not crucial, although the mineral assemblage plays important role in the neutralizing reactions. 17 This can be understood considering that
634
m a n y of these minerals consist of the same building blocks, namely a l u m i n u m oxide, silicon oxide, calcium oxide and so on but in different proportions and with different cations. The choice of the secondary minerals is thus to some extent subjective but in most cases we have studied the simulated titration results are quite similar.
3
RESULTS
AND
3.1
Long-term
acid
DISCUSSION neutralizing
capacity
of
the
steel
slag
The experimental results for the long-term acid neutralizing capacities of the steel slag are s h o w n in Figure 1. T w o obvious features can be seen in this figure. One is that the changes of the A N C with time mainly appear in the experiments until about 500 hours. Another feature is that most of the acid neutralizing capacities are found in the pH range above 8.5. This means that there is a large buffering capacity in the relatively high pH range. It is less in the weakly alkaline or neutral pH range.
l0
Steel Slag i....-/
, ...................................................................................................................
. .~..';::.:.: ...............................
~i"-_~'~ ........................................................................
/
r
.,.,o- ...................
pH=9.0 ". . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
./
4
,~
pH=7.0
! pH=9.5 f. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2
~1~
pH=10
]... . . .
t ....
0
I ....
I
t ....
1000
....
I
, ....
2000
....
I
, ....
3000
4000
0a)
Time
FIGURE 1. The acid neutralizing capacities (ANC) of the steel slag for different pH levels and reaction times.
100 -&,
80 -
"',~ ..,%
6 0 - ~- ....... 5teel Slag
~
40. ~
24 hours ......... 168 hours - - 500 hours
20.
0-
I ....
7.0
I ....
7.5
I ....
8.0
I ....
8.5
I ....
9.0
I ....
9.5
I
10.0
pH
FIGURE 2. The amount of ANC needed as a function of the pH levels for different reaction times for the neutralizing of the steel slag
635 Because the acid neutralizing capacities for various pH levels only have very small changes after 4000 hours as shown in Figure 1, we may take these values as the approximately long-term neutralizing capacities of the steel slag. The relationships among the reaction time, the pH levels and the acid neutralizing capacities are shown in Figures 2 to 4. Figure 2 quantitatively indicates the importance of the titration time on the determination of the ANC. The 24-hour experiment only measured less than 60 percent of the long-term ANC for this kind of solid waste even using a sample which has very small particle size. If we wish to ensure that 95 percent of the long-term neutralizing capacities is obtained, the experiments need to run more than 500 hours for the pH range below 9. The experimental results for 24 hours, one week (168 hours) and 500 hours are shown in Figure 3. In order to evaluate the neutralizing processes of the slag, the pH titration curves for 24- and 168-hours were used to compare with the geochemical modeling for short time and relatively long time neutralizing processes since the 168-hour curve is close to that of the 500-hour.
S~ag
Steel
9-" "%~ %
8--
..........168 hours - - - 500 hours
7
_ I
. . . .
\
~
I ....
i ....
2
\ ', ~
\
I ....
\ ',
i ....
I ....
i ....
4 6 ANC (mmol H+/8 sla~)
I ....
I ....
I
8
10
FIGURE 3. The pH levels as a function of the ANC for different reaction times for the neutralizing of the steel slag
=
4000
s
Steel S~g / { ?
3000 -~
=~
.
~
75% ANC ........... 95% ANC
2000-
~
/ ?
-
i
/
s s ? 2
~ o.
1000 "
o
?
/ ?
.J
i
7.0
. . . .
t
7.5
. . . .
t
8.0
. . . .
t
8.5
. . . .
i
9.0
. . . .
s"
I . . . .
9.5
t
10.0
pH
FIGURE 4. The reaction time as a function of the pH levels for different amounts of the ANC needed for the neutralizing of the steel slag It was also found that the consumption rates of the ANC are pH dependent in the high pH range. Figure 4 provides the consumption times for 75 % and 95 % of the long-term ANC under different pH levels. The time needed to obtain the ANC increases
636 considerably above pH 9. Thus the neutralizing reaction rates increase with decreasing pH in this range. Below pH 9, the reaction time becomes less pH dependent though the reaction rates were fast. These phenomena should be related to the mineralogy of the slag and kinetics of the neutralization reactions The detailed discussion will be given later. 3.2
Evaluation
of the neutralizing
processes
3 .2 .1 Simulations of the neutralizing processes Two cases were studied by simulation of the neutralizing processes. One case is to simulate the 24-hours batch titration experiment (Case 1), and the other is the 168-hours experiment (Case 2). The final parameters used in the simulations are listed in Table 2, the reactive fractions of main elements are given in Table 3, and the potential secondary minerals which may precipitate are shown in Table 4. The simulation results are compared with the experimental data. The comparisons of pH titration curves for both short-term (24-hours) and relatively long-term (168-hours) experiments are presented in Figure 5. It is seen that a very good agreement between the modeling and the experiment has been obtained. Besides the proton, the distributions of other main cations in the aqueous phase were also compared with the simulations for different pH values (Figure 6). A general agreement between the simulations and the experiments has been obtained particularly for the most important cation, calcium in the 168-hour experiments. From these comparisons, it has been verified that the reaction path modeling works well in the simulation of neutralizing processes for the steel slag. The differences between the short and relatively long term neutralizing processes of the slag have also been distinguished by the modeling. This is clearly found from the final parameters of the simulations (Table 2 and 3).
10.5 10.0
n
,'0~,~~' "., ,. Case2(168hours)
Experiments[ "--- Simulations
9.5 9.0
~R
8.5
-
8.0
-
7.5
=
7.0
Case
l (24
hours~"
~
.b
~ I ! ,,.
=
I
:h,,,l,,,,n
1
.... I ....
i,,,,l
2
....
n .... I ....
3
4
n .... I,,,~,,,I
5
I
.... i ....
6
I .... i,,~l
7
.... i
8
ANC (mmol H+/g slag) FIGURE 5. Simulations of the neutralizing processes of the steel slag TABLE 2. Modelin~ parameters used in the simulations Reactive concentration of main components of the slag (mmol/g)
Case 1
CaO*
MgO*
A1203*
2.745
0.056
0.004
SIO2" CaCO3 Na2SO4 K2SO4 0.021
0.01
0.003
The ratio of rate constant**
0.001 0.8000 - 0.9029
Case 2 3.572 0.336 0 . 0 1 2 0.105 0.01 0.008 0 . 0 0 4 0.6250 - 0.6925 Case 1 means the 24-hours batch titration experiment, and case 2 the 168-hours experiment. * Major modeling parameters including the ratio of rate constant. The others usualy keep constant. ** The ratio of rate constant of dissolutuion of slag to the acid addition
637
TABLE 3. Reactive fractions of main elements in the neutralizing processes
Case 1
Ca
Mg
Na
K
A1
Si
S
0.50
0.03
0.30
0.20
0.01
0.01
0.10
Case 2 0.65 0.18 0.80 0.80 0.03 0.05 0.30 Case 1 means the 24-hours batch titration experiment, and case 2 the 168-hours experiment.
1000 -
Ca Stee~ S][ag + ..............................."§.............."7"...................................~...........
r
100 ............................................................... Mg 10
A
"...,.
............................ ~
"'*"
A Na [] ............................... . ~ .................................................... g~ ............"..- A .............................................
0.1
] I I I I
--
0.01 "t'~ 7.0
,,~I
.... ~ ~ n "
Simulations expe.-Ca expe.-Mg expe.-Na expe.-K
,,,~
7.5
I,,
8.0
[ ] ....... . ...... ~..... ,o.....
~ .... % ....... ,o,,.. "O,-oo.Q
K
~x
........... -...-.~ ~
,I
....
8.5
I ....
9.0
"-,j"..~',,~ ~ " . . "-.." .... :" " . . " - . " "'-.?.
I ....
9.5
I,, l i 10.0
pH FIGURE 6. Comparison of simulation with experiments for dissolved species in aqueous phase during the titration of steel slag for 168-hour experiments
TABLE 4. Potential secondary minerals that appeared in the simulations Minerals
Formulas*
Brucite
Mg(OH)2
Calcite
CaCO3
Chrysotile
Mg3Si205(OH)4
Clinochlore- 14A
Mg5A12Si3010(OH)8
Dolomite-ord
CaMg(CO3)2
Gibbsite
AI(OH)3
Kaolinite
A12Si20 5(O H)4
Mesolite
Na0.676Ca0.657 All.99Si3.01010 ~
Montmor-Ca
Ca0.165Mg0.33 A11.67Si4O 10(OH)2
Saponite-Ca Ca0.165 M g 3A10.33Si 3.67~ 10(OH)2 * Based on EQ3/6 Thermodynamic Database (GEMBOCHS. V2-EQ6DATA0.COMoR2, 1995-09-02)
3.2.2 Characteristics of the neutralizing processes of the slag The final parameters of the geochemical simulations (Table 2) indicate which c o m p o n e n t of the slag plays an important role in the neutralizing processes. It is obvious that Ca-, M g and S i -c o n t a i n i n g mi n e r a l s are i m p o r t a n t for the neutralizing reactions. T h e c a l c i u m c o m p o n e n t p r o v i d e d the largest a m o u n t of acid neutralizing capacity in the titration. It
638 should be noted that the carbonation seems to have no large influence on the ANC because the buffer capacity is small below pH 8.5. The differences between the long- and short-time neutralizing reactions appeared mainly in the reactive fraction of the slag and the reaction rate. Compared with short-time titration processes, the reactive fraction in a long-time titration process showed much larger increase in Si and A1 than other elements. These elements are usually considered as main elements of the slag matrix. This indicates that matrix dissolution resulted in the difference between the long and short time titration processes. This feature can also be found from the ratio of reaction rate of slag to the acid addition. These ratios can be considered as the relative reaction rates of the slag in the neutralizing processes. From Table 2, it is found that the faster reaction rates corresponded to the short-time processes, and the slower rates exhibited in the long-time processes. This means that the short time titration experiments take account of the relatively fast reactions though the reactive fraction of the slag is smaller. Mineralogy and dissolution kinetics of the slag considerably affect the neutralizing processes and the ANC. The major mineral phases of the electric arc steel slag may include calcium-silicate (dicalciumsilicate, tricalciumsilicate and dicalciumferrite), calciowustite, magnesiowustite, calcium aluminate, free lime and periclase. 23, 25 The hydration of the mineral phases is important for the neutralizing processes. The main reactions of the hydration are: 25 for free-CaO and free-MgO, CaO + 1420 ---) Ca(OH)z
(1)
M g O + H20 ---) Mg(OH)2
(2)
for calcium silicates and calcium aluminate, 2Ca2SiO 4 + 4 H 2 0 ---) 3 C a O . 2SiO 2 ~ 3H20 + Ca(OH)2
(3)
2Ca3SiO 5 + 6 H 2 0 ~ 3CaO 92Si02~ 3H20 + 3Ca(OH)2
(4)
CaO 97Al203 + 121-120 ---) CaO 9Al203 961-120 + 6Al203 91-120
(5)
According to the experimental results (Figure 4), both hydration reaction equilibrium and kinetics may have influence on the neutralizing processes of the steel slag. Hydration reactions of free-CaO and free MgO should result in a high pH value in the aqueous phase. These reactions may be fast and easily reach an equilibrium. However, some of the neutralizing reactions may be controlled by kinetic factors. In the relatively high pH range, the reaction time for certain ANC is reduced with the decrease of pH and the reaction rates are slow in comparison to those at lower pH. On the other hand, the hydration reactions appeared to be less sensitive to the proton and fast in the relatively low pH range. The pHdependent reaction behavior of the slag may be attributed to the dissolution kinetics of the matrix minerals of the slag because this behavior more easily can be found from the long time experiments than from the short time experiments. It is obvious that the kinetic behavior of the slag is different from that of aluminosilicate minerals and silicate glasses. For normal silicate minerals and glasses, the dissolution rates are usually fast in both high and low pH ranges, and are slow in a neutral pH range. 26-27 Because the kinetic behaviors of the main mineral phases of the slag are not available, further assessment of the dissolution of the slag matrix will require more detailed empirical knowledge of these processes. A general agreement between the simulations and experimental measurements in the distributions of the main cations in the aqueous solution (Figure 6) indicates that the assumption of local equilibria among the aqueous species and secondary minerals is reasonable for the relatively long-term neutralizing processes of the slag. The local equilibria may also result in that the neutralizing reaction rates became very low, because
639 the dissolution of main mineral phases of the slag should be inhibited by affinity effect when saturation state of solution increases just as in a long time experiments. However, if the mechanisms of the neutralizing reactions or the reaction system are changed, the local equilibria will not be upheld, and the ANC of the slag could be different in a changed situation. For example, the ANC of the slag may be different if the neutralizing reactions are carried out in a flow-though system. We therefore need to consider the changes of reaction mechanisms and reaction system in order to have a comprehensive understanding for the neutralizing processes of the steel slag especially for a long-term consideration. 4
CONCLUSIONS
The determination of acid neutralizing capacity of the electric arc steel slag depends on reaction time. A short-time titration experiment may lead to a large underestimate of the ANC, particularly at low pH. About 500-hours experiment may be needed to account for about 95% of the ANC for the most interesting pH range. The acid neutralizing capacities of the steel slag are different for different pH levels. Most of the capacities will be consumed in a relatively high pH range (above pH 8.5). In a low pH range, the slag may only provide limited ANC for neutralizing reactions in the time scale of thousands of hours. The major contribution for the ANC comes from the Ca- (and Mg-) containing constituents of the slag. Comparison of long-time neutralizing process with the short-time one, the differences are in the reactive fraction of the elements and average reaction rates of the slag. For a long-time process, there were larger reactive fractions in Si and A1 elements and smaller average reaction rates to account for dissolution of the slag, than that for a short-time process. The pH dependence of the ANC may be explained by the hydration reactions of main mineral phases and dissolution kinetics of the slag matrix. The geochemical modeling works well in the evaluation of long-term ANC for the steel slag. It provides valuable information about the neutralizing reactions in both mineral and aqueous phases in this reaction system, which otherwise may not be easy to obtain from experiments. The contribution to the ANC of Si and A1 increases considerably with time. It is probable that the aluminosilicate minerals can give an additional significant ANC for very long times, i. e., in exess of 10's of years, this would give additional ANC in the near neutral range.
Acknowledgments Financial support for this research from the Swedish Environmental Protection Agency is gratefully acknowledged. 5
REFERENCES
1. AFR Utilization of Steel Slags--Possibilities and Environmental Impacts. Swedish Waste Research Council (AFR), November 1, Stockholm (1994). 2. Chandler A. J., Kosson D. S., Eighmy T. T., Sawell S. E., Hartl6n J., Van der Sloot H. A., Hjelmar O. and Vehlom J. An international perspective on characteristation and management of residues from municipal solid waste incineration, summary Report, International Ash Working Group (1994). 3. Morel F. M. M., and Hering J. G. Principles and Applications of Aquatic Chemistry, WileyInterscience (1993). 4. Stumm W., and Morgan J.J. Aquatic Chemistrym Chemical Equilibria and Rates in Natural Waters. Third edition, Wiley-Interscience(1996). 5. Belevi H., St~impfli D. M. and Baccini P. Chemical behaviour of municipal solid waste incinerator bottom ash in monofills. Waste Manage. Res. 10:153-167 (1992). 6. Johnson C. A., Brandenberger S., and Baccini, P. Acid Neutralizing capacity of municipal waste incinerator bottom ash. Environ.Sci Techn. 29:142-147 (1995). 7. Stegemann J. A., Schneider J., Baetz B. W. and Murphy K. L. Lysimeter washing of MSW incinerator bottom ash. Waste Manage. Res. 13:149-165 (1995).
640
8. Zevenbergen C., Van Reeuwijk L.P., Bradley J.P., Keijzer J., and Kroes R. Leaching of heavy metals from MSW incineration bottom ash in a disposal environment. Proceedings of 5th International Landfill Symposium, Vol. III: 369-377, Cagliari, Italy (1995). 9. Talbot R. W., Anderson M. A. and Andren A. W. Qualitative model of heterogeneous equilibria in a fly ash pond. Environ.Sci Techn. 12:1056-1062 (1978). 10. Roy W. R., and Griffin R. A. Illinois basin coal fly ashes. 2. Equilibria relationships and qualitative modeling of ash-water reactions. Environ.Sci Techn. 18:739-742 (1984). 11. Theis T. L., and Gardner K. H. Dynamic evaluation of municipal waste combustion ash leachate. The 5th International Conference on Ash Management and Utilization, Arlington, VA (1992). 12. Comans R. N. J., and Meima J. A. Modeling Ca-solubility in MSWI bottom ash leachates. In Environmental Aspects of Construction with Waste Materials, Goumans J. J. J. M., van der Sloot H. A.and Aalbers Th. G.(eds), pp. 103-110, Elsevier Science B.V. (1994). 13. Zevenbergen C. and Comans R. N. J. Geochemical factors controlling the mobilization of major elements during weathering of MSWI bottom ash. In Environmental Aspects of Construction with Waste Materials, Goumans J. J. J. M., van der Sloot H. A.and Aalbers Th. G.(eds), pp. 179-194, Elsevier Science B.V. (1994). 14. Eighmy T. T., Eusden Jr. J. D., Marsella K., Hogan J., Domingo D., Krzanowski J. E. and St~impfili D. Particle petrogenesis and speciation of elements in MSW incineration bottom ashes. In Environmental Aspects of Construction with Waste Materials, Goumans J. J. J. M., van der Sloot H. A.and Aalbers Th. G.(eds), pp. 111-136, Elsevier Science B.V. (1994). 15. Eighmy T. T., Eusden J. D.Jr. Krzanowski J. E., Domingo D. S., St~impfli D., Martin J. R., and Erickson P. M. Comprehensive approach toward understanding element speciation and leaching behavior in municipal solid waste incineration electrostatic precipitator ash. Environ.Sci Techn. 29: 629-646 (1995). 16. Yan J., Neretnieks I. and Moreno L. Influence of glass phase dissolution kinetics on the leaching processes of solid wastes. Proceedings of 5th International Landfill Symposium, Vol. III: 407-424, Cagliari, Italy (1995). 17. Yan J., Neretnieks I. and Moreno L. Neutralizing processes in leaching of solid wastes: Modeling of interactions between solid waste and strong acid. (submitted for publication) (1997). 18. Helgeson H. C. Evaluation of irreversible reactions in geochemical processes involving minerals and solutionsmI. Thermodynamic relations. Geochim. Cosmochim. Acta 32:853-877 (1968). 19. Helgeson H. C., Garrels R. M., and Mackenzie F. T. Evaluation of irreversible reactions in geochemical processes involving minerals and solutions--II. Applications. Geochim. Cosmochim. Acta 33:445-481 (1969). 20. Helgeson H. C., Brown T. H., Nigrini A., and Jones T. A. Calculation of mass transfer in geochemical processes involving aqueous solutions. Geochim. Cosmochim. Acta 34:569-592 (1970). 21. Wolery T. J. and Johnson J. W. EQ3/6, A Software Package for Geochemical Modeling, Version 7.2b. Lawrence Livermore National Laboratory (1995). 22. Wolery T. J. and Daveler S. A. EQ3/6, A Computer Program for Reaction Path Modeling of Aqueous Geochemical Systems: Theoretical Manual, User's Guide, and Related Documentation (Version 7.0). Lawrence Livermore National Laboratory (1992). 23.Bialucha R. and Geiseler J. Use of by-products--A German view. Swedish Waste Research Council (AFR), Utilization of Steel Slag--Possibilities and Environmental Impacts, November 1, Stockholm (1994). 24. Ye Guozhu Metallurgical treatment of steel slag. In Utilization of Steel Slag--Possibilities and Environmental Impacts, Swedish Waste Research Council (AFR), November 1, Stockholm (1994). 25. Sutio Hideaki The present state and trend of research and development for utilization of steelmaking slags in Japan. In Utilization of Steel Slag--Possibilities and Environmental Impacts, Swedish Waste Research Council (AFR), November 1, Stockholm (1994). 26. Sverdrup H. and Warfvinge P. Weathering of primary silicate minerals in the natural soil environment in relation to a chemical weathering model. Water Air Soil Pollut. 38:387-408 (1988). 27. Nagy K. L. Dissolution and precipitation kinetics of sheet silicates. Chemical Weathering Rates of Silicate Minerals. In Rev. Mineral. 31, White A. F. and Brantley S. L. (eds), pp. 173-233 (1995).
Goumans/Senderdvan der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved. REUSING WATER TREATMENT BRICK MANUFACTURING
PLANT
L. Feenstra, TNO Environment, Residual Substances Division,
SLUDGE
Eenstroom,
AS
SECONDARY
RAW
MATERIAL
IN
Energy and Process innovation, Apeldoorn, The N e t h e r l a n d s
J.G. ten Wolde, Reststoffenunie Nieuwegein, The Netherlands C.M.
641
Waterleidingbedrijven
Boral Industrie B.V.,
Tolkamer,
B.V.,
The N e t h e r l a n d s
ABSTRACT
In c o n j u n c t i o n with Boral and TNO Environment, E n e r g y and Process innovation, Reststoffenunie has m a n u f a c t u r e d on an industrial scale a trial p r o d u c t i o n run of bricks (70,000) from clay to w h i c h d r i n k i n g - w a t e r sludge was added. The bricks were a s s e s s e d in terms of p r o d u c t i o n technique and environmental impact (leaching behaviour). The results of the study were then taken as a basis for closer evaluation of the feasibility of this option. The Dutch g o v e r n m e n t is fostering the useful a p p l i c a t i o n of waste substances and therefore part financed the project (DROP s u b s i d y scheme from the Department of Public W o r k s / R o a d and H y d r a u l i c E n g i n e e r i n g D i v i s i o n and the m i n i s t r y of Housing, Spatial P l a n n i n g and the Environment). The study showed that water iron is ideally suited as a redc o l o u r i n g agent in brick production. From the process technique point of view, water iron is relatively simple to incorporate. In terms of leaching behaviour, the bricks satisfy the r e q u i r e m e n t s of the B u i l d i n g Materials Decree and this a p p l i c a t i o n is e c o n o m i c a l l y attractive both to the water companies and the b r i c k industry. The process has now become reality. R e s t s t o f f e n u n i e has signed a 5-year contract with Boral Industrie B.V. on the basis of these study results.
642 Outline
of
the p r o b l e m s / i n t r o d u c t i o n
D r i n k i n g - w a t e r s l u d g e that is r i c h in iron (also r e f e r r e d to as w a t e r iron) is p r o d u c e d at d o z e n s of d r i n k i n g - w a t e r sites in the N e t h e r l a n d s . U n t i l r e c e n t l y , the b u l k of the s l u d g e p r o d u c e d was s t o r e d in w h a t are t e r m e d f l u s h i n g ponds. A l a r g e a m o u n t of h i s t o r i c a l p r o d u c t i o n is still s t o r e d in t h e s e b a s i n s . Reststoffenunie (Residual S u b s t a n c e s Union) was set up in 1995 for the p u r p o s e of p r o c e s s i n g the r e s i d u a l s u b s t a n c e s r e l e a s e d in an e n v i r o n m e n t a l l y a c c e p t a b l e m a n n e r a n d at a r e a s o n a b l e p r i c e . A m a r k e t s u r v e y r e v e a l e d that the b r i c k i n d u s t r y w o u l d be an i n t e r e s t i n g o p t i o n for d r i n k i n g - w a t e r s l u d g e in t e r m s of q u a l i t y a n d q u a n t i t y alike. The i r o n c o m p o n e n t c o u l d be u s e d as a r e d - c o l o u r i n g a g e n t a n d the o t h e r a n o r g a n i c c o m p o n e n t s to r e p l a c e p r i m a r y raw materials. Trial
brick
production
A f t e r p o s i t i v e r e s u l t s h a d b e e n a c h i e v e d on a small s c a l e in u s i n g i r o n - r i c h d r i n k i n g w a t e r sludge, a trial b r i c k - p r o d u c t i o n r u n was o r g a n i z e d to i n v e s t i g a t e w h e t h e r the use of w a t e r iron was a c t u a l l y f e a s i b l e t e c h n i c a l l y a n d e n v i r o n m e n t a l l y , w o u l d l e a d to a s a l e a b l e p r o d u c t a n d w o u l d p r o v i d e a use for w a t e r i r o n that was s a t i s f a c t o r y to all p a r t i e s . The trial was c a r r i e d out w i t h ten tons of w a t e r i r o n t h a t c o u l d be r e g a r d e d as r e p r e s e n t a t i v e of n a t i o n a l w a t e r - i r o n p r o d u c t i o n in t e r m s of c o m p o s i t i o n . The w a t e r i r o n h a d b e e n d r i e d in a s l u d g e d e p o t to a r o u n d 50% d r y m a t t e r a n d c o n t a i n e d 35% Fe and 150 p p m a r s e n i c . A 5% i n j e c t i o n b y v o l u m e was t a k e n as the d o s a g e (3% d o s a g e b a s e d on d r y m a t t e r ) . E x p e r i e n c e h a d a l r e a d y b e e n g a i n e d w i t h this d o s a g e in the p r e l i m i n a r y i n v e s t i g a t i o n . The a d d i t i o n of this q u a n t i t y r a i s e d the i r o n - l i m e f r o m 0.85 to 1.05. This r a t i o is d e c i s i v e for the c o l o u r of the brick. M i x i n g w a t e r i r o n w i t h c l a y was p e r f o r m e d in a s i m i l a r w a y to m i x i n g v a r i o u s clays. This m e a n s that the w a t e r i r o n in the r a w - m a t e r i a l s d e p o t at the b r i c k f a c t o r y was a d d e d to the clay. The m a n u f a c t u r i n g p r o c e s s d i d not r e q u i r e a n y m o d i f i c a t i o n s . The trial s e t - u p was s e l e c t e d to p r o d u c e b r i c k s o v e r a few hours. A total of some 70,000 b r i c k s w e r e p r o d u c e d o v e r this p e r i o d . The t r i a l r e v e a l e d that u s i n g w a t e r iron h a d a p o s i t i v e e f f e c t on d r y i n g p r o p e r t i e s . T h e r e was less b r e a k a g e t h r o u g h d r y i n g t h a n was c u s t o m a r y a n d s h r i n k a g e t h r o u g h d r y i n g was a l s o s l i g h t l y less. I n s p e c t i o n of the p r o d u c t i n d i c a t e d that a b e t t e r red c o l o u r i n g h a d b e e n a c h i e v e d . One n e g a t i v e a s p e c t to e m e r g e was that c o m p r e s s i o n s t r e n g t h h a d d e c l i n e d slightly. The r e a s o n l a y in the w a t e r iron p r o d u c i n g a l e a n e r mix. It is a n t i c i p a t e d that m i n o r a d j u s t m e n t s to the p r o c e s s can m o d i f y the n e g a t i v e i m p a c t on c o m p r e s s i o n s t r e n g t h . T h a t said, the b r i c k s d i d still c o m p l y w i t h the s t a t u t o r y r e q u i r e m e n t for c o m p r e s s i o n s t r e n g t h . Effect
on composition
of r a w m a t e r i a l s
and emissions
A c o m p a r i s o n of the m a c r o - a n d m i c r o - c o m p o n e n t s of w a t e r i r o n and t y p e s of c l a y u s e d in the N e t h e r l a n d s r e v e a l e d that on a v e r a g e it was o n l y the Fe, Mn, As, Zn, Cu a n d c h l o r i d e l e v e l s in w a t e r iron that w e r e c l e a r l y h i g h e r t h a n in clay. As far as the o t h e r c o m p o n e n t s are c o n c e r n e d , the l e v e l s in w a t e r i r o n w e r e the same or
643 l o w e r t h a n the l e v e l s in the c l a y s . T h e e f f e c t of a d d i n g the w a t e r i r o n o n the c o m p o s i t i o n of the r a w m a t e r i a l w a s m i n i m a l at a d o s a g e of 5 v o l - % w i t h the e x c e p t i o n of the i r o n a n d m a n g a n e s e c o n t e n t . T h e i r o n c o n t e n t r i s e s f r o m 4.0 to 5 . 8 % (a 45% i n c r e a s e ) a n d the manganese c o n t e n t f r o m 0 . 0 8 % to 0 . 1 3 % (a 60% i n c r e a s e ) . S u c h a r i s e in the m a n g a n e s e l e v e l is n o t a p r o b l e m . It w i l l c o m e as n o s u r p r i s e to l e a r n t h e r e f o r e that s u c h d o s a g e s h a v e l i t t l e i m p a c t on the e x p e c t e d e m i s s i o n s to air. Impact
on
composition
of
raw materials
and
emissions
Component
C h a n g e in i n p u t composition
Effect
As
r i s e in concentration from 12 to a r o u n d 15 mg/kg
Zero
Fe203
r i s e in to 5 , 8 %
CaO
no c h a n g e level
MnO
rise 0.08
from
Zero
Cu
r i s e in concentration from 13 to a r o u n d 16 mg/kg
Zero
Zn
r i s e in concentration from 67 to 70 m g / k g
Zero
s l i g h t r i s e in concentration
Zero
m i n o r c h a n g e in concentration
Zero
d e c l i n e in l e v e l f r o m 6.3 to 6.2%
Zero
Other
heavy
Halogens Organic content
(F,
metals Cl,
Br)
substances
in to
level
from
4
Zero
in 3 . 2 %
level 0.13%
to
air
on e m i s s i o n
644 Leachinq trial L e a c h i n g t r i a l s r e v e a l e d that the a v a i l a b i l i t y for l e a c h i n g of the c o m p o n e n t s f r o m the w a t e r i r o n b e i n g s t u d i e d was at the d e t e c t i o n b o u n d a r y of the a n a l y s i s m e t h o d . The a v a i l a b i l i t y f r o m w a t e r iron w a s c o m p a r a b l e w i t h that f r o m c l a y a n d if a n y t h i n g was less and not m o r e a v a i l a b l e for l e a c h i n g . It was e s t a b l i s h e d for As, Cr, V, F a n d S c o m p o n e n t s that the b a k i n g p r o c e s s i n c r e a s e s a v a i l a b i l i t y . A p o i n t to n o t e is that this is n o t a c o n s e q u e n c e of the a d d i t i o n of w a t e r i r o n but of c o m p o u n d s (such as a r s e n a t e s , c h r o m a t e s a n d v a n a d a t e s ) f o r m e d t h r o u g h o x i d a t i o n d u r i n g the b a k i n g p r o c e s s , w h i c h are b e t t e r s o l u b l e a n d b e t t e r a v a i l a b l e for l e a c h i n g t h a n the c o m p o u n d s of t h e s e c o m p o n e n t s that o c c u r in clays. As far as m o s t of the o t h e r m a i n e l e m e n t s are c o n c e r n e d , the a v a i l a b i l i t y for l e a c h i n g d e c l i n e s as a r e s u l t of the s i n t e r i n g of the p r o d u c t d u r i n g the b a k i n g process. It is t r u e of all the c o m p o n e n t s i n v e s t i g a t e d that l e a c h i n g from the b r i c k b y the a d d i t i o n of w a t e r iron leads to i m m i s s i o n s that are l o w e r t h a n the m a x i m u m p e r m i s s i b l e (mpi) for a c a t e g o r y IA a p p l i c a t i o n in the B u i l d i n g M a t e r i a l s D e c r e e . Of the c o m p o n e n t s i n v e s t i g a t e d , As (0.5 t i m e s mpi), V (0.4 t i m e s mpi) a n d F (0.2 times mpi) w e r e the m o s t c r i t i c a l . T h e s e i m m i s s i o n s are c o m p a r a b l e w i t h t h o s e of b r i c k s w i t h o u t the a d d i t i o n of w a t e r iron. No i m p a c t on i m m i s s i o n to soil was t h e r e f o r e i d e n t i f i e d f r o m u s i n g w a t e r iron. T h e l e a c h i n g to soil f r o m a b r o k e n b r i c k was a l s o l o w e r t h a n the mpi for c a t e g o r y 1 a n d 2 a p p l i c a t i o n s . T h i s m e a n s that the b r o k e n b r i c k c a n be u s e d at a s e c o n d p h a s e of its life as a n o n - m o u l d e d c o n s t r u c t i o n m a t e r i a l in a c a t e g o r y 1 a p p l i c a t i o n . Immissions calculated w i t h o u t w a t e r iron.
with
Elements
standard
use
of b u i l d i n g
Immission (mg/m 2)
bricks
mpi
with
and
(mg/m 2)
13 b u i l d i n g bricks without water iron
red Rhine/Waal clay brick
building brick with water iron
As
4-213
50
75
435
F
664-2258
1627
i,I00
14,000
V
39-697
266
280
2400
Mo
15-169
ii
150
Evaluation/use in practice G i v e n the p o s i t i v e e x p e r i e n c e s of the trial p r o d u c t i o n , this use of s l u d g e has m e a n w h i l e b e e n i m p l e m e n t e d in p r a c t i c e . One m a j o r b o o s t h e r e is that the M i n i s t r y of Housing, S p a t i a l P l a n n i n g a n d the E n v i r o n m e n t has i s s u e d w h a t is t e r m e d a d e c l a r a t i o n of nona p p l i c a b i l i t y . W a t e r i r o n c a n be r e g a r d e d as a full r a w m a t e r i a l in its o w n right. The n e g a t i v e label of w a s t e s u b s t a n c e is no l o n g e r
645 applicable. T h e r e is no p r o b l e m in t e r m s of sales v o l u m e , as the t o t a l q u a n t i t y of w a t e r i r o n that can be u s e d in b r i c k p r o d u c t i o n e x c e e d s s e v e r a l f o l d the t o t a l a n n u a l p r o d u c t i o n of w a t e r i r o n of a r o u n d 2 0 , 0 0 0 t o n (dry m a t t e r ) . The w a t e r c o m p a n i e s a n d the b r i c k i n d u s t r y w i l l h a v e to take a n u m b e r of m e a s u r e s in u s i n g w a t e r iron. For e x a m p l e , the b r i c k i n d u s t r y w i l l h a v e to c r e a t e a s u p p l y of w a t e r i r o n in a depot, b e c a u s e the s u p p l y of w a t e r i r o n does not m a t c h d e m a n d as a r e s u l t of s e a s o n a l f a c t o r s a n d o t h e r i n f l u e n c e s . No a d d i t i o n a l m e a s u r e s n e e d be t a k e n for this storage. S t u d i e s b y the K I W A r e v e a l e d that the l e a c h i n g of w a t e r i r o n s a t i s f i e s c a t e g o r y 1 r e q u i r e m e n t s u n d e r the B u i l d i n g \ M a t e r i a l s D e c r e e . To e n s u r e c o n s i s t e n c y of i n p u t q u a l i t y , the v a r i o u s c o n s i g n m e n t s of w a t e r i r o n in the d e p o t s h o u l d be m i x e d i n t o a h o m o g e n o u s w h o l e p r i o r to use. As the w a t e r i r o n o r i g i n a t e s f r o m v a r i o u s w a t e r s u p p l y sites, the w a t e r c o m p a n i e s are c o n s i d e r i n g s e t t i n g up a c e r t i f i c a t i o n system. The c e r t i f i c a t i o n p r o c e s s w i l l l a y d o w n r e q u i r e m e n t s for the m a c r o c o m p o s i t i o n ( m o i s t u r e c o n t e n t , s a n d p e r c e n t a g e , i r o n a n d m a n g a n e s e content, o r g a n i c s u b s t a n c e s , etc.) a n d the m i c r o c o m p o s i t i o n ( i n c l u d i n g the a r s e n i c c o n t e n t ) . T h e s l u d g e t r e a t m e n t ( t h i c k e n i n g a n d d e w a t e r i n g ) w i l l a l s o h a v e to be optimised. A n e c o n o m i c a n a l y s i s has s h o w n that the c o s t s of the b r i c k o p t i o n ( i n c l u d i n g d e w a t e r i n g ) are l o w e r b y a f a c t o r of 2 t h a n the t r e a t m e n t c o s t s w i t h o t h e r o p t i o n s s u c h as use as a f l o c c u l a t i o n a g e n t for d e w a t e r i n g s e w a g e s l u d g e or use as a d e p h o s p h a t i n g agent. U s a g e as a r a w m a t e r i a l for the b r i c k i n d u s t r y is a t t r a c t i v e to the w a t e r c o m p a n i e s (saving on d u m p i n g costs) a n d the b r i c k i n d u s t r y (saving on the cost of r a w m a t e r i a l s ) alike. T h i s is t h e r e f o r e a win-win situation. M e a n w h i l e all the w a t e r c o m p a n i e s h a v e c o n f i r m e d that this as the m o s t a t t r a c t i v e o p t i o n for the m e d i u m a n d
they regard l o n g term.
This Page Intentionally Left Blank
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
ASSESSMENT
OF CHEMICAL
SENSITIVITY
647
OF WAELZ
SLAG
Hae-Ryong BAE 0), Radu BARNA (1)(2), Jacques MIEHU (2), Hans van der SLOOT (3), Pierre M O S Z K O W I C Z 0), Christian D E S N O Y E R S (4) (1) LAEPSI, INSA Lyon, 20 av A Einstein, 69621 Villeurbanne, France (2) Polden, INSAVALOR, BP 2132, 69603 ViUeurbanne, France (3) ECN, Petten, The Netherlands (4) Metaleurop Recherche, Trappes, France 1. A b s t r a c t In the recycling industry, the recuperation of zinc from Electric Arc Furnace dust by the Waelz process generates important quantities of slag. This slag presents good mechanical properties, and for the most siliceous slag, a high stability which would enable its use by total or partial substitution of certain granulates in civil engineering. Our study (within the framework of a european programme, cofunded by the European Commission - DGXII) concerns the physico-chemical and mineralogical characterization and leaching behaviour of several types of Waelz slag. The leaching tests used are regulatory tests and specific characterization tests of leaching behaviour. They take into account the influence of several main parameters of the valorization scenarios envisaged for the slag (e.g. pH, Redox potential, chemical nature of the leachant, type of contact - liquid/solid etc). 2. I n t r o d u c t i o n 2.1. Objective In the European Union, the steel industry produces about 700000 tons/year of Electric Arc Furnace (EAF) dusts. The EAF dusts can be recycled by the Waelz process which converts the dusts into an impure zinc oxide, called Waelz oxides, which is reprocessed in metallurgical plants (i.e. Imperial Smelting Process). A slag is also produced. The recycling of all the EU EAF dusts could lead to the production of about 500000 tons/year of Waelz slag. This slag is basically an iron reduced slag. In order to reuse it, i.e. as aggregates or as filling material in civil engineering (specific road and construction applications), in total or partial substitution of natural materials, important research programmes concerning its characterization and long term behaviour will have to be carried out. Our study is a contribution to the chemical and mineralogical characterization of different Waelz slags. We have also observed the influence of different parameters (i.e. pH, EH, leachant composition,...) on the release of different chemical species during different leaching tests. There is a wide choice of parameters due to the variety of disposal and reuse scenarios Of the slags. The aim of our paper is to gain a better understanding of the slag properties in relation to different reuse scenarios.
648 2.2. General presentation of the programme partners The research programme was carried out with the participation of two main industrial companies operating the Waelz process within the European Union: BUS Metall GmbH and METALEUROP. Two independant research laboratories, POLDEN-Insavalor and ECN carried out the programme in the laboratory and on a pilot scale, coordinated by Metaleurop Research (MER). 2.3. Description of the Waelz process technological parameters
and identification of the most relevant
The Waelz process allows to treat EAF dust (polluting waste from steel industry) in order to recycle zinc and to obtain a steel residue. After mixing of EAF dusts with coke breeze (reducing agent) and other additives (e.g. smceous sand), this feed is continuously introduced into a rotary kiln. The temperature in the furnace, about 1100~ allows reduction and vaporization of zinc and other volatil metals (lead, cadmium) to recover ofter oxidizationcondensation an impure zinc oxide (Waelz oxides). The molten residue, containing low zinc and lead contents, is water cooled at the oudet of the kiln to form granular Waelz slag. According to the mixture and composition of the material on entering the furnace, the operating conditions of the furnace (e.g. temperature or residence time, etc) and the characteristics of the slag cooling, two different types of slag are obtained: silica rich slags and lime rich slags. 2.4. Case history of local valorization of slags For many years, low amounts Waelz slags have been used locally as basic material for road construction, sportsgrounds and dykes. However, for use in compliance with the ever stricter environmental regulations, it has become necessary to characterize the slag more precisely. Particular attention must be paid to long term leaching behaviour of slags in order to estimate their long term behaviour under the conditions of different valorization scenarios. 2.5. General presentation of the experimental study The experimental study can be divided into three parts. The ftrst part is to determine the constituents and the structure of the slags (chemistry and mineralogy). The second part is to apply the French and Dutch regulatory tests (results not presented here). Finally, the third part is to characterize the leaching behaviour of slags using tests specially designed and adapted to the objectives of the study presented below. 3. E x p e r i m e n t a l
study
3.1. Presentation of the studied reference slag samples In the following table we present the elementary composition (% mass), basicity index 031) and the codes used for the five Waelz slags under study. The slags come from five different plants. The technological operating conditions used for sampling of the slags in each installation was characterized. Homogenization, carried out at the end of the process ensured the representative character of the slag sampling. The granulometric fraction studied was between 0-20ram.
649
Code B.I. Zn Pb Cd As Cu Fe metal FeO SiO2 CaO K20 Na20
2-FG 2-DU 1-FQ 1-OK 0.39 0.48 0.4 2.6 2.38 0.24 3.77 3.2 1.07 0.36 1.55 1.6 0.0007 0.0011 0.0002 0.0015 0.0141 0.0141 0.089 0.04 0.49 0.32 0.4 0.29 17.2 24.9 23.2 23.3 6.62 3.47 6.75 21.5 31.4 37.3 26.9 6 7.58 15.35 7.7 13.1 0.29 0.11 0.22 0.23 0.41 0.39 0.51 0.42 0.97 0.8 2.03 0.75 Table 1 :Chemical composition of slags (% mass)
3-FG 3.5 0.35 4.22 0.0015 0.0107 0.32 6.21 39.7 7.77 23.5 0.23 0.92 1.12
The slags 2-DU, 1-OK and 1-FQ are slightly alkaline (BI about 0.4 to 0.5) whereas the slags 2 and 3-FG are more alkaline (BI greater than 2.5). The pollutant content, especially Zn and Pb, varies within quite large limits (more than ten fold) according to the slag characteristics. 3.2. Characterization of leaching b e h a v i o u r 3.2.1. C h e m i c a l sensitivity test: p H 5 a n d p H 12.5 (controlled with N a O H Ca(OH)2 s a t u r a t e d solution)
or with
The leaching tests at controlled pH aim to estimate pollutant release (at equilibrium and dynamically) from slags which undergo prolonged aggression as far as pH is concerned. The choice of pH 5 corresponds to acidic aggression which is quite classical (due, for example, to acid rain or oxidized sulfur compounds), pH 12.5 corresponds to alkaline aggression (for example, water having been in contact with a material containing hydraulic binders). In the case of acid aggression, the effect of pH is evaluated as such without using acids whose anion would influence the solubility (such as CO2 or organic acids). As for the alkaline medium, we have tried to evaluate the effect of pH alone controlled by a simple species (NaOH) and also the effect of a calcium containing medium close to the hydraulic binder scenario. The choice of this type of test involves the use of a pH regulation loop, consisting of a pH electrode, an electronic regulator and a pump. The main experimental parameters of the controlled pH test are presented in table 2. Particular care was taken to avoid carbonation of the leachates. 3arameter initial mass ratio L/S 3article diameter stirring renewal time pH
conditions 10 <4ram continuous after 1 day, 2, 3, 4 and 7 days 5, controlled using H N O 3 1N 12.5, controlled using N a O H saturated solution of Ca(OH)2 0.45 btm filtration Table 2 9Experimental parameters of chemical sensitivity test
1N
or
650 In the following table we present the cumulated release (mg/kg of slag) after 7 days of leaching for certain elements (Pb, Zn, As and Fe) according to the pH. 1-OK 1-FQ pH=5 128.5 20.9 pH= 12.5,NaOH 4.1 ND pH=12.5,Ca(OH)2 2416.4 634.6 pH=5 201.7 6.9 pH=12.5,NaOH 5.2 ND pH=12.5,Ca(OH)2 141.8 8.1 pH=5 ND ND pH=12.5,NaOH 22.1 4.1 pH=12.S,Ca(OH)2 8.1 6.3 pH=5 258.9 3235.7 pH=12.5,NaOH 10.8 ND pH=12.5,Ca(OH)2 ND ND Table 3 : Cumulated release (7 days) for
Pb
Zn
As
Fe
2-DU 14.9 91.7 1372.5 420.1 33.6 24.7 ND 0.5 ND 255.9 6.0 3.2 different types
2-FG 19.6 111.3 3353.8 2336 25.9 125.2 ND 3.7 ND 7376.2 ND ND of leaching
3-FG 97.6 2294.3 2316.2 650 18.1 14.7 ND 0.3 ND 8401.2 3.0 3.3
The following graphs show the cumulated release of lead, zinc, arsenic and iron after 7 days of leaching for each type of slag and for each test (controlled pH 5, pH 12.5 by NaOH or in saturated lime solution). In a general manner, it can be noted that there is a greater cumulated release for the most a l k a l i n e s l a g s , 2 and 3 FG, which are therefore less stable.
According to the s p e c i f i c i t y (figure 1): -
of metals,
two types of behaviour can be demonstrated
metals whose release is greater under acidic conditions (controlled pH 5) : Zn and Fe for example ; metals whose release is greater in alkaline conditions : Pb and As for example.
The influence of the chemical nature of the alkaline leachant (lime or NaOH) is particularly important in the case of release of lead (strongly leached by a lime solution), of zinc and of arsenic. The behaviour of arsenic shows that it is not solubilized at pH 5 and very slightly solubilized at alkaline pH (saturated lime solution). This phenomenon could be explained by the integration of certain solubmtydata (Nishimura) : arsenic is only slightly soluble at acid pH in the presence of certain heavy metals (Zn, Cu, Ni and Co) and also only slightly soluble at alkaline pH in the presence of alkaline earth metals (Ca, Mg, Sr and Ba). As regards lead and zinc, a distinct destabilization of these elements in a saturated lime solution can be noted.
651
Pb
~
As
10000
25
1000
E 20
10o
~ 10 0
1
2DU
1OK
1FQ
[o..,,
2FG
9-~-
3FG
]
Ca(O =
Fe
g
"O
N
NI~ND
N
1OK
1 FQ
2FG
ND._ ND
3FG
[EI pl'15 [] NaOH [] Ca(OH)2 I
Zn
10000
10000 u
.ND
2DU
_E
1000
lOO
== ~
E
--
2DU
=
,
=
11~
,
1FQ
=
,
=
2FG
1000
lO
I:tl
1
I
2DIJ
3FG
1 OK
1 FQ
/ 2FG
3FG
II2] pH5 9NaOH [] Ca(OH)2 ]
F1pl-15 [] NaOH [] Ca(OH)2 ]
Figure 1 9Cumulated release of metals according to pH during 7 days leaching The release dynamics are represented as the evolution of the cumulated extracted mass against the cumulated contact time. Three types of behaviour of the leached species can be distinguished" -
slow dissolution without apparent saturation of the solution (see figure 2). The concentration is therefore proportional to contact time. This is the case for zinc (slags 2DU and 2-FG at pH 12.SNaoH, 2-DU at pH 5), of copper (2-DU at pH 5 or pH 12.5Ca(OI~2), of lead (slags 1-FQ at pH 12.5c~(o~2).
35
E 30 25
g
20
15 ~ 10 E
5~ o~----
o
o A
~f 2
4
6
2DU
2FG
8
days
Figure 2 9Dissolution behaviour of zinc for slags 2-DU and 2-FG at pH 12.5NaOH after successive leachings
652 - rapid initial dissolution with a slow progression of the phenomenon, asymptotic (figure 3). his may be due to a mass dissolution of certain grams but also to a surface dissolution of the grains, followed by low mass transport towards the solution. This type of behaviour can be observed for lead in the case of slag 1-FQ at pH 12.5Ca(OH)2 and for arsenic in the case of slag 1-OK at pH 12.5Ca(OH)2. 3500 E
3000 2500 "~ ~ 2000 1500 1000 500
1OK / 1FQ / 2FG I
o
-
t
0
,., . . . .
0
3FGJ
,-,
#
2
i.........................
4
6
days
Figure 3" Dissolution behaviour of lead in slags at pH 12.5Ca(OH)2 during the successive leachmgs - dissolution with apparent control by saturation of the solution. This is the case for lead at pH 12.5Ca(OH)2 for slags 2-FG, 1-OK (figure 3) and of zinc for slags 1-OK and 2FG at pH 12.5Ca(OH)2 (figure 4).
E
150 T
=. lOO} E ~ ~
+
I
50 Ol
0
_
+ 2
~ 4
1OK 2FG
6
days Figure 4: Zinc dissolution be haviour for slags -1-OKand 2-FG at pH 12.5Ca(OH)2 during successive leachings 3.2.2. Redox properties of Wealz slag Using a recently developed test procedure (Draft NVN 7348, 1995) to assess the reducing properties of a material, the Wealz slags have been tested. The ftrst step in the procedure is to identify, whether the material being tested has reducing properties ( conditions : L/S=2, particle size < lmm, degassed demmeralized water, 24 hours contact time; measurement of redox potential against redox potential for normal oxidized water at same end pH). If this proves to be the case the release of reducing substances and the reducing capacity of the solid phase is assessed. The reducing substances in the leachate and the reducing substances in the solid phase are oxidized by an excess of CelV, the surplus is back-tit_rated with Fe II (Hoede et al, 1993). From the tests, it follows that the change in redox potential is less than 50 mV at the normal pH of slag leachate (pH 7-12).
653
This indicates that the slags do not affect their surrounding by imposing reducing conditions (less than 0.04 Mol O2/kg). Still the slag do possess reducing properties as Fe(0) and Fe(2+) and Mn(2+) are present. This is illustrated by the oxygen consumption measured after Ce titration 1.8-2.0 Mol 0 2 / k g . ,uu i - - ~ / ~[-'-'~-.}~
Domainof reducing properties
600
I
500
" - --- -_
,7
BF slag
o
Ph0sph. Slag
D
Steel Slag
V
demiwater >
400
-
O
@
D @ % ~ ; ~ Q ' ~ ~
o
DU2
o
FG3
|
FQ1
m 300
O
ca
200
-
! 00
-
Ca~ ( D O
| V
% "~
OK1 0 6
I
I
I
I
I
I
7
8
9
10
11
12
13
o
FG2
pH
Figure 5 9Redox potential in function of the pH The measured redox potential in leachates has been placed in perspective to other slags by plotting the redox potential measured as a function of pH in comparison with oxidized, demineralized water (figure 5). The Wealz slag is reducing, but it is only appaerent at pH < 6. This is related to the solubility of reduced Fe and Mn below pH 6. The release of reducing substances above pH 6 is however limited. This is probably caused by the rather low leachable S content. From other work S -species have been identified as the most important carhers for imposing reducing conditions on the slag surroundings. The presence of metallic Fe can lead to stresses in the slag matrix, when water can reach these iron phases through cracks. 3.2.3. A N C test The objective of the study was to follow the concentration modifications at equilibrium of the main elements in the slags according to the different levels of acid attack of the leachant. The study also allows characterization of the slags according to their acid neutralizing capacity. The protocol of the test is as follows (table 4) : 3arameter mass ratio L/S ~article diameter stirring temperature time to reach equilibrium pH control filtration
conditions 10 <4ram continuous 23 _~2~ 48 hours
solutions of HNO3 ( from 10 -4 M to 2 M) 0.45 bun Table 4" Experimental parameters of ANC test
The range of acid concentration allowed us to obtain, at the end of the contact time, a pH between (denoted final pH) 2 and 12. Two types of graph present the test results ; the evolution of the final pH of the leachates according to the acid added (fig 6) and the species solubility according to final pH of the leachates (fig 7).
654
The dependance of the final pH of the slag solutions on the quantity of nitric acid added (following figure) shows that the slags have different neutralizing capacities, in logical agreement with the progression of the basicity index" 2-DU < 1-OK < 1-FQ < 2-FG < 3-FG. 14
12 10
8 6 4 2
0 0,01
0,1
1
10
100
mmol acidlg slag
......
-Figure6 i-Ev0iuti0n o f p H accori~ingtoacid added. .......
The decrease in pH for acid leaching of the slags 2-DU, 1-OK and 1-FQ is relatively linear with increased acidity of solutions until about 5 mmol. A more important decrease then occurs for 2-DU and 1-OK, which could indicate a chemical transformation of a phase. The two most alkaline slags of the series (2-FG and 3-FG) show an important decrease in their buffering capacity especially in the interval 0.4 to 2 mmol HNO3/g of slag. The following figures present the concentrations (mg/1) of certain elements (Pb, Zn, As et Na) in the extraction solutions according to their final pH.
,ooo: x i, x
lOOOO r m
E
100 10,
1{
lOOO
Zn
x
E ~Z~O 5
pH
1
x
o,1
O
10
X
0,01
15
t
0
As
lO
E
0
pH
, 5
pH
-~ 10
~O2DU [] 1OK_A 1FQ X 2FG * 3r-G]
10
~
i
1 I 15
j Na
loo
x
15
x'F"
E
xn
0,01
lO
1ooo
zx z~Dx
o,1
5
I
[o 2DU [] 1OK A 1FQ X 2R3 , 3FG
1
Pb
,o
z~
0,1: 0,011 0
x
lOO
0
zx<>
no
o
-4-
~
5
10
pH
& ........
-o 2DU [] 1OK Z~ 1FQ x 2FG ~ 3FG t
Figure 7 9Concentrations of metals at equilibrium against pH
15
655
Solubilization of the species according to the nitric acid added shows characteristic behaviour : -
amphoteric behaviour for solubilization of zinc and lead. Their release is low for pH 7 to 11, which are in fact the pH values obtained by application of the test X31-210. - greater dependance on the slag specificity for the solubilization of arsenic and certain (Ca, Cu and Cd). A greater dependance on slag nature can be noted in an acid medium. low dependance solubility/pH for alkaline elements (Na and K). The agreement between ANC test results and controlled pH test results can be noted, especially at pH 5 and pH 12.5 controlled by NaOH. 3.2.4. p H
stat
This pH static test (Van der Sloot et al, 1993 .... ) provides information on the pH sensitivity of leaching behaviour of the slag. This information is important as it forms the basis to identify the solubility controlling phases. Furthermore the test gives information on the sensitivity to pH in specific field scenarios, where the pH of the slag may change due to external influences. The acid neutralization capacity (ANC) derived from the test is a useful additional property in this respect as it allows a quantification of the buffering of the system when exposed to externally imposed pH changes. Finally, data from many different leaching tests can be placed in perspective using the pH test data as a reference. To assess the behaviour of slags as a function of pH, a pH stat test is carried out on granular material (d< 2mm) using a liquid to solid ratio of L/S = 10 and pH control in the range from pH = 4 - 12. A pH controller is steered by computer dosing HNO3 or N a O H depending on whether acid or base is required. The dosing equipment is developed in such a way as to minimize the volume increase. Results of pH stat tests on Wealz slag are presented in figures 8.
656
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In figures 8 the data are given. Results are presented for a silica rich slag (2 DU) and for a lime rich slag (3 FG). To illustrate the agreement between pH stat information and other test methods, the results of the single pH controlled tests and the ANC test data are included for comparison. As can be seen from the graphs there is no relation between chemical composition and leachability, as leachability is largely controlled by pH in solution. The agreement between different tests is reasonable to good with some specific exceptions, where the addition of for instance Ca or CO2 affects directly the solubility of a specific element (e.g. Cd, As). 3.2.5. L e a c h i n g modelling: identification of the solubility controlling phase Using the geochemical speciation model MinteqA2, the leaching data obtained from the pHstat experiments on Wealz slags have been modelled to identify the potential solubility controlling phases. This information is relevant for decisions on possibilities for modification to improve slag leaching characteristics. It also helps to identify the influence of basicity and the degree of amorphous glass versus crystalme phases. The leaching data obtained for major, minor an trace elements have been transformed to tool/1 concentrations and used to calculate saturation indices for all possible mineral phases. Based on the evaluation of the saturation indices, those phases that are likely controlling phases due to a good match between measured concentrations corrected for activity (Debye-Huckel) and solubility of specific phases over all or a part of the pH range studied have been selected. Below the behaviour of a few individual elements is discussed: Cadmium - The leachability of Cd is apparently not controlled by otavite (CdCO3), as the actual solubility is 2 orders of magnitude less than predicted for otavite. Cd-silicate does not appear to be a relevant controlling phase either. In this case, solubility may be correlated with matrix dissolution. The Cd/Si ratio of leached concentrations amounts to 0.0008 (st dev. 0.0006), when a few extreme values are omitted. This is remarkably constant for a trace element, so this mechanism may apply here.
Calcium - The solubility of Ca is reasonably well explained for the lime rich slags, where gypsum is a controlling phase. There are many possible calciumsilicate phases that may be controlling release in the silicate rich slags.
- In the low pH region Fe II minerals are relevant solubility controlling phases. The mineral siderite proves to be a likely controlling phase. In the neutral to alkaline pH domain Fe leachability can be controlled by ferrihydrite- Fe(OH)3. This will be particularly relevant upon aging of the slag, when due to environmental influences (oxidation and carbonation) a coating of ferrichydroxide is formed on the slag surface. Iron
The most relevant solubility controlling phase for Pb solubility from Wealz slag is Pb(OH)2. PbSiO4 is not of importance. In the longer term the important controlling phase appears to be cerrusite - PbCO3. Only in case of very fresh slags, which have not undergone sufficient aging - conversion of Pb to PbO will occur, then upon contact with water this will change to Pb(OH)2 and finally upon contact with the atmosphere to PbCO3 . Depending of the state of aging control of Pb solubility by one of the intermediate phase may be most prominent. The leachability of Pb as function of pH is very similar for all slags. The silica rich slag show generally a higher leachability at high pH than the lime rich slags. Lead-
658 - The solubility of Zn is different for the different slags. In some cases the ZnSiO3 solubility matches quite well with the measured data points (3 FG, 2 DU for part of the pH range). In case of 2 FG and the high pH range for 1 OK, willemite (Zn2SiO4) appears to be the major solubility controlling phase. The leachability of the lime rich slags is significantly higher than that of the silica rich slags. This is particularly relevant in the low pH domain. Zinc
The geochemical modelling has shown that for the majority of the elements possible solubility controlling phases can be identified. The lime rich and silica rich slags can be distinguished in terms of their leaching properties for several elements. Based on the modelling it can be concluded that slag surface mineralogy controlling leachability is different from bulk slag mineralogy as obtained by XRD and similar techniques. This is particularly important when the slag can be considered as a stable material with very limited susceptibility to dissolution. The changes in the slag surface due to exposure to environmental conditions will largely control what the environmental impact from slags in a given scenario will be. From this work it can be concluded that application in an acidic environment (e.g. peat soils, acid sandy soils) should be avoided. 4. C o n c l u s i o n s The experimental programme aims to characterize the intrinsic physico-chemical properties of slags studied as well as their leaching behaviour. It aims to demonstrate the influence of certain main parameters (such as pH). Several important points can be made : 9 The release of polluting elements during leaching tests using the X31-210 standard for example is below the regulatory thresholds for slags 2-DU, 1-OK, 1-FQ and 2-FG. This is not the case for 3-FG. 9 The data obtained with different leaching tests compare generally well with the data obtained with the pH stat test and ANC test, which implies that these tests can be regarded as a basic characterisation test of leaching behaviour. The pH stat data can be used as a basis of reference for quality control testing. It also provides crucial information on potential risks of exposure to other pH conditions in the field, which can not be obtained from a single step leaching test. 9 The geochemical modelling has been placed in relation to slag characteristics such as lime-rich and silica-rich properties. This has revealed interesting general characteristics of Wealz slags and leads to a better identification of possible solubility controUing phases. The changes in the slag surface due to exposure to environmental conditions will largely control what the environmental impact from slags in a given scenario will be. The leaching behaviour of the different Wealz slags is largely solubility controlled in the pH domain ranging from pH 5 to 12. This can be concluded from the similarity in leaching as function of pH in spite of differences in slag basicity and level of crystalization. _ Complex interactions can be observed between the chemical nature of the leachants and the physico-chemical specificity of each slag. The valorization scenarios by partial or total substitution of granulates by slags in cement with high limestone content may pose the problem of the high availability of lead, observed experimentally.
659
_ It can be noted that more alkaline slags such as 2-FG and 3-FG are more sensitive to the influence of pH than slags 1-FQ, 1-OK and 2-DU. The acid-base capacity and the release of elements are dependant on the characteristics of each slag. The low buffer capacity of the silica rich-slags implies that application or disposal of untreated slag in acid environments (pH 4-6) should avoided. To optimize integration of slags in materials, it will be necessary to orient research towards less alkaline binders than classical CPA cements, while offering sufficient protection against acid waters. A concrete with low limestone content (slag, alummates), where the alkalinity in the porewater would have a less corrosive effect on the vitreous phases of slags, would seem more appropriate. _
5. R e f e r e n c e s M E H U J.. M O S Z K O W I C Z P.. BARNA R.. PHILIPPE P. et MAYEUX V. French qualification procedure for solidification processes. :Environmental Aspects of Contruction with Waste Materials. Proceedings of international conference. WASCON'94. edited by Goumans.J.J.M..Van Der Sloot.H.A..Albers. Th.G.. Elsevier.Amsterdam.1994. pp. 281 - 292. BARNA R. Etude de la diffusion des polluants dans les d6chets solidifi6s par liants hydrauliques. ThlEse de doctorat. INSA de Lyon. 1994. 210p. S A N C H E Z F. Etude de la lixiviation de milieux poreux: application au cas des d6chets so fidifilEs par liants. ThlEse de doctorat. INSA de Lyon. 1996. 245p. NISHIMURA T, I T O H C.T. T O Z A W A K. Stabilities and solubilities of metal arsenites and arsenates in water and effect of sulfate and carbonate ions on their solubility. Metallurgical Soc Inc, ISBN 0- 87339-037-7. 1987 H O H B E R G . I. et RANKERS. R. Leaching properties of cement-bound materials. In:Environmental Aspects of Construction with Waste Materials. Proceedings of international conference. WASCON'94. edited by Goumans.J.J.M..Van Der Sloot.H.A..Albers. Th.G.. Elsevier.Amsterdam.1994. pp. 387 - 396. VAN D E R SLOOT H. A.. COMANS R.N.J. et HJELMAR O. Similarities in the leaching behaviour of trace contaminants from waste, stabilized waste, construction materials and soils. ECN-R_X-94-070 et Waste Quality Institute (VKI). Denmark. 1994.vol 40. D U T R E V., VAN D E CASTEELE C. Solidification/stabilisation of hazardous arsenic containing waste from a copper refining process. Journal of Hazardous Materials. 1995. pp. 55 -68. J O H N S O N C. A.. B R A N D E N B E R G E R S. et BACCIN P. Acid neutralizing capacity of municipal waste incinerator bottom ash. Environ. Sci. Technol.. 1995. pp. 142 - 147. N E N 7341. Leaching characteristics of building and solid waste materials - Leaching tests Determination of the availability of inorganic components for leaching. Draft 1993 (previously part of N V N 2508); Netherlands Normalisation Institute; the Netherlands N E N 7343. Leaching characteristics of building and solid waste materials - Leaching tests Determination of the leaching of inorganic components from granular materials with the column test. Draft 1993 (previously part of N V N 2508); Netherlands Normalisation Institute; the Netherlands Draft N V N 7347. Leaching characteristics of building and solid waste materials - Leaching tests - Determination of the leaching behaviour of inorganic components from compacted granular building materials and waste. Draft 1994 , Netherlands Normalisation Institute; the Netherlands N E N 7349. Leaching characteristics of building and solid waste materials - Leaching tests Determination of the leaching of inorganic components from granular materials with the cascade test. Draft 1993 (previously part of N V N 2508); Netherlands Normalisation Institute; the Netherlands
660 Draft NVN 7348 Determination of reducing properties of materials. H.A. van der Sloot en D. Hoede. Laboratorium onderzoek naar de invloed van reducerende eigenschappen op het emissie-gedrag van industrieslakken in oeverbeschermmgen. ECN-C-94-094. 1994. Felmy, A.R., D.C. Girvin, and E.A. Jenne, MINTEQ--A computer program for calculating aqueous geochemical equilibria, EPA-600/3-84-032, U.S. Environmental Protection Agency, Athens, (1984). Regulation for Construction Materials. Staatsblad van het Konmkrijk der Nederlanden, 1995, 567. H.A. van der Sloot. Developments in evaluating environmental impact from utilization of bulk inert wastes using laboratory leaching tests and field verification.1996.Waste Management, 16 (1-3), 65-81.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
661
Immobilisation of heavy metals in contaminated soils by thermal treatment at intermediate temperatures C. Zevenbergen ~, A. Honders 1, A.J. Orbons ~, W. Viaene 3, R. Swennen 3, R.N.J. Comans 4, and H.J. van Hassel: 1 IWACO B.V., P.O. Box 8520, 3009 AM Rotterdam, The Netherlands 2 Centre for Soil Treatment, Europalaan 250, 3526 KS Utrecht, The Netherlands 3 University of Leuven, Geology, Celestijnenlaan 200 C, B-3001 Heverlee, Belgium 4 Netherlands Energy Research Foundation, P.O. Box 1, 1755 ZG Petten, The Netherlands 5 Scarabee, Couwenhoven 62-31, 3703 HN Zeist, The Netherlands
Abstract Thermal treatment at intermediate temperatures is widely considered as an effective technology to remove organic contaminants from soil by volatilization and/or destruction. Average operating temperatures of commercial soil treatment systems generally vary between 500 and 650~ Thermal treatment also alters the physical and chemical properties of the soil, and thus affects the leachability of co-contaminants such as heavy metals. In the present study the effects of thermal treatment on heavy metal leaching from soil have been examined. The results presented in this paper and other work suggest that thermal technologies in the intermediate temperature range offer an opportunity to destroy simultaneously organic contaminants and to immobilize heavy metals in soil in one unit operation. However, further work is necessary to determine the capabilities and limitations of these technologies for soils contaminated with both heavy metals and organic pollutants. The results of this study seem to justify emphasis on the role of iron and clay minerals in thermally treated soil and on the conditions of thermal treatment which affect the behaviour of these constituents in soils. 1. INTRODUCTION Thermal treatment in a rotary kiln at intermediate temperatures is widely considered as an effective technology to remove organic contaminants from soils by volatilization and/or destruction. In the Netherlands about 1,000,000 tonnes/year of contaminated soil is thermally treated. Thermal treatment also alters the physical and chemical properties of the soil and thus may affect the leachability of co-contaminants like heavy metals. At present, there is a lack of knowledge on heavy metal leaching from thermally treated soil and its relation with soil characteristics and operating conditions. Average operating temperatures of thermal soil treatment plants vary between 500 and 650~ At these temperatures heavy metals remain in the soil except for mercury. This heavy metal is removed from the soil during thermal treatment by volatilization (Van Hasselt, pers. comm.). Modern thermal soil treatment plants are equipped with flue gas treatment to meet stringent air pollution standards (Van Hasselt, 1996). In the kiln the soil is usually directly heated by a gas burner. The atmosphere in the kiln is generally reducing (Van Hasselt, pers. comm.). Once the pre-selected temperature has been reached, it is held only a few minutes and then drops to a temperature just below 100~ In this cooling step the hot treated soil is quenched either with water or indirect with water and air. The treated soil is kept in storage in a stockpile during several days to several weeks. During storage the temperature may increase slightly due to exothermic weathering reactions.
662
Recent research has indicated that heavy metals exhibit a lower leachability in soils after thermal treatment at intermediate temperatures than the original untreated materials. Eddings and co-workers (1994) have revealed that a significant fraction of cadmium, chromium, and lead reacts with the glassy aluminosilicates during thermal treatment of montmorillonite clay at a temperature of 650 to 980~ to form compounds that are extractable only by HF digestion techniques. At lower temperatures (150 to 500 ~ Wei (1995) has demonstrated that soils spiked with cadmium and lead after treatment in bench-scale thermal experiments generate lower TCLP leachate concentrations than the untreated soil. They also observed that an increase in thermal treatment temperature enhanced immobilisation of these metals. In the present experimental study the effects of thermal treatment at intermediate temperatures on heavy metal leaching from soils have been examined using petrographical and chemical techniques, electron microscopy, and leaching tests. Essential features of the thermal alteration products are emphasized, along with their effect on the soil leaching properties. For comparison, leaching data derived from soils before and after treatment in Dutch thermal soil treatment plants are reported. These data have been accumulated by the authors in the last three years as part of a larger study on the leaching behaviour of contaminated soils, which is reported elsewhere (Heynen et al., 1997). 2. MATERIALS AND METHODS A series of laboratory experiments were performed in which three different soil types were heated with different residence times and temperatures, ranging from respectively 10 to 30 min. and from 550 to 750~ (abbreviated hereafter as min/~ Differential thermal analysis (DTA), X-ray diffraction (XRD), selective extractions, and electron microscopy were used to obtain information on thermal alteration products and heavy metal distribution. Other important properties as allophane content and cation exchange capacity (CEC) of the untreated and treated materials were included as well. Heavy metal leaching was determined using a column test (Dutch standard NEN 7343). The experimental results were compared with data derived from full-scale treatments. Materials Three (air-dried) soil samples were used in the experimental study: (i) a residue from wet soil treatment (sample I), (ii) a clayey soil sample (sample II), and (iii) a fine loamy soil sample (sample III). The residue from wet soil treatment used in this study comprises the < 63 #m fraction of a contaminated soil separated by means of a wet mechanical separation method in a soil washing plant. The soil samples were selected to cover a wide range of physical and chemical properties. Physical and chemical properties of the soil samples are listed in Table 1. Thermal experiments The thermal experiments were carried out in an electrically heated ceramic laboratory furnace. The temperature of the furnace was adjusted using a thermocouple which was placed in the centre of the furnace. When the test temperature was reached, crucibles filled with about 2 kg of air-dried material were introduced into the centre of the hot furnace. No attempt was made to determine the amount of metal volatilization during the treatment. Loss of heavy metals by volatilization for all thermal experiments was expected to be less than 5 % (Wei, 1995). The heated material was subsequently quenched with air. After cooling the treated samples were rewetted to approximately their original moisture content. Finally, the samples were cured in a climate room at 28~ during three weeks to simulate the weathering conditions during storage in a stockpile. The samples lost about 6 to 13 % of their original total weight due to the thermal treatment.
663
DTA, XRD, CEC, and allophane content Differential thermal analysis (DTA) was used to study under controlled conditions the thermal changes and reactions in the samples which occur in the temperature range between 80 and 750~ The untreated samples were heated at a rate of 10~ together with thermally inert alumina. The temperature differentials between the sample and the alumina indicate physical and chemical processes occurring in the soil. The reactions are typically graphed from an arbitrary base line as a function of the temperature. Exothermic reactions show upward curves and endothermic reactions downward curves from the baseline. Mineralogical properties of the clay minerals in the < 2/zm fraction of the untreated and thermally treated samples were studied by (well-oriented) X-ray diffraction (XRD). Cation exchange capacity and allophane content of the untreated and thermally treated samples were determined according to Mizota and Van Reeuwijk (1989). Table 1. Physical and chemical properties of the untreated samples (in % (w/w)).
SiO2 TiO2 A1203 Fe203 MnO MgO CaO Na20 K20 P205 BaO LOI sum As* Cd* Cr* Cu* Pb* Ni* Zn* fraction < 63 #m organic matter CaCO3 dry matter
sample I 53.75 0.7 11.05 5.75 0.11 1.76 6.87 0.7 1.95 0.29 0.05 16.9 99.88 23 2.2 51 120 870 41 750 29 12 11 52.3
sample II 56.13 0.84 13.3 6.03 0.1 2.26 5.58 0.49 2.4 0.22 0.03 12.00 99.39 12 <0.1 35 13 24 25 93 55 10 <0.1 67.5
sample III 82.88 0.31 4.17 2.14 0.02 0.35 1.49 0.38 1.18 0.07 0.04 5.6 98.63 5.4 3.5 24 1,700 2,000 26 2,300 3.3 3.9 1.9 84.3
LOI loss on ignition * in mg/kg Electron microscopy Advanced electron microscopy equipped with a solid state X-ray detector provides detailed information on alteration features and the distribution of heavy metals in soil and soil-like materials (Bates et al., 1992; Zevenbergen et al., 1996). The untreated and treated sample III (10 min/650~ were examined using an analytical (transmission) electron microscope (AEM) (JEOL 2010). Sample III was selected for its relatively high heavy metal content (see Table 1). Size-fractionated samples of dried and disaggregated soil were embedded in epoxy and thinsectioned using an ultramicrotome. The thin sections ( > 100 nm thick) were examined using
664
brightfield/darkfield imaging, lattice fringe, and surface area electron diffraction (SAED). Compositional trends were determined using energy dispersive x-ray spectroscopy (EDS). Selective dissolution with oxalate and dithionite Both oxalate and dithionate dissolution procedures for the extraction of A1 and Fe have widely found application in soil classification and in studying soil genesis (Mizota and Van Reeuwijk, 1989). Acid oxalate extraction was used to determine the heavy metal fraction that is associated with the 'active AI' and 'active Fe' components. This includes allophane (amorphous hydrous aluminosilicate), A1- and Fe-humus complexes, and amorphous or poorly ordered oxides such as ferrihydrite (Mizota and Van Reeuwijk,1989). The extraction with dithionite was used to determine the heavy metal fraction that is associated with the so-called 'free iron oxide'. This fraction consists of ferrihydrite and crystalline Fe, such as goethite, hematite, and magnetite particles up to about 50 #m in diameter. The extraction procedures were carried out on sample III. Leaching test Column tests were carried out according to NEN 7343. This test is considered to simulate leaching in the short- and medium-term by relating the emission (expressed in mg/kg) to the liquid to solid (L/S in 1/kg) ratio and is adopted in Dutch legislation.
3. RESULTS AND DISCUSSION DTA The DTA analysis of the three samples shows a first change at temperatures up to 200~ which reflects the loss of pore water, adsorbed water and part of the interlayer and lattice water of the clay minerals. These reactions are endothermic. A large exothermic peak between 200 and 400~ arises in all samples from the oxidation of organic material. As the temperature rises, the chemically combined water within the clay minerals begins to be lost (dehydroxylation). The clay minerals remain intact until a temperature of 450 to 600~ is reached, and then break down into an amorphous mass. All the samples show a continuous loss of water due to dehydration and dehydroxylation of the clay minerals upon heating. As a consequence a distinct endothermic peak cannot be discerned in this temperature range. The small endothermic peak around 570~ marks the inversion of alpha-quartz to beta-quartz. Sample III exhibits an endothermic peak at 740~ presumably due to the dissociation of calcium carbonate in CaO and CO2 (see Fig. 1). 322"C
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665
Electron microscopy Comparison of the untreated and treated sample III shows that a significant fraction of the clay minerals has melted (vitrified) during the thermal treatment (see Fig. 2). The melted particles have a porous appearance and Fe appears depleted in these particles. Some of the particles have conserved their original structure, which indicates that they have not been subjected to the pre-set temperature (see Fig. 3). In both samples, the heavy metal content in the clay structures is very low. In the untreated sample the heavy metals are found as major elements in discrete, submicrometer metal, alloys, and primary oxide grains. Occasionally they are found as major constituents in coatings on silicate grains like quartz and clays. The untreated sample also contains abundant grains rich in Fe and O. The major elements in all these grains are O en Fe and their crystal structures range from poorly crystalline to wellordered polycrystalline grains. The poorly crystalline grains consist of ferrihydrite while the well-ordered grains are hematite. Admixtures of these two minerals are common and both contain heavy metals (see Fig. 4). Unlike the untreated sample, discrete metal alloys and primary oxide grains are rare in the treated sample. Virtually all the heavy metals found in the treated sample are associated to the Fe and O rich grains. However, the grains are no longer ferrihydrite or hematite but rather magnetite and hematite. Heavy metals are sometimes present as major constituents of Fe-rich grains with the magnetite structure (see Fig. 5). The magnetite is presumably formed during the thermal treatment both from ferrihydrite and from Fe released by the clay minerals, and appears to have incorporated most of the heavy metals into its spinel structure. T29
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Figure 3. Partially melted clay particle clay in the thermally treated sample III (10 min/650~ (1 cm = 300 nm).
666
Figure 4. Fe-rich grain embedded in a clay matrix of the untreated sample III. EDS spectrum from the Fe-rich grain indicating that this phase contains Zn (1 cm = 300 nm). Selected area diffraction pattern (SAED) from the Fe-rich grain showing classic ferrihydrite diffraction characteristics.
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667
Mineralogy, cation exchange capacity and allophane content Results of the X-ray diffraction analysis, cation exchange capacity and allophane content determination are summarized in Table 2. The X-ray diffraction results suggest that during the thermal treatment the bulk of the clay minerals are decomposed resulting in the formation of more poorly ordered structures. In general the allophane content increases as the temperature and/or residence time increases, which indicates that the decomposition of the clay minerals is not yet fully completed under the prevailing experimental conditions as mentioned above. This suggestion is consistent with the earlier noted TEM observations. The CEC decrease is mainly a result of the destruction of organic matter. Table 2. Relative abundance of the clay minerals ( < 2 /zm fraction), CEC, and allophane content of the untreated and treated samples. sample
treatment
sample I
untreated 10 min/550~ 30 min/5500C 10 min/650~ untreated 30 min/650~ 10 min/750~ 30 min/750~ untreated 10 min/650~ 30 min/650~ 30 min/750~
sample II
sample III
relative abundance of the clay minerals ++ + +/++ o
++
CEC (meq/lOOg) 29 17 13 15 48 16 18 12
allophane content (% (w/w)) 1.3 1.2 2.0 1.5 1.8 1.2 1.8 3.8 0.6 1.2 1.5 3.0
+ + = abundant (> 10%); + = moderate (5-10 %)" +/- = traces (5%); - = non detectable (< 5%)
Selective dissolution with oxalate and dithionite The results of the selective dissolution experiments are given in Figure 6. The results indicate that the dithionite extractable heavy metals fraction and to a lesser extent the oxalate extractable fraction are affected by the thermal treatment. The dithionite extractable heavy metal fraction is in general significantly higher in the thermally treated samples than in the untreated sample. Moreover, the dithionite extractable heavy metal fraction generally increases when the samples are treated at longer residence times. With the exception of Cu, a relatively small amount of the heavy metals, compared to the amount released in the dithionite extraction, is extracted in the oxalate extraction. The increase of the dithionite extractable heavy metal fraction after thermal treatment most likely results from the transformation of amorphous iron hydroxides to more crystalline forms. This conclusion is consistent with the earlier noted TEMobservations that during the thermal treatment a significant fraction of these metals is scavenged by magnetite structures. In addition, the results reveal that Cu and to a lesser extent Cd, Pb, and Zn are also retained by thermal decomposition products of the clay minerals (allophane-like materials). Along similar lines, compositional mapping using electron microscopy recently indicated that these heavy metals were incorporated into newly formed hydrous aluminosilicate rims on glassy grains of weathered municipal solid waste incinerator ash (Zevenbergen et al., 1996 and references therein). In conclusion, the formation of magnetite and allophane-like materials may significantly reduce leaching of those metals from thermally treated soil.
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i
30 min; 650c
Figure 6. Amount of dithionite and oxalate extractable heavy metals (in mg/kg) from the untreated and the thermally treated sample III (10 min/650~ Leach testing Figure 7 summarizes the column test results derived form the untreated and treated samples. The results are presented as cumulative emission (mg/kg) at a L/S ratio of 10. As the temperature and/or residence time increase, the emission of the heavy metals generally decreases. This effect is most obvious for Cd, Cu, Pb, Ni, and Zn. As exhibits no systematic trend. In sample I a longer treatment time seems to enhance the leachability of this oxyanion, while the reverse is observed in samples II and III. The leachability of Cr appears to be slightly affected by the thermal treatment with the exception of the sample which has been treated at the highest temperature and longest residence time (sample II, 30 min/750~ In this sample the leachability of Cr is significantly higher than in the untreated sample. For comparison, CEC, pH, and leaching data of As, Cr, Cu, Ni, Pb, and Zn derived from different soils before and after treatment in thermal soil cleaning plants (closed symbols) are plotted together with the experimental results (open symbols) in Figure 8. These results reveal that the pH generally increases from values around 8 to values around 10, while the CEC decreases after thermal treatment. Both alterations are probably a result of the removal of organic matter by thermal destruction. The leaching data derived from the samples which have been treated in the thermal treatment plants are consistent with the above reported experimental leaching data. Recent studies have revealed that the leachability of heavy metals like Zn, Cd, Pb, and Cu from thermal residues (e.g. coal fly ash and MSWI ashes) and contaminated soils typically attains a minimum value in the pH range between 7 and 10 (Van der Sloot, 1996; Meima and Comans, 1997). Since the pH of both untreated and thermally treated soils roughly coincides with this pH-range, it is not likely to assume that the observed decrease in leachability of these contaminants after thermal treatment is exclusively due to a pH-effect.
669
sample I
~ A
[] untreated
0.1
E
[] 10 min; 550 C B 10 min; 650 C
"E 0"01
II 30 min; 550 C
0.001 As
i
Cr
Cu
Pb
Ni
Zn
sample II
,.,,,
E
Cd
[] untreated
0.1
930 min; 650 C B 10 min; 750 C
0.01
930 min; 750 C 0.001 As
Cd
Cr
Cu
Pb
Ni
Zn
sample III E
[] untreated
0.1
[] 10 min; 650 C 0.01
B 30 min; 650 C
0.001
I
As
Cd
Cr
Cu
Pb
Ni
Zn
Figure 7. Leachability (cumulative emission at L/S = 10 in mg/kg) of heavy metals from the untreated and thermally treated samples.
670
i
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I~ 9
. ..-- 9
E i ~9 1.000 .
.
.
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....
,....
0 o ......
.---..o
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...-"
...."
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"
i .
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l
~
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"
.
. . t, 0.010 0,100 1 .000 10.000 emission before thermal treatment (mglkg)
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9
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.
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O /
.-
e i
.. - " ....
...........
..... 9
.
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..
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.
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9
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.-
"
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,~
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i
0.001 0.001 0.010 0.100 1 .000 10.000 emission before thermal treatment (mglkg)
S
I
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"
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t ~
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| 0.001
0.001 0.01 0 O. 100 1 .000 emission before thermal treatment (mglkg)
0.0(!1
I,
t, '
0.001 0.010 O. 1O0 1 .000 emission before thermal treatment (mglkg)
Figure 8. CEC (meq/100g), pH and leachability of As, Cd, Cr, Cu, Ni, and Zn (cumulative emission at L/S = 10 in mg/kg) before and after thermal treatment (open symbols = thermal experiments" closed symbols = soil thermal treatment plants).
671 60
r
o
E
5o
"E
40
9~r
30
- 9 .,~
{
20
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o
14
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o
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20
30
40
50
60
before thermal treatment (meqllOOg)
4
6
8
10
12
pH before thermal treatment
Figure 8. Continued 4.
C O N C L U S I O N S
Recent research has revealed that heavy metals exhibit a lower leachability in soils after thermal treatment at intermediate (550 to 750~ temperatures than the original materials. Our experimental data are consistent with these observations. However, we also observed that thermal treatment had little or even an inverse effect on the leachability of As and Cr. The experimental data reveal that thermal treatment results in a drastic change of the mineralogical an chemical nature of the soil. The thermal alterations which may contribute to the observed decrease in leachability of these heavy metals are: an increase of the pH; formation of crystalline magnetite structures from ferrihydrite and from Fe expelled by the clay minerals during the treatment. These magnetite structures appear to have scavenged heavy metals during the treatment; formation of reactive, amorphous aluminosilicates from clays. It should be noted that the pH of thermally treated soil will gradually decrease to approximately its original pH due to weathering and accumulation of organic matter. At present, it is unknown to what extent these processes will affect the leachability of these heavy metals. Further study is needed to assess the effects of weathering on leaching behaviour of thermally treated soils on the longer term. The results presented in this paper and other work suggest that thermal technologies in the intermediate temperature range (500 - 650~ offer an opportunity to simultaneously destroy organic contaminants and to immobilize heavy metals in soils in one unit operation. However, further work is necessary to determine the capabilities and limitations of these technologies for soils contaminated with both heavy metals and organic pollutants. The results of this study seem to justify emphasis on the role of iron and clay minerals in thermally treated soil and on the conditions of thermal treatment which affect the behaviour of these constituents in soils. R E F E R E N C E S
Bates, J.K., Bradley, J.P., Teetsov, A., Bradley, C.R., and Buchholtz ten Brink, M., 1992. Colloid formation during waste form reaction: Implications for nuclear disposal, Science, Vol. 256, p. 649-651. Eddings, E.G., Lighty, J.S., and Kozinski, J.A., 1994. Determination of metal behaviour during incineration of a contaminated montmorillonite clay. Envirn. Sci, Technol. 28, 17911800.
14
672 Heynen, J.J.M., Comans, R.N.J., Honders, A., Frapporti, G., Keijzer, J., and Zevenbergen C. (these proceedings). Development of enhanced testing procedures for the determination of leachability of heavy metal contaminated soils. Meima, J.A. and Comans, R.N.J., 1997. Geochemical modelling of weathering reactions in municipal solid waste incinerator bottom ash. Environmental Science and Technology (in press). Mizota, C. and van Reeuwijk, L.P., 1989. Clay mineralogy and chemistry of soils formed in volcanic material in divers climatic regions. Soil Monograph 2, ISRIC, Wageningen, The Netherlands, 186 p. Van Hasselt, H.J., 1996. Emission control system to meet Ducth stack emission standards on soil thermal cleaning plants. International Incineration Conference, Seatle, Washington, p. 177-181. Van der Sloot, H.A., Comans, R.N.J., and Hjelmar, O., 1996. Similarities in leaching behaviour of trace contaminants from wastes, stabilized wastes, construction materials, and soils. Sci.Tot.Env. 78, p. 111-126. Wei, Y., 1995. Leaching study of thermally treated cadmium-doped soils. Hazardous Waste & Hazardous Materials. Volume 12, No 3, p. 233-242. Zevenbergen, C., Van Reeuwijk, L.P., Bradley, J.P., Bloemen, P., and Comans, R.N.J., 1996. Mechanism and conditions of clay formation during natural weathering of MSWI bottom ash. Clays and Clay Minerals, Vol. 44, No 4, p. 546-552
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997Elsevier Science B.V. All rights reserved.
673
INVESTIGATION STRATEGIES FOR CONTAMINATED SOILS IN FINLAND Hanna-Liisa J~irvinen Geological Survey of Finland Espoo, Finland
ABSTRACT Geological Survey of Finland has completed a project, the objective of which was to find reliable investigation strategies and methods for contaminated soils. The applicability of the ISO-standard Draft, the Dutch prestandard and U.S. EPA recommendations to Finnish geological and hydrogeological environment was studied. During the preliminary assessment, standard drafts turned out to be complicated and expensive to carry out in practice. However, the basic principles in drafts were found applicable also in Finland. During the project a new Finnish practice was created. Investigations on contaminated sites are divided into three phases: preliminary survey, field investigation and additional investigations. Based on the preliminary survey the site is whether "probably uncontaminated" or "potentially contaminated". The distribution of the contaminants is always supposed to be heterogeneous in Finnish geological environment. Quality control samples reveal the quality of sampling and analyzes. The local baselines (background values) must always be verified in any assessment of contaminated sites. The sampling pattern is determined after the preliminary survey and it is designed primarily caseby-case. The sampling intensity is recommended for both probably uncontaminated or potentially contaminated sites.
1 INTRODUCTION This research project was carried out by Geological Survey of Finland and Technical Research Centre of Finland. It was partially financed by The National Environmental Geotehnics Program (organized by The Technology Development Centre of Finland). The objective of this project was to recommend reliable investigation strategies and methods for contaminated sites as well as provide quality specifications for various remediation methods. In the final project report the recommendations are given for a number of soil and water samples, their type and locations, quality control, analytical methods etc. The report includes a short review of geophysical methods, which are an increasing field of investigation. The recommendations are based on the experiences gathered from investigation projects made by Geological Survey of Finland and Finnish consultant companies. So far, widely differing approaches and research intensities have been used in the completed remediation projects. Investigations have generally been insufficient. Usually remediation decisions are based limited risk assessment if any, which may lead to unnecessarily intensive measures. The aim of this project was to unify research procedures to improve risk assessment and remediation solutions. The applicability of the ISO-standard Draft (ISO 10381-5 Version 6, draft) Ref.l., the Dutch prestandard (NVN 5740) Ref.2. and U.S. EPA recommendations Ref.3. to Finnish geological and hydrogeological environment was studied. However, they turned out to be complicated and expensive. Thus, the project published a new Finnish practice Ref. 4.
674 2 RESEARCH PHASES Investigations in contaminated sites are divided into three phases like in ISO-standard Draft Ref.l.: preliminary survey, field investigation and additional investigations. During the preliminary survey, information on the past and present activities on the site as well as basic information on the soil stratification and hydrogeology is gathered. The hypothesis of the situation at the site is formulated like in ISO-standard Draft Ref. 1. and the Dutch prestandard Ref 2. The site is whether "probably uncontaminated" or "potentially contaminated". However, the distribution of the contaminants is always supposed to be heterogeneous in Finnish geological environment. Soil and groundwater sampling is carried out during the field investigation phase. The chemical analyzes of samples will reveal if the site is contaminated. Additional investigations are applied if needed. The objectives of additional investigations are the following: to provide detailed information on the geological condition of the site and its impact on surrounding area, to give detailed information on contamination (3D) for risk assessment as well as for remediation design, cost estimate and performance. The intensity of additional investigations depends on geological and hydrogeological conditions, the reliability of earlier investigations, the future activity on the site and considered remediation technologies. After investigations the risk assessment is carried out to evaluate the hazards on the site and in the surrounding areas. Health risk and ecological risk assessment provide information to decision makers as the consequences of possible actions.
3 SUBSTANCES FOR CHEMICAL ANALYZES When potentially contaminated sites are investigated, soil and groundwater samples should be analyzed for the contaminants that are probable according to the preliminary survey. In addition, a small number of samples should be analyzed for a wider spectrum of substances. The most common contaminants should be analyzed when probably uncontaminated sites are in question. Ref. 1. Typically in Finland soil has been contaminated by the following inorganics in industrial or other activities: arsenic, chromium, copper and mercury in wood-processing industry; copper, nickel, zinc and cyanides at mines, metal smelters and shooting ranges; zinc and chromium in surface treatment; mercury and chromium in leather and fur industry. The typical organic contaminants are creosote oil, chlorinated phenols, dioxines and furans at wood preserving facilities; PCBs in chemical wood-processing industry; oils, solvents, dioxines and furans in chemical, textile, metal and machine building industry as well as at waste management facilities; oils, solvents, and gasoline at gas stations. Ref.5.
4 QUALITY CONTROL SAMPLES AND BASELINE SAMPLES Quality control (QC) samples are taken near to the point from which the original soil samples were taken. QC samples are divided into two parts which both are analyzed. The differences between an original sample and a QC sample reveal the quality of sampling. The differences between QC subsamples reveal the quality of analyzes. The recommended number of QC samples is 10% of the number of original samples. At least two QC samples should always be taken.
675 To get more reliable information on contamination in groundwater, samples can be taken from several groundwater pipes/wells and several times. Geochemical baseline samples give the natural concentrations of elements at the region. Natural concentrations of several elements exceed the guide values designated for contaminated soils in many areas in Finland. The local baselines (background values) must always be verified in any assessment of contaminated sites. Ref. 6. The baseline samples have to be taken near the site of concern, from the same kind of geological environment that is not contaminated.
5 SAMPLING PATrERN Samples should be taken from both probably uncontaminated sites and potentially contaminated sites. Sampling pattern is determined after the preliminary survey. It is designed primarily case-by-case. The locations of sampling points and sampling intensity are based upon the knowledge of site conditions, such as geological variabilit, the area of the site, contaminant concentrations and migration directions. Also systematic or random sampling can be used if there is a specific reason for the applicability. Groundwater monitoring pipes/wells should be located upgradient and downgradient of the site as well as at the contaminated site.
6 SAMPLING INTENSITY The number of samples to be taken depends on the area of the site, topography and geological conditions. Every other sampling point is extended deeper -- in other words at every other point both topsample and subsample are taken. Topsamples are taken from the topsoil layer which existed when contamination occurred. Topsamples are taken from the surface to one meter deep (0 - 1 m). Subsamples are taken at the level where human activities have not extended (natural soil layers). If there is a soil layer with very low permeability (clay, clay rich in organic material or silty till) it is recommended to take subsamples from the surface of this layer or from the soil layer above bedrock. It needs to be very careful not to penetrate soil layers with low permeability. In addition, the possibility of perched groundwater needs to be considered. Drilling holes can be filled up with bentonite slurry to avoid migration. It is economical to take several samples at the same time, although all of them would not be analyzed. During the field investigation primarily individual samples are analyzed. Mixed samples may be useful in same cases, but their applicability has to be considered carefully. In practice, it is possible to mix at the most five individual samples. Groundwater samples are always analyzed individually. Table 1 shows the recommended sampling practice when the site is probably uncontaminated. When the site is probably uncontaminated, groundwater samples are taken from the nearest groundwater wells. Surfacewater samples are taken toward the flowing direction from a ditch, lake, pond or river.
676 Table 1. Recommended sampling practice for probably uncontaminated sites.
~
S
o
i
l
samples S a m p l i n g grid in h o r i z o n t a l p l a n e e.g.
(ha)
Number sampling points
Number of a n a l y z e d topsamples
Number of a n a l y z e d subsamples
3 4-9 6-13 7 - 16 8-18
3 4-9 6-13 7-16 8-18
2 2-4 3-6 3-8 4-9
9 10 11 12 13
9 - 20 1 0 - 22 11 - 24 12 - 25 13 - 27
4-10 5-11 5-12 6-12 6-13
Number of QC samples
mxm <1
2 3 4
50 60 x 60 65 x 65 70 x 70
5 6 7 8 9 > 9
75 x 75 75 x 75 80 x 80 80 x 80 90 x 90 100 x 100
1
50
x
20 22 - 24 - 25 - 27 -
If the hypothesis of the site is potentially contaminated the investigations have to ensure the contamination of different subsites and strata. In addition, it is necessary to find the distance from where on no contaminants are detected or their concentration is lower than the trigger concentration. In the table 2 is shown the recommended sampling practice for potentially contaminated sites.
Table 2. Area
(ha)
<1 2 3 4 5 6 7 8 9 >9
Recommended sampling practice for potentially contaminated sites. Soil s a m p l e s
Sampling g r i d in horizontal p l a n e e.g. mxm 25 30 35 35 40 40 40 45 45 50
x x x x x x x x x x
25 30 35 35 40 40 40 45 45 50
Groundwater samples Number of sampling points
Number of a n a l y z e d topsamples
Number of analyzed
Number of QC samples
Number of monitoring pipes/wells
subsamples
16 25 28 32 36 39
-
25 36 43 50 56 61
42 - 66 45 - 71 48 - 75
16 25 28 32 36 39 42 45 48
-
25 36 43 50 56 61 66 71 75
8-12 1214 16 18 19 21 22 24 -
18 21 25 28 30 33 35 37
3-4 3-4 3-4 4-6 4-6 4-6 4-6 4-6 4-6 >6
677 Near "hot-spots" smaller grids are recommended, e.g. 10 m x 10 m. "Hot-spots" are usually located at the places where the polluting agents are used: near the cylinder and sink at wood impregnation plant, around oil tanks etc. When the site is very large, it is economical to use large grids at first. In addition, geophysical methods can be useful before locating sampling points. The geological conditions have to be considered carefully when deciding number and location of groundwater wells.
7 CROSS-CONTAMINATION DURING SAMPLING Cross-contamination is a severe problem during sampling process. Sampling should be started from the cleaner part of the site and then proceeded to more contaminated parts. To avoid contamination the sampling equipment can be cleaned mechanically or chemically. Usually mechanical cleaning (drying, brushing, blowing, rinsing) is adequate. When contaminants are not water soluble, special cleaning agents or solvents can be used. Organic contaminants can be cleaned with such chemicals as methanol, acetone, hexane or isopropanol. Chemical cleaning is recommended before a new sampling project is started. Usually mechanical cleaning is sufficient between the sampling points. Ref. 7.
8 DISCUSSION Even though it is difficult, almost impossible to give general instructions for investigations some guidelines have to be established. In the beginning there is a strong tendency to underestimate the need of detailed data. Later, during remediation the deficiencies turn out to be expensive. The applicability of these recommendations is reevaluated in future when more experience is gathered. Currently the Geological Survey of Finland is working on research project including intensive sampling on contaminated wood impregnation plants.
9 REFERENCES
1. Draft International Standard, ISO 10381-5 Version 6, 18 July 1994. Soil quality Sampling -Part 5: Guidance on Procedure for the Investigation of Urban and Industrial Sites with Regard to Soil Contamination. ISO/TC 190/2/PG 5 Working Draft. 2. NVN 5740. 1991. Bodem. Onderzoeksstrategie bij verkennend onderzoek. Delft: Nederlandis Normalisatie-Instituut. 3. Mason, B.J. Preparation of Soil Sampling Protocols: Sampling Techniques and Strategies. PB92-220532, U.S. Environmental Protection Agency, Las Vegas, Nevada (1992). 4. Mroueh, U-M., Jarvinen H-L. and Lehto O. Saastuneiden maiden tutkiminen ja kunnostus. Teknologiakatsaus 47/96, The Technology Development Centre of Finland, Helsinki (1996). 5. Puolanne, J., Pyy O. and Jeltsch U. (eds.). Saastuneet maa-alueet ja niiden k/isittely Suomessa. Muistio 5/1994, Ymp~iristOminister6, Helsinki (1994).
678 6. Salminen, R. Raskasmetallien luonnolliset taustapitoisuudet eri maalajeissa. Environmental geology applications, The Finnish Environment 71, Finnish Environmental Institute, Helsinki (1996). 7. Naturv~.rdsv~.rket. V~.gledning for miljOtekniska markunders6kningar. Del II: F~.ltarbete. Raport 4311. Naturv~.rdsv~irket, Lindk6ping (1994).
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
679
Development of fast testing procedures for determining the leachability of soils contaminated by heavy metals J.J.M. Heynen a, C. Zevenbergen a
R.N.J.
Comans b,
A.
Honders c,
G.
Frapporti a,
J.
Keijzer a,
IWACO B.V., Consultants for Water and Environment, P.O. Box 8520, 3009 AM Rotterdam, The Netherlands ECN, Netherlands Centre for Energy Research, P.O. Box 1, 1755 ZG Petten, The Netherlands SCG, Centre for Soil Treatment, Europalaan 250, 3526 KS Utrecht, The Netherlands
Abstract In the Netherlands, the use of mildly contaminated soils, both treated and untreated, in civil and public works is regulated by the Dutch Act for Building Materials ("Bouwstoffenbesluit") [1]. By law, the leachability of inorganic contaminants and heavy metals, as determined by the column test (NEN 7343), is not permitted to exceed certain values. Drawbacks of the prescribed test procedure are the relatively high costs and the long time needed for testing. The latter adds to the logistics of soil handling and utilization. Therefore, faster and less costly testing procedures for determining the leachability of soils are required. Over the last 3 years, an extensive study has been performed on the leachability of Cd, Cu, Zn, Pb, Hg, Ni, As and cyanide in treated and untreated soils and some dredged sediments. In this study and earlier studies 230 samples of untreated and treated soils and dredged sediments were analysed for specific soil parameters, contaminant contents and leachability. The analytical data were accumulated, statistically interpreted and evaluated. Some of these soil samples were selected for geochemical speciation modelling [2], some other soil samples were selected for studies on the effects of thermal treatment on leaching behaviour [3]; the results are reported in separate papers in these conference proceedings. The study resulted in: (1) a deeper level of understanding of the chemical and physical processes governing the leachability of contaminants from (re-usable) soils; (2) the description of a less time-consuming procedure for assessing the leachability of contaminants from (re-usable) soils. The latter result can reduce the need for temporary storage and consequently can lower the handling and logistic costs considerably.
680 1. INTRODUCTION In the Netherlands, the use of secondary materials like mildly contaminated soils, both treated and untreated, is regulated by the Dutch Act for Building Materials ("Bouwstoffenbesluit")[1]. About 1.5 to 2 million tonnes of mildly contaminated soil per year are re-used in the Netherlands and an additional 1.5 to 2 million tonnes of contaminated soils are treated by various processes (mainly wet processes, e.g. classification and flotation, and thermal processes). After treatment, the soil should be suitable for re-use. If a soil is to be re-used, the leachability of inorganic contaminants and heavy metals from that soil, as determined by the column test (NEN 7343), is not permitted to exceed certain values, prescribed by the Dutch Act for Building Materials [1]. Drawbacks of the prescribed test procedures are the relatively high costs and long time required for testing (about 5 weeks). The latter slows down soil handling and delays utilization, resulting in the need for costly temporary storage. Therefore, attemps are being made to develop faster and less costly testing procedures for determining the leachability of soils. Also more insight is required into methods for optimizing soil treatment with respect to leachability. The objectives of the study described in this paper are: 1. the development of fast testing procedures for determining the leachability of inorganic contaminants from treated and untreated contaminated soils. These testing procedures need to be evaluated against the testing procedures prescribed in the Dutch regulations; 2. the assessment of the effects of soil treatment processes on the leachability of the soil. Clearly, the second objective overlaps with topics of related papers in these conference proceedings on speciation modelling [2] and leaching behaviour of thermally treated soils [3]. Therefore, the second objective will not be discussed extensively in this paper. A large number of samples of treated and untreated contaminated soils and dredged sediments have been investigated using laboratory analyses (determination of leaching characteristics with the standard column leaching test and fast leaching tests and leachingrelated parameters). The laboratory data gathered have been combined with leachability data from earlier studies [4,5,6] and have been statistically analysed.
2. TESTED SOILS AND DREDGED SEDIMENTS 2.1. Selection of soils to be studied For this study, samples of suitable treated and untreated contaminated soils and dredged materials were provided by the Dutch Centre for Soil Treatment (SCG), the Dutch Development Programme for Treatment Processes for Contaminated Sediments (POSW) and several soil treatment contractors and soil distribution centres. About 250 soils to be treated via the Centre for Soil treatment were screened on certain intake parameters. According to a prescreening in 1993, the most frequently appearing inorganic contaminants were: Cu, Zn, Pb, Cd, Hg and cyanide. Initially the study was focused on sets of treated and untreated soils and dredgings which were contaminated with the above-mentioned contaminants (in dredgings Ni was analysed instead of cyanide). The soils studied were selected on the basis of their contaminant levels. Later on, in order to
681 fill in the leachability bandwidth to be studied, an effort was made to obtain soils with elevated leaching levels (around the legal limits for leachability). From then on, nontreatable soils as well as soils re-usable without any treatment were added to the study; soils were selected on the basis of the CEN (TC 292) two-step batch leaching test and arsenic was added to the set of parameters. An overview of the origin and number of samples included in this study is given in table 1. Table 1 Origin and number of samples included in the full research programme (one combination of untreated and treated soil samples) Technology Number Soil untreated/treated wet 19 sets (n = untreated/treated thermal 7 sets (n = re-usable soil 14 samples (n = non-treatable soil 6 samples (n =
set is a
38) 14) 14) 6)
Dredged sediments
untreated/treated wet 2 sets untreated/treated biological 1 sets untreated/ripened 2 sets ripened sediments 2 samples remark: 1 two outputs (coarse and fine granular) and one input (3 samples per set).
(n (n (n (n
= = = =
61) 2) 4) 2)
2.2 A d d i t i o n a l data from other studies
The data set obtained from selected soil samples and laboratory testing was expanded using data from other studies on leaching characteristics of different types of soils. These soils were: natural soils [4]; contaminated soils from the Rotterdam [5] and Amsterdam [6] region. the leaching database of the Centre for Soil Treatment (SCG). In this way a database containing data on contaminant concentration and leachability of a total of ca. 230 samples of soil and dredged material was compiled.
3. L A B O R A T O R Y TESTING The following parameters were determined for 86 samples investigated: contaminant concentrations: Cu, Zn, Pb, Cd and Hg. Cyanide was also determined in soils treated by wet techniques (classification, flotation, etc.). Ni was determined in dredged sediments; As was determined mainly in non-treatable and re-usable soils; leachability of the above-mentioned contaminants using the column test (NEN 7343). Leachability was determined only for contaminants with soil concentrations
-
682 above Dutch background levels; leachability of the above-mentioned contaminants using the CEN TC 292 two-step batch test; availability of the above-mentioned contaminants for leaching (NEN 7341). Availability was determined only for contaminants with soil concentrations above Dutch background levels; analysis of specific soil parameters: lutum, fraction < 63 #m, organic matter, cation exchange capacity (CEC), chloride, sulphate, carbonate, phosphate, iron, sulphide, pH and electric conductivity.
4. STATISTICAL EVALUATION AND RESULTS The data gathered were stored in a database-structure and were statistically analysed with SPSS 7.0 for Windows", using non-parametric statistics. Correlations between soil parameters and leachabilities (determined by several different leaching tests) have been calculated using the Spearman rank correlation coefficient. The use of non-parametric statistics was preferred, because of the possible existence of outliers and a large number of leachabilities below the detection limits. Most striking observations are: at near neutral soil pH-values and soil concentrations below Dutch intervention levels, heavy metal leachabilities are below Dutch legal re-use standards in more than 95 % of the cases; at pH < 5, the pH of the soil is strongly correlated with the percentage of the total amount of heavy metal leached (relative leachability of heavy metals); leachability of cyanide nearly always exceeds the Dutch legal maximum limits for re-use; there is no correlation between typical soil-characterizing parameters (e.g. calcite, organic matter) and leachability; wet soil treatment (classification/flotation) and, in particular, thermal soil treatment generally lower the leachability. leachability and contaminant soil concentration are not correlated within the relevant concentration bandwidth (background level to intervention level) [7] . Figure 1 shows the leachability (column test) versus contaminant concentration for copper. the results of the column-test (NEN 7343) and the 2-step batch test (CEN TC 292) show a positive correlation. This is illustrated in figure 2 which shows the correlation between these two leaching tests for copper, which has the highest correlation. However, this correlation has a very large bandwidth, which reduces the prediction precision of the CEN-test in relation to the column-test L/S= 10 (which is the reference test in the Netherlands). a short column test L/S= 1 is also positively correlated with the complete column test L/S= 10 (NEN 7343) (illustrated for copper in figure 3). Although the bandwidth is somewhat smaller, extrapolation from L/S= 1 to L/S= 10 introduces an additional inaccuracy in the predictability, because part of the leaching occurs in the interval between L/S = 1 and L/S= 10. Remark: the leachability or emission limit (U1) for the re-use of contaminated materials, -
-
-
-
-
-
683 as stated in the Dutch Building Material Act [1], differs according to the application depth of the contaminated material. Throughout this paper the value of U1 is related to an application depth of 0.7 m. 12.6 9
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Figure 1. Cu-leachability (column test L / S = 10) versus Cu soil concentration (labels are pH-values).
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684 4 E ~UI
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Figure 3. Column test leachability for copper.
5.
(L/S=10)
leachability
versus
short column
test
(L/S=1)
FAST P R O C E D U R E S F O R D E T E R M I N I N G L E A C H A B I L I T Y
Because the legal maximum leaching limits in the Netherlands are based on the column test L / S = 10 (NEN 7343), a fast procedure can only be used as a tool for predicting the results of the column L / S = 10 test. Available options are: 1. a faster test; 2. an empirical model, based on the relation between soil parameters and leachability; 3. testing at maximum contaminant concentration and pH bandwidth.
5.1 Faster leaching test -
For this option, two tests are taken into consideration: the CEN TC 292 test; the column test L/S = 1. Both tests save about 3 weeks, as compared with the full ( L / S = 10) column test.
There is a risk that a fast leaching test result will be below the legal limit whereas the mandatory test result will exceed the legal limit (U1). This risk can be set at an acceptable level: e.g. 5 percent 9 The results of the fast test should consequently be compared to
685 newly adjusted, derived limits (U*) with a 95% reliability that the result of the mandatory test will not exceed the legal limit (U1). This is illustrated in figures 2 and 3, by the 5 % risk (upper limit) line of the linear regression (data are log-transformed to achieve normality). Table 2 gives the derived leaching limits, based on the available dataset, and the percentage of results below the derived leaching limit. Table 2 Derived leaching of the mandatory (U1). metal U1 (mg/kg)
limits (U*) for the CEN and column L/S= 1 test, below which the result column test with 95 % reliability will not exceed Dutch leachability limits results < U1
(%) As Cd Cu Hg Pb Ni Zn Remark: Material
U* CEN (mg/kg)
results < U* CEN
U* L/S= 1 (mg/kg)
(~)
0.88 92 0.025 41 0.032 94 0.00017 79 0.72 94 0.12 60 0.018 90 0.00035 57 1.9 99 0.23 86 1.1 93 0.017 68 3.8 94 0.7 71 U1 (and derived U*) are based on emission limits as Act [1] for an application depth of 0.7 m.
results < U* L / S = I
(%) 0.006 0.0012 0.056 0.00032 0.04 0.018 0.38 stated in the Dutch
44 69 83 62 90 73 82 Building
The percentages below U* are comparable for both datasets. Note that the percentage below U* is lower than the percentage below U1. This is due to the chosen risk limit of 5 %. Soils with a fast test leachability above U* have a risk higher than 5 % of exceeding U1 (Dutch leaching limit for re-use) and should be tested with the mandatory column test (L/S= 10). Sufficient data have now been gathered to justify the conclusion that both fast tests can be used for Cd, Pb, Zn and Cu. With regard to As, Hg and Ni and other metals not included in this study, more data are necessary to obtain a more precise prediction. There is no statistical preference for the one or the other studied option for this fast leaching test. The CEN-test uses the same L/S-10-ratio as the full L/S= 10 column test, which means that time-dependent leaching behaviour can be compared more easily. Further the CEN-test is probably better compatible with (developing) European regulations. However, if the result of the fast test does not comply with the derived limits, a full mandatory column test L/S= 10 (NEN 7343) should be carried out. The advantage of the short column L / S - 1 test is that it is the first step of the full column test and can simply be prolonged.
5.2. Empirical model The anticipated relation between contaminant soil concentrations and leachability could not be confirmed by statistically significant correlations. A combination of concentrations and soil specific parameters (calcite, organic matter etc.) did not yield significant correlations with leachability either. Therefore it is concluded that an empirical leaching model that uses the soil characterization parameters analysed in this study is not feasible.
686 5.3. Maximum
levels
An examination of the overall database (230 soils including earlier studies [4,5,6], containing As, Cd, Cr, Cu, Hg, Ni, Pb, Zn and cyanide) reveals that only 0 (Cr) to 9 (As) % of all observations on heavy metals and arsenic exceed the Dutch legal leachability limits (U1). This percentage can even be reduced when soils comply with two preconditions: the contaminant concentration is below Dutch intervention level; the pH of the soil is above 5. For all soils that complied with these two conditions this percentage decreased to 5 % or less. This implies that over 95% of all soils (that comply with the above-mentioned preconditions), meet the standards that Dutch legislation has set for re-use in civil and public works. This finding is in agreement with the results of other recent studies, e.g. a study on the leaching behaviour of zinc in contaminated Meuse sediments [8]. It would seem therefore that leaching tests on soils that comply with mentioned preconditions may be skipped, since in 95 % of the cases their leachability will meet Dutch legal standards. Exceptions to this rule may be: thermally treated soils with respect to Mo, Sb, Se; these metals may have a higher leachability in thermally treated soils [3]; ripened dredged sediments which may have elevated leaching of sulphate, chloride and bromide [9]. However, leachability of these anions is strongly correlated with their concentration, which probably gives a good indication of the result of the column L/S= 10 leaching test (NEN 7343). soils with cyanide concentrations above background level.
-
6. CONCLUSIONS In 95 % of the 230 cases the soil samples contaminated with the following heavy metals: As, Cd, Cr, Cu, Hg, Pb, Ni and Zn, with contaminant levels below Dutch intervention levels and pH above 5, showed leaching levels below the legally set Dutch limits for reusable contaminated soils. It would seem therefore that treated and untreated soils, with heavy metal concentrations below intervention level and pH above 5, in fact comply with the Dutch Building Material Act [1]. However, soils containing other contaminants or soils which do not meet the preconditions for pH and metal concentration must still be tested for leachability. A fast leachability test can be used to obtain a prescreening. This fast test could be: (1) the 2-step CEN TC 292 batch test or (2) the short column L / S = I test. The results of these fast tests can be translated into the mandatory column L/S= 10 (NEN 7343) test by stating an adjusted leachability level below which the mandatory column test with 95% confidence level will not exceed the legal limits. There is no statistical preference for the one or the other fast leaching test. From the practical and regulatory point of view, both tests have their own advantages and disadvantages. Apparently, in most cases, a fast leachability determination procedure is possible. There will then be no need for extended temporary storage while awaiting test results. This will result in significant cost reductions for handling and logistics.
687 7. RECOMMENDATIONS A leaching protocol for treated and untreated contaminated soils and dredged materials needs to be developed within the framework of a quality assurance system. In this protocol guidelines must be given for situations where no leachability tests are required. This protocol can be based on the results of this study. The precision of the observed relations and derived reliabilities can be further improved by continued systematic gathering of data of soils and dredged materials, that become available in the future. Special attention should be given to possible critical parameters: oxy-anions in thermally treated soils and sulphate, chloride and bromide in (ripened) dredged materials and cyanide. The data currently available are reliable enough for applying a fast leaching determination procedure, as described in this paper, to treated and untreated soils contaminated with Cu, Pb, Cd and Zn.
8. REFERENCES Bouwstoffenbesluit Bodem- en Oppervlaktewaterenbescherming, Staatsblad 1995, 567. (Dutch Building Material Act). R.N.J. Comans, Modelling processes controlling metal leaching from contaminated and remediated soils, these proceedings. C. Zevenbergen et al., Immobilization of heavy metal contaminated soils by thermal treatment at intermediate temperatures, these proceedings. P.G.M. de Wilde, J. Keijzer, G.L.J. Janssen, Th.G. Aalbers, C. Zevenbergen, Uitloogkarakteristieken en chemische samenstelling van referentiegronden, RIVM report nr. 216402001, August 1992. IWACO, Uitloogonderzoek aan verontreinigde grond afkomstig uit de Rotterdamse regio, IWACO report nr. 1028100, December 1992. IWACO, Uitloogonderzoek aan verontreinigde grond afkomstig Amsterdamse regio, IWACO report nr. 1037720, November 1993.
uit
de
J. van Leeuwen, A. Orbons, E. van Gent and Th.G. Aalbers, Uitloging van zware metalen uit stads- en natuurgronden, Bodem. Vol.4, nr.2, pp. 83-86, May 1995. A.L. Hakstege, J.J.M. Heynen and H.P. Versteeg, Beneficial use of contaminated sediments within the Meuse river-system, these proceedings. J. Heynen, S. Ouboter and P. Kroes, Chemische aspecten bij het rijpen en nuttig toepassen van verontreinigde baggerspecie, Rijkswaterstaat DWW reportnr. W-DWW-96-043, May 1996.
This Page Intentionally Left Blank
Goumans/Senderffvan der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997Elsevier Science B.V. All fights reserved.
689
E L E C T R O K I N E T I C TRANSPORT IN NATURAL SOIL CORES Douglas I. Stewart, L. Jared West, S. Richard Johnston and Andrew M. Binley Abstract
ElectroMnetic transport in natural soils has been investigated by applying a constant voltage across 500mm long by approximately 200mm diameter natural soil cores for periods of up to 8 weeks. Contaminant ions were circulated through afluidfilled reservoir between the anode and the soil and distilled water was circulated through a similar reservoir adjacent to the cathode. During the experiments electrical current, voltage along the core, water flow rate, and anolyte and catholyte p H were monitored at regular intervals. Periodically, the electrical supply to the power electrodes was switched off and detailed electrical measurements were made using 68 monitoring electrodes positioned around the soil core, in order to produce three dimensional electrical resistivity images of the cores. After testing the cores were dissected and analysed for contaminant content, pore fluid composition, and pH. Data are reported that show that zinc tracer transport is initially strongly retarded, with the zinc predominately sorbed to the soil. Initially zinc enters the anode region mainly by electromigration. This cannot change the pore fluid ionic strength due to charge balance constraints, and hence zinc influx is limited by initial ionic strength. However, after 8 weeks of testing, the ionic strength of the pore fluid in the anode half of the cores were significantly elevated by co-diffusion and electroosmotic advection of the anolyte into the core. Introduction Electrokinetic treatment is a developing technology for treating contaminated land. An electric current is passed through the soil causing migration of charged species towards collection wells. In fine grained soils pore water flow is also induced. Many laboratory studies have been performed to characterise electrokinetic transport, and the dominant processes have been identified (e.g. Eykholt and Daniel, 1994; Hamed et al., 1991; Yeung and Mitchell, 1993), and insitu field decontamination has been attempted (e.g. Lageman, 1993).
This paper reports a study investigating electrokinetic transport in natural soil cores. The aims were to characterise the interactions between natural soil and the contaminant, to investigate the influence of organic matter on electrokinetic transport, and to establish the usefulness of electrical resistivity tomography (ERT) for monitoring electrokinetic transport in the soil cores. ERT is a non-invasive technique where the spatial variation of resistivity is determined from measurements taken with an array of electrodes, which during insitu decontamination would be located in boreholes or along the ground surface. Methodology
Core preparation The cores used in this study were taken from the Lancaster University Hazelrigg field station. The sub-soil at this site is an orange/brown clayey loam containing rounded gravel particles up to 15cm in diameter. The mineralogy of this soil determined by quantitative XRD using a position sensitive detector (M. Batchelder, pets. comm.) is 46% quartz, 18% kaolinite, 14% montmorillonite, 11% illite, 5% feldspar, and 6% aluminium and iron hydroxides. The
690
Figure 1: Soil core during electrokinetic testing topsoil is brown/grey silty loam, 30-40cm deep, with most of the root material in the upper 10-20cm. The moisture content of these samples were all around 20% by weight. Soil cores were carefully excavated by hand. The upper horizon of top-soil was removed (1520cm, sufficient to remove visible plant mass). Each core was exposed by careful hand excavation, aiming for a sample diameter of 200mm (actual diameters were typically 240mm). Samples of the soil trimmings were taken for analysis every 10cm over the depth of the core. An electrode shell and geotextile were then placed on the top, and the entire core and electrode shell were coated with fibre-glass resin and fibre glass matting was applied. This was repeated twice more, and a final coat of resin was applied. The core was then left overnight for the resin coating to dry. The next day, it was under-cut and inverted, and sealed in polythene to prevent moisture loss. Once in the laboratory, each core was trimmed, a second reservoir shell was positioned and sealed with glass fibre and resin. Next, 68 monitoring electrodes were inserted by drilling through the fibre-glass. Finally, the power electrodes and reservoir packing (a drainage geogrid) were inserted into the electrode shells, and the core was saturated by upward flow under approximately 1m head. Prior to testing, each core was positioned horizontally and the reservoirs were filled with the electrolyte solutions (see Figure 1).
Electrokinetic testing During testing a constant voltage of 30V was applied between the power electrodes, and reservoir header-tank weights, electric current, and the voltage along the core were recorded at regular intervals. Test durations were nominally 2 weeks, 4 weeks, and 8 weeks. Periodically data were gathered for ERT image reconstruction (while electrical supply to the power electrodes was disconnected) using a computer controlled earth resistance meter. A total of 1860 'four-electrode' measurements were taken using combinations of the 68 monitoring electrodes, in each of four radial planes and two axial planes. Each 'four-electrode' measurement consists of driving current between two electrodes and measuring the resulting potential difference between the other two electrodes. Electrode polarisation is minimised by using a low frequency alternating current. ERT data acquisition took approximately 4 hours. After testing, the cores were sliced into 50mm thick layers. Each slice was then photographed from the top and bottom, before the slice was homogenised. The pH of each slice was
691 measured by inserting a pH penetration electrode directly into the soil. Samples of each slice were then oven dried at 105~ to determine the moisture content, and the samples were analysed for elemental composition using X-ray Fluorescence Spectroscopy. Samples for XRF analysis were ground and compressed into pellets for analysis. Pore water was extracted by centrifuge displacement (at approximately 50,000g) using an immiscible organic solvent (technical grade tetrachloroethylene). The pore water samples were passed through a 0.45 micron filter, and sub-samples to be analysed for metal content were acidified. The pore fluid pH and conductivity were measured. Pore fluid cation and silica concentrations were measured by Induction Coupled Plasma Atomic Emission Spectroscopy (ICP-AES), and anion concentrations were measured by Eluent Suppressed Ion Chromatography (using a carbonate eluent). Total carbonate concentrations in the pore fluids were measured separately using the flow injection analysis procedure described by Hall and Aller (1992).
Electrical resistivity tomography data processing. A resistivity distribution was found for each set of four-electrode measurements by numerical non-linear inversion. The specimen was represented by a three-dimensional finite element mesh of 2288 triangular prismatic elements. This includes additional layers of elements at both ends of the specimen to account for the change in resistivity in the reservoir electrolyte. The resistivity of each finite element is treated as a variable for the inversion process. The data inversion employs a weighted least squares approach with regularisation. No constraining of the procedure using other measurements such as reservoir fluid conductivity was employed. More details of the three-dimensional ERT image reconstruction can be found in Binley et al. (1996). Results Data summarising each test is presented in Table 1. Voltage profiles taken from two rows of monitoring electrodes showed that the voltage applied across the cores were dropped steadily along their length. Variations in core conductivities were not discernible from this data. After problems with electrode degassing had been resolved, the electrical current through the cores was between 50 and 65mA and did not vary much during testing.
The moisture content in the anode half of the cores did not change much from field value of 20%, but in the upper, organic rich horizon it increased to between 30-40%. This increase in the moisture content was probably the result of pre-test core saturation. Electroosmotic flow data for test HR1 were erratic during optimisation of the test conditions. After optimisation, flow rates in specimens HR2 and HR3 varied between 100 and 300ml/day and tended to reduce during the test. Table 1" Test details Test [[ Duration
U
(days)
Charge passed (kC)
HR1 16 32 HR2 29 124 HR3 57 277 Note: Applied voltage was 30V in all three tests
Energy (k J) 968 3,820 8,464
Average EO permeability (cm2/V.s) 1.6xl 0 .5 0.Sxl 0-5
692
[-.._~,__ HR1 ~
HR2
~
HR3
5000 _ "~ 4000_ O'}
E
3 0 0 0 __
.m (9
> 2000 u .c_ 1 0 0 0 . N 0
0
0.2
'" 0.4
Normalised
-
: -0.6
-
I'0.8
--
', 1
dist. f r o m a n o d e
Figure 2: Total zinc level after core testing Figure 2 shows total zinc profiles from tests HR1-3 (the background zinc level was below 40mg/kg everywhere). The anolyte was originally Zn(NO3)z at a concentration of 1,000mg/1 of zinc. Figure 2 shows that zinc was transported into the cores from the anolyte. After 2 weeks of voltage application, the zinc level had risen to over 2000mg/kg in the anode slice, with a much smaller increase in the next 50mm slice. After 4 weeks, the level in the anode slice had reached about 4000mg/kg, and zinc levels were also elevated in the next two slices. After 8 weeks, zinc levels were elevated to around 4000mg/kg in all but the cathode slice (where it had only increased to about 500mg/kg). The initial soil pH decreased with depth from about 6.5 to 5. The soil pH after testing was typically about 5, with some increase towards the cathode, reaching pH 8 in the cathode slice of the core tested for longest (HR3). The pH of the catholyte was about 11 at the end of all the tests. Any Zn 2+ approaching the cathode reservoir is therefore likely to have been precipitated. Figure 3 shows the conductivity of the pore fluid from the background samples, and that from cores HR1-3 after testing (the values for normalised distances of 0 and 1 represent the reservoirs). In the background soil, the conductivity was highest (1600~tS/cm) near the ground surface (corresponding to the cathode end of the cores), dropping over 250mm to a value of 200~tS/cm. Pore fluid conductivity was steady at about 200ktS/cm over the next 250mm (corresponding to the anode half of the cores). After 2 weeks of testing, there was little change in pore fluid conductivity. After 4 weeks, the pore fluid conductivity in the anode slice had increased to 700~tS/cm, but with little change over the rest of the core. After 8 weeks of testing, the pore fluid conductivity in the anode half of the specimen had risen to nearly 1500ktS/cm, with an even larger increase in the anode slice (3000~tS/cm). The conductivity near the cathode was below the background value although this may reflect natural variability. Figure 4 shows the pore fluid composition of the background samples, and at the end of tests HR1-3. In the background soil, the pore fluid has elevated concentrations of calcium and nitrate near the ground surface (corresponding to the cathode end of the cores). Other ions present in the pore fluid from background samples were sodium, potassium, chloride, sulphate
693 o 14000
~"
12000
bg
=
HR1 ~ H R 2
.__ HR3
t
10000 8000 6000
4000 o
2000 0
0
0.2
0.4
0.6
0.8
1
Normalised dist. from anode
Figure 3" Conductivity of the extracted pore fluid and carbonate. The elevated concentrations of Ca 2+ and NO 3- near the surface may reflect fertiliser application. The data from core HR1 show that, after 2 weeks of testing, the pore fluid compositions are similar to those from the background soil. There were small increases in Zn 2+ in the anode slice and of natural cations (K § and Na +) in the cathode reservoir (relatively high calcium and nitrate concentrations in core HR1 probably reflect natural variability). The carbonate present in the cathode reservoir was probably HCO 3- produced by interaction of electrolytically produced OH- with atmospheric CO 2. The data from core HR2 shows that, after 4 weeks of testing, the pore fluid compositions remained similar over most of the core, although there was an increase in Zn 2+ in the 100mm nearest to the anode, and there was a significant increase in Ca 2+ in the cathode reservoir. The data from core HR3 show that, after 8 weeks of testing, there were elevated Zn 2+ levels in the two-thirds of the core closest to the anode and relatively low calcium and nitrate levels near the cathode. However, the pore fluid zinc levels remained much less than the total levels, which shows that most of the zinc was sorbed on contact with the soil. Figure 5 is a plot of the Zn 2+ in pore fluid against sorbed zinc, for those parts of all three cores which were composed of orange brown clay loam (i.e. excluding the topsoil). Despite the scatter, it can be seen that Zn(II) exhibits Freundlich sorption behaviour where the sorbed concentration is less than 4,000 mg/kg. Sorbed concentration did not exceed this value for higher pore fluid Zn 2+ levels. Figure 6 shows ERT images of core HR2 at intervals during testing (the elements representing the fluid reservoirs are included on the images). The main feature that is evident from the images is an increase in the catholyte conductivity due to electrolytic decomposition of water. The zone of conductivity increase in the images does not correspond exactly with the reservoir. This smearing is typical of ERT reconstruction for zones outside the electrode
694 ;
Zn
=
Ca
=
Na
,t
K
e
NO3
o
CO2
o
CI
A SO4
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--
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i
i
0.2
-
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-
,
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1
(b)
lO 0
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(c)
=-
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,
~
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,
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0.2
0.4
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1
7o 6o
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~ 40 "~ 30 "~ 20 o
lO o
(d)
60 50 40 30 20 10 0
N o r m a l i s e d dist from a n o d e
Figure 4" Composition of pore fluid (a) background, (b) HR1, (c) HR2, and (d) HR3
695 10000 9
m
A 9
b
9
1000
o'J
E
G =
. , -
N o9
100 9
L_
o
10
9
, HR1 1 0.01
I 0.1
9HR2
9HR3
]
I
I
I
I
I
I
10
100
1000
10000
P o r e Fluid Zinc, mg/I
Figure 5" Pore fluid zinc concentration versus sorbed zinc for the clayey loam arrays. Apart from this effect, the images show that the conductivity of the soil changed little over four weeks of testing. Discussion
The measured pore fluid anions and cations in cores HR1 and HR2 are approximately in charge balance, indicating that all major ionic species have been identified. The measured ionic concentrations indicate a large excess of positively charged species in most of core HR3, and in the cathode reservoirs at the end of tests HR2 and HR3. The ion chromatograph results for these regions have unidentified peaks, the most significant of which had a relatively long residence time in the ion chromatography column. It is inferred that interaction of dissolved species with either the equipment or the soil released (as yet) unidentified anionic species, which migrate towards the anode. Ideal solution conductivities were calculated from the ionic compositions using the dilute solution ionic mobilities (e.g. see Reiger, 1994). The ideal solution conductivities were typically higher than the measured pore fluid values, with the difference being largest where the ionic strength of the pore fluid was highest. This is consistent with the non-ideal behaviour of real electrolytes. The fractional contribution of each ion to the extracted pore fluid conductivity (its transport number) was calculated by assuming that (i) deviation from ideal behaviour within the pore fluid from each slice was the same for all ions and (ii) that all the major ionic species had been identified (this was not the case in core HR3). Within the cores, other effects such as diffusion and electroosmotic flow will modify transport numbers from those calculated for the extracted pore fluid (Alshawabkeh and Acar, 1996). Figure 7 shows the major contributors to pore fluid conductivity (defined here as ions with transport numbers greater than 0.1) in the cores. Each sub-region indicated in Figure 7 has the same major contributors, although ionic concentrations varied substantially within subregions. These diagrams, although an approximation to the main charge carriers in the cores, are a useful for identifying the major processes during electrokinesis.
696
Anode reservoir
Cathode
Soil c o r e
......... ; L
.a~,,~P-ffs;r;oir
q o cm
t=0
t = 7 days
t = 28 days
t = 14 days
Resistivity (ohm m) 10
20
30
40
50 60
70 80
Figure 6: ERT images of core HR2 The pore fluid ionic strength after two weeks of electrokinesis had not changed much from background values, despite the increased ionic strength in the reservoir (anions produced by electrolysis and cations from the core raised the level in the cathode reservoir). The major processes that have occurred are electromigration of zinc from the anode reservoir, replacing the natural pore fluid calcium near the anode, and electromigration of HCO 3- from the cathode reservoir, replacing nitrate and sulphate near the cathode. After 4 weeks, the pattern remains similar, although increases in pore fluid strength are now apparent in both the anode and cathode slices. These local increases probably result from codiffusion of anions and cations from the reservoirs, and in the anode slice, electroosmotic advectiono Charge balance constraints prevent electromigration alone from causing these ionic strength variations (i.e. zinc can only enter the anode region by the electromigration mechanism at the same rate that natural cations are leaving it).
697 Background inn
Ca 2+
Ca2+
NO 3SO42-
NO 3-
Anode Res.
Soil
Cath. Res.
Core HR1 .|
ill
Zn 2+ Ca 2+
Ca 2+
NO3- ] NO3SO42-
SO42-
SO42- t NO 3" NO 3- SO42-
Zn 2+ | Zn 2+
Zn 2+
Zn 2+ Na +
Ca 2+
Ca 2+
Ca 2+
Ca 2+
NO 3- NO 3-
SO42C1-
SO42-
SO42-
NO 3SO42-
NO 3-
HCO 3-
Zn 2+
Ca2+! Ca 2+
Ca 2+
Ca 2+
K
+
Na +
NO 3-
NO 3- HCO 3-
Core HR2 Ca 2+
_
HCO 3-
Core HR3 Zn 2+
Zn 2+
NO 3- NO 39-
Zn 2+
?-
Zn 2+
Ca 2+
SO42- HCO 3HCO 3-
?- indicates a region where there are unidentified anions. Figure 7: Distribution of major contributors to extracted pore fluid conductivity
Ca 2+
9-
698 After 8 weeks, the ionic strength in the anode half of the core had risen substantially, but had reduced slightly in the cathode half. The ionic strength in the anode half rose as anions accumulating in this region (due to diffusion and electroosmotic advection) are not used up in precipitation or electrode reactions. This allowed higher concentrations of Zn 2+ to enter from the reservoir The pore fluid data show that the requirement for continuity of current and charge balance within the cores prevents electromigration alone from altering ionic strength (it can only substitute ions for others of equivalent charge). Changes in ionic strength at a specific location are therefore the result of advection or diffusion of electrolytes to that location, or dissociation/neutralisation and dissolution/precipitation reactions at that location.
Conclusions O Initially, most of the zinc tracer entering the core was sorbed to the soil, whereas after 8 weeks of testing the sorption capacity had been exceeded across 60% of the specimen, permitting larger pore fluid zinc concentrations. | Initially, there was little change in pore fluid conductivity. This is because the dominant transport process is electromigration, which alone cannot alter ionic strength. | Later, co-diffusion and electroosmotic advection of the anolyte into the cores progressively increased the ionic strength of the pore fluid. Acknowledgements The authors would like to acknowledge the support of the U.K. Engineering and Physical Sciences Research Council through grant GR/K57770. M Batchelder, Natural History Museum is thanked for the quantitative XRD analysis. Appendix. References Alshawabkeh, A.N. and Acar, Y.B. (1996). Electrokinetic remediation. II: theoretical model. ASCE Journal of Geotechnical Engineering, 122(3), 186-196. Binley, A, Pinheiro, P. and Dickin F., (1996), Finite Element based Three-Dimensional Forward and Inverse Solvers for Electrical Impedance Tomography, In: Proc. Colloquium on Advances in Electrical Tomography, Computing and Control Division, lEE, Digest No. 96/143, p6/1-6/3, Manchester, June, 1996. Eykholt, G.R. and Daniel D. E. (1994). Impact of system chemistry on electroosmosis of contaminated soils. ASCE Journal of Geotechnical Engineering, 120(5), 797-815. Hall, P.O.J. and Aller, R.C. (1992). Rapid small volume flow injection analysis for total CO2 and ammonium in marine and freshwaters. Limnology and Oceanography, 37, 1113-1119. Hamed, J., Acar, Y.B. and Gale, J.G. (1991). Pb(II) removal from kaolinite by electrokinetics. ASCE. Journal of Geotechnical Engineering, 117(2), 241-271. Lageman, R. (1993). Electroreclamation: applications in the Netherlands. Environmental Science and Technology, 27, 2648-2650. Reiger, P.H. (1994). Electrochemistry. Chapman and Hall, 2nd ed. Yeung, A.T. and Mitchell, J.K. (1993). Coupled fluid, electrical and chemical flows in soil. Geotechnique 43(1), 121-134.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
699
Re-use of Sieve sand from demolition waste E v e r t Mulder
T N O - Waste Technology Division P.O.Box 342, 7300 AH Apeldoorn The Netherlands Phone: +31 55 5493919, Fax: +31 55 5493287
Abstract Sieve sand originates from activities as sorting and/or breaking of demolition waste. In a breaking process the first step is a sieve step, to remove the fines. Next, the coarse material is broken and upgraded to a secondary raw material, which can be used as a coarse aggregate in concrete, or as a road construction material. The fine material (sieve sand) cannot be applied without due consideration. The sieve sand may be contaminated with Poly cyclic Aromatic Hydrocarbons (PAHs) to such an extent that, according to the Dutch Building Materials Decree, the sieve sand is not applicable as a granular building material. For this reason Van Bentum Recycling Centrale ordered TNO to carry out a research into the possibilities of stabilising/solidifying the sieve sand in such a way that the PAHs are fixed (immobilised). In the Building Materials Decree the organic contaminants are assessed on the basis of total concentration, in mg/kg. However, from an environmental point of view, not the total concentration, but the leaching is relevant. For this reason, the impact on the environment of the application of stabilised sieve sand has been assessed in terms of mg leached per square meter surface area by TNO. The results of the leaching tests performed, show that even highly contaminated sieve sand (containing up to 1,000 mg/kg PAils) can be sufficiently stabilised. Only 0.7 mg/m 2 (being 0.002 % of the total concentration) is being leached during the 64 days lasting Dutch diffusion test.
Introduction In The Netherlands most of the demolition waste is being upgraded to a secondary raw material and re-used as a coarse aggregate in concrete, or as a road construction material. To obtain a material of high quality the demolition waste is sieved first, to remove the fines. This fine material (sieve-sand) should not be disposed of, but utilised, both from an economic and environmental point of view. On the other hand, it cannot be applied without due consideration. The sieve-sand may be contaminated with Poly cyclic Aromatic Hydrocarbons (PAHs) to such an extent that, according to the Dutch Building Materials Decree, the sieve-sand is not applicable as a granular building material (as will be explained in the next paragraph). For this reason Van Bentum Recycling Centrale ordered TNO to carry out a research into the possibilities of stabilising/solidifying the sieve sand in such a way that the PAHs are fixed (immobilised). The intention is to use the stabilised sieve sand to heighten a piece of land for use
700 as an industrial area. The stabilised sieve sand has to be environmentally assessed as a monolithic material, on the basis of leaching. In this paper the characteristics of sieve sand will be described first. Then a description will be given of the stabilisation process, that was used to immobilise especially the PAHs. After that the results of a leaching test on the stabilised sieve sand are given and the stabilised material is environmentally assessed. Finally the paper ends with some conclusions and a recommendation.
Sieve sand and its leaching characteristics Sieve sand originating from a breaker of demolition waste, consists for the greater part of sand and small concrete and ceramic brick particles. Besides, it may also contain small particles of wood, roofing material or plastics. Though these kind of "physical" contaminants are relevant for the strength development of the stabilisation, environmentally the "chemical" contaminants are of more importance. The most critical chemical contaminants are PAHs and sulphate. The PAHs are present in tar-containing particles, originating from for instance roofing material, chimneys or tar containing asphalt concrete. The sulphate originates from mortar and plasterboard. For the research, described in this paper, a sieve sand sample was used containing high concentrations of contaminants, as a worst case. Only the PAH content was measured. The sample contained 1000 mg/kg PAH (the 10 of the Dutch Ministry of VROM). This is much more than is allowed by the Dutch Building Materials Decree. As for the environment, not the total content of contaminants in a material is important, but only that part of the contents that will leach out, in its application. For that reason in The Netherlands the environmental assessment of building materials is primarily based on leaching (at least for inorganic contaminants). Because of the fact that for organic components standardised leaching tests are not available yet, the assessment of organics is still based on total content. Nevertheless, in this paper also PAHs will be environmentally assessed on the basis of leaching. For this reason part of the sieve sand sample was leached in a column test for organic components, in accordance with the draft standard NVN 7344. In this column test, up to a liquid / solid (L/S) ratio of 10 1/kg, totally 0,26 mg/kg PAHs were leached. Even if the Dutch Building Materials Decree should assess the utilisation of granular (unbound) sieve sand on the basis of leaching (instead of total concentration), this sample of sieve sand would not be allowed to be utilised. The calculated immission of PAHs would be two times higher than the limit value of a category 2 application [1]. However, from an other investigation it follows that a sample with a lower total concentration of PAHs (490 mg/kg) has leaching characteristics that would allow it to be utilised as a category 2 construction material [2]. Not only the leaching of PAHs was determined, but also the leaching of inorganic contaminants. The leaching of the heavy metals that were investigated (for instance arsenic, barium, molybdenum and antimony) was far below the limit values of the Building Materials Decree (even the category 1 limit values). On the other hand, the leaching of sulphate was beyond the category 1 limit value, but below the category 2 limit value. From these investigations it can be concluded that highly contaminated sieve sand should not be utilised as such, from an environmental point of view. According to the Building Materials Decree, the total concentration of PAHs is too high (even though this is environmentally less relevant). Nevertheless, also an assessment based on leaching indicates that this sieve sand
701 should not be used as such, because of the leaching of PAHs. Besides, the leaching of sulphate exceeds the limit value of a category 1 application. Therefore stabilisation is a must.
Stabilisation of PAH-containing sieve sand A research has been carried out to find the best binding agent for the stabilisation / solidification of the sieve sand, described in the previous paragraph. Four different agents were tested in two addition percentages (5 % and 10 %): 9 blast furnace slag cement; 9 geo-cement (a cement produced from secondary raw materials); 9 a mixture of blast furnace slag cement and an additive; 9 a special cement, specifically developed to bind high concentrations of organic components. Three criteria were used in finding the best binding agent. First of all the mechanical strength after 28 days of hardening had to be beyond 5 MPa. The specimens with 5 % binding agent added, did not fulfil this criterion. Of the specimens with 10% binding agent 3 kinds had a pressure strength of more than 5 MPa. The only one that did not, was geo-cement. The second criterion was the leaching of PAHs from the stabilised material, determined by means of the diffusion test (tank leaching test). The leaching of PAHs of the remaining three types of specimens differed from 0.7 mg/m2 surface area (blast furnace slag cement with additive) to 3.6 mg/m2 (special cement). The third criterion was the price of the binding agent. The blast furnace slag cement was much less expensive than the special cement. Concerning the additive, this was little more expensive than the blast furnace slag cement itself, but the mixture performed better than the cement as such, so the mixture of blast furnace slag cement and the additive was chosen as the best binding agent. In the experiments mentioned in the following paragraphs, stabilised sieve sand was investigated, that was mixed up with 9% blast furnace slag cement and 1% of an additive. The material was compacted well in cubes of 10 * 10 * 10 cm. After 28 days of hardening the density of the material was 1900 kg/m3. The pressure strength was 6 MPa.
Leaching of PAHs from stabilised sieve sand The leaching characteristics of a stabilised, monolithic material are in The Netherlands determined by means of the diffusion test (or tank leaching test). This test is standardised for inorganic components (NEN 7345). In this standardised test the specimen (or product) is immersed in five times its volume leachant, consisting of demineralised water, beforehand acidified to a pH of 4 with nitric acid. This leachant is renewed at 0.25, 1, 2.25, 4, 9, 16 and 36 days after the start of the test. At 64 days the test is finished. The eight eluate fractions are filtered, measured (pH and conductivity), conditioned and analysed on relevant components. From the analysis results for each component the quantities leached at the eight different times are calculated (expressed in mg/m2 surface area). These emissions are then plotted against time (on a double logarithmic scale). If the leaching indeed is diffusion controlled, this should yield a straight line with a regression coefficient of 0.5. If so, a diffusion coefficient can be calculated from the leaching data. This diffusion coefficient can be used to predict a diffusion controlled leaching (emission) in course of time by extrapolating the leaching with time.
702 In the research, described in this paper, the leaching behaviour of all inorganic components, mentioned in the Building Materials Decree, was investigated. Only a small number of components could be detected (Ba, Cu, Mo, Ni, Sb, V, Zn, CI, SO4). The results (emissions) are given in table 1. In the Building Materials Decree the organic components are assessed on the basis of total concentration, in mg/kg. However, from an environmental point of view, not the total concentration of a contaminant in a building material is important, but its leaching from the building material. For this reason, the impact on the environment of the application of stabilised sieve sand has been assessed in terms of mg leached per square meter surface area by means of the diffusion test. In principal this test was performed in accordance with NEN 7345, but additionally, some pre-cautions were taken to prevent the degradation and/or the absorption of leaching PAHs. The pre-cautions were: 9 The leaching vessel was made of glass and covered to avoid evaporation of the more volatile PAHs. 9 The leaching vessel was packed in aluminium foil, to prevent degradation of PAHs by ultraviolet radiation of sunlight. 9 The eluates were filtered in teflon filter devices, by means of pressure filtration, to avoid absorption of PAHs in the device. 9 The eluates were put in brown flasks in between the time of sampling and the time of analysis, again to avoid degradation of PAHs. The leaching of PAHs, determined in this way (according to NEN 7345, with additional precautions), proved to be diffusion controlled, on the analogy of the inorganic components. The diffusion coefficient was very low (1,5 * 1015 mE/sec). The emission in 64 days was 0.68 mg/m 2.
Environmental assessment of PAH-leaching In table 1 not only the emissions (in terms of mg/m 2 product surface area) are given, but also calculated immissions, in terms of mg/m 2 soil surface area. These immissions are calculated in order to be able to compare the leaching test results with the limit values of the Building Materials Decree. These limit values (Maximum Allowable Immissions into the soil), are based on the principle of "Marginal Burdening" of the soil. This means that the upper meter of soil may not be contaminated by leaching from building materials in there application by more than 1% of the target values for soil. The Building Materials Decree distinguishes two categories of applications, one without provisions to prevent rain water from coming into contact with the building material (Category 1) and one with those provisions (Category 2). These provisions can be a non-permeable clay liner or a plastic liner.
703 Table 1" Leaching characteristics of stabilised sieve sand Component Emission Immission Immission (calculated) Cat. 1 application Cat.2 application in mg/m 2 in mg/m 2 in mg/m 2 As Ba Cd Co Cr Cu Hg Mo Ni Pb Sb Se Sn V Zn Br C1 CN F SO4 PAHs
< 0.33 16" < 0.07 < 0.67 < 0.33 0.6 < 0.66 1.4 1.1 < 0.67 0.38 < 0.66 < 1.3 5.1 10 < 66 2400* < 20 < 66 51000" 0.68*
< 1.5 170 < 2.2 < 7 < 3.4 6.3 < 1.5 3.6 11 < 7.0 4.0 < 1.5 < 3.0 49 100 < 370 4100 < 150 < 690 85000 7.1
< 0.5 54 < 0.7 < 2.2 < 1.1 2.0 < 0.5 1.1 3.6 < 2.2 1.3 < 0.5 < 1.0 15 33 < 120 1300 < 47 < 220 27000 2.3
Max. Allowable Immission in mg/m 2 435 6300 12 300 1500 540 4.5 150 525 1275 39 15 300 2400 2100 300 30000 75 14000 45000 15
* = calculation based on a diffusion coefficient From table 1 it can be learned that the calculated immissions for most inorganic components are below the maximum allowable immission values, even for a category 1 application, except for SO4. For the anions Br and CN this is no sure, because the analysis techniques are not able to determine such low concentrations yet. The sulphate immission for a category l application is higher than the limit value, whereas the calculated immission for a category 2 application is still below the limit value. So, the stabilised sieve sand should be considered a building material that can be utilised in category 2 applications only (because of the leaching of sulphate). Also for PAHs immissions have been calculated, even though the Building Materials Decree does not give a maximum allowable immission for PAHs (because the Building Materials Decree assesses PAHs on the basis of total content). So, to be able to assess PAHs on the basis of leaching, a "maximum allowable immission" for PAHs had to be derived. This was done, starting from a target value of soil for PAHs of 1 mg/kg and following the same route as was done in the Building Materials Decree for inorganic components. In that way a maximum allowable immission of 15 mg/m 2 was calculated for PAHs.
704 The leaching of PAHs if very low, compared with the total amount. Only 0.002 % of the PAHs present in the stabilised sieve sand are leached during the 64 days lasting diffusion test. From table 1 it can be learned that, if PAHs would be assessed on the basis of leaching, the stabilised sieve sand would not have any problems to fulfil the criteria, even for a category 1 application. The results of this leaching research show that the stabilisation process is capable to decrease the leaching of PAHs to such an extent that that it can be considered harmless to the environment. This is in contradistinction to the conclusion of an assessment on the basis of total content of PAHs. For this reason it is highly recommended to environmentally assess organic components in building materials on the basis of leaching, on analogy to inorganic components.
Conclusions 9 Highly contaminated sieve sand may not be utilised, neither if assessed on the basis of total content of Poly cyclic Aromatic Hydrocarbons (PAHs), nor if assessed on the basis of leaching of these PAHs. 9 This highly contaminated sieve sand (containing up to 1,000 mg/kg PAHs) can be stabilised well by adding 9% of blast furnace slag cement and 1% of an additive. Only 0.7 mg/m2 (being 0.002 % of the total concentration) is being leached during the 64 days lasting Dutch diffusion test. 9 If the stabilised material has to be assessed on the basis of total content of PAHs the material may not be utilised still (not the presence has been effected, but its mobility). 9 However, if the stabilised sieve sand would have to be assessed on the basis of leaching, the material could be utilised in a category 2 application (because of the relatively high leaching of sulphate). The leaching of all other components (PAHs inclusive) is below the limit values of category 1 applications. Recommendation It is highly recommended to environmentally assess organic components in building materials on the basis of leaching (on analogy to inorganic components) and not on the basis of total content. Literature [1]
Zijlstra, R.K., and E. Mulder, Comparison of the shake test (CEN) and the column leaching test (NEN), TNO report No. TNO-MEP - R96/450 (in Dutch), December 1996.
[2]
Zijlstra, R.K., and E. Mulder, Determination of the leaching of PAHs with the column leaching test, TNO report No. TNO-MEP- R96/400 (in Dutch), November 1996.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997Elsevier Science B.V. All rights reserved.
705
Organic substances in leachates from combustion residues I. Pavasarsa, A.-M. FAllmanb, B.
Allard a
and H. Bor6na
aDepartment of Water and Environmental Studies, Link6ping University, SE-581 83 Linkoping, Sweden bSwedish Geotechnical Institute, SE-581 93 Link6ping, Sweden
Abstract
The release of water soluble organic substances from combustion residues (bottom ashes from municipal solid wastes and from wood) has been followed in laboratory batch leaching experiments. Leachates have been sampled during 70 days (single-step procedure) or frequently during first 24 hours (starting after 10 min), followed by a change of leachate solution and continued sampling during 70 days (two-step procedure). The leachates were analysed for conductivity, pH, Eh, TOC and metals (Cu and Cr). High concentrations of dissolved organic matter were obtained already within 24 hours (up to 30 mg/1 of TOC for the solid waste bottom ash). However, a rapid loss of the TOC from the solution (half-life of 50 days) was also observed. The release of copper appeared to be governed by the simultaneous release of organic carbon, while the release of chromium was independent of TOC.
1. INTRODUCTION The technical use of ashes and slags from combustion of various materials (municipal solid waste, wood etc.) is limited by the potential environmental effects due to releases of hamafial components from these products as a result of leaching by percolating precipitation. Most efforts to characterize and quantify the leaching properties of ashes and slags have been focussed on the release of inorganic components under various conditions, not considering the simultaneous release of soluble organic agents. However, the total contents of organic carbon in solid ashes from combustion of e.g. mixed municipal wastes and wood are usually several percent, assessed from measurements of loss of ignition. Total concentrations of organics in leachates in the 10-100 mg/1 range have been reported [ 1, 2]. There are also reports on the composition of the hydrophobic fraction, leached by organic solvents [3]. The aim of the present study is to assess concentrations and eventually metal complexing properties of readily soluble hydrophilic organics in leachates from combustion residues (municipal solid wastes and wood). Effects on the release of copper and chromium related to the simultaneous release of organic matter will be quantified.
706
2. MATERIALS AND METHODS 2.1. Materials Bottom ash from municipal solid waste incineration (BA) and wood (primary wood chips and secondary wood materials) firing (WA) were obtained from energy production plants at LinkOping, Sweden. All ash samples were dried at 50~ and crushed to a size of <1 mm prior to storage and laboratory testing. The total element composition was analysed by ICP-AES or ICP-MS (samples digested in lithium borate or nitric acid; see Table 1). The organic carbon content was assessed from the measurements of the loss on ignition (LOI) at 550~ Table 1 Elemental composition of the ash samples [4]
Element Si Fe Ca Al Na K Mg S Ti P Cu Zn Ba Mn LOI
Bottom ash ~/kg 209 108 87.9 56.9 26.4 14.3 11.3 8.56 6.20 4.50 3.40 3.08 1.66 1.36 43
Wood ash 8/kg 222 28.3 82.5 48.6 12.2 40.3 10.9 5.44 2.90 5.00 0.089 1.84 1.32 4.80 170
Element Pb Sr Cr Zr Ni Sn V W Co As Mo Nb Cd Be Hg
Bottom ash mg/kg 737 285 274 200 138 130 58.7 33.9 19.1 16.0 16.0 13.1 5.80 1.78 <0.4
Wood ash m~/kg 253 407 94.5 168 40.0 64.7 53.9 <14 11.1 38.2 <6 13.5 4.50 2.30 <0.4
2.2. Leaching procedures Approximately 400 g of dried and crushed ash samples and appropriate amounts of MilliQ-water (liquid/solid (L/S) ratio of 5) were put into 2 1Duran glass flasks. Two different leaching procedures were followed. In the single-step procedure the flask containing ash sample and MilliQ-water was agitated for 24 hours (end over end rotation) after which the first sample was taken. The bottle was put on an oscillating shaking table, and sampling was continued with logarithmic time intervals for 70 days. In the two-step procedure the first leachate sample was taken already after 10 min, and sampling then continued with logarithmic time intervals for 24 hours. The leachate was removed after 24 hours by centrifugation at 3000 rpm for 0.5 h and replaced with flesh leaching solution (MilliQ-water). Leaching and sampling continued as in the single-step procedure for 70 days. Two replicates were measured. In addition, a blank sample (without ash) was included in the single-step series.
707
2.3. Leachate analyses The withdrawn samples were filtered (0.45 ~tm cellulose acetate filter) and acidified prior to analysis (TOC and metals). Organic carbon was analyzed by high-temperature catalytic combustion (Shimadzu TOC-5000 analyzer). Concentrations of copper and chromium were determined by atomic absorption spectrophotometry (Perkin Elmer 1100B equipped with a HGA700 oven). Conductivity was measured with a Radiometer CDM210 conductivity meter and pH was determined with a Metrom 632 pH-meter. Redox potential was measured with a Pt-calomel electrode combination. A fractionation procedure was tested on a separate ash system (leaching of B A with 0.1 M NaOH for 24 h) in order to divide the TOC into arbitrarily defined fractions by passage through columns with XAD-8 and XAD-4 in sequence [5]: 9 Hydrophobic acids; adsorbed on XAD-8 at pH 2 and desorbed with 0.1 M NaOH 9 Other hydrophobic agents; adsorbed on XAD-8 at pH 2, but not desorbed with 0.1 M NaOH 9 Hydrophilic acids; adsorbed on XAD-4 at pH 2 but not on XAD-8; desorbed with 0.1 M NaOH 9 Other hydrophilic agents; neither adsorbed on XAD-8 nor on XAD-4 at pH 2
3. RESULTS The 10 dominating elements (besides oxygen) in the ashes account for around 53 and 46% of the mass of BA and WA, respectively. The additional 18 other elements that were determined account for around 10%. LOI indicates maximum organic matter contents of 4.3 and 17% of BA and WA, respectively. This illustrates significant differences between the ashes in terms of the contents of oxygen, as well as organics. The BA generally has higher levels of heavy metals (Fe, Cu, Zn, Pb, Cr, Ni, Sn, Mo etc.) than the WA, while the concentration of alkali and alkaline earth metals are similar in both ashes. Most pronounced are the differences in concentrations of Cu, Fe and Cr (highest in B A), as well as Mn and organic carbon (highest in WA).
3.1. Conductivity The conductivity reflects the total dissolved solids in the leachate [6]. The increase of the conductivity with time is given in Figure 1. Conductivity in step two (change of leachate after 24 hours in the two-step procedure) is added to the final value after 24 hours in step one. The conductivity after 64 days reached the level 450-500 mS/m for both the BA and WA. The final conductivities in the single-step procedure would correspond to some 7 and 10% of the total inventory of monovalent cations (K, Na) in BA and WA, respectively, neglecting contributions from other cations.
3.2. pH Initial pH of 9 was obtained in both the BA and WA systems, Figure 2, however with different developments vs time for the two systems. Both systems exhibited higher pH in the two-step than in the single-step procedure. A maximum pH of 9.8 was obtained during the second step in the WA system, and pH of 9.5 or above was maintained after 64 days. A decrease of pH towards 8.5-8.8 was obtained in the B A system for both procedures.
708
o
600
Bottom ash
500
E --400 09 E 300
~~(a)
(b)
200 "o
" 100
o
o
0 0,1
600 O ?o 500 E
OO
1
10 100 Time, h
"o
"o o
10000
Wood ash
x
400
E .~ 300 >
1000
~
b)
200 100 0
t
0,1
1
t
t
10 100 Time, h
t
1000
10000
Figure 1. Conductivity as a function of time in BA (top) and WA (bottom) leachates (a) Single-step leaching (b) Two-step leaching
3.3. Redox potential Redox potentials of-100 to -150 mV corresponding to pe + pH of 6.5 to 7.5 were obtained at the start of the BA leaching, but increased to around 200 mV after 24 hours (pe + pH of 12). A new low level of around -100 mV (pe + pH of 10) was observed after the change of leachate solution in the two-step procedure (see Figure 3). This level increased slowly with time to a stable level of around 400 mV (pe + pH of 15-16). The redox potential in the closed bottle of the single-step procedure was-450 mV after 24 hours (pe + pH of 2), increasing with time to the same level as in the two-step procedure (pe + pH of 1516). Initial potentials around 400 mV were obtained in the WA system with minor changes with time (constant pe + pH of 15-16). Potentials corresponding to pe + pH of 15-16 (in all WA leachates as well as in the BA leachates after a long time) simply reflect a system in contact with the atmosphere (E ~ of 0.95 V).
709 10,0
Bottom ash
9,5 -r" ct.
9,0 8,5 []
8,0
i
0,1
10,0
I
1
I
o
Q.
~
9,0
~
u
1000
~
....~ ~ - - . -
9,5 "I-
I
10 100 Time, h
Wood ash
(a)
~.,
10000
(b)
--
8,5 8,0
I
0,1
1
I
t
i
10 100 Time, h
1000
10000
Figure 2. pH as a function of time in BA (top) and WA (bottom) leachates (a) Single-step leaching (b) Two-step leaching 20
Wood ash
15"I-
Bottom ash
(a, b) ~. (b"
/~
Q.
+ 10
/
0
I
0,1
1
I
(a) I
10 100 Time, h
Figure 3. pe + pH as a function of time in BA and WA leachates (a) Single-step leaching (b) Two-step leaching
I
1000
10000
710 3.4. Release of organic carbon The accumulated release of carbon, determined from the measured TOC-concentration, is given in Figure 4. TOC-concentration in step two (change of leachate after 24 hours in the two-step procedure) is added to the final value after 24 hours in step one. There were immediate releases of organic carbon into the leachates of both systems, giving TOClevels of 20 and 4.5 mg/l for BA and WA, respectively, already after 10 minutes of exposure. The TOC-concentrations had reached levels of 30 and 5.5 mg/1 after 24 hours, corresponding to a release of around 150 and 30 mg/kg from the BA and WA, respectively. 180 Bottom ash ,~n 160 m 140 "~ 120 E 100 d 80 i Wood ash 60 ttj
~
t~
n-
40
20 0
~
~~-.'~r 0,1
1
~-~=~"='-~
+
*
..... ~
(b)
~a) (b)
=- - ~ ~
~~ " ~ ' ~ -
10 100 Time, h
1000
(a) 10000
Figure 4. Organic carbon release as a function of time in BA and WA (a) Single-step leaching (b) Two-step leaching The carbon release continued after the change of leachate in the two-step procedure, and leveled out at a total release of around 170 mg/kg for the BA system, based on the TOC-concentration. Changes in TOC-concentrations were minor in the WA system in the second step of the two-step procedure. The high levels of the released organic carbon in the BA system represented only some 0.8% of the maximum organic carbon inventory (assessed from the LOI). The released fraction from the WA system was only some 0.03%, and the concentrations much lower than for the BA system, despite the fact that the carbon content was almost four times higher in the WA. The TOC-concentration was decreasing with time, after the initial fast in-growth, in the singlestep leaching for both the BA and WA systems. Half of the maximum TOC-concentration was lost alter about 50 days in the BA leachate, but already after 4 days in the WA leachate. Preliminary results (BA only) indicate the following distribution of leachable (by 0.1 M NaOH) organic matter: 25-30% hydrophobic acids, 5-10% other hydrophobic agents, 10-15% hydrophilic acids and 45-55% other hydrophilic agents. 3.5. Release of Cu and Cr
The accumulated releases of copper and chromium, determined from the measured concentrations, are given in Figures 5 and 6. The release in step two (after change of leachate after 24 hours in the two-step procedure) are added to the values after 24 hours in step one.
711 1,4 ~
Bottom ash
1,2
~
9aa
a ~ [] (b)
13) ~'-~ 9 1,0 o3
~
E =- 0,8 o 0,6 "o (!.}
m n,'
'
~
~
(a)
Wood ash
0,4
o
0,2
.8-
o
A
o ~
(a, b)
o
0,0 0,1
1
10 100 Time, h
1000
10000
Figure 5. Copper release as a function of time from BA and WA (a) Single-step leaching (b) Two-step leaching 0,5 O3
0,4-
03
Wood ash
jm..-~ [] [ ] ~ ] = r ~ o
E 0,3t...-
o -o 0,2ttl
m
0,1-
n,'
Bottom ash
(a, b)
0,0 0,1
1
10 100 Time, h
1000
10000
Figure 6. Chromium release as a function of time from B A and WA (a) Single-step leaching (b) Two-step leaching There was an immediate release of copper into the leachates of the B A systems, giving concentrations of around 120 mg/1 already after 10 minutes and 190 mg/l after 24 hours. Concentrations significantly above the blank values (10-20 rag/l) were not observed in the WA leachates. The copper release continued after the change of leachate in the two-step procedure, and leveled out at a total accumulated release of 1.3 mg/kg for the BA system. This represents around 0.04% of the inventory. The copper concentration was decreasing with time, after the initial fast in-growth, in the singlestep leaching for the B A system. A reduction to half of the maximum copper concentration was observed after about 10 days in the B A leachate. There was an immediate release of chromium into the leachates of the WA systems, giving concentrations of 40 mg/1 already after 10 minutes and 50 mg/1 after 24 hours. Concentrations significantly above the blank values (2-4 rag/l) were not observed in the BA leachates. The
712 chromium release continued after the change of leachate in the two-step procedure, and reached a total release of 0.4 mg/kg for the WA system. This represents around 0.4% of the inventory. The chromium concentration was decreasing slightly with time, after the initial fast in-growth, in the single-step leaching for the WA system.
4. DISCUSSION
The increase of conductivity with time indicates a similar dissolution process for the two ashes. The high conductivity obtained already after 10 min indicates a rapid release of soluble salts. Notable is the fact that most of these salts were released during the initial 24 hours (above 90% of the conductivi'ty after 64 days in the single-step procedure), in fact, largely within the first 10 minutes (60 and 80% of the final conductivity for BA and WA, respectively). The pH-development of the WA system indicated a continued release of hydroxide which was not compensated by the generation of acids, including CO2 contribution from the atmosphere. In the BA system, however, a pH-maximum of 9.2 was obtained already within one hour. A pH decrease towards 8.5 and below in this system after 64 days reflects an inflow of CO2 from the atmosphere into the system and, probably, also a release of acidic organic material. There is an approach towards a CaCO3-dominated system at constant CO2-pressure, which would have an equilibrium pH of 8.3 8.4. The decrease in pH was more pronounced in the single-step leachate, where the readily released TOC was not removed as in the two-step procedure. The considerably higher release of TOC from the BA in comparison with the WA could be one of the reasons for the different pH-developments. The differences in other pH-controlling systems (Ca, Mg, as well as Na-K) indicate possibly a higher content of alkali hydroxides in the BA-systems, which otherwise would give a higher pH in the BA than in the WA leachates, disregarding the potential effects of organics. The low redox potential initially obtained in the B A leachate, particularly in the single-step procedure after 24 hours of exposure in a closed system, indicates the presence of elements in their reduced state in the solid BA. The lowest measured potential (-450 mV) corresponds to an E~ of around 0.1 V, which could be representative of an Fe(III)/Fe(II)-couple. A significant fraction of the iron in the BA would be Fe(II), while no reducing capacity is indicated for the WA systems (with considerably lower total content of iron than the B A). As a consequence, chromium would be expected to exist primarilly as Cr(III) in the BA, while a significant or dominant fraction of Cr(VI) can not be excluded in the WA. The reduction of the TOC-concentration in the single-step leaching after the initial fast release of carbon indicates either a loss of volatile organics or adsorption on solid surfaces or possibly a degradation to carbon dioxide. The presence of a large fraction of leachable volatile organics is not likely in high-temperature combustion residues. A substantial adsorption or binding of organic agents directly after a release during the leaching is not probable. A degradation through microbial processes seems to be a likely explanation to the rapid loss of TOC. This degradation would contribute to the pronounced pH-decrease observed in the single-step systems in contrast to the second step in the two-step procedure, where the initially released TOC-fraction is removed by change of leachate solution. It is evident that the organic carbon fraction is of different nature in the two materials, reflected by the differences in leachability and degradation rate. Only a minor fraction of the high organic content of the WA is leachable by water. The maximum copper concentration (280 mg/l, corresponding to 4.4x10 -6 M) is 1-2 orders of magnitude above expected total solubility, considering complexation and potentially solubility
713 limiting secondary solid phases (hydroxide and possibly hydroxy carbonates). The similarity in leaching behaviour and concentration change with time between copper and TOC is striking. The appearent over-saturation can be explained, assuming that around 1% of the TOC represents a strong complexing agent with a complexing capacity of 6-7 meq/g, which is not unreasonable. Similar enhanced releases related to the presence of organics have previously been claimed [7]. The reduction in appearent copper concentration with time in the single-step procedure can either be due to a decreasing solubility or a loss of copper due to adsorption. The decreasing pH (c.f. Figure 2) could actually lead to an enhanced adsorption of an organic complex. The loss of organic carbon assumed to be due to microbial degradation and a related reduced total solubility is, however, more likely as an explanation to the decreasing copper concentration. The leaching behaviour of chromium in the WA system has some similarity with the behaviour of copper in the BA system. The slightly decreasing concentration of chromium with time in the singlestep leaching could be due to interactions with organics analogous to the copper system, although less pronounced. The absence of significant chromium releases from the BA system, however, indicates a predominant dependence on the redox conditions (c.f. Figure 3) and the oxidation state of chromium in the solid matrix. The existence of chromium predominantly as Cr(III) of low solubility in the B A would lead to a slow leaching-rate and low over-all chromium concentrations. In the WA, which apparently has a minor reducing capacity, chromium may exist partly as Cr(VI), which would be more mobile and soluble as compared to Cr(III).
5. CONCLUSIONS An immidiate release of organic carbon compounds was observed by the exposure of the BA and WA to water. The readily released organic fraction was partly lost from the systems with time, possibly as the result of microbial degradation. The long-term leaching of the organic inventory of the ashes and subsequent TOC-degradation in solution should be further analysed. The potential metal solubilizing effects of organic matter released from the ashes by leaching has been demonstrated for copper. Further studies of the nature of these organic agents (composition, complexing properties and chemical stability) are required in order to allow an assessment of the over-all importance of the organic fraction for release and mobilization of metals. The possible existence of strong complexing agents that may significantly affect both leaching rates and solubilities of certain metals is of particular interest.
6. ACKNOWLEDGEMENTS Financial support from the Swedish Waste Research Council (AFR), as well as from the Swedish Association of Waste Management (RVF) is gratefully acknowledged. Mr. J. Rogbeck and Mr. L. Larsson, Swedish Geotechnical Institute, both participated in the early planning of this project.
7. R E F E R E N C E S
1 P.H. Brunner, M.D. Mueller, S.R. McDow and H. Moench. Total organic carbon emissions from municipal incinerators. WasteManagement & Research (1987) 355-365
714 2 B.S. Shane, C.B. Henry, J.H. Hotchkiss, K.A. Klausner, W H. Gutenmann and D.J. Lisk. Organic toxicants and mutagens in ashes from eighteen municipal refuse incinerators. Arch. Environ. Contain. Toxicol. 19 (1990) 665-673 3 H. Belevi, N. Agustoni-Phan and P. Baccini. Influence of organic carbon on the long-term behaviour of bottom ash monofills. In Proc. Forth International Landfill Symposium, Cagliari (1993) Environmental Sanitary Engineering Centre, Cagliari 4 A.-M. F~illman and J. Hartl6n. Leaching of slags and ashes - controlling factors in field experiments versus in laboratory tests. In J.J.J.M. Goumans, H.A. van der Sloot and T. Aalbers (eds), Environmental Aspects of Construction with Waste Materials, Elsevier, Amsterdam, (1994) 39-54 5 I. Pavasars, A.-M. Fallman, B. Allard and H. Bor6n. Work in progress 6 Standard Methods for the Examination of Water and Wastewater. American Public Health Association, American Water Works Association, Water Pollution Control Federation, Washington (1985) 7 R.N.J. Comans, H.A. van der Sloot and P.A. Bonouvrie. Geochemical reactions controlling the solubility of major trace elements during leaching of municipal solid waste incineration residues. In J Kilgroe (ed.), Municipal Waste Combustion Conference, Air and Waste Management Association, Pittsburg (1993) 667-679
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
715
Leaching Behavior of P C D D / F s and PCBs from Some Waste Materials
S. Sakai, S. Urano, H. T akatsuki Environment Preservation Center, Kyoto University, Yoshida Honmachi, Sakyo-ku, Kyoto-city, 606-01, Kyoto, Japan
Abstract Although it is well known that some waste materials and their incinerator residues contain persistent organic pollutants (POPs) such as PCDD/Fs and PCBs, little attention has been paid to the leaching behavior of these chemicals because of their low leachability. Due to the co-existence of surfactants in wastes, however, leaching concentration of POPs may increase. Therefore, leaching tests with and without those substances were conducted in order to understand the influence of surfactant-like substances on POPs leaching. In those tests, LAS (Linear Alkylbenzene Sulfonate) and humic acid was used as surfactant-like substances. Shredder residues from car/electrics recycling and fly ash from a municipal solid waste (MSW) incinerator were used in content analyses and leaching tests. Furthermore, an experiment was carried out to understand the influence of fine particles to the leaching concentration of POPs. The results of the leaching tests indicate that surfactant-like substances increase the leaching concentration of POPs, and fine particles related closely to the transporting behavior of POPs.
1. Introduction Although it is well known that some waste materials and their incinerator residues contain PCDD/Fs and PCBs, the leaching behaviors of these persistent organic pollutants (POPs) have not been studied well because they seldom solute with water due to their low solubility. It has been reported that the leaching behavior of POPs with co-existent substances, which are surfactant-like substances, solute salts, and so on, is not the same as their behavior without them ':). Clarifying the influences of these co-existences and gaining an understanding of the maximum leaching concentration, or so called availability of POPs, is therefore considered to be imPortant. Since surfactants increase solubilities of hydrophobic substances, it is thought that the leaching behavior of POPs will be affected by these chemicals at disposal sites. In this study, content analysis and leaching tests with LAS and humic acid solutions have been conducted in order to understand leaching behaviors and availabilities of PCDD/Fs and PCBs from automobile shredder residues and MSW incinerator residue. Furthermore, the influence of fine particles on POP leaching concentration has also been investigated. From these results, a leaching test procedure for POPs is discussed.
716
2. Experiments 2.1 Samples In this study, shredder residues and fly ash from a munid pal solid waste indnerator (MSWI) were used for content analysis and leaching tests. Shredder residues, the remains of valuables recovered from waste automobiles and discarded electronic goods, total more than one million tons per year in Japan. Most of them have been landfilled in non-controlled landfill sites where they stayed until March of 1996. About five hundred thousands tons of shredder residues have also landfilled illegally on Teshima Island in Kagawa Prefecture, some of which have been burned in the open. Shredder residues contain PCBs used in capacitors and some other materials until 1972 in Japan. It has been reported that openly burned shredder residues contain PCDD/Fs z~. Three types of shredder residues, two kinds from junked waste automobi le shredder residues and one discarded electronic goods shredder residue, were used in this study. Shredder residue K and Y, originated by automobiles, were sampled in 1988 and 1995, respectively. Shredder W, shredder residue from electronic goods, was sampled in 1995. Table 1 shows the results of physical fraction analysis. After classifying 1 kg of residue according to its physical fractions, the weight of each fraction was measured. Automobile shredder residues consist of high percentage of fine materials, mainly glass and others, and 90 percent of the electronic goods shredder residue was plastics. MSWI fly ash was a sample from a continuous type incinerator with an incineration capacity of 200 ton/clay, and equipped with a mechanical stoker and electric precipitator. Table 1 Physical Fractions of Shredder Residues Mass Percntage Fractions
Shredder residue K from automobile
Shredder residue Y from automobile
Shredder residue W from electronic goods
Hard Prastics
17.1
7.5
36.6
Soft Prastics
4.1
8.9
54.0
Rubber and Leather
11.6
10.2
0.0
Fine Materials (Soil, Glass, Wood, Paper, Texture)
61.6
51.6
0.01
Metals
5.6
21.8
9.4
2.2 Con tent Analysis From 1 kg of shredder residue, the weight of samples were reduced as uniformly as possible, and then PCDD/Fs and PCBs were measured. 20 g of MSWI fly ash was picked, and after one hour ultrasonic treatment in 2 m mol/l HCI, PCDD/F and Co-PCB congeners, PCB homologue were measured.
717
2.3 Leaching Test 2.3.1 Test Series Table 2 Solvent in Leaching Test Table 2 shows solvents in the leaching test. LAS and humic acid, leachate from an Solvent Concentration [mg/l] industrial waste landfill site, and distilled water Distilled Water as a control experiment were used in the 10 leaching test. LAS, Dodecylbenzene sulfonic LAS (DBS) acid sodium salt (DBS), one of the main 1000 ingredients used in a home-use synthetic 10 Humic Acid detergent, was also used in this study. Although 200 the biodegradability of LAS is relatively high, Leachate from Humic Acid 8.4 it has been detected widdy in the water Indutrial Waste environment. At landfill site, LAS may be LAS 0.33 Landfill Site contained in industrial waste and sewage sludge. On the other hand, humic acid, which is a very stable substance formed in nature by degradations of organic substances with high molecular weight, has been detected in landfill leachate in Japan ranging from tens to hundreds mg/1. Although humic acid doesn~ have a designated structural formation generally, it has many benzene rings and various function groups LAS : Tokyo Chemical Industry Co., Sodium Dodecylbenzenesulfonate (D1238), soft type, 65 % in water and humic acid : Tokyo Chemical Industry Co., Nitrohumic Acid (H0161 ) were used in this study. Experimental solvent conditions set up high and low concentrations in the leachant. The low one has a detectable concentration in the environment, and the high one is set up considering with Critical Micelle Concentration (CMC) . It is well known that if the micelle is formed, leaching concentration of hydrophobic substances increases surrounded by surfactants "~ 2.3.2 Leaching Test Method In the leaching tests for shredder residues, each residue was picked up according to its original mass percentages of fraction, and after cutting all samples under 10 mm in a diameter, each component was mixed for use in leaching tests. In order to prevent a sampling variation, twice the weight of each component was taken and cut it in half almost completely in a series of leaching tests. Furthermore, leaching tests were carried out three times, and after centrifuge of the leachates respectively, supernatants were mixed. Although the leaching test time for waste material is regulated to 6 hours in Japan, it was set up for 24 hours of horizontal shaking for the purpose of reaching equiliburium. In the leaching test for MSWI fly ash, 100 g of a sample, whose particle size was under 4 mm, was mixed with leachant on the condition of L/S (liquid per solute ratio)=l 0 with 200 mg/1 humic acid solution, with 24 hours of horizontal shaking. Separation of solid and liquid after shaking was carried out with a centrifugal separator in this study because a large amount of soil and fine particles obstructed filtration. Therefore,
718
after an obtained leachate was filtered with glasswool to remove large materials, the leachate was centrifuged at 700 G, 10 minutes. This conditi on was determined by the sedimentation velocity formula to remove over 0.45 ~m particles. Particle density was required for this calculation, and it was measured with a pycnometer.
2.3.3 Experiment of the Influence of Fine Particles This experiment was carried out in order to understand the influence of fine particles on POPs leaching concentration. Fine particles are suspended particles, not those sedimented with gravity. In this experiment, a leachate obtained from another leaching test was divided into two fractions. Each half of the leachate was centrifuged and filtered, respectively and then PCDD/Fs and Co-PCBs congeners, PCBs homologue, and particle distribution were measured. From these results, the influence of fine particles on POPs leaching concentration was studied. This leaching test was carried out by shredder residue Y under conditions of I/S= 10 with 500 mg/1 LAS solution. 2.4 Analytical Method 2.4.1 PCDD/Fs and PCBs PCDD/Fs and PCBs were measured by HRGC/HRMS. Clean-up procedure was followed according to the Analytical Manual of PCDDs/PCDFs in Waste Management in Japan '~. GC columns were Spelco SP-2331 for low-chlorinated PCDD/Fs (from tetra to hexa) and J & S Science Co. DB-5 for PCBs and high-chlorinated PCDD/Fs (hepta and octa). 2.4.2 Particle Size Distribution Measurement Particle size distributions were measured using a Shimazu Particle Distribution Meter with a Centrifugal Separator (SA-CP3). Since the leachate contained LAS, it was washed three times with distilled water, and measured with the Particle Distribution Meter on after the addition of a dispersion reagent. This meter has a variable velocity centrifugal separator, which can measure particle concentration based on absorbances.
3. Results and Discussion 3.1 Con tent Analysis Table 3 shows the contents of PCB homologue in shredder residues. PCBs concentration in shredder residue K and Y ranged from 1,800 to 11,000 ng/g and 15,000 to 24,000 ng/g, respectively. Even the same shredder residue contained a rather varied PCBs content due to a difference in fractions. On the other hand, PCBs concentration in the electronics shredder residue was 1,200 ng/g, which shows low concentration in comparison with automobile shredder residues. As for PCDD/Fs in shredder residues, those were all under quantity limit, though the detectable limit is 0.1 ng/g, due to oily chemicals. On the other hand, PCDD/Fs were detected in the previous leaching tests on the LAS 1,000 mg/l series, which shows shredder
residues contain PCDD/Fs, though these were not detected in the content analysis because high detectable limit. Our recent experiments on a different shredder residue, however, indicate that it contained 0.25 ng-TEQ/g of PCDD/Fs. Table 3 Results of PCBs Contents Analysis for Shredder Residues [ng/g] Shredder Residues S.R. K Homologue
first
S.R. Y second
first
S.R. W second
M~CBs
2.1
9.6
19
27
1.3
~CBs
200
800
1300
1600
80
T3CBs
720
4200
4600
6800
320
T4CBs
450
3100
3700
5300
310
PsCBs
280
1600
3400
6400
340
H6CBs
140
770
1600
3200
170
H~CBs
21
120
280
600
25
O~CBs
2.1
28
70
1.8
N9CBs
0.2
1.1
3.1
6.0
0.2
D,~---'B Total CBs
0.3 1800
N.D. < 0.1 11000
N.D. < 0.1 15000
N.D.<0.1 24000
N.D.<0.1 1200
3.1.2 Municipal Solid Waste Incinerator Fly Ash Table 4 shows the result of PCDD/Fs content analysis in MSWI fly ash. The actual total concentration was 780 ng/g and the toxicity equivalent concentration was 10.3 ng-TEQ/g. As for PCBs, the total PCBs concentration was 26 ng/g and actual Co-PCBs concentration was 3.3 ng/g, with toxicity equivalent concentration being 0.053 ng-TEQ/g. This fact indicates that MSWI fly ash contains high PCDD/Fs and low PCBs in comparison to shredder residues. 3.2 Leaching Test 3.2.1 Shredder Residues Table 5 shows the results of PCBs leaching concentration on shredder residue K. As to the series using humic acid 10 mg/1 for leachant, PCBs leaching concentration was almost the same as when using distilled water. However, another series demonstrated an obvious increase in leaching concentration. Especially on the LAS 1,000 mg/l series, this tendency was demonstrated clearly. Compared to the results with distilled water, PCBs leaching concentration using LAS 1,000 mg/1 solution increased about 70 times. This result indicates that a sulfactant will strongly affect the PCBs leaching concentration. On the other hand, humic acid also increased the PCBs leaching concentration. In the case of 200 mg/1 humic acid solution, the concentration of PCBs increased about 4.6 times higher compared to the results from using distilled water. Although the leachate from the industrial waste disposal
Table 4 PCDD/Fs in MSWI Fly Ash and Leachate Fly Ash
l
Leachate
Toxicity Toxicity Actual Actual Concentration Equivalent Concentration Equivalent Concentration Concentration [ng/gl [ng/ll [ng-TEQ/g] [ng-TEQ/I]
2,3,7,8-T4CDD
0.56
T4CDDs
10
1,2,3,7,8- P~CDD
2.3
PsCDDs,
0.56
2.4 1.15
26
1,2,3,4,7,8-H6CDD
N.D.<0.05
,
0
N.D.<0.05 N.D.< 0.05
3.3
0.33
N.D.<0.05
1,2,3,6,7,8-H6CDD
6.5
0.65
N.D.<0.05
1,2,3,7,8,9-H6CDD
5.3
0.53
N.D.<0.05
0
i
I-kCDDs
71
0 i
0
N.D.<0.05
i
1,2,3,4,6,7,8-H~CDD
85
0.85
0.06
0.0006
!
~CDDs
150
0.13
!
O, CDD
210
0.21
0.40
0.00040
Total PCDDs
470
4.28
2.9
0.0010
2, 3,4 ,7 , 8- P.sCDF
4.3
2.15
N.D.<0.05
0
P~CDFs
58
I
N.D.<0.05
I
1,2,3,4,7,8-H6CDF
7.9
0.79
N.D.<0.05
0
1,2,3,6,7,8-H&~DF
8.2
0.82
N.D.<0.05
0
1,2,3,7,8,9-H6CDF
12
1.20
N.D.<0.05
2,3,4,6,7,8-H6CDF
1.0
0.10
N.D.<0.05
I-kCDFs ~
78
!
1,2,3,4,6,7,8-HTCDF
!
N.D.<0.05
0
44
0.44
0.09
0.0009
11
0.11
N.D.<0.05
0
0.05
0.16
I
1,2,3,4,7,8,9-HTCDF ~CDFs
85
0.09
~'~
!
OsCDF
52 !
!
0.00016 i
Total CDFs
310
6.02
0.25
0.0011
Total PCDDs and PCDFs
780
10.30
3.2
0.021
PCDD/Fs Leaching quantity ratio [%] Co-PCBs
3.3
Co-PCBs Leaching quantity ratio [%] PCBs
J
26
PCBs Leaching quantity ratio [%]
0.053
0. 0041
0.0020
0.91
0.00024
0.28
0.0045
37 1.4
site contained few PCBs, the result indicates that PCBs leaching concentration will ,. increased by this substance in leachate. Since humic acid concentration in this leachate was 8.4 mg/1 and LAS concentration was 0.33 mg/l as shown in Table 2, it is considered that these surfactants affect PCBs leaching concentration. From this result, the leaching quantity of PCBs was only a low percentage contained in shredder waste. Compared with the release ratios, the leaching quantity of PCBs from shredder residues K was 0.006 % at maximum content with distilled water and 2.6 % at minimum content with LAS 1,000 mg/1. Therefore, it is necessary to understand the maximum leaching quantity which is so called the availability in leaching test for heavy metals. Table 5 Result of Leaching Test on PCBs for Shredder Residue K [ng/1] Solvents Homologue
Distilled Water
Leachate from Industrial landfill Site
MICBs
0.12
D2CBs
8.8
LAS
Humic Acid
10 mg/1
200 mg/l
10 mg/1
1000 mg/l
0.14
N.D.<0.1
N.D.<0.1
0.15
5.1
16
5.7
18
12
360
74
1500
57
900
28
990
28
650
!
31
110
28
27
91
15
7.5
51
T3CBs
26
52
T4CBs
16
i i
P~CBs
8.1
|
H6CBs
6.9
12
6.2
38
HTCBs
2.3
2.5
1.2
7.4
9.9
130
O~CBs
0.15
0.24
0.20
1.4
0.75
23
NgCBs
N.D.<0.1
N.D.<0.1
N.D.<0.1
0.22
0.10
2.0
~oCB
N.D. < 0.1
N.D. < 0.1
N.D. < 0.1 N.D.<0.1
Total CBs
68
130
79
310
~
0.10
0.52
210
4600
3.2.2 Municipal Solid Waste Incinerator Fly Ash Table 4 also shows the results of the leaching test for MSWI fly ash with 200 mg/l humic acid solution. Although the leaching quantity equaled only a small amount (0.002%) of its content, PCDD/Fs obviously leached . In general, high-chlorinated PCDD/Fs which are contained in high concentrations in the fly ash, leached conspicuously in addition to T4CDDs. The leaching quantity of PCDD/Fs was only 0.002 % of the content in the fly ash. On the contrary, Schramm has reported that PCDD/Fs leached out about 31 ng/1 with LAS 1,500 mg/1 leachant 6). In this case, the leaching quantity ratio will be 2.5 % of its contents. These results indicate that the leaching concentration of PCDD/Fs will be different from that of the leachant, PCDD/Fs contents and others. It is therefore necessary to understand availability through something like the serial batch test using surfactants.
..3 Fine Particle and POPs Leaching Concentration It has been reported that a fluctuation was brought about in the leaching test for POPs 7). These organic substances have a strong adsorptive tendency, that is POPs may be adsorbed easily on fine particles or colloids Because of this adsorption, POPs leaching concentration will be changed by filtration. Therefore, each half of the leach_ate was centrifuged and filtered, removing over 0.45 lam particles in order to understand the relation between the distribution of particles and PCBs leaching concentration. Table 6 PCBs Concentration and Table 6 shows a result of this experiment for Separation Method shredder residue Y. The result demonstrated PCBs Concentration [ng/1] that PCBs leaching concentrations between Homologue Centrifuge Filtration filtering and centrifuging were significantly M~CBs 14 2.6 different. PCBs leaching by filtering was only 30 % that of centrifuging. ~CBs 1000 240 Particle distributions of shredder residue T3CBs 4900 1600 leachates are shown in Figure 1 and 2. Only T4CBs 5400 1600 particles under 10 l~m are shown in the Figure. The particle distributions are P.~Bs 5000 1500 different from each other. Although the ~CBs 2100 610 untreated sample had a peak of particles ~CBs 500 140 from 6 to 10 ~m, the centrifuged one is O, CBs 94 27 mostly occupied under 0.5 ~tm particles. The particles around 2 lam maybe due to N9CBs 9.6 2.6 reaggregation. On the other side, the filtered D,~-'B N.D. < 0.05 N.D. < 0.05 sample was out of measurement range because of low absorbance, but the particle distribution in the filtered sample was supposed to be under 0.08 ~tm. These results demonstrate PCBs leaching concentration will be strongly affected by fine particles. Therefore, it is necessary to pay attention to the behavior of the particles, when leaching behavior is to be taken into account. G.A.Rood et al. ~ have reported that the results of leaching tests for POPs centrifuged at 3,470G, 90 minutes were suitable for the filtered sample with a 0.45 lam membrane filter, though the concentration of the centrifuged sample was from 2 to 25 times higher compared to the results for the filtrate. Our experiment indicates that almost all fine particles, on which PCBs adsorbed, were removed by filtration with a 0.45 ~m membrane filter. This is an extremely strong centrifugal force and a long time is required to correspond with the results of the filtered sample. Which size should be removed is an open discussion for the next step, and the fact that the leaching concentration of POPs is affected by fine particles should also be taken into account at that time.
723
,--, E
10.00 8.00 6.00 5.00 4.00 3.00 2.00 1.50 1.00 0.80 0.60 0.50 0.30 0.20 0.15 - - - - 0.10 0.00 0 .
.
.
10.00 8.00 .
.
.
|
..................................................................
= "~ O
"=~ -~ 9 "6 "~
. . . . . . . . . . . . . . . . . . . . .
.........................................
3100 2.00 1.50
I
|
1.00
|
"-'---'1
..........
-~ a; ~ !
50 100 I%] Figure 1 Particle Distribution in Untreated Leachate
Since it is expected that the higher chlorinated PCBs have a stronger adsorptive force, homologue distributions "E of centrifuged and filtered samples are & shown in Figure 3. Homologue compositions for PCBs were almost the same for both samples. This demonstrates _o that homologue compositions do not ~ O
0.40
8: 8 o. 15
................................ ,......................................................................
0.10 0.08 0.00
iiiii
0
!
!
ii
100
50
Figure 2 Particle D i s t ~ u t i o n in Centrifuge Leachate 1 0.8
0.6
O1)
O
depend on separation methods , although ~ rj they do affect leaching concentration. As for the result of Co-PCBs, the concentrations in centrifuged and filtered samples were 0.44 ng-TEQ/1 and 0.13 ngTEQ/1 respectively, a ratio of concentrations between centrifuged and filtered was almost the same with a ratio of homologue concentrations.
0.4 0.2 0 centrifuged
filtered
[7
M1CB
I'] D2CB
II
!i
T4CB
i~ P5CB
E~] H6CB
I--ITCB
II
~
O8CB
T3CB
N9CB
!"! D1003 4. C o n d u s i o n s The following conclusions were derived from this study :
Figure 3 PCBs Homologue Composition in Cntrifuged and Filtered Leachate
1) PCBs contents in shredder residues used in this study range from 1.8 ~g/g to 241xg/g, and it is indicated that the difference in PCB s concentrations is large among the same kinds of shredder residues. Even though the same residue from a shredding plant were used, there are great differences depending on the original waste materials.
724
2) Leaching concentrations of PCDD/Fs and PCBs from shredder residues and MSWI fly ash are strongly affected by suffactants. In the results of PCBs for shredder residues, it was indicated that PCBs concentration when using LAS 1000 mg/l leachant was about 70 times higher than that of distilled water. It was also indicated that the concentration with humic acid 200 mg/1 solution was about 4.6 times higher compared to the results from distilled water. PCDD/Fs were detected from MSWI fly ash with humic acid 200 mg/l leachant and automobile shredder residue with LAS 1,000 mg/l leachant. 3) Compared with the release ratios, the leaching quantity of PCBs from shredder residues was 0.006% (at maximum contents) with distilled water and 2.4 % (at minimum contents) with LAS 1,000 mg/1. Therefore, it is necessary to understand the maximum leaching quantity, availability, for various waste materials. 4) The results found between the centrifuged sample and the filtered one operated on the condition of removing the particles under 0.45 ~m, and were quite varied. The result of the particle distribution indicated that the centrifuged leachate usually contained the particles under 0.45 lam against the filtered leachate, probably containing particles under 0.08 tam. The difference in particle distributions would cause a fluctuation in the leaching concentration for POPs. Compared to the PCB homologue and Co-PCB congeners composition in the samples, however, the compositions hardly changed at all. References
1) K.-W. Schramm, W. Z. Wu, B. Henkelmann, M. Merk, Y. Xu, Y. Y. Zhang and A. Kettrup ( 1995): Influence of Linear Alkylbenzene Sulfonate (LAS) as Organic Cosolvent on Leaching Behavior of PCDD/Fs from Fly Ash and Soil, Chemosphere, 31 (6) , pp. 3445-3453 2) S. Sakai, S. Urano, H. Takatsuki, K. Shiozaki, K. Gokita (1996) : Influences of Humic Acid and Linear Alkylbenzene Sulfonate (LAS) on Leaching Behavior of PCDDs/PCDFs and PCBs from Shredder Residues, Organohalogen Compounds, 28, pp. 11-15 3) M. Hanashima, H. Takatsuki, O. Nakasugi (1996): A Case Study of Environmental Contamination Caused by Illegal Dumping of Hazardous Waste, Waste Managonent Research, 7(3), pp. 208-219 [in Japanese] 4) N. Shinozuka, Chang Lee (1982): Surface Active Properties of Marine Humic Acids, Marine Chemistry, 33, pp. 229-241 5) Japan Waste Research Foundation ( 1991) : Analytical Manual of PCDDs/PCDFs in Waste Management [in Japanese] 6) K.-W. Schramm, M. Merk, B. Henkelmann and A. Kettrup (1995) : Leaching of PCDD/F from Fly Ash and Soil with Fire-extinguishing Water, Chemosphere, 30 (12), pp. 22492257 7) G. A. Rood, M. H. Broekman, Th. G. Aalbers (1994) : Investigating a leaching test for PCBs and organochlorine pesticides in waste and building materials,
Environmental
Aspects of Construction with Waste Materials, Elsevier Science B.V., pp. 271-280
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
725
E N V I R O N M E N T A L Q U A L I T Y ASSURANCE SYSTEM F O R USE OF C R USHE D M I N E R A L D E M O L I T I O N WASTES IN E A R T H C O N S T R U C T I O N S Wahlstr6m, M. 1), Laine-Ylijoki, J.~, M/i/itt/inen, A. 2), Luotoj~irvi, T. 3) & Kivek~is, L. 3) I~VTT Chemical Technology, Environmental Technology, P.O. Box 1403, FIN 02044 VTT, Finland
2) Lohja Rudus Environmental Technology Ltd, M/ikirinteentie 19, FIN-36220 Kangasala, Finland 3)Lohja Rudus Environmental Technology Ltd, Pronssitie 1, FIN 00440 Helsinki, Finland Abstract
Annually 1 million tons of mineral demolition wastes mainly consisting of concrete and bricks, is produced in Finland. The crushed materials have in field studies on test roads showed favourable geotechnical properties for use in road constructions. The test samples from two crushing plants were chemically characterised and the leaching behaviour was studied by using column, two-stage batch leaching and pH static tests. Only sulphate and chromium leaching from the crushed material was detected. There was a good agreement between column and batch leaching tests. The contents of harmful organic compounds were very low. Based on experience and the results of the experimental study, a practical sampling and testing strategy for an environmental quality assessment system was developed. A two-stage batch leaching test was chosen for the quality control of demolition waste. Preliminary target values for leaching of sulphate, Cr, Cd, Cu and Pb were set. Both geotechnical and environmental properties of the crushed material indicate that the use of demolition waste in road constructions is acceptable and can be recommended to replace landfilling of this material. However, a detailed demolition plan is most important in order to have an acceptable material for utilisation in earth constructions.
1. I n t r o d u c t i o n
Annually 1 million tons of mineral demolition wastes consisting mainly of concrete and bricks, is produced in Finland. According to current practice, most of mineral demolition waste is landfilled. A proper demolition plan is most important in order to have an acceptable material for utilisation in earth constructions. Selective demolition of buildings and constructions will in the future improve the quality of waste materials. Also due to introduction of taxes on landfilling of wastes, utilisation of mineral demolition wastes will be encouraged. Crushed demolition concrete can be recycled as an aggregate in concrete or used in road bases, the latter of which is of particular interest in Finland. Contents of harmful compounds must be low and the leaching behaviour of trace metals must be acceptable in materials to be utilised in earth constructions. Potentially harmful materials should be identified and removed before demolition. Examples of potentially harmful materials are preserved wood, electronic equipment, plastics, adhesives, sealing materials, asbestos, fluorescent lamps, gypsum materials, and tarry material. A list of harmful compounds, which may be present in materials or equipment used in the construction industry, is presented in Table 1. The list does not include asbestos, because due to Finnish regulations asbestos always needs to be removed from the buildings before demolition. The content of harmful materials in building wastes has been estimated in a Swedish report (Sigfrid 1993) to be about 1 % of total mass. The age and the use of the building to be demolished significantly influence the amount of harmful materials. Cadmium for example was used in PVCplastics as stabiliser and colour pigment during the 1960"s and 1970"s. The amount of cadmium in plastised PVC was 0,03-0,06 weight-% and in hard PVC 0,012-0,36 weight-%. Another example is the use of sealing compounds for wall elements containing polysulphides. The PCB-contents in these compounds were at least 20 %.
726
Table 1. Harmful compounds in mineral demolition wastes.
Harmful compound Organic bound cadmium Metallic cadmium Metallic mercury Metallic lead Organic lead PCB-compounds CFC-compounds Oil and PAH Chromium (molybdenum) Copper Phenolic compounds
Application Stabilisers in plastics Surface finishing agents Indicators and switches in electrical installations Sealing in sewer drains Additive in plastics Sealing compounds, condensers and anti-slip floor covering Freezing agents and polymeric insulating materials Oil spills, e.g. from machines, in felt roof, tarry compounds Concrete, mortar Wires, copper water pipes Insulating materials, adhesives
2. Materials and methods 2.1 Materials During 1995 the environmental impacts of two different types of demolition wastes were studied in a Finnish project. One of the demolition materials was hollow core slab waste, which had been abandoned due to manufacturing process requirements. In the actual production of hollow core slabs, moulds used are treated with a mixture containing mineral and plant oil before moulding. The oil content in the crushed hollow core slab was estimated from the oil consumption to be a maximum of 110 mg/kg. The other demolition material was mineral building demolition waste, mainly concrete and bricks, from several domestic buildings. In both cases the material of about 10.000 t was crushed to a grain size of less than 70 mm. In these cases the test samples obtained as the crushed material was falling from a conveyor belt to a storage heap. Three test samples about 70 kg each were taken with about one week interval. The grain size distribution of the samples is shown in Figure I. The test samples were milled and divided by a riffle box to representative laboratory samples of grain sizes of 4 and 20 mm.
Grain size distribution Grain size distribution Building demolition waste Hollow core slab % passing through % passing through lOO 1O0 ..~ L Sample 1 I Sample 11i 90 IZ' [, 8O 80 / '!" Sample 2 Sample 2 I. .... --[, 70 /,,,:~ Sample 3 Sample 3 I 60 60 i - '-/"7........... / 79 SO ,',~50 ...... ./ 40 ~;;-40 -", 30 30 /' 20 20 ....~s / < -f'-....;../"" ..10 . . . . ~.~......7 .. ...... lO -~"~h----~ ...., ,, ...... , ,, ..... 0 , ~ ; ~ l , ...., , ...... , , t .... 0 0,030,1 0,3 1 3 10 30 100 0,030,1 0,3 1 3 10 30 100 Particle size (mm) Particle size (mm)
9o_
i
I
Fig.1. Grain size distribution in test samples (the coarse fraction > 32 mm was excluded from the test samples).
727
The chemical analysis of the laboratory samples were performed as follows: content of metals was measured after aqua regia digestion by ICP- or AAS-technique. phenol-index was determined using 4-aminoantipyrine spectrometric method after distillation PAH-compounds were determined by SIM-technique with a gas chromatograph (GC) equipped with a mass selective detector. Three internal standards (d-pyrene, ]]]]-binapthyl- and indeno(1,2,3-cd)fluoranthene) were added. The sample extracts were cleaned with Silica| before analysis. PCB-compounds were determined with gaschromotograph techniques (GC-ECD) from the hexane extract. 2,4,6-tribromibiphenyl was used as an internal standard. The results from the chemical analysis is presented in Table 2. As comparison also values from literature are shown. During the summer 1996 additional quality controls of three other crushed demolition material were performed. The materials tested wastes were crushed railway sleeper waste, crushed bricks and crushed building demolition waste. These materials where not characterised chemically.
Table 2. Chemical composition of two demolition wastes and reference values from literature (all concentrations except loss of ignition in mg/kg). Building demolition waste
A1 Cd Cr Cu Hg Mo Pb Zn Phenol-index PAH PCB Sample 1-3 Sample 1 Sample 3 Loss of ignition
Hollow core slab waste
12900 2,6 20 15 <0,02 <2 11 59 <0,2 0,042
14400 3,3 29 18 <0,02 <2 10 44 <0,2 0,96
0,066 0,085 0,020 2%
<0,015 <0,015 <0,015 2%
Concrete granulates from d e m o l i t i o n w a s t e
(CROW 1994) Mean Range value 25300 0,3 0,13-1 104 30-160 15/21 7-20 0,10 0,10-0,10 2/3 25 11-50 70 51-100
Crushed concrete
(K~ilvesten 1996) Sample A
2,75 + 0,14 87,9+ 9,0 9,82 + 0,37 <0,0433 <6,18 6,36 + 0,19 52,3+ 0,7
3,8 %
Sample B
0,137+0,002 55,5+5,0 7,63+0,15 <0,0438 <6,18 5,91 +0,03 41,8+1,1
3,3 %
2.2 Leaching tests
The leaching behaviour of the hollow core slab waste, demolition building waste and also railway sleeper waste were evaluated by column tests. To study the variation of leaching behaviour and correlation between column and batch leaching test, several batch leaching tests were performed. The batch leaching test is more suitable for quality control than column test. The results were interpreted on the basis of liquid (L) to solid (S) ratio achieved in the tests, in which assumptions about the percolation rate at the site and the geometrical dimensions of the construction were made. The column test to L/S 10 usually depicts the leaching behaviour on the site for hundreds of years.
728
Column tests were conducted with crushed materials (< 4 mm) according to the Nordtest method ENVIR 002 (similar to the Dutch standard NEN 7343) to L/S 10. The water flow rate was kept as low as possible to L/S 2, after which the flow rate was speeded up. Distilled water acidified to pH 4 with nitric acid was used as leachant. Leachant is pumped from the bottom up to the top of the column, where the eluate fractions are collected based on the L/S ratio. Batch leaching tests (CEN-test) were performed according to prEN 12457 with two leaching steps. During the test procedure the sample was mixed for 6 hours with water at L/S 2 and for 18 hours subsequently at L/S 8. The eluates collected from the column tests and batch leaching tests are analysed. The pH environment may change with time, e.g. due to carbonation, pH static tests were carried out to characterise the pH-dependent leaching behaviour for selected metals. The tests were performed with crushed material (< 4 mm) at L/S 5 under pH-controlled conditions using automated pH control equipment with nitric acid addition. The tests were carried out for 24 hours at pH-values from 4 to 12. Leaching is usually dictated by diffusion in road constructions which are isolated with pavements. Therefore diffusion tests provide a more realistic information of the leaching in isolated road constructions. For diffusion tests performed according to the Dutch standard NEN 7345 demolition wastes were mixed with 2 % cement. A test specimen was immersed in water for 64 days and at certain time intervals the leachate was renewed and analysed.
3. Technical properties Lohja Rudus Environmental Technology Ltd has developed a processing system which enables the reuse of demolition waste in road construction. The concept is based on the self-hardening properties of crushed materials which contain unreacted cementious compounds. The crushed demolition waste, Betoroc, has a 2-3 times better bearing capacity than gravel or crushed rock allowing the reduction of the road base thickness. However, the improved bearing capacity cannot be fully utilised in the Finnish climate because of the frost periods. Crushed demolition wastes have been used in about ten road constructions since 1994. The bearing capacities measured have been in accordance with planned values or have exceeded them. The Emodulus of crushed demolition waste have varied between 600-1500 MN/m 2, whereas the corresponding reference value for gravel is 280 MN/m 2. The hardening properties of the crushed demolition waste is checked by preliminary tests. Test specimen are prepared using an ICT Gyratory compator, which simulates the action of a roller generating vertical and horizontal shear deformation in the material, after which the E-modulus of the crushed concrete waste is estimated. Also the optimal water content and maximum dry density are measured for construction design. The handling techniques, transportation, spreading and levelling, resemble those used for gravel. However, the compaction is more demanding, as at least five compactions with a vibrating roller of 10 tons are needed. The water demand is clearly higher, the optimum water content is in the range of 810 %. The crushed concrete needs also to be watered one month after construction or until paving. The design of a construction with railway sleepers and hollow core slabs is shown in Fig 2. The bearing capacity of the construction was in accordance with planned capacity. The E-modulus of the crushed demolition waste was 1000 MN/m 2, while the optimum water content was 9 % and the maximum dry density was 1950 kg/m 3. The compression strength of test specimen after 28 days of hardening was 2 MN/m 2. On sites without risks for frost periods the construction layer could have been reduced with 20 cm compared to the alternative construction with gravel. The total cost of the construction was identical to the ordinary construction with gravel, but the technical properties, especially the bearing capacity, were much better.
729
Betoroc construction ~
0,1mI
S
o
~
;3>
I"-'~ .~.z. v
I ~~~,
f (!
t ~
asphalt concrete U[
...... ,........
~.~@1Crushed concrete waste
(E = 2500 MPa)
Base course + Sub-base:
Base course: G r a v e l 0 - 65 mm (E = 280 M P a )
0,25
~71 0-50 mm
,.t . L.'
with gravel
Soft asphalt concrete
0,1
I (E = 2500 M P a )
L--j Z ....~l
0,35 ~ ~ ~ ' , ,. ,,
Alternative construction
(E = 1000 M P a )
z/.:.~(.-~ ~...~. )..~.:-'.c.~..~Sj Sub-base: ,..j ~-.:_., .... ~ ~--,~ ~a "" G r a v e l 0 - 100 mm 0,3 ml (-] %-)d.];~'-.,6T-:, C~ (E = 280 MPa) 0,55
F i l t e r sand (E = 50 M P a )
0,35 ~
~
Filter sand (E = 50 M e a )
30 MPa
T ' 30 MPa
Fig 2. Design of a construction with crushed railway sleepers and hollow core slabs compared to al alternative construction with gravel.
4. R e s u l t s 4.1 B a t c h l e a c h i n g t e s t a n d c o l u m n test
The pH-values of the eluates collected in column tests and CEN batch leaching test were all alkalii usually the pH-values were near 12,5. The high pH-value of the eluates indicate that the pH-valu were mostly dictated by the leaching of CaO which in water solutions gives pH-value of 12 Concrete contains free lime which is liberated during crushing. Furthermore the eluates from t batch leaching tests with coarse materials (< 20 mm) resulted in high pH-values. The leached amounts from the demolition wastes were low, only sulphate and chromium were leach (Tables 3 and 4). The leaching of sulphate was much higher from building demolition waste than fro hollow core slab waste and railway sleeper waste. The leaching of sulphate from the buildi demolition waste was significantly high in the first eluates of the column test. This indicates a was off effect of sulphate. As expected leaching of metals from crushed hollow core slab waste was low than from building demolition waste, because the origin of the hollow core slab waste. The colur test and CEN-batch leaching test results were comparable. In this case the maximum grain size did r influence the leaching behaviour, probably due to small difference in pH-values of the eluates and d to a very low leaching generally. Both results were also comparable to refence values (see Table 5). cumulative leaching, mg/kg
3 000 1 000 ~ 300 100/
Sulphate (S04)
-
cumulative leaching, mg/kg
1I ~
u,o
0,1 . . . . . . . . "~
0,1 0,2
0,5 1 2 L/S, I/kg
, i
5
, i
10
Demol.build,waste Column --I-- test Demol.CEN build, waste test
~
--e--
0,03 ~ - ~i ~..~..~-~"_~..~ "___~ ;
0,01 /-~ 0,003
1
C h r o m i u m (Cr)
~
0,001 0,0003
"
,* ;:i;! ~:~Ji
.-" "
~ ... ..
l
0,1 0,2
. *
. . . .
0,5 1 2 US, Ilkg
5
Hollowcoreslab Column --4F-- test Hollowcore slab CENtest Railwaysleeoer C01umn test Railwaysleeoer CEI~.-test
10
Fig. 3. Leaching of sulphate and chromium from three different types of demolition wastes.
Table 3. Accumulated leaching of metals and sulphate in batch leaching tests and column tests at L/S 2 and L/S 10 (concentrations in mg/kg).
Table 4. Accun~ulatedleaching of metals and szrlphate in batch leaching rests at L/S 10 (concentuafions in mg/kg).
731
Table 5. Accumulated leached amounts from crushed concrete and building demolition wastes presented in literature (concentrations in mg/kg) Crushed concrete (K/ilvesten 1996)
Demolition waste from a building
(Johansson et al 1996)
Concrete granulate from demolition
(CROW 1994) Column test
Column CEN-batch leaching test (2 steps) CEN-batch test leaching L/S 2 L/S 2 L/S 10 L/S 2 L/S 10 L/S 10 Sample Sample Sample Sample Sample A A B A B Cd <0,0002 <0,0002 <0,0002 <0,001 <0,001 0,01 0,01 0,002 Cr 0,046 0,0269 0,0294 0,148 0,157 0,16 1,71 0,11 Cu 0,0084 <0,002 0,004 <0,01 0,0121 0,07 0,12 <0,12 Ni <0,07 <0,04 <0,04 0,02 0,06 <0,1 <0,02 0,36 Pb <0,002 0,003 0,002 0,011 0,010 O,04 0,05 <0,04 Zn 0,043 0,01 0,01 0,31 0,31 0,01 0,05 <0,5 1) pilot-scaleresearch carried out in test bins which were placed outside under normal weather conditions
Concrete destruction debris
(Mulder 1991) Lysimeter ~) L/S 5
0,039
0,16
4.2 Diffusion test
The diffusion test results are shown in Table 6. The eluates were alkaline due to the calcium compounds. The leaching of sulphate and metals were low. The leached amounts were comparable to other cement solidified products reported in literature.
Table 6. Leaching of test specimen prepared from demolition waste containing 2 % cement in the diffusion test (cumulative leached amounts expressed as mg/m 2/64 d).
8042C1Cr Cu A1
Specimen from building demolition waste
Specimen from hollow core slab waste
Concrete road stone
1700 58 0,8 0,6 450
490 32 1,0 0,08 500
2085 <17 <3 <1,5
(CROW 1994)
4.3 p H static tests
The influence of pH changes on the leachability of metals and the buffer capacity of the test materials were studied using pH static test. The results for hollow core slab waste are presented in Fig. 4. The nitric acid consumption to achieve pH 10 was 0,6 mol/kg. The leaching of cadmium and lead was very low at all studied pH-values. Also the leaching of chromium from building demolition waste was very low. In the case of hollow core slab the decrease of pH influenced the increasing leaching of sulphate, copper, chromium and lead. The leaching of metals and sulphate was lowest at pH-values near 12. The leaching of metals was at no studied pHvalue over 1 mg/kg.
732
pH Static Test Hollow Core Slab
10000 ....................................................................... 1000 t~ ~-
~"
,~.
a~'"'"~ _ ~ ~
100 10
0,I
O,Ol
"O'-9 Cadmium (Cr) Chromium (Cr) Copper (Cu) X Lead (Pb) ,W.. Sulphate (SO4)
-
- ~ , , ~-~'~-:--2- - ~ ,v,~-
v
v
v
0,001 pH
Fig. 4. Leaching from hollow core slab waste in pH-static tests.
5. E V A L U A T I O N O F T H E E N V I R O N M E N T A L I M P A C T S O F D E M O L I T I O N W A S T E S The results from the leaching tests can be compared to 9 airborne emissions in wet depositions (e.g. sulphate) 9 reference materials normally accepted as construction materials 9 trigger values for estimation of contaminated soil (especially the content of organic compounds) 9 background concentrations in different environmental surroundings (e.g. risk assessment based on the calculation of metal transport by water in typical utilisation sites). The correlation between column tests and batch leaching test was very good. The leached amounts from the demolition wastes were low, only sulphate and chromium were leached. The leaching of metals from the demolition wastes studied was most cases in the range of variations observed for noncontaminated concrete (see Tables 4- 6). The influence of pH change on the leachability was estimated as low. Only in the case of hollow core slab waste an increase of leaching of chromium was observed at pH 10. However, the neutralisation capacity of the mineral demolition wastes is high. The environmental impacts of leaching of metals is estimated to be very small for the studied materials especially in isolated road constructions. The airborne emissions of sulphate in Finnish wet depositions varied in the year 1994 on the range of 146-539 mg/m 2. Assuming that the water flow through an isolated road construction of demolition waste from buildings is 10 mm/year and the thickness of a layer would be around 0,5 m and the density of 2 t/m 3 , it can be roughly calculated that LIS 0,1 might be achieved in about hundred years. According to column test results the total accumulated release of sulphate from building demolition waste after the first 100 years is around 0,3 kg/m 2. The assumed release of sulphate is of course higher than the contribution to soil from wet deposition, but the leaching of sulphate is not regarded as significant. The results from column tests can also be compared to concentrations in natural water. According to diffusion test results the leaching of sulphate was lower than from concrete road stone. The content of PAH and PCB analysed from the demolition wastes can be compared to Finnish trigger values for evaluation of contaminated soil (Ministry of the Environment 1994):
no restriction for use, (mg/kg) only restricted use accepted (e.g. non-residential areas)
Phenols 10 40
PAH 20 200
PCB 0,05 0,5
733
Despite the used trigger values currently being under examination and possible changed in the future, these values can be used for comparative purposes in order to get an indication of the degree of contamination in the demolition waste. The PAH and PCB contents in the hollow core slab waste material were far below the values for unrestricted use for soils. The PCB content in the building demolition waste was approximately equal to the lower value for unrestricted use, but far below the value for use of soils in non-residential areas. The demolition waste is quite alkaline due to the liberation of CaO during crushing. The upper layer of the demolition waste is anticipated to be carbonated with time. The alkalinity is also typical for concrete constructions and no restrictions of use of concrete constructions are known. However, on sensitive water areas where the water flows are not mixed with big water volumes, risk assessment of the alkaline waters on aquatic life might be necessary. Quality requirements for mineral demolition wastes were recommended for the leaching of some metals and for the total content of PCB and PAH. The content of PCB and PAH should not exceed Finnish trigger values for unrestricted or restricted use depending on the site. The target values for the leaching of the sulphate and some metals in the CEN-test with two steps are given in Table 7. The chosen metals are the most common metals which might be present in building wastes. The target values were estimated to be achievable for non-contaminated demolition wastes. Moreover, the values are also in agreement with the proposed Dutch values for maximum leaching in column test which have been calculated based on an acceptable immission of leached compounds from secondary raw materials which are used in earth constructions (Aalbers et al 1993). The target value for sulphate is the same as given in the Netherlands for mineral demolition waste for unrestricted use.
Table 7. Proposed target values for leaching from mineral demolition waste at L/S 10. Proposed test method is CEN-batch leaching test with two leaching; steps.
Element
Target (mg/kg)Value
Chromium Copper Lead
0,02 0,5 0,5 1,0
The conclusion of the study on the demolition wastes studied were following: 9 Crushed hollow core slab wastes may be used without restriction, but for use in the vicinity of sensitive ground water areas special considerations on environmental effects are needed. 9 The building demolition wastes are not recommended to be used on ground water areas. Building demolition waste are also recommended to be isolated with a water impermeable layer for example asphalt in order to avoid any risks with organic compounds from improper demolition of building.
6. A R E C O M M E N D A T I O N F O R A Q U A L I T Y A S S U R A N C E S Y S T E M F O R D E M O L I T I O N WASTE FROM BUILDINGS These recommendations are developed for mineral demolition wastes arising from a selective demolition of buildings and constructions. A proper demolition plan of buildings is the most important. Mineral demolition wastes should not be recovered from constructions contaminated by industrial activity or spills. Crushed demolition wastes are recommended to be used outside sensitive areas, e.g. ground water areas. The material shall be isolated with a water impermeable layer, such as asphalt.
734
In this work it was concluded after several discussions with experts (also including environmental authorities) that the quality assurance system of the crushed material can only be a spot control and therefore a simple sampling strategy was chosen. If crushed demolition wastes is used without any restrictions a more extensive test protocol needs to be created for quality control. The proposed recommendations for the quality assurance system are as follows: 9 the test samples are taken after crushing, preferably from a falling stream 9 the samples are collected in a vessel with dimension preventing particles to jump out of the vessel and with a width exceeding the conveyor belt. The vessel is to be only half full. 9 three separate test samples are taken at the same time and of these three samples one is randomly chosen for analysis. The two remaining samples are stored until the examined test sample is proved to be acceptable. If the examined test sample fails the quality requirements, the both other test samples are examined. 9 the weight of each test sample is about 50 kg. 9 at least 500 t materials shall be crushed between two subsequent sampling moments 9 the sampling frequency is the following:
"Amountof crushed material 0n the crushing site
Sampling frequency
l < 10.000 t/year > 10.000 t/year
1 sampling/for every starting 2500 t of crushed material 1 sampling/for every starting 2500 t until 10.000 t and after this 1 sampling/for every startin~ 5000 t of crushedmaterial
9 the particle size of the sample is reduced to less than 4 mm, after which the sample is divided into a laboratory sample (1 kg) e.g. using a riffle box 9 the laboratory sample is studied using the CEN-test with two leaching steps (target values presented in Table 7) ACKNOWLEDGEMENT This work was carried out with support of Lohja Rudus Environmental Technology Ltd. References. 1. Aalbers, Th.G. et al. Milieuhygienische kwaliteit van primaire en secundaire bouwmaterialen in relatie tot hergebruik en bodem- en oppervlaktewaterenbescherming. Rijksinstitutuut vor Volksgezondheid en Milieuhygiene, RIVM-rapport no 771 402 006. The Netherlands (1993). 2. CROW. Uitlogen op karakter, Handboek Uitloogkarakterisering II Materialen, The Netherlands (1994). 3. Johansson, H.G., Ydrevik, K. & Arvidsson, H. Crushed concrete - a material for use in construction of roads, VTI Notat 1-1996, Link0ping, Sweden (1996) (in Swedish). 4. K~lvesten, E. Milj0m~issig karakterisering av v~igbyggnadsmaterial. (Laboratory study of the leaching from road building materils) Examensarbete LiU-IFM-Kemi-Ex504, Univerisity of Link0ping, Sweden (1996) (in Swedish). 5. Ministry of the Environment. Contaminated soil sites and their managament in Finland, Contaminated soil site survey and remediation project, Final report. Memorandum 5/1994, Ministry of the Environment in Finland, Department for environmental protection (1994) (in Finnish). 6. Mulder, E. The Leaching behaviour of Some Primary and Secondary Raw Materials Used in Pilot-scale Road bases. Presented in: Waste Material in Construction. Ed. J.J.J.R. Goumans, H.A. van der Sloot & Th.G. Aalbers, 1991 Elsevier Science Publishers B.V., 255-264. 7. NEN 7343: Leaching characteristics of building materials and solid waste material - Leaching tests - Determination of leaching characteristics of inorganic components from granular and building waste materials. NNI, Delft, The Netherlands (1992). 8. NEN7345: Leaching Characteristics of soil and stony building and waste materials - Leaching tests - Determination of the leaching of inorganic components from building and monolithic waste materials with the diffusion test. NNI, Delft, The Netherlands (1995). 9. NT ENVlR 002: Solid waste, granular inorganic material: Column test. Nordtest, Espoo, Finland (1995). 10. prEN 12457, Characterization of waste - Leaching - Compliance test for leaching of granular waste materials. Determination of the leaching of constituents from granular waste materials and sludges, Draft (1996). 11. Sigfrid, L. Milj0st0rande material i rivningsavfall (Harmful materials in building wastes). Reforsk FoU 81, Sweden (1993).
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
735
ENVIRONMENTAL CERTIFICATION OF BOTTOM ASHES FROM COAL FIRED POWER PLANTS AND OF BOTTOM ASHES FROM MUNICIPAL WASTE INCINERATION
F.J.M. Lamers 1, J.W. 1KEMA, P.O. Box 9035, 2Dutch Fly Ash Corporation, P.O. 3Waste Processing Association, P.O.
van den Berg 2 and J.G.P. Born 3 6800 ET Arnhem, the Netherlands Box 301, 3730 AH De Bilt, The Netherlands Box 19300, 3501 DH Utrecht, The Netherlands
ABSTRACT
Environmental certification is the preferable method for building materials producers to demonstrate that their products comply with the relevant environmental legislation. A procedure for environmental certification according to the Dutch Building Materials Decree is described for granular materials. The operation of this procedure is shown for two granular products, namely bottom ashes from conventional coal fired power plants and bottom ashes from municipal waste incineration. 1
INTRODUCTION
In the Netherlands, an aspect of growing importance for building materials in their utilization is the environmental quality. In 1995 the so called Building Materials Decree has been issued by the Dutch Ministry of Housing and the Environment. In this Decree, that is part of the Soil Protection Law, rules are set towards the environmentally responsible utilization of building materials. For this purpose, demands are set regarding the immission of 21 inorganic components from a building material into the soil. The immission into the soil is calculated from the results of laboratory leaching tests. Based on the level of measures against infiltration into the soil of leachate, two utilization categories are distinguished, with different emission boundaries: Category 1 - no restrictions and an estimated infiltration of 300 mm/year Category 2 - restricted utilization, estimated infiltration of 6 mm/year. Apart from that, compositional limits are set regarding the content of a set of organic components. In table 1 an overview is given of the calculated emission demands for Category 1 and Category 2 granular building materials, for a height of respectively 1.5 m (relevant for bottom ashes from coal fired power plants) and 50 m (relevant for bottom ashes from municipal waste incineration). For granular building materials the column test is the compulsory test method.
736
The emission and composition boundaries will be implemented in 1998. The building materials producers have had the opportunity then to make sure that they can meet with the demands. In the explanatory note to the Building Materials Decree, the Ministry of VROM states that environmental certification is the main acceptable proof that the demands of the Building Materials Decree are met. Table 1
Calculated leaching limits for Category 1 and Category 2 Building Materials, for granular materials, for a utilization height of respectively 1,5 and 50 m (relevant for CF-bottom ashes and MWI-bottom ashes)
Component
As
Ba Cd Co Cr Cu Hg Mo Ni
Pb Sb Se Sn V Zn Br CI CN-complex CN-free F SO4
Category 2 Category 1 Granular material,leaching values in Granular Materials leaching values in mg/kg mg/kg 0.86 3.4 0.026 0.30 0.69 0.47 0.018 0.21 0.85 1.3 0.035 0.036 0.15 1.3 2.9 2.7 580 0.032 0.006 7.2 1100
7.0 55 0.061 2.3 11.7 3.3 0.075 0.84 3.5 8.1 0.41 0.094 2.3 32 14 4.0 8800 0.35 0.07 96 22000
A general framework for the environmental certification of granular materials is described and subsequently illustrated for bottom ashes from coal fired power plants (CF-bottom ashes) and bottom ashes from municipal waste incineration (MWI-bottom ashes). Both the Dutch Fly Ash Corporation (CF-bottom ashes) and the Waste Processing Association (MWI-bottom ashes) are developing certification procedures for their materials.
737 2
FRAMEWORK FOR ENVIRONMENTAL CERTIFICATION
2.1
General
For a certificate of environmental quality, two possible schemes can be followed: -
lot by lot inspection; every lot is tested and evaluated separately
-
production certification; a certificate is issued regarding the expected quality of the production process.
The production certification is based on the so-called moving
average. If the moving average exceeds the limits of the category in which the building material is utilized than correction measures are taken to improve the quality of the building material. An example of a moving average chart is given in figure 1.
1.6
leaching / demand leaching limit
1.4 1.2
........... ~
...... [..........................................................................................................................
1 0.8 0.6
0.4 0.2 0
line connecting individual analyses 3
4
5
6
7
8
moving average line
9 10 11 12 13 14 15 16 17 18 19 20
sample number (successive in time) I-indiv. analyses
Figure 1
9moving average I
Example of a moving average chart for quality control of granular building products
For most building materials producers production certification is the preferable method for a certificate of environmental quality. Production certification safeguards a stable supply to the market and makes long term supply contracts possible.
738 The following aspects play an important role with production certification of building materials in general: -
reduction of errors due to variation in the total analytical procedure
-
representative sample taking and sample preparation
-
selection of critical elements a test scheme that meets with the quantities that are produced of the specific material and
-
with the risks that the emission limits are exceeded -
statistically based interpretation of data, based on consumers risks and producers risks
-
the use of short tests to predict the results from the compulsory test.
2.2
Reduction
of errors
The total standard error in leaching results is built up from the following elements: ./ +S 2leachinganalysis Stotal=yS 2sample+S 2preparation where" Stotal
= the total standard error in leaching analyses
Ssample
- the standard error during sampling (both due to inhomogeneity of the lot and due to the influence of the sample size)
S preparation
- the standard error during sample preparation (milling, comminution, preparation of a laboratory sample)
Sleaching analysis
= the standard error during the leaching procedure and subsequent chemical analysis of the leachate.
By increasing the number of increments per sample, by using a sample splitting apparatus and by analysing several samples, the influence of the standard error on the final result can be lowered. 2.3
Representative
sample
taking
and
sample
Representative sampling is dependant of the following factors -
grain size of the material
-
percentage of the material that contains the critical element
-
variation inside the lot.
preparation
739 Based on the formulas for sample size derived from NEN 7301 a minimum sample size is calculated. Variation inside the lot can be counterbalanced by taking one sample in several increments. The necessary increment size is dependant on the grain size of the building material. If possible, the increments are taken from a moving stream. 2.4
Selection of critical elements
Based on a statistically relevant number of samples (minimum 10) a selection is made of the critical elements. For as specific building material critical elements are those elements that lead potentially to exceeding of the relevant limits. They are described as elements with Xaverage,leaching> 25% of the relevant limit of the Building Materials Decree. For production certification only the critical elements should be analyzed on a regular basis. For the critical elements the distribution of Xleaching is established (normal, log normal or undefined) to be able to choose a method of assessment of the leaching results against demands from the Building Materials Decree. 2.5
Testing scheme
Based on the quantity of building material that is produced and based on the chance that a limit is exceeded the testing scheme is chosen. In the testing scheme thenumber of leaching analyses per quantity of building material is established. 2.6
Producers risk and consumers risk
The consumers risk and the producers risk are defined as: -
the risk that a defect lot is incorrectly approved (consumers risk)
-
the risk that a correct lot is incorrectly rejected (producers risk).
During production certification both the producers risk at a specific percentage of defects (higher than 50%) and the consumers risk at a specific percentage of defects (lower than 50%) are brought down to an acceptable low level.
740 The risks are determined by the number of samples that is analyzed and by the rejection limit that is chosen. The rejection limit can be generally written for normal distributions
as
Xr,jection
= limit Building Materials Decree + k'standard error (k can both be positive and negative). For log normal distributions the formula is slightly different.
2.7
Short tests
The compulsory column test takes 21 days. This makes it impossible to take quick correcting measures after the start of deterioration of the quality of a granular building product. The use of short tests such as the test developed by CEN TC 292 or the Dutch Cascade Batch test can offer possibilities to reduce the necessary time-span for analyses. The possibilities to do that are largely dependant on Xav,r,ge for the most critical component in relation to the limit from the Building Materials Decree and on the relation between the short test and the compulsory test.
BOTTOM ASH FROM COAL FIRING (CF-BOTTOM ASHES) 3.1
General
CF-Bottom ash is the bottom ash that originates from coal combustion in dry bottom boilers. It exists of coarse ash particles that have been coagulated and sintered and subsequently quenched in the water basin that acts as a water lock for the boiler. Yearly about 80,000 tons of CF-bottom ashes are produced. CF-bottom ash is almost exclusively used in granular form as embankment material or road base material. The material is used up to a hight of 1.5 m as a light embankment material in weak soils and because of that comes into contact with the ground water. This means that it can be only utilized as a category 1 building material. If category 1 is exceeded only utilization in concrete is possible for which momentarily no market exists in the Netherlands.
3.2
Critical elements
In 1990 and 1991, the composition and leaching behaviour of 40 CF-bottom ashes was determined. Between 1991 and 1996 quality improvement measures were carried out and two new power plant lines were started.
741 In 1996 a new testing series was started to get actual information about leaching of the CFbottom ashes per power plant. In table 2, the leaching in the column test of 40 CF-bottom ashes (1990-1991) is shown. As can be seen in table 2, the variation in leaching behaviour is large, with variation coefficients running up to 275%. This is partly due to the fact that leaching from the CF-bottom ashes is relatively low. From table 2 it can be deduced that barium, molybdenum, selenium, antimony, sulphate and vanadium are potential critical elements. Table 2
Component As Ba Cd Co Cr Cu Hg Mo
Ni
Pb Sb Se Sn V Zn
S04
The leaching behaviour of 40 Dutch CF-bottom ashes
Average leaching (mg/kg) 0.113 3.145 0.002 0.039 0.013 0.044 0.0025 0.120 0.0855 0.015 0.015 0.053 0.025 0.253 0.241 538
Standard deviation (mg/kg) 0.0929 2.862 0.003 0.109 0.007 0.117 0.0005 0.097 0.195 0.005 0.0063 0.052 0.005 0.169 0.301 607
Limit Category 1 BMD (mg/kg)
Variation coefficient
0.86 3.4 0.026 0.30 0.69 0.47 0.018 0.21 0.85 1.33 0.035 0.036 0.15 1.3 2.9 1100
82.24 91.01 148.26 276.49 54.49 262.74 2O 80.8 228 33.3 40.7 97.8 20 67 124.45 112.7
(%)
The classification of CF-bottom ashes is therefore uncertain. From table 2 it appears that about 50% will fall in category 1 - which means unhindered utilization and 50% will fall in category 2 (isolation demands). Specifically for selenium the leaching often lies at the boundary of category 1 / category 2.
742 3.3
Procedures
for certification
Based on the data from 1990-1991 a provisional assessment guideline for certification was drafted in 1996. This consists of the following elements: -
method of sampling
moving stream
-
sample size
25 kg
-
number of increments
2O
-
sample frequency
1 sample per 2000 tons
-
certification method:
lot by lot
-
rejection limit
every analysis has to comply with the limits from the Building Materials Decree
-
3.4
critical elements Actual
Ba, Mo, Sb, Se, SO 4, V status of certification
Testing of the possibilities for certification was started with the CF-bottom ashes of two power plants. The CF-bottom ashes of one of those two Dutch power plants comply with the demands for certification. These ashes are supplied to the market with a certificate. For the CF-bottom ashes from the other power plant quality improvement is necessary. Momentarily research is carried out regarding the quality of the CF-bottom ashes from the other power plants. Preliminary results show that for a number of power plants quality improvement of the CF-bottom ashes is called for. 3.5
Further development
of certification
The following aspects for CF-bottom ash certification are under development; -
development of a short test. The possibilities of the CEN TC 292 test as a short test for quality control are tested. Preliminary results for selenium leaching show a good correlation between the column test and the CEN-TC 292 test as shown in fig. 2
-
quality improvement measures for the other power plants
-
actualization of the provisional scheme for certification. Attention will be given towards the statistical basis for evaluation of the analytical results (can we make use of a moving average system?)
743 0.25 0.2
y = 0.9183x + 0.0005 R2 = 0.885 []
B []
9
v
~ 0.15 9
[]
~ r
= datapoint CEN / column regression line CEN / column
[]
[]
0.1 0.o5
i
t
I
I
0.05
0.1
0.15
0.2
0.25
leaching in CEN TC 292 test (mg/kg)
Figure 2
Leaching behaviour of selenium from CF-bottom ashes of 6 power plants, tested both in the compulsory column test NEN 7343 and in the CEN TC 292 short Batch test BOTTOM ASH FROM MUNICIPAL WASTE INCINERATION (MWI-BOTTOM ASH)
4.1
General
MWI-bottom ash is the solid residue from combustion of municipal waste or in a Municipal Waste Incineration Furnace. Often MWI-bottom ashes have been subjected to a post treatment consisting of magnetic separation of iron and sieving and comminution of particles > 40 mm. Fly ashes from Municipal Waste Incineration are kept separate from the MWIbottom ash. In the Dutch situation it is forbidden to prepare mixed ashes from fly ash and bottom ashes. In 1996 800,000 tons of MWI-bottom ash were produced in the Netherlands. The last years MWI-bottom ash is utilized for 100%, primarily in granular form as embankment material up to a hight of 10 m or more or as a road base material.
744 MWI-bottom ashes are supplied to the market with a certificate for its technical and environmental behaviour. The environmental part of this certificate is based on "old" legislation. MWI-bottom ashes up to now always comply with the demands for environmental certification. The Building Materials Decree enforces more severe demands than the present regulations. Because of that a large part of the MWI-bottom ashes does not comply with the demands from the Building Materials Decree. To safeguard its outlet to the market the Dutch Ministry of the Environment has developed a "Special Category for MWI-bottom ashes". In this category MWI-bottom ashes can be utilized under a set of isolation measures. With the Municipal Waste Incineration sector the appointment has been made to pursue steady quality improvement of its byproducts so that MWI-bottom ashes can be utilized as Category 2 Building Materials in future. 4.2
Critical elements
Since 1987 MWI-bottom ashes have been subjected to a regular quality control from which the environmental part is based on the serial batch test NEN 7349. This test is however not the compulsory test for the Building Materials Decree. Since 1991 all the Dutch Municipal Waste Incineration plants have also carried out column tests (the compulsory test for the Building Materials Decree) on their MWI-bottom ashes. The purpose was to build up sufficient leaching data to be able to prepare environmental certification according to the Building Materials Decree and to show the extent of quality improvement that has been realized in the run of years. In table 3 the leaching data for 1996 are shown.
745 Table 3 Component
As Cd Cr Cu Mo Ni Pb Sb Zn Br CI SO4
Leaching data for 26 column tests from MWI bottom ashes during 1996 Average leaching (mg/kg) O.054 0.003 O.O77 1.96 2.09 0.17 0.16 0.20 0.16 10.4 3040 5360
Standard deviation (mg/kg) 0.195 0.003 0.22 1.27 2.91 0.30 0.42 0.15 0.24 6.0 B.a. B.a.
Limit Category 2 BMD (mg/kg) 7.0 0.061 11.7 3.27 0.84 3.5 8.2 0.42 14 4.0 8800 22000
Variation coefficient
(%)
361 105 291 65 139 176 260 72 149 58 B.a. B.a.
From table 3 it can be concluded that the following critical elements exist for MWI-bottom ashes: Cu, Mo, Sb and Br. Based on the average leaching of chloride > 0.25 * U2, chloride could also be considered as a critical element. However, during the total period 1991 - 1996 the leaching limit for category 2 has only once been exceeded. A significant reduction in copper leaching has been effected between 1991 and 1996. For the critical elements the distribution of the leaching data has been established. The data can both be described by a normal and a log normal distribution. Because the fitting for a normal distribution seemed slightly better, the testing criteria have been based on a normal distribution of the results. Based on the leaching behaviour of bromide, presently all MWI-bottom ashes should be considered as MWI-bottom ashes. For most Municipal Waste Incineration plants also the leaching of molybdenum exceeds the Category 2 limits.
746 4.3
Procedures
for certification
An assessment guideline for certification is under development. Two classes for certification are distinguished, viz. Category 2 MWI-bottom ash and Special Category MWI-bottom ash. A definite assessment guideline is expected in 1997. The assessment guideline exists of the following aspects: -
method of sampling
moving stream
-
sample size
180 kg
-
number of increments
2O
-
sample frequency
1 sample per 2, 3 or 6 weeks, depending on the yearly quantity of MWI-bottom ash per MWI plant
-
-
-
certification method category 2 MWI-bottom ash:
production certification
special category MWI-bottom ash
production certification
limit correcting measures category 2 MWI-bottom ash
Xmoving average, 8
critical elements
Cu, Mo, Sb, Br
~>limit Category
2
The chosen method of certification is production certification. Based on the historical percentage of lots that fully meet with the demands for Category 2, the MWl-bottom ash of a specific Municipal Waste Incineration Plant is integrally classified as Category 2 MWI-bottom ash or as Special Category MWI-bottom ash. More than 75% has to meet with the demands for category 2 to be classified as Category 2 MWI-bottom ash. After the first occasion that the moving average for a Category 2 MWI-bottom ash exceeds the limits of category 2, correcting measures are taken to improve the quality. If within a set number of successive analyses the moving average is not brought down below the category 2 limit, the production certificate for category 2 can be (temporarily) withdrawn. 4.4
Actual
status
of certification
All MWl-bottom ashes are supplied with a certificate according to the existing regulation. Certification according to the Building Materials Decree will be introduced after the enforcement of the leaching limits from the Building Materials Decree. This enforcement is set for 1998.
747 MWI-bottom ashes of all Municipal Waste Incineration plants should be considered then as Special Category MWI-bottom ash. This is based on the leaching behaviour of bromide (always) and molybdenum (for most Municipal Waste Incineration plants) during 1996. 4.5
Further development of certification
The following aspects for MWI-bottom ash certification are under development; -
development of a short test. The possibilities of the CEN TC 292 test as a short test for quality control are considered. A comparison of the results from the column test and the serial batch test (from which a large set of data is available) has been carried out in 1996. More than 50% of the datasets batch-test / column test were not fit for comparison. A translation serial batch - column seems possible, but appears not reliable enough in the case of data near a leaching limit
-
demonstrations with quality improvement measures are carried out (for instance accelerated aging).
5
DISCUSSION
Both CF-bottom ashes and MWI-bottom ashes show a leaching level for the most critical element around one of the leaching limits. Because of that aspect, production certification is preferred; lot by lot inspection may lead to a high percentage of rejection, whereas a production certificate safeguards a more continuous supply. Leaching levels exceeding the limit set by the Building Materials Decree are allowed, provided they are compensated by other parts of the production, showing a lower leaching level. The use of short tests as an alternative for the compulsory column test is under development. The inherent lower reliability that is caused by the translation of one test to another may be "smoothed out" in the case of a moving average method.
748 LITERATURE
LAMERS, F.J.M. and BERG, J.W. VAN DEN, 1995. Environmental certification of pulverized fuel ash and bottom ash. Proceedings Power Gen Europe '95, Penwell, pp. 581-598 MINISTRY OF HOUSING AND ENVIRONMENT, 1995. Building materials Decree for the soil protection and surface water protection (in Dutch). NNI, 1995. NVN 7301, Leaching characteristics of solid earthy and stony building and waste materials. Sampling. Sampling of granular materials from streams. NNI, 1995a. NEN 7343, Leaching characteristics of solid earthy and stony building and waste materials. Leaching tests. Determination of the leaching behaviour of inorganic components from granular materials with the column test (in Dutch). NNI, 1995b. NEN 7349, Leaching characteristics of solid earthy and stony building and waste materials. Leaching tests. Determination of the leaching behaviour of inorganic components from granular materials with the cascade test (in Dutch).
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
WASCON 97 June 4 - 6, 1997 Houthem
St G e r l a c h
The Netherlands
W o r k s h o p DF2 -Quality Control and Certification
Q u a l i t y A s s u r a n c e in t h e i.,,~u,,. o h . ~ .,.~,,, ~ - ~ . :, A n a l y s i s o f Soils
Leslie Heasman M J C a r t e r Associates Station House Long Street Atherstone Warwickshire UK. CV9 IBH Tel +44 1827 717891 Fax +44 1827 718507
749
750
Quality Assurance in t h e
Analysis of Soils
Land is a resource that is scarce in many countries. In the industrialised countries land which becomes available for development in or near to the industrial and business centres of cities and towns has usually been used before and land which remains under use has frequently been subject to a variety of previous uses. Historically little consideration was given to the potential environmental impacts of industrial activities hence the controls over processes and emissions were not as careful as they are today. In order to minhnise the use of green field land for new developments, companies are encouraged to redevelop previously used sites. This results in a need to measure the contamination present at the site and to assess the associated environmental risk. Based on an assessment of risk proposals are made for the remediation deemed necessary to return the site to beneficial use. The entire process of site assessment and remediation frequently involves a number of different professionals with very different backgrounds and experiences. These include: 9 9 9 9 9 9 9 9 9 9 9 9
landowners developers financial advisers bankers insurers land agents/surveyors legal advisers/lawyers planners architects/engineers regulators environmental advisers chemists/laboratories
The different backgrounds and degree of knowledge of each of the parties involved in the redevelopment of potentially contaminated sites can create a lack of confidence by each of the various parties in the capabilities, knowledge and skill of the others. It is only when some degree of confidence is achieved between the different professional interests and some degree of trust is established that the main objectives of brown land redevelopment are likely to be achieved. The area of the analysis of contaminated soil is one in which it is vital to ensure confidence in data as it is the baseline on which risk assessments and remediation programmes are founded. Little work has been carried out on methodology for the analysis of contaminated soils and there is generally a poor understanding by the users of laboratory generated data of the confidence which can be attached to the data generated by any one laboratory. Many examples are cited of extreme discrepancies between data generated by different laboratories asked to analyse the same samples and between duplicate samples analysed in the same laboratories. A number of initiatives are underway in the UK to improve confidence and understanding between the laboratories which analyse contaminated soils and those who use the data in an assessment of the environmental impact of the sites from which the soil was sampled.
751 A report has been prepared by the UK Environment Agency ~ on quality assurance in the analysis of soils from contaminated land. The report is intended for users of laboratory generated data and seeks to describe and explain the quality control procedures used in analytical laboratories. The report reviews the existing application of quality assurance in the area of contaminated soil analysis. The different quality assurance procedures are explained and their benefits outlined. Recommendations are made for best practice regarding quality assurance in this technical field. Aspects of the general process that are addressed include sampling, transportation of samples, sample preparation, analysis and reporting of data. A report is in preparation by the UK Environment Agency 2 on the available laboratory methods used for the analysis of contaminated soil. The report is intended to inform users of laboratory data of the existing choice available in analytical methods and the different data which is generated. The effect of the use of different extraction procedures is explained as is the effect of using different analytical techniques. The report identifies the benefits and limitations of each of the methods used currently in the UK for the analysis of priority contaminants in soils. There is much discussion and confusion regarding the use of different analytical methodologies for the analysis of the same parameter. It is considered by some that consistency and comparability can only be achieved by the selection of single analytical methodologies for specific contaminants. This is not generally the view of the laboratories as this approach would eliminate the potential to develop and improve methods and would reduce flexibility. It is generally agreed by laboratories and the infor,ned users of laboratory data that provided that a method can be validated according to an accepted protocol with specified acceptance criteria standard methods need not be used. Work has been carried out funded by the UK Environment Agency3 to develop a validation protocol for methods used in the analysis of contaminated soil. The protocol provides guidance on criteria which can be used to assess whether analytical method performance is satisfactory and on the approach laboratories can take to provide data which can be assessed against these criteria. The validation protocol will provide a means by which non standard methods can be validated or standard methods can be validated for different soil types. Work has been carried out funded by the UK Environment Agency4 to identify and coordinate research related to the laboratory analysis of contaminated soil. It is concluded that lhnited research and development work is carried out in a number of areas related to the laboratory analysis of contaminated soils however the work is disparate and not well coordinated. A report on the work considers the drivers for improved methods of analysis for contaminants in contaminated soils which include the need for reliable data which is fit for purpose: to inform policy to support regulation for effective management of contamination 9 to reduce the commercial risks associated with land transfer and development
752 The report identifies the key organisations involved in funding and managing research into the laboratory analysis of contaminated soils and in identifying and improving existing analytical methodologies. Existing or planned relevant research progrmrunes are identified. The report includes proposals for the communication of research outputs and for liaison and coordination between the different parallel research programmes. A further initiative which has developed in the UK to improve confidence in the assessment and redevelopment of contaminated land is the formation of a body, The Forum on Contamination in Land (FOCIL), which coordinates the approach of the main professional institutions to this technical area. The FOCIL mission statement is to enhance the
understanding of and facilitate improved coordination between professionals dealing with contaminated land thus benefiting the process of assessing and managing risks associated with contamination, both commercially and environmentally. Key objectives of FOCIL include: to expand awareness, knowledge and competence among professional advisers and develop an integrated approach to contamination in land to identify, disseminate and encourage the use of best practice to assist market confidence, enhance environmental benefits and encourage sustainable development The activities of FOCIL include: 9
dissemination of information on available best practice guidance co-ordination and communication meetings with other groups involved with contaminated land
9
involvement in development and preparation of best practice guidance
9
hosting seminars and workshops on relevant topics and issues provision of a unified professional response to Government and other bodies on contamination issues
9
maintenance of listings of specialists in contaminated land matters
In order to improve cormnunication and understanding between the users of laboratory data and those who request analyses and use the data in their assessment of potentially contaminated sites FOCIL produced the sheets presented in Annex 1 to ensure that both parties involved in the analytical process are aware of all the relevant facts and have all the information necessary to ensure that the analyses carried out are fit for purpose and that the resultant data are of a known quality.
753 References
Laboratory of the Government Chemist, M J Carter Associates, Acer Environmental, H B Berridge and Partners. Quality Assurance Associated with the Analysis of Soil from Contaminated Land. UK Environment Agency. In press 1997. .
.
.
Methods for the Chemical Analysis of Soils from Potentially Contaminated Land. UK Environment Agency. In preparation 1997. Hyder Environmental, M J Carter Associates, H B Berridge and Partners. Validation of Methods for the Analysis of Soils from Potentially Contaminated Land. Recommendations from the Working Group on Soil Chemical Analysis. UK Environment Agency. In preparation 1997. M J Carter Associates, Hyder Environmental, H B Berridge and Partners. Framework for the Identification, Prioritisation and Coordination of Research Related to the Laboratory Analysis of Contaminated Soil. UK Environment Agency. In preparation 1997.
This Page Intentionally Left Blank
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved. W . M . A . J . W i l l a r t , I n i s t r y of The Hague, T h e N e t h e r l a n d s .
Dutch policy materials
as
incentive
Housing,Spatial
to
environmentally
755 Planning
and
controlled
the
Environment,
re-use
of
waste
Introduction The b u i l d i n g i n d u s t r y has a l o n g h i s t o r y o n the r e u s e of o l d m a t e r i a l s . T h e f i r s t c h r i s t i a n c h u r c h e s w e r e o f t e n b u i l t o n the s i t e of o l d t e m p l e s , w i t h the b u i l d i n g materials of t h e s e t e m p l e s . In o l d c i t i e s buildings and b u i l d i n g m a t e r i a l s h a v e b e e n r e n o v a t e d a n d r e - u s e d . T h e u s e of i n d u s t r i a l w a s t e s as b y - p r o d u c t s is of a m o r e r e c e n t date. B l a s t f u r n a c e s l a g a n d f l y a s h are u s e d in cement, r o a d b a s e s o f t e n w e r e m a d e w i t h s l a g or b u i l d i n g rubble. A f t e r the s e c o n d w o r l d w a r r u b b l e was u s e d in G e r m a n y a n d T h e Netherlands as a g g r e g a t e for the b u i l d i n g of n e w h o u s e s . In t h i s p e r i o d f r e s h b u i l d i n g m a t e r i a l s w e r e s c a r c e a n d the r u b b l e h a d to be r e - u s e d or landfilled. E n v i r o n m e n t was n o t an i s s u e at t h a t time. S o m e t w e n t y y e a r s ago the reuse of waste materials became politically interesting. The a s p e c t s of l e s s e n i n g the a m o u n t of f r e s h l y d e l v e d m a t e r i a l s on one h a n d a n d the a r e a o c c u p i e d b y l a n d f i l l on the o t h e r h a n d w e r e the d o m i n a n t f a c t o r s . The f i r s t activities were directed to l a r g e w a s t e streams as b u i l d i n g r u b b l e a n d f l y a s h f r o m coal f i r e d p o w e r s t a t i o n s . T h e u s e of f l y a s h w a s s t i m u l a t e d b y a b a n on l a n d f i l l for f l y ash. T h e c e m e n t i n d u s t r y w a s p u t under pressure to u s e f l y a s h in the n e g a t i o n s for e x t e n s i o n of t h e i r m i n i n g c o n c e s s i o n . O n this p r o d u c t the p r o d u c e r s of the a s h a n d the user, the c e m e n t i n d u s t r y h a d a m u t u a l i n t e r e s t a n d all f l y a s h in The N e t h e r l a n d s is u s e d mow. B u i l d i n g r u b b l e w a s a t o u g h e r s t r e a m to t a c k l e . The number of p a r t i e s i n v o l v e d was l a r g e r a n d the q u a l i t y of the p r o d u c e d m a t e r i a l s v a r i e d l a r g e l y . '~ogether w i t h the i n d u s t r y the g o v e r n m e n t w o r k e d on standards, certification of different materials. A very positive influence had the p r e s c r i p t i o n to u s e the m a t e r i a l in g o v e r n m e n t and m u n i c i p a l p r o j e c t s . At this t i m e r o u g h l y 90% of the b u i l d i n g w a s t e s a r e reused a n d s i n c e j a n u a r y this y e a r we h a v e a b a n o n the l a n d f i l l i n g of reusable building wastes. Policy At f i r s t t h e r e was no i n t e g r a t e d p o l i c y t o w a r d s w a s t e m a t e r i a l s a n d t h e i r re-use, m a n a g e m e n t was j u s t s o l v i n g a c t u a l p r o b l e m s . Later a new policy was d e v e l o p e d in w h i c h a p r e f e r r e d s e q u e n c e of a c t i v i t i e s c a m e f o r w a r d . The "steps" in this p o l i c y are: - p r e v e n t i o n of w a s t e - r e u s e of p r o d u c t s -reuse of m a t e r i a l -burning with energy recovery -landfill. To a t t a i n the a i m s of the p o l i c y a n u m b e r of p u s h a n d p u l l m e a s u r e s w e r e taken. P u s h i n g m e a s u r e s u s e d are: -a l a n d f i l l b a n for d e s i g n a t e d c a t e g o r i e s of w a s t e - t a x a t i o n on l a n d f i l l e d w a s t e s - t a r g e t s in the e n v i r o n m e n t a l p e r m i t s of p r o d u c e r s - d e p o s i t s on n e w p r o d u c t s to p r o c e s s t h e s e p r o d u c t s
when
defunct
P u l l i n g f a c t o r s are: - p r o v i d e r e g u l a t i o n s for the u s e of s e c o n d a r y m a t e r i a l s - s t i m u l a t e s t a n d a r d i z a t i o n a n d c e r t i f i c a t i o n of s e c o n d a r y m a t e r i a l s - p r e s c r i b e the u s e of s e c o n d a r y m a t e r i a l s for g o v e r n m e n t b u i l d i n g s a n d roads - s u p p o r t for r e s e a r c h a n d d e v e l o p i n g n e w t e c h n o l o g i e s for p r o c e s s i n g w a s t e s
756 to u s e f u l
products
and
realisation
of p i l o t
projects.
Results There has n e v e r b e e n an e v a l u a t i o n of the i n f l u e n c e of the d i f f e r e n t m e a s u r e s t a k e n but I c a n g i v e y o u an o v e r v i e w of r e c e n t d e v e l o p m e n t s . i.
2.
3.
4
L a n d f i l l ban: was i n t r o d u c e d in 1995 for o v e r t w e n t y d e s i g n a t e d w a s t e c a t e g o r i e s r a n g i n g f r o m car w r e c k s to h o u s e h o l d w a s t e s . The a i m is to p r e v e n t the l a n d f i l l of u s e f u l r e s o u r c e s a n d if not r e - u s a b l e t h a n r e c o v e r the e n e r g y or u s e the o r g a n i c w a s t e s in a c o m p o s t p r o c e s s i n g plant.As new techniques of w a s t e processing become available new t y p e s of w a s t e are a d d e d to the l a n d f i l l ban. R e c e n t a d d i t i o n s are b u i l d i n g a n d d e m o l i t i o n waste, b l a s t i n g g r i t a n d w a s t e wood. S i n c e the i n t r o d u c t i o n of the l a n d f i l l b a n the q u a n t i t i e s of l a n d f i l l e d wastes have dropped sharply. Taxation on l a n d f i l l e d wastes: this taxation serves several p u r p o s e s . The f i r s t is s i m p l y to g e t money. S e c o n d a r y r e a s o n s are to d i m i n i s h the f i n a n c i a l g a p in c o s t s b e t w e e n l a n d f i l l a n d b u r n i n g in an M W C P a n d b y r a i s i n g the p r i c e of d i s p o s a l , the f i n a n c i a l i n c e n t i v e to p r o c e s s a n d r e - u s e w a s t e grows. Targets in e n v i r o n m e n t a l permits. B y a d d i n g t a r g e t s on w a s t e processing in e n v i r o n m e n t a l permits the p r o d u c e r s is f o r c e d t a k e a c t i o n w i t h r e g a r d to his waste. It is n o t s i m p l y a m a t t e r of costs, he has to p r e v e n t or r e - u s e his w a s t e to c o n t i n u e his activities. In s o m e c a s e s a g r e e m e n t s of the g o v e r n m e n t with branch-organisations are made aimed at reducing waste or r e p r o c e s s i n g waste, e.g. p a c k a g i n g a g r e e m e n t . S o m e c o m p a n i e s h a v e t h e i r o w n p o l i c y on e n v i r o n m e n t a l targets with respect to reuse. A large company in this region, aims at r e d u c i n g the w a s t e l e a v i n g t h e i r p r e m i s e s to zero. T h e y h a v e a s y s t e m to s o l i d i f y a n d r e - u s e their own wastes in r o a d s o n t h e i r site. Deposits: It is a f a i r l y r e c e n t d e v e l o p m e n t , f i r s t introd1~ced to s t a r t a s e l e c t i v e d e m o l i t i o n of o l d cars. On new cars a d e p o s i t is p a i d to p a y for the f i n a l d e m o l i t i o n . The s y s t e m is q u i t e a s u c c e s s a n d h e l p e d to i m p r o v e the q u a l i t y s t a n d a r d s in the car d e m o l i t i o n b r a n c h .
So far the p u s h i n g f a c t o r s , p u s h i n g a l o n e is not s u f f i c i e n t to r e a c h the goals. P u l l i n g f a c t o r s are a l s o n e e d e d to r e a c h a g e n e r a l l y a c c e p t e d r e - u s e of m a n y p r o d u c t s . The a b o v e m e n t i o n e d p u l l i n g f a c t o r s are: 1
2
3.
4.
Regulations. The p u t t i n g of n e w p r o d u c t s on the m a r k e t is n o t e n o u g h to c o m e to the u s e of them. A g g r e g a t e s n o w can c o n s i s t of s e c o n d a r y m a t e r i a l s , a n u m b e r of y e a r s a g o y o u w e r e o n l y a l l o w e d to u s e n a t u r a l s a n d or g r a v e l . J o i n t r e s e a r c h f r o m all i n v o l v e d p a r t i e s l e d to n e w r e g u l a t i o n s w h i c h i n c l u d e d s e c o n d a ry m a t e r i a l s . Standardisation and certification. R e - u s e is m a d e p o s s i b l e b y a d a p t a t i o n of r e g u l a t i o n s . S e c o n d a r y m a t e r i a l s h a v e to c o m p l y with functional and environmental specifications, standards h a v e to be d r a w n up so that p r o d u c e r s as w e l l as u s e r s k n o w the q u a l i t y of the p r o d u c t s t h e y d e l i v e r or buy. To e n s u r e the c o m p l i a n c e w i t h the s t a n d a r d s c e r t i f i c a t i o n s y s t e m s are used. U s e of s e c o n d a r y m a t e r i a l s . To p r o m o t e the u s e of s e c o n d a r y m a t e r i a l s it is e s s e n t i a l that g o v e r n m e n t an m u n i c i p a l a u t h o r i ties lay d o w n an e x a m p l e b y p r e s c r i b i n g s e c o n d a r y m a t e r i a l s in t h e i r own b u i l d i n g s a n d roads. Support in the d e v e l o p m e n t of n e w t e c h n i q u e s and pilot projects. T h e r e are p r o g r a m m e s to s u b s i d i z e n e w t e c h n o l o g i e s a n d to p r o m o t e p i l o t p r o j e c t s for n e w p r o d u c t s m a d e of waste. O n e of t h e s e p r o g r a m m e s was d i r e c t e d at the r e - u s e of d r e d g i n g spoils, a m a j o r w a s t e s t r e a m in this c o u n t r y . A n o t h e r p r o g r a m m is d i r e c t e d at s t a b i l i s a t i o n of wastes, techniques have been d e v e l o p e d , n o w the a i m is the r e a l i s a t i o n of p i l o t p r o j e c t s .
Goumans/Senden/van der Sloot, Editors
Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
757
EVOLUTION OF REGULATIONS AND STANDARDS FOR STABILIZED HAZARDOUS INDUSTRIAL FINAL WASTE MANAGEMENT IN FRANCE Anne-France Didier (SDPD Mmist~re de rEnvironnement) -Jacques Mfhu ( P O L D E N -INSA LYON Ddveloppement) - Valfrie Mayeux (ADEME)
1. M a i n p r i n c i p l e s
and definitions
1.1. D e f i n i t i o n o f H a z a r d o u s I n d u s t r i a l W a s t e s
The regulatory status of H1W is defined in the decree which transposes into French law, the European list of hazardous wastes including the definition of hazardous wastes (mainly based on the 14 criteria in appendix III of the Directive 91/689. The decree establishes the relationship between what were previously called in France "wastes with harmful effects" and then "special industrial wastes". Hazardous wastes therefore include : - Special Industrial Wastes (which are the object of regional disposal plans, and on whose
landfilling a tax is applied which contributes to finance remediation of contaminated sites), - Dangerous Wastes (from household waste), - Health Activity Wastes (with risk of infection). Since the beginning of 1997, France has actively been working on the development of practical modalities of application of hazardous criteria (H1 - H14) to allow extension of the hazardous waste list to new wastes from the European Waste Catalogue. For this, Association R E C O R D (cooperative research network on waste) and the Ministry of Environment have launched joint research programmes concerning the application of hazardous criteria to a wide range of wastes, including stabilized wastes. The organizations in charge of the research are INERIS (H1-H3 "physical hazards"), INRS (H4 -H12 "hazards for human health) and P O L D E N (H14 "environmental hazard"). The methods should be operational at the beginning of 1998. The French Ministry of Environment also intends to publish a text concerning the general classification of all wastes (non-hazardous or municipal like, inert) based on the European Waste Catalogue. Such a comprehensive text would therefore greatly facilitate the work of local authrities and industry on site.
758 1.2. T r e a t m e n t / d i s p o s a l of SIW (Special Industrial Waste) Annually, France produces about 150 million tons of waste including 7 million tons of SIW. Apart from possible valorization scenarios (recycling, reuse, extraction of materials, substitution of materials, energy recovery,...) SIW can either be treated prior to disposal: physico-chemical treatments or incineration (1 558 329 tons in 48 installations (ADEME-1994)), or "disposed of" in a class I landfill, if they can be considered as "final" and "stabilized" (727 696 tons in 13 landfills (ADEME- 1994)).
1.3. N o t i o n of Final waste The law of July 13th 1992 concerning all wastes (hazardous, municipal like, inert) determines that in 2002, only "final" wastes can be admitted for landfill. According to this law, a final waste is "a material which can no longer be treated under present day technical and economic conditions, particularly by extraction of the valorizable fraction or reduction of its hazardous orpollutant character". This notion can of course evolve. Furthermore, as it is not based on quantitative criteria, it is often the object of many debates concerning its application particularly on household waste.
1.4. N o t i o n if Stabilized Final Special Industrial Waste (SF-SIW) The bye-laws (18/12/92) define two categories of Final SIW which must be stabilized prior to landftllmg in class I. From March 30th 1995, for category A (APC residues, steel dusts, used catalysts...) From March 30th1998 for category B (industrial waste water treatment sludges, metallurgy slag (excluding those from salt baths-based process), waste from polluted soil treatment...) The SF-SIW is considered as such if it satisfies a certain number of physico-chemical criteria defined by the bye-laws (18/12/92 and 18/2/94) which concern the immediately leachable fraction obtained from compliance tests (X31-210, X31-211...) in the process of standardization on a European scale (WG2 of C E N / T C 292). Progressively, the notion of long term behaviour of wastes according to the exposure scenario will be introduced in the French regulations for landfill (it is already the case for studies in progress concerning inert waste in class III landfills. It is the object of the French standard X30407, now awaiting standardization on the European level (WG6 of the C E N / T C 292). These new requirements will probably modify the notion of stabilized waste and therefore what can be expected from stabilization process performances. Meanwhile, Final SIW can be considered as stabilized either as they stand, if they meet the requirements of the bye-laws (18/12/92) or after application of a stabilization process.
759
1.5. Stabilization
p r o c e s s e s
Three main techniques are now used or are being developed in France : stabilization using mineral binders (industrial stage), encapsulation in organic binders (pilot stage), - vitrification (under development, first industrial unit in 1997).
-
-
Stabilization using mineral binders represents all the industrial installations to date, either near or on the class I landfill site. 9 of the 13 Final SlW French landfills have a stabilization facility for an annual capacity of about 600 000 tons.
2. E v o l u t i o n of regulations and standards (Stabilized) Industrial Waste (SF-SIW) management in France
Final
Special
The application (first partial then total) of the new regulation concerning Final SIW landfilling has brought to light the radical increase in management costs (stabilization + landfillmg). This has led industry to reconsider waste generating processes (reduction at the source, pretreatment) in order to reduce either the quantity or the polluting character of the wastes (authorizing, for example, admission as they stand). In certain extreme cases, this has led to questioning the process itself, leading to the abandon of the considered waste. As far as management of the "unavoidable" Final SIW is concerned, 3 fields of reflection are now open to industry: 1 Delisting of the stabilized waste (either classified as "M" assimilated to household waste, allowing landfilling in class II sites, or potentially as "I" inert, allowing landfilling in class III sites). 2 Valorization of the stabilized waste, thereby economizing the cost of landfill and even fixing a sale price for the material. 3 Valorization of the waste as it stands, thereby further economizing the cost of stabilization.
2.1. Delisting of (S)F-SIW This alternative was first offered by the regulation within the framework of article 14 of the byelaw 25/01/91 concerning household waste incineration. A project for an application circular of this article stipulating acceptance conditions of stabilized APC from MSW incineration in class II was elaborated and discussed with industry. This perspective was at the origin of new projects for stabilization process development, a number of which had to be reconsidered due to the withdrawal of the circular. Even today, it is clear that certain stabilization techniques are situated in the delistmg perspective. This is the case for processes including an extraction. This position can be justified on the one hand by the fact that the composition (and therefore the potential pollution) of the waste is
760
greatly modified (by extraction of the most soluble fraction and part of the leachable metals) and on the other hand by the fact that, in certain cases, part of this salt fraction can be valorized. This is also the case for vitrification, whose cost alone is much greater than that of stabilization by mineral binders + landfilling. A comprehensive evaluation procedure for vitrification processes is now in the validation phase. Its application could allow delistmg or even banalization of vitrified materials under certain conditions and respecting certain thresholds which remain to be defined. This is apparently not the case for processes using mineral binders which are at present used in waste stabilization for class I landfilling. It must be noted here that as stabilized wastes are not considered hazardous in the European list, the French regulations which stipulate their landffllmg as hazardous waste (according to the typology of the Landfill Directive draft) could be reviewed on a case by case basis, by considering the technical and scientific aspects, in particular for wastes having undergone an extraction or a reduction of the polluting potential.
2.2. Valorization of (S)F-SIW The possibility of valorizing certain stabilized wastes in Civil Engineering is being considered mainly for vitrified materials. Association R E C O R D has funded a study on the general approach concerning banalization of materials elaborated from waste, leading to the definition of technical specifications (use criteria) and environmental specifications (present and future) to be taken into account. A certain number of conclusions can be drawn from this study: - I t is practically excluded that civil engineers envisage the use of monolithic materials elaborated by stabilization processes (blocks of solidified waste for example), -
To have a chance of being valorized, a waste must be able to substitute a material already used (sand, granulate, filler) and must respect the technical specifications like granulometry, mechanical strength and reactivity towards the other componants (e.g. problem of compatibility between hydraulic binders and vitreous matrixes, which has led several vitriflers to orient their technique towards slow cooling which enhances crystallization),
-Furthermore, the cost must be lower than the material to be substituted and the quantitative and geographical availability must be at least equivalent, - Environmental specifications are cruelly lacking on the French level. The development of the standard X30-407 should compensate for this in the long run (see later), -Valorization which consists of simply diluting the waste in a construction, the waste not participating by its properties in the specifications of the construction, is to be excluded, - T h e simple fact that a waste has cost too much to be stabilized for landffllmg does not confer any value to it as regards valorization in civil engineering. Valorization of stabilized waste is therefore to be studied on a case by case basis according to demand, technical and environmental specifications in a given scenario.
761
2.3. Valorization of waste as they stand This is a real problem at the moment for wastes which do not follow the normal disposal routes but are stored internally on site (in-situ landfill, mmmg waste, metallurgy slag heaps). In most cases, a strict respect of the regulations (stabilization + landfilling) is not economically compatible with the considered industrial activity. Maintaining the activity (and the associated employment) implies research for acceptable solutions (including for the environment). It may entail modification of the site and landfilling conditions on site or the valorization scenario. A certain number of these wastes are already valorized in civil engineering locally. This is sometimes carried out "unofficially" which does not necessarily mean it is a clandestine activity, but rather on the basis of a temporary authorization while waiting for further data allowing delivery of an official authorization. To emerge from this unsatisfactory situation, the metallurgy industry has launched research programmes with the aid of the European Commission concerning feasibility of valorization of different types of metallurgy slag in civil engineering. The most promising scenarios are in road construction. For such wastes; local use is a crucial issue given the cost of transport. As far as valorization conditions in civil engineering are concerned, the limits are the same as those described in 2.2. The main problem is still the environmental specifications to be defined according to the scenario. To compensate for this lack of specifications, the practically systematic reference of most promotors in civil engineering for the use of waste is the circular of May 9th 1994 concerning MSW bottom ash. The Ministry of Environment states clearly that the regulatory conditions proposed here do in no way allow appreciation of the behavioural parameters nor prediction in the long term. Even if the economical and social context for MSW Incineration has needed a short term regulatory text, its extension to all types of material, for which landfill wants to be avoided, is abusive and not relevant. The Ministry of Environment and the Ministry of Civil Engineering are preparing a charter to propose a joint evaluation procedure. At first, this charter would apply to non hazardous waste (particularly MSW incineration bottom ash) and then would perhaps be extended to certain (S)FSIW according to their long term behaviour and in a well-defined scenario.
3. E v o l u t i o n
of environmental specifications
The future of stabilization and its development towards other wastes or other scenarios is direcdy related to the evolution of the environmental specifications. These will be carried out on two complementary and successive levels : the verification of stabilization performances and the long term behaviour of the obtained material in the specified scenarios.
3.1. Performances The arrival of a new regulatory context with strict constraints concerning the waste streams (theoretically covering all the F-SlW) including new wastes, relatively difficult to stabilize but for which a budget exists (APC from MSW incineration) has led to the development of more ambitious and more effective processes.
762 The (logical) drawback is that, taking into account the considerable extra cost of application of these techniques and their practically statutory situation on class I landfill site entrance for those managed by site managers, the demands of the authorities and producer industries are justifiably greater. For this situation, the A D E M E has initiated studies to develop "Comprehensive Evaluation Procedures of S/S processes - (CEP)", specific to the techniques used (mineral binders (POLDEN), vitrification (CEA), polymers (LNE), bitumen (forthcoming). Theae comprehensive procedures, whose application would be a condition for obtaining funding for the industrial scale processes, are being implemented for mineral binders, and are in the final stages for the others. The objective of these CEP is to establish the nature of pollutant retention by the technique considered (encapsulation, micro-encapsulation, inclusion in the porous structure, integration in the phases of the matrix, change in mmerology), the mechanisms of leaching behaviour (washoff, congruent dissolution, coupling of dissolution and diffusion, shrinking front,...) its quality (resistance to physical, chemical and biological attacks) and its durability (resistance to ageing under constraints). Even if the objective here is process-oriented instead of material-oriented, therefore situated before the behavioural evaluation of material in specific, landfill or valorization scenarios, a certain number of experimental procedures will be the same.
3.2. L o n g term behaviour The draft Landfill Directive mentioned 3 levels of waste evaluation : level I : basic characterization, long term leaching behaviour level 2 : compliance tests, verification of long term behaviour parameters level 3 : rapid tests - control on site. Level 2 corresponds to compliance tests which exist in most European countries for landfill acceptance (X31-210, X31-211 in France). Working Group 2 of CEN/TC292 is in charge of standardization of these tests on the European level. The stage of enquiry is now completed for the first test concerning fragmented waste. Working Group 6 (whose French mirror is the AFNOR commission X30Y) is in charge of level 1, which is of strategic importance as it supposedly conditions the other levels. The field of application of these studies covers much more than just the problem of (S)F-SIW landfflling. It also concerns other types of disposal (the project in progress as regards regulations for landfilling of different types of inert wastes integrates this notion), valorization in civil engineering (bottom ash in road construction for example) and in a general manner any evaluation of pollutant release from a defined source in the environment under specified conditions (mechanical, geotechnical, climatic, biological, site use, risk factors...) at a given time scale.
763
To prepare these evolutions, France has elaborated a methodological standard (X30-407), whose outlines, after amendment, were adopted by the participants in the European workshop. Standardization of the tests to be applied to given couples waste (or materials)/scenarios, is in progress. One of the first priorities that the Ministry of Environment has fixed at the X30Y commission, concerns acceptance of stabilized waste in class I landfzll. Logically, regulations should integrate these notions in the more or less long term, and propose rules to be respected which will probably be in the form of thresholds, but corresponding to the nature and to the objective of the tests carried out for this evaluation of long term leaching behaviour. For example, it is probable that notions of pollutant availablity according to the chemical context (pH in particular) and pollutant flux per unit surface for monolithic waste will be introduced. We think that this will have three types of consequence on stabilization processes : - A certain selection, which will favour the best processes, i.e. those for which a real stabilization (control of pollutant flux) will be proposed ; - A selection within the different waste sources which can be potentially stabilized towards those which can reach a real stable state. For the others (bio-evolving, oxidizable, soluble and not retained...) pretreatments may be deemed necessary. We could therefore see S/S processes evolve towards waste streams which, due to their physico-chemical nature, logically correspond to them. We would therefore change from (( a market to conquer )) logic to a "treatment logic". The so-called secure landfill could then play its role to the full : protecting "unstable" stabilized waste from external attack. ; -Justified and supported opening up of other horizons to stabilized wastes based on appropriate measures and models ; stabilization of contaminated sites, old stocks or slag heaps, "inert" wastes in class III,...Here, the waste streams, techniques and especially the environmental specifications remain to be defined. The X30-407 (or its european equivalent) will not be sufficient. The evaluation of ecocompatibility i.e. the cross checking of pollutant flux emitted and the acceptable flux in the environment will be necessary. It will consist of defining adapted scenarios (geotechnical, hydrodynamic, bio-physico-chemical,..) for stabilized wastes, i.e. generating flux considered as being in exchange equilibrium with the environment. INSA Lyon is organizing an international congress in September 1998 : " W A S T E STABILIZATION AND ENVIRONMENT" which aims to demonstrate the necessary complementarity of the three consecutive stages of evaluation : - control of pollutant emission from stabilized wastes according to their intrinsic character and scenario conditions, - transport and evolution of pollutants from these wastes in the environment, - i m p a c t of these pollutants on man and the environment. The final objective is of course to provide information to the regulatory and industrial organizations responsible for stabilized waste management.
764
Main reference works
9
des d6chets : p r o g r a m m e de recherche pluriannuel de I'ADEME, coordonn6 par P O L D E N - W G 6 du CENTC 292 -Groupe de travail "Mise en d6charge des d6chets mertes" Mimst~re de l'Environnement, POLDEN -Etudes d'application des crit~res de danger de la Directive 91/689 : Mmist~re de l'Environnement, Association RECORD, INERIS, INRS, P O L D E N - Groupe de travail, valorisation des d6chets en BTP : Mimst~re de l'Environnement - PEAs de PSS (liants min6raux, vitrification, polym~res) : A D E M E , P O L D E N , CEA, LNE - Stabilisation des DIS, Etude prospective : ADEME, A L G O E management, P O L D E N -Eco-compatibilit6
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
765
T e s t m e t h o d s and criteria for t h e a s s e s s m e n t of i m m o b i l i z e d w a s t e Dr.lr. G.J.L. van der Wegen Intron, institute for materials and environmental research B.V. P.O. Box 5187, 6130 PD SlTTARD, The Netherlands on behalf of Centre for Civil Engineering Research and Codes (CUR) P.O. Box 420, 2800 AK GOUDA, The Netherlands Abstract The Dutch environmental policy is aimed at a drastic reduction of its wastes as well as the minimalisation of the use of natural raw materials. It focuses on prevention and re-use of wastes. Immobilization of (hazardous) wastes can contribute to this policy either by producing building materials from wastes or, if the former is not achievable, by producing less leachable wastes for landfilling. A reliable set of test methods and related criteria is indispensable to apply these techniques in an environmentally safe and justified manner. Moreover, these test methods are also an expedient to further development of immobilization techniques as well as the quality assurance during full-scale application of these techniques.
Introduction The Dutch environmental policy is aimed at a drastic reduction of its wastes as well as the minimalisation of the use of natural raw materials. It focuses on prevention and re-use of wastes. Immobilization of (hazardous) wastes can contribute to this policy either by producing building materials from wastes or, if the former is not achievable, by producing less leachable wastes for landfilling. It is important to know the effect of immobilization techniques on the leaching behaviour as well as on the durability of immobilized wastes. Test methods and related criteria to assess the characteristics of immobilized wastes have been selected and/or developed within the frame-work of CUR-committee D24. These instruments for assessment are indispensable for the development of the immobilization techniques, for proving their fitness for purpose and for quality assurance during full-scale application of these techniques.
Definition and aim of immobilization Immobilization is defined as a technological treatment in which the physical and/or chemical properties of a waste are changed in such a way that the spreading of the pollutants by leaching or erosion is reduced adequately on the short as well as on the long term. Immobilization is a treatment on a particle size level. Storing waste in big bags or steel drums will prevent spreading of the pollutants, at least on the short term, but is not considered as immobilization. Immobilization has two aims: 1 Treatment of (hazardous) waste in such a way that it can be applied as a building material.
766 Producing less leachable wastes which can be landfilled under less severe isolation measures for soil protection (i.e. changing a hazardous waste into a socalled controlled leaching waste). The first aim, application as a building material, is preferred because of much lower costs of removal of the waste, less trespassing on capacity of landfilling and the reduction in use of natural raw materials. However, if this first aim cannot be achieved the second aim is still of importance. The costs of removal will still be lower (although not as much as for the use as a building material) and the very limited capacity of landfilling hazardous wastes in The Netherlands will be spared.
Regulations In The Netherlands, the Building Materials Decree sets regulations under which building materials may be applied in or on the soil or in surface water in an environmentally safe and justified manner. This decree classifies building materials according to their content of organic pollutants and to their leaching of anorganic pollutants. Depending on their classification different regulations are applied [1]. Test methods for the determination of the leaching behaviour of anorganic compounds are the socalled diffusion test for monolithic materials (Dutch Standard NEN 7345) and the socalled column test for granular materials (Dutch Standard NEN 7343), which is a percolation test. For organic compounds in building materials no suitable leaching test have yet been developed. Therefore, composition values for organic pollutants are determined instead of immission values, which undervaluates most immobilization techniques. The Building Materials Decree requires that the building materials possesses an adequate durability. This property is not fully translated into test methods and criteria. If a waste cannot be upgraded to a building material, in most cases it has to be dumped, according to the Dutch regulations set forth in the Disposal of Solid Waste Materials Decree. This decree classifies waste materials according to their leaching behaviour. The test method used for this classification is the above-mentioned column test, however with a lower liquid/solid ratio (1 instead of 10). The material to be investigated in the column test must be crushed to a particle size less than 4 mm according to NEN 7343. Crushed monolithic materials will therefore result in a substantial higher leaching rate than in practice because of the substantial higher specific surface due to the crushing. Immobilization techniques based on monolithic products are therefore undervaluated. Some immobilization techniques require binders and additives. They contribute to an increase of mass and often in an increase of volume compared to the untreated waste. In the case of landfilling this is an undesired side-effect of the treatment. Dutch regulations require that such an increase in volume must be less than 25%. Durability of immobilized waste to be landfilled is less essential than in case of use as a building material. Usually, the waste is covered by other waste or a layer of sand within a few days or weeks preventing further influence of atmospheric conditions.
Additional test methods In addition to the Dutch regulations described above, CUR-committee D24 has developed test methods and criteria to fully assess immobilization techniques for hazardous wastes. With respect to environmental aspects the following properties of the immobilized wastes have
767 been considered: Water soluble salts At higher contents this might lead to (partial) solution of the matrix of the immobilized waste. This will influence the leaching behaviour as well as the durability in an adverse way. Digestible organic matter This might lead to enhanced leaching due to adsorption and complexation reactions with pollutants. Redox-conditions Under reducing conditions leaching of some heavy metals is substantially less than in an oxidized state (e.g. lead and copper in the presence of sulphide compared to sulphate). The reducing capacity of the immobilized waste is important with respect to the leaching behaviour on the long term. Efficiency of immobilization In order to assess the merit of immobilization techniques or to compare different techniques a diffusion test similar to NEN 7345 has been developed for untreated waste. Comparison of the leaching behaviour of the untreated and treated waste will give the environmental gain of the immobilization techniques. The test methods developed for the above-mentioned properties together with those postulated in Dutch regulations are summarized in table 1. The durability of immobilized wastes depends on its own structure as well as the climatic conditions it will be exposed to. In The Netherlands the most relevant mechanisms of degradation are: Wet/dry Under alternating wet/dry conditions cohesion failure might occur due to recrystallization of salts in and/or shrinkage of the immobilized waste.
solution
and
Freeze/thaw Above a critical degree of saturation with water every material is susceptible to freeze/thaw conditions. This critical degree of saturation depends on the pore structure and strength of the material. Erosion In order to prevent spreading of pollutants by wind erosion, the immobilized waste should have a certain resistance against this mechanism of degradation. Oxidation A combined interaction of oxygen, moisture and solar radiation (activation energy) can degradate organic binders usually leading to a brittle and 'dusty' surface appearance. Test methods for these four durability properties are shown in table 2.
768
Conclusions Immobilization of hazardous waste plays an important role in the Dutch environmental policy. It can contribute to the reduction of wastes to be landfilled, to an increasing re-use of wastes thereby saving natural raw materials. A reliable set of test methods and related criteria is indispensable to apply these techniques in an environmentally safe and justified manner. Moreover, these test methods are also an expedient to further development of immobilization techniques as well as the quality assurance during full-scale application of these techniques.
References
Hendriks, Ch.F. and Raad, J.S .... Principles and background of the Building Materials Decree in the Netherlands", Materials and Structures, vol. 30, January-February 1997, pp. 3-10. CUR-report no. 183, ,,Guideline for the assessment of immobilized wastes", 1995 (in Dutch, with English summary).
Table 1. 'Environmental' test methods
I
property
description of test method
standard
leaching
diffusion test (monolithic) column test (granular) = percolation test 'granular' diffusion test redox-potential/reducing capacity soluble salts at liquid/solid ratio of 10 extraction with 0.1N NaOH
NEN 7345 NEN 7343 NVN" 7347 NVN" 7348
efficiency redox soluble salts organic matter NVN = pre-NEN
Table 2. Durability test methods property
description of test method
wet/dry resistance
6 wet/dry cycles: 5 hours in water of 20~ + 42 hours drying at 70~ monolithic: critical degree if saturation granular: 20 freeze/thaw cycles sand blasting test, originally designed for wear resistance Xenon-test chamber
freeze/thaw resistance erosion resistance oxidation resistance
standard modified ASTM D559 RILEM 4CDC3 NEN 5924 modified NEN 2875 ISO 4892
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All fights reserved.
769
Inorganic immobilisation of waste materials F.Felix, A.L.A. Fraaij and Ch.F. Hendriks Delft University of Technology, Faculty of Civil Engineering ABSTRACT Inorganic immobilisation uses cement to solidify certain waste materials. Water is added to the cement in a ratio between 0.4 and 0.6. The calcium-silicates of the cement react with the water to a calcium-silicate-hydrate-gel (CSH-gel), lime and heat. This gel is able to cling aggregates together. The result is a hardened product, including some pores. Pores larger than 20 nm are called capillary pores, smaller than 20 nm are called gel-pores. The concrete is characterised by a high pH about 13. The effectiveness of immobilisation techniques is a relation of immobilisation and leaching mechanisms, degradation mechanisms, economic and environmental merit and application potentialities. Because of the high pH most heavy metals show good immobilisation DroDerties, except molybdene and chromium (VI). Other components of the waste material are hard to immobilise, such as anions and organic compounds. Therefore additives may be used. Oxidants can start a REDOX-reaction with the anions and organic compounds are used to bind the organic components. Once the product is put into practice, the effectiveness of the technique depends on leaching mechanisms: surface release, diffusion and dissolving. Especially diffusion is important for immobilisation. Diffusion is caused by the presence of water in the pores of the material and concentration variations between the material and the medium. Degradation mechanisms, important for inorganic immobilisation, are erosion, wet/dry-periods, temperature changes and freeze/thaw-periods [25]. These influences can cause the formation of cracks. Economic merit is not often given. It can be divided in benefits and costs of the used and produced materials, disposal, transport, labour, energy, aftercare and investments. Like economic merit ,the environmental merit is not simply given. An important factor is the priority list of Lansink. The benefit of the technique is the replacement of primary materials by secondary materials. Furthermore other aspects have to be considered, such as the use of energy, the release of components and the use of materials. The effectiveness of immobilisation is last but not least dependent on the aDDliCation potentialities. The application of immobilised products, both the disposal and useful application, needs after-care. This means, control of the product is required. Furthermore, the effectiveness of inorganic immobilisation depends on the used waste material. Some material have more opportunities than others. ~Desired properties of the waste material are: high content of heavy metals, low content of anions and organic compounds. Moreover, it is desirable to minimise the pretreatment of the waste. Currently used waste materials for inorganic immobilisation are: filter dust, coal ashes, MIP-fly ash, residues from burning of coal, sewage sludge, lime sludges, fluorgypsum waste, arsenical waste, slag and dust from steel making. It can be concluded that inorganic immobilisation can show good encapsulation properties for certain waste types. Especially these waste types, containing heavy metals. Degradation of the product is mostly caused by the formation of cracks. Therefore, after care of the product is required.
INTRODUCTION Inorganic immobilisation is based on cement as solidifying agent. In addition of water cement hydrates to a cement gel, that is able to cling aggregates together. The result is a hardened product, which can chemical and physical encapsulate certain hazardous waste
770 materials. This technique is widely developed in several countries, such as the United States of America and Japan. However, in the Netherlands, little applications of this technique are known due to legislative restrictions in the past. Current changes in legislation are helping to put developed processes into practice and thereby requiring a research to the advantages and disadvantages of inorganic immobilisation techniques. The scope of this article is a 'state of the art' of immobilisation based on cement. The scope of paragraph 1 are properties of concrete relevant to inorganic immobilisation. In paragraph 2 the effectiveness of inorganic immobilisation techniques is given. This paragraph contains encapsulation mechanisms, degradation mechanisms, application potentialities, costs and environmental aspects. Paragraph 3 deals with required properties for the waste materials. Chemical and physical properties of the waste material will be related to the effectiveness of the immobilisation. The last paragraph presents the advantages and disadvantages of inorganic immobilisation.
1 RELEVANT PROPERTIES OF CONCRETE
Composition
Concrete is made up of cement, aggregate, water, air and additives. Currently used cement types in the Netherlands are: Portland cement, blast furnace slag cement and Portland fly ash cement. The clinker consists of the following minerals: 3CaO.AIO 3 , 3CaO.SiO2 and 2CaO.SiO2. In the presence of water, CaSO4.2H20 (gypsum) and/or CaSO4, 3CaO.AIO3 reacts to CaO.AI203.3CaSO,.32H20 (ettringite). This ettringite forms a layer on the cement aggregates and thereby delays the hydration reaction of cement. Water is added to the cement to react with the calcium-silicates and form a calciumsilicate-hydrate-gel (CSH-gel) and lime. An example of the reaction of 3CaO.SiO 2 with water is: 2 (3CaO.SiO2)
+
6 H20
->
3 CaO.2SiO2.3H20
+
3 Ca(OH)2
During the reaction, the content of unhydrated cement decreases and the content of CSHgel increases, which leads to a decrease in volume. The second objective of the addition of water is the increase in the workability of the concrete. Water surrounds the cement particles and aggregates. In the final product, water can be present in the concrete under several conditions. In the capillary pores 1 water can be present as water vapour. Above the saturation pressure condensation will occur. The saturation pressure increases with decreasing pore size. Some water is not physically bound to the surface of the solid material. This water is called 'free water'. Free water exists in the larger gel pores and the capillary pores due to condensation. In addition, water is often adsorbed to solid materials in the smaller gel pores. Water can also form a layer between components of the CSH-gel. Aaareaates, such as river gravel and river sand are added to the cement paste. Important factors of the aggregates are the weight, shape and particle size distribution. In some cases, additives are used to improve the workability of the concrete or to decrease the Water/Cement-ratio. After casting of the concrete, the concrete will be vibrated for proper compaction. Air decreases the strength of the concrete, by forming pores in the concrete. However, in special cases air is consciously brought into the cement gel by means of air entraining agents. Concrete thereby becomes more resistant to freeze/thaw(salt)-periods, due to these pores.
1An explanation of capillary and gel pores will be given later in this paragraph.
771
Strength
Strength develops in time and continues to increase for years. This process will be accompanied by a gradual change in pore size distribution.
Pores
Pores in the CSH-gel take approximately 25% of the total volume of CSH-gel, the gel pores. They are defined as pores smaller than 20 nm. Pores larger than 20 nm are called capillary pores. During the hydration reaction the capillary pores diminish, due to the growth of cement hydrates. An increase in capillary pores is responsible for a decrease in strength and durability, as well as for an increase in permeability of the final product. However, capillary pores will always be present in cement. The pores are due to the presence of water.
W/C-factor
An important factor in the properties of cement is the Water/Cement-ratio. Especially, the strength depends on the W/C-factor. A higher W/C-factor includes a lower strength. The reason for this is that cement of W/C-factor above 0.38 will always consists of capillary pores. Besides that, too little water will make mixing difficult [17]. A typical W/C-factor lies between 0.4 and 0.6.
Hydration heat
The hydration reactions are exothermic and thereby the setting of cement is accompanied by the formation of heat. Heat is transported from the centre of the matrix to the surface. This results in temperature gradients, which can cause cracks in the case of bulk concrete.
pH
The pH of the cement product is high, approximately 13. The pH depends on the cement type. Furthermore, the development of OH-concentration in the pore water is a function of the W/C-factor [8]. A high W/C-factor results in a lower pH, as shown in the following figure. mg O H ' / l i t e r 11000.
L 9o00 [ 8800
,
, ,
~-
Cement p.c.-A wcf.: x OJ.O
/
oo.s6
II
7700 I6600
,
,~O.L5
. . . .
11111 ilili11~.3.
!1111 Ilill
11 ilIF~
pH
! !
l~~]J~ll.3 66 l 1LLLII
ssoo| I IIIIIIII I I !11~/~rllll.3s. .oo I I II111111 I I l~Z.ii/i i 1111111 I i i111111.32g 33ooi --I I l..llllll:.iI b ~1~ -1i?111 I I I111111 22001 i ! III~_~ IIII!! ! I l llllll.zs. .oo I ~ I - I oi, I l lllllll ,I I ii1iii ! !11 ii11 I
1o
--
1oo
1ooo
- Time (hours|
Figure 1: development of the OH-concentration in the pore water as a function of the W/Cratio
772 2 EFFECTIVENESS OF INORGANIC IMMOBILISATION The effectiveness of immobilisation can be measured by the encapsulation of components of the waste material. The more components are encapsulated, the more effective the technique is. Therefore, immobilisation and leaching mechanisms will be described. Furthermore, the environmental and economic merit and application potentialities contribute to the effectiveness of inorganic immobilisation. Immobilisation mechanisms The effectiveness of encapsulation depends on the immobilisation mechanisms during and after the cement production. Firstly, the physical mechanisms will be considered. Secondly, the chemical mechanisms are described and finally the use of special additives is related to the desired encapsulation of components.
Physical mechanisms can be divided in hardening of the concrete and porosity of the concrete. Fine particles (less than 74/Jm) weaken the bounding between waste particles and the cement by coating the larger particles [17]. This coating inhibits chemical binding of the contaminants. Pretreatment of the waste may be required to reduce the fine particle concentration. As said before, pores are always present in the final product. The leachability of components is increased by the increase of the porosity. Furthermore, a high content of soluble salt in the matrix may lead to an increase of the porosity of the matrix, due to the leaching out of these contaminants [6]. Chemical mechanisms of immobilisation of the waste are due to the chemical attraction of the surface atoms and of the precipitation of hydroxides. Most cations, such as heavy metals are insoluble at pH 8-10. They are immobilised by the precipitation of metalhydroxides. Above and below this pH range the solubility increases. Figure 2 shows the solubility of cations in relation to the pH. At the high pH of concrete, most cations are not in its least soluble state. Ions, such as molybdene and chromium (VI) are mobile at high pH, due to a low chemical retention, see figure 4. Although chromium and molybdene show high leachability, some heavy metals can be bound to the CSH-gel by adsorption and chemisorption. Metal ions may be incorporated into the crystalline structure of the cement [17]. Anions are difficult to immobilise, f.e. cyanide, chloride and bromide. They do not precipitate with hydroxide and often are not bound into the crystalline matrix [22]. Exceptions of anions that can be immobilised are sulfates and sulfites. Organic compounds may delay the setting time. Besides that, organic compounds show less fixation in the matrix [22]. Moreover, organic compounds can increase the mobility of other compounds [6]. Organic compounds may be decomposed at high pH. This will lead to soluble organic compounds. Trace elements, both organic and inorganic, can be bound to the organic compounds and thereby leave the matrix [6]. Volatile organic compounds, but also inorganic volatile compounds such as mercury, can release the matrix during mixing. Oil and grease cause the coating of waste particles and thereby weaken the bond between particle and cement [17]. Pretreatment may be required. Additives Because of the described immobilisation mechanisms, some cations, anions and organic compounds are hard to immobilise. Therefore, additives are used in order to optimise the immobilisation. In the following, the common additives are given and related to the desired encapsulation mechanism. In order to optimise the hydration reaction, pozzolanic materials, such as blast furnace slag and cement-kiln dust are added. Pozzolanic materials need an additional calcium source
773 100 -
,,,
ii~
.
.
.
.
.
.
.
0.001
~oo1!
0,0001
6
7
8
9
10
11
12
pit
Figure 2" typical solubility curve for metals over a range of pH in order to start the hydration reaction. Therefore and for the formation of ettringite, an external calcium source is required, like CaCO 3, CaCI2 or CaS04. NaOH and MgOH may be used for to optimise the hydration reaction. Ohter additives are useful for the immobilisation of inorganic compounds: soluble silicates, reducing agents, oxidants and clay. Soluble silicates, such as silica gel (Na2Si4.xH20), can chemically bind inorganic contaminants. However, a common disadvantage of these additives is an increase of volume [17]. Reducing and oxidising agents, such as Fe2§ Fe3+, Mn 4§ S042 are added in order to precipitate salts [14]. A REDOX-reaction will lead to less soluble salts and thereby diminish the leaching out of salts through the pores. An example is the reduction of ferric (11) to ferric (111)and the oxidation of chromium (VI) to chromium (111). ~P~L Leiil-OrlliLIl-I-I
IIIll IIIl/
:-I;~
I Iq~o
I..LI-;~uiUUi11:;
thc:lll
ch'---:"--iuiiiiuiii
il~llvi ) .
Pl.~..~..,ir
:'~h.~ r
4...~(q./ i[;;ll~.}q.ll'i~a"P~rkl~ i . . . .
~u,
. ~i .p . . . .
pounds and excessive free water. It is also used as absorbent of organic compounds. Organic compounds can be JmmobJlJsed in the cementious matrix by addition of certain absorbents, such as the previous mentioned clay and active coal [10,18]. By means of clay, organic compounds are absorbed in the matrix by non-ionic forces. Soluble organic compounds with carboxyl and hydroxyl groups, can be fixed at the Ca2+-ions. They thereby become insoluble. Organic, particularly aromatic, substances can be JmmobJlJsed by absorbJng on a layered clay mineral. The clay is therefore modified with alkyl ammonium, which increases the adsorption surface. The surface has no longer hydrophJlic but organophJlJc conditions. After adsorption of the organic compounds, the clay is mixed with a hardenable inorganic binder, f.e. Ca(OH) 2 or Ca2+-contaJnJng cement. Other techniques consists of the addition of an absorbent, others than clay [3,16]. A possibility is mixing the waste with a binder and then granulate. The granulates are contacted with solvent to extract hydrocarbons. Or the organic waste may be dispersed in water containing a cationic ammJne as an emulsifier. Subsequently, the emulsion is mixed with cement. The last technique for organic waste materials is to saturate the waste material with water and mixing it with an inorganic hydraulic binder [21 ]. Because of the high pH some organic compounds will vaporJse.
774
Leaching mechanisms
Once the immobilised product is put into practice, the isolation of contaminants from the environment can not totally be guaranteed. The mechanisms that contribute to the release of these contaminants are caused by the presence of water in the pores or on the surface of the product, due to the following mechanisms: surface release, dissolving and diffusion. Figure 3 shows these three mechanisms. ImmoblllsedproduGt I
~
Water Dissolving
--~ .......
9 o.,u.,..
.....
....
I Goomot,i(:alurfgioo
Figure 3: leaching mechanisms Surface release is characterised by a short leaching out of contaminants in the first period of the application of the product. High soluble components at the surface of the product dissolve into the surrounded media. However, this leaching mechanism is important in organic immobilisation techniques and vitrification, inorganic immobilisation shows little surface release. Over a longer period, the rate of dissolving of contaminants is almost constant. An example is the release of Ca 2. from stabilised gypsum. However, this mechanism does not contribute much to the total leaching out of contaminants. The most important mechanism is diffusion [6]. Diffusion is caused by the presence of water in the pores of the material and concentration variations between the material and the medium. As shown in the following formula, the diffusion capacity can be related to the tortuosity and the chemical retention.
L= L f Do R T
f2. Do R.T
flux (m2/s) availability (-) average diffusion coefficient (m2/s) chemical retention (-) tortuosity (-)
The tortuosity of an immobilised product indicates the path length of the ion through the pores [6]. A low tortuosity means a small diffusion path and thereby a high leaching rate of the component. Some immobilised products show low tortuosity compared to most building materials, such as concrete and bricks [6]. While tortuosity is a measure for the physical resistance of a component to diffusion, chemical retention is a measure of the hindrance of diffusion due to chemical interactions. Chemical retention is highly correlated to the pH in the pores. A typical relation between pH
775
and some metals is shown in figure 4. 10'
,~. ~o _~ *
,, ....
,,**o
10'
~.~....-~"--"
_...,t 10"
~
/
~
~
Z,
........
.o
F
.. ....... ..'" . . . . " " " '
........................................................ ...
10" 10"
". . . . . . . . . . .
ee . I . ~ ,~"
~, rr
, o ~ 1 7 6. . . . . . . ,'
I0"
Me
7
-
9-
-
, 8
. . . .
, 9
.....
, "'" 10
"
"
, " 11
"
"'"
, 12
pH
Figure 4: relation between chemical retention of certain elements and pH from the environment
Leaching test
To test the leaching behaviour of immobilised products, several test methods are suggested. Dutch legislation concerning building materials and landfill prescribes column tests, diffusion tests and availability tests. The first step in the column test is to fractionise the product to particles smaller than 4 mm and to put them in a column [11]. Acid water (pH =4) is lead from the bottom to the top of the column. After certain periods, the liquid is tapped of. These periods correspond to specific water/material ratios (L/S= 0.1, 0.2, 0.5, 1.0, 2.0,5.0 and 10.0). This column test gives insight into the leaching behaviour of a fractionised waste over a short period (5 year) and moderate period ( < 5 0 years). Another test, the cascadetest, gives information about the long term leachability. A column, water/material ratio of 100 kg/kg, is therefore shacked. The disadvantages of these tests is that the material has to be fractionised, while the aim of immobilisation is to encapsulate hazardous components in a compact matrix. The second test, diffusion test, is specific for materials larger than 40 mm, like monoliths of immobilised waste. The product is surrounded by water of pH =4. At certain time intervals, corresponding to L/S = 0.25, 1, 2.25, 4, 9, 16, 36, 64, the water is tapped of and purified. The concentrations of the percolates are determined. The results of the tests give insight into the contribution of diffusion to the total release of components. The aim of the third test, the availability test, is insight into the leaching of inorganic components of a material under extreme circumstances [24]. The material is first pulvuris~d until 95% of the material is smaller than 125pm. It is presumed that diffusion can take place through the whole material. During 3 hours, the material is washed with water (pH = 7) and after that washed with acid water (pH =4). The ratio between water and material is each time 50. The composition of the percolate is determined and translated to the amount of leached components. These test methods are not capable of measuring the leaching behaviour of organic compounds. Therefore, Dutch legislation is based on the organic composition of an immobilised product. However, these composition demands in Dutch legislation contradict the aim of immobilisation. Not the composition of the product but the leaching behaviour is a measure of the effectiveness of immobilisation.
Degradation mechanisms
Besides leaching mechanisms, other degradation mechanisms may effect the immobilised product. This gives an indication of the durability of the product. Durability is defined as the
776 resistance to chemical, physical and mechanical influences. A product is durable when no intolerable decrease of relevant properties is caused by the influences. Figure 5 shows these external influences. Inorganic immobilisation products are mostly sensitive to erosion, wet/dry-periods, temperature changes and freeze/thaw-periods [25]. Erosion is defined as the degradation of material due to moving media, such as wind, rain and rivers [5]. The wind contains fine particles, that deteriorate the surface [6]. Also, human activities can effect the surface, by walking and driving. The erosion sensitivity depends on the strength level of the product. Products with high strength have high resistance to erosion. Wet/dry p~riods are characterised by the transportation of water between the environment and the immobilised product [6]. This may lead to (re)crystallisation. In addition, wet/dry-periods causes shrinkage or swelling of the material. Shrinkage is due to the release of water from the pores and the release of adsorbed water. Because of increased surface tensions, due to the release of water, the surfaces attract each other. This leads to shrinkage. The extent of shrinkage highly depends on the W/C-factor and the cement contents. Shrinkage is from major importance for products with high amounts of clay. This can cause the formation of cracks on the surface of the product. The opposite mechanism appears in the presence of water, the wetting periods. Although this mechanism is much faster than shrinkage, shrinkage is the major cause for the formation of cracks during wet/dry-periods.
/ o
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Figure 5: degradation mechanisms
Temperatur~ ~hanges has been mentioned before in relation to the formation of heat during hydration. The same argumentation can be followed here. Temperature changes of the environment causes temperature gradients in the immobilised material. This can lead to the formation of cracks. Freeze/thaw-periods can also contribute to micro- and macro-cracks formation. Due to freezing periods, the water in the pores can be frozen. Because of the increase of volume, the pores will expand. Water that is not frozen is pressed into smaller pores and causes press-tensions. Another mechanism is the crystallisation of the water vapour in the concrete
777 to form ice. Due to the lower vapour tension of ice in comparison to water, water vapour moves to the ice. This leads to the growth of the ice in the larger pores and the decrease of water in the smaller pores. This mechanism also leads to the formation of cracks in the material. A specific case of freeze/thaw periods is the presence of thawing salt. Thawing salts lead to the decrease of the freezing point of water, depended on the concentration of salts. In the centre of the material the concentration of salts is lower than on the surface. Temperature gradients and these concentration gradients may causes the surface and the centre to freeze. The isolated water, between the ice zones, presses on the surface, which will lead to degradation of the surface [2]. Application po ten tiafities Before a useful application can be chosen, the product has to be tested on legislation concerning building materials. The aim of legislation concerning building materials is to promote the recycling of secondary building materials combined with the aim to control the leaching out of hazardous components. In order to determine the suitability of building materials, the leaching behaviour of the inorganic components and the organic composition of the material must be determined. If application is allowed, the destination of the product partly depends on the strength of the product. Compared to 'traditional' concrete, compressive strengths of immobilised products have decreased by a factor 10 [25]. The immobilised products may be used in f.e. sound walls. However, most of the immobilised products are not allowed to be used as building materials. In that case, legislation concerning the disposal of waste becomes active. This legislation is also based on the leaching behaviour of inorganic components in the waste. C2-waste: waste that is very harmful. At this moment, there is one C2-1andfill and in the near future no new disposal will be built. Some C2-waste materials are or will be prohibited to be disposed of, such as fly-ash from Municipal Incineration Plants (MIP). With permission of the specific foreign country, the waste may be exported. C3-waste: waste that has high leachability of inorganic harmful components. The waste must be isolated from the soil. C4-waste: this waste has moderate leachability of inorganic components. Some adaptations are needed to minimise the leaching out of components to the soil. Both useful application and disposal require more or less after-care of the immobilised product. In case of disposal, isolation, control and managing activities are required. When a product is usefully applied, the demolish of the object needs to be controlled. Costs Economic merit depends on the benefits and costs of the used and produced materials,
ned if the immobilised products can be usefully applied. Costs of transport, energy and labour are almost the same as for primairy materials. The disposal of the immobilised products may give some benefits. The benefit of the promotion of the waste material from C2disposal to C3-disposal is approximately f75.-/ton, from C2- to C4-disposal f2OO,-/ton. The total cost of the inorganic immobilisation of MIP-fly ash is estimated between f60,- and f180,- /ton fly ash [27]. In most cases the economic merit for both disposal and useful application is low or negative. Environmental aspects Inorganic immobilisation techniques are meant to contribute to the environmental merit of the waste. Therefore, a comparison is needed between other possible treatments of the waste material and inorganic immobilisation. In Dutch legislation the priority list of Lansink gives a first indication of the environmental merit of immobilisation. This priority of waste managements is as follows:
778 1. prevention 2. recycling; 3. useful application; 4. treatment; 5. combustion; 6. disposal. Immobilisation techniques can be placed on the fourth level and in most cases have to be compared with disposal of the waste. Most hazardous waste materials, such as hazardous sewage sludge, can not be treated otherwise or directly usefully applied. They are placed on a so called C2-, or C3-1andfill. An useful application of this immobilised waste material therefore seems to contribute to the environmental merit. However, a few comments have to be made. If the immobilisation technique does highly effect the environment, it is not clear that immobilisation is the right waste management. In case of inorganic immobilisation, degradation of the environment is caused by the use of energy, the use of primairy materials, the release of hazardous components during the process and the leaching out of hazardous components. Benefits for the environment are the replacement of primairy materials, such as gravel, by immobilised products and upgrading the waste material on the previous mentioned priority list of Lansink. Inorganic immobilisation show little release of components during the process and some use of additives and energy. The environmental merit of the technique can be the replacement of primairy materials by secondary materials. So, from environmental point of view, inorganic immobilisation is a good option, when the leaching out of components can be minimised and the properties of the product fulfill demands of legislation concerning building materials. Otherwise, an environmental merit is not simple to give. 3 PROPERTIES OF RELEVANT WASTE MATERIALS In the previous, the effectiveness of inorganic immobilisation has been considered. The effectiveness partly depends on the used additives. Besides that the properties of the waste material are an important factor in the effectiveness. Some components are good, other are hard to immobilise. In this paragraph, desired properties of the waste material are given. Inorganic immobilisation is partly based on chemical reactions. Therefore, the following properties of waste materials are desired. Waste materials containing heavy metals, especially Pb, Cd and As are mentioned in patents. Figure 4 shows that Mo and Zn are hard to immobilise. In addition Cr 8+ is mobile at high pH. It is therefore desirable to immobilise waste materials with high content of heavy metals, apart from Mo, Zn and Cr 8§ Secondly, the presence of anions is in most cases not desirable. Exceptions of anions that can be immobilised are sulfates and sulfites. Thirdly, low concentrations of organic compounds are desirable. Especially Volatile Organic Compounds can not be fixed in the solidifying matrix and can cause degradation of the immobilised product. The effectiveness depends not only on the chemical reaction, but also on the physical encapsulation. Some properties of the waste have negative or positive influence on this encapsulation. F.e., the presence of pozzolanic or hydraulic components in the waste is preferable. On the contrary, particles with size less than 74pm, are not preferable. They weaken the bond between particle and cement. From environmental point of view, the so called C2-waste materials can contribute more to environmental merit than C3-waste materials. Moreover, some C~ -waste materials are or will be prohibited to be disposed of, such as MIP-fly-ash. Furthermore it is desirable to minimise the pretreatment of the waste. Pretreatment requires additional energy, additional primairy materials and/or may lead to a residue, that has to be treated. Pretreatment may be required f.e. when the waste material contains oil, grease and particles less than 74pm.
779 Other considerations for the effectiveness of immobilisation are: variations in composition, other treatments for the waste material and the amount of the production of waste. Currently, some inorganic immobilisation techniques have been developed. Accordance to patents, the following waste materials are common: filter dust, coal ashes, MIP-fly ash, residues from burning of coal, sewage sludge, lime sludges, fluorgypsum waste, arsenical waste, slag and dust from steel making.
CONCLUSIONS In this last paragraph the advantages and disadvantages of inorganic immobilisation will be mentioned, which are deduced from the description of the effectiveness of inorganic immobilisation in the previous paragraphs. Because of the high pH of the product most heavy metals show good results, except from Mo and Cr 8+. Other components of the waste material are hard to immobilise, such as anions and organic compounds. These elements can leave the immobilised product through diffusion. Therefore, most inorganic immobilised products do not fulfill the legislation concerning building materials. Additives or pretreatment may be required to improve the immobilisation. Furthermore, immobilisation requires severe after-care of the product. However, in special cases this may lead to a decrease of economic and environmental merit of the technique. Inorganic immobilised products are highly sensitive to the formation of cracks. Especially erosion, wet/dry-periods, temperature changes and freeze/thaw-periods [25] influence the product. Besides these disadvantages, inorganic immobilisation can be a good solution for the amount of certain hazardous waste materials. REFERENCES [1] Bolier, D., 'lmmobilisatie mag weer', in: Milieumarkt, l Oe-jaargang, nr.5, 1996, p.27-29 [2] Bijen, J.M.J.M., Fraaij, A.L.A., Rooij, M.R., 'Materiaalkunde, collegedictaat',Delft, Technische Universiteit Delft, 1995 [3] Chevron Research Company, "Cleanup of oily wastes', WO-C-91/00900, 1991 [4] Conner, J.R., 'Chemical fixation and solidification of hazardous waste', New York, Van Nostrand Reinhold, 1990 [5] Civieltechnisch Centrum Uitvoering Research en Regelgeving, 'Handleiding voor het beoordelen van
immobilisaten' Gouda, CUR, 1995
[6] Civieltechnisch Centrum Uitvoering Research en Regelgeving, 'Beoordeling van immobilisaten, een voorstel voor criteria en testmethoden', Gouda, Civieltechnisch Centrum Uitvoering Research en Regelgeving, 1993 [7] Ecoserdiana, "Process for stabilizing and solidifying wastes from aluminium processing by means of an inorganic matrix', EP-C-O561746, 1993 [8] Fraaij, A.L.A., "Fly ash a pozzolan in concrete', Delft, Technische Universiteit Delft, 1990 [9] "Grenswaardennotitie, storten gevaarlijk afval', mei 1993 [10] Haese, R., e.a., "Verfahren zur Bindung yon organischen und anderen Stoffen, verbunden ,it einer Absenkung der Eluat-Werte bezOglig TOC und anderer Stoffe bei der Deponierung von Abf~llen', DE-B4405323, 1995
780 [11] Hendriks, Ch.F., 'Dictaat Bouwstoffenbesluit', Delft, Technische Universiteit Delft, 1995 [12] Interstock Nederland, "Process for the preparation and lech-resistant solidification of filter dusts and reaction products of flue-gas purification of waste and seage sludge incineration plants', WO-C-92/22512, 1992
[13] Lea, F.M., "The chemistry of cement concrete', Londen, Edward Arnold Publishers, 1970 [14] Lopat Industries, inc., "Compositions to encapsulate chromium, arsenic, and other toxic metals in wastes', EP-B-0352096, 1990 [15] Means, J.L., Smith, L.A. (e.a.), 'The application of solidification/stabilization to waste materials', Boca Raton, Lewis Publisher, 1995 [16] Noakes, John, E., "Solidifation of organic waste materials in cement', WO-B-92/15098, 1992 [ 17] Noyes, R., 'Unit operations in environmental engineering', Park Ridge, Noyes Publications, 1994 [18] Pelt&Hooykaas B.V., "Process for immobilizing environmentally noxious metals and organic substan-
ces', EP-B-0398410, 1990
[19] Pelt&Hooykaas B.V., "Toxic waste fixant and method for using same', EP-A-0535758, 1992 [20] Pelt&Hooykaas B.V., "Method for fixing waste material', EP-C-0547716, 1993 [21] Pelt&Hooykaas B.V., ",4 method for the immobilisation of waste material contaminated with organic chemical compounds', EP-B-0590711, 1994 [22] Projectgroep Ontwikkeling Saneringsprocessen Waterbodems, 'lnventariserend onderzoek naar "state of the art" van immobiliseren', Lelystad, Directoraat Generaal Rijkswaterstaat, 1992 [23] Rheinische Baustoffwerke GMBH&CO.KG, "Kompaktierung von Industriest~ube und Deponie der
Kompaktate', EP-B-0380713, 1990
[24] Stichting Postacademisch Onderwijs Civiele techniek en Bouwtechniek, 'Cursus-map immobilisaten', Delft, PAO, 1996 [25] Technotrans, "Conferentie immobilisatietechnieken', Ede, 1997 [26] Tuijn, J. van,'VROM is om: immobiliseren is weer bespreekbaar', in: Milieumarkt, 7'-jaargang, nr.6, 1993, p.18-21 [27] Vereniging van Afvalverwerkers, "Beoordeling van Koude en Thermische Immobilisatietechnieken voor verwerken A Vl-vliegas', Nationaal Onderzoeksprogramma Hergebruik van afvalstoffen, Utrecht, 1996 [28] Wastech, "Treatment of hazardous waste material', WO-A-91/05586, 1991
Goumans/Sendergvan der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All fights reserved. PHYSICAL P R O P E R T I E S C O N T A M I N A T E D SOIL
AND
LONG
Pirjo Kuula-V~iis~inen
781
TERM
STABILITY
OF
STABILIZED
Keijo Kumila
Tampere University of Technology, Tampere, Tampere University of Technology, Tampere, Finland Finland Hanna-Liisa J~irvinen
Geological Survey of Finland, Espoo, Finland ABSTRACT By far, Finnish authorities have applied Dutch procedures when accepting stabilization as a remediation solution. However, the local conditions should play an important role when quality standards are given. In this study one of the main interests was to evaluate the long term properties of stabilized materials in the harsh climate conditions. The investigated material (As, Cr, Cu contaminated) was sampled from the surface layer of an old impregnation plant in Eastern Finland. The binding agents used in stabilization were ordinary Portland cement, fly ash and gypsum. The mixes with different binding agents were planned to achieve two different compressive strengths, 5 MPa and 1 MPa. The compressive strength results indicated a slower early strengthening speed with the samples containing fly ash or gypsum. The measured water permeabilities varied between the values of 10-7 and 104 m/s. The results from the pore size distribution measurements showed a rather large amount of gravitation pores 20...30 % which result supports the rather large water permeability. The long term stability of samples were tested with freeze-thaw test (ASTM D 4842). A Finnish standard (SFS 5447) for testing the freezethaw resistance of concrete specimens was also used. The ASTM-freeze-thaw test results indicated that the samples of low compressive strength didn't withstand the stress as well as the firmer samples. All the samples were partially destroyed in the SFS freeze-thaw test.
INTRODUCTION In Finland there are several old saw mills and wood impregnation plants where the soil is ,-,.,.~.,,.,+,-.,,,~;~,-.,++,.,I k,,:~+ " u] ,,.+. o ~ , l ~ . , . , ,
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groundwater areas wherea grounwater is situated under postglacial gravel and sand deposits with large water permeability. There is a urgent need to isolate the contaminant from the surrounding environment either by stabilization or by other remediation methods. Stabilization with the cement based binder agents is the most common remediation method for inorganic contaminants in soils. Obviously the utilization of stabilization is based on physical encapsulation and chemical fixation of contaminants allthough there are several different theories discribing the binding mechanism of contaminants in cement matrix (Cockel). The utilization of fly ash and waste gypsum as binder agent in stabilization is can be done for economical reasons to replace the more expensive binder agent cement. But also there usage of fly ash can decrease the leachability of heavy metals from stabilized materials by decreasing the pH value of the cement matrix (Cote2). The most important physical properties of stabilized soils are low water permeability and sufficient compressive strength. The values of compressive strengths mentioned in the literature vary in the large range from 0.1 MPa to 10 MPa In Finland the weather conditions are rather hard compared to most parts of Europe. The utilization of stabilization in our climate conditions should be tested before the remediation
782 project is done. The purpose of this research project was to evaluate the physical properties of stabilized contaminated soils and also evaluate the long term resistance of stabilized samples against freezing and thawing using different standardized methods. One of the main interests was to evaluate the effect of different compressive strengths on physical properties and long term stability.
MATERIALS AND METHODS The contaminated soil was originally from a surface layer of an old impregnation plant in Eastern Finland. The soil was contaminated by a impregnation agent contaning arsenic, chromium and copper. The average amounts of these contaminants were 650 mg/kg for As, 480 mg/kg for Cr and 590 mg/kg for Cu. The average pH value of the material was 5.4. The total amount of soil delivered to our laboratory was 1000 kg. The equal quality of the material was guaranteed by controlled cross mixing method using a large laboratory mixer. During the mixing it was noticed that the contaminants were concentrated to a certain "hot spots" which couldn't be broken totally during mixing. The soil contained also small parts of wood. The grain size distribution of the soil sample is presented in Figure 1. The average grain size of the contaminated soil was 0.45 mm and the amount of fine fractions (<0.074 mm) was 4 %. According to the Finnish soil classification the material was determined as sand. ..................... 77
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G rain size (m m) Figl 1. The grain size distribution of the contaminated soil The binder agents used in stabilization were Ordinary Portland cement, fly ash and a mixture of calcium sulfate and fly ash. The fly ash was originally from a coal burnig power plant and the gypsum came from a suplhur refining process of a power plant, gypsum was calcinated at 200 ~ before the use. The amount of binder agent and water was evaluated according to the Finnish standards for soil-cements mixtures. The mixtures were planned to achieve certain compressive strength values at the age of seven days (1 MPa and 5 MPa). After the premium tests the amount of binder agent of 6 % was selected for 1 MPa samples and 9 % for samples of 5 MPa. The water binder ratio w/c was 0.8. The compressive strengths were measured at the ages of 7, 28 and 91 days. The total amount of samples prepared for this project was nearly 300. The test specimens were compacted with a gyratory compactor to a specific density of 90 % of the maximum proctor density. The water permeabilities were measured by an equipment using changeable hydraulic gradient (2-4 m) in a measurement cell with flexible walls.
783 The mineral composition of the original and stabilized samples were analysed with a X-ray diffraction analyser. The main minerals of the contaminated soil were quartz, plagioglace, feldspar, hornblende and chlorite. The pore size distributions were analysed by a mercury intrusion porosimetry. The chemical analysis of the samples and the leachability of contaminant are published later in other forum. The long term stability of 91 days old specimens was tested in three different methods: according to ASTM standards for wetting and drying (D 48433 ) and freezing and thawing (D 48424 ) and also with a Finnish SFS 5 standard for freeze-thaw resistance of concrete specimens. Both ASTM tests contain 12 testcycles. The ASTM freeze-thaw test temperatures are -20 ~ and +20 ~ The results of the ASTM test are presented as a relative mass loss of tested specimen compared to control specimens. After our opinion the the 12 cycles of ASTM-test for wetting and drying is corresponding one summer in Finland.. The ASTM-test for freezing and thawing is corresponding one year in Finnish weather conditions if the stabilized material is uncovered. The SFS-test contains 100 cycles of freezing (-20 ~ and thawing (+30 ~ The SFS test corresponds rather harsh climate stress during several tens of years.
RESULTS The results of the compressive strength measurements are presented in Table 1. The specific density of all the samples was on average 1,9 g/cm 3 and the moisture content was on average 4 %. The measured strengths were slightly higher than those that were originally planned. The early compressive strengths of specimens with fly ash or gypsum were smaller than the values detected with pure Portland cement. The hydraulic reactions are starting slightly slower when fly ash or gypsum are used. When the curing time increased to 91 days the compressive strengths reached the same level as with Portland cement. The amount of binder agent clearly controlled the level of the compressive strength.
Table 1. The compressive strengths of the stabilized specimens at the ages of 7, 28 and 91 days. Sample =
C6 C9 C6FA20 C9FA20 C6FA10G10 C9FA10G10 C6 FA20 G10 -
Compressive strength _.~..~.1_ ~-~ ...1 ~tt u z c age U . L 7 u a y ~ (MPa) 2,15 4,85 1,22 4,08 1,02 3,54
Compressive strength Compressive strength ~_, t t ,L_ uz~., a ~ u..~,-,o l 1 . . o day ~3 r a ~ t h e age e,f 91 d a y s ~ ~ (MPa) (MPa) 2,28 1,90 4,81 5,32 2,05 1,90 4,47 5,67 1,36 2,10 4,26 4,23
the amount of binder agent (cement) is 6 % 20 % of the binder agent is fly ash 10 % of the binder agent is gypsum
The water permeability test results are shown in Table 2. The water permebilities varied between 10-5 and 10 -7 m/s. Obviously the organic wood pieces and also the unbroken contaminant hot spots are increasing the water permeability values. The porosity results are also showing a rather
784 large amount of gravitation pores which clearly indicate a large water permeability. The total porosity measured by mercury porosimetry varied from 20 to 30 %. A smaller amount of pores was detected with the samples having higher compressive strength. The amount of gravitation pores (diameter > 10 lam) was on average 60 % when the amount of binder agent was 9 % and 80 % when the amount of binder agent was 6 %.
Table 2. The water permeabilities of specimens at the age of 91 days. Sample C6 C9 C6FA20 C9FA20 C6FA10G10 C9FA10G10
Water permeability (m/s) 2.5.10 -5 2.0.10 .5 2.2.10 -5 1.2"10 .6 6.8"10 .6 7.0"10 .6
The results from the ASTM wetting and drying test showed that during the test no remarkable changes was noticed the samples were as firm as before the test. The mass loss of the specimens were under 1%. The relative mass losses during the ASTM freeze-thaw test are shown in Table 3. The maximum weigth loss during the test was 10 %. The results indicate clearly that the resistivity of samples containing 6 % of binder agent is weaker than the resistivity of samples contaning 9 % binder agent. During the test the changes in sample structure was also noticed (Fig 2). The SFS freeze-thaw test results (Fig 3) indicate a partial collapse of the sample structure.
Table 3. The relative loss of mass during the ASTM D4842-90 freeze-thaw test. Sample C6 C9 C6FA20 C9FA20 C6FA 10G 10 C9FA10G10
The relative loss of mass (%) 5.04 0.28 9.81 0.25 10.23 0.2
Fig 2. Specimens after the ASTM D4842-90 freeze-thaw test.
785
riii! .... ~':~::~i!
ii!i!i~,!ii~,i~i~,i~,'~i~i,~,i'~i
:
i~
'~:~
Fig 3. A sample after the SFS freeze-thaw test.
The X-ray diffraction analysis showed that the samples contained ordinary cement minerals no prove was found of mineral compounds containing Aa, Cr or Cu. The chemical and minerological composition of stabilized materials will be tested more accurately in the future.
CONCLUSIONS The early compressive strengths of specimens contaning fly ash or gypsum as binder agents were lower than the values of those samples containing cement as binder agent. The compression strengths of all samples reached the same level after 91 days curing when the amount of binder agent was the same. The ASTM test for wetting and drying didn't show any changes in the sample structure and no differences between the samples. The ASTM test for freezing and thawing showed clearly that the samples with lower (1 MPa) compression strength had a lower resistance for freezing and thawing. Allthough after the literature (LaGrega et. al. 6) the relative weigth loss of 15 % would be acceptable. On the other hand the ASTM tests are rather soft for Finnish weather conditions. The SFS standard test showed clearly that the samples cannot resist hard weather conditions. After our opinion the ASTM freeze-thaw test should be done for all materials used in stabilization. The pore size distribution results and the water permeability results correlate well with each other. The pore size distribution test could be a premium test when stabilization mixture is planned. The chemical leaching tests and more accurate structrural analysis for all the tested specimens will be done in the future.
REFERENCES 1. Cocke, D. The Binding Chemistry and Leaching Mechanisms of Hazardous Substances in Cementitious Solidification/Stabilization Systems, Journal of Hazardous Materials, vol 24 (1990). 2. Cote, B. Contaminant Leaching from Cement-Based Waste Forms under Acidic Conditions, Ph.D Thesis, McMaster University, Ontario, Canada (1986).
786 3. ASTM D4843-88. Standard Test Method for Wetiting and Drying Test of Solid Wastes (1988). 4. ASTM D4842-90. Standard Test Method for Detrmining the Resistance of Solid Wastes to Freezing and Thawing (1990). 5. SFS 5447. Concrete. Durability. Freeze-Thaw Resistance (1988), in Finnish. 6. LaGrega M.D., Buckingham P.L. & Evans J.C. Hazardous Waste Management. McGraw-Hill (1994).
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
EVALUATION FOR
OF CONTAMINANT
787
RELEASE
STABILIZED/SOLIDIFIED
MECHANISMS
WASTES
F. Sanchez ~,2, A.C. Garrabrants ~, T.T. Kosson ~, J. Mdhu 2 and D.S. Kosson ~ 1Rutgers, The State University of New Jersey Dept. of Chemical and Biochemical Engineering P.O. Box 909 Piscataway, NJ 08855-0909 USA Tel: 1.908.445.4346 Fax: 1.908.445.2637
21nstitut National des Sciences Appliquees de Lyon LAEPSI/POLDEN CEI - BP 2132 - 27 bd du 11 Novembre 69603 Villeurbanne cedex France Tel: +33.(0)4.78.89.51.65 Fax: + 3 3 . ( 0 ) 4 . 7 2 . 4 3 . 9 8 . 6 6
ABSTRACT Understanding mechanisms which control release of inorganic contaminants from stabilized/solidified wastes is important for predicting the long-term leaching of contaminants in the field.
Frequently, release is assumed to be controlled either by solubility
or diffusivity as limiting cases. These assumptions also dictate the interpretation of laboratory test methods. The consistency of testing and interpretation methods was evaluated using a synthetic solidified matrix spiked with lead chloride, using both batch extractions on size reduced materials, where equilibrium was anticipated, and tank leaching tests, where diffusion controlled release was anticipated.
Differing approaches to determining
availability, solubility and dynamic release are compared.
1.
INTRODUCTION The development of stabilization/solidification processes using hydraulic binders has
resulted from increasingly stringent regulations in the field of environmental protection.
The
characterization of the leaching behavior of these materials is crucial in the environmental assessment of disposal or re-use scenarios. Numerous leaching tests with different purposes have been developed: simple tests used for regulatory compliance, and more "elaborate" tests used to better understanding of the physical-chemical phenomena that occur during and leaching and provide a basis for long-term prediction. One possible classification of leaching tests is to distinguish between equilibrium and mass transfer leaching tests.
Equilibrium
leaching tests, which typically are conducted on crushed materials, aim to determine an available release potential (e.g., Availability Test [1]), constituent solubilities (e.g., by varying pH [2, 3]), or matrix alkalinity.
Mass transfer leaching tests carried out on
788 monolithic samples over a long period (from one week to 3 months or more) aim to determine release rates and account for chemical and physical properties of the matrix. Their interpretation is generally carried out by coupling results obtained with the equilibrium leaching tests. A comparison of leaching tests applied to stabilized/solidified wastes and their interpretation was carried out within the framework of the international scientific collaboration between the Association RE.CO.R.D (Waste Research Cooperative Network, France) and the HSMRC (Hazardous Substance Management Research Center, US). Results discussed in this paper are the outcome from two parallel studies conducted by the research groups at the National Institute of Applied S c i e n c e s - LAEPSI/POLDEN - - (INSA), Lyon, France [4] and Rutgers, The State University of New Jersey (RU), USA [5]. The specific objectives of the study were to compare testing results from varied conditions for measurement of (i) matrix alkalinity, (ii) constituent solubility as a function of pH and (iii) dynamic release. A synthetic solidified waste spiked with lead chloride was used for the initial phase of this study.
2. MATERIALS AND METHODS 2.1
Solidified
Matrix
Preparation
The synthetic solidified waste was prepared by mixing 34 wt% Ordinary Portland Cement, 11 wt% water, 44 wt% sand, 10 wt% lead chloride, and 1 wt% sodium chloride. Although the INSA and RU samples were prepared using the same mixture, they were prepared at two different times using different lots of cement. Samples for INSA were cast as blocks of 15x20x10 cm 3. Samples for RU were cast as 10 cm diameter by 10 cm height cylinders using polyethylene molds. Samples were vibrated in the molds for one minute before being stored at room temperature in sealed plastic bags. After 28 days of curing, samples for INSA were cored from the cast blocks as 4 cm diameter cylinders with 1, 2 or 8 cm heights. Fragments of the blocks were saved in sealed plastic bags as source material for tests on crushed materials.
After 6 months of curing, samples for RU were removed from the molds
and used for testing.
2.2.
Measurement
of Matrix
Alkalinity
and
Lead
Solubility
Determination of matrix alkalinity and lead solubility from the matrix as a function of pH was carried out using modifications of the Acid Neutralization Capacity leach test [6]. RU samples were crushed and jar milled until > 85% passed through a 300 #m sieve.
In
parallel, ten 5 g aliquots of the size-reduced matrix were extracted using nitric acid solutions
789 of varying acidity.
Deionized water and nitric acid were added in varying proportion to each 5
g sample to achieve acid additions ranging from 0 to 10 meq/g solid and a final liquid addition of 30 ml (liquid to solid (LS) ratio of 6:1 ml/g).
Each case was carried out in triplicate.
After liquid addition, the mixtures were agitated for 24 hr.
Subsequently, the solid and liquid
phases were separated using vacuum filtration through 0.45 l~m pore size polypropylene filters. The filtered extract was analyzed for pH and lead. INSA samples were size reduced by grinding to pass through either a 160 l~m or 315 ~m sieve [4]. Two different size reduction criteria were used to assess the impact of size reduction on observed results.
In parallel, twenty 15 g aliquots of the size-reduced matrix
were extracted using nitric acid solutions of varying acidity.
Deionized water and nitric acid
were added in varying proportion to each 15 g sample to achieve acid additions ranging from 0 to 10 meq/g solid and a final liquid addition of 150 ml (LS of 10:1 ml/g).
After liquid
addition, the mixtures were agitated for either 24 hr or 5 days to assess the time required to achieve equilibrium for both the < 160 and < 315 l~m material. The case using < 315 l~m material with agitation for 24 hr was carried out in triplicate, while the other cases were carried out without replication.
Subsequently, the solid and liquid phases were separated
using vacuum filtration through 0.45 l~m pore size polypropylene filters.
The filtered
extract was analyzed for pH and lead. The acid neutralization behavior of the materials was evaluated by plotting the pH of each extract as a function of milli-equivalents of acid added per gram of dry solid. Lead concentration in each extract is plotted as a function of extract final pH to provide solubility as a function of pH. The acid quantities required to reach pH 11.9 and pH 9 were observed because of the nature of the studied materials (cement matrix).
Indeed, pH 11.9 is the pH
theoretically reached after the complete neutralization of the portlandite (pH of transition between the predominance of CSH (i.e., gel formed of hydrated calcium silicates) and portlandite (i.e., Ca(OH)2) and the predominance of rich calcium silicate CSH gels) [4]. pH 9 is the pH for which the solubility of most amphoteric metals is minimum and corresponds also to the stability limit of the main hydrate phases of a cement matrix. The concentration of calcium hydroxide produced during the hydration reactions of the cement, can also be estimated from the curve.
2.2. Mass Transfer
Leaching Tests
Short duration monolith leaching tests were carried out on the RU samples in triplicate.
Each monolithic sample (10 cm diameter by 10 cm height) was immersed in 4.6 I
of deionized water, equivalent to 10 ml of leachant/cm 2 exposed surface area. The leachant
790
was refreshed with an equal volume of deionized water, following a
2N
progression of 3, 6, and
12 hours, 1, 2, 4 and 8 days. The leachant to surface area ratio and refreshing intervals were selected to minimize solubility limitations of lead in the leachant.
After each extraction
interval, the resulting extracts were filtered through 0.45 pm pore size polypropylene filters and preserved with nitric acid to pH < 2 for chemical analysis. The concentration of constituents in each leachate was measured with flame atomic absorption spectrometry. The observed diffusivity for each replicate was calculated as the mean of the observed diffusivity for each leaching interval.
The observed diffusivity for each leaching interval was
calculated as Dobs, i = ~
2pA
(t 65 - t ~ o
i+l
[m2/sl )
where B~ =
mass released per unit surface area [mg/m 2] during the leaching interval [ti,
ti+ 1]; p
=
Ao =
sample density [kg/m3]; constituent availability determined by the Availability Leach Test [1] or initial concentration [mg/kg]; and
ti
=
leaching time interval i [s].
Only intervals during which the slope of mass released as a function of the log of the time interval was between 0.35-0.65 were included in calculation of the mean diffusivity. Long-term monolith leaching tests were carried out on INSA samples.
Cylinders 4 cm
diameter and 1, 2 or 8 cm height were used to provide variable external surface area and to observe the depletion of very soluble species (e.g., sodium and chloride) during the testing. Three replicates were carried out on the 2 and 8 cm height samples. Deionized water was employed as the leachant based on a liquid-to-solid ratio of 10 ml of leachant per gram of solid. The leachant was refreshed with an equal volume of deionized water at time intervals of 3, 5, 16, 24 hours, 2, 3, 4 days, 1, 2, 3, 4 weeks and thereafter every month up to a cumulative leaching period of 8 months.
After each extraction interval, the resulting extracts
were filtered through 0.45 pm pore size polypropylene filters and preserved with nitric acid to pH < 2 for chemical analysis. The pH of the filtered leachates was measured prior to preservation.
Leachates also were analyzed for Na, Ca, Pb with ICP-AES and CI with ion
chromatography.
Complete results have been reported elsewhere [4].
The initial leachable concentration, Co, (i.e., availability) and the observed diffusivity of very soluble species (e.g., sodium or chloride) in the matrix were calculated using a 3-
791
dimensional diffusion model [7]. These parameters are simultaneously identified to account for depletion of the species in the solid core. A model coupling dissolution and transport by diffusion was used for species whose solubilities exhibit a strong dependence on pore water pH [4]. This model was used because of large pH gradients which existed between the matrix interface with the leachant and within the solid matrix. This model can be divided into several stages: 9 Release of portlandite using a shrinking front model; 9 Calculation of the induced pH profile, assuming thermodynamic equilibrium occurs in the pore water; 9 Determination of local lead solubility by calculation assuming the main equilibria or from experimental results (e.g., equilibrium leaching tests); and, 9 Description and calculation of lead transport by diffusion in the pore water. The coupled dissolution/diffusion model requires the knowledge of several parameters: matrix porosity, solid phases concentrations of constituents of interest (e.g., Ca, Na, Pb), constituent solubility as a function of pH, and observed diffusivity within the porous medium for each species of interest. The values of these parameters were initialized using experimental data and then adjusted by successive simulations until the simulated results coincided with the observed experimental results. The equilibrium concentration of lead as a function of pH was determined using a simplified representation of the cement pore water solution assuming two principal components, calcium hydroxide and lead hydroxide. The resulting chemical equilibria considered were Ca (OH) 2 r
Ca 2+ + 2 O H
10-5.2=[Ca 2+] [OH-] 2
(If solid Ca(OH)2 is present)
Pb(OH) 2 ~
Pb 2+ + 2 O H
K2= [Pb2+] [OH-]2
(If solid Pb(OH)2 is present)
P b ( O H ) 3 ~:, Pb 2+ +3 O H
K 3 = [Pb 2+] [OH-] 3 [Pb(OH) 3]
2 H20 ~
HBO++OH -
l 0 -]4 = [H 3 0 + ] [ O H - ]
The solubility product for calcium was provided by the literature [8].
The solubility
constants for lead also could be found in the literature, but may be far removed from the experimental conditions. Thus, solubility constants for lead were estimated from the experimental solubility curve given by the equilibrium leaching tests and adjusted so that the simulated results were consistent with the results of the monolithic leaching test.
792 3.
RESULTS
3.1
Acid
Neutralization
Capacity
and
Lead
Solubility
Acid neutralization capacity curves for each experimental condition are provided in Figure 1. Contact time (24 hours or 5 days) did not significantly influence acid neutralization capacity for the INSA samples size reduced to < 160 ~m and yielded the same results as the RU samples. RU samples size reduced to less than 300 ~m appeared to exhibit similar behavior as the < 160 ~m INSA samples because of the more aggressive size reduction procedure employed, which most likely resulted in a finer particle size distribution than 300 ~m. Contact time did effect the acid neutralization capacity observed between pH 11 and 3 for INSA samples size reduced to < 315 ~m.
Initial neutralization of portlandite (to pH 11.9)
appeared to be the same for all samples, but dissolution of the CSH gel at pH < 11.9 appeared to be incomplete when the larger particle size sample was contacted for only 24 hr. This is probably a consequence of less exposed surface area for the larger particles and thus slower dissolution.
For a given particle size, the contact time required to reach 90% of constituent
solubility increases with increasing fractional solubility and pD (Figure 2; Crank [9]) 1,2 Thus, the contact time required to reach 90% of solubility for a constituent with a Pgobs of 14.5 and fractional solubility much less than 10% (i.e., MoJMo< 0.1), is greater than 24 hr. The acid quantities per gram of solid dry required to reach pH 11.9 and pH 9 are presented in Table 1 according to the size reduction and contact time. The concentration of calcium hydroxide produced during the hydration reactions of the cement can be estimated at 230 kg/m 3 of porous medium (ca. 21% of the hydrated cement paste) based on the acid quantity required to reach the pH 11.9. Lead solubility as a function of pH for each experimental condition is presented in Figure 3.
1
pD=-Iog(D) where D is in units of m2/s.
2 The diffusion of a constituent from a sphere into an infinite solution can be modeled as Mo
/t 2 =
, l_ Oo,,.n'.,'.t/
~exp
a2
where, Mt
--
M o
=
gobs
=
M t/Mo = t a Moo
= = =
constituent mass released from the particle size in time t [mg]; initial constituent mass in the particle (availability) [mg]; fractional release; observed constituent diffusivity in the particle matrix [m2/s]; contact time or test duration [s]; particle radius [m]; and constituent mass released from the particle size in an infinite time [mg].
793 Overall, test results obtained from all experimental conditions were in good agreement. Neither contact time nor extent of particle size reduction influenced the observed lead solubility.
This is in contrast to the particle size effect observed for acid neutralization
capacity but consistent with the dissolution of Ca(OH)2 from the solid matrix. The impact of liquid-to-solid ratio appeared to be slight.
For pH < 5, the solubility obtained with the ratio
of 6:1 ml/g was slightly greater than the results obtained with the ratio of 10:1 ml/g; for pH > 5, the results of solubility obtained with the ratio of 6:1 ml/g were slightly less than those obtained with the ratio of 10:1 ml/g. The minimum solubility was observed between pH 8 and 10 which suggests that lead hydroxide is the solid phase controlling solubility [8].
3.2 Assessment of Dynamic Release 3.2.1
Leaching behavior of soluble species (sodium and chloride) Cumulative release of sodium and chloride as a function of time for all experimental
are presented in Figures 4 and 5. A comparison of the leachant to surface area and volume ratios, as well as sample surface area to volume ratios are useful for interpreting release data (Table 2). The release of sodium was rapid with > 80% of the total content released from the long-term monolithic leaching test samples but only 22% from the short-term test samples after one week of leaching. The release of chloride during one week of leaching was ca.
50% of the total chloride from the long-term monolithic leaching test samples and 4%
from the short-term leaching samples.
The smaller fraction of sodium and chloride released
from the short-term test samples over this time period was attributed to the smaller sample surface area to sample volume ratio. Release over 8 months from the long-term samples was consistent with the sample surface area to volume ratio for each case. Depletion occurred most rapidly for the sample with the greatest surface area to volume ratio. flux for each element was similar for all cases.
Initial release
However, there was greater initial release of
sodium from the RU samples which may be attributable to surface washoff as a consequence of sample molding. Values of availability and diffusivity for sodium and chloride determined using the different interpretation techniques are compared in Table 3.
Availability for the long-term
release test cases was estimated using the 3-dimensional diffusion model, while availability for the short-term release test cases was measured using the Availability Leach Test. The available sodium content was different for the long and short-term test cases resulting from the different batches of the solidified matrix.
However, the fraction of total sodium which was
available was consistent between the two approaches (92 vs. 95%).
Availability
794 determinations for chloride also provided consistent results for the two approaches (67 vs. 63%). Observed diffusivity values for sodium were consistent for both approaches. However, the observed diffusivity for chloride using the short-term testing approach was sixfold less than that obtained from the long-term testing approach. A comparison of the long-term flux measured and predicted by the 3-dimensional diffusion model indicates that good agreement is obtained up to
ca.
500 hr but measured fluxes are greater than predicted at longer intervals
(Figure 6). The residual flux observed after 500 hr may be a result of dissolution of the constitutive phases of the solidified matrix.
3.2.2
Leaching behavior of lead The coupled dissolution/diffusion model was used to simulate results obtained from
both long-term and short-term monolith leaching tests.
Parameter estimates used in the
model and a comparison of model results with long-term leaching of lead and calcium are presented in Table 4 and Figure 7, respectively. A comparison of model results with shortterm leaching for lead are presented in Figure 8. Overall, the parameters estimated from the short-term testing are consistent with those from the long-term testing.
The equilibrium
constants derived from the model fitting to the experimental data also permit simulation of lead solubility in the pore water as a function of pH. This simulated lead solubility is compared with the experimental results from equilibrium testing described earlier (Figure 9). The lower predicted solubility as compared to the measured solubility at pH > 8 may from differences in the aqueous ionic strength and chloride complexation between the experimental conditions.
Within the pore matrix during long-term testing, ionic strength and chloride
complexation with lead may be less than in batch experiments because of the rapid removal of sodium and chloride from the matrix. An observed diffusivity for lead also can be calculated from the short-term testing results using the interval method. The resulting Dobs, equal to 1016 m2/s, also can be related to a retardation factor and effective diffusivity according to Dobs = De. R If we consider as the effective diffusion coefficient the identified coefficient obtained by using the coupled dissolution/diffusion model (i.e., D e ---10 9"1 m2/s) the retardation factor R is then equal to 8x10 6. However, this does not clearly correlate with the equilibrium constants according to the relation
795 AVLT * Solubility= L ~oao*
p].m
q
R-1
where AVLT
=
constituent availability or initial concentration [mg/kg]; sample density [kg/m3] 9
p
11
=
porosity" and
R
=
retention factor.
The coupled dissolution/diffusion model also indicates that no depletion of lead in the solid phase occurs at the solid-liquid interface during the time scale of the long-term laboratory testing. liquid interface.
Thus, lead release is controlled by solubilization phenomenon at the solid-
In this case, the coupled dissolution/diffusion model can be simplified.
The
diffusional transport of lead within the matrix can be neglected and the flux of lead can be expressed in terms of a mass transfer coefficient and the difference between the lead saturation concentration and the concentration in the leaching solution. Consequently a shrinking core model can be used to describe the release of calcium and pH changes at the solid-liquid interface.
Lead release then can be described using a constant mass transfer
coefficient and changes in lead solubility as a function of pH [4].
3. C O N C L U S I O N S Based on the experimental results, modeling and previous observations, the following conclusions can be made: 9 Measurement of acid neutralization capacity is sensitive to the extent of particle size reduction and extraction period.
Use of particle size reduction to < 165 ~m in conjuction
with a 24 hr agitated extraction has been demonstrated to achieve equilibrium.
However,
determination of lead solubility in conjunction with acid neutralization capacity was not sensitive to extent of particle size reduction or extraction period over the range of conditions examined. These relatively simple tests provide a rapid method to measure the buffering capacity of solidified materials and the solubility of specific metals as a function of equilibrium pH.
Results also provide information on the chemical speciation of the
pollutants considered and allow prediction of the amount of acid required to neutralize a certain alkalinity and to decrease the pH to a certain point. This information can be used to estimate how long the buffering capacity will last [10] and consequently when the solubility of amphoteric metals may increase dramatically.
796 Evaluation of leaching parameters did not appear to be sensitive to curing intervals in excess of 28 days.
Short-term (8 days) and long-term (8 months) dynamic leach tests on monolithic samples provided consistent estimates of leaching parameters (observed diffusivity, leachable concentration or availability) for very soluble species.
Thus, short-term
testing may be used to evaluate long-term release when the principal mechanisms of release are well understood. Evaluation long-term leaching should consider the effect of the depletion of very soluble species from the solid matrix on leaching parameters (e.g., porosity).
Determination of constituent availability based on direct experimental measurement and two parameter modeling of dynamic release provided consistent results. Lead appeared to be less soluble in the pore water of the solidified matrix than observed during determination of solubility using a batch extraction on size reduced material. This effect may have resulted from rapid initial release and depletion of very soluble species (sodium and chloride) during monolith leach testing.
797 Table 1. Acid neutralization capacity from experimental conditions evaluated. mEq/g dry
INSA samples < 160 i~m 24 hr 5 days
Contact time
RU samples
< 315 Ilm 24 hr 5 days
< 300 l~m 24 hr
pH 11.9
2.2
2.3
2.6+0.005 a
2.8
2.4+0.6 a
pH 9
5.9
5.8
3.2+0.08a
5.7
6.5+0.2a
a
Standard deviation of data points from their mean (three replicates).
Table 2. Comparison between sample surface area, volume and leachant volume for experiment conditions evaluated. INSA samples Ratios
d 4 c m , h 1 cm
d4cm, h2cm
d4cm, h8cm
d 10 cm, h 10 cm
3
2
1.25
0.6
8.4
12.5
20
10
10
10
10
2.4
/ Sampl._.__ee surfac.__eear_.ea/ (cm.1) Sample volume ) /
Leac_..hatevolum_ee / ( cm ) Sample surface area)
( Leachatev~ mass
j
RU samples
d = diameter, h = height.
Table 3. Comparison of the leaching parameters for monolith leaching tests. Long term monolithic leaching tests (8 months of leaching) 3-D di'f'fusional model
Diffusivity test method (One week of leaching)
Interval diffusion coefficient method
Availability test - Interval diffusion coefficient method
Co (kg/m a)
Co/Ci (%)
- log Dobs (m2/s)
Co (kg/m a)
-log Dobs (m2/s)
Co (kg/m 3)
Co/C~ (%)
-log Dobs (m2/s)
Sodium
9.0
92
10.2
9.0 b
1 0.5
17+0.8a
95
10.3+0.02 a
Chloride
52.7
67
10.5
52.7 b
10.9
58.7+0.3 a
63
11.3+0.01a
Elements
Standard deviation of data points from their mean (three replicates). b The availability test was not carried out on these samples. Therefore, the values used for the available potential release are the identified values obtained by using the 3-D diffusional model. a
798 Table 4.
Parameters values from the coupled dissolution/diffusion model.
Porosity Solid phases concentrations (kg/m 3 of porous medium)
Equilibrium constants
Effective diffusion coefficients: -log De (m2/s)
22% a Ca(OH)2
230 b
Pb(OH)2
217 c
Product of solubility: [Pb2+][OH-] 2 [Pb 2+ ] [OH-] 3 Constant of complexation: [Pb (OH);] Ca2§ Pb2§ PbIOHla
INSA results 10 16s
RU results 10-16-7
10 1398
10 -13"98
9.15 9.1 9.1
9.15 9.3 9.3
a The experimental porosity measured by mercury intrusion was 11% on unleached samples. However, the identified porosity was 22%. This difference may be due to the variation of the porosity during the leaching according to the release of soluble species and precipitation of phases (i.e., pursuit of the hydration reactions, carbonation). As the structural evolution of the matrix during the leaching is not included in the model, the porosity is adjusted by successive simulations. b Estimated from the acid neutralization results. c Calculated from the initial quantity incorporated.
[]< 160 pm, 24 hr- INSA 9< 160 pm, 5 days - INSA O< 315 pm, 24 hr- INSA ~,< 315 pm, 5 days - INSA O<300pm, 24hr- RU
14 12 10
1000
.J=.
f
f
i
:~ 100
8
fJf
Particle Diameter = 300 t~m
60 hrs
I 0 . 1 < M~/Mo <0.9
[] 18 hrs
6 4
o ~ o
10
o
1
6 hrs
i
0
2
4
mEq/g
6
8
10
12
dry solid
Figure 1. Acid neutralization capacity curves from experimental conditions evaluated.
I
13
14
15
'
16
pO (-LOG [mg/m2])
Figure 2. Required test contact time to achieve 90% solubility as a function of constituent solubility as a function of constituent diffusion coefficient with a 300 ~m diameter particle.
799
-1
a"
A
~-2 E
n< 9< O< O<
bo -
160 160 315 300
~m, 24 hr - INSA lLtm, 5 days - INSA pm, 24 hr - INSA ~um, 24 hr - RU
"~wolm
od\
.~-3
LS = 6:1 ml/g - RU LS = 10:1 ml/g - INSA
-()~'0
o~-4 o
,-,Q 0 m
-5
o
mb
$
-6 0
2
4
6
8
10
12
14
pH
Figure 3.
L e a d solubility as a function of pH m e a s u r e d using batch e x t r a c t i o n s on size r e d u c e d
material. 5.5
6
&" 5
~'5.5
E
o
m
0
z
o
4
-
l:l
0 0 000000
E --
Q
o
B
3,5
0
0
4.5
W~
9
0 ~,d4cm []d4cm Figure 4.
1 log h 1 cm-INSA h8cm-INSA
2 t (hr)
3
<>d4cm h 2 c m - I N S A O d l 0 c m h 10 cm- RU
C u m u l a t i v e r e l e a s e of s o d i u m as
a f u n c t i o n of time.
4
0
1
log
2 t (hr)
3
4
Ad4cm
h 1 cm-INSA
<>d4cm
[]d4cm
h8cm-INSA
Od 10cm h 1 0 c m - R U
Figure 5.
h2cm-INSA
C u m u l a t i v e r e l e a s e of c h l o r i d e as
a f u n c t i o n of time.
800 1
A
(R
E
0
O~
E
Z
O
-2
"EL
0
O~
E
ri-
1
A
/
-2
J
O
O
x
X-3
-3
m
ILL O~ O
I,I.
-4
O~ O
dP
-5 -1
0
1 log t
2 (hr)
3
m
-4 -5
4
-1
0
1 log
Figure 6. Comparison of diffusional model prediction and experimental data. chloride are shown.
A
E
0.5
A
u}
o
~-o,s,~ 0
TM
2 t (hr)
3
4
The flux of sodium and
-1
~-1.5
El
E
-2
.............
~. - 2 . 5
-1
0
0
x -1.5 _=
....
-3
~ -3.5 ~ 0
_o -2.5
-4
-4.5
1
0
1
2
log t (hr)
3
4
1
0
1 2 log t (hr)
3
Figure 7. Comparison of coupled dissolution/diffusion model prediction and experimental data. flux of calcium and lead during the long-term monolith leaching tests are shown.
4
The
801
[] 0 A
04
E
O'}
0
-0 5
-1
-1
~'-2
E -1.5 I~.
-2
a.
-3
-5
_~ 6
.r/ O
-7
-4
-4.5 -1
0
1 log t (hr)
2
Figure 8. Comparison of coupled dissolution/diffusion model prediction and experimental data. The lead flux during the diffusivity test (i.e., one week of leaching) is shown.
4.
-
.=-4
w. - 3 . 5 o'J o
---
~o_ 3
"-2 " 5 0 x
< 160 l~m, 24 hr- INSA 9 < 160 pm, 5 days - INSA O < 315 #m, 24 hr- INSA O <300#m, 24hr-RU Simulation
-8
0
2
4
6
8
10
12
pH
Figure 9. Comparison of simulated and experimental data of lead solubility.
ACKNOWLEDGMENTS
This research was supported by the Association RE.CO.R.D. (Waste Research Cooperative Network, France) and the HSMRC (Hazardous Substance Management Research Center, USA), an Advanced Technology Center of the New Jersey Commission on Science and Technology.
5.
1.
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803
Response of Various Solidification S y s t e m s to Acid Addition J.A Stegemann a, C. Shi, and R.J. Caldwell Wastewater Technology Centre, now operated by Water Technology International Corporation P.O. Box 5068, Burlington, Ontario, L7R 4L7 aPresent address: Imperial College of Science, Technology and Medicine, London SW7 2BU
Abstract Experiments were conducted to examme the responses of five different solidification systems, with and without waste, to acid addition. The chosen cementing systems were: portland cement, portland cement with silica fume, alkali-activated blast furnace slag, lime and coal fly ash, and high alumina cement with lime and gypsum. The solidified products were tested at several ages using the Acid Neutralisation Capacity test, a series of batch extractions of ground wastes with varying amounts of nitric acid, which allows a titration curve to be plotted. Experimental results indicate that different cementitious systems vary in their response to acid addition; the location of the pH plateau of the titration curve depen&s on the nature of the hydration products formed by the binder system and is affected by waste components.
1. INTRODUCTION Stabilisation/solidification is used to convert wastes into non-hazardous materials, which can potentially be used in construction applications. Conventional cement-based solidification processes provide a high pH environment in which heavy metal contammants have a low solubility. Contaminants can also be captured within the physical structure of a monolithic cement-based matrix. Acid attack can result in contaminant release by 1) reducing the pH to a range where heavy metal contaminants become more soluble, and 2) corroding the physical structure of the cement-based matrix. Evaluation of the efficacy of solidification processes usually emphasises the first mechanism. For example, the USEPA Toxicity Characteristic Leaching Procedure (U.S. Federal Register, 1986) examines the solubility of metals upon addition of a limited amount of acid, and the Acid Neutralisation Capacity test, recommended by the Wastewater Technology Centre (1991), focuses upon the amount of acid required to achieve a pH of 9, below which the solubility of many metals increases. However, a non-monolithic, soft-like waste product will have a high permeability, so that even a high acid neutralisation capacity can readily be depleted by environmental influences such as carbonation, acidic ram or ground water. Thus, production of a dense, durable monolith is the key to successful containment of contaminants. Acid resistance is an important aspect of the durability of a monolithic solidified waste product, and the pH at which dissolution of the solidified waste matrix occurs may not be the same as that at which precipitated metals dissolve. Resistance of a cement matrix to acid attack will depend on 1) the matrix morphology, mcludmg the density and porosity, and 2) the ability of the matrix components to neutralise acid. Passivation by deposition of reaction products on the surface of the monolith may also play a role. The present work was undertaken in order to examine the response of different solidified waste matrix components to acid addition.
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2. BACKGROUND
2.1. Solidification Systems Although some wastes are alkaline and have their own acid neutralisation capacity, this is usually minimal. Thus, the acid resistance of a solidified product is normally associated with the binder system. The binder system may comprise from 20 to 60% of the solidified product. While portland cement is still one of the more frequently used binders, other cementing materials are being increasingly utilised in the solidification industry. These materials are often industrial by-products, which may be available at a lower cost than portland cement, and may even result in a superior solidified product if they are used appropriately. Some common binders for waste solidification include portland cement, cement kiln dust, Class C or Class F coal fly ash, and blast furnace slag. Additives such as silica fume may be used for matrix densification; high calcium lime may be added for pozzolanic activation; dolomitic lime, or even limestone may be added primarily to provide acid neutralising capacity. All of these binder systems are principally composed of oxides of calcium and silicon, with lesser amounts of aluminum, iron and sulphate. Minor quantities of other elements, such as sodium, magnesium, potassium, carbonates, and chlorides, may also play a role in these systems. The portland cement manufacturing process is controlled to result in a relatively consistent product, consisting of approximately 45% tricalcium silicate, 27% dicalcium silicate, 11% tricalcium aluminate, 8% tetracalcium aluminoferrite, and 3% gypsum 1. Upon reaction with water, these crystalline compounds produce a matrix which is approximately 50% calcium silicate hydrate (CSH) and 20% calcium hydroxide; the remainder of the matrix is composed primarily of hydrated calcium sulphoaluminates, calcium sulphoaluminates, including ettringite (Ca6Al~(SO4)~(OH)1~.26H20), which has been touted for its uptake of contaminants (McCarthy, Hassett and Bender, 1992 and Bambauer et al., 1988) and calcium aluminoferrites. By contrast, industrial by-products may vary. considerably in their characteristics, both within and between the facilities that produce them. The most hydraulically reactive materials of this type are composed of aluminosilicate glass, with significant concentrations of calcium, iron and other modifiers. Some reactive crystalline materials may also be present. Often, the reactivity of industrial by-products can be enhanced by the addition of chemical activators, such as sodium silicate, or potassium or sodium carbonate or hydroxide. Because of the varying composition of industrial by-product binders, and the vitreous nature of some, the reactive species can not easily be expressed as stoichiometric compounds, as can those in portland cement. However, the hydration products of industrial by-product-based binders are also mainly CSH, with hydrated calcium alummates, calcium sulphoalummates, and calcium aluminoferrites. By contrast with portland cement, silica-rich industrial by-products react with calcium hydroxide to produce CSH; therefore, calcium hydroxide is not a reaction product, although excess calcium hydroxide may be present mitially, and some may remain after hydration is complete.
2.2. Cement Hydration Products and pH Although there are cement and concrete structures which are thousands of years old, conventional cement and concrete products are not generally designed to maintain their integrity over the very long time spans which would be desirable for durability of solidified products. Hence, acid resistance of cements is not of equal mterest for construction purposes, and has not been extensively researched. Nevertheless, review of the literature yields information about the chemistry of cement hydration products which is relevant to the consideration of acid neutralisation capacity. 1The difference from a total of 100% is accounted for by variations in these percentages, and the presence of other minor components.
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Lime and Calcium Silicate Hydrate In investigating the structure of CSH, several researchers have studied the pH of CSH formed in pure calcium oxide-silica-water systems at different Ca/Si mole ratios. Figure 1 plots data generated by Greenberg and Chang (1965), and data from a paper by Grutzeck, Benesi and Fanning (1989), which are also in agreement with data generated by earlier researchers. With a small shift in pH between the two studies, the data appear to indicate three pH regimes, controlled by different types of CSH, in combination with silica gel or lime. At a Ca/Si ratio of 2 to 3, such as is found in portland cement, calcium-rich CSH, with a Ca/Si ratio of approximately 1.7, coexists with lime. The lime is highly soluble, and controls the pH at 12.3. Taylor (1993) found that the CSH in portland cement blended with industrial by-product binders has a lower Ca/Si ratio, and falls over time, due to the consumption of lime by silica. For example, the mitial Ca/Si ratios of the CSH in systems contaming fly ash and silica fume are 1.55 and 1.5, respectively, and fall to 1.45 ano 1.0 over time. Aqueous solubility of calcium and silica from this CSH is relatively low (Atkmson, Goult and Hearne, 1985). For Ca/Si ratios from 1.7 to 1.1, the pH falls from 12.3 to ll.9; at a Ca/Si ratio of 1.1/1.0, there are mdications of pH control at I 1.9 by co-existmg calcium-rich and silica-rich CSH; then, for Ca/Si ratios from 1.0 to 0.65, the pH decreases from 11.9 to 9.9. At. pH 9.9, silica-rich CSH with a Ca/Si ratio of 0.65 coexists with silica gel (Grutzeck, Benesi and Fanning, 1989). Due to the presence of mcreasing quantities of silica gel, which takes on water, and has a relatively high solubility, a structurally stable matrix can not exist in this low pH range. Mathematical modelling of portland cement-based wasteform durability (Atkinson, Goult and Hearne, 1985) has shown that, under certain specified leaching conditions, it would take in the order of l05 years for the pH to drop from 12.5 to 12.0, corresponding to a decrease in the Ca/Si ratio from 3 to 1.7, due to leaching of lime, but, due to the low solubility of CSH, it would take another more than 10~ years for the pH to drop from pH 12 to 10.5, corresponding to a decrease in Ca/Si ratio from 1.7 to 0.8. The significance of solidified waste pH control by soluble alkaline components is further illustrated by earlier data for a heavy metal platmg sludge solidified with four different binder systems, at the Wastewater Technology Centre (CSt~, 1986). Table 1 summarises the decrease in leachate pH and specimen mass over a two-year modified ANS/ANSI 16.1 test. The mass losses recorded for specimens containing fly ash were significantly lower than those for portland cement-based specimens not containing fly ash, even though the fly ash systems showed a greater decrease in pH. It is postulated that both the loss of mass, and decrease in pH in the portland cement specimens was due to leaching of alkali and lime. By contrast, the pH of the fly ash systems may have decreased partly due to reaction of lime to form less soluble calcium silicate hydrate. Table l" Summary of 2-Years of PH and leachin ~ data System pH Soluble silicate/p0rtland cement Clay/port!and cement . Coal fly ash/portland cement Coal fly ash/lime
decrease 0.9 0.4 !.3 1.3
,
Mas s lost 13% 7% 3% 3%
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Calcium Sulphoaluminates Depending on the tricalcium aluminate and gypsum content of ordinary portland cement, the maximum ettrmgite content of the hydrated product at early ages is approximately 15%; cementing systems containing industrial by-products, such as fly ash, with lime and gypsum may have a higher ettrmgite content of, for instance, 20% (Solem and McCarthy, 1992); expansive cements based on calcium aluminates could have an ettringite content of up to approximately 50%. Because of the potential for uptake of contaminants by ettringite mentioned above, use of high ettrmgite systems for waste management has been advocated. However, such systems can be expected to have a different pH response to acid addition than cements based primarily on CSH. Although the pH stability of ettrmgite has not been specifically addressed, some studies, both in the areas of cement and concrete, and waste management, have touched on this issue. Day (1992) cites existence of a crystalline ettringite in the pH range from 11.5 to 11.8, and a non-crystalline phase with a similar composition from pH 12.5 to 12.8. Ghorab and Kishar (1986) found the pH of an ettringite solution to be 11.2, whereas a group of workers based at the University of North Dakota have variously observed pH values from 9.8 to 12, although their most recent work states that ettrmgite is not stable below pH 11 (McCarthy, Hassett and Bender, 1992, Kumarathasan et al., 1990, and Hassett et al., 1989). Literature values of the solubility product, for ettrmgite range from 10-3~ to 10-4~ (Day, 1992, and Deng and Tang, 1994), from which the theoretical pH of a saturated solution may be calculated as ranging between 11.0 and 11.6.
3. M E T H O D S AND MATERIALS
3.1. Preparation of Solidified Wastes The present work exammed the response of five different solidification systems, with and without waste, to acid addition. The chosen cementing systems were: portland cement, portland cement with silica fume, alkali-activated blast furnace slag, lime and coal fly ash, and high alumma cement with lime and gypsum. The formulations used to prepare laboratory batches are labelled 1 to 5 in the heading for Table 2. "W" indicates the batches containing waste. The first four systems were selected for their differing compositions, particularly their Ca/Si mole ratios (shown in the last row of Table 2), and were used to solidify a plating sludge containing heavy metals; the latter is a high ettrmgite system, which was used to solidif)' a hazardous waste incinerator ash. This ash contained high levels of chloride and sulphate and was 69% soluble in water. Gypsum was omitted when the ash was solidified, to cause the sulphate and chloride in the ash to form part of the ettringite matrix. The laboratory batches without, waste were all prepared at a water to cement ratio of 0.4, whereas a water to solid ratio of 0.5 was required for thorough mixing of the samples containing waste. All specimens were cured for 7, 28 and 56 days at 22oC in a moisture chamber before testing, with the exception of the fly ash/lime samples, which were cured at 46oC to accelerate curing.
3.2. Measurement of Acid Neutralisation Capacity After each of the three curing periods, the solidified products were subjected to the Acid Neutralisation Capacity (ANC) test (Wastewater Technology Centre, 1991). This test revolves batch extraction of a series of 5.0 g subsamples of finely ground (<0.2 mm) solidified material with 30 mL of nitric acid for 48 hours, in sealed plastic bottles. The concentration of nitric acid in each subsample is adjusted so that plotting of the final extract pH values yields a titration curve. Total dissolved solids contents were determined by drying aliquots of selected extracts at 105oC, to obtain an indication of the solubility of the solidified waste matrix at different pH's.
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Table 2. Formulations used in acid neutralisation capacity exI)eriments Component CaO SiO2 Percentage of dry mix % % 1 1W 2 2W 3 3W 4 Portland cement 63 21 10 40 80 32 . 0 Silica fume 0.39320 8 0.6 96 Blast furnace slag 33-44 33 . . . . . 92.5 37 38 Sodium metasilicate 0 47 . . . . 7.5 3 80 Class F coal fly ash 3.7 47 . . . . . 20 High calcium lime 74 0 . . . . High alumina cement Gypsum 60 60 60 Metal plating sludge Hazardous waste ash 40 50 40 50 40 50 40 Water mmi m i 3 3 1. 1.4 0.5 0.5 0.5 Ca/Si ratio 4 ..
4W
5W
o
48 12 -
l0 60
-
30
40
50 0.5
-
10 30 -
-
60
40
50
4. RESULTS AND D I S C U S S I O N For each of the above binder systems, titration curves of pH as a function of equivalents of nitric acid added per kg of dr), cementing material were plotted. Selected data which illustrate particular features of these plots are shown in Figures 2 to 6. Total dissolved soli&s concentrations, corrected for the amount of nitric acid added, were used to estimate the fraction of the matrix which was dissolved by acid addition. The soluble fraction of the binder without waste addition has been plotted in Figures 2 to 5. Figure 6 shows the soluble fraction of the ettrmgite-based matrix, both with and without waste addition. For each system, the amount of acid necessary to achieve a pH of 9, and the fraction of the matrix dissolved at pH 9 were determined graphically and have been summarised in Table 3. 4.1.
Portland
C e m e n t
Figure 2 shows the titration curves for the portland cement system with and without addition of the platmg sludge after 56 days of curing. Because portland cement hydrates quickly, the titration curves for the samples cured for 7 and 28 days were not significantly different from those shown. From the pH data observed for pure systems summarised above in Section 2.2, the portland cement system, with a Ca/Si ratio of 3, would be expected to exhibit pH plateaus at 12.3, 11.9 and 9.9. In fact, while it is possible to distinguish plateaus at 12.3, and possibly 11.9, the titration curve also flattens at appro.ximately 10.9, and 9. There is no evidence of a pH plateau at 9.9. Apparently, the presence of impurities in a real cement cause the formation of different CSH phases than are observed in a pure system.
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Figure 3. Titration curve for the portland cement and silica fume binder system, with waste after 56 days curing, and without waste after 7 and 5(3 days curing, showing also dissolution of the portland cement/silica fume mat,rLx as a function of acid addition.
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Figure 4. Titration curve for the blast furnace slag and sodium silicate binder system, with waste after 56 days curing, and without waste before mixing and after 56 days curing, showing also dis~lution of the activated slag matrix as a function of acid addition.
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Figure 5. Titration curve for the coal fly ash and lime binder system, with and without waste after 56 days curing, showing a ] ~ dissolution of the lime/fly ash cement matrix as a function of acid addition.
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Figure 6. Titration curves and matrLx dmsolutlon as a tunctlon ol acIO aomtmn for the high alumina cement and lime binder system, with gypsum, and with hazardous waste incinerator ash after 56 days curing.
811 Table 3. Dissolution of various binder s },stems at pH 9 Matrix components Amount of acid to pH 9 tecl/k~ of dr~ cement) Portland cement 16.0 Portland cement/waste 12.6 Portland cement/silica fume 11.0 Portland cement/silica fume/waste 9.5 Blast furnace slag/sodium silicate 6.2 Blast furnace sla6/silicate/waste 0.5 Coal fly ash/lime 3.6 Coal fly ash/lime/waste 2.8 High alumina cement/lime/~]~sum 8.1 High alumina cement/lime/haz, ash 4.8 i
i
Percent dissolved 50% 15% 35% 15% 10% 5% 7% 5% 1% 45% i
i
As expected, the first two pH plateaus, at 12.3 and 11.9, coincide with low solubility of the matrix, as indicated by total dissolved solids measurements. Matrix solubility appears to increase below pH 11.9. The portland cement system containing plating sludge does not exhibit a pH plateau at 12.3; it appears that the waste has consumed the lime, or altered hydration so that lime was not generated. The pH drops directly from 12.3 to 8, with a slight flattening at pH 10.5. Interestmgly, the fractional solubility of this sample appeared to be much lower than that of the pure cement system. A detailed analysis of the plating sludge was not performed, but it may have contained a high proportion of insoluble components. 4.2. Portland C e m e n t with Silica F u m e Replacement of 20% of the portland cement with silica fume decreased the Ca/Si ratio from 3 to 1.4. Accordmgly, it was expected that the pH plateau at 12.3 would disappear when the lime produced by hydration of the portland cement reacted with the silica fume. Indeed, the pH 12.3 plateau was apparent in the sample without waste at 7 days of curing, as shown in Figure 3, but had disappeared by 28 and 56 days of curing.
From 28 days of curing, the titration curves for the portland cement and silica fume samples were very. similar, with and without addition of the plating sludge. These curves were also similar in shape to the portland cement and plating sludge sample (Figure 2), exhibiting a plateau between pH l0 and 11. The matrix solubility indicated by the total dissolved solid measurements was lower than for the pure portland cement samples, increasing beyond the pH plateau. Again, the solubility of the sample containing waste was lower than that of the sample without waste. 4.3. Activated Blast F u r n a c e Slag The binder system produced by activation of blast furnace slag with sodium metasilicate has a low overall Ca/Si ratio of 0.5. Based on the earlier discussion ofpH data for a pure system, this would predict the coexistence of low Ca/Si CSH and silica gel, but only a relatively small proportion of the slag hydrates, so the Ca/Si ratio of the CSH is higher, and silica gel is not formed.
The pH vs. acid addition data for the blast furnace slag samples for 7 and 28 days resembled the titration curves plotted at 56 days in Figure 4. The curve for the slag system without addition of waste shows an initial pH of 12.5, droppmg to a pH plateau at approximately 9. A titration curve for the unreacted mixture, calculated based on the acid neutralisation capacities of the individual components, has also been plotted in Figure 4. Comparkson of the unreacted and reacted systems clearly demonstrates that a reaction takes place, with the products having a lower initial pH, and a higher and more distmct pH plateau, than the reactants. The total dissolved solids measurements indicate that the solubility of this system is low, even under acidic conditions. However, it does appear that the solubility increases at the start of the pH 9 plateau. The pH 9 plateau may represent CSH
812
coexisting with silica gel, its theoretical position having been altered by the presence of other components in the slag, such as Mg. The plating sludge inhibited the set of the slag. These samples did not develop physical strength, and pH plateaus for reaction products associated with strength development are absent from the titration curve. It seems likely that the plating sludge reacted with the sodium silicate, and prevented activation of the slag. In spite of the lack of hydraulic reactions, the solubility of the slag/waste system was still low.
4.4. Coal Fly Ash and Lime As coal fly ash alone has very little acid neutralisation capacity, and lime would control the pH at 12.3, the pH plateau between 11 and 12 for the fly ash and lime without waste in Figure 5 again confirms that pozzolanic reactions, creating CSH, have occurred. The data shown is for the sample cured for 56 days; the pH plateau in the sample cured for only 7 days was closer to 12, while the titration curve at 28 days was the same as that shown. Again, the Ca/Si ratio was approximately 0.5, but, as was the case for the slag system, a significant proportion of the fly ash has remained unreacted, so that the Ca/Si ratio of the CSH formed is higher. In this case, the development of strength by the samples containing plating sludge indicates that the presence of waste in the system did not inhibit the set. However, the titration curve for the sample containing waste is considerably different, exhibiting a rapid drop from pH 12 to 10.5, and then a linear slope to pH 4, with no discernible plateaus. Examination of total dissolved solids data shows low solubility for this system, with and without waste, although it does appear that the solubility of the pure system jumps when the pH drops below 9, while that of the system containing waste increases steadily as the pH drops.
4.5. High Ettringite System The titration curves for the two calcium sulphoalummate-based systems at 56 days have been plotted in Figure 6. The curves for 7 and 28 days had similar features. Two distinct pH plateaus are visible in Figure 6. The pH plot for the system containing gypsum rather than hazardous waste incinerator ash levels off at pH 12.4 and 10.8. That for the system containing ash levels off at pH 12 and 10. The overall acid neutralisation capacity for the sample containing incinerator ash is lower because the ash contains components which do not participate in the reaction, whereas the gypsum reacted fully. The pH 12/12.4 plateau is attributed to tetracalcium alummate hydrate (4CaO.Al2Oa.13H~O) which would be expected to be the other main hydration product in this system, in addition to ettringite, which results in the plateau at pH 10/10.8. Unreacted lime may also contribute to this plateau. The presence of impurities in the hazardous waste incinerator ash clearly affects the position of the pH plateaus, as compared with the gypsum system. Chloride may be incorporated in the ettringite, or form calcium chloroaluminate (3CaO.AhOa.CaCl:. 10H20). Other researchers have found this product to be stable between pH 11 and 12.5 (Ben Yair, 1971). Total dissolved solids measurements show the solubility of the ettrmgite system containmg gypsum rather than incinerator ash to be extremely low. The solubility data plotted in Figure 6 show that the use of ash rather than g3.j)sum in the matrix increases iLs solubility drastically. If it is assumed that the solidification process does not affect the solubility of the ash (69%), a matrix solubility due to ash dissolution of 41% may be calculated, based on 60% of ash in the matrix. In fact, a slightly higher matrix solubility is observed.
813
5. CONCLUSIONS Except for the pH plateau at 12.3 caused by excess lime, the pH plateaus observed in response to acid addition for real binder systems do not exactly correspond to those observed for the pure CaO-SiO2-H20 system. Different hydraulic cements exhibit different pH plateaus, which probably reflect differences in the structure and composition of the CSH formed. In general, plateaus seem to occur at pH levels lower than those anticipated. The addition of waste to a cementing system also appears to lower the pH plateau. Solubility of all matrices was low above pH l l.5; depending on the binder system, matrix dissolution appeared to increase at pH values ranging from 9 to 11.5. The following conclusions are drawn for the specific systems studied: 9
For portland cement, the amount of acid required to achieve a pH of 9 was 16 eq/kg cement, but greater than 10% matrix dissolution was observed at an acid addition of 6 eq/kg cement, and a pH of less than 11.5. 9 Addition of silica fume to portland cement appeared to lower the pH at which significant matrix dissolution was observed to approximately 10.5, with an acid addition of 6 eq/kg cement. 9 The activated blast furnace slag binder system showed increased matrix dissolution at pH 9, with an acid addition of 4 eq/kg cement, but overall low solubility over the pH range from 12.5 to 5. 9 The solubility of the coal fly ash and lime system was also low over the pH range from 12.5 to 4. An increase in solubility was observed after addition of 4 eq/kg of acid/kg cement, below pH 9. 9 The solubility of a high ettringite cement system, using gypsum to form calcium sulphoaluminates, is very low at pH values above 10, however, use of a high sulphate and chloride waste in place of gypsum results in a matrix with very high solubility.
6. R E C O M M E N D A T I O N S AND S U G G E S T I O N S FOR F U R T H E R W O R K 9
9
9
9 9
Rather than using a fixed pH as the criterion for performance in an acid neutralisation capacity test, it may be advisable to consider the pH at which a particular waste/bmder system appears to undergo a significant increase in solubility. As waste materials are generally not alkaline, adjustment of binder formulations containing alkaline components such as lime or sodium silicate to compensate for their consumption by the waste material should be investigated. From tests on ground samples, the actual effect of acid addition on the structural integrity of a solidified matrix is not easily apparent, because disintegration of structural matrix components may occur without chemical dissolution. Also, the physical structure of the matrix, including density and porosity, as well as its chemical stability, will influence its acid resistance. Thus, the acid resistance of monolithic samples should be investigated. Chemical changes continue to occur in cements over long time periods, therefor, the acid neutralisation capacities of different binders at ages of several years should be investigated. While micromorphological studies may confirm inclusion of waste components in calcium sulphoaluminate phases, leaching studies should be conducted to evaluate the stability of these compounds in the environment.
7. R E F E R E N C E S
Atkinson, A., Goult, D.J., and Hearne, J.A. (1985), "An Assessment of the Long-term Durability of Concrete in Radioactive Waste Repositories", Material Research Society, proceedings of symposium, Vol. 50.
814
Bambauer, H.U., Gebhard, G., Holzapfel, T., Krause, C., and Willner, G. (1988), "Schadstoff-Immobilisierung in Stabilisaten aus Braunkohlenaschen und REA-Produkten - I. Mmeralreaktionen und Gefiigeentwicklung: Chlorid-Fixierung", Fortschritte der Mineralogie, Vol. 66, pp. 253-279. Ben-Yair, M. (1971), "Studies on the Stability of Calcium chloroaluminate", Israel Journal of Chemistry, Vol. 9, pp. 529-536. Cot6, P.L. (1986), Contaminant Leaching from Cement-based Waste Forms Under Acidic Conditions, Ph.D. Thesis, McMaster University, Hamilton, Ontario. Day, R.L. (1992), The Effect of Secondary Ettringite Formation on the Durability of Concrete: A Literature Analysis, University of Calgary, Department of Civil Engineering, Research Report No. CE 92-2. Deng, M. and Tang, M. (1994), "Formation and Expansion of Ettringite Crystals", Cement and Concrete
Research, Vol. 24, pp. 119-126. Ghorab, H.Y., and Kishar, E.A. (1986), "The Stability of the Calcium Sulphoaluminate Hydrates in Aqueous Solutions", Proceedings of the 8th International Congress on the Chemistry of Cement, Rio de Janeiro, Vol. 5. pp. 104-109. Greenberg, S.A. and Chang, T.N. (1965), "Investigation of the Colloidal Hydrated Calcium Silicates. I]. Solubility Relationships in the Calcium Oxide-Silica-Water System at 25~ Journal of Physical Chemistry, Volume 69, Number 1, pp. 182-188. Grutzeck. M., Benesi, A., and Fanning, B. (1989), "Silicon-29 Magic Angle Spinning Nuclear Magnetic Resonance Study of Calcium Silicate Hydrates", Journal of the American Ceramic Society, Vol. 72., No. 4, pp. 665-68. Hassett, D.J., Pflughoeff-Hassett, D.F., Kumarathasan, P., and McCarthy, G.J. (1989), "Ettringite as an Agent for the Fixation of Hazardous Oxyanions", Proceedings of the Twelfth Annual Madison Waste Conference on Municipal and Industrial Waste, Madison, WI, September 20-21. Kumarathasan, P., McCarthy, G.J., Hassett, D.J., and Pflughoefl-Hassett, D.F. (1990), "Oxyanion Substituted Ettringites: Synthesis and Characterisation; and their Potential Role in Immobilisation of As, B, Cr, Se and V", Materials Research Society Symposium Proceedings, Volume 178, pp. 83-104. McCarthy, G.J., Hassett, D.J., and Bender, J.A. (1992), "Synthesis, Crystal Chemistry and Stability of Ettringite, A Material with Potential Applications in Hazardous Waste Immobilisation", Materials Research Society Symposium Proceedings, Vol. 245, pp. 129-140. Solem, J.I~, and McCarthy, G.J. (1992), "Hydration Reactions and Ettringite Formation in Selected Cementitious Coal Conversion By-Products", Materials Research Society Symposium Proceedings, Vol. 245, pp. 71-78. Taylor, H.F.W. (1993), "Nanostructure of C-S-H: Current Status", Advanced Cement-based Materials, Vol. 1, No. 1, October. U.S. Federal Register (1986), Appendix I, Part 268, Toxicity Characteristic Leaching Procedure (TCLP), Vol. 51, No. 216, November 7. Wastewater Technology Centre (1991), Proposed Evaluation Protocol [or Cement-Based Solidified Wastes, Environment Canada Publication EPS 3/HA/9, Ottawa.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
815
C o n t a m i n a t e d soil - c e m e n t stabilization in a d e m o n s t r a t i o n project. J. van Leeuwen, A. Pepels and G. van Emst Gemeentewerken Rotterdam (Public Works), Engineering Division. P.O. Box 6633 3002 AP Rotterdam, the Netherlands
Abstract
Public Works Rotterdam has carried out a feasibility study on the application of contaminated soil-cement stabilization layers (ref. 1). Conclusion was that contaminated soil can be benificially used as a soil-cement stabilization in a road construction. As a follow-up, a full scale demonstration project (1.000 m 2) was realized in 1995: a . (sub)base of a soil-cement stabilization layer with soil that should have been landfilled (ref. 2). Samples of the soil-cement layer have been tested on both environmental and physico-mechanical aspects. Conclusions were that the base is environmentally safe and no isolating provisions are required. The compressive strength is sufficient according to the standard of 1.5 N/mm 2. However, the need for licences and an environmental impact assessment and the procedure time involved will be the bottle-neck for the on-site immobilization of (hazardous) waste materials in constructions.
Introduction
In the city of Rotterdam large quantities of polluted soil are excavated as a result of soil sanitation and/or infrastructural works. The policy aims of the government and the municipality of Rotterdam are in the first place the beneficial use, secondly to clean and thirdly to landfill the polluted soil, depending on the type of soil combined with the type and content of contaminants. In infrastructural works, the base is mostly constructed of primairy sand - cement stabilizations. The question was whether the (clean) sand could be replaced by polluted, sandy soils. The policy of beneficial use of the municipality of Rotterdam is being implemented by stimulating the use of contaminated soil in infrastructural works. Public Works Rotterdam has carried out a feasibility study on the application of contaminated soil-cement stabilization layers (ref. 1). Conclusion was that contaminated soil can be benificially used as a soil-cement stabilization in a road construction. The question was whether the immobilisation of strongly polluted soil is executable on a full-scale project. In 1993 the national government started a tender-program on immobilization: Technology 2000 - Immobilization. The goal of this program is to stimulate the development and implementation of immobilization techniques on 22 selected hazardous materials, including non cleanable polluted soil. As a follow-up of the feasibility study and within the framework of the program . Technology 2000 - Immobilization, a full scale demonstration project was realized in 1995: a (sub)base of a soil-cement stabilization layer with soil that should have been landfilled (ref. 2). The stabilization is the base of the temporary storage site of DOPNOAP, a landfill for polluted soil in Rotterdam. The goals of this project were: the demonstration and testing of the beneficial use of polluted soil immobilized with cement as a road base. to check the feasibility and practical aspects of the immobilization process.
816
to determine the performance of the soil/cement stabilization under practical use. Figure 1 shows a cross-section of the construction. AFPHALT-LAYER
\ SUBSOIL
CONTAMINATEDSOILC~k~_~E MENT STABILIZATION
Methods and materials
The project consisted of the following steps (see figure 2): 1) 2) 3) 4) 5) 6)
the selection of a batch of polluted soil, which has to be landfilled (physical and environmental criteria) the characterisation of environmental and physical quality of the soil *content of contaminants and leaching behaviour *grain-size, pH, organic matter the determination of the optimal cement addition. the determination of practical aspects of the construction *mixed in place *mixed in plant the sampling of cylinders after execution and hardening the determination of environmental and physical quality of the soil/cement stabilization *compressive strength, durability. *leaching behaviour
In order to landfill the polluted soil, a large number of batches are registered at the Soiland Residuebank of Rotterdam. Some of these batches are suitable for immobilization. As a result of the feasibility study some initial criteria were derived for the selection of batches of soil. * organic matter < 10% dry weight * fines (< 2 um)< 5% dry weight * moisture content < 10% The batch which best met these criteria, was selected and stored at the site. Then samples were taken in order to characterize the batch and several tests were carried out in the laboratory. The leaching of the soil has been determined according to NEN 7343. One of the most important questions was the optimal addition of cement and the question whether other additives had to be added. The physical tests were carried out according to the Dutch RAW-standards. After 28 days of hardening, the compressive strength of the proctors was determined. If they met the standard (5.0 N/mm2), the proctors were tested in the tank leaching tests (NEN 7345), to determine the leaching of the immobilized product.
I
batch 1
siteinvestigation mixed in place / mixed in plant
initial criteria
selection and storage
Figure 2: The scheme of the project
characterization and investigation of cement content
execution
afler 28 days
I
+
afler 8 months
testing of : environmental and -physic0 mechanical quality durability (only 23 days)
818
When a batch is qualified as suitable and a large scale infrastructural work is planned, the go/no go decision can be made. The local geology of the site determines the method of realisation ('mixed in place' or 'mixed in plant'). After the realisation samples were taken in order to determine the environmental and physico-mechanical quality of the immobilized product. After respectively 28 days and 8 months of hardening, samples (cylinders) were taken out of the base. The cylinders were tested on both physico-mechanical and environmental aspects. Compressive strength, durability and leaching of the cylinders were determined after 28 days. After 8 months the compressive strength and leaching of the samples were determined again. Durability is defined as the resistance agains wetting, drying and freezing and is investigated by wet-dry and freeze-thaw cycles. The cylinders were exposed to wet/dry and freeze/thaw-cycles, according to Dutch RAW-standards. After those cycles the compressive strength has been measured again.
Results and discussion Characterisation of the batch
A batch of about 500 tons of soil, contaminated with heavy metals, mineral oil and poly aromatic hydrocarbons (pah's) and supposed to be landfilled, was selected for the base. in table 1, the results of the most critical components are shown. In the national Dutch legislation for the use of raw and secundairy materials (Bouwstoffenbesluit, the Dutch Building materials Decree) criteria for the use of soil are included. For both granular and immobilized materials standards on content and leaching have been developed. Table 1:
environmental quality
component
content
Leaching
Range
Trigger value
L/S=10
Copper (mg/kg)
95,2- 181
87,5
0,0660
0,58
Lead (mg/kg)
46,5- 125
325,5
0,0500
1,6
Zinc (mg/kg)
124 - 274
278,0
0,327
3,3
Trigger value
According to the mentioned legislation for benificial use, the soil is not suitable because of the high content of copper and the presence of oil and pah's. However, for immobilized soil and/or other materials, there aren't any standards for the content of heavy metals (and other anorganic components). On the other hand, the products or materials have to meet the standards for leaching for these anorganic components. For the organic components, no leaching tests have yet been developed, so the standards are based on the content of these components. From table 1 can be concluded that the leaching of copper, lead and zinc meet the standards for unisolated application. The addition of cement should be sufficient to produce a product which is environmentally safe.
819
Next, the physico-mechanical aspects of immobilization were determined. The results are presented in table 2. Table 2:
physical quality and demand of cement
Moisture content Fineness modulus (test 18) pH-test (test 23.2) Suitability test (test 22.1 )
17,6 % 21,6 % 1,50 1,90
10% cement 12,11 12,13
10 % cement 10 % cement 10 % cement
1,6 N/mm = 1,6 N/mm = 1,5 N/mm 2
Determination content of cement (test 22.2) 8 % cement 10 % cement 12 % cement
2,7 N/mm = 5,0 N/mm = 3,7 N/mm 2
The optimal content of cement is derived from the results in table 2 and the fea.sibility study with several batches of polluted soil (ref. 1). In the feasibility study an emperical relation between moisture content and cement content has been observed. It is concluded that the water-cement ratio should be 1:1. The moisture content of the batch was in the range 17,6 - 21,6%, so, according to the feasibility study the cement-content should be about 19%. The standard for the compressive strength of proctovs is 5,0 N/ mm 2. From table 2 a cement-content of 10% was deduced. In the demonstration-project, the mean of these two contents, 14,5% of cement, was used.
Practical aspects of realisation
The most easy way to realize the soil-cement stabilization layer is the mixed-in-place technique. Because of the instable subsoil at the site, it was not possible to get a proper mixture and condensation of the soil-cement. By using the mixed-in-plant proces, it was possible to get a good mixture of soil, cement and water. Next, the mix was put into the site and rolled with a road-roller and covered with an asphalt layer. The area of the soilcement stabilization is about 1.000 m 2.
Physico-mechanical and environmental aspects
The results of the physico-mechanical tests are displayed in table 3. According to Dutch standards the compressive strength of cylinders should at least be 1,5 N/ram 2. From tabel 3 it can be concluded that the compressive strength of all samples meets this
820
standard. From our experiences with traditional sand-cement stabilizations, the distribution in the results is in the same order of magnitude. After wet-dry and freezethaw cycles, the cylinders keep their strength, so the sustainability is judged to be good. Table 3:
physico-mechanical aspects
physico-mechanical aspect
Volumic weigth (kg/m 3,)
average
median
minimum
maximum
28 d
8m
28 d
28d
8m
28d
2000
1972
1998
1933
1935 2079
2,8
1,6
3,0
8m 2000
(n=16) (n=3) Comp.strength (N/mm 2 )
3,1
4,7
6,2
7,1
(n=16) (n=3) After freeze/thaw-cycles (n=8) - comp. strength
3,5
2,9
1,6
8,2
- volumic weigth
2013
2007
1960
2090
After wet/dry-cycles (n=8) - comp. strength
4,5
3,2
1,9
7,8
- volumic weigth
1944
1919
1835
2043
The environmental research consisted of tank leaching tests. After 28 days and after 8 month 3 samples were tested. These tests should be carried out for a period of 64 days, according to NEN 7345. As the granular soil already met the leaching standards, the leaching was tested for control. Most of the leaching of the components takes place in the first 4 days. A test for the period of 4 days normally gives a sufficient indication of the leaching behaviour. In table 4 the results of the leaching tests are given. The leaching is much below the trigger values. The base is environmentally safe and no isolating provisions are required. Table 4:
environmental aspects, range of leaching in mg/kg
sample after 28 days (leaching during 4 days) sample after 28 days (leaching during 64 days) sample after 8 month (leaching during 4 days) triggervalue reuse (leaching during 64 d)
Cupper
Lead
Zinc
1,93- 2,66
1,92- 3,26
1,89- 1,92
4,96
3,26
3,24
3,10 -4,36
1,74- 1,81
1,74- 1,94
51
120
200
821
Legislation
In the Dutch legislation, the Environmental Management Act, a licence is required for the treatment of waste-materials. For the treatment of hazardous waste, an Environmental Impact Assesment has to be submitted. In that way, the on-site immobilization of waste materials in case of the beneficial use in infrastructural works, also needs a licence. In this demonstration project the environmental licence to perform the project was no obstacle, because it was part of a larger project which already demanded a licence. In following immobilization projects, permission can be an obstacle because of the long procedure time and the large amount of information which should be submitted. In many cases an environmental impact assessment study will be demanded.
Costs
The costs of the soil-cement stabilization are comparable to the costs of a standard sand-cement stabilization and a base of granular, broken debris (so-called 'Repak'). The positive gain of the benificial use of the contaminated soil are the costs of disposal in the landfill (about 80 Dutch Guilders a ton, which do not have to be paid). The negative gains are the costs of additional cement, additional handling, environmental and physico-mechanical research. The differences in costs of the soil-cement stabilization layer are about f 1 ,-/m 2 in relation to repak and f 40,-/m 2 in relation to sand-cement. Soil-cement is the economical most benificial option. The use of cement is a critical aspect for an economic advantage. A lower content of cement will give a substantial reduction of the costs. The cement content in the soilcement stabilization is 4 to 6% more than in a regular sand-cement stabilization. In case of a small batch of polluted soil the costs of environmental and physicomechanal research are relatively high. The use of contaminated soils for a cement stabilization in a (road-) base seems to be economical beneficial when the total amount of soil is at least 500 tons.
Conclusions
The conclusions of the project were: The leaching of metals is in accordance with Dutch standards. The base is environmentally safe and no isolating provisions are required. The compressive strength is sufficient according to the standard of 1.5 N/mm 2. The distribution in the results is equal to those of sand-cement stabilizations. The durability is judged to be good. Considering the costs of landfilling, a minimal batch size of 500 tons of soil is economical beneficial, when it is used in infrastructural works. The site, including its base, is still in use, and no problems are being encounterd. In future projects a critical review on the optimal content of cement is recommended. The results of standarized tests in a laboratory with each batch of soil are preferable to the use of the empirical relation of the water-cement ratio. The need for licences and an environmental impact assessment will cause problems for the on-site immobilization of hazardous materials in constructions. Certification of processes and contractors (quality management) are recommended as important alternatives.
822
References 1:
-
2:
-
Kroes P.J. and J. van Leeuwen, Contaminated Soil Cement Stabilizations for Application as a Construction Material in "Environmental Aspects of Construction with Waste materials" (WASCON-proceedings), (1994). Gemeentewerken Rotterdam, Ingenieursbureau Milieu, Immobilization of non-cleanable soil for the beneficial use as a foundation (in Dutch), (1996).
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997Elsevier Science B.V. All fights reserved.
823
S T A B I L I Z A T I O N OF A G A L V A N I C S L U D G E B Y M E A N S OF C A L C I U M SULPHOALUMINATE CEMENT R. Cioffi 1, M. Lavorgna 2, M. Marroccoli 3 and L. Santoro 2 I Dipartimento di Ingegneria dei Materiali e della Produzione, Universith di Napoli "Federico II", P.le Tecchio 80, 80125 Napoli, Italy. 2
Dipartimento di Chimica, Universit~ di Napoli "Federico II", via Mezzocannone 4, 80134 Napoli, Italy.
3
Dipartimento di Ingegneria e Fisica dell'Ambiente, Universith della Basilicata, via della Tecnica 3, 85100 Potenza, Italy.
ABSTRACT A solid waste containing heavy metals from galvanic treatment has been stabilized by means of a binding matrix containing [3-2CaO.SiO2, 4CaO-3AI203.SO 3 and CaSO 4 able to give calcium silicate and trisulphoaluminate hydrates upon hydration. Experiments have been carried out with mixtures containing up to 60% waste under three different points of view, as follows. The influence of the waste on the technical properties of the stabilized products, the leaching behaviour under four different conditions and the effect of the leaching medium on the binding matrix have been studied.
1. INTRODUCTION In the light of the most recent directives of European Community regarding solid wastes management, stabilization/solidification processes will play in the near future a more and more important role because it will be only allowed to dispose of inert or stabilized residues. These directives have been acknowledge by the Italian Government by means of a specific act dated 5 February 1997. This act gives high priority to the development of technologies addressed towards recycling and reuse of solid wastes as well as recovery of raw materials and energy from the wastes themselves. Among the available stabilization technologies, the most frequently applied are cement-based. They rely on the formation of a calcium silicate hydrate matrix and make use of a number of systems such as ordinary portland cement, blast furnace slag and mixtures of lime and coal fly-ash (or other pozzolanic materials). A wide range of residues is currently stabilized by means of these inorganic processes, that is industrial solid wastes containing heavy metals, nuclear wastes, municipal solid wastes incineration ashes, sludges from wastewater treatment
824 plants and so on. In addition, this technology can also be applied to the stabilization of contaminated soils and sediments. In these cases the addition of bentonite can be helpful to reduce the amount of binding/stabilizing matrix. Once a specific stabilization process has been selected as the most appropriate for a specific solid waste, a number of questions must be answered in order to assess its environmental feasibility. These questions belong to three different technical fields, as follows. First of all, some interactions will arise between the binding/stabilizing matrix on one side and the waste components on the other. These interactions must be properly studied in order to understand how the binder performance changes in response to the waste admixture. This is particularly important if the ultimate scope is the reuse of the final product. Furthermore, the process must be environmentally acceptable which means that the release of contaminants due to leaching must be studied in different conditions in order to get an understanding of what the long term behaviour of the stabilized product in the environment will be. The test conditions in which leaching should be carried out will have to be designed keeping in mind the ultimate scope of this aspect of the of the research. Last but not least, and in addition to what just stated, emphasis should be given to the interactions which may arise in the environment between leaching media and the binding matrix itself. This aspect has been generally disregarded by the researchers in the field, but is particularly important because the long-term exposure to leaching media may result in complete release of contaminants if the stabilizing matrix undergoes substantial modification. In this paper a solid waste from galvanic treatment containing heavy metals, mainly Cd, Cr, Cu, Ni, Pb and Zn, has been stabilized by means of a novel cementitious matrix based on calcium silicate fl-2CaO.SiO 2 and sulphoaluminate 4CaO'3AI203"SO 3. Upon hydration this matrix forms calcium silicate hydrate and calcium trisulphoaluminate hydrate (ettringite). It was tested in previous work for both physico-mechanical and stabilizing properties proving to be suitable for the application under study. The study referred to in this paper deals with physico-mechanical properties of stabilized samples, release of metals in different leaching tests and matrix behaviour during leaching.
2. EXPERIMENTAL The components fl-2CaO.SiO 2 and 4CaO.3AI203-SO 3 of the binding matrix were synthesized in the theoretical ratio 1:1.5 by firing a raw mixture of CaCO 3 (45.37%), bauxite fines (31.09%), zeolitic tuff (13.27%) and CaSOn-2H20 (10.27%) at 1200~ for 90 min. The binder was obtained by adding anhydrous C a S O 4 to the fired mixture in the ratio 1:2.5. The chemical composition of the waste, bauxite fines and zeolitic tuff is reported in Table 1. Binder-waste mixtures were prepared containing 0 (pure binder), 20, 40 and 60% of waste and hydrated at 25~ 100% RH, and water/solid ratio equal to 0.4, 0.46, 0.5 and 0.56 for the systems containing 0, 20, 40 and 60% waste, respectively. These values were chosen in order to get constant workability. The hydration time ranged between one hour and 28 days (672 hours). Small samples (about 3 g) of each system have been used to study the kinetics of hydration. This part of the study has been carried out by determining the amount of chemically combined water by ignition at 1000~ for the time required to reach constant weight. In addition, the formation of hydrated products has been monitored by differential thermal analysis (DTA).
825 Table 1 Chemical composition of waste, bauxite fines and zeolitic tuff (wt%) LOI*
Waste 34.60
SiO2 A1203 CaO
48.12 7.62
K20 Na20 SO3 Fe203 MgO Cr203 NiO CdO ZnO CuO PbO MnO *LOI = Loss on ignition
Bauxite fines 25.40 4.70 51.50 0.02 -
0.15 0.44 6.57 0.98 0 04 0.04 0.02 001 0.03
15.3 0.04
Zeolitic tuff 9.93 52.93 17.21 3.54 7.26 2.97 0.13 3.71 1.44
_
Samples of the three mixtures containing 20, 40 and 60% of waste have been submitted to the following three leaching tests: (a) the dynamic TCLP test [ 1],with pH 4.94 acetic acid/sodium acetate buffer, liquid/solid ratio equal to 20 ml/g and leachant renewals at 1, 3, 8, 14, 24, 48, 96, 168, 376, 672 and 1344 hours on monolithic cylindrical samples dxh=2x3 cm2; (b) the same test as (a) but with pH 3.86 CO2-saturated solution instead of acetic buffer and renewals up to 672 hours; (c) the static US-EPA test [2] with pH 5 acetic acid solution and ratio liquid/solid equal to 16 mug on granulated sample (size < 9.5 mm) and (d) the availability test [3] carried out on 3 g of pulverized sample (size < 180 ~tm) with 150 ml of pH 7 nitric acid solution for 3 hours plus additional 150 ml of pH 4 HNO 3 solution for 3 hours. Following these tests, the leaching solutions have been analysed by means of atomic absorption spectroscopy and the solids before and after leaching have been characterized by means of DTA and scanning electronic microscopy (SEM).
3. RESULTS AND DISCUSSION Figure 1 shows the amount of normalized chemically combined water. The absolute values have been divided by the fraction of binder present in each mixture to get the normalized values. It is seen that only the mixture containing 20% of waste behaves like the pure binder, while in the cases of the 40 and 60% mixtures the waste does not simply dilutes the binder but its components inhibit the binder hydration. This effect increases as the content of waste in the
826 mixtures increases. 40 The hydrated samples have been t~ submitted to DTA -~ 30 stowing that in all the c d~ systems the main E o hydration product is .L~20 ettringite. As in similar [] Pure binder systems studied O 20% waste previously [4-6], other A 4 0 % waste 4 1 0 V 60% waste hydration products are calcium silicate hydrate, aluminium hydroxide I 0 ~l I I I I I I gel and traces of 672 13 7 1424 72 168 336 calcium monosulphate Time of hydration, hours (root scale) hydrate. Significant Figure 1. Normalized chemically combined water for the differences were only pure binder and the three mixtures binder-waste. observed at short hydration times as shown by the thermograms of Figure 2, relative to 1 hour hydration. In this figure the thermogram relative to the pure binder shows the main endotherm at 82~ that reveals the presence of ettringite. The three minor endotherms at 128, 205 and 260~ are related to dehydration of CaSO4"2H20, calcium monosulphoaluminate hydrate and aluminium hydroxide gel, respectively. The thermograms / / '~J/ /~ 60% relative to the systems containing 20, 40 and 60% of waste show that neither ettringite, nor calcium monosulphate hydrate form at 1 hour hydration when the system contains the waste. In these thermograms the endotherms related to dehydration of CaSOn'2H20 show that the waste catalyses the conversion of / \ G 0% anhydrous CaSO 4 to CaSOn-2H20, according to previous findings relative to similar systems [7]. The endotherms relative to , I i I i 1 t 1 , I dehydration of aluminium hydroxide are of 0 100 200 300 400 500 Temperature, ~ higher intensity in the cases of the wastebinder mixtures because this compound is simultaneously a hydration product and a Figure 2. Thermograms of samples aged component of the waste. The particular 1 hour. E: ettringite G: gypsum; M calcium baseline shape of the thermograms relative to monosulphate hydrate; A: aluminium the waste-binder mixtures is due to the hydroxide.
827 presence of waste. The values of unconfined compressive strength are 25 MPa for the pure binder and 21, 16 and 2 MPa for the mixtures containing 20, 40 and 60% of waste, respectively. These values are such that recycling the stabilized products in the field of building materials is possible for waste content up to 40%. The value of 2 MPa observed for the mixture containing 60% waste exceeds the value of 0.44 MPa recommended by Stegemann and Cot6 for segregated landfill disposal [8]. Figures 3 and 4 report the results of the 100 dynamic leaching tests carried out with the ~ Ca 10 acetic buffer and CO 2saturated solution, respectively. Table 2 ---_-2_-2 ....... shows the results of US,r _. . . . . -o---N~ O EPA leaching test and 0.1 availability test. The [] 20% waste O 40% waste data are relative to the Zk 60% waste 0.01 metals Cd, Cr and Ni, rj whose presence is of greater environmental 1 i I I 0.001 concern, and are 0 500 1000 1500 Time, hours expressed as percentages of the initial Figure 3. Results of TCLP test. quantity present in the sample. These results show that the leaching 10 Cd behaviour depends strongly on the nature of the metal as Cr, Ni and ::::::::::::::::::::::::: ............[] ................... Cd are released in -~.. ]~: ..[]..... steeply increasing
~
.
.
.
.
.
.
--lk amount in each test. The 0 " I ~) IE, P ! .....O -,a- . . . . . _Nj. . . . . . . .. . . ~ . . _ .-ZX--. . . . . . _ ..... chemical nature of the leaching medium has ~0.01 ~ V1 20% waste also a strong effect ~' O 40% waste inasmuch as quite larger amounts are released in TCLP test compared to 0.001 0 150 300 450 600 750 CO2-saturated solution Time, hours test. To this regard, it is important to point out Figure 4. Results of CO2-saturated solution leaching test. that pH does not have the effect that one would expect. In fact each step in the CO2-saturated solution test starts with a pH value of 3.86 and ends at about 5. These values, compared to the value of 4.94 of the acetic buffer make clear that pH is not the most important factor that characterize the leaching
828 medium. Table 2 Results of US-EPA leaching test and availability test (wt%) Type of test and waste content US-EPA 20%
Availability
40%
60%
20%
40%
60%
Cd
2.29
1.78
3.54
58.93
56.96
55.66
Cr
0.16
0.05
0.05
0.70
0.47
0.56
Ni
0.30
0.22
0.42
17.47
10.76
10.37
The results of Figures 3 and 4 show that the percentage of metal released increases with the amount of waste in the mixtures and this means that the binder does not simply physically segregate the waste from the attack of the leaching media, but that chemical interactions take place between the binder and the waste components whose extent decreases as the binder content decreases. Finally, the data of Table 2 confirm that the results of a leaching test depend strongly on the chemical nature of both the metal and the leaching medium, and also on the physical nature of the solid sample. The effect of the leaching medium on the binder in the stabilized samples has been studied only in the two dynamic tests carried out with the acetic buffer and CO2-saturated solution. The monolithic cylindrical samples submitted to these tests underwent significant modification which could be evidenced by cutting the samples themselves orthogonally to the cylinder axis. In this way it was possible to distinguish between an external leached layer of increasing thickness and an internal unleached shrinking core. The thickness of the leached layer increased with time and with the amount of waste in the mixture and reached the values of about 2, 3, and 7 mm at the end of both tests in the cases of 20, ._._b_b~ 40 and 60% waste, C respectively. Figure 5 shows the results of DTA I i I I i I carried out on samples 0 200 400 600 800 Tempcrature,~ of the leached layer and the unleached inner core Figure 5. Results of DTA carried out on samples from in the case of the external layer (a) and inner core (b) of system containing mixture containing 40% 40% waste after CO2-saturated solution leaching test.
829 waste submitted to the CO 2 -saturated solution test. It is seen that ettringite undergoes substantial decomposition to gypsum, aluminium hydroxide and calcium carbonate. This observation can also explain why lower amounts of metals are leached in the CO 2 saturated solution test. The formation of metal carbonates can limit metal ions solubility even if the pH is lower than in the TCLP test. This chemical modification of the system implies also substantial morphological modification, as it can be seen in Figure 6 where two micrographs of samples of the external leached layer and the inner unleached core are shown. The system of Figure 6 is that containing 20% waste submitted to TCLP test. The needlelike ettringite crystals visible in micrograph (a) completely disappear after leaching, as seen in micrograph (b). This observation, relative to the sample submitted to TCLP test, is not in contrast with the results that ettringite decomposes in CO2-saturated solution giving calcium carbonate among the other products. In fact, in a previous work carried out on systems similar to those studied in this paper [5], the acetic buffer attack caused ettringite decomposition into calcium acetate pentahydrate, gypsum and Figure 6. SEM micrographs of samples of external layer (below, aluminium hydroxide. b) and inner core (above, a) of system containing 20% waste after TCLP test.
830 4. CONCLUSIONS The results of the experiments can be summarized as follows with respect to the three aspects of the study emphasized in the introductory part. From the point of view of binder performance, the addition of waste to an extent greater than 20% inhibits the hydration process. On the other hand, mechanical performance is reduced starting from 20% waste addition. The leaching behaviour is strongly influenced by the chemical nature of both the metal and leaching medium. Other parameters that greatly influence metal release are the physical nature of the stabilized samples and type of liquid-solid contact. Of the two tests carried out under the same conditions, namely TCLP an CO2-saturated solution, the latter gives rise to lower amounts of metals released. Finally, it has been found that the attack of the leaching medium to the stabilized samples causes severe modification of the binder. This modification has been characterized from both the chemical and morphological points of view, showing that ettringite decomposes to gypsum, aluminium hydroxide and a calcium compound. The possibility that metal carbonates precipitate during the CO2-saturated solution test can explain the lower metal ions release observed in this test in comparison to TCLP test.
REFERENCES 1. 2. 3. 4. 5. 6. 7. 8.
Toxicity characteristics leaching procedure, Federal register, Vol. 51, No. 261 (1990). U.S- EPA Test method for evaluating solid waste, SW-846. Washington, DC: Office of Solid Waste and Emergency Response (1986). NEN 7341 (formerly NWN 2508), Determination of leaching characteristics of inorganic components from granular (wastes) materials. NNI, Delft (1993). J. Beretka, B. de Vito, L. Santoro, N. Sherman and G.L. Valenti, Cement and Concrete Research, 23 (1993) 1205. V. Albino, R. Cioffi, M. Marroccoli and L. Santoro, Journal of Hazardous Materials, 51(1-3)(1996)241. R. Berardi, R. Cioffi and L. Santoro, Journal of Thermal Analysis, in press. V. Albino, R. Cioffi, L. Santoro, and G.L. Valenti, Waste Management and Research, 14 (1996) 29. J.A. Stegemann and P.L. Cot6, Science of the Total Environment, 178(1-3) (1996) 103.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
831
Reuse of Secondary Building Materials in Road Constructions T. Berendsen The Environmental Engineering Department of Public Works Rotterdam, P.O. Box 6633, 3002 AP Rotterdam, The Netherlands.
Summary
Slags, asphalt and coal-ash, have long been used as secondary building materials. They are released when roads are reconstructed. Their reuse is presently prescribed in the Provincial Rules 'Working with Secondary Building Materials'. Mid1998, a new national decree will become effective: the 'Building Materials Decree'. It lays down the guidelines for design, realisation and maintenance of constructions, aiming at avoiding contamination of the soil by secondary building materials. The implementation of this policy requires road constructors to adopt a new approach. They have to do environmental research, and pay attention to design, isolation measures, maintenance and possible monitoring measures. In practice this means that adhering to a schedule concerning environmental research is essential. The civil engineer needs to implement environmental aspects into his basic planning, design and realization process. The Environmental Engineering Division of Public Works Rotterdam has developed a working model which incorporates the Building Materials Decree into the current engineering planning process. This working model is an instrument used by the Port of Rotterdam.
1.
INTRODUCTION
The city of Rotterdam produces large quantities of waste. Limited space, not only in Rotterdam but also in the rest of the Netherlands, prohibits waste material landfilling. So when secondary building materials, such as slags, asphalt, coal-ash and fly-ash, are released, they are usually reused. The municipality of Rotterdam also stimulates the reuse of building materials, because it saves on new raw materials and on the increasing tariffs for landfilling. Project leaders ought to keep in mind this economic benefit. Technical road reconstructions will confront engineers and workers with secondary building materials from the past, both in foundations and asphalt layers. Handling of these materials is outlined in rules and regulations, such as the 'Building Materials Decree' and the memorandum 'Working with Secondary Building Materials'.
832
In this paper special attention will be paid to secondary building materials because nearly 10 % of road foundations in the city and port of Rotterdam contain secondary building materials, a quantity large enough to deserve special attention.
2.
PROBLEM ANALYSIS
How to put into practice the policy of reusing secondary building materials when reconstructing roads? When we studied the rules and regulations in view of the practice, we found a few bottlenecks. The most obvious one is that the quality of the secondary building materials released during road reconstructions is unknown. Furthermore, the necessary contamination- and leaching tests take a lot of time and cause stagnations, and in some cases involve unexpectedly high costs of extra measures. Another problem is how to deal with small-scale projects, in which the reuse of secondary building materials is not financially interesting. To make sure that the reuse of secondary building materials is successful in the whole region, it is important to set up a good organization with a good infrastructure and knowledge of the market. The Public Works Department in Rotterdam is aware of these problems. Therefore, in order to avoid the bottlenecks, its Environmental Engineering Division developed a working model called: Pragmatic application of secondary building materials in road constructions.
3.
STARTING POINTS
As starting points for this working model two aspects have been studied. In the first place the applicable rules and regulations and in the second place the current practice of road constructions in Rotterdam. The working model described in this paper integrates these two aspects.
4.
RULES AND REGULATIONS
Five official decrees form the basis of the working model. The three most important are: (1) the Building Materials Decree, (2) the Project Decision Building Materials Decree, and (3) the memorandum Working with Secondary Building Materials. Less important are (4) the Waste Dumping Ban and (5) the Installation and Licenses Decree. These rules and regulations in relation with the process of road reconstructions are summarized below: ad(1) The Building Materials Decree sets limiting conditions for using primary and secondary building materials in civil projects in land- and waterbottom. It aims at establishing a national, general protection level for soil and at stimulating the reuse of secondary building materials. It applies to granular (unmoulded)
833
or stony materials, such as ashes, slags, soil and matured harbour sludge, used in the open air. Wood, steel, clean soil, and materials with parameter concentrations below the prescribed leaching level are excluded The Building Materials Decree divides building materials into three categories on the basis of composition values for organic compounds and immission values for inorganic compounds in building materials. Category 1 are building materials that do not exceed the composition and immission values. The chance of diffusion is very low. Therefore they can be used without special conditions, such as isolation. Category 2 are building materials exceed the composition values, but exceed the immission values without isolation. These materials need to be isolated from percolation and groundwater to minimize the risk of diffusion of chemical pollution, and are also subject to specific management and maintenance rules. The third is a special category, such as bottom ash from waste incineration and tarry asphalt granulate. Reusing those requires more isolation constructions. ad(2) The Project Decision Building Materials Decree (effective as per 6 December 1995) is a practical elaboration of the former. It helps you to decide on the provisions and management rules for category 2 and the special category. It deals, for instance, with determining the distance between the average highest groundwater level and the secondary building material, and with the difference between moulded and unmoulded waste. It also contains guidelines for leaching tests and isolation constructions, and a checklist for inspection and maintenance. For example, the surface of the road could function as an isolation for rain percolation. It is necessary to keep this intact by means of inspections and maintainance works. ad(3) The Building Materials Decree will become effective in phases. Its full implementation is expected in 1998. Until then an interim policy of the associated Dutch provinces: 'Working with Secondary Building Materials' is valid. Its most important requirements are: Determine the application category by contamination- and leaching tests. Take special isolation measures (e.g. liner) and environmental control measures, in planning and realization. Registration of the nature of the materials and the exact location is essential. When the location gets another function, it is obligatory to remove the secondary building materials. In general this interim policy is quite similar to the Building Materials Decree. ad 4) The 'Waste Dumping Ban' of 27 June 1995 lists all the building materials that are not allowed to be dumped, for example fly-ashes, building debris, sieve sand, purification sludge, contaminated soil, household and industrial waste.
834
The memorandum of implementation further explains the features of the parameters. ad 5) The 'Installation and Licences Decree' became effective on 1 March 1993. It applies to installations using more than 50 M 3 building materials from outside in or at the soil. These applications are subject to a license from the competent authority, unless the building materials are directly used in a civil project in an environmentally acceptable setting.
In review: What are the requirements and duties in relation to the different categories CATEGORY
REQUIREMENT/DUTY 1
2
B o t t o m ash f r o m
TAG
w a s t e incineration d u t y of reporting to authorities for soil > 50 m3 d u t y of reporting to authorities other than for soil
*
*
making a plan of e n v i r o n m e n t a l m a i n t e n a n c e
*
*
d u t y of taking back
*
*
d u t y of removal
*
*
possible e x e m p t i o n of d u t y of r e m o v a l
-
*
W a t e r pollution act
*
minimum amount 1 , 0 0 0 t o n n e s (roads) 1 0 , 0 0 0 tonnes (large civil projects)
-
isolation measures
-
*
2x
*
supervision and m a i n t e n a n c e
-
*
*
*
-
*
*
*
m a x i m u m a m o u n t in surface w a t e r
111
~
C U R R E N T P R A C T I C E OF R O A D C O N S T R U C T I O N S IN THE M U N I C I P A L I T Y OF R O T T E R D A M
Apart from the limiting conditions, i.e. rules and regulations, the practical feasibility also influences our model. Therefore, it is important to have insight into the road constructing process. We used the process developed by the Management and Maintenance Department of the Harbour Authority of Rotterdam, the so-called Procedure of Programmed Maintenance d.d. 2. June 1992. It outlines the different stages of road reconstructions. The programming phase is dedicated to planning. The main task is to realize a cost prognosis expressed in an estimation. The participants will form a project team. The next step is the phase of initiative. The most important part is to collect the
835
necessary information and to discuss with the other participants the ingredients for the program of requirements. Also the drawings of initiative will be made, and attention can be paid to the influence on the budget because of global knowledge of the quality of the soil and the released building materials. In the design phase the definitive design of the project will be completed. Subactivities are: plan development, definitive program of requirements agreed with the project team, sending the technical program of requirements to all departments and people involved, make and check the concept drawings, make credit estimation and financial proposal. In the technical preparing phase the specification with plans will be made and submitted. The budget estimation will be prepared and the order will be placed out conform this specification. The realisation phase sees the supervision of the realisation process, and sometimes change the realisation order. In the final phase, called transfer, the civil project will be transferred to maintenance management. Environmental maintainance can be placed under the same management, but in practice it will be done by rational road mantainance (for instance inspections of road damages). The most important activities are: inspection of the object, is everything constructed conform the plans, considering the costs by transfer, storing all information, such as drawings, into the archive or data base. These aspects determine the further amount of money, which will be involved in the maintenance of the project. A one-year intake period of transfer is advisable. An adequate transfer contract containing a liability clause for the constructor is essential.
m
IMPLEMENTATION OF RULES AND REGULATIONS IN PRACTICE: THE MODEL "PASM"
The activities within the framework of active building material management, conform the policy of reusing secondary building materials, are based on the usual working process of road reconstructions and have been condensed in a model called PASM: Pragmatical Application of Secondary building Materials. The model is a process schedule including the usual stages of programming, initiation, planning and design, technical preparation, execution, and transfer.
INITIATION
Planning
I
Model PragmaticApplication of Secundairy Waste Materials.
1Collectinformation~ of secundary| waste materials ~
No
836 , Yes temporary landfill
characteristicsof materialsto be used
No
Yes ....
12 Weeks
(4 Carryingout---~ characterising ~
I I
Yes I I
t
Technical preparation
I
design I constructionas to requirements category2 J ./
9 Weeks
environmental f planl+PBT S
_
C0nstruction
.......
I
7
._
6i -- " ~-i Informcompetent~ L authority
Building ~, management
Building 1 management
nvironmental Transfer
I ~-
Transfer
environmental care
)
~
Sg
[-
V T.... fer J
,
10' Maintenance
Finish.
/
J
837
7.
E X P L A N A T I O N OF THE S C H E D U L E
The activities in the schedule are further explained below. They have been arranged according to the stages of regular road reconstruction. When it is proven that the building material belongs to category 1, the shortcut on the right-hand side is applicable
Initiation The program of requirements of the present and the desired situation is drawn up in this stage. In view of the design of the new road construction, the following is important: 1. Collect general information about the quality of the soil at the construction site, about the level of pollution, and about the categories of the building materials to be released. It is essential to have insight into the categories. Expected extra work and extra costs should be reported to your principal and your partners.
Planning/design In this phase a program of requirements will be made and the definitive design will be agreed on. Will material from the old construction be used? If No, continue with step 3. .
.
,
Determine whether released materials need to be dumped temporarily. Temporary landfill on location could be arranged with the authorities. Determine the need to examine secondary building materials for establishing the right category. In most cases a leaching test is necessary (furnace slag, asphalt granulate, sieve sand etc). Building materials certified by the Ministry of Housing, Physical Planning and the Environment need not be examined. In the latter case, proceed to step 6a. Define the application category of the final product. Generally, cement will be added to the released secondary building materials in order to obtain a goodquality final product. This final product should be subjected to leaching tests, and its composition must be assessed in order to determine the application category. These tests take about 12 weeks. If the result is category 1, the only required procedure is notifying the authorities.
838
Technical Preparation In this phase the construction drawings will be prepared in accordance with the functional program of requirements. The technical programm of requirements will be detailed in view of the project's realization. .
.
6a.
Design adaptations in conformity with the application requirements for category 2 secondary building materials. These materials must be isolated from rain- and groundwater. Submit licence (PBT) one month before the start of the project. Draw up the environmental maintenance plan phase 1, describing tasks and responsibilities for supervision, inspection, and maintenance. Report to the competent authority. For category 1 two days before starting will do. Continue to step 9.
Realization stage In this stage the project will be realized. ~
,
Building management is the main task. It is important to make correct drawings of the isolation constructions and the installation. Add last corrections to the environmental maintainance plan. The subsequent final plan will describe the exact location of isolation measures and the maintenance tasks.
Transfer In this stage the object will be transferred to the manager of the technical maintenance department. In view of long-term and current road maintenance, the following is essential: .
10.
Transfering of the object. Final inspection, and collecting the final construction drawings and other information such as the type, location, amount, and characteristics of the secondary building materials and the isolation measures. It is important to store this information in an adequate data base and keep it for at least five years. The technical maintenance manager is responsible according to the rules and regulations. This step involves environmental maintenance, such as inspections, substitution, and repairs, for which the responsibility can be transferred from the owner to the technical maintenance manager.
10a This step is only applicable to category 1 secondary building materials. Normal maintenance of the surface of the road is sufficient.
839
8.
CONCLUSION
To avoid infraction of rules and regulation, as well as stagnation and high unexpected costs it is important to plan environmental research of the secondary building materials and the soil at the location at an early stage. The civil engineer needs to implement environmental aspects into his basic planning, design and realization process. The Environmental Engineering Division of Public Works Rotterdam has developed a working model which incorporates the Building Materials Decree into the current engineering planning process. This working model is an instrument used by the Port of Rotterdam.
9.
DISCUSSION
Environmental research, and particularly leaching tests, takes usually a lot of time and can be the critical path of the project. In many cases it may cause much stagnation and high unexpected costs. Another bottleneck for managers developing road constructions, is uncertain estimates if information about the building materials is lacking. As we have seen, the quality of the secondary building materials will determine the final costs of the extra measures. A third bottleneck is the situation when less than 1000 tonnes of secondary building materials becomes available. The costs of environmental research and special isolation measures are then out of balance with the total costs of the whole road construction project. Further, stagnation of small-scale road constructions in a city centre, with its many underground mains and busy traffic, should be avoided. In the Public Works Department of Rotterdam we try to tackle these bottlenecks by using the described working model in the first place. Aiming at more accurate cost estimations we are undertaking a network research of secondary building materials by checking old archives and doing field reseach. For various reasons it is important to collect all this information in a data base. The main one is that the information can be used as a tool for the market: When is what and where becoming available? At the centre of collected information a special organization could function as a 'broker' in secondary building materials, and in fact is already operating in Rotterdam. Having gained much knowledge and experience about the characteristics of secondary building materials, it should be possible to come to agreements with the authorities over the necessity to speed up the procedures for small-scale applications of secondary building materials.
840
10. REFERENCES
1) 2) 3) 4)
5) 6) 7)
Dassen W.G, Piersma W, Schelwald R, Vries I.M.J, Re-use of waste metrials in constructional works; experiences in the city of Rotterdam,the Netherlands, Waste Materials in construction, Elsevier 1991; Berendsen T, Kooman J, Handleiding bij reconstructie en groot onderhoud van wegen, environmental engineering division of Public Works Rotterdam, 13 juni 1996; Berendsen T, Kooman J, Achtergronddocument Milieubeheer bij reconstructie en groot onderhoud van wegen, environmental engineering division of Public Works Rotterdam, 13 juni 1996; Bouwstoffenbesluit bodem- en oppervlaktewaterbescherming. Besluit van 23 november 1995. Staatsblad, 1995 567. Provincie Zuid-Holland. Nota 'Werken met secundaire grondstoffen;. Den Haag, mei 1995. Stortbesluit Bodembescherming. Besluit van 20 januari 1993, houdende regels inzake het storten van afvalstoffen. Uitvoeringsregeling Bouwstoffenbesluit. Directoraat-Generaal Milieubeheer, Directie Bodem. 's-Gravenhage, 20 december 1995.
Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
841
MSWI RESIDUES IN THE NETHERLANDS PUTTING POLICY INTO PRACTICE
Jan G.P. Bom, Ralph A.L. Veelenturf Service Centre MSWI residues, c/o Waste Processing Association, P.O. Box 19300, 3501 DH Utrecht, The Netherlands
Abstract
The Dutch policy with regard to the residues of Municipal Solid Waste Incineration (MSWI) aims towards maximization of useful application and minimization of required volume for disposal of these residues. This policy has been put into practice successfully for MSWI bottom ash. During recent years, virtually all bottom ash has found a useful application in road construction and embankments. In the case of MSWI fly ash, the policy led to the use of this material as a filler in asphalt for road construction. The demand for asphalt fillers containing MSWI fly ash, however, is limited. As a result only 20 - 30 % of the MSWI fly ash produced has been usefully applied as an asphalt filler. For the residues from flue gas scrubbing, no feasible useful application has been found to date. As a result, the entire production has been disposed of in landfills. Long term policy, however, also aims towards the development of uses for flue gas scrubbing salts.
1. INTRODUCTION The principle goal of MSW incineration is reducing the amount of space required for disposal of wastes in landfills. Incineration of MSW is preferred in the Netherlands. This preference is also an integral part of the regulations currently applicable in this respect. Incineration of MSW results in residues, typically occupying about 10% of the original volume (and 25% of the weight) of the incinerated waste. Useful application of these residues further increases the waste reduction, and therefore reduced the amount of space required for landfills. The Dutch policy not only takes into account the limited space available for land-fill, it also aims towards preserving natural resources. The total annual consumption of building materials in this country amounts 140 million tons. This consumption - primarily sand, gravel, clay and marl - results in a considerable loss of natural resources. These two aims lead to an increase of the useful application of residues, including those from MSW incineration. Like virtually all residues that can be used as secondary building materials, MSWI residues are contaminated, posing a potential risk of soil pollution. We are therefore faced with a paradox: use of residues is beneficial to the environment in the sense that it preserves natural resources, while in itself this use poses the potential environmental threat of leaching contaminants into the soil. To resolve this dilemma, regulatory measures have been stipulated for the use of residues, including MSWI residues.
2. GENERAL POLICY FOR MSWI RESIDUES In the period from 1990 through 1997, environmental legislation has changed drastically in the Netherlands. This has also affected useful application and disposal of MSWI residues. Briefly, the high standard of re-use of MSWI residues must be consolidated despite increasingly stringent environmental legislation. To this end, the government and the other parties involved are attempting
842 to work together as much as possible. This cooperation resulted in the creation of a 'Policy Paper MSWI residues' (March 1995). In this document (in Dutch: 'Implementation-plan AVI-reststoffen' [1 ]) a common course is set for the future. In order to coordinate the 36 proposals laid down in this Policy Paper, the 'Service Center MSWI residues' (acronym in Dutch: ACR) was established in mid1995. The ACR is a temporary organization located at the office of the Waste Processing Association. The latter will be responsible for acting on most of the proposals. The objective of the ACR is to finalize all proposals as agreed by 1998.
3. POLICY F O R MSWI B O T T O M ASH
The 'Policy Paper MSWI residues' summarizes the policy for MSWI bottom ash as follows: 1. Continuation of the current re-use figure (almost 100%) despite increased production. 2. Application of MSWI bottom ash preferably as replacement for sand in embankments in projects of at least 10,000 tons. Incentive for application in projects of 100,000 tons or more. 3. Bottom ash that does not meet common regulatory environmental quality standards (N2-status of the Dutch Building Materials Decree) is only to be used in government-controlled projects. 4. Improvement of the environmental quality in order to achieve N2-status, preferably by 1997.
4. PRACTICE OF MSWI B O T T O M ASH 4.1 Continuation of the current re-use figure (almost 100%)
In the Netherlands, raw MSWI bottom ash is upgraded prior to useful application. This upgrading has been described in detail elsewhere [2]. In short, upgrading consists of magnetically removing scrap material, and crashing and sieving to remove all components with a diameter larger than 40 ram. The non-combusted material present in the fraction > 40 mm is normally recycled back to the incinerator. I000
800 t 600 '
. . .
o
-
400_ "Z6
200 0
-1988
"'" ~ " ~ 1989
[
1990
1991
~-~ Production ~
1992
1993
Utilisation
/
1994
1995
1996
Disposal
Figure 1 Production and use of MSWI bottom ash in The Netherlands By means of this upgrading, the resulting MSWI bottom ash is transformed into a secondary building material. All MSWI bottom ash that has been produced over the past ten years has therefore found
843 useful application, primarily in embankments and road-base layers. The yearly production and use of MSWI bottom ash is depicted in Figure 1. Quality control and quality assurance can be regarded as one of the reasons for the present success of MSWI bottom ash in The Netherlands. In recent years, certification has become an important activity, nowadays covering the majority of MSWI bottom ash (Figure 2).
T
800 700 600 500 400 300 200 100 0
,Certi.ef:. . . . certi.e
. . . . . . . . . . . . . . . . . . . . . . . . . . .
m mm 1991
1992
1993
1994
1995
1996
Figure 2 Certification rate of MSWI bottom ash Untreated MSWI bottom ash contains about 10% retrievable ferrous metals. Annually some 65,000 to 75,000 tons of recovered steel scrap is recycled in the steel industry. Because the retrievable ferrous metals give a positive revenue, some of the processing costs are compensated. Most Dutch MSWIs (will) recover non-ferrous metals by applying Eddy-Current techniques as well. In 1996 some 1,700 tons of non-ferrous metals were recovered. Within a few years, this figure is expected to rise to about 3,500 to 5,000 tons (based on an expected annual production of more than 1,000,000 tons of MSWI bottom ash), depending on the methods for non-ferrous retrieval to be implemented. It should be noted that non-ferrous separation produces a better quality bottom ash and that the revenues of non-ferrous are such that it is considered a worthwhile investment. It can be concluded that every effort is being made to upgrade MSWI bottom ash to a generallyaccepted secondary building material. Over the past then years this effort has been rewarded by the market, in the sense that virtually all upgraded bottom ash has found useful application. 4.2 Preference for use in large projects Until the early eighties, MSWI bottom ash was used in predominantly small projects. The majority of these projects consisted of road-base construction and simple use as a paving material for (farm) yards. The latter projects had an average quantity of only 30 tons. This type of project has always been regarded as unsuitable by the government. The proposed Building Materials Decree (to be fully implemented in 1998) therefore prohibits the use of MSWI bottom ash in projects involving less than 10,000 tons. In the mean time, preferential use of MSWI bottom ash in large projects has been stimulated by all parties involved. This policy presents logistical problems, however. MSWI bottom ash is continuously produced while large projects that are suitable for MSWI bottom ash occur only incidentally. Most MSWI are able to stockpile the amount of bottom ash that is produced in one year, which appears to be sufficient for a majority of the projects. Recently an embankment was constructed using almost 1,000,000 tons of bottom ash. The size of this project required an intermediate stockpile near the project itself. In order to verify whether the agreed preference for large projects is implemented in practice, the ACR monitors the projects in which MSWI bottom ash is used. Figure 3 depicts the market share of
844 large (> 100,000 tons), intermediate (between 10,000 and 100,000 tons) and small scale projects (< 10,000 ton). In recent years, more than 2/3 of the bottom ash has been used in large-scale projects. In addition, projects involving 10,000 ton or less now comprise less than 5% of the market (Figure 3). One of the targets of the policy for MSWI bottom ash has therefore been implemented in practice: a majority of the material is used in large-scale projects. Moreover, in the near future the use of MSWI bottom ash in projects smaller than 10,000 tons will be prohibited by the Building Materials Decree.
1000 ~
< 10 kto n
I
10-100kton
~>100kton
I
'
800
"= •
600 400 z
~
200
1990
1991
1992
1993
1994
1995
1996
Figure 3 Size of projects in which MSWI bottom ash is used. 4.3 Preference for use in projects by order of the government The Dutch governmental structure consists of three layers: national authorities (ministry of Public Works), provinces and municipalities, which all order projects for road construction and embankments. With respect to large projects, the ministry of Public Works is the most likely candidate for ordering these projects to be constructed with MSWI bottom ash. Although this ministry is a strong advocate of use of secondary building materials in large projects, it holds the view that the provinces and municipalities also should take their responsibility in using MSWI bottom ash in their projects. 1000
800 600 ~"
400
r
200
90 I
91
92
93
Non-Government ~ Unknown
r-n Provinces
94
95
96
~ Municipalities
c--n Ministrie of PW
Figure 4 Ordering of MSWI bottom ash. The ACR monitors the actual market for MSWI bottom ash itemized per category of Government. In Figure 4 the results are shown, indicating that over 75% of the bottom ash is used by municipalities
845 and the ministry of Public Works. This result approaches the objective of 100% that has been determined in the policy for MSWI bottom ash. Yet non-government users are still needed to ensure that a considerable fraction of the production need not be dumped in landfills. It should also be noted that off-take by the ministry of Public Works can vary widely from one year to the next. In short, additional effort is required from all parties involved at this point in time if the desired level of 100% sale to government-controlled users is to be reached.
4.4 Improvement of the environmental quality The current successful use of MSWI bottom ash will be continued in the future. The strict environmental demands as presented in the 'Building Materials Decree' require the development of new techniques in order to improve the environmental quality of MSWI bottom ash. The current environmental quality of bottom ash does not meet the requirements as formulated in the 'Building Materials Decree' with respect to the leaching behavior of Copper, Molybdenum, Antimony and Bromide. Although practical uses will remain possible within the boundaries of future regulatory demands, the required precautions (such as the use of polyethylene and sand-bentonite liners) will probably weaken the market competitiveness of MSWI bottom ash. By improving the environmental quality of MSWI bottom ash, its market-share can be secured. The benefits of improved environmental quality are recognized by the 'Policy Paper MSWI residues', which has adopted this as one of the objectives for MSWI bottom ash.
%
t m Copper "0
... Molybdenum
-. 0
.
o
r n ..... ~ - - . - - - - ~ - - - - - 4 ~ - - - - - - n
,,..,
_._ Antimony
..~iii..Bromide
I
..
....
.
.
. . . - u - . . . . m ..... ~ 7 . . / - . ~ .- ~ . . ~ I . .-~
.
.
" ....-.,~,\
.
.
,r . . . . . e- ....
............................................................................................................................................................................ i~--i.~
. ~ ..
.
e.
*- .~.-.......... .:.": ~:'"~~;
~,
~"
........................................................ ~ . ~ : z - ~ H . < : : .
..................................................................
"'M
~
.m .... g
...... m .....
.... I
i
0,1
i
1991
i
i.
i
i
1992
i
i
i
i
1993
i
i
1
i
1994
I
I
i
i
1995
i
i
i
i
i
i
i
1996
Figure 5 Leaching of MSWI bottom ash related to the demands for granular materials (N2) of the 'Building Materials Decree' Briefly, the leaching of Antimony, Copper, Molybdenum and Bromide must be reduced in order to meet the general quality standards as laid down in the Building Materials Decree (Figure 5). During the period from 1992 until 1996, laboratory research and studies have been performed in search of applicable techniques for improving the leaching behavior. At the end of 1996 it was decided that both scrubbing MSWI bottom ash and accelerated aging of the materials as well as the use of additives were techniques worth testing on a larger scale. In 1997 accelerated aging of MSWI bottom ash will be tested on a pilot-plant scale (batches of 50 tons). In addition, process-integrated scrubbing of MSWI bottom ash using the quench will be tested at two different MSWIs. Depending on the results, a decision concerning implementation of one - or both - of these techniques will be made
846 early in 1998. Based on this time schedule, the required quality improvement should be reached at the end of the year 1999. Reduction of the leaching is regarded as a tool for maintaining the high level of use of MSWI bottom ash; it is not an objective in itself. In fact, the 'Policy Paper MSWI residues' aims towards 100% use of MSWI bottom ash, even if it does not meet the desired environmental quality. Should the techniques explained above fail to achieve the desired quality, or if they succeed only against high financial or environmental costs, it can as yet be decided that the techniques will not be implemented. In any case, the results of the large-scale experiments will probably lead to some debate about the question of whether the advantages of an improved quality justify the financial costs and environmental side-effects (e.g. discharge of waste water in case of scrubbing).
5. POLICY FOR MSWI FLY ASH
The 'Policy Paper MSWI residues' formulates the policy for MSWI fly ash as follows: 1. Continuation of the application as a filler in asphalt pavements. 2. Re-use of the remaining fly ash fraction by applying solidification techniques 3. In the case of disposal, quality improvement to a C3-class hazardous waste. In addition to the 'Policy Paper MSWI residues', other regulatory factors affect the manner in which policy for MSWI fly ash is put into practice. Due to its leaching behavior, MSWI fly ash is categorized as a C2-class hazardous waste. Disposal of C2-class (i.e. untreated) MSWI fly ash in landfills will be prohibited as from 1998. Improvement of the environmental quality of fly ash is therefore required in order to increase its uses and to make it possible to continue to dispose of the remaining fraction in landfills.
6. PRACTICE OF MSWI FLY ASH 6.1 Continuation of the application as a filler in asphalt
After years of decline, the production of MSWI bottom ash rose once again in 1996. This increase was brought about by the raise of incineration capacity in The Netherlands. The gradual decrease of the amount of fly ash produced can be explained by the modernization program of the MSW incineration capacity over the past years. Not only the flue gas emissions were reduced as a result of this modernization, but the amount of fly ash per ton of waste was almost halved (1990: 3.0%, 1996 1.6%) as well. Since the early eighties, MSWI fly ash has been used as a filler in asphalt in the Netherlands. On average, 20 kton of fly ash is used this way every year. Due to the limited capacity of the market for asphalt - and thus for asphalt fillers - in the Netherlands, the re-use figure for fly ash cannot increase substantially unless other uses are developed. MSWI fly ash replaces part of the lime present in asphalt-filler. About 30% of the filler mixture consists of fly ash. The water repellent properties of bitumen ensure low leachability of contaminants. The resulting asphalt meets the environmental demands formulated in the 'Building Materials Decree'. This is a result of the fact that bitumen encapsulates the fly ash particles and only 2% of fly ash is present in the asphalt.
847
100
1988
1989
1990
1991
1992
1993
I ~ Produktion ~ Utilisation
1994
/Disposal
1995
1996
1
Figure 6 Production and use of MSWI fly ash in The Netherlands 6.2 Treatment of MSWI fly ash and subsequent use or disposal As mentioned above, landfilling with MSWI fly ash will be prohibited as from January 1998 unless the leachability is reduced. In order to develop alternatives for landfilling, an extensive research program has been launched. The objective of this program is to explore options for reducing the leaching behavior of MSWI fly ash so as to produce a secondary building material or a C3-class waste material. In addition, the possibility of use of fly ash as an additive in concrete has been investigated.
Solidification Solidification involves the fixation of heavy metals, usually by employing cement and additives, in order to reduce their leachability. This quality improvement would result in a shift in landfilling category: C3-status instead of C2. Under more favorable conditions, solidified waste materials could be used as building materials. In recent years numerous solidification techniques and formulae have been tested. Until now, however, none of these have produced the required class-C3 landfill quality. Application as a building material is also impossible. The problems are caused by insufficient fixation of the soluble salts, such as chlorides and bromides. It should be noted that the failure of solidification techniques in treating MSWI does not apply to other types of waste materials which may contain less soluble salts. Soluble salt can be removed easily by applying washing techniques. This option, however, introduces contaminated waste-water. Only a limited number of Dutch MSWIs are permitted to discharge waste water at all. Thus the combination of solidification and washing is a vimaally unfeasible option for most of the MSWIs in The Netherlands.
Melting of MSWIfly ash Melting is regarded as a suitable technique for transforming MSWI fly ash into building materials. And yet the high operating temperatures of these melting processes - at least 1300~ - require a great deal of energy and expense. It is therefore generally accepted that melting techniques are only costeffective when the products can be used in a practical way. The quality of the melting products is considered, therefore, in conjunction with the standards for building materials rather with than those for landfilling. Since 1993, the Waste Processing Association has commissioned several laboratory research projects and feasibility studies concerning melting of MSWI fly ash. However, it was ultimately concluded that larger scale experiments were essential before any definite conclusions could be made. On behalf of one of the members of the Waste Processing Association two pilot-scale
848 melting experiments have been performed using different melting techniques. The resulting granular products have been compared with the Dutch 'Building Materials Decree'. It appeared that the leaching of Antimony did not meet the standards for unrestricted useful application within the framework of the legislation mentioned. The members of the Waste Processing Association therefore believe that the over-all feasibility (quality of the product, energy consumption, resulting residues) do not justify continued development of this technique. Use as an additive in concrete
In 1995 and 1996 the 'Service Center MSWI residues' commissioned two studies concerning the feasibility of use of fly ash as a puzzolanic additive in concrete. MSWI fly ash can be used in concrete in three ways: 9 as filler to obtain high density concrete; 9 as partial cement replacement; 9 as a high value additive after further size reduction (micronized ash, wet grinding). The application of MSWI fly ash in concrete is similar to its use as a filler in asphalt. The technical specifications for concrete, however, are more stringent. Based on the technical, environmental en economical properties, this option appeared technically feasible providing that the fly ash is scrubbed prior to pretreatment (milling). Again, the cause for this inevitable scrubbing step is the presence of soluble salts in MSWI fly ash. Without scrubbing before use of the ash, the leaching of the resulting concrete exceeds the limits for bromide as formulated in the 'Building Materials Decree'. Filler containing MSWI fly ash can not be used in re-enforced concrete due to the corrosive action of the chloride it contains. Removal of chlorides is required when use in reinforced concrete is considered. As stated above, scrubbing the fly ash introduces the problem of waste water. Taking into account that the balance of revenues and total costs are not much better than the current practice of disposing of untreated MSWI fly ash in landfills, the over-all feasibility of use in concrete is probably marginal. Use in German salt or coal mines
In 1996 useful application of MSWI fly ash in German coal and/or salt mines emerged as an attractive option. Specifically, the application in concrete for construction of walls in coal mines is an economically and environmentally favorable option that suits the Dutch policy with respect to MSWI fly ash (useful application). Early in 1997 a request for an export permit for this option was granted by the Dutch government. However, the government regards the use of MSWI fly ash in salt mines as a material for filling obsolete mine galleries as disposal. Consequently, no permits for the export MSWI fly ash for use in salt mines in this manner will be granted.
7. POLICY FOR MSWI FLUE GAS SCRUBBING RESIDUES Again we refer to the 'Policy Paper MSWI residues', in which the following policy for MSWI flue gas scrubbing residues is formulated:
1. At present no options for practical use. 2. Research of the options for re-use of the salt fraction in APC residues. 3. Explore the options for quality improvement so that disposal in landfills as a C3-type waste is made possible. At the present time, no regulatory demand for quality improvement has been formulated. The high costs for disposal of C2 wastes, however, represents a financial incentive for improving the quality to C3 waste.
849 8. PRACTICE OF MSWI FLUE GAS SCRUBBING RESIDUES 8.1 Production and subsequent landfill In 1989 the Dutch incineration emissions guideline 'Richtlijn Verbranden '89' was formulated. Subsequently, this guideline was changed into the 'Decree Air Emissions Waste Incineration'. This Decree came directly into force for new MSWI facilities on February 21, 1993. Existing plants were given until January 1, 1995 to comply with the emission limits stated in this directive. In order to meet these regulations, the flue gas scrubbing system in a typical Dutch MSWI facility currently contains: 9 one or more dust removal systems (usually ESP, sometimes fabric filters), 9 a 2-stage wet scrubbing system, 9 an activated cokes or activated carbon system 9 either a catalytic or a non-catalytic DeNOx-process. As a result, heavy metals are washed out of the flue gases using water (and the chemicals dissolved therein) and end up in a filter cake: ca. 3 kg per ton of waste. Acid components in the flue gases such as HC1 and SO2 are neutralized and subsequently discharged into surface water. Alternatively, this waste water is spray dried, producing approximately 15 kg of dry salt. In other words: the implementation of flue gas scrubbing is directly reflected in the production of solid residues. This is depicted in Figure 7.
35 30 .................t [ ~ Filtercake 25
o
I
~
Spray dry salt
]].......................................................................................... ~
...........
.....................................................................................
.................................................................................................................. . . . . . . ...........
1988
1989
1990
1991
.....
1992
1993
.....
1994
..... N N i
1995
..........
1996
Figure 7 Production of flue gas scrubbing residues 8.2 Treatment of flue gas scrubbing residues To date, there is no practical use for flue gas scrubbing residues. This is in part a result of the fact that this relatively new material is unknown. Moreover, the residues contain high levels of leachable contaminants and lack a matrix that has the physical strength to be exploited. Because these residues are the precipitated reaction products resulting from water treatment (water discharged by the wet flue gas scrubbers), their leaching behavior is only moderate compared to that of MSWI fly ash. Flue gas scrubbing residues are therefore categorized partly as C2 and partly as C3 hazardous waste.With respect to treatment of MSWI flue gas residues, the following options are considered:
850
Solidification The current initiatives for treatment of MSWI flue gas residues aim towards reducing the leaching behavior of the fiker cake using solidification techniques with subsequent disposal at C3 landfill sites. A description of this technique has been given for fly ash. Currently, only the disposal of treated residues is considered.
Purifying and discharging the salts Those MSWIs that are not allowed to discharge their waste water make use of spray-drying techniques. Spray-drying results in a product that consists primarily of salts. Upon spray-drying, the waste water is injected into hot (raw) flue gas. The raw flue gases contaminate the resulting dry salt with heavy metals. A subject of current investigation is the manner in which this product could be separated into a clean salt fraction and a contaminated filter cake containing heavy metals. Materials with a high soluble salt content, such as flue gas scrubbing residues, can be desalted by washing. Two options are available: 9 washing of salts and simultaneous or subsequent removal of specific contaminants to allow discharge of a clean salt solution into the sea; 9 selective washing of salts and beneficial practical application of the salt (CaCl2). The feasibility of both of these options is not yet clear. In any case, it is crucial that an acceptable solution for the discharge of salt water be found. It also has become apparent that there is a global abundance of CaC12. Any production of purified GaG12 out of flue gas scrubbing residues will probably result in additional discharge of this salt into the sea somewhere else.
9 CONCLUDING REMARKS In the Netherlands, the policy for MSWI residues has been agreed upon by all parties involved. This policy has been formalized in a 'Policy Paper MSWI residues'. The current and future practice with respect to use, disposal in landfills and treatment of MSWI residues is monitored in order to verify that it develops in accordance with the policy as determined. The monitoring is performed by a specific project organization (ACR), which also facilitates frequent meetings of the parties involved. During these meetings, the parties responsible for acting on proposals laid down in the Policy Paper report their progress to the other parties. In summary, not only has a policy been put forward but also an organization has been founded to ensure that this policy is put in practice, or modified where necessary based on a possible new and common understanding of all the parties involved.
10 REFERENCES
1.
2.
3.
Implementatieplan AVI-reststoffen / Policy Paper MSWI residues (1995), Publikatiereeks Afvalstoffen, hr. 1995/22, Report edited by the Ministry of the Environment, Zoetermeer, The Netherlands (in Dutch). Jan G.P. Bom, Antonius C.G. van Beurden, Emile A. Colnot, Ruud H. Keegel (1997), High standard upgrading and utilization of MSWI bottom ash, financial aspects. Paper presented at the Fifth annual North American Waste-to-Energy Conference, April 22-25, 1997, Research Triangle Park, NC, USA. Development of new technologies for MSWI residues (1997), brochure edited by the Netherlands Agency for Energy and the Environment, Utrecht, The Netherlands.
Goumans/Senderffvan der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
851
The Materials and Energy Potential Method for the Quantitative Distinction Between Waste Valorization and Elimination in the Cement Industry J.A. Zeevalkink TNO Institute of Environmental Sciences, Energy Research and Process Innovation, PO Box 342, NL-7300 AH Apeldoom, the Netherlands, e-mail [email protected]
Abstract A quantitative method is proposed to distinguish between the valorization and elimination of waste in a cement kiln. Examples are presented to illustrate the consequences of the developed approach. These examples are related to the process conditions in the kiln in the dry- and the wet-cement process. The Materials and Energy Potential (MEP) method which is presented in this report is based on the recognition that a specific waste can contribute to the cement-making process as an alternative raw material and, at the same time, as a source of energy. The paper is based on a report prepared for Febelcem, the Federation of the Belgian Cement Industry.
Introduction A quantitative method is proposed to distinguish between the valorization and elimination of waste in a cement kiln. Examples are presented to illustrate the consequences of the developed approach. These examples are related to the process conditions in the kiln in the dry- and the wet-cement process. Valorization is defined as the processing of a waste in a cement kiln to substitute raw materials and/or fuels. In this case, the waste contributes, in a positive way, to the cement production process. Waste combustion in a cement kiln without any substitution or process improvement and with the sole purpose of final waste processing is defined as elimination. The differentiation between elimination and valorization is of importance as regulations distinguish between waste elimination and valorization. For instance, directives of the European Union allow the export of waste for the purpose of valorization.
Proposals to distinguish between valorization and elimination A review of earlier proposed methods to define valorization shows that most approaches are based on the comparison of the waste with a fuel and that a clear appreciation of both the energy and the raw material value of a waste does not yet exist. Examples of conditions on calorific value or raw materials content are: 9 In Germany, according to the "Kreislaufgesetz", the energy content has to be larger than 11 MJ/kg and the fuel efficiency must be at least 75 %. Conditions on raw materials content have not been published. The Ministry of the Environment (VROM) in the Netherlands sets a calorific value limit of 15 MJ/kg and states that only liquids can be processed (valorized) properly in a cement kiln (i.e. no sludges
852
and no solids). In a former paper, a limit had been proposed of 18 MJ/kg or a useful ash content exceeding 50%. 9 In France, based on EC Directive 94/67, energy recovery for the cement industry is recognized from 5 MJ/kg. 9 In a proposition to B U W A L and in an OVAM paper, it is proposition that processing of a waste can only be regarded as valorization if the calorific value exceeds 25 MJ/kg and the contaminants in the waste do not exceed the given concentration limits o___rthe calorific value exceeds 15 MJ/kg and the concentration of the contaminants in the waste does not exceed the limits, and the total concenti'ation of Ca, Si, A1 and Fe is larger than 10 %.
General conditions for waste processing In order to have an acceptable treatment of waste in a cement kiln, some general conditions have to be met: 9
permit conditions and emission standards must be met; the quality of the cement must fulfil limits with respect to its structural capabilities and its environmental compatibility;
9
the production process must not be impaired and the safety of the workplace must be ensured; an environmental assessment should show that the cement process must be the best way of handling the waste materials. In this assessment, the cement option should be compared with alternatives such as reuse, recycling, incineration in specialized waste combustion facilities or other facilities;
9
the waste materials should not be mixed in order to reach the maximum allowable limits of contaminants in the waste.
These requirements result in criteria which limit the quantity of secondary materials used or can exclude specific wastes entirely. Several criteria have been formulated in the literature and are related to gaseous emissions, cement quality, health standards, and reactor maintenance. These criteria are necessary conditions for the application of waste in general, but do not determine the difference between elimination of waste or valorization. When these conditions are not met, the waste considered cannot be treated in a cement kiln: processing is not acceptable. Valorization or elimination The method which is presented in this paper is based on the recognition that a specific waste can contribute to the cement-making process as an alternative raw material and, at the same time, as a source of energy. This is a specific advantage of waste processing in the cement process which is expressed in the assessment method: the Materials and Energy Potential (MEP) method. Essential steps in the development of the proposed method are: division of the waste in a raw materials fraction and the rest or energy fraction which is separately evaluated as a source of energy;
853 quantitative measures for the raw materials content and the value of the energy fraction are developed; based on these measures, an assessment of waste processing as valorization or elimination is proposed. Below, the decision scheme is shown to decide upon valorization or elimination of a waste in a cement kiln following the MEP method. Another essential aspect of the proposed method is the interpretation of the term "'source of energy". In this study, a "source of energy" is distinguished from a "fuel" with calorific values of 15 MJ/kg up to 40 MJ/kg (wood, coal, oil). The starting point chosen is that any energy contribution (to the cement process) is sufficient for the classification "energy source".
Definition
o f
r a w
materials fraction
First, the raw materials part is established. This fraction contains the components that are useful to (functional in) the cement process: CaO (CaCO3), SiO2, A1203, Fe203 and SO3. The other inorganic components (including water) in the waste are allocated to the raw materials fraction up to maximum values, mwaf and maif, by which the fraction functional components is allowed to contain an equivalent amount of water and non-functional components as occur in natural raw materials. If Ca occurs as CaCO3, the CaCO3 quantity is allocated to the raw materials fraction. The following expression is used to calculate the measure M for the raw materials value of the waste: M = usmf/(
1 - minw ) ( 1 - mini )
wherein: usmf fraction of useful materials in waste as such waf water fraction in waste as such inf -- fraction of inert, non-functional components in the waste mwaf maximum water fraction allowed in raw materials fraction maif maximum inert fraction allowed in raw materials fraction mini - minimum value of inf and maif minw - minimum value of war and mwaf. -
-
-
-
Example 1: For the dry-cement process, the raw materials fraction can contain up to 15 % water. In this report, a maximum o f 10% is used as an example f o r the non-functional part o f the raw materials. So, f o r the dry-cement process m w a f = 0.15 and maif = 0.10. Example 2: For the wet-cement process, up to 30 % water atwl the same percentage, 10, o f nonfunctional (inert and trace) elements are allocated to the raw materials fraction, comparable to the natural raw materials. Again, as an example a maximum o f 10% non-functional components is used f o r the non-functional part o f the raw materials fraction. So, for the wet-cement process m w a f = 0.30 and maif = 0.10.
854
A s s e s s m e n t o f w a s t e as a s o u r c e o f e n e r g y
Secondly, the energy value of the rest or energy fraction ( - waste minus raw materials fraction) is expressed in a measure E. It is proposed to consider the combustion of a material as energy valorization if the autothermal combustion temperature, calculated for the actual conditions in the cement kiln, exceeds a required minimum process temperature, For the measure for the energetic value of a material in a process, E, the following expression is introduced: E = ( Tcomb - T o )
/
(Tref - T o )
wherein: Tr~f -
an essential reference temperature level in the process to be reached (~
Wcomb --.
the autothermal combustion temperature of the considered material under the prevailing process conditions (~
To =
an initial temperature level in the process to be considered as the starting temperature for the heating process (~
E expresses relatively the extent to which the required temperature level, Tr~', is reached or exceeded by the combustion of the energy fraction. This being the case, the material is able to contribute to the energy needs of the process. To expresses a basic temperature to be used as the initial temperature for calculating Tcomb.For example, To could be the combustion air temperature at the inlet of the kiln. As a consequence of the above, a material with the composition of the energy fraction is valorized as a source of energy if:
Ell Example: For the dry- as well as for the wet-cement process Trey is set at 1500 ~ exceeding the minimum required temperature f o r clinker formation of 1450 ~ . The process conditions to calculate the combustion temperature are: an oxygen concentration o f 3 %, an inlet temperature o f the air of 800 ~ (= To) and an energy efficiency o f 75 %. Thus, the E measure is calculated as: E = (Tcomb- 800)/(1500- 800)
855
Decision scheme for waste valorization in a cement kiln.
R e m a r k s :
no
Calculate the raw materials fraction of the waste M waste M
Acceptable with regard to: - health risks - emissions - technical product quality - environmental product quality
Raw materials fraction M: sum of CaO, CaCO 3, SiO2, ./~203, Fe203 and S O 3 - corrected for moisture content and non-funcUonal components in natural raw materials -
Calculate composition of rest fraction
Calculate concentrations and heating value based on 100 % rest fraction
Calculate max. temperature attainable when combusting rest fraction: Tcomb ~
Process conditions in cement kiln: - process temperature min. 1500 ~ ( = Tref) - 75 % energy efficiency - 3 % oxygen content air inlet 800 ~ (To) - M = raw materials fraction
Calculate M = raw materials measure E = energy measure
- E = (Tcomb - To)/(Tref- To) - T O = initial temperature (e.g. air inlet) - Tref = reference temperature, required in process
856
Generalized assessment of a waste as a source of r a w materials and energy For the general assesment of processing a waste with a raw materials and an energy part, the Materials and Energy Potential of the waste, defined as the sum of M and E, is proposed as a measure. It follows from the starting points referred to above that processing a waste with E> 1 or M - 1 in the cement kiln is a case of valorization. It is proposed generally to consider processing of a waste in a cement kiln as valorization when
E+M>I This relation is the basis for the Materials and Energy Potential method presented in this study. E is calculated from the energy fraction, M from the raw materials fraction. Examples are presented to show the consequences of this method that enables a quantitative distinction between valorization and elimination. For wastes with an M value of nearly 1, the formulated condition may be too strict. The result of the appreciation of the raw materials aspect is that TNO's MEP method favours processing of wastes with a raw materials component in the cement kiln. The allocation of (part of the) water in the waste to the raw material fraction is favours the processing of wet wastes in the wet-cement process. Generally, however, from the results of the calculations for actually applied as well as for artificially composed wastes, it is concluded, that in many cases the conclusion is the same for the wet process as for the dry process. In the following table, some calculations are presented as example. $
--~ waste
characteristics LHV * water ash
(MJ/kg) (%) (%)
Organic
Filtration
Artificial
Filter
LD
solvent
earth
waste
cake
stag
25 20
-
12.5 20 50
3.4 50 20
6 50 20
0 5 95
1873 0 1.53 1.53 YES
1912 0.59 1.75 2.45 YES
1151 0.24 0.50 0.74 NO
1400 0.24 0.86 1.09 YES
1.0 1.0 YES
1873 0 1.53 1.53 YES
2023 0.70 1.75 2.45 YES
1212 0.29 0.59 0.88 NO
1476 0.29 0.96 1.25 YES
1.0 0 1.00 YES
Dry-cement process Tcomb (excl. raw materials fraction) (~ a (-)
E (-) E+M (-) Valorization Wet-cement process Tcomb (excl. raw materials fraction) (~ i (-) E (-) E+M (-) Valorization
*
Lower Heating Value of waste as such
857
Conclusions
9 The main types of criteria for waste treatment in a cement process discussed in literature are conditions for emission standards, limits on concentrations of contaminants in the waste and limits with respect to cement quality. These aspects do not distinguish between valorization and elimination; cement processes in which wastes are used have to respect these limits whether valorization or elimination is at stake. 9 Generally, it can be concluded that in Germany, Belgium and The Netherlands, the issue of valorization and elimination has not been worked out on process technological considerations only, which explains the widely different ranges of criteria. Proposed conditions are mainly based on limits to heating values. Raw material aspects are hardly discussed. 9 The MEP method is based on the recognition that a specific waste can contribute to the cementmaking process at the same time as an alternative raw material and as a source of energy. This is a specific advantage of waste processing in the cement process. 9 The MEP method favours processing of wastes with a raw materials component in the cement kiln. 9 Non-functional compounds (Mg, P, Na-, K components and trace elements) are allowed in the raw materials fraction up to a preliminary maximum of 10 %. A better justified value should result from a study of quantities occurring in natural raw materials.
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Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
USING ENVIRONMENTAL
859 ECONOMICS
IN DECISION MAKING AND POLICY FORMULATION FOR SUSTAINABLE CONSTRUCTION
WASTE MANAGEMENT
A.L. Craighill and J.C. Powell Centre for Social and E c o n o m i c R e s e a r c h on the G l o b a l E n v i r o n m e n t ( C S E R G E ) , University o f East Anglia, N o r w i c h and University C o l l e g e L o n d o n , U K Abstract
The UK Government is aiming to increase the amount of construction and demolition waste that is reused and recycled as part of its commitment to a sustainable waste management strategy. Reusing and recycling construction waste reduces the need for raw materials and energy, with corresponding reductions in environmental emissions, aesthetic impacts and damage to natural ecosystems. However, the use of secondary materials also gives rise to environmental and social impacts, particularly in the transport and reprocessing stages. Lifecycle assessment can be used to compare alternative options for the sustainable management of construction waste.
Within the assessment, economic valuations of
environmental and social impacts provide weightings to enable this comparison.
In the past, waste management decisions have been based primarily on financial cost, and there has been no mechanism for taking environmental and social costs and benefits into consideration.
Industry and local authorities are increasingly having to take account of a
broader range of criteria, and the discipline of environmental economics provides a means by which these 'external' impacts can be quantified and included in decision making alongside financial costs.
860 Introduction
An estimated seventy million tonnes per year (17%) of the UK's waste arises from the construction and demolition industries.
Although a large proportion (63%) of the waste
created is already 'recycled' most is used for low grade purposes such as access roads within landfill sites, and only 4% is used to replace primary aggregates in more demanding construction uses. As part of a commitment to a sustainable waste management strategy the UK Government has recently introduced a target to increase the reuse and recycling of aggregates in England from 30 to 55 million tonnes per year by 20061.
A number of recent reports examine the potential for recycling construction industry wastel' 2, 3 These confirm that the disposal route taken is usually the one with the lowest financial cost.
There is much potential for increased recycling of construction waste, but there is
currently no economic or technical incentive to do so. If it could be demonstrated that the environmental and social benefits outweigh the costs of recycling, and if these benefits could be included in the decision making process, then this would provide a greater incentive to recycle. This effect has been illustrated by the UK landfill tax, which has already gone some way towards providing such an incentive. Industry and local authorities are increasingly turning towards waste reduction and recycling strategies in order to reduce their disposal costs.
Using secondary materials for construction in place of primary materials displaces the environmental and social impacts which would have arisen from primary material extraction and processing. The reduced need for raw materials and energy for extraction and processing correspondingly reduces environmental emissions. aesthetic impact and damage to natural ecosystems.
Reduced extraction also lessens the
861 Although local authorities and the construction industry are being encouraged to recycle, this is not without its own environmental and social impacts, which arise particularly from transporting the materials and in reprocessing them.
Transport can be over significant
distances, and gives rise to environmental emissions, road congestion and casualties from road traffic accidents.
A technique that can be used to compare the overall level of impacts created by reusing or recycling, with those from landfilling construction waste, is lifecycle assessment (LCA). LCA examines environmental impacts over the entire lifecycle; from obtaining the raw materials, manufacture, distribution, use, re-use/recycling, to final disposal.
The overall
environmental impact can be determined and alternative options can be compared. We are developing a lifecycle assessment model which takes into consideration the impacts discussed above.
The application of economic damage costs to impacts is explored, which allow
environmental and social costs and benefits to be included in the decision making process alongside financial costs.
Both public and private bodies are being encouraged to take
account of a broader range of criteria, and the economic valuation of impacts can thus aid policy formulation for sustainable waste management.
Lifecyele assessment methodology
Lifecycle assessment (LCA) has been used successfully for examining the environmental impacts of products and materials 4. A limited amount of work has been carried out in applying LCA to waste management 5'6. However, there has been very little work done in the application of LCA to construction and demolition waste.
862 The main stages of LCA are goal and scope definition, inventory analysis, impact assessment and the interpretation of results. The goal and scope definition sets the study's boundaries and aims.
Setting the appropriate boundaries is not straightforward because each process
within the lifecycle is connected to several other processes, which has some impact on the main system. Sensitivity analysis can be used to determine which of tiaese have a significant effect on the main lifecycle and should be included in the LCA.
The inventory analysis is a detailed compilation of all environmental inputs and outputs at each stage of the life cycle. These are presented in terms of quantities of materials and energy required, and outputs of gaseous emissions, liquid effluent and solid waste. In our study we are collecting this data from case studies. Qualitative information, such as visual impact, or raw materials scarcity is difficult to include at this stage.
The inventory data are further analysed in the impact assessment stage.
The impact
categories are chosen (e.g. greenhouse gases) and the data is aggregated.
The relative
contribution of each input or output to each environmental impact is then quantified using carbon dioxide equivalents, for example. This results in a 'balance sheet' of impacts for each lifecycle, which provides the basis for comparison.
Unless the outcome is obvious and one lifecycle is better than the other for every impact, it is necessary to apply a system of valuation to the results. In the valuation stage of the impact assessment relative weights are assigned environmental impacts. This enables a comparison of impacts that have been quantified in different units.
Valuation remains the most
contentious aspect of LCA because the weighting factors involve a subjective element and trade-offs are required between different environmental problems.
863
Valuation methodologies
There are various alternative methodologies available for weighting the impacts in the impact assessment stage. The four main approaches can be classified as distance-to-goal techniques, environmental control costs, economic damage approaches or scon.'ng approaches.
More
details on this can be found in Powell et al. 7.
In our study, we make use of economic damage costs. Economic values are available for a number of different impacts including gaseous emissions, road accident casualties and road congestion (Table 1). They are based on factors such as the number of working days lost through illness, or the cost of repair to acid-rain damaged buildings.
They also usually
include an assessment of cost based on contingent valuation. This is a technique used by environmental economists involving a questionnaire survey which asks people how much they would be willing to pay to avoid a particular impact occurring.
The valuation
methodology is then to multiply the economic value by the emission, or the nurnber of expected casualties for example, arising from the entire lifecycle.
Alternative options can
then be compared.
Conclusions
Government pledges to increase the sustainability of waste management means that it is necessary for industry and local authorities to base decision making on a broader range of criteria, including environmental and social costs in addition to financial ones.
For
construction waste, the impacts of alternative management options can be compared within the framework of a lifecycle assessment. Nevertheless, a valuation methodology is required to provide relative weights for the impacts, to be able to make a decision.
864 Economic valuation is a useful weighting technique, enabling the comparison of different types of impact, such as carbon dioxide emissions versus methane emissions, or even carbon dioxide versus road traffic accidents. Unfortunately, economic valuations do not yet exist for all impacts. Although the valuation of environmental impacts has been used for some time by environmental economists, this technique is new in the field of LCA.
However, this
economic valuation may appeal to industry because it places environmental costs and benefits on the same scale as financial ones, thus making it easy to include them in the processes of decision making and policy formulation.
Acknowledgements CSERGE is a designated research centre of the Economic and Social Research Council (ESRC). This research is being funded by the Engineering and Physical Sciences Research Council (EPSRC).
References
1. Department of the Environment. Guidelines of Aggregates Provision in England, Min Planning Guidance (MPG6). HMSO, London (1994). 2. Bnmner, P.H. and Stampfli, D.M. Material balance of a construction waste sorting p
Waste Management and Research 11 (1): 27-48 (1993). 3. Shaw, J.M. Recycling in theory and practice: the case of highways construction. Mira
Planning: 14-17 (1995). 4. Habersatter, K. and Widmer, F. Ecobalance of Packaging Materials, State of 1 BUWAL Report, Federal Office of Environment, Forests and Landscape, B (1991). 5. White, P.R., Franke, M. and Hindle, P. Integrated Solid Waste Management: A Lifec
Inventory. Blackie Academic and Professional, Glasgow (1995). 6. Powell, J.C., Craighill, A., Parfitt, J.P. and Turner, R.K. A Lifecycle Assessment Economic Valuation of Recycling. Journal of Environmental Planning
Management 39 (1): 97-112 (1996). 7. Powell, J.C., Pearce, D.W. and Brisson, I. Valuation for Life Cycle Assessment of W Management Options. CSERGE Working Paper WM 95-07, Centre for Social Economic Research on the Global Environment, University of East Anglia University College London (1995). 8. Fankhauser, S. Evaluating the Social Costs of Greenhouse Gas Emissions CSEF Working Paper GEC 94-01, Centre for Social and Economic Research on the Glq Environment University College London and University of East Anglia (1994). 9. European Commission. ExternE: Externalities of Energy. European Commission DG Luxembourg (1995).
866 10. Department of Transport. Highways Economics Note No.l, 1993 Valuation of Road Accidents. Department of Transport, London (1994). 11. Newbery, D.M. Pricing and Congestion: Economic Principles Relevant to Pricing Roads.
Oxford Review of Economic Policy 6 (2): 22-38 (1990).
867 Table 1. Economic Damage Costs
Emission
(s
Road
(s
Casualties CO2
0.40
Mortality
744,060
CO
0.60
Serious injury
84,260
CH4
7.20
Minor injury.
6,540
802
258.40
Road
(pence/PCUkm b /HGVUkm c)
Congestion 127.00
Motorway
0.26
0.52
N20
61.40
Non central
12.30
24.60
PMIO a
898.00
Rural
1.50
2.99
NOx
Notesl"a Particulates less than 10 lam diameter; b Passenger car unit kilometre; c Heavy goods vehicle unit kilometre. References: 8,9,10,11
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Goumans/Senderffvan der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
869
Application of waste materials a success now, a success in the future ir. J. Th. van der Zwan Road and Hydraulic Engineering Division Directorate-General for Public Works and Water Management Ministry of Transport, Public Works and Watermanagement Delft, The Netherlands tel: +31152699391 fax: +31152611361 E-mail: [email protected]
Summary The recycling or reuse of secondary materials is a nowadays practice in the Netherlands. At this time more than 10% of all granular materials used in he building industry is replaced by secondary materials. Especially in infrastructurale works large quantities are being applied. The drive for he use of secondary materials is sustainable development. Thanks to the government's policy and entrepreneurship successes have been scored. At this time nearly all streams of granular waste streams or industrial byproducts (f.i. building and demolition waste, milled asphalt, municipal incineration bottom ash, coal fly ash, steel slag phosphorus slag, blast furnace slag) are being reused completely. Over the years there is a perceptible change in the questions related to the use of those materials. Where in the beginning the questions were mainly technical, today they are dealing with market forces, economics etcetera. From a governmental point of view, it is necessary to spot in time bottle necks that can be prohibitive for the successful application of those materials. The role of the government is very subtle. On the one hand in the Netherlands there is a strong believe in a free market economy, on the other hand the interests of the government are in having environmental acceptable applications and in striving for high grade use. This means that the government has to set coals and to create conditions for a free market that will achieve those coals. In the paper the way the Directorate-General of Public Works and Watermanagement deals with this subject will be explained. A special attention is paid to the succesfactors that have been decisive for the successful introduction of the secondary materials. A study into these succesfactors has been performed in order tot enable the government to increase its efficiency in the implementation of the policy. Also the question of sustainable use of materials will be dealt with. Not always reuse of a material is a synonym for or in line with sustainable development. Given the fact that the first use of the secondary material of course fulfils technical and environmental criteria, it is necessary to take into account the reuse and re-reuse of secondary materials. In this kind of life cycle approach not only technical or environmental conditions have to be set, bus also labor conditions, actual control of the material and other questions influence the acceptability of the first application. If these questions are not taken into account than it is possible that a solution now creates a larger problem in the future.
870 Application of waste materials: a success now, a success in the future.
1.
Introduction
Recycling of materials in construction has been a normal matter in the Netherlands for many years. In [1], the situation in the Netherlands a number of years ago has already been examined. In recent years, a further development has taken place whereby more insight into market tendencies has arisen. The market for secondary materials has been professionalized whereby, in a number of cases, the differentiation between primary and secondary materials is beginning to fade. However, watchfulness remains the precept. Changing social, economic, scientific and technical insights can influence the current recycling possibilities. The government, with responsibility for collective standards and values, has a task to intervene and/or make adjustments where necessary. This paper will explain factors which influence the application of secondary materials, and the manner in which the government can work as effectively as possible proceeding from policy responsibility. 2.
Umbrella policy lines
In brief, the policy framework in which the recycling of secondary materials occurs will be stated. The Netherlands is a densely populated and relatively affluent country. This means that there are many people living per surface unit who all have their requirements for space and comfort. In addition, there is a high level of activity in many economic sectors. All these factors are accompanied by a continual need for the use of the limited space available in the Netherlands. Every year, very large quantities of land are used for dumping wastes, on the one hand, and for extracting surface minerals such as gravel, sand, clay, and limestone, on the other hand. The realisation that this way cannot be continued any longer is deeply anchored in Dutch society. The closing of material cycles, resulting in the need to dump less, and less exhaustion of non-renewable raw materials, is a consequence of that. The application of secondary materials derives from the national administration's policy as articulated in, among others, the National Environmental Policy Plan (NMP). [2]. The NMP states: "This NMP contains the strategy for the environmental policy for the medium long term. The strategy was developed against the background of the wish to solve or to control environmental problems within the life of a generation." For environmental control, striving for a sustainable development is the starting point. A sustainable development is a development that supplies the needs of the current generation without endangering the possibilities for future generations to also supply their needs. "Sustainable development takes shape through feedback at the sources aimed at a combination of: closing of cycles in the chain of raw material - production process - product - waste, and the accompanying emissions; saving energy, together with increasing efficiency, and the deployment of durable energy sources; promotion of the quality (above quantity) of products, production processes, raw materials, waste and environment, with a view to longer usage in the economic cycle." The quotations above from the NMP outline a policy-directed framework within which the recycling of materials takes place. The NMP-plus [3] published later does not give any change of policy, but rather a speeding up of it. A further realisation of the general policy is taking place in about three (for this subject) relevant policy lines.
871 2.1 The waste materials policy [4] The waste materials policy, for a significant part, is based on the so-called Lansink ladder. (the Dutch Lower Chamber motion from 1979). This ladder indicates a priority ranking for the waste materials problem. Lansink * * * *
Ladder: prevention recycling burning dumping
This ranking shows that prevention, or avoiding waste in accordance with policy scores highest, followed directly by recycling (in this context, by recycling is also meant the reuse of materials). In the stream of priorities, there are a number of materials which, in terms of nature and quality, are suitable for application in civil engineering. 2.2 Soil protection policy Protection of the soil is another policy line which influences the deployment of secondary materials. In a general sense, it can be said that secondary materials, due to their composition, can have other environmental effects than conventional materials. In the framework of the Soil Protection Law, a General Administrative Order (AMvB) was published which states preconditions to the application of materials on or in the soil. The Soil and Ground Water Building Materials Degree (BSB) [5] is intended to give the environmental-hygienic preconditions proceeding from soil and ground water protection to the use of secondary and primary materials on or in the arable soil or in ground water or on or in the soil under surface water. The BSB limits its operational sphere to granular (stony) materials applied outside. In applying construction materials, the concept of marginal burdening of the soil is used. Marginal burdening of the soil entails: a very minor increase in the proportions of contaminated materials in the compact layer of the soil, and protection of the ground water at the level of ground water target values. Marginal burdening of the soil is numerically filled in as: a burdening of the soil as a result of the extraction of surface materials, which mathematically leads to an increase in the compact layer of the soil of at least 1% of the proportions of contaminated materials in comparison with the soil target values in 100 years, averaged over a meter of standard soil considered to be homogeneous. The BSB assumes an emissions model which can be determined based on the emission from the surface material in a specific application. In order to continue existing recycling of materials, the maximum permitted emission level for a few materials has been increased. The emission applies to inorganic components. NEN-standards have also been developed for this in order to be able to determine the emission. A good extraction test for organic components is still lacking; because of that, a composition requirement was assumed for organic components. The BSB will come into effect in phases, the first for soil. Mid-1998, the BSB should be fully in effect. 2.3 Surfaceminerals policy [6] In order to supply the need for raw materials for construction, a policy in this domain has been developed by the government. Long-last development implies, among other things, integral control of the chain of raw materials. This entails closing the chain of raw materials in construction as much as possible, preventing degradation of the quality of raw materials, and limitating the production of waste. For the supply of raw materials, this particularly means the frugal use of raw materials and the responsible recycling of waste materials as secondary raw materials. This leads to, among other things, less excavation and dumping of waste. The main objective for the supply of raw materials in construction is: "The policy of the national government with respect to the supply of raw materials for construction has as goal to supply the need of private individuals, businesses and government for construction raw materials (in a socially
872 responsible manner) by: encouraging the use of raw materials as sparingly as possible; \ stimulating the deployment of secondary raw materials in a responsible manner as often as possible; supporting more deployment of replaceable raw materials; and ensuring the timely excavation of an adequate portion of surface minerals from Dutch Soil in the total supply of construction raw materials." The policy lines above indicate that it can be a matter of a synergy aimed at the application of secondary materials.
3.
Generalpreconditions and assumptions
In order to get the application of secondary materials in civil engineering off the ground, it is necessary to satisfy a large number of preconditions and assumptions. The most elementary principle is that there should be a market; in short, a supplying party and a demanding party who can come to terms based on a transaction that is attractive for both of the parties. In the Netherlands, the principle of the free market economy rules, which is also the prevailing opinion for the application of secondary materials. This means that the government is of the opinion that a very important role has been reserved for private industry. However, only in a few instances is private industry the producer of the waste material to be recycled. In nearly all cases, the government itself is a producer of waste materials, resulting from its function or its policy. (E-fly ash as a result of the energy policy, municipal waste incineration bottom ash (AVI-bottom ash) as a result of the policy of burning waste materials, construction and demolition waste from infrastructural works, dredging spoils from maintenance of waterways, etc.). This means that if the government wants to use private industry as a resource for solving its own problems, it must be attractive for business to operate in this market. In the following section, a number of essential preconditions will be explained.
3.1
Unambiguous policy
3.2
Private Industry
3.3
Engineering-technical parameters and rules
For the application of secondary materials, it is necessary that there be an unambiguous government policy that has taken shape in unambiguous rules that are fixed for a long time. The application of secondary materials is surrounded by risks and uncertainties. The risks concern both material-technical risks and environmental-hygienic risks. As far as the environmental-hygienic risks are concerned, it is noted that in the past, the lack of an unambiguous frame of reference of what is permitted under which preconditions, and the various ways in which the competent authorities (particularly provinces) deal with this, have frightened off potential customers. The various ministries responsible for policy (and responsible management within ministries) have not always worked together in an exemplary way in years past. At the same time, an application now may not lead in the future to a situation in which the current user is punished for his use. To realise government objectives, an active business community is necessary. It is the business community that through investments and implementation, ultimately sees to it that the policy objectives are realised. The business community is always ready to invest if the prospects are sufficiently attractive. That means that the government must pursue an investment-friendly policy, and should be reliable as legislator and lawgiver. There are examples where the government has not appeared to be too reliable a partner. That the business community has become cautious after that should be clear. To be reliable also means that the government should be ready to take risks itself by, for example, stimulating, as a customer, the application of secondary materials in its works. In a country where nearly every application is set down in standards, it is practically impossible to achieve a general application if there is not a setting down of the application of secondary materials in
873 a standard or a similar document in private law. The inclusion of secondary materials in standards and rules is an important precondition. This means that in the general sense, the effects of the application of a secondary material on the engineering-technical parameters of the construction segment must be known. In sum, the price/performance ratio of a secondary material or of a product manufactured from it should be known. In general, it is a time-consuming affair. Working for years with a limited choice of primary materials has led to, on the one hand, empiricism being based completely upon it and, on the other hand, standards and the like from that empiricism being valid only for those materials. In addition, the civil sector seems to be conservative. It takes a long time before a new material has proven itself in the market.
3.4
Economicsof the application
It has already been stated above that the price/performance ratio of an application should be known. One may assume from this that in the Dutch market economy, a material will only get a chance via market-conforming mechanisms. This means that the price/performance ratio of a secondary material must at least be comparable with a traditional material (and preferably, more suitable still).
In a general sense, it concerns an "artificial" market. Recycling or useful application is often not cheaper. This is partly due to the fact that not all cost factors which determine the price of primary materials are charged. That is a question of time-dependence is shown by the fact that bitumen, once a waste product, is now a raw material that has a market value of approximately f 350 per ton. If many applications are looked at, then it appears that these only get off the ground if favourable measure are introduced from the government's side which influence the economics of the product. The following will make that clear. The application of granulates in road construction has become a successs because dumping fees have increased sharply. The granulates have certain engineering-technical qualities which make application as a base course material possible. However, in order to have a chance in the marketplace, those materials should at least be cheaper than traditional materials. The application of a base course leads to a reduction in asphalt thickness. This means that the granulate base course must at least be cheaper than the price for the quantity of asphalt saved. In other words, the asphalt price determines the maximum price of the granulates. This means that a demolition waste recycling plant (because of the fixed and variable costs) must ask an acceptance-price in order to be able to sell the finished product competitively. Only with a dumping fee higher than the acceptance-price of the plant will the flow of demolition waste be altered toward recycling. Also, making changes in a package of products occurs only if this is economically attractive. The demolition granulate materials for road construction are in principle also suitable for application in concrete. This, however, requires an extra treatment step (sifting and washing), whereby extra costs are incurred, but an extra waste product (silt) also remains. This silt will have to be dumped at a very high cost due to the pollution level. Despite the fact that the granulate can yield a higher price (comparable with gravel), the net result is negative due to the extra treatment and the costs of processing silt. This means that without the creation of preconditions by the government in a market-conforming manner (e.g., a special dumping charge for silt), no altering of the stream of waste can be obtained.
3.5
Chainconcept approach
An important concept connected to unambiguous policy is the chain concept approach. In a general sense, this means that application of materials now may not stand in the way of future recycling. It is noted that the policy on that subject has not yet fully taken shape in an operational sense. The application of the chain concept means that a designer or a materials expert should ask himself if application of a secondary material now stands in the way of future recycling. Many factors play a role in this. A number of these will be explained. In the first place, it certainly concerns a civil engineering evaluation. If an application now leads to the
874 conclusion that the material soon cannot be recycled, the cart is being put before the horse. Indeed, in the application, the secondary material is often mixed with new materials, whereby the waste problem then only grows larger over time. In addition to a civil engineering reason, the same can also apply for environmental-hygienic and labour-hygienic aspects. Another aspect concerns enforceability. Due to preconditions, it could be necessary to allow a specific secondary material only in a specific application. One can think of the application of materials with lightly increased radioactivity only in concrete in an outside environment. (as might be the case with phosphorus slag). Future recycling of this concrete, then, must take place in such a way that this material does not come directly into concrete that is applied in an interior environment. If this cannot be guaranteed, it should seriously be considered not trying to pursue the first application. A careful weighing of risks and application possibilities is necessary for this. Here, an important responsibility lies with the government to make these choices and to create the necessary preconditions.
3.6 Quality For a customer, it is important to obtain certainty about the quality of the materials to be applied. More than with primary materials, secondary materials are surrounded by a negative image. (secondary - second-hand and thus, (in the mind of some clients) inferior with respect to primary materials). Supplying materials with a good quality guarantee (certificate) is of importance in order to get the application further off the ground. It concerns, then, a certificate which covers both the materialtechnical and the environmental-hygienic parameters. 3.7 Market To sell materials, a market is needed. On a macro scale, there appears to be no problem. Every year in the Netherlands, there is a need for approximately 140 ml tons of granulated raw materials, while approximately 30 ml tons of secondary materials can be brought to the market. At this moment, some 15 ml tons of secondary materials is being sold. It appears that in some market segments, there is saturation of the market, whereby growth in sales remains behind. In the Netherlands, it seems that the market for granular (unbound) base courses is saturated, whereby substitution effects could arise by further growth in production [7]. It should be realisecl that the market for secondary materials is very inelastic. Indeed, the production is largely determined by the size of the flow of waste and no._.~tprimarily by the demand for the product. 3.8 Market acceptance Satisfying the conditions mentioned above is not yet sufficient for market acceptance. There appear to be multiple forces at work. The customer has an important role in this. In time it must be so that the differentiation between secondary materials and primary materials fades. Think of the example of bitumen given, but one can also think of blast-furnace cement or Portland fly ash cement. Before it comes to that, it is necessary for customers to take their responsibility and open their works for the application of secondary materials. At this moment, it is the large authorities in particular that are fulfilling an exemplary function (Rijkswaterstaat (Directorate-General for Public Works and Watermanagement), the municipality of Rotterdam). Putting forward experiences about this is a means by which to achieve further spreading. Of importance is that managements of organisations themselves form a clear policy for deploying secondary materials. At this moment, the decision to apply secondary materials or not is often taken at a low level without clear guidelines from management at the foundation. 4.
Current situation
Figure 1 shows which materials in what quantities are produced and recycled. It is good to state here that recycling has various degrees, from low-quality application such as embanking material to highquality application such as, for example, replacement of scarce primary materials (application in
8?5 concrete or asphalt recycling). Different countries use different definitions of reuse and have different objectives. This does not mean that overall recycling figures from different countries can be compared with each other just like that. The degree of high quality and application in various market segments is a better measure for this.
demolition waste sand polluted ground steel slag l Recycling IE~ blast-furnace slag3 Production waste materiel phosphorous gypsum mmm phosphorous slag EC fly ash concrete and brickwork dredge spoil 1) ~ ~ ~ ~ m dredge spoil 2) n u n n n un nn dredge spoil 3) m m m m m m mum mmnnnnn n n waste incineration bottom ash asphalt rubble others , ,
m////ons o f
tons p e r y e a r 0
1
2
3
4
5
6
7
8
9
10 11
12 13 14 15
1) heavily contaminated 2) moderately contaminated 3) lightly contaminated
Figure 1: production and recycled amounts o f waste materials
In this article, this aspect will not be explained in further detail. At this moment, suffice it to say that in the Netherlands, and particularly from the standpoint of surface minerals policy, there is an effort to apply secondary materials at as high a quality as possible. The table shows that it is, of course, clear that recycling is successful in the Netherlands. However, circumstances can change, whereby it could be possible that a current application would come under pressure. Although the objective in the Netherlands is to reach a market-conforming sale of secondary materials and, thus, there is a very important role reserved for private market parties to anticipate developments, it may have become clear from the previous paragraphs that certainly for secondary materials, the role of the government is great. Therefore, the government deemed it necessary to find out what the success factors were in the past which saw to it that the current market situation arose, in order to learn lessons for policy to be developed in the future. In what follows, a number of important learning points from that study will be explained. 5.
Definition of success
In order to recognise success factors, it is necessary to define the concept success in detail. The definition of success strongly depends on the point of view of the observer. Success for one is failure for another. From the standpoint of the government, the following definitions could be applied.
876 5.1 Degreeof market acceptance The national government feels the need to obtain more insight into the market acceptance of secondary raw materials. ~ In the last decade, partly due to an active policy of various authorities, a large number of bottlenecks were overcome in order to be able to apply secondary raw materials in construction. The market acceptance of the various secondary raw materials differs, though. Some, such as granulated blastfurnace slag, are completely established; for others, conversely, market acceptance and appreciation is (more) uncertain. (figure 2)
no no
acceptation ap
dredge spoil class II and III
cleaned ground
rket cep
waste incineration bottom ash
iation
concrete and brickwork
blast furnace slag
D E G R E E OF S U C C E S S
Figure 2: representation of success
Crudely put, there are three groups of materials to differentiate: Accepted and preferred: Many secondary materials for which the (domestic) demand is greater A. than the supply are characterised by the fact that these materials are (often) preferred above primary materials in their specific application. Acceptance is determined by the unique quality/product characteristics, whereby price is less important.
Accepted: For many secondary materials, supply and demand are not always "in balance"
(usually oversupply). Then, application often depends on the relative price ratios between the various suitable materials. The price instrument is, thus, the success factor here in terms of sales (large price sensitivity). The quality differences in the supply (differentiating capability) are minor. Demand is often stimulated directly or indirectly by government. The material does have sufficient specific characteristics that, due the influence of price, lead to application. C.
Not yet accepted: For the secondary materials belonging to this group, sales and application are especially a result of pressure; for example, mandatory application or the policy-driven, explicit choice for application by specific customers (e.g., Rijkswaterstaat), without leaving it to market forces. Sensitivity to policy, then, is also great. Despite pressure, it must be concluded that in a number of cases, application is still not a viable proposition (material is dumped).
877 It is particularly groups B and C for which it is important to have insight into the factors which determine and/or can lead to such market acceptance that in the long run, it will also be a matter of establishment (appreciation). Not all secondary raw materials, however, have the potential to become fully established (appreciation). The degree of market acceptance/appreciation, however, does indicate to what extent a market-conforming sale is achievable and has a relationship with the extent to which the government should be involved in the sale. This will be discussed further. 5.2 Percentageof utilisation The percentage of utilisation is a simple instrument for ascertaining whether recycling of secondary materials has been realised. It gives quantitative information which can also make a development over time more visible. Figure 1 is an example. However, without testing on other objectives (e.g., prevention, market conformity, closing of the cycle chain approach), the instrument is too limited to be able to determine the degree of success. 5.3 Realisation of wishes and objectives Here, the perception of the actor plays a very strong role. Indeed, objectives and wishes can vary greatly from actor to actor. Thus, policy authorities will often have abstract objectives (degree of building cycle closing, long-lasting development, reduced volume of dumping, etc.), and the business community, operational objectives (continuity, market volume, revenue, profit, etc.). 6.
Sales of secondary materials and success factors
Because success here is a combination of these three definitions and is partly dependent on the actor, giving a formula for success is an impracticable affair. Due to the dynamics of time as stated earlier, circumstances change and other success factors are determining. The study shows that it is also possible to make this visible over time. To be emphasised is that what follows is indicative of the Dutch market and the Dutch situation. It certainly does not have to be true that in other countries with other societal, economic and social structures, the same end-evaluation needs to be reached. 6.1 Evaluation in time at the macro level Over time, there are four periods to be differentiated at the macro level (figure 3). Each period has its own impediments, but also its own success factors. In the following diagram, the most important impediments and success factors are summarised. The last period, 1995 to 2005, is characterised by long-lasting development. In addition to sharpening standards for asbestos, dust and radiation, an expected impediment is a greater reluctance to invest in new recycling techniques; for some materials, a greater dependence on a few processors and purchasers (environmental rules too complicated) and the falling away of national boundaries.
878
Period in the recycling industry
Most important impediments
Most important success factors
1. Trade in valuable secondary raw materials to 1970
9lack of good reference projects 9unknown reprocessing technology + application 9dumping and discharging cheap
9utilising unique product characteristics 9saving transport costs 9little thought for environmental risks
2. Trade + (threatening) scarcity of primary raw materials 1970-1985
9insufficient processing capacity/infrastructure 9suppliers split 9specifications, rules and regulations, product requirements not adjusted 9sharpened environmental policy (various compartments sometimes inconsistent) 9cost/return ratio
9technology development, incl. improved construction 9adjustment of product rules/specification standards 9threatened shortage of primary raw materials 9building up of (logistical) infrastructure
3. Trade + control of waste flow 1985-1995
9poor management of quality/ quality systems (to about 1990) 9authorities reticent as customer (to about 1990) 9unclear/environmental policy/ enforcement/provincial borders 9no continuity of supply and/ or balance of supply and demand in the short term
9high dumping charges 9quality improvement + certification 9suppliers bundle together 9consistent market policy 9authorities (RWS) serve as example
4. Sustainable development (closing of cycle + functional utilisation or fitting in) 1995-2005
9sharpening standards, Occupational Health & Safety, radiation, asbestos 9(competent) authorities more decentralised 9harmonisation supply + demand middle long term 9falling away of national borders
9ban on dumping + charges 9clarity as to what is allowed (environmentalhygienic/long-lasting context) 9(special)large-scale projects + multi-year planning projects 9long-last designs e.g. material decision lists, new technologies
Figure 3: Overview of bottlenecks and success factors, 1970-2005
6.2
Evaluation at the meso level In practice, in addition to influence from social groups and attitudes of the end users/consumers, there are three groups of actors to be differentiated:
879 government as lawgiver suppliers of secondary raw materials users of secondary raw materials. These actors are initiators of success factors and/or take measures aimed at sales promotion/improvement of secondary raw materials. Government measures again influence the activity of suppliers or users and thus, indirectly, sales (see paragraph 3). It appears that just as for "ordinary products", secondary raw materials also have a Product Life Cycle (see figure 4). Some secondary raw materials that are not saleable now or difficult to sell were previously sold without problems; e.g., dredging spoils and AVI-bottom ashes.
Different
phases
of
development
amount
#
no
turi y'
intere
GlYl~m,Jr'n
c k~=Ig o c l m m ~ l l t l o n opoll vwm_.oto elnd
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paper and synthetic mixtures
was'L'o inclnoratlon bottom aoh
oonor~'L'e ~ brlol<worl<
glass selected by colour
tim
EC i f l y ~ I
I , decrease ~ O U O slao I
blaat 'Fur'r'e,,~ olag
paper glass not selected
Figure 4: Product Life Cycle In the market dynamics, we a see a large number of products, each in its own phase of the Product Life. Social, economic and technical developments (both national and international) are making the PLC shorter, in general. In addition to the business community, the government should also have insight into this dynamic, both from the quantitative and the qualitative aspect. Indeed, it could be necessary to influence the market if, due to autonomous market tendencies, undesirable movements from a policy standpoint arise (for example, substitution or product development which is less desirable from the standpoint of chain control). It may be stated that due to the dynamic in policy, the market is still turbulent at this moment.
880 In table 1 are the most important impediments mentioned in the study [8] Table 1 : The 7 most frequent* impediments 1. 2. 3. 4. 5. 6. 7.
No quality control, know-how (insight into composition) No economic basis/dumping cheaper or (international) competitive position too weak Demand fitful, sometimes saturated market Poor image, acceptance Product standards/requirements not adjusted Insufficient (treatment) processing capacity Problems with radiation
The success factors are given for these seven most important impediments. The five most important success factors for the business community were: 1. Active selling/consistent selling policy 2. Investments in reprocessing capacity 3. Readinessto solve the social problem 4. Quality improvement and certification 5. Seek high-value market-technical The five most important success factors for the government were: 1. Helping with development of recycling technology 2. Exemplary function of government 3. Clarity created as to what is allowed 4. Dumping policy (dumping charges, bans) 5. Authorities were also problem-owners (road construction, AVI-bottom ash, polluted soil) It clearly follows from this that the market is directed more towards the operational aspects, and the government clearly must be active in the preconditions domain. For that matter, for the government in the Netherlands, a bundling of market forces by acting together with the business community and thus, being a good (and reliable) partner both at the policy level and the technical level (standards, certification, price setting, etc.) may not go unmentioned. 7.
Looking ahead
On the basis of the aforementioned, recommendations for future policy are to be drawn up. Knowledge of the market is an essential condition for acting with initiative and stimulation in the correct manner with a policy framework. The future policy of the government could be directed toward: - unique applications follow the market - specific applications direct/expand the market - amorphous applications influence the market directly With respect to the first two applications, the government should determine the preconditions in particular. For unique applications, the material itself already has an adequate price/performance ratio. The preconditions, then, will often only be set at the environmentaltechnical and labour-hygienic level. For the specific applications, it is often necessary that economic preconditions also be influenced if the material is to be sold in conformity with the market. Dumping policy can give a clear stimulus here, but certainly deviant behaviour should
881 also be counteracted through enforcement. As far as the "amorphous applications" are concerned, for example, moderately polluted soil and poorer qualities of secondary raw materials, applications undesirable from a policy standpoint should be avoided by playing a more active role. For these applications, in a number of cases, the government will itself have to act imperiously for recycling by consciously creating market demand and favouring these materials with respect to others in the application. If there is no clear policy being pursued, these materials will come off worst with respect to other materials, or the price should be so low (pay a lot for sales) that undesirable side effects can arise. These types of side effects will work negatively on the desired environmental quality and other policy objectives. In a general sense, for both the government and the business community, it is important to have insight into which phase of the Product Life Cycle the product is in. Depending on the bottlenecks with which the product is faced, it can be ascertained based on experiences from the past which success factors are suitable for removing the impediments. In figure 3, all of this is shown. The basis in all cases is a clear answer to the question "what do I want". Conclusions
The market for secondary materials is influenced to an important extent by the government which, in order to realise policy objectives, should be active in terms of preconditions. Due to the importance of the government for a good application of waste products as secondary raw materials, it is necessary that the government have knowledge of the market conditions and the market dynamic and act accordingly and take its responsibility. Through good interaction between the government and the business community based on common interests, a successful selling of secondary materials can also be ensured in the future based on knowledge of the market dynamic (Product Life Cycle) and knowledge of the nature of the application (unique, specific or more amorphous).
LITERATURE
R. van Winden, J.Th. van der Zwan, J. Zeilmaker, Applications of waste materials at infrastructural works, Waste Materials in construction, Studies in Environmental Science 48, Wascon '91. Nationaal Milieubeleidsplan, Tweede Kamer, vergaderjaar 1988-1989, 21 137, nrs 1-2 Nationaal Milieubeleidsplan-plus, Tweede Kamer, vergaderjaar 1989-1990, 21 137, nrs 20-21 Notitie inzake preventie en hergebruik van afvalstoffen, oktober 1984. Bouwstoffenbesluit bodem- en oppervlaktewaterenbescherming, Staatsblad van het Koninkrijk der Nederlanden, jaargang 1995, nr 567 Structuurschema oppervlaktedelfstoffen, Deel 1Ministerie van Verkeer en Waterstaat. 1994
Ontwerp
planologische
kernbeslissing,
Opname capaciteit van de wegenbouw voor secundaire materialen, publicatiereeks grondstoffen nr1997/09, Dienst Weg- en Waterbouwkunde, Delft, 1997 Marktacceptatie secundaire grondstoffen, huidige succesfactoren leerpunten overheid voor de toekomst, Publicatiereeks grondstoffen nr 1997/05, Dienst Weg- en Waterbouwkunde, Delft, 1997
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Goumans/Senden/van der Sloot, Editors Waste Materials in Construction: Putting Theory into Practice 9 1997 Elsevier Science B.V. All rights reserved.
883
Sustainable building and the use of raw materials in the Civil Engineering Sector. (by Ing H. Wever MSc, Road and Hydraulic Engineering Division, Ministry of Transport, Public Works and Water Management, Delft, The Netherlands). Abstract: One of the objectives of the policy of the Directorate-General of Public Works and Water Management (RWS) to be pursued is 'Sustainable building'. Sustainable building can be given meaning through the attention paid to raw materials consumption (reduction and optimal reuse of materials), reduction of waste, reduction of environmental pollution and fitting the physical infrastructure into the landscape. In 1995 RWS has prepared the "Sustainable building in the GWW-sector" programme for internalization sustainable building in projects by RWS. Sustainable building The Ministry of Housing, Spatial Planning and the Environment (VROM) has set up, within the framework of the National Environmental Policy Plan (NMP), a plan of approach to "Sustainable Building: Invest in the Future". This plan contains the environmental objectives from the environmental policy for the "building" sector in The Netherlands. The Ministry of VROM directs its attention mainly towards, for its activities in the field of sustainable building, the building of dwellings and utility building. The Directorate General of Public Works and Water Management (RWS) of the Ministry of Transport, Public Works and Water Management, as a government body, has a function to provide an example and take the lead in sustainable building in the Civil, Road and Hydraulic Engineering sector (GWW sector). RWS is working, with other parties from the GWW sector, on a further extension of the concept of sustainable building for the GWW sector.
Objectives of sustainable building Integral chain management
Minimizing the use of primary raw materials, stimulation of the reuse of secondary materials and encouraging the use of renewable raw materials.
Energy Minimizing the use of energy from non-renewable sources, maximizing the use of disintensification sustainableenergy sources, i.e. wind and sun energy.
Improving the quality Maximizing the quality of the surroundings of the building work, by of design and siting adaptationto the landscape, limiting the use of space and also improving the quality of building materials used and the welfare of the people.
Sustainable building at the Directorate-General of Public Works and Water Management (RWS) RWS has as its principal tasks the protection of The Netherlands against water, monitoring the water quality and building, maintaining and managing the main infrastructure in The Netherlands. The RWS has prepared the "Sustainable building in the GWW sector (DuBo-GWW)" programme. The objective of the programme can be seen as attempts to create a (more) sustainable infrastructure, with sustainable building as an opportunity to contribute to a sustainable society. In other words sustainable building is the bringing about of the concept of sustainable development (Bruntland Committee 1987). The infrastructure should be adapted, designed, built, maintained and demolished in such a way that the future generations are not placed in the danger of being unable to provide for their needs. For the RWS sustainable building is concerned with, in particular, roads, waterways, dams, dikes, structures (viaducts, bridges and locks) and electrical installations. The DuBo-GWW programme activities can be divided into transfer of knowledge (extending support), development of knowledge (by carrying out demonstration projects at RWS) and developing the instruments. Assistance and guidelines are produced so that evaluations/choices can be made in the field of sustainable building.
884 Sustainable building is more than a longer technical life (durable building) There is only one translation in Dutch for the words "durable" and "sustainable", i.e. "duurzaam". The word "duurzaam" was used frequently in the past for the technical life of an object. Sustainable building, however, means the sustainability from the environmental point of view. Sustainable building is seen often as building so that the effects of building have, throughout the entire life cycle of the object, the least possible impact on the surroundings, raw materials, energy etc., from the plan to build until demolition. The table below shows for each area of attention the objectives that can be aimed for in projects carried out by RWS.
Areas of attention and terms of reference for sustainable building Raw materials savings in the use of primary raw materials stimulating the use of secondary raw materials aiming for a raw material balance in projects use of renewable and not primary raw materials use of less environmentally damaging materials Waste prevention of the release of waste materials selective demolition and segregated collection of waste careful removal and processing of wastes reuse of building and demolition wastes in projects III. -
Energy energy-saving designs for objects and the infrastructure energy savings in carrying out infrastructural works energy management in existing electrical installations stimulating the use of sustainable energy sources energy disintensification of raw materials (savings in energy consumption during producing, transport, earthmoving, execution)
IV.
Design and space limiting the use of space (research into the possibilities of building underground) limiting the effects on the surroundings: limiting nuisance: noise and vibrations, subsidence, dust, odours and loss of ground water adapting the infrastructure to the landscape nature-friendly designs retaining the potential for the future (possibilities for adaptation)
Primary raw materials: raw materials extracted from the earth. Secondary raw materials: raw materials originating from various processes. Renewable raw materials: wood, coconut mats, fascine mattresses of reeds, shells.
Sustainable building aspect 'Raw materials' Sustainable building is a broad subject. Many of the activities carried out in the building cycle are involved with sustainable building. In this paper, however, the raw materials aspect is central.
Sustainable supply of raw materials The Ministry of Transport, Public Works and Water Management is responsible for the area of policy concerning the supply of raw materials for building in The Netherlands, as defined in the Structural Scheme for Surface Minerals. The aim of this policy is to meet the needs for building materials of individuals, businesses and the government in a socially acceptable way. RWS is responsible not only for the formulation of policy concerning the extraction and having sufficient stocks of raw materials, but it also stimulates the reuse of raw materials arising from the residues of industrial processes, waste incineration and energy production, in its own works and those of other principals in the GWW sector.
885 The Road and Hydraulic Engineering Division (DWW)
DWW is one of the technical and scientific divisions of the RWS that perform research and give advice about, amongst others, the use of raw materials in road building and hydraulic engineering projects. The Checklist Materials and the Environment appeared in the DWW in 1996 within the framework of sustainable building. This gives an order of preference of which (secondary) materials in which applications, from the point of view of sustainability, have the best scores. The evaluation has taken place on both the environmental and policy aspects.
The environmental aspects on which the materials were evaluated are: pollution, exhaustion (running out of raw materials and fossil fuels) and attack on the landscape (ecological, abiotic, visual, culturalhistoric, geographic). Attention is given to the environmental aspect 'energy' in the attempts to reach energy disintensification of materials. Materials used in building are evaluated by the amount of energy that it costs to extract/produce the material, to transport it to the building site, to use it and the possibility of its reuse after demolition of the work. A choice, based on the environmental evaluation and the material's energy analysis, can be made of the materials to be used and the possibility of highquality reuse of the material. In addition to the environmental aspects the business policy of RWS concerning the use of raw materials also has an important role in the use of secondary raw materials (for category 2 and the special category of building materials; using the category classification of the Building Materials Degree). The four main lines of this policy document are: - Only large scale applications of secondary raw materials (preference: > 100,000 tons). - The use of secondary raw materials should be, at least, not more expensive when compared with the use of primary raw materials. - In principle, the RWS should make the materials available to the contractor (directorate supply). - No materials originating from abroad should be used. By means of the checklist and other instruments used at the RWS, materials that are a regional or a national problem for society can be used in an environmentally-justifiable way as secondary raw materials in infrastructural works.
Sustainable building in the building process Sustainable building is directed towards all the phases of the building process, the phases in the life of infrastructural work. From the initial planning (route study/environmental impact report) to building, use and demolition and the reuse of the materials. Sustainable building can be said to have been achieved when the opportunities to prevent pollution, exhaustion of resources and attacks on the landscape in all these phases have been used optimally. With it the balance is struck between the functionality, the environment, nature and the landscape, welfare and the economy. Each phase in the building process is a preparation for the next. Choices made in earlier phases have influences upon the environmental effects in later phases.
Manuals and guidelines for the implementation of sustainable building Road and Hydraulic Engineering Division (Dienst Weg- en Waterbouwkunde): - Checklist Materials and the Environment - RWS Guideline for waste materials / Environmental specification provisions - RWS Maintenance products environmental instruction manual - Decision making manual for the main infrastructure / Scouting studies manual (route study) - Preliminary Nature compensation manual / Guide for measures for the fauna near roads and water - Nature-friendly embankments handbook (jointly with the CUR, Gouda, The Netherlands) Building Division of the RWS (Bouwdienst-Rijkswaterstaat): - Guideline for Sustainable Design / Guideline for Energy-saving Design (Engineering structures)
Programme Bureau Sustainable Building in the GWW sector
c/o Building Division Rijkswaterstaat PO Box 20.000 3502 LA Utrecht, The Netherlands Help desk: +31 - (0)15 - 2699262 / (0)30 - 2857971 Fax: +31 - (0)30 - 2897418
886 Sustainable building in the building cycle of a road project M a j o r m a d project i Description. building cycle, " " q " " phases ! .
.
.
.
.
.
.
i~'
.
.
" ....
. . . . . . . . . . . . . . . . . . . .
, " ~
......
, ! ..................................................
Bottleneck for the accessibility objective of the Dutch traffic policy. Scouting p h a s e
Selection at strategic level: the utility / necessity for the construction/adaptation of the infrastructure is examined. A choice needs to be made of which aspects of sustainable development will be included in
Level Of sustain- ..... able devel ~oPment :.,
: .................
, ................
I. Sustainable society
the scouting study.
Planning phase (route study/e.i.r)
Various solutions (road variants) are worked out in the researchreport. Environmental terms of reference and objectives for sustainable building should be formulated during this phase. The environmental impact of the building / adaptation of the infrastructure should be shown. A first road design will be made in this phase. The choice between the use of primary or secondary raw materials should be made (the choice has an influence upon the design and the environmental impact).
sustainability
The decision to build will be taken
II. Sustainable building
Design phase
Selection at operational level: the width, depth, height and shape of the work will be decided. These dimensions influence the quantities of raw materials needed for the work. The choice between the use of primary or secondary raw materials for the various parts of the design is made during this phase (adapting the design to the chosen materials). - Low maintenance structures/road designs - Technical durability of the design - Flexibility in the design of the work
sustainable designing
Preparation phase
Selection at detail level: selection of raw materials. Determine the environmental impact of the materials to be used.
Specification of requirements completed/Work contracted out to a contractor. Construction phase
Negative effects upon the environment as a result of construction should be minimized. The contractor carries out the work according to the specification. Environmental care may be connected with that for health and safety of the workers and quality assurance.
sustainable construction / environmental care
Utilisation phase
Choices in the manner of management and maintenance. When secondary raw materials have been used monitoring of the work will have to be taken into account. When the work is no longer adequate study the possibilities for improvement/adaptation, before deciding on demolition.
sustainable management / environmental care
Demolition phase
Environmentally-friendly (selective) demolition. Examine the opportunities for reuse of the materials made available. After demolition the site can be used for the new function.
sustainable demolition / environmental care
![Environmental aspects of construction with waste materials: proceeding[s] of the International Conference on Environmental Implications of Construction Materials and Technology Developments, Maastricht, the Netherlands, 1-3 June 1994](https://epdf.tips/img/300x300/environmental-aspects-of-construction-with-waste-m_5a3f46fab7d7bc2e4b1631da.jpg)
