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ENVIRONMENTAL ASPECTS OF CONSTRUCTION WITH WASTE MATERIALS PROCEEDING OF THE INTERNATIONALCONFERENCE ON ENVIRONMENTAL IMPLICATIONS OF CONSTRUCTIONMATERIALS AND TECHNOLOGY DEVELOPMENTS, MAASTRICHT, THE NETHERLANDS, 1-3 JUNE 1994
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Studies in Environmental Science 60
ENVIRONMENTAL ASPECTS OF CONSTRUCTION WITH WASTE MATERIALS PROCEEDING OF THE INTERNATIONAL CONFERENCE ON ENVIRONMENTAL IMPLICATIONS OF CONSTRUCTION MATERIALS AND TECHNOLOGY DEVELOPMENTS, MAASTRICHT, THE NETHERLANDS, 1-3 JUNE 1994
Edited by
J.J.J.M. Goumans Netherlands Agency for Energy and the Environment (NOVEM), P.0. Box 8242, 3503 RE Utrecht, The Netherlands
H.A. van der Sloot Netherlands Energy Research Foundation (ECN), P.0.Box 1, 1755ZG Petten, The Netherlands
Th. G. Aalbers National Institute of Public Health and Environmental Protection (RI VM), P.O. Box 1, 3720 BA Bilthoven, The Netherlands
ELSEVIER Amsterdam
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London
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NewYork
- Tokyo
1994
ELSEVIER SCIENCE B.V. Molenwerf 1 P.O. Box 21 1,1000 AE Amsterdam, The Netherlands
ISBN 0-444-81853-7
01994 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), Salem, Massachusetts. 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 publisher. 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
Studies in EnvironmentalScience Other volumes in this series 1 2
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33
Atmospheric Pollution 1978 edited by M.M. Benarie Air Pollution ReferenceMeasurement 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 PotentialIndustrialCarcinogensand Mutagens by L. Fishbein IndustrialWaste Management by S.E. Jergensen Tradeand Environment: ATheoretical 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 Eliminationof Wastes edited by G. Milazzo Bioengineering,ThermalPhysiologyand Comfort edited by K. Cena and J.A. Clark Atmospheric Chemistry. Fundamental Aspects by E. Meszaros Watersupply and Health edited by H. van Lelyveld and B.C.J. Zoeteman Man under Vibration. Sufferingand Protection edited by G. Bianchi, K.V. Frolov and A. Oledzki Principlesof Environmental Science and Technology by S.E.Jergensen and I. Johnsen Disposalof RadioactiveWastes by Z. Dlouhy Mankind and Energy edited by A. Blanc-Lapierre Quality of Groundwater edited by W. van Duijvenbooden, P. Glasbergen and H. van Lelyveld Educationand Safe Handlingin PesticideApplication edited by E.A.H. van Heemstra-Lequin and W.F. Tordoir PhysicochemicalMethods for Water and Wastewater Treatment edited by L. Pawlowski Atmospheric Pollution 1982 edited by M.M. Benarie Air Pollution by NRrogenOxides edited by T. Schneider and L. Grant Environmental Radioanalysis by H.A. Das, A. Faanhof and H.A. van der Sloot Chemistry for Protectionof the Environment edited by L. Pawlowski, A.J. Verdier and W.J. Lacy Determination and Assessment of PesticideExposureedited by M. Siewierski The Biosphere: Problems and Solutions edited by T.N. VezirMlu Chemical Events in the Atmosphere and their Impacton 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 Chemistryfor Protectionof the Environment 1985 edited by L. Pawlowski, G. Alaerts and W.J. Lacy Acidification and its Policy Implicationsedited by T. Schneider Teratogens: Chemicals which Cause Birth Defects edited by V. Kolb Meyers Pesticidechemistryby G. Matolcsy, M. Nadasy and Y. Andriska Principlesof EnvironmentalScience and Technology (secondrevisededition) by S.E. Jergensen and I. Johnsen
34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 51 52 53 54 55 56
57 58 59
Chemistry for Protectionof the Environment 1987 edited by L. Pawlowski, E. Mentasti, W.J. Lacy and C. Sarzanini Atmospheric Ozone Researchand its Policy Implicationsedited by T. Schneider, S.D. Lee, G.J.R. Wolters and L.D. Grant 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 BioenvironmentalStudies: The HanfordExperience, 1944-1984 by C.D. Becker Radon in the Environment by M. Wilkening Evaluationof Environmental Data for Regulatory and ImpactAssessment by S. Ramamoorthy and E. Baddaloo EnvironmentalBiotechnology 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 Pollutionedited 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 inThe Netherlands edited by G.J. Heij and T. Schneider Handbook of RadioactiveContamination 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: Evaluationand Policy Applications edited by TSchneider Biotechniques for Air PollutionAbatement and Odour Control Policies edited by A.J. Dragt and J. van Ham EnvironmentalScience Theory. Concepts and Methods in a OneWorld, Problem-OrientedParadigm by W.T. de Groot Chemistry and Biology of Water, Air and Soil. Environmental Aspects edited by J. Tolgyessy The Removalof NitrogenCompoundsfrom Wastewater by B. Halling-Serensen and S.E. Jergensen EnvironmentalContamination 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 GeologicalDisposalof 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. Sequiera
vii
mREw0RD The Organizing and Scientific Committees of the Second International Conference on Environmental Implications of Construction Materials and Technology Developments, WASCON '94, present herewith the Proceedings of this Conference, which will be held from 1-3 June, 1994 in Maastricht, The Netherlands.
SCOPE OF THE CONFERENCE The concept of "Sustainable Development", implicating the protection of soil and groundwater, the limitation of waste production and the re-use of solid waste materials is still the leading theme of WASCON '94. Although it is clearly recognized in most countries that products derived from solid waste materials can be applied as construction materials, research is still needed to assess various environmental problems. With respect to the prediction and control of release of contaminants, there is a trend towards consensus on the usage of the various leaching tests which have been developed. In addition, scientific research regarding, e.g., speciation of chemical elements is generating results which can be transferred to technical studies and also to more general studies dealing with environmental and life cycle analysis. As is reflected in the amount of contributions, two topics are currently the subject of study in many countries. Firstly, there is the problem of municipal waste, stabilization of waste materials and their use in civil works is the subject of research and legislation. Within these fields it can be seen that transfer of knowledge and know-how is helping to find environmentally acceptable solutions, also with respect to establishing criteria and standards. Secondly, as far as technology development is concerned, it can be seen that research has encompassed a wider variety of materials. Furthermore, quality control and standardization has been applied to these types of materials. The incorporation of environmental quality standards is the next step to be taken, and is already the subject of some studies. In general, technical studies are often already accompanied by environmental aspects as an integral part of research. The Organizing Committee hopes that WASCON '94 will contribute to the solution of environmental and technical problems concerning the re-use of waste materials and, thus, to sustainable development in building practice.
SCIENTIFIC COOPERATION
The International Society for Construction with Alternative Materials (ISCOWA) was founded as a result of the First Conference, held in 1991. To date over 110 members from 15 different countries have been registered, and the Society is still growing. ISCOWA is actively participating in the scientific committee and in the organization of WASCON '94. Many ISCOWA members have submitted contributions to the Conference, arid the Chairmen of the ISCOWA working groups have prepared state-of-the-art documents.
viii
ACKNOWLEDGEMENT Organizing an international conference means a lot of work for many people, we therefore wish to express our thanks to the following: The members of the Honorary Committee and the Scientific Committee. The Dutch National Institute of Public Health and Environmental Protection (RIVM), the Netherlands Energy Research Foundation (ECN), Environment Canada, The United States Environmental Protection Agency (USA-EPA), Community Bureau of References, Commission of the European Community (BCR), The Danish National Agency for Environmental Protection, European Association for Use of the By-products of Coal Fired Power Stations E.V. (ECOBA), the Center for Applied Research, Norway, and the Netherlands Agency for Energy and the Environment (NOVEM) for supporting the Conference. The Netherlands Ministries of Housing, Physical Planning and the Environment and of Economic Affairs, which have sponsored the Conference. Van Namen and Westerlaken Congress Organization Services, De Boer and Van TeyIingen Public Relations and the staff of Elsevier Science B.V. All Authors, Participants of the Conference, and all others who have contributed to WASCON '94. On behalf of the Organizing Committee, Dr J.J.J.M. Goumans Chairman of ISCOWA Utrecht, The Netherlands, 21 March 1994
ix CONTENTS Preface
...................................................................................................
vii
SECTION 1: Opening The International Ash Working Group: A Treatise on Residues from MSW Incinerators .............................................................................................. S.E. SAWELL, A.J. CHANDLER, T.T. EIGHMY, J. HARTLEN, 0. HJELMAR, D. KOSSON, H.A. VAN DER SLOOT and J. VEHLOW
3
International Progress in Solid Waste Management J.H. SKINNER
..............................................
7
Life Cycle Analyses; Results of Some Case Studies J. CRAMER
.............................................
17
SECTION 2: Environmental Aspects Chemical Processes Controlling the Mobility of Waste Material Contaminants in Soils L.G. WESSELINK, P.M. DEKKER and T.G. AALBERS
..
31
Leaching of Slags and Ashes - Controlling Factors in Field Experiments versus Laboratory Tests ....................................................................................... A.-M. FALLMAN and J. HARTLEN
39
Validation of Leaching Model on Actual Structures ............................................. G . VAN DER WEGEN and C. VAN DER PLAS
55
Intercompanson of Leaching Tests for Stabilized Waste ...................................... H.A. VAN DER SLOOT, G.J.L. VAN DER WEGEN, D. HOEDE and G.J. DE GROOT
63
Immobilisation Potential of Cementitious Materials F.P. GLASSER
...............................................
Coal Fly-Ash Leaching Behaviour and Solubility Controlling Solids R. GARAVAGLIA and P. CARAMUSCIO Modelling Ca-Solubility in MSWI Bottom Ash Leachates R.N.J. COMANS and J.A. MEIMA
........................
87
....................................
Particle Petrogenesis and Speciation of Elements in MSW Incineration Bottom Ashes T.T. EIGHMY, J.D. EUSDEN, JR., K. MARSELLA, J. HOGAN, D. DOMINGO, J.E. KRZANOWSKI and D. STAMPFLI
77
103
..
111
X
An Approach to the Assessment of the Environmental Impacts of Marine Applications of Municipal Solid Waste Combustion Residues ................................................. 0. HJELMAR, E.A. HANSEN, K.J. ANDERSEN, J.B. ANDERSEN and E. BJ0RNESTAD Quality Assessment of Granular Combustion Residues by a Standard Column Test: Prediction versus Reality ............................................................................. M. JANSSEN-JURKOVICOVA, G.G. HOLLMAN, M.M. NASS and R.D. SCHUILING
137
161
Geochemical Factors Controlling the Mobilization of Major Elements during Weathering of MSWI Bottom Ash .................................................................. C. ZEVENBERGEN and R.N.J. COMANS
179
Leaching Behaviour of Building Materials with Byproducts under Practical Conditions ............................................................................................... P.J.C. BLOEM, F.L.M. LAMERS and L. TAMBOER
195
FGD Gypsum Definitions and Legislation in the European Communities, in the OECD and in Germany ........................................................................................ F. WIRSCHING, R. HULLER and R. OLEJNIK
205
In-situ Utilization of Waste Bentonite Slurry N. UCHIYAMA and S. HORIUCHI
.....................................................
217
..................................
227
The use of MWI Fly Ash in Asphalt for Road Construction J.B.M. HUDALES
Enhanced Natural Stabilization of MSW Bottom Ash: A Method for Minimization of Leaching ............................................................................................. J.J. STEKETEE and L.G.C.M. URLINGS Immobilization of Slag Material by Foam Bitumen J.H. DIJKINK
233
.............................................
239
..........................
247
Immobilisation of Phenol and PAH by Special Hydraulic Binders P. VOGEL and M. SCHMIDT
Leaching of Organic Contaminants from Contaminated Soils and Waste Materials ...... M. WAHLSTROM, H. THOMASSEN, J. FLYVBJERG, A.C. VELTKAMP, C. OSCARSSON, J . - 0 . SUNDQVIST and G.A. ROOD
257
Investigating a Leaching Test for PCBs and Organochlorine Pesticides in Waste and Building Materials ................................................................................ G.A. ROOD, M.H. BROEKMAN and T.G. AALBERS
27 1
French Qualification Procedure for Solidification Processes ................................... J. MEHU, P. MOSZKOWICZ, R. BARNA. P. PHILIPPE and V. MAYEUX
281
xi Utilization Status, Issues and Criteria Development for Municipal Waste Combustor Residues in the United States ....................................................................... D.S. KOSSON, B.A. CLAY, H.A. VAN DER SLOOT and T.T. KOSSON
293
........................
305
Validation of Dutch Standard Leaching Tests Using NEN-IS0 5725 G.J. DE GROOT and D. HOEDE
The Laconia, New Hampshire Bottom Ash Paving Project ..................................... C.N. MUSSELMAN, M.P. KILLEEN, D. CRIMI, S. HASAN, X. ZHANG, D.L. GRESS and T.T. EIGHMY
315
Application of Fly Ash and other Waste Materials for the Construction of an Off-Shore Island Opposite the Coast of Tel-Aviv ............................................................. Y. ZIMMELS, G. SHELEF and A. BOAS
329
Fly Ash Utilisation in Civil Engineering .......................................................... J.G. CABRERA and G.R. WOOLLEY
345
High Pressure Mixing: A New Technology to Re-use Waste Materials Containing CaO and/or MgO ...................................................................................... R. HAVERKORT, W. DEKKER and J. SENDEN
357
Environmental Compatibility of Cement and Concrete ......................................... S. SPRUNG, W. RECHENBERG and G. BACHMANN
369
.................................................
387
European Standardization of Additions for Concrete ........................................... J.M.J.M. BIJEN
397
State of the Art of Waste Characterization on European Level ............................... A. TUKKER. M. VAN DEN BERG and H.A. V A N DER SLOOT
409
Leaching Properties of Cement-bound Materials I. HOHBERG and R. RANKERS
Leaching Behavior Assessment of Wastes Solidified with Hydraulic Binders: Critical Study of Diffusional Approach ............................................................ P. MOSZKOWICZ, R. BARNA, J . MEHU, H. VAN DER SLOOT and D. HOEDE Burning of Hazardous Wastes as Co-Fuel in a Cement Kiln Does it Affect the Environmental Quality of Cement? .......................................... K.H. KARSTENSEN Approach towards International Standardization: A Concise Scheme for Testing of Granular Waste Leachability ..................................................................... H.A. VAN DER SLOOT, D.S. KOSSON, T.T. EIGHMY, R.N.J. COMANS and 0. HJELMAR Speciation of As and Se during Leaching of Fly Ash E.E. VAN DER HOEK and R.N.J. COMANS
..........................................
42 1
433
453
467
xii Measurement of Redox Potential During Standardized Column Tests J. KEIJZER and A.J. ORBONS
.......................
477
The Influence of Reducing Properties on Leaching of Elements from Waste Materials and Construction Materials ........................................................................... H.A. VAN DER SLOOT, D. HOEDE and R.N.J. COMANS 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 ..................
483
49 1
R. MEIJ and H.P.C. SCHAFTENAAR Role of Facilitated Transport in the Emissions of Secondary Raw Materials J.J. STEKETEE, J.C.M. DE WIT, G.J. VAN ROSSUM and L. G. C.M. URLINGS
...............
507
Immobilization of Heavy Metal Ions by the Alkali Activated Slag Cementitious Materials ................................................................................................. J. MALOLEPSZY and J. DEJA
5 19
Integrated Treatment of MSWI-residues: Treatment of Fly Ash in View of Metal Recovery ......................................................................................... B. LAETHEM, P. VAN HERCK, P. GEUZENS and C. VANDECASTEELE
525
Life Cycle Assessment of a Road Embankment in Phosphogypsum: Preliminary Results .................................................................................... J.W. BROERS, F.E.T. HOEFNAGELS and H.L. ROSKAMP
539
Co-combustion of Coal and Waste Wood, Consequences for the By-product Quality M.L. BEEKES, C.H. GAST and A.J.A. KONINGS
.....
543
SECTION 3: Technical Aspects Use of Demolition Concrete to produce Durable Structural Concrete P.J. WAINWRIGHT and J.C. CABRERA
.......................
553
Improvement of Portland Cement/Fly Ash Mortar Strength using Classified Fly Ashes ............................................................................................... J. PAYA, V. BORRACHERO, E. PERIS-MORA, A. ALIAGA and J. MONZd
563
Ground Fly Ashes: Characteristics and their Influence on Fresh and Hardened Mortars .................................................................................................. J. PAYA, V. BORRACHERO, J. MONZO, E. PERIS-MORA and A. ALIAGA
571
Development of Cementitious Products using Industrial Process Wastes as Sources of Reactive Sulfate and Alumina ......................................................... G. BELZ, J. BERETKA, R. CIOFFI, L. SANTORO, N. SHERMAN and G.L. VALENTI
579
xiii Potentials for Utilisation of PFBC Ash ............................................................ J. ROGBECK and P. ELANDER
589
..................................
599
Recycling of Magnesium Slags in Construction Block Form M. COURTIAL, R. CABRILLAC and R. DUVAL
Improving the MSWI Bottom Ash Quality by Simple In-Plant Measures I. SCHNEIDER, J. VEHLOW and H. VOGG
...................
Potentials in Quality Improvement of Processed Building Rubble by Demolition and Treatment Technics .............................................................................. J.O.V. TRANKLER and I. WALKER Quantities and Qualities of Municipal Waste Incinerator Residues in the Netherlands J.G.P. BORN
605
62 1
....
Upgrading Techniques for the Quality Improvement of Municipal Waste Incineration Residues ................................................................................. F.J.M. LAMERS and J.G.P. BORN Re-use of Colliery Spoils in Construction Materials using Fluidized Bed Combustion J.J.M. HEYNEN, H.N.J. A. BOLK, G.J. SENDEN and P.J. TUMMERS
...
633
645
655
Recovery of Raw Materials from Reclaimed Asphalt Pavement .............................. E. MULDER, C. DE GROOT, C. JONKER and J. VAN DER ZWAN
665
.............................
673
Applications for Coal-use Residues: An International Overview L.B. CLARKE
Specifications and the Use of Wastes in Construction in the United Kingdom ............. R.J. COLLINS and C.J. ATKINSON Overview of Coal Ash Use in the USA S.S. TYSON
..........................................................
687
699
Environmental Life Cycle Analysis of Construction Products with and without Recycling ................................................................................................ M.S. A.M. SCHUURMANS-STEHMANN
709
Assessment of the Environmental Compatibility of Industrial By-products and Recycled Materials .................................................................................... R. BIALUCHA, J. GEISELER and K. KRASS
719
...................................
727
Environmental Management in Large Construction Projects E.K. LAURITZEN
A Concept for the Environmental Evaluation of Waste Management Benefits A. TUKKER and D.J. GIELEN
.............
737
......
749
Technological and Environmental Properties of Concretes with High PFA Content H.A.W. CORNELISSEN and R.E. HELLEWAARD
xiv Towards Sustainability with Construction and Demolition Waste in Belgium? G. DESMYTER, B. LAETHEM, B. SIMONS, J. VAN DESSEL and J. VYNCKE Disintegration of Fly Ashes i? the Rotary-vibration Mill J.SIDOR and M.A. WOJCIK
.............
.......................................
759
775
Release of Heavy Metals from a Municipal Solid Waste Incineration Residue Stabilized in Non-traditional Matrices ............................................................. V. ALBINO, R. CIOFFI, B. DE VITO, M. MARROCCOLI and L. SANTORO
789
Applications of By-products from Coal Gasification Power Plants: Quality- and Environment-Related Aspects ........................................................................ M.L. BEEKES, J.W. VAN DEN BERG and A.J.A. KONINGS
80 1
Quality Improvement of MSW Fly Ash and APC Residue from MSW Incinerator Amsterdam-West using Different Iinmobilisation Processes ................................... H.T.M. VAN DE LAAR, J. SLAGTER, R.F. DUZIJN and J.H. DE ZEEUW
81 1
Certification System for Aggregates Produced from Building Waste and Demolished Buildings ................................................................................................ C.F. HENDRIKS
82 1
Sampling and Sub-sampling of Primary and Secondary Building Materials: A Statistical Treatise .................................................................................. A.M.H. VAN DER VEEN and D.A.G. NATER
835
Industrial Scale Application of the Alkali Activated Slag Ceinentitious Materials in the Injection Sealing Works ...................................................................... W. BRYLICKI, J. MALOLEPSZY and S. STRYCZEK
84 1
The use of MSWI Bottom Ash in Asphalt Concrete ............................................. M.M.T. EYMAEL, W. DE WIJS and D. MAHADEW
85 1
How to Prevent Expansion of MSWI Bottom Ash in Road Constructions? ................. M.M.C. ALKEMADE, M.M.T. EYMAEL, E. MULDER and W. DE WIJS
863
Microstructure of Concretes Containing Artificial and Recycled Aggregates J.A. LARBI and P. STEIJAERT Frost Susceptibility of Recycled Aggregate M.M. O’MAHONY
................
877
.......................................................
889
...............................
897
Use of Crushed Tile and Concrete as Filling in Pipe Trenches J. FOLKENBERG
XV
Use of Ashes from MSW Incineration in Cementitious Building Materials A. GERDES and F.H. WITTMANN
................. 905
Effect of Grain Size Composition of the Calcium-sulphate Fly Ashes on the Properties of Autoclaved Building Materials ..................................................... Z. PYTEL and J. MALOLEPSZY Sulphate and Acid Attack on Concrete i n Ground and Landfill C. PLOWMAN
...............................
909 917
Contaminated Soil Cement Stabilizations for Application as a Construction Material .................................................................................................. P.J. KROES and J . VAN LEEUWEN
925
The Assessment of a Pollutant Charge of Dredged Sediments as a Tool to Minimize Adverse Environmental Effects ...................................................................... E. PERIS-MORA, J. MONZO, J. PAYA and J.M. MESA
929
Minestone Substraturn,Behaviour under Loading ................................................ K.M. SKARZYNSKA and E. ZAWISZA
939
Ecological and Energy-saving Advantages and Benefits of Building with Earth .......... H. HOUBEN
94 1
Fly Ash and Slag Reactivity in Cements - TEM Evidence and Application of Thermodynamic Modelling ........................................................................... H.S. PIETERSEN and J.M.J.M. BlJEN
949
SECTION 4: Closing State of the Art Report: Use of Waste Materials in Construction Technological Development .......................................................................... G.R. WOOLLEY
963
...............................
979
A Unified Approach to Leaching Behavior of Waste Materials T.T. EIGHMY and H.A. VAN DER SLOOT
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SECTION 1: Opening
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Environmental Aspects of Construction with Waste Materials JJJ.M. Goumans, HA.van der SIoot and Th.G. Aalbers (Editors) @I994 Elsevier Science B. V. All rights resewed.
3
THE INTERNATIONAL ASH WORKING GROUP A TREATISE ON RESIDUES FROM MSW INCINERATORS
S.E. Sawell', A.J. Chandler*, T.T. Eighmy3, J. HartlCn4, 0. Hjelmars, D. Kosson6, H.A. van der Sloot' and J. Vehlow8 'Compass Environmental Inc., Burlington,Ontario; 2 k J . Chandler C Associates Ltd., Willowdale, Ontario; Wniversity of New Hampshire, Durham, N H 4SwedishGeotechnical Institute, Linkoping, Sweden; SDanishWater Quality Institute, H0rsholm, Denmark; 6RutgersUniversity, Piscataway, NJ; 'Netherlands Energy Research Foundation,Petten NL; and *KernforschungszentrumKarlsruhe, Germany. INTRODUCTION Historically, public opposition to the siting and operation of MSW incinerators has been based on concerns over emissions of contaminants to the atmosphere. During the late seventies and early eighties, much of this concern was justifiable in light of the poor performance of inadequately designed and operated facilities. In addition, the enhanced ability of scientists to detect trace contaminants a t very low concentrations placed further pressure on the industry to reduce emissions. During the last decade, the industry has responded positively. Most of the concerns related to emissions have been addressed by optimizing incinerator design and operation through improved combustion control. This and the implementation of new air pollution control technologies has dramatically reduced atmospheric emissions. The advent of more efficient incinerator operation and better designed air pollution control systems has resulted in concern shifting from air emissions to the disposal of the captured residues. Today, ash disposal is one of the major issues limiting the acceptance of new incinerator facilities.
The Incinerator Ash Issue In 1988, several individuals actively involved with ash research were prompted to suggest that the global data base on incinerator ash needed to be compiled, reviewed and critically evaluated. The major concern over incinerator ash centred on the potential for human exposure to contaminants, specifically trace metals in the ash. The potential exposure routes were deemed to be either through the inhalation of fugitive dust or the ingestion of contaminated groundwater. While the industry was moving to address these issues, environmental groups continued to push for more stringent regulations for the disposal of the residues. Several factors contributed to this situation:
o The heterogeneity of the ash characteristics often resulted in contradictory laboratory results; o Residue sampling and analytical protocols were not well established, thereby compounding the variability of the data; o Different combustion technologies and air pollution control systems produce additional variations in the quality of the residues, and thus the data bases exhibit large variations;
4
o Cursory groundwater studies provided an inadequate data base on the long-term effects of monofills on groundwater; and, o The management of these residues had to consider the contentious issue of potential long-term effects, particularly in view of the changing nature of the MSW stream. All of these factors contributed to a highly fragmented data base, which in turn, had resulted in much confusion and a preponderance of misconceptions. The researchers thought that a critical review of the data would provide an opportunity to clarify many of the issues. Thus, in April of 1989, a proposal entitled: "MUNICIPALWASTE INCINERATION: AN INTERNATIONAL PERSPECTIVE ON THE CHARACTERISTICS, DISPOSAL, TREATMENT and UTILIZATION OF RESIDUALS"
was submitted to the International Energy Agency (IEA) for consideration and this led to the establishment of an International Ash Working Group (IAWG). OBJECTIVES While the general purpose of the project was to put the MSW incinerator ash issue into proper perspective, the project had some specific objectives. 1. Provide a compilation of existing knowledge concerning MSW incinerator residues;
2. Define uniform protocols for the sampling and full characterization (including defining chemical, physical, and leaching properties) of incinerator residues; 3. Describe the fate and behaviour of contaminants during the incineration process including documenting the effects of different incinerator designs, air pollution control systems, incinerator operations and refuse feedstocks; 4. Evaluate and develop recommendations with regard to current or proposed disposal, treatment
utilization and recovery practices; 5. Provide regulators with recommended standard performance criteria for the residues under
different management scenarios; 6. Identify areas for further research and development.
APPROACH
The IAWG has formally met 12 times over the past four years. In addition, international seminars have been held on several occasions to bring together regulators and researchers to discuss economically and environmentally sound ash management practices. For example, the IAWG has held two major seminars on leaching protocols, one in Europe and the other in North America. Most of each meeting was devoted to review of specific topics as selected from the original outline. Members of the group prepared for the meetings by gathering together information on the selected subject(s) and provided this material to the group at the meetings. The discussions that ensued lead
5
to the development of draft chapter outlines, which were used to develop the full chapters. Draft chapters were then distributed to the sponsoring agencies for review and comment. In addition to the original mandate of the Group, the IAWG has been designated as a full Activity under the International Energy Agency’s Bioenergy Agreement - Conversion of MSW to Energy Task XI. Consequently, the IAWG now has a mandate to meet through 1994 to consider revisions to the final report based on the review of any new evidence from current research topics.
Benefits
The IAWG is committed to providing the sponsors with the information necessary to make sound decisions regarding MSW incinerator residue management. 1. Outline criteria for data gathering to assist other researchers with project development and generation of quality comparable data; 2. Develop a framework on which to build characterization protocols for assessing incinerator residues;
3. Provide a bench-mark for other researchers by compiling the data and identifying areas requiring further research; 4. Enable regulators, the public and industry to make decisions on the environmental acceptability
of alternative management options for the residues; 5. Develop a framework for international collaboration on managing other wastes.
Sponsorship In order to ensure objectivity, the IAWG sought out funding based on equal contributions from a large number of government agencies, private organizations and companies. The IAWG gratefully acknowledges the support provided by: Maior Soonsors Asea Brown Boveri Ltd Danish Ministry of Energy Energy, Mines & Resources Canada Environment Canada European Economic Community KernforschungszentrumKarlsruhe International Energy Agency Integrated Waste Services Association Minor SDonsors American Society of Mechanical Engineers Greater Vancouver Regional District Northeast Waste ManagementOfficiabAssociation
International Lead Zinc Research Organization LAB France
NOVEMlRIVM Netherlands Swedish National Board for Industrial & Technical Development United Kingdom Department of the Environment United States Environmental Protection Agency Wheelabrator EnvironmentalSystems Inc.
New Jersey - Solid Waste Administration Waste Processing Association Netherlands
6 REPORT OUTLINE In order to develop better options for the management of MSW incinerator ash, full characterisation of the different residues and knowledge of the factors which influence those characteristics are necessary. Consequently, the scope of the report was expanded far beyond that originally anticipated. The current document now includes discussions on a wide range of topics directly related to ash management, including:
o the physical and chemical composition of municipal solid waste, and a summary of how it is managed in the various countries;
o the variations in incinerator and air pollution control technology and how these will influence the characteristics of the residues; o the regulations governing the operation of incinerator facilities and residue disposal; o recommended sampling and analysis methodologies for the residues; o the characteristics of various residue streams based upon the recommended characterisation protocol; o the fate of elements during incineration as a function of waste feed input, incinerator/APC technologies and operation; o the leaching of residues, including a discussion of fundamental physical and chemical aspects, as well as interpretation of appropriate test methods, leaching modelling and detailed discussions of laboratory and field leaching data; o the potential fate of elements in the environment as a function of management practices; o the treatment and modification of residues including a generic outline of possible alternatives and detailed discussions of solidification/stabilization,separation, and thermal treatment. o the chemical, physical and leaching properties of products made from treated incinerator residues; o the disposal or utilization of residues, including a review of the alternatives available and the influence of fate and transport systems, i.e., short and long term impacts; and o identification of areas for further study.
Overall, the report represents a comprehensive examination of the complex issues pertaining to MSW incinerator residues. The IAWG has prepared a summary of the full report which will be distributed in conjunction with this conference. It is anticipated that the full document will be published in hard-cover form and made available to the public in the near future.
Environmental Aspects of Conshuction with Waste Materials JJJM Goumans, HA.VM der S I w t and l71.G.Aalbers (Editors) @I994 Elsevier Science B.K AN rights reserved.
7
International Progress in Solid Waete Management John H. Skinner, Ph.D. President of ISWA the International Solid Waste Association Bremerholm 1, Copenhagen K, Denmark DK 1069
Introduction: Solid Waste Management, The Environmental Issue of the '90's. Solid waste management has moved to the forefront of the environmental agenda. The level of activity and concern by citizens and governments worldwide have reached unprecedented levels. Nations are considering restrictions on packaging and controls on products in order to reduce solid waste generation rates. Local and regional governments are requiring wastes to be separated for recycling, and some have even established mandatory recycling targets. Concerns about emissions from incinerators and waste-to-energy plants have resulted in imposition of state-of-the-art air pollution controls. Landfills are being equipped with liners, impervious caps and leachate collection systems, and gas and groundwater is being routinely monitored. There is wide scale public opposition to siting of new solid waste treatment and disposal facilities. As a result, the costs of solid waste management are increasing rapidly. Previously considered a local issue, it is now clear that solid waste management has international and global implications. Concerns about transboundry shipment of hazardous waste has led to the adoption of the Base1 Convention by the United Nations. Recognizing the interrelationship between solid waste standards and economic development, the European Community is moving forward to harmonize waste disposal requirements in member countries. Around the globe countries are discovering thousands of sites where hazardous wastes have been spilled, dumped or otherwise discarded resulting in contamination of soils, surface waters and ground water. The economic costs of clean-up these sites will stress national economies and at the same time offer enormous international business opportunities. Solid waste management in countries with developing economies poses a special set of problems. In these countries quite often financing is not available for the construction of waste treatment facilities, and there is a lack of trained personnel to operate waste management systems. Also, there are generally no regulations or control systems, no administrative body responsible for solid waste control and no obligation for industry to dispose of wastes properly. The United Nations Environment Programme has focused on solid waste management in developing economies as a priority concern. More than ever before, solid waste management policy makers world wide need sound and reliable information on the technical performance, environmental impact and costs of solid waste
8
collection, recycling, treatment and disposal systems. ISWA, the International Solid Wastes and Public Cleansing Association is putting forward a number of programs that are trying to address that need.
The Mission and Organization of ISWA. The objective of ISWA is to promote the adoption of effective and economically sound solid waste management practices that protect the environment and conserve materials and energy resources. ISWA is a professional association open to members from all countries in the world. Its activity is solely in the public interest through professional development of its members; it does not pursue any commercial or political aims. ISWA is truly an international organization in that its governing body, the General Assembly, is made up of National Members from 20 countries around the world. Most countries with an established solid waste management infrastructure hold National Membership in ISWA. National Members must be national organizations representing all professional activities related to solid waste management in the member country. National Members are encouraged to form national committees of solid waste professional associations within their countries to assure a broad representation in ISWA. It is this international network of National Member organizations that provides ISWA the ability to reach thousands of solid waste professionals throughout the world. ISWA also has over 700 individual and organizational members in over 60 countries. Recognizing the special solid waste management problems in developing countries, ISWA also provides a Development Membership category pending the establishment of a fully functioning National Member organization.
ISWA Programe. ISWA carries out its mission through a series of efforts to collect and disseminate information to its members. The ISWA Journal, Waste Management and Research is published six times a year by Academic Press and has a ten year history of successful issues containing high quality peer reviewed articles. Our newsletter, the ISWA Times is published quarterly and provides practical and useful information to its readers. The ISWA Yearbook, the International Directory of Solid Waste Management and Public Cleansing, provides extensive listings of companies and organizations in the solid waste field, as well as a wide range of articles summarizing activity throughout the industry. ISWA sponsors and cosponsors a number of conferences, workshops and symposia. Important ISWA conferences and congresses for the next several years include: 1994 ISWA Annual Conference, in conjunction with the UK Institute of Wastes Management, Torbay, UK, June 14-17, 1994.
9
ISWA 25th Anniversity Congress, Vienna, Austria, October 16-20, 1995. ISWA Quadrennial Congress, Yokohama, Japan, October 27November 1, 1996. In order to provide the opportunity for the development of specialized ISWA activities, working groups on the following seven subjects have been established: Hazardous Waste Sanitary Landfill Incineration Recycling and Waste Minimization Collection and Transport Sewage and Water Works Sludge Biological Waste Treatment. ISWA members can belong to these working groups and engage in practical information exchanges with members from other countries. Through these working groups ISWA holds many specialized symposia and workshops and has developed an international solid waste professional book and report series.
Integrated Solid Waste Management. ISWA members and most other solid waste management professionals recognize that there is no single, simple solution to solid waste problems. Instead an integrated approach is necessary combining the elements of several techniques. In the United States, the Environmental Protection Agency published The Solid Waste Dilemma: An Agenda for Action, which outlines an integrated set of strategies for dealing with solid waste management. These strategies are very similar to those recommended by the European Commission, the United Nations Environment Programme and countries around the world. Integrated solid waste management is a comprehensive strategy involving four key elements applied in a hierarchial manner: 1.
Reducing the volume and toxicity of the solid waste that is generated,
2.
Recycling or reusing as much as possible of what is generated,
3.
Recovering energy from the remaining waste through combustion systems equipped with the best available pollution control technology, and
4.
Utilizing landfills with adequate environmental controls.
In the following sections each of the elements of this strategy will be discussed in turn. Also recent data on U.S. practices will be presented for purposes of illustration.
10
Waste Reduction. Waste reduction activities are important to halt or slow down the increasing rate of waste generation per-capita. For example, the most recent data from the U . S . indicates the total amount of municipal solid waste increased from 180 million tons in 1988 to 196 million tons in 1990, which represents an increase in the per capita generation rate from 1.82 to 1.95 kg. per person per day. Waste reduction has several aspects, all of which should be addressed. One is toxicity reduction, in which the nature of waste is changed by reducing manufacturer's use of toxic materials in consumer products. Another is volume reduction-cutting the amount of waste generated by using less material in the first place. A prime example of this is a reduction in packaging. Waste reduction also includes encouraging the production of products that can be recycled more easily, such as shifting from multimaterial to one-material packaging. Other options to reduce wastes include the redesign of products, material use changes, and restrictions on specific product types. The approach to reducing waste must be broadly based incorporating actions that can be taken by industries, individuals, commercial enterprises and governmental agencies. Industry can reduce waste through raw material substitution and redesign of products and processes. Individuals, commercial enterprises and agencies can use their purchasing power to create a demand for low waste products or items produced from recycled materials. Governments should investigate the use of economic and other incentives to encourage waste reduction. Waste reduction efforts also need to focus on consumer behavior. Education and information dissemination programs can be effective means of causing desired behavioral and attitudinal changes. There are many cases of successful reduction of wastes produced by industrial processes. Experience has shown that modifications to industrial processes that reduce waste also result in lower raw material, energy and waste disposal costs. Productivity is often enhanced and liabilities related to release of hazardous substances are reduced. The fact that waste reduction quite often pays has been demonstrated repeatedly.
Recycling. There are two basic approaches to recycling solid wastes. The first involves separating recyclable materials by the waste generator and separately collecting and transporting these materials to recycling markets. The second involves collecting mixed wastes or commingled recyclable materials and separating them at a central processing facility. In the U.S., through a combination of these practices the percentage of the municipal solid waste stream recovered for recycling or composting increased from 13 percent in 1988 to 17 percent in 1990. Prior separation of recyclable materials has the advantage that the materials are not contaminated by other wastes.
11
However, this requires the waste generator (e.g. householder) to separate the wastes correctly and store them in separated form. Also, the generator needs to transport the separated material to recycling centers or separate or compartmentalized collection vehicles need to be used. Key factors in success of preseparation efforts are the cooperation and willingness of the generator to participate in the program over the long tern, and the additional collection and transport costs that may be required. Mixed solid waste can be separated for recycling at local processing centers or materials recovery facilities (MRFs). Inn the U.S. for example, there are over 200 MRFs in operation, construction, or advance planning stages. Some plants process segregated recyclables; others separate mixtures of glass bottles, aluminum cans and steel cans; still others process mixed residential or commercial wastes, separating the recyclable materials. The success of these plants depends on the processing costs and the quality of the recyclable material produced. A major factor affecting recycling economics is the difference in cost between disposal and recycling. In many locales this cost difference is narrowing. For example, in the U.S. the disposal fee for landfills and waste-to-energy plants has increased dramatically over the past 10 years. Today, on the average, a solid waste management system in the U.S. can avoid $25 to $40 per ton in disposal costs for every ton it recycles, whether or not it gets paid for the recycled material. In some locations the savings are even higher. A major recycling impediment is the question of continued viability and availability of secondary materials markets. Can manufacturers expand markets so they can accept all of the material that is being collected by the new residential programs? Topping the list of problematic waste material markets is the market for old newspaper. In the late 1980s, there was dislocation in markets due to an oversupply created by the large number of municipal collection programs that were all bringing new supplies to markets simultaneously. Many U.S. municipalities were forced to pay to recycle collected newspapers. Current market figures show that the value of old newspaper varies from $40/ton to a -(negative) $4O/ton. Problems are also being experienced in other recycled material markets, including those for glass, plastic and for compost produced from yard waste and mixed municipal solid waste. There are some encouraging trends that suggest the problem of oversupply of old newspaper could be reduced. Some newsprint producers in the U.S. and Canada have announced plans for new facilities to make use of recycled fiber. Others have undertaken feasibility studies for new facilities. It is important to understand that separation of materials from the solid waste stream in itself does not constitute recycling. Recycling only occurs when these materials are incorporated into products that enter commerce. Therefore requirements to separate certain fractions of materials from waste may produce a supply of materials, but these requirements
12
in themselves will not ensure recycling. In fact, if markets for these materials are not found, and the materials are subsequently disposed of, all of the costs of recycling are experienced with none of the benefits. Similarly, requirements to incorporate separated waste materials in products will not result in recycling unless these products are of a quality and price that they successfully compete in the marketplace. To analyze the economic feasibility of recycling one must consider the price received for the recycled material, the solid waste collection and disposal costs avoided and the costs of separation, collection and processing the separated materials. In making these cost comparisons it is important that all environmental costs and benefits are internalized. Also, the benefits to future generations in terms of natural resources conserved or landfill space conserved must be considered. Any virgin raw material subsidy that artificially drives down the price must be accounted for so that virgin materials and recycled materials compete in an equitable manner. Similarly, procurement specifications that arbitrarily discriminate against recycled materials should be eliminated. In order to effectively carry out successful recycling programs, solid waste managers must operate in a business-like manner as raw material suppliers. They must treat the users of their materials as customers. This means they must produce recyclable materials meeting the customer's material quality requirements, and offer recyclable materials at a price competitive with other material supplies. They must operate their separation, collection and processing systems to produce competitively priced, quality materials at the lowest possible costs. The elements of success of a recycling operation are the same as for any successful business; staying close to the customer, understanding and meeting their quality needs and operating in a cost effective manner to produce a competitively priced product.
Combustion with Energy Recovery. Waste-to-energy facilities can achieve an 85% volume reduction in the waste burned. In the U.S. these plants have increased their handling of solid wastes from a negligible percentage of the municipal solid waste stream in the early 1980s to almost 16% of municipal solid waste today. Waste-to-energy plants have faced two main problems in their fight to win public acceptance: air pollution concerns and the heavy metal content of the ash generated in the combustion process. On January 14, 1991, the U.S. EPA issued regulations for new municipal (New Source Performance Standards, or NSPS) and guidelines for existing plants. These standards incorporate good combustion practices, emissions monitoring and highly efficient air pollution control systems to control organic emissions (dioxins and furans), metals, acid gases and other pollutants. The standards are similar to those used in other countries to regulate incinerators. EPA estimated that in 1994 the national costs of these rules will be $170 million a year for new
13
facilities and $302 million a year for existing facilities. Therefore in the U.S. there will be a substantial financial investment to upgrade the environmental performance of municipal incinerators. Another environmental concern that has developed over the past several years involves the disposal of ash residues from municipal waste incinerators. Usually significant amounts of lead, cadmium, zinc, mercury, arsenic, and other metals are found in incinerator ash, especially fly ash. The environmental concern is the potential for these metals to leach out of the residue when disposed of with other wastes in a sanitary landfill. This has led to the utilization of monofills or landfills used solely for ash disposal. In September 1992, the U . S . EPA issued an opinion that ash generated by solid waste-toenergy incinerators is not considered a hazardous waste under Federal law and that the new requirements for solid waste landfills will ensure that ash is disposed of in a manner that protects human health and the environment. Also, technologies have been developed to chemically extract metals from incinerator ash or to solidify and stabilize the ash by adding cement or kiln dust to create a concrete like substance. While these technologies are effective in removing or stabilizing metals, they do result in added disposal costs. Some of these costs can be offset if the ash is treated to the extent that it can be used safely and sold as an aggregate or building material. In the U.S. over 8 million tons of incinerator ash are produced annually.
Landf i 11s. Landfill technology has advanced very rapidly over the past decade. Today's state-of-the-art landfills are equipped with leachate collection systems, liner systems, systems for control of landfill gas, groundwater monitoring, closure and post-closure care and much more. The objective is to ensure that landfilling is performed in a manner that greatly reduces the change of environmental degradation--and also, that any degradation that occurs is quickly detected and remediated. In the U . S . the number of landfills continues to decrease, two main consequences are seen: first, communities face longer transport distances to deliver their solid waste to disposal sites; secondly, several large facilities, designed to serve a limited number of communities for a given number of years, are seeing their lifespans drastically foreshortened by the influx of waste from outside their service areas. Due to more stringent landfill regulations, many small facilities will shut down because they will be unable to meet the new requirements. A hoped-for-result is a decrease in opposition to landfills, stemming from greater public faith in the environmental soundness of facilities that are allowed to operate.
14
Some observers believe the combination of continued strong public opposition and tougher landfill rules will result in a system of large, remotely located regional landfills. Signs of this can be seen already.
A Strategy for Continuous Improvement. Over the past 20 years there has been substantial progress in addressing solid waste problems. However, many problems still exist and we understand them to be very complex. To deal with them, the strategies that have been used in the past will not be enough. As we move towards the 21st Century, a number of forces must come together to lead to continuous improvement in solid waste management. These include:
Continued, Rigorous Enforcement of Environmental Laws and Environmental standards must be rigorously Regulations. enforced in order to assure the public that our solid waste systems are operated in ways that protect human health and the environment. Enforcement must create an incentive for compliance with environmental standards. It must level the playing field so that violators are not at a competitive economic advantage to the good citizens that comply. Waste Reduction as the Strategy of Choice. The traditional approach to solid waste management has been a pollution control strategy where wastes are collected and treated or disposed of after they are generated, or waste is cleaned up after it has occurred. A waste reduction strategy is different, it means not creating the waste in the first place. This can be accomplished through changing product designs, increasing process efficiencies, and extending product lifetimes. Waste reduction results in reduction in waste treatment and disposal costs, reduced liability for environmental damages, lower raw material costs and process efficiencies. Risk-Based Decision Making. Solid waste management decision-making must be based on a comparative analysis of the relative environmental risks of the various options available. Quite often there is public opposition to a particular facility because of concern about environmental risk. While the public expresses a preference for recycling over waste-to-energy or landfill, it is often forgotten that the recycling process itself produces waste or residuals that must be managed or disposed of (eg. waterborne wastes produced from the deinking of recycled newsprint or increased air pollution from additional collection vehicles). In order to make an informed decision, the risk of one option must be compared to the alternatives. Priorities must be based on relative environmental risk. In order to do this we need to develop better and more reliable risk assessment methodologies and put them to use. Significant advances need to be made in our capabilities to assess the risks to ecological systems. An investment in risk assessment research will certainly pay off. Public Information to Encourage Voluntary Action. Providing data and information to those who make or influence decisions can
15
lead to voluntary actions with significant environmental benefits. A good example is the Toxic Release Inventory (TRI) in the US. Each year industries are required to publish the total release of certain toxic wastes to the environment and make this information publicly available. When the public for the first time realized the total environmental releases from all of these plants and facilities, they demanded that something be done about it. This led to the establishment of the 33/50 Program. Under this program companies voluntarily agree to reduce their waste discharges of by 33 percent by the end of 1992 and 50 percent by the end of 1995. Over 700 companies have made written commitments which will reduce the discharge to the environment of 150,000 tons of toxic chemicals by 1995. Information is a powerful tool which can stimulate real results.
Environmental Education. As the above example shows an informed public can be an effective force in environmental protection. However, professionals in the field must do a much better job in explaining to the public the true nature of environmental risks and what can be done about them. The National Environmental Education Act which was passed in the US in 1991 provides some excellent vehicles for doing this including (1) support for environmental curriculum development, (2) assistance for teacher training and ( 3 ) scholarships and fellowships for environmental science and engineering. It is very important to increase environmental literacy to build public support for environmental programs and train future generations of environmental professionals. Economic Incentivee. Market based economic incentives can be used as an alternate to regulation or as a means of making regulations more effective. For example, the liability standards under the US Superfund legislation make a waste generator liable for environmental damages caused by that waste. This produces a very strong economic incentive for waste reduction and on-site waste treatment. Other economic incentives such as pollution charges and deposit systems should also be evaluated for future solid waste management policies. Research and Development. A sustained, long term research and development effort is necessary to improve our understanding of the environmental impacts of solid waste management systems and develop solutions. What are the health effects of environmental releases from solid waste management systems. How do pollutants move through the environment and change in their physical and chemical form? What are the routes of exposure for human populations and ecological systems? How can we monitor and detect pollutant levels in real time? What are the most cost effective approaches to waste reduction, recycling, combustion nd disposal. These are just a few of the questions that research must address. However, research should not be limited to technological and physical science issues. Research into the social and economic aspects of solid waste management is necessary to understand and better design economic incentives and information and education programs.
16
-
Technology Transfer Domestic and International. Research and development alone is not enough, the results must be transferred into the field as new and improved solid waste management systems are developed. Therefore, outreach efforts to apply the results of research are essential. This is especially true on an international basis where there are potentially large market opportunities for cost effective environmental technologies. Technology transfer to countries with developing economies is especially important, if we expect these countries to be able to participate effectively in improving the global environment. Integration of Solid Waste Management Policy With Other Policies. Other national and international policies can have as strong or stronger influence on solid waste management as can environmental policies. Consider the effect of: (1) energy policy on the incentives for waste-to-energy facilities, (2) transportation policy on freight charges for recycled materials, (3) agricultural policy on the uses of sludges as fertilizers or soil conditioners. Other examples include the effect of financial policy on investment into environmental technologies and military policy's effect on clean-up of defense installations. Solid waste management professionals must play a role assuring the solid waste management implications of these policies are assessed in national and international forum. These are the issues that will be facing the solid waste professional of the future. There remains a tremendous opportunity to improve waste management through technological development. However it will be necessary to combine technical and engineering skills with risk assessment, market forces, public information and education, enforcement strategies, pollution prevention, research and development and technology transfer. Solid waste management professionals must show leadership in developing broad based strategic initiatives to bring about continuous improvement in integrated solid waste management. To find out more about ISWA programs and activities including membership information contact the General Secretariat in Copenhagen Denmark.
Environmental Aspects of Construction with Waste Materials JJJ.M. Goumans, H A van der Sloot and Th.G. Aalbers (Edtors) 01994 Elsevier Science B.V. All rights resewed.
17
LIFE CYCLE ANALYSES; RESULTS OF SOME C A S E STUDIES
J . Cramer, TNO Centre for Technology and Policy Studies,
P . 0 . Box 541, 7300 AM Apeldoorn, The Netherlands
INTRODUCTION
Environment-oriented product policy has become a central focus of Dutch environmental policies. The aim of this policy is Yo prevent, or at least reduce, the effects on the environment caused by products in the various phases of the product life cycle'. Up to now the government has placed the emphasis mainly on developing methods of analyzing the environment-friendliness
of
products
(environmental life
cycle
analyses). Experience has been gained in the application of life cycle analysis for various purposes. Of particular interest a r e four different applications :
-
the use of life cycle analysis in 'eco-design'
-
the use of life cycle analysis in 'integrated chain management'
-
the use of life cycle analysis in information exchange between
the use of life cycle analysis in product comparisons
producers in a product chain
In this contribution the experiences with these four types of application will be discussed on the basis of case studies. Many of these case studies have been conducted on behalf of the Netherlands Company for Energy and the Environment (NOVEM) (often together with the State Institute for Public Health and Environmental Hygiene (RIVM))
.
18 LIFE CYCLE ANALYSIS
Before dwelling upon the various applications of life cycle analysis I shall first explain in more detail the principles of life cycle analysis. A
life cycle analysis of a product is the systematic analysis of the environmental impact of a product, calculated over the entire life cycle of a product ('from cradle to grave'). Life cycle analysis is often combined with an assessment of the environmental impact of a product (together called 'life cycle assessment' ( L C A )
.
A full life cycle assessment is made up of five components which
together form a comprehensive structure. These components are (Heijungs et al. , 1992):
*
goal definition, including a consideration of the type of potential application;
* *
inventory analysis of environmental interventions ; classification and modelling of environmental interventions on the basis of their potential environmental effects;
*
evaluation of the product based on its potential environmental effects ;
*
improvement analysis aimed at changes which are desirable on environmental grounds.
There are various bottlenecks standing in the way of the further development and application of the LCA method, principally relating to its complexity. In practice, setting up an extensive LCA is extremely time-consuming , especially for individual companies. The development of environmental indicators has been promoted as a way of reducing the complexity of the LCA method. Environmental indicators are index numbers that express the environmental impact of
19 particular products in terms of a limited number of relevant units, specifically raw materials,
energy,
emissions, nuisance,
waste,
recyclability, reparability and the product's life. They a r e intended for communication purposes, for initiating an increase in environmental awareness and for regulation purposes (TAUW Infra Consult, 1992). Four main stages can be distinguished in the process of establishing environmental indicators (TAUW Infra Consult, 1992). Stage one involves setting the objectives : what a r e we establishing environmental indicators for? Stage two consists of the stock-taking: the entire chain is described and checked for impacts on the environment. Stage three
involves the classification: all the pieces of information obtained a r e classified according to workable categories. Finally, stage four consists of the evaluation: the data collected is evaluated, assessed and translated into environmental indicators.
I n principle, the purpose of environmental indicators is to present the
data from a LCA in a clear and simple form and so to simplify the exchange of information between the parties in the market. In a number of recent trial projects, for instance, environmental indicators have been drawn up for paper, plant trays, margarines, printer ribbons, paints, light sources, insulating plates and a particular make of high performance heating boiler.
It proved possible, in principle, to present L C A data in terms of environmental indicators. However, i t did not usually reduce the workload. In order to establish environmental indicators one still needs some form of life cycle assessment. A s a result, environmental indica-
20 tors a r e regarded as the results of a particular type of L C A , although the approach adopted for the LCA may vary.
THE USE OF LIFE CYCLE ANALYSIS IN PRODUCT COMPARISONS Up to now, a great deal of time and energy has been devoted to applying L C A to product comparison. For instance, what a r e the effects on the
environment of, respectively, a wooden, an aluminium or a plastic window frame or of an ordinary light bulb or a long-life light bulb or the milk carton as opposed to the milk bottle throughout the entire product life cycle ? Such comparisons are made, for example, in the context of ecolabelling or of debates elicited in society about the environmentunfriendliness of particular products. The results of these product comparisons have frequently been the subject of debate. This is due to the fact that various assumptions have to be made in assessing the environmental performance of a product. For instance, the results depend on how the boundaries of the object of study are defined. Moreover, the results will vary according to the technical qualities of the two products to be compared. For example, comparing a milk bottle with a milk carton requires assumptions about the technical qualities of both types of products. Finally, a lack of information on specific environmental aspects forces one to assess on the basis of incomplete data.
Thus, the value of LCA in product comparison is very much dependent on the acceptance of the data used, the system boundaries and the methodological aspects of
data interpretation, classification and
evaluation. In order to minimize the degree of subjectivity in assessing the environmental performance of products, attempts are being made to
21 establish an internationally accepted methodology. At the international level, SETAC (Society of Environmental Toxicology and Chemistry) is the leading forum for the development of such an accepted methodology.
THE USE OF LIFE CYCLE ANALYSIS I N 'ECO-DESIGN' The LCA approach ha5 not generated a great deal of experience in the systematic improvement of products. It does, however, play a role in
' environment-oriented p r o d u c t development i n companies', commonly known as 'ecodesign'
. Environment-oriented product development can
be described as: designing and developing a product in such a way that environmental criteria a r e given equal weight to other criteria.
This
method seeks to limit, or if possible prevent, deleterious effects on the environment throughout the product's life cycle.
A number of demonstration projects have already been conducted in the
Netherlands in the field of environment-oriented product development, especially in the context of programmes such as Ecodesign and Milion. The designs have included an environment-friendly office chair, a reusable plant tray for flower auctions and a new package for dried coffee creamer. There has also been a growth in international demand for environmentally aware product design and development. This was also evident from the first European 'Design for the Environment' Conference which was held in Nunspeet, the Netherlands between 21-23 September 1992 (National Reuse of Waste Research Programme, 1992).
To date, the demonstration projects have been instituted mainly in cooperation with interested individual companies. The major problem in the implementation of these demonstration projects has been a lack of
22 environmental information. This could be a lack of information about the environmental properties of substances and materials that might be used,
for instance. Nor was there always adequate information
regarding the application of materials in products, the use of products or the disposal of the waste produced by them.
In view of this lack of information, it proved difficult in practice to use an extensive LCA. Moreover, resource limitations and time constraints led to the application of a simplified, 'streamlined' LCA. This implies that only the major environmental bottlenecks in the product chain were identified and improvements to the product were concentrated on these. The selection of these improvement options is still largely based on an ad-hoc approach. Improvement options are not systematically compared.
As a result,
environment-oriented product development initially
developed separately from the LCA method. This is gradually changing, however. On the basis of the experience already gained with environment-oriented product development there a r e a number of follow-up steps that could be taken (Cramer ( e d . ) , 1993). The first step would be to extend the attention to environment-oriented product development within the company as a whole. Up to now the projects have concentrated mainly on the design process in a company, which only involves one department.
A second advance would be to intensify the linkage between process and
product improvement. In the context of waste prevention and emissions, companies still concentrate mainly on process improvement; in this respect, the product is generally ignored.
23 On the other hand, with environment-oriented product development the production process is generally unaffected.
A third step would be to stimulate product improvement throughout the
entire chain o r column of companies involved in the manufacture of a certain product. This requires intensive cooperation between the various suppliers and customers in a given chain. The basis for such collaboration is the effective exchange of information between producers and within a product chain.
In response to the three problems raised above a third type of application has been initiated: the application of integrated chain management within companies.
THE USE OF LIFE CYCLE ANALYSIS IN 'INTEGRATED CHAIN MANAGEMENT' Integrated chain management has become a central concept in Dutch environmental policies. It is not a familiar concept abroad. In other countries, for example, the terms 'lifecycle management' o r 'productstewardship' a r e used in much the same way as the term 'integrated chain management'. Integrated chain management aims to ensure that substance cycles in product chains a r e managed in an environmentally, socially and economically responsible manner.
To realize such a 'sustainable
substance cycle' the relevant product must generate the least possible emissions and consume the least possible amount of energy in all phases of that product's life cycle (from extraction of the raw materials, the manufacture of the material and the product, via use by the consumer,
24 up to the disposal of the product in the waste s t a g e ) . This is why 'integrated chain management' transcends the individual company level and includes the whole product chain from cradle to grave. To encourage the adoption of integrated chain management in industry, in association with three companies TNO has developed a concrete methodology for i t s practical application. The participating companies were: AKZO Chemicals, EPON ( t h e Electricity Production Company for East and North Netherlands) and Van den Bergh Nederland (Cramer et a l . , 1993). In principle, this methodology does not differ fundamentally from that
.
of 'ecodesign' In practice, however, the present approaches diverge. The 'ecodesign' approach is to generate and select improvement options in a rather ad-hoc way, while 'integrated chain management' attempts to do this in a very structured way. Central to the method TNO and the three companies have developed is the identification and comparison of various options for bringing about environmental improvements both in the company itself and throughout the rest of the product chain. In this way, the company taking the initiative for chain management can act as a catalyst for change throughout the entire product chain.
The first step in the methodology for implementing integrated chain management is to select one major environmental bottleneck to be improved. This identification of environmental bottlenecks is based on a simplified, 'screening' life cycle analysis. The next step is a structured comparison of various improvement options on the basis of an ecological assessment.
25 To provide an indication of the extent to which a certain option constitutes an improvement from an environmental perspective, the "concept of environmental merit'lis introduced. "Environmental merit" is the difference in the environmental impact before and after the implementation of a given improvement option (including both the primary and secondary environmental impacts of the improvement option). After comparing the potential improvement options on the grounds of a purely ecological assessment, the social, technological and economic feasibility of the options and their suitability for the specific company a r e assessed. The option finally adopted will be determined by weighing up the ecological assessment and the social, technological and economic feasibility and the company-specific feasibility. The resulting step-by-step plan for implementing integrated chain management in individual companies has been tested in the participating companies. Although the approach followed should be further developed and evaluated, the experiences of the companies involved have been positive.
THE USE OF LIFE CYCLE ANALYSIS IN INFORMATION EXCHANGE BETWEEN PRODUCERS I N A PRODUCT CHAIN In order to bring about actual changes in products and production processes the information exchange between producers in a product chain is crucial. To date no structural system for such an exchange of information exists. In this context, the introduction of 'environmental product profiles' ( E P P ) is being investigated within the framework of
.
the Economic Commission for Europe ( E C E ) An environmental product profile is defined as an objective set of relevant environmental
information relating to a product and aspects of its life cycle, especially for producers and professional users (Weterings et al., 1993). The question, however, is which type of information should be exchanged among producers. To answer this question three case studies were carried out by TNO and TAUW Infra Consult in close cooperation with manufacturers. Three (groups of) products were selected: printer ribbons, roofing plates and carpets. The aim of the experiments was to test which type of information exchange was considered useful and advisable. The information presented was based on a 'screening' life cycle analysis. The quantitative data were presented by means of environmental indicators, such as the consumption of raw materials, energy consumption, greenhouse gas emissions, emissions of gases contributing to acidification, etcetera. In order to compare different forms of presentation, two versions with different levels of data aggregation have been developed: -
.
a long version (approx 10 to 20 pages) in which the environmental data consist of a classification of the (potential) environmental effects per step in the product chain and an evaluation of the most important environmental effects;
-
a short version (approx. 5 to 10 pages) in which the environmental data are limited to the evaluation of the most important environmental effects.
In response to these two sets of data those interviewed criticized the way in which environmental data were presented by means of environmental indicators. The amount of oil, gas and coal used, and the volume of several categories of emissions and waste products were quantified per product unit. Although, in general, quantitative data
27 were regarded as useful, most respondents felt the way in which the data w e r e presented was meaningless. They lacked the expertise to understand the quantitative indicators without a lengthy explanation of the reasoning behind the methodology and about the reliability of the resulting data. A few respondents even preferred one grade for all environmental aspects together. Based on the experiments it was concluded that authorities should take a cautious step-by-step approach in implementing E P P s. The content and presentation of environmental product profiles should be further developed
in
organisations
close
cooperation with
manufacturers
and
their
.
CONCLUSIONS
The above examples show the importance of life cycle analysis for various applications. Some bottlenecks in the use of life cycle analysis come to the fore as well. Due to the complexity of life cycle assessment ( L C A ) (including life cycle analysis) an extensive L C A is very time-
consuming. Moreover, an internationally accepted methodology of LCA i s still in development. In practice, therefore, simplified, so-called screening L C A ' s a r e often used. While the approaches used in various applications have been developed separately, there now seems to be a demand for closer attuning of these various approaches. This is a hopeful sign. Ultimately, coherence is needed in the methodologies for assessing and improving the environmental performance of products. This growing consensus will
28 stimulate both government and industry to put environment-oriented product policy into practice.
REFERENCES Cramer, J . et a l . , Theorie en praktijk van integraal ketenbeheer (Theory and practice of integrated chain management), NOVEM/RIVM, TNO- Apeldoorn, 1993.
Cramer, J. (ed. )
. , Productgericht
milieubeheer (Product-oriented
environmental management), Advisory Council for Research on Nature and Environment ( R M N O ) , publication nr. 78, Rijswijk, 1993.
Heijungs, R . et al., Environmental life cycle assessment of products, Guide and Backgrounds, NOVEM/RIVM, Centrum voor Milieukunde, Leiden, 1992.
National Reuse of Waste Research Programme, First NOH European conference; 'Design for the environment',
21-23 September 1992,
Nunspeet, the Netherlands, Utrecht , 1992.
TAUW Infra Consult, Environmental indicators; An evaluation of the pilot projects, NOVEM/RIVM, TAUW Infra Consult, Deventer, 1992.
Weterings, R . et al., The exchange of environmental product profiles between professional users : three case studies in the Netherlands, Discussion paper to the ECE-Seminar on low-waste technology and environmentally sound products, Warsaw, Poland, 24-28 May 1993, TNOApeldoorn, 1993.
SECTION 2: Environmental Aspects
This Page Intentionally Left Blank
Environmentol Aspects of Consttuction with Wmte Moterials JJJM Goumons, H A von &r SImt and l71.G. Aalbers (Edtom) a1994 Elsevier Science B. K All rights reserved.
31
Chemical Processes Controlling the Mobility of Waste Material Contaminants in Soils
L.G. Wesselink', P.M. Dekker' & Th. G. Aalbers' 1) National Institute of Public Health and Environmental Protection, P.O. Box I 3720 BA
Bilthoven, the Netherlands.
ABSTRACT
Effects of solid waste materials on the chemistry of underlying soils are investigated experimentally and through modelling of fundamental soil chemical processes.
INTRODUCTION
Solid waste materials are increasingly re-used in construction works [ 11. Release of contaminants from these waste materials may affect the chemistry of underlying soils and groundwaters. Metal ions are known to bind to soil organic matter and (hydr)oxides [ 2 ] , which strongly reduces their mobility in soils. Current knowledge of the fundamental mechanisms underlying these binding processes has been compiled in the ECOSAT model [3, 41. ECOSAT combines chemical speciation calculations with surface complexation- and transport models. Here, we report laboratory column experiments in which effects of leachates from steel- and phosphorous slag (P-slag) on four sandy soils were studied. The ECOSAT model will be applied to these experiments to identify and quantify processes that control contaminant mobility in soils under waste materials. To assess the potential hazards of waste materials to soils and groundwaters, knowledge of these processes is indispensable 11 I.
32 MATERIALS AND METHODS
Column studies
Steel- and phosphorous slag (P-slag) were percolated in large columns, up-flow with synthetic rain water @H ~ 4 . 5 )The . columns (30 cm diameter) held 90 kg of slag, and a flow rate of 0.2 US (l/kg) per day was maintained. Subsequently, slag effluents were percolated up-flow through four different columns holding 1 kg of soil, also at a flow rate of 0.2 L/S per day. The chemical composition of percolates at L/S= 0.1, 0.5, 1-5, 7.5, 10, 12.5 and 15 was analyzed. General characteristics of the soils are given in Table 1. Table 1. Soil characteristics
a is surface soil, b is sub soil, ox refers to oxalate extractable Fe aid Al
Model calculations
ECOSAT [3,4] calculates chemical equilibria involving speciation in solution, mineral
dissolution/precipitation and sorption on organic matter and hydroxide surfaces. In addition, transport (convective and dispersive) of dissolved species is calculated. Proton and metal binding to humic substances can be described by several variable or non-variable charge modules within ECOSAT. In this study, metal and proton binding was modelled using the multicomponent Langmuir-Freundlich equation (1). which considers the natural heterogeneity in binding properties of humic ligands [3, 51:
33 where O,., is the fraction of total available surface groups
(a,,,) covered with metal or proton
i, Ki the metal or proton surface complexation constant and m (O<m
RESULTS AND DISCUSSION
PH Percolates from steel- and P-slag were strongly alkaline with pH around 13 and 9.5 respectively (Fig. 1). Soils under P-slag were able to buffer the pH around a value of six throughout the experiments whereas soils under steel slag approached relatively quickly to the extreme high pH governed by steel slag (Fig. 1). The pH breakthrough observed in the soils under steel slag appeared in the order of soils D, A, C and B (extreme cases B and D shown in Fig. I), which is probably related to organic carbon content of these soils (Table 1).
Calcium Calcium was the dominant cation in both P- and steel-slag percolates, with extreme high concentrations in the latter (see Fig. 2). Calcium concenuations in steel slag percolates were close the Ca(OH), equilibrium, whereas Ca in P-slag had a lower solubility. Calcium was strongly retained in all soil columns (Table 2). Surprisingly, most soils maintained Casolubility below that of CaCO, (Fig. 2). The observed pH buffering around a value of six and Ca retention in the soils may be explained by Ca-proton (Ca-H) exchange on soil organic matter [3, 51. This is supported by observed Ca-H exchange ratios of around 0.6 (as derived from the slope in the pH-pCa diagram left of the CaCO, line in Fig. 2). This value is typical for Ca-proton exchange on humic substances [3, 51. Notice, that organic carbon contents in our soil are rather low (Table 1). Yet, even a organic carbon content of 1% may provide around 100 meq/kg-soil of carboxylic (-COOH) potential binding groups 161. In soils under steel slag, Ca inputs strongly exceeded this estimated potential binding capacity (Table 2). which explains the rapid breakthrough of pH in these soils. In Column-B (2.2% C)
,
5
14
steel-slag
12
I I
I
I D
I
W
P
I I I I I I
0
4
I
I I
I
I
I
B -
10
m
++
3
+
0
Ia
Y
c31
0 -
8
'
I I I D P-slag ..,."'.._,,I1 ..' ................................. ,... ..'..._.. .I,__ '.I...::."' __.. .._....'. .. ". ....,__.. " 't /,.: I I I II
I
i
I
'.'
,
I I jCaCO3 I I I I I
1
... 10
0
20
L/S ratio W k g soil)
,
CaCHi
I
0
4
0
8
II I
B
6
a
2
2
4
8
6
1 0 1 2 1 4
PH
Fieure 1: =-
Figure 2:
pH with dme in percolates of steel- and P-slag and soils (A, D) underlying these slags.
Calcium activities in percolates from slags and soils. Calcium ahvines were calculated with ECOSAT [3]. Srabiliry lines for CaOH, and CaCO, (at pC02=2.5) arc indicated. (+) arc pacolates from s t 4 slag and soils under steel slag and ( ) are percolates kom P-slag and soils under P-slag.
~~
35 approximately 250 meqkg of Ca was retained in the soil column until the moment of pHbreakthrough. This appears to match well with the estimated potential binding capacity of soil organic matter in soil B (220 meqkg). After the pH breakthrough, Ca solubility quickly approached that of Ca(OH),, as governed by steel slag (Fig. 2). In soils under P-slag, Ca retention during the experiment remained well below the estimated potential binding capacity. It is therefore expected that pH-buffering through Ca-H exchange in these soils will continue over prolonged time periods.
Table 2. Input-output balances of soil columns at L/S-soil=lS. Inputs are from slag percolates, outputs from soil percolates. Shown is input minus output in mmol/kg of soil material. Soils under P-slag
Ca A1 Fe
K Mg
A.
B.
C.
D.
-2.40 0.45 0.20 0.73 0.25
-3.13 0.47 7.30 0.89 1.24
-6.91 0.14 0.32 0.77 0.63
-3.13 -0.09 0.01
1.27 0.23
Soil under steel slag
________________________________
Ca A1 Fe
K Mg
-30 0.49 0.04 0.96 0.04
-153 1.07 2.37 0.80 0.47
-86 1.13 0.05 0.40 0.06
-61 1.17 0.023 0.64 0.01
Heavy metals
Generally, the pH is regarded as an important control on metal solubility 121.This is for our study illustrated in Fig.3 which shows that pH buffering toward a value of six decreases heavy metal solubility. Cadmium, Pb and Zn concentrations in our soils (below pH
7) were undersaturated with respect to mineral phases. Similar to Ca, observed pH-pM(etal) relationships between 0.5 and 1 (Fig. 4) can be explained by binding to soil organic matter surface groups [3,51.
d r
N
36
7
rl d
r
0
m ro
0
d
0 0 0 0
o n 0
I
h
N
0
I
co
M
d
37
Modelling Here, ECOSAT [3] will be applied to calculate the chemistry of the soil columns. For this purpose, a 'model' soil column is proposed in which metals and proton binding to soil organic matter is calculated. Binding to soil organic matter is described with the LF formulation (Eq. 1). The total binding capacity will be estimated from total carbon content, as discussed previously. Initial binding of metals to soil organic matter will estimated from pyrophosphate extractions (in prep.). Estimates for binding constants of metals and protons on humic compounds, and their heterogeneity, are available from recent studies [3, 5,7J.Thus it will be tested whether an independent set of model parameters can be applied to our experiments.
CONCLUSIONS
Laboratory column experiments indicated that exchange processes of H, Ca and heavy metals on organic matter control the solubility of these components in the investigated soils. Recent developments in soil science allow mechanistic modelling of these processes, which is of great importance to properly assess the potential hazards of waste materials to soils and groundwaters. Modelling results will be included in a full paper to be available at the WASCON '94 meetings.
REFERENCES
[l]
Aalbers, Th. G.et al., 1993. Milieuhygienische kwaliteit van primaire en secundaire bouwmaterialen in relatie tot hergebruik en bodem- en oppervlaktewaterbeschedng. RNM-report no. 771402006.
[2]
McBride. M.B., 1989. Reactions Controlling Heavy Metal Solubility in Soils. Advance in Soil Science, Volume 10.
[3]
de Wit, H. 1992. Proton and metal binding to humic substances. Phd-thesis, Agricultural University Wageningen, the Netherlands
38 [4]
Keizer, M.G. et al., 1992. ECOSAT, Technical note, Department of Soil Science and Plant Nutrition, Wageningen Agricultural University, the Netherlands.
[5]
Beneditti. M.F. et al., 1994. Metal ion binding to hurnic substances: experiments and modelling. J. Coll. and Int. Science. in press.
[6]
B u m a n , P. 1985. Carbon/sesquioxide ratios in organic complexes and the transition albic-spodic horizon. J. of Soil Science, 36. 355-260.
[7]
Tipping, E., and Hurley, M.A. 1992. A unifying model of cation binding by hutnic substances. Geochirn. Cosrnochirn. Acta 56, 3627-3641.
Envimnmental Aspcts of Construction with Waste Materials J.J.J.M. Goumans, H A . van der Sloot and Th.G. Aalbers (Editors) @I994 Elsevier Science B.V. All rights reserved.
39
-
Leaching of slags and ashes controlling factors in field experiments versus in laboratory tests A-M. Fallman and J.Hartlen Swedish Geotechnical Institute, S-581 93 Linkoping, Sweden, Abstract Significant differences in pH and redox potential were obtained in field experiments in comparison with the values obtained in laboratory column tests. Tested materials were blast furnace slag, steel slag, municipal waste incineration bottom ash (MSWI BA) and wood ash. All but the MSWI BA showed significantly lower pH in the lysimeter leachates. A result of the different controlling conditions in the field experiments was significant changes in leached amounts of some metals in comparison with laboratory tests. Other laboratory tests such as availability, oxidised availability and pH static tests were used to investigate the complex influences of pH and oxidation on leachate composition. 1. INTRODUCTION
Residues are the non-profitable end products in many industrial processes and energy production utilities. In combustion and smelting processes, slags and ashes are produced which in many cases have properties similar to geological materials. These materials can be used as filling and construction materials instead of being landfilled if the net environmental impact is acceptable. However, many residues have not been considered for utilisation purposes, while others have been discussed widely but are not yet accepted mostly because of possible environmental impact. It is important to identify the factors that classify a residue as usable or unusable. In the assessment of the suitability of such materials, not only the technical characteristics need to be considered but also the assessment of the possible environmental impact. During recent years, the processes governing leaching have increasingly become the focus in the discussions of leaching from residues. The actual leachate concentrations are regarded as a result of primary processes, the so-called controlling factors. van der Sloot et al (1991a) divided the factors into chemical, physical and biological factors. The chemical factors listed were: pH, redox, complexation of other components and reaction kinetics. The main physical factors relate to the water flow, while biological activities relate to changes in redox or concentrations of complexing components. Among the chemical factors influencing leaching processes and leachate concentrations, pH is so far the one studied most. Results from batch tests on coal combustion residues and MSWI residues show that leachate concentrations are highly pH-dependent for many substances (van der Sloot et al 1991b, DiPietro et al 1989). It was also shown that the same component gave different pH behaviour for different materials. van der Sloot (1992b) showed that pH dependency also changes with the amount of leachant
40
used, A study made by Hjelmar (1991a) on coal fly ash showed good agreement between laboratory column tests and lysimeter results. Some disagreement was seen for reduced ashes, where redox conditions changed during the lysimeter test. In the field of mine waste remediation, the problem of acid and metal contaminated mine drainage is of major concern. The problem is related to the instability of sulphides in atmospheric conditions. The sulphides are oxidised, producing both sulphuric acid and dissolving metals from the minerals (Steffen Robertson and Kirsten 1989). Almost all oxidation of sulphur species results in production of protons and thus a decrease in pH (Lindsay 1979). The solubility of the metals is also increased by the decrease in pH. Under these conditions, the factors controlling the leaching process are both redox and pH. Laboratory tests regularly used for investigating the leachability of residues, except for mine waste, are mostly relate leached amounts to the amount of percolated water (Fallman and Hartlen 1993). In these tests, pH and redox in the leachates are mainly dictated by the residue itself under water saturated conditions. Minor quantities of air are dissolved in the leachant and also minor amounts of acid when the leachant is acidified to simulate acid rain. Coarse grained material will rarely be water saturated under field conditions. Changes in the controlling factors in the field situation in comparison to what is obtained in the laboratory may change the entire pattern of leaching from a residue. Therefore there is a need for studying the influence of outdoor atmospheric exposure of different materials and the changes in the values of the factors controlling leaching. At the Swedish Geotechnical Institute (SGI) a project focussing on the factors controlling leaching from residues from smelting and combustion processes was started in 1992. Materials in this study have been chosen based on the following criteria: the materials must have physical properties relevant for utilisation in constructions, they must possess environmental properties that are not obviously environmentally non-acceptable and they must be produced in considerable quantities. Four materials were chosen: blast furnace slag (BF slag), steel slag, sorted municipal solid waste incineration bottom ash (MSWI BA) and wood ash, all of which are regarded as potentially secondary raw materials in constructions. BF slag has been widely used in road constructions in Sweden and sorted MSWI BA is discussed for utilisation purposes. In this paper, the results from laboratory leaching experiments are compared with the composition of the leachate obtained during almost one year run of the lysimeters. The comparisons focus on the differences in the values of the controlling factors pH, redox and flow in the laboratory tests versus the field tests and the consequences of these differences for the leachate composition. 2. MATERIALS AND METHODS 2.1 Materials
The samples for laboratory analyses and tests were taken from the material used in the lysimeters. The samples were dried at 105 "C and divided into laboratory samples before storage. The composition was analysed by using two different digestion methods: lithium borate melt for the main components and nitric acid in Teflon bomb and micro wave oven for certain trace elements. The analyses were made on ICP-AES and with ICP-MS for low detection by SGAB.Table 1 shows the composition of the materials.
41
Table 1 Total composition of the slags and ashes used. The materials were digested in lithium metaborate melt for main components and with nitric acid in Teflon bomb for certain trace elements (I). All elements were analysed on ICP-AES. ICP-MS was used for low concentrations. Results are means for duplicates. M S W BA Wood Ash Substance Unit BF Slag Steel Slag -
Al
%
Ca Fe K Ms
YO YO
Mn Na P S Si Ti Asi Ba Be Cdi CO' Cr CU' Hg' M O
Nb
Ni' Pb' Sn Sr V W Zn' Zr LOI, 550°C
6.48 22.4 0.17 0.49 11.8
2.19 22.1 24.2 0.046 4.52
5.69 8.79 10.8 1.43 1.13
4.86 8.25 2.83 4.03 1.09
0.41 0.43 0.013 1.71 14.6
3.93 0.037 0.46 0.126 5.77
0.136 2.64 0.45 0.856 20.88
0.48 1.22 0.50 0.544 22.2
PPm PPm PPm PPm
1.57 <0.7 406 7.0 <0.40
0.28 5.3 728 <1.2 0.45
0.62 16.0 1660 1.78 5.8
0.29 38.2 1320 2.3 4.50
PPm PPm PPm PPm PPm
0.5 1 34.6 5.7 C0.40 C6.0
5.8 7760 166 <0.41 20.6
19.1 274 3400 C0.39 16.0
11.1 94.5 88.7 <0.39 16.0
PPm PPm PPm PPm PPm
35.2 7.3 C0.40 14.5 324
198 45.0 21.5 17.5 178
13.1 138 737 130 285
13.5 40.0 253 64.7 407
PPm PPm PPm PPm
353
1210 410 244 55.2
58.7 33.9 3080 200 4.3
53.9
% %
YO %
YO YO % %
Yo
<14
1840 168 17
Blast furnace (BF) slag is one of the residues from ore based iron production and is mainly composed of silica and lime. Sulphur is the highest content among the four materials and comes mainly from the coke used in the process. From Table 1, it can be seen that the BF slag
42 contains low concentrations of trace elements. The melted slag from the blast furnace is poured into basins, left to cool, then dug out, sorted and stored before utilisation mainly in road bases. The slag in this study has a particle size of 0-300 mm. The steel slag used in the study is generated in scrap based electric arc krnace steel production. The process results in a slightly reduced slag mainly composed of iron and calcium, and containing 0.7% chromium and 0.1 % vanadium. Among the other trace elements, copper and zinc may be mentioned. The slag is sorted by magnetic separation and consists of the particle sizes 0-300 mm. The bottom ash from municipal waste incineration (MSWI) and wood firing ash both come from Linkoping and are sorted by magnetic separation and screening to a fraction of 2-35 mm. The incineration plant burns 200,000 tonnes MSW annually and has an advanced air pollution control system. The fly ash and APC residues are treated and deposited separately from the bottom ash. In the wood firing plant, both primary wood chips and secondary wood materials are used. Therefore the ash content of trace elements is higher than if only primary wood chips are used. The zinc content is probably originates from galvanised nails in demolition waste. 2.2 Lysimeters The lysimeters were built at the lysimeter field outside the SGI building. The field contains installations to facilitate the collection and sampling of leachates. One lysimeter was built for each residue. The dimensions were dictated by the criteria used for columns, i.e. that the largest particle size must not be larger than 1/10 of the column diameter (Hjelmar 1991b). The largest particle size was 300 mm, resulting in a lysimeter size of 3.0 x 3.0 x 1.2 m3, The walls of the lysimeters are made of plywood and covered with pre-formed HDPE liner. A geotextile with a fixed synthetic draining layer covers the bottom liner of the lysimeter and prevents particles entering the leachate collecting system. The percolate is collected at the centre of the bottom and flows into the draining system by first passing a water seal that prevents the entry of air into the system. The percolate flows are continuously measured by a tipping bucket system, all the equipment being made of noncontaminating materials placed indoors in the basement of a nearby building. Samples are collected proportional to the flow by every second tipping of the bucket. After sampling, leachates are stored under argon gas in a confined system with a minimum of contact with the atmosphere to prevent C 0 2 take-up as well as oxidation. No surface run-off from the materials is expected and the wall of each lysimeter is higher than the top surface of the filled material. In addition to the collection of leachate at the bottom of the lysimeter, cups are installed at two levels in the lysimeter material. The cups are placed at 1/3 and 2/3 of the hight from the bottom, with four cups at each level. The cups are reached by a thin pipe and the water is sucked up. A meteorological station for continuous collection of data on wind, temperature, humidity and precipitation is permanently established on the lysimeter field. The first and second lysimeters containing blast hrnace slag and steel slag were built in December 1992 and the third and fourth lysimeters with bottom ash from MSW incineration and wood firing were built at the beginning ofFebruary 1993. The lysimeters contain 12 tonnes BF slag, 21 tonnes steel slag, 14 tonnes MSWI BA and 10 tonnes wood ash respectively. Leachates are collected as closely as possible to predefined liquid to solid ratios (L/S-ratios). The L/S-ratios increase logarithmically: L/S=O.OOl, 0.003, 0.01, 0.03, 0.1, 0.3 and 1.0. The L/S-ratios are based on the actual amount of material in the lysimeter and collected amounts of water will thus vary between the lysimeters.
43 2.3 Laboratory tests The materials were tested in the laboratory by availability and column test. These test methods are regularly used when characterising waste at the SGI laboratory. In addition, oxidised availability tests and pH-static batch tests were conducted. By using these tests, leaching under different controlled conditions can be measured. The availability test is designed to give the total amount of element concentration that will be leached without the material previously being weathered. In this test, factors limiting leaching such as buffer capacity, solubility constraints and large grain size are eliminated. The elements bound in silicate minerals are not expected to be leached in the test (Groot et al 1989). The availability test used follows with some modifications the procedure proposed in the Dutch draft standard procedure NVN 7341. Finely ground material, 95% 4125 pm, is leached twice in serial batch at L/S=lOO. pH is held constantly at pH 7 for 3 hours in the first step and at pH 4 for 4 hours in the second step by the addition of nitric acid. The two leachates are combined before analysis. The oxidised availability test is used to examine the redox dependency of the leaching process for certain substances and minerals and to estimate the potential for leaching under oxidised conditions. The test is conducted in the same manner as the ordinary availability test, except that the redox potential is kept at the level of fully oxidised conditions of demineralised water for the actual pH by the addition of peroxide. The column tests were conducted on materials crushed to 420 mm to match the column size. The columns were made of transparent polyacrylate with a diameter of 0.19 m and a height of 1.0 m. The experiments were conducted under up-flow water saturated conditions at a constant rate of L/S=O.Vday and run until L/S=4 was reached. As leachate, a synthetic rain water was used, consisting of demineralised water acidified with nitric acid to initial pH=4. The leachates were filtered directly by an on-line filter system with 0.45 pm filters, before storage under nitrogen atmosphere. Samples were taken out at accumulative L/S=O. 1, 0.3, 0.7, 1.0, 2.0 and 4.0. pH-static tests were carried out to characterise the pH-dependent leaching behaviour for different substances under the redox conditions dictated by the residue. L/S=5 was used to perform a leaching test where chemical equilibrium may be obtained and the amount of acid added is still realistic. This L/S-ratio was also used by DiPietro et al (1989) and van der Sloot (1992b). The materials were crushed to <4 mm. Leaching tests were carried out for 24 hours and at pH4, 6, 8, 10 and 12.
2.4 Leachate analyses The leachates were analysed on the basis of general parameters such as pH, redox, electric conductivity and total dissolved solids, metals and salts. The metals were analysed on ICPA E S and ICP-MS for low concentrations. The leachates from the BF slag could not be analysed with the normal procedure on the ICP because gaseous hydrogen sulphide was driven off at the acidification of the samples, giving an excessive detection of sulphur. These samples had to be analysed unacidified. 3. RESULTS
The residues have been exposed to natural weather conditions in the lysimeters since the end of 1992 and beginning of 1993 respectively. The precipitation and percolation flows, as presented
44
in Figure 1, show the dynamics in the lysimeters where evaporation and storage play an important role in reducing the percolate in comparison to the precipitation. The response in percolation in January and February for BF slag and steel slag comes from occasions when the accumulated snow melted in a short time. 1993 was warmer and drier than normal during the first half of the year and wetter and colder than normal during the second half of the year (Vhder och Vatten 1993). As a consequence, very little percolate was produced during spring and early summer 1993, see Figure 1. The steel slag gave the earliest and largest response to rainfall. This is a compact slag with little porosity. BF slag and MSWI BA gave a slower response due to greater storage and evaporation capacity. BF slag is more porous than steel slag. Wood ash is finer than the other residues and has a higher content of organic material. These factors play an important role since the wood ash has only produced small amounts of percolate during the year and the ash does not seem to have become saturated until November after the evaporation has decreased. The lysimeters have been more or less unsaturated during long periods in 1993, especially during the spring, and the residues have been more or less freely exposed to atmospheric conditions.
120
100 Precipitation 80
BF slag
60
--
40
-.
IIIU Steel slag
v
MSWI BA Wood ash
-
R .
Jan
Feb
Mar
April May
June
July
Aug
Sep
Oct
1993
Figure 1 . Measured monthly percolation in the lysimeters during 1993 compared to precipitation measured at the Malmslatt, Linkoping weather station (Vader och Vatten 1993). The percolate samples from the lysimeters had been accumulated to target L/S-ratios, as mentioned previously. These US-ratios were reached at individual times in the different lysimeters. The development of pH in the percolates from the lysimeters differs more or less strongly from the pH obtained in the laboratory column tests, as can be seen in Figure 2. The pH in the leachate from wood ash and steel slag lysimeters is not strongly alkaline as in the column tests, but is about pH 9. This is probably a result of the exposure of the material under unsaturated conditions to carbon dioxide. The wood ash samples were collected in October
45
and November.The first and second steel slag samples were collected during the snow-melt in January and February and the third one not until the beginning of August. The first BF slag percolate was collected after a snow-melt in February and showed a somewhat greater pH decrease than the steel slag, but in the next two samples, after 6 and 7.5 months respectively, the pH had decreased to around pH 4. The decrease in pH is similar to what has been seen in lysimeters with mine waste and old copper slag (Fallman 1991) and is probably the result of oxidation of the reduced sulphides and other sulphur species in the BF slag. Only the leachate production in the MSWI BA lysimeter resulted in an overlap between the lysimeter and the column test results, which in this case show good agreement in pH. The sampled percolated leachates from the lysimeters are stored under argon gas and are thus exposed neither to oxygen nor carbon dioxide. The only contact the water has with the atmosphere is from dripping from the pipe until tipping of the bucket, which is probably of minor importance. The measured pH values may therefore be regarded as representative for the constructed field situation. The measurements have shown that pH in the leachates from laboratory column experiments and the field-like situations may differ considerably for alkaline materials and reduced sulphur containing materials.
-
BF slag, column A
-*-
BF slag, column B BF slag, lysimetcr
Steel slag, column
--c
Stecl slag, lysiineter
--ct.
MSWI BA, column MSWI BA, lysimeter Wood ash, column
------i
0,001
0,Ol
031
1
10
Wood ash, lysimeter
us Figure 2. pH in leachates from the lysimeters and the laboratory column tests as a function of the L/S-ratio (leachate/solid material). Redox measurements have been made in all tests in the laboratory and on the leachates from the lysimeters. A material that shows reduced conditions in laboratory leachate will eventually be oxidised when exposed to the atmosphere, as in the field situation. The presentation of redox potential related to the pH provides information on the redox situation in a material and on the possible behaviour of redox sensitive minerals and substances. All materials showed more or less reduced conditions in the leachates from the column and availability tests, as can
46
be seen in Figures 3 and 4. The most reduced conditions were obtained in the leachates from BF slag. Even the lysimeter leachate shows reduced conditions in spite of the full exposure to air and also through the change in pH. Leachate sampled from the cups in this lysimeter shows somewhat less reduced conditions. Even after one year, the lysimeter still smells of sulphur. The leachate from the MSWI BA lysimeter also shows slightly reduced redox potential, which is comparable to the level obtained in the columns. In the other two lysimeters, more or less oxidised conditions are obtained in the later lysimeter leachates. In the steel slag, measurements in the cups show reduced conditions.
Redox in leachates, BF slag -X-
Demin water, ox
1
Availability Availability, ox
O
-
Column A
-A-
Column B
pH-stat 2 -400
4
6
8
B
1
Lysimeter, cups oct
+
Lysimter, cups dec
PH
Redox in leachates, Steel Slag 800 -x-
-
600
s
Demin. water, ox Availability Column
400
+- pH-stat
i3
200
Lysimeter +
0
Lysimeter, cups oct
I
0
2
4
6
8
1 0 1 2 1 4
Lysimeter, cups dec
PH
Figure 3 . Redox measurements as a function of pH in the laboratory tests and lysimeter leachates for BF slag and steel slag.
47 The use of oxidised availability renders a possibility to estimate the change in leachability due to oxidation of the material. In the ordinary availability test, oxidation may occur due to the amount of dissolved oxygen in the water at high L/S-ratio and the open test system. The comparison between the tests thus shows the difference between a somewhat oxidised test and a full oxidation test. Triplicates have been made for the ordinary availability test and duplicates for the oxidised test. Reproducibility was good for the ordinary test but less satisfactory for the oxidised test. In the oxidised test with wood ash, organic material may become oxidised. The consequences of this for leachability complexing organics have not been investigated.
Redox in Ieachates. MSWI Bottom Ash
8oo
1
Demin. water, ox
-X-
Availability
-
+
2oo
PH-stat
1
Lysimeter
O J
0
Column
1
2
4
6
8
+
Lysimeter, cups oct
1 0 1 2 1 4
PH Redox in leachates. Wood Ash 8oo
1
Dcmin. watcr, ox
-x-
600
5
Availability
400
--
200
+
-----t
pH-stat
i5
Lysimetcr
+
01 0
Column
Lysinieter, cups oct
I
2
4
6
8
1 0 1 2 1 4
PH Figure 4. Redox measurements as a hnction of pH in the laboratory tests and the lysimeter leachates for MSWI BA and wood ash.
48 Comparisons between the results from the two tests are shown in Figures 5 and 6 . The main components were generally little influenced by the oxidation, except for iron, which was clearly reduced in the leachates by the oxidation in all materials except for wood ash. In the bottom ash leaching, also silica is reduced. Among the trace elements, copper, lead and cadmium, which are known to form sulphides with low solubility, gave the highest response to the oxidising conditions. Zinc would be expected to behave in the same manner but did not
Availability, Blast Furnace Slag
100000
2 8 2
v
B
4
10000 1000 100
h
10
~
1 0,1
0,001 0*01
Innl-luull Ca Fe K Mg Na
I
S
Si
II Zn
Al As Ba Cd Co Cr Cu Ni Pb
V
0Availability, oxidised
Availability
Availability, Steel slag
1000000
-
38 Y
3
<8
100000 10000 1000 100 10 1
0,1 0,Ol Ca Fe
K Mg Na
I
S
Si
Availability
A1 As Ba Cd Co Cr Cu Ni Pb
-
1
V
Zn
0Availability, oxidised
Figure 5. Availability for leaching of main constituents and trace elements in BF slag and steel slag under redox potential dictated by the material and oxidised conditions respectively. Values below detection limit are marked with <,
49
respond differently in the two tests. This shows that the sulphides have probably been oxidised in the ordinary availability test. Chromium, arsenic and vanadium, which all form complexes with different solubility depending on the oxidation state, show response to the oxidation. The leachability is increased for chromium in all materials, for vanadium in all but one and for arsenic in the ashes where the results were above the detection limit.
Availability,MSWI bottom ash 1000000 100000
3
c!$
10000 1000
v
3
100 10
4
1 0,1
0,Ol
Ca Fe K Mg Na
S
Si
A1 As Ba Cd Co Cr Cu Ni Pb
V
Zn
V
Zn
0Availability, oxidised
Availability
Availability, Wood ash 1000000
P
p
v
3
100000 10000 1000 100
cl
a
4
10
1 0.1 0.01
Ca Fe
K Mg Na
I
S
Si
Availability
Al
As Ba Cd Co Cr Cu Ni Pt
0Availability, oxidised
1
Figure 6. Availability for leaching of main constituents and trace elements in MSWI bottom ash and wood ash under redox potential dictated by the material and oxidised conditions. Values below detection limit are marked with c.
50
The influence of pH changes on the leachability of the materials was studied in the pH static tests. These were conducted in the range of pH 4 to 12 in steps of two units. The results from the pH static tests are applied to the pH levels obtained in the column tests and the lysimeter leachates, or as closely as possible. The leachabilities at the chosen pH levels are compared, see Table 2. This shows the likely pH effect on the leachability of the materials in the lysimeters, in comparison with the column tests. The decrease in pH leads to a higher leachability for most substances, with some exceptions for substances that were not affected or reduced in leachability. In the BF slag, the highest increase in leached amounts was obtained, as well as the largest pH change in the lysimeter. Table 2 Change in leachability with pH as obtained in the pH-static tests. Symbols indicate change in leached amounts in an interval of pH change representing the difference between the pH in the leachate from column test and that obtained in the lysimeter leachate: +++ = >lOO-fold increase, ++ = 10-100-fold increase, + = 2-10-fold increase, 0 = rt 2-fold, - = 2-10-fold decrease, - - = >lO-fold decrease, < = both values are below detection limit. BF Slag Steel Slag MSWI BA Wood Ash Substance pH, column 12,l-11,4 12,2-11,s 7,3-8,0 10,s-11,2 8,s-9,3 7,6-8,0 9,s-8,7 pH, lysimeter 7,7-4,1 Discussed pH-change 12 --> 4 12 --> 8 No difference 10 --> 8
Al Ca Fe K Mg Mn S Si As Ba
Cd co Cr cu Ni
++ +
__
m
0
I,
0
0
I,
0
__
+++ +++
')
++ 0
+ + ++ ++ <
++
++
++ ++ _
Not applicable
0
I,
+
I,
I, 0,
++ ++ +
,,
0
I,
+
,I
I!
,I
0
8,
<
II
-
I,
Pb V Zn ++ _ I ) Unsuccesshl analyses of sulphur at pH=lO and 12.
++ ++ 0
0
+ + + + 0 0
+ ++
51
The differences in the leachate composition between column tests and lysimeters can not be directly compared because by November 1993, none of the lysimeters had produced so much percolate that the L/S-ratio overlapped the first samples from the column tests. Instead a comparison is made between the extrapolation of the curve of accumulated amounts from the lysimeter and the values obtained from the column tests. The extrapolation is made from the two latest L/S data points and with an extension of the L/S-ratio by a factor of approximately 3. The comparisons are presented with the precision of magnitudes in Table 3. Table 3. Comparisons by the magnitude of the accumulated leached amounts from the lysimeter, extrapolated from the latest two data points, and from the column tests. ++ = > I 00 times more leached in the lysimeter, + = 10-100 times more, f = 10 times more to 10 times less, - = 10-100 times less leached in the lysimeter, - - = >lo0 times less, < = values below detection limit. Substance BF slag Steel slag MSWI BA
Al
f
__
-
Ca Fe Mg Mn
f
f
f
<
<
<
++ ++
+
f
-
S
f
f
Si As Ba Cd
'>
f f
<
+ +
f
f
f
<
f
+
*
f
<
f f
f f
f
-
f
++
=k
co Cr cu Ni Pb
*
++
V + Zn + I ) No values obtained for column tests.
f
f
< f
<
+
Wood ash results are not compared because of the small amounts of leachates so far from the lysimeter. 4. DISCUSSION
The exposure of the alkaline and more or less reduced residues, BF slag, steel slag and wood ash, to atmospheric conditions clearly influences the controlling factors of pH and redox potential in the leaching process. This is seen in the comparisons, Figure 2, between the pH in column test leachates and the lysirneter leachate. The lower pH values obtained in the lysimeter leachates are responses to the exposure of the material to air, i.e. oxygen and carbon dioxide.
52
The loose filling of the material and low water retention in the coarse materials in the lysimeters provide good exposure to air. A more water retaining and less air permeable structure can be obtained by the use of compaction techniques and a content with smaller particle sizes. This may reduce the rate of the pH change to a level lower than indicated by these field experiments. Lysimeter experiments conducted on a mixture of steel slag and BF slag covered with a sealing of asphalt and grass (Mulder 1991), showed maintained highly alkaline conditions in the leachates, pH 12-12.5. In that study, the material was not readily exposed to air. The actual air and water saturation conditions for a residue in a utilisation or deposition object are thus important for prediction of the leaching conditions. Non-water saturated conditions and free exposure to the atmosphere may lead to a lowering of the pH values, as found in these lysimeter tests, and leaching conditions different to those predicted in laboratory tests. The carbon dioxide reservoir in the atmosphere reduces the pH in an alkaline solution to pH 8-9 (Stumm and Morgan 1981), which is the same pH as obtained in the lysimeters with steel slag and wood ash. The MSWI BA leachate was approximately neutral in the column test, and the lysimeter leachate held about the same pH. The decrease to pH 4 in the leachate from the BF slag lysimeter is probably a combination of the oxidation of the considerably reduced material containing sulphur and the buffering of carbon dioxide. Although pH is one of the most important factors controlling leachate composition, the change in pH in a leachate may in itself be caused by redox reactions. The composition of the leachate from these materials is influenced by oxidation, as shown in Figures 5 and 6, to increased or decreased leachability. Metals bound to minerals unstable under oxidising conditions were released, such as cadmium, copper, nickel and lead bound in sulphides. Substances such as chromium and vanadium formed more soluble complexes after oxidation, while iron decreased due to precipitation. Barium concentrations were lower in the steel slag lysimeter leachate than in the column leachate, probably due to barium sulphate precipitation after oxidation of sulphur species to sulphate, see Table 3 . This was not seen in the BF slag, where a slight increase, within one order of magnitude, in leached amounts of barium was seen in the lysimeter. Redox conditions in the BF slag lysimeter still had sufficient reducing potential to prevent the reduced sulphur species from being oxidised to sulphate. The combination of the effects of oxidation and pH change may produce a larger or smaller net impact on the leachate composition than the effect of the individual changes. The larger extrapolated leached amounts of copper from BF slag in the lysimeter in comparison with the column test are the result of increased availability due to oxidation and increased solubility due to pH decrease. This increase is of more than two orders of magnitude (concentrations of 5 - 10 mg Cu/l). The application of the results from the pH static tests on the pH differences between column tests and lysimeter tests in Table 2 show that lower pH in the leachate from the lysimeter compared to the column mainly results in increased solubility. In some cases, decreased solubility is obtained. This is seen in the pH static tests on steel slag for the decrease in pH from 12 to 8. In these lysimeter leachates, the increased availability due to oxidation of sulphides and reduced solubility due to pH decrease le to virtually no difference in the leached amounts of copper and to a decrease in leached amounts of lead in comparison with the column results. The magnitude of the influence of pH and redox, see Figures 5 and 6 and Table 2, show that pH influences solubility to a greater extent than oxidation.This comparison is based on the oxidised availability test and the pH static tests. However oxidation makes the metals readily available for pH influence.
53 Leached amounts from the MSWI BA lysimeter and column tests are mainly in good agreement, which is expected due to the small difference in pH and the similar redox appearance. The effects of changes in controlling factors on the leaching of main components is as important as the trace element (im)mobilisation. A change in the leachability of these elements may change the physical properties of the materials. Hitherto, the changes in controlling factors have not significantly changed the matrixes of the materials. However, leaching of calcium from the BF slag is so high that an increase in leaching rate within one magnitude may influence the physical properties of the material. The interactions between pH and redox potential caused cartain problems in the tests applied in the laboratory. The pH static tests were used to investigate pH influence on the leaching process. As can be seen from Figures 3 and 4, the redox potentials were changed more by the pH than the curve for demineralised water. The results cannot be interpreted as solely pH dependent without further evaluation. The results used in the evaluation in Table 2 for BF and steel slags are less influenced by redox changes and are regarded as reliable. The wood ash results showed oxidised conditions at pH 10. The availability and oxidised availability tests show the influence of redox on leachability. However, The ordinary availability test may be somewhat oxidising as the tests are carried out at high L/S-ratio in demineralised water with normally dissolved amounts of air and exposure to the atmosphere. This implies that the true redox influence on leachability may be greater than observed in Figures 5 and 6 . The test procedures for both pH static tests and oxidised availability need to be further developed regarding redox control. Special interest needs to be paid to the oxidation of organic substances in the samples in the oxidised availability test. 5. CONCLUSIONS
The comparisons between the laboratory tests and the one year lysimeter tests have so far shown that significant changes in controlling factors may occur within a time interval that must be considered as short in comparison with 100 years of utilisation or 1,000 years of deposition. The effects of these changes after one year, with some extrapolations, were a more than 100fold difference in the leached amounts for a few substances and a more than 10-fold difference in the leached amounts of a number of substances. The lysimeter tests at SGI, in comparison with laboratory tests, show that it should be possible to use laboratory tests for prediction of field behaviour. However, laboratory tests must be designed to reflect the parameters influencing leaching, such as pH and redox. Field data are important for further studies of the impact of other factors, such as organic complexing substances and flow conditions. It can be clearly concluded that one single batch test would not have been sufficient for testing these materials.
ACKNOWLEDGEMENTS This paper contains results from a research project focusing on the leaching processes of residues with a potential for utilisation. The financial support is provided by the Swedish Waste Research Council.
54 REFERENCES de Groot, G.J., Wijkstra, J., Hoede, D. and van der Sloot, H.A. (1989): Leaching characteristicsof selected elements from coal fly ash as a hnction of the acidity of the contact solution and the liquid/solid ratio. Environmental Aspects of Stabilization and SolidificationofHazardous and Reactive Waste, ASTM STP 1033, P.L. CBt6 and T.M. Gilliam, Eds. ASTM, pp170-183. 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. International Conference on Municipal Waste Combustion, April 11-14, 1989, Hollywood, Florida, session 2B, pp 21-43. Fallman, A-M. (1991): Uppsamlingsanlaggningar- varp och slagg i Falun. (Lysimeters - waste rock and copper slag in Falun.) The Dalalven Commission Mine Waste Project 4:2.SGIVaria 323, SGI, Linkiiping (in Swedish). Fallman, A-M. and Hartlen, J. (1993): Karakteriserhg och klassificering av avfall. Genomglng av olika landers system (Characterisation and classificationof waste. Survey of the systems in different countries.) Rapport 4226, Naturvlrdsverket, Solna (in Swedish). Hjelmar, 0. (1991b): Personal communication. Hjelmar, O., Aagaard Hansen, E., Larsen, F. and Thomassen, H. (1991a): Leaching and soil/groundwater transport of contaminants from coal combustion residues. EFP 1323/86+1323/86-19+1323/89 Water Quality Institute, Hsrsholm. Lindsay, W.L. (1979): Chemical Equilibria in soils. John Wiley and Sons, New York. Mulder, E. (1991): The leaching behaviour of some primary and Secondary Raw materials used in pilot-scale road bases in "Waste Materials in Construction",Proceedings of the international conference on environmentalimplications of construction with waste materials, Maastricht, The Netherlands, 10-14 November, 1991 ( Goumans, J.J.J.M., van der Sloot, H.A., Aalbers Th.G. eds), pp. 255-264. Studies in Environmental Sciences 48, Elsevier, Amsterdam. NVN 7341: Leaching characteristics of building and solid waste materials - Leaching tests determination of availability of inorganic components for leaching. Draft, June 1992. Steffen Robertson and Kirsten (BC) Inc. (1989): Draft acid rock drainage technical guide. Volume 1. Prepared for the British Columbia Acid Mine Drainage Task Force, Vancover, British Columbia. van der Sloot, H., de Groot, G.J., Hoede, D. and Wijkstra, J. (1991a): Mobility of trace elements derived from combustion residues and products containing these residues in soil and groundwater. ECN-C--91-059, Netherlands Energy Research Foundation (ECN), Petten. van der Sloot, H., E.E. van der Hoek, de Groot, G.J. and Comans, R.N.J. (1992b): Classificationof pulverized coal ash: Part 1. Leaching behaviour of coal fly ash. ECN-C-92-059, Netherlands Energy Research Foundation (ECN), Petten. van der Sloot, H., Hoede, D. and Bonouvrie, P. (1991b): Comparison of different regulatory leaching test procedures for waste materials and constructions materials. ECN-C--9 1-082, Netherlands Energy Research Foundation (ECN), Petten. Vader och Vatten (1993), SMHI, Norrkoping.
Environmental Aspects of Construction with Waste Materials J.J.J.M. Goitmans, H A . van der Slmt and l3.G. Aalbers (Editors) 81994 Elsevier Science B.V. AN rights resewed.
55
Validation of leaching model on actual structures G. van der Wegen" and C . van der Plasa aIntron, institute for materials and environmental research B . V . , P.O. Box 5187, 6130 PD Sittard, The Netherlands
Abstract In The Netherlands a standard set of leaching tests has been developed to assess the environmental impact of building materials, especially if waste materials are incorporated. For monolithic building materials a diffusion test is used for this purpose. The results of a research programme on the validation of the diffusion-controlled leaching model for actual cement and asphaltic concrete structures are presented. For fully submerged concrete the concentration profiles measured can be explained by diffusion processes. Exposed to atmospheric conditions there is an outer zone in the concrete which is dominated by moisture movement due to wet/dry cycles while the behaviour in the zone at greater depth is dominated by diffusion processes. 1. INTRODUCTION In The Netherlands legislation with respect to the environmental impact of building materials is in a well-developed stage [1,2]. In order to assess the leaching behaviour of building materials, standarized test methods as well as related acceptance criteria are necessary. Such standarized test methods [3-51 and criteria [2] have been developed in The Netherlands. For pollutants in monolithic materials, these criteria are based on a diffusion controlled leaching mechanism. From tank leaching experiments over a period upto 3 years [6] a diffusion controlled release of pollutants from stabilized waste materials was observed. Moreover, it is well known that penetration of chloride ions from seawater into submerged concrete over a period of decades can be described by a diffusion controlled process [7]. On the other hand exposure of stabilized coal-waste blocks in seawater for a period of 8 years did not show concentration profiles purely determined by diffusion [8]. In order to validate the assumed leaching model for monolithic building materials (diffusion) it was decided to determine concentration profiles of relevant anorganic pollutants in actual structures exposed to leaching for a long period.
2. SELECTION OF SUITABLE ACTUAL STRUCTURES Actual structures exposed to leaching for a long period were selected on the following criteria:
56
1. a cement or asphaltic concrete structure; 2. containing a homogeneously distributed waste material (powder), having a sufficient high pollutant concentration level; 3 . an age of at least 10 years and preferably more than 20 years; 4. an accurate knowledge of the climatic conditions, especially the time of wettness; 5. permission from the owner to drill cores for experimental research. After an extensive search for candidate structures meeting above-mentioned requirements, the following two were selected: A lock for ships in the Dortmund-Ems canal at Alte Rheine, Germany built in 1975. In the concrete walls powdercoal fly ash (EFA-Fuller, RM) has been used. The concrete composition was documented as 290 kg/m3 portland blast furnace slag cement, 50 kg/m3 of above-mentioned fly ash and a water cement ratio of 0.55. The lower parts of the wall have been wetted continuously (except for a few short maintenance works), whereas the upper parts have been exposed to atmospheric conditions. Asphaltic concrete as a protective top layer on the shore of the Veersemeer, a lake in The Netherlands. This asphaltic concrete was applied in 1966 and contains powdercoal fly ash as a filler (about 5% m/m). The water level of the Veersemeer is artificially controlled. Each year on 1. April the water level is raised to 0 m. NAP, whereas on 1. November the water level is lowered to -0.7 m. NAP. No parts of the asphaltic concrete top layer have been wetted continuously. Hence, only samples could be obtained from the 'tidal' zone (i.e. 7 months continuously wet, followed by 5 months of atmospheric conditions).
3. EXPERIMENTAL PROCEDURES 3.1. Sampling From the concrete wall of the lock at Alte Rheine cores 100 mm in diameter and about 150 mm in length have been drilled in December 1992 (age = 17 years). Two cores have been drilled in the atmospheric zone and two cores in the zone below the lowest water level (continuously wet zone). From the asphaltic concrete top layer on the shore of the Veersemeer two cores 100 mm in diameter and about 100 mm in length have been drilled in November 1992. The age of the asphaltic concrete top layer at that time is about 26 years, which corresponds to a total submerged period of about 15 years (7 out of 12 months submerged). The moisture content of the cores have been preserved by adequate sealing in plastic bags until the time of specimen preparation. 3.2. Specimen preparation Starting from the top surface (i.e. external surface of actual structure) of the cores, layers of about 1 mm in thickness each have been grinded to dust. The dust generated has been collected in a cyclone connected with a vacuum cleaner. Pretests have shown that in this way about 95% of the grinded cement concrete and about 85% of the grinded asphaltic concrete is retained by the cyclone. The asphaltic concrete cores have been frozen down to -20°C prior to grounding to obtain brittle behaviour (i.e. better grinding performance). After grinding each layer and collecting the dust generated, the full equipment has been cleaned to prevent 'contamination' of the next layer to be sampled.
57
3.3. Chemical analysis About half of each of the subsamples obtained by the preparation technique described above was used for determination of the cement respectively the bitumen content. The cement content was calculated from the loss-of-ignition and the insoluble residue in hydro-chloric acid according to the British Standard BS 1881:part 124. The bitumen content was estimated from the loss-of-ignition at 950°C. 'Total' concentration of the elements investigated in the grinded subsamples have been measured after dissolution in a strong acid and subsequent analysis by atomic absorption spectroscopy. In addition, for the cement concrete cores obtained below the lowest water level (continuously wet) also the concentration of pollutants extracted by saturated lime water (representative for the pore water in the concrete considered) has been determined. 4. EXPERIMENTAL RESULTS
4.1. Cement concrete cores The 'total' concentration of the pollutants Cu and Mo as a function of distance to concrete surface is shown in Figure 1 for both cores drilled in the atmospheric zone of the concrete wall. For both elements in both cores two clear maxima in concentrations, i.e. at the surface and at a depth of about 20 mm respectively, can be seen. Although the elements As and V were also investigated, no clear concentration profiles were obtained because these concentrations were near or below the detection limit of the analysing technique. 3 m -
-.-core
-bT 2
1
* care 2
25w'
m
Figure 1. 'Total' concentration profiles of the pollutants Cu and Mo in both cement concrete cores drilled in the atmospheric zone. The 'total' concentration of the above-mentioned pollutants for both cores drilled in the zone below the lowest water level showed an almost constant value, i.e. almost no dependence on distance to concrete surface. The concentration of these pollutants extracted by saturated lime water, however, did clearly show a marked decrease near the
58
concrete surface as illustrated in Figure 2 for the elements Ba and Mo. - 0.- 6.E-14(d/r) A
0
15
30
core ci Core CZ
45
A
corec1 Core CZ
60
distance to surtace (mm)--t
0
distance to surface (mm)-+
Figure 2. Concentration profiles (extracted by saturated lime water) of the pollutants Ba and Mo in both cement concrete cores drilled in the zone below the lowest water level. The solid line is the calculated profile based on Fick's second law of diffusion.
4.2. Asphaltic concrete cores The 'total' concentration of Cu and Mo relative to bitumen content as a function of the distance to the top surface of the asphaltic concrete is shown in Figure 3 for both cores drilled. From this figure one can observe a decreased concentration level for the outer 5 mm for both elements and both cores. No such concentration profile was observed for Na . 3.m
.
distance to surtace
(mm)--.
dlatance to surtace
(mm) --.
Figure 3. 'Total' concentration profiles of the pollutants Cu and Mo in both asphaltic concrete cores.
59
Comparing Figure 3 with Figures 1 and 2 it shows that the scatter in results is much higher for the asphaltic concrete cores than for the cement concrete cores. This is probably due to the greater inaccuracy of the determination of the bitumen content compared to the cement content.
5. DISCUSSION 5.1. Cement concrete cores Atmospheric zone The concentration profiles of the elements Cu and Mo in both cores drilled in the atmospheric zone of the concrete wall (see Figure 1) show a constant concentration level at greater distances from the surface (for Cu: > 60 mm; for Mo: > 30 mm). This concentration level is most likely the original content of these pollutants in the concrete. Hence, in the atmospheric zone no leaching out of Cu and Mo has occurred. This can be explained by the absence of physical contact of the concrete surface with external water (except for rain and splash water which is of minor importance). On the contrary Cu and Mo are enriched in the outer zone of about 60 mm and 30 mm respectively. The concentration profiles for both elements and cores are very typical: two maxima, one near the concrete surface and one at a depth of about 20 mm from the concrete surface. The concentration profile at a depth of 20 mm and larger can be explained by a diffusion controlled process as is illustrated in Figure 4. In this figure the concentration profiles for two ’surface’ concentrations (i.e. both peak values at a depth of 20 mm in Figure 1) of Cu have been calculated by Fick’s second law of diffusion [9]. In these calculations the age of the concrete (= 17 years), the original Cu content (= 200 ppdbinder) and an estimate of the effective diffusion coefficient of Cu = 1.1O-l’ m2/s [ 10,111 has been substituted. The calculated profiles compare well with the corresponding part of the measured profiles.
--CO=1800 mglkg
0‘ 20
40
0
dlstance to surface (mm) --t
Figure 4. Concentration profiles for Cu according to Fick’s second law of diffusion. Effective diffusion coefficient = 1.10-’3m2/s; diffusion time = 17 years.
60
According to Reference 12 the depth in concrete over which wet/dry conditions of the atmosphere are noticable, is about 20 mm irrespective of type of cement or water cement ratio of the concrete. This compares very well with the first part of the concentration profiles shown in Figure 1. Under dry atmospheric conditions the concrete will dry out, i.e. the water in the concrete pores will evaporate leaving non-volatile components behind. This process starts at the concete surface and proceeds by moving the moisture front to increasing depth in the concrete. Due to the evaporation of the water in the concrete pores the concentration of the (non-volatile) pollutants in the remaining pore water will increase. This explains the building up of a concentration maximum at a depth of about 20 mm. At greater depth no moisture variations occur and hence diffusion processes are the main transport mechanism. Subsequently by splash water as well as by capillary suction of pore water from the lower submerged parts of the structure, canal water, containing amongst others Cu (concentration = 4 pg/l), is absorbed in this zone causing rewetting and enrichment with pollutants like Cu. This ’pumping’ mechanism due to wetldry cycles explains the enrichment of these pollutants as well as the concentration profile observed in the outer layer of the concrete structure. Submerged zone The ’total’ concentration profiles of As, Cu, Mo and V did not show any significant leaching or enrichment. In addition to the fact that the concentrations of As and V were near the detection limit of the analysing technique, this is probably due to the fact that under these conditions most of these pollutants are chemically bounded in the cement matrix. Only a small fraction of the total concentration is free available in the pore water for diffusion processes [ 111. Therefore, additional measurements have been performed in which concentration profiles have been determined based on extraction with saturated lime water (i.e. simulation of pore water). The results given in Figure 2 clearly show leaching of Ba and Mo from the concrete into the canal water. In this figure the solid lines represent the concentration profiles calculated by Fick’s second law of diffusion, using an effective diffusion coefficient of 5*10-14m2/s and 3.1014 m2/s for Ba and Mo respectively. These values obtained by curve fitting compare well with literature data [10,11], taking into account the effect of redox conditions present in portland blast furnace slag cement [13].
5.2. Asphaltic concrete cores The concentration profiles of the elements Cu and Mo in both asphaltic concrete cores drilled in the shore protection layer of the Veersemeer show within the scatter a more or less constant concentration level at depths greater than about 5 mm from the surface. This concentration level is most likely the original content of these pollutants in the asphaltic concrete. The decreased concentrations in the zone between the surface and 5 mm depth is due to leaching of these pollutants into the water of the lake. The extend of this leached zone corresponds very well with that of the concentration profile calculated by Fick’s second law of diffusion, using an effective diffusion coefficient of 1.10-” m2/s for both elements in asphaltic concrete [ l l ] , see Figure 5. The apparent increase in concentration of both elements near the surface is probably due to the evaporation of polluted water in the rough surface texture (due to erosion) of the material.
61
-
dlaancs lo aufface (mm)
Figure 5. Relative concentration profiles for Cu and Mo according to Fick's second law of diffusion. Effective diffusion coefficient = 1*10-'*m2/s; effective diffusion time = 15 years.
6. CONCLUSIONS The leaching of cement concrete which is continuously submerged in water is diffusion controlled, at least for the pollutants investigated. The leaching of cement concrete surfaces which are exposed to atmospheric conditions is determined by two mechanisms: an outer zone of about 20 mm (for climatic conditions as in Western Europe) is dominated by moisture movements due to wet/dry cycles, while the zone at greater depth is dominated by diffusion processes. The measured leaching depth of the elements Cu and Mo in asphaltic concrete can be explained by Fick's second law of diffusion using literature values for the effective diffusion coefficient. For Na no consistent leaching behaviour in the asphaltic concrete was observed. 7. ACKNOWLEDGEMENT
The authors are grateful for the financial support from NOVEM and the Dutch Ministry of Public Health, Physical Planning and Environment, which made this research programme possible to perform.
8. REFERENCES 1 'Pre-decree building materials soil and surface water protection', Staatscourant No. 121, June 1991, The Netherlands (in Dutch).
62 2 Th.G. Aalbers, et al., 'Environmental quality of primary and secundary building materials in relation to reuse and soil protection'; RIVM-report No. 771402005, June 1992, Bilthoven, The Netherlands (in Dutch). 3 Pre-NEN 7341, 'Leaching characteristics of building and solid waste materials Leaching tests - Determination of the availability of inorganic components for leaching', October 1992, Delft, The Netherlands (in Dutch). 4 Pre-NEN 7343, 'Leaching characteristics of building and solid waste materials Leaching tests - Determination of the leaching of inorganic components from granular building and waste materials', October 1992, Delft, The Netherlands (in Dutch). 5 Pre-NEN 7345, 'Leaching characteristics of building and solid waste materials Leaching tests - Determination of the leaching behaviour of inorganic components from building materials, monolitic waste and stabilized waste materials', October 1992, Delft, The Netherlands (in Dutch). 6 H.A. van der Sloot, 'Leaching behaviour of waste and stabilized waste materials; characterization for environmental assessment purposes', ECN-report No. 89-185, December 1989, Petten, The Netherlands. 7 CUR-report No. 100, 'Durability of marine structures', April 1981, Gouda, The Netherlands (in Dutch). 8 D.E. Hockley and H.A. van der Sloot, Environ. Sci. Technol., 25, No. 8 (1991) 1408-1414. 9 J. Crank, 'The mathematics of diffusion', Oxford, U.K., 1990. 10 H.A. van der Sloot, et al., 'Environmental aspects of stabilization and solidification of hazardous and radioactive wastes', ASTM STP 1033 (1989), p. 125-149. 11 G.J. de Groot, et al., 'Characterization of the leaching behaviour of products', Mammouth-report No. 9, ECN-report No. C-90-007, March 1990, Petten, The Netherlands (in Dutch). 12 CUR-report No. 90-3, 'Carbonation, corrosion and moisture', June 1990, Gouda, The Netherlands (in Dutch). 13 H.A. van der Sloot, et al., 'Influence of redox conditions on the leaching behaviour of waste materials', ECN-report No. C-93-037, June 1993, Petten, The Netherlands (in Dutch).
Environmental Aspects of Construction with Waste Materials J.J.M. Goumans, H A . V M &r SImt and Th.G. Aalbers (Editors) el994 Elsevier Science B. K AN rights reserved.
63
INTERCOMPARISON OF LEACHING TESTS FOR STABILIZED WASTE H.A. van der Sloot” , G.J.L. van der Wegenb , D. Hoedea, G.J. de Groot’. a
Netherlands Energy Research Foundation, P.O. Box 1, 1755 ZG Petten, The Netherlands INTRON B.V., P . 0 Box 5187, 6130 PD Sittard, The Netherlands.
Abstract The emphasis on treatment of waste by solidificatiodstabilization has led to the need for leaching tests to assess the environmetal benefits of such treatment processes. In this paper the intercomparison of leaching tests for stabilized waste. In this study Municipal Solid Waste Incinerator fly ash was used. The sample preparation, the testing for homogeneity of sample batches and the verification of the repeatability of leaching in one laboratory using the tank leaching test used as common method to be carried out by all participants is described. The influence of prolonged hardening of the cement-based stabilization product is discussed as well as influences of forced air and carbondioxide exposure on the release. An explanation of the observations and the implications for testing are given.The repeatability and reproducibility standard deviation of analysing a Standard Leachate for Na proved to be resp. 2.5 and 5.3 % (mean value: 56.85 mg/l). The repeatability and reproducibility standard deviation of the effective diffusion coefficient of Na obtained form the tank leaching test (expressed as pD, = log D,) proved to be resp. 0.071 and 0.095 (in pD,units) at a mean value of 11.46 (N=l6). 1. INTRODUCTION
The potential environmental hazards caused by waste materials varies strongly between wastes from different sources. Although elimination and minimization of waste streams have the highest priority, it is obvious that significant waste streams remain. These have to be dealt with in an environmentally acceptable manner. Some bulk wastes and treated wastes can be applied beneficially in construction [I]. Several waste streams will require treatment, such as stabilization, before disposal to minimize adverse environmental effects. At present, proper methods to address potential environmental effects from monolithic waste forms are not implemented in regulation. Therefore, proper performance criteria for the evaluation of the effectiveness of the immobilization technologies are lacking. Current regulatory test procedures based on single extraction of crushed material at a fixed liquid to solid ratio are inadequate to assess environmental impact from these type of solid specimen [2]. The mechanism controlling release from monolithic specimen is neglected when crushed materials are used. In addition, for quality control purposes a need for reference materials exists in this field. This paper describes pre-normative work in the form of an intercomparison of testmethods for stabilized waste materials, which focusses on the factors controlling release
64 from monolithic materials. In view of the increased need for treatment of waste, the methodologies evaluated in this intercomparison will be highly relevant for the Community Directive on Landfill of Waste Materials and future regulations in the field of waste minimization, treatment and utilization. The work consists of - preparation of a sufficiently large number of representative samples of stabilized waste (cement-based), - testing of homogeneity and repeatability of leaching within one laboratory, - intercomparison of test results from all participating countries for one common test (tank leaching test) in comparison with other tests currently applied at the national level for these type of materials, - data interpretation and statistical evaluation of test results, - dissimination of findings to respective national bodies and to CEN. More than 25 laboratories from EEC and EFTA countries participate in this intercomparison. 2. EXPERIMENTAL 2.1 Preparation of tests specimen
Test cubes of 4 x 4 ~ 4cm composed of very rapid hardening Portland cement, municipal solid waste incinerator (MSWI) fly ash, natural sand and water were produced to meet requirements with respect to homogeneity, sufficient durability to avoid loss of cohesion during the leaching test and a low permeability. To ensure specimen homogeneity the fly ash was sieved to remove material larger than 0.5 mm. The fraction < 0.5 mm was homogenized using a spinning riffler and subsequently divided in 25 subsamples of 2 kg each. The intended composition needed modification due to an unacceptable degree of swelling and retardation of setting of the mortar. The following composition was used: Portland cement class C (540 kg/m3), MSWI fly ash (210 kg/m3), natural sand (0.5-1 mm ; 560 kg/m3), silver sand (125-500 vm; 560 kglm3) and water (300 kglm3). After two days of curing in the molds at 20 "C and > 95 % relative humidity (R.H.) the specimen have been stored in airtight plastic bags at 20 "C and > 95 % R.H. A total number of 140 batches of 12 cubes were prepared. 2.2 Material properties The physical homogeneity of the samples was checked by measuring the apparent density and the vacuum porosity. From seven out of 14 production dates one batch was selected at random, of which 4 cubes were tested following RILEM CPC-I 1.3 [3]. Based on an analysis of variance one batch proved slightly different for both density and vacuum porosity. The average apparent density proved to be 1968 kg/m3 with a standard deviation of 7.9 kg/m3. The one batch deviates less than 0.7 % in density from the others. The average vacuum porosity amounts to 22.2 % VJV with a standard deviation of 0.2 % VJV. The deviation of the one batch is less than 2.7 % from the overall average. This leads to the conclusion that except for a small difference in one batch the cubes can be considered homogeneous with respect to the physical properties apparent density and vacuum porosity. The compressive strength was measured after 7, 28, 90 and 182 days. The results are: 7 days: 48.1 * 1.7 MPa; 28 days: 54.4 * 3.5 MPa; 90 days: 60.7 * 0.9 h4Pa and 182 days: 64.0 1.6 MPa. This level of compressive strength indicates that no problems with the durability are to
65 be expected. The permeability of the specimen was tested by a water penetration test, which led to a value of 4.9 * 10 - I 2 m/s. 2.3 Chemical composition To verify the homogeneity of the specimen between cubes and within one cube out of 7 production dates a single cube was selected at random and analysed after size reduction to less than 125 pm and complete homogenization. A subsample was taken for analysis by ICP, Atomic Absorption Spectrometry and Ion Chromatography. One cube was selected at random and cut in equally sized slices, which were size reduced to below 125 pm, homogenized and analysed with the same techniques. The samples were brought in solution by bomb destruction for Cu, Li, Mo, S and Cd using a HNO3 ,HF,HC104 mixture. For Ca, Mg, K, Na, Ba, Sr, Zn, Pb, Co and Si a fusion with lithium metaborate was applied. 2.4 Availability for leaching To be able to define a driving force for release by diffusion the availability for leaching is determined using a method described in NEN 7341 (formerly NVN 2508[4]), which is based on leaching a fine ground sample under pH control at pH =7 using a liquid to solid ratio (L/S) of 50 Vkg and subsequently at pH=4 using again L/S=50. The extracts are combined and analysed using the same methods as for the analysis of solid samples. 2.5 Tank leaching test (common procedure)
The common procedure used in the intercomparison is a tank leaching test similar to the NEN 7345 (formerly NVN 5432 [5]), which in its general principle resembles the ANS16.1 procedure [6]. This method allows distinction of release mechanisms, such as wash-off effects, dissolution and the main process diffusion controlled release. In addition, physical restriction (tortuosity) and chemical retention of individual constituents can be calculated, when the release of an inert constituent (no retention in the matrix) is measured. Often Na, K or CI can be used for that purpose. A short description of the method is given below. The procedures deviates from the NEN procedure as the procedure has been shortend to 16 days to limit the occurrence of possible depletion of mobile constituents from the 4 cm cubes [7]. Procedure: After rinsing the tank or vessel with acid and washing with water the specimen is placed in the tank on a support. The vessel is filled with water using 5 times the volume of the speciment to be tested. The specimen must be submersed completely. The leachate is removed and replaced by fresh leachant after 2 , 8 , 2 4 , 4 8 , 72, 102, 168, and 384 hours (16 days). The proposed renewal time series follows from the formula: tn = n2 . t l using t l = 0.083days (2 hours) and n= I , 2, 3, 5, 6, 7, 9 and 14. The leachate is filtered and after measurement of pH and conductivity acidified pH=2. Part of the sample in kept unacidified for analysis of e.g. sulfate, bromide, chloride. To demonstrate the reproducibility of the leaching procedure the leaching test was carried out in one laboratory on 10 cubes randomly selected from the different batches. To demonstrate the consistency of the shorter procedure (16 days) in comparison with the full procedure (64 days) the test was also carried out on specimen of lOxlOxl0 cm. To be able to identify possible changes in leachability with time due to further curing of the specimen the test was also carried out three month and seven month after the first test.
66 2.6 Calculations From the leachate analysis data the release (ms/m2) in each time interval is calculated. Based on the assumption that this release is entirely based on diffusion an effective diffusion coefficient is calculated from:
D,i = z (Ei)2/ (4. Uavaii.p)2.(dti -
in m2/s
In which De,i is the effective diffusion coefficient for a component calculated for fraction I; E; is the measured release of the component in fraction i in mg/m2 ; Uavail.is the availability for leaching according to NEN 7341 in mg/kg dry matter; p is the density of the product in kg/m3 ;ti is the time of liquid renewal of fraction I in s and ti.l is the liquid renewal time of fraction I1. The effective diffusion coefficient is usually expressed as pD. = - log D,. By averaging the thus obtained effective diffusion coefficients (De in m%) for an interval of several points, in which the first and last cycles may be omitted due to resp. wash-off effects or depletion, an average D, is obtained, which can subsequently be used to calculate release for other geometries and other time intervals. 2.7 Other test methods applied
Apart from the common method other leaching tests have been: Tank leach test as described but in stead of using a closed vessel air was bubbled through the solution causing an almost neutral pH in the leachate. Tank leach test as described but in stead of using a closed vessel carbon dioxide was bubbled through the solution resulting in a pH between 5 and 6 in the leachate. The German standard DIN 38414 S4 [8], the new proposed CEN procedure for waste [9], the Swiss TVA procedure [lo], the US - EPtox and TCLP method [ll], a modified availability test with pH control at pH =12.5, a pH static procedure [2,12] and the AFNOR method for stabilized waste [ 131. 2.8 Standard leachate A standard leachate has been prepared by extracting a 1 : 1 mixture of MSWI fly ash ( metals: Pb, Cu , Cd and Zn) and coal fly ash (oxyanions: Mo, B, V) at liquid to solid ratio (LS) = 10 using pH control at pH = 4.This leachate needs to be diluted 10 times before use as a standard leachate. The aim of the leachate is to provide a common solution to be analysed by all participants, which will allow conclusions on the degree of agreement between analytical methods used. Therefore the concentration level of constituents is chosen such that with the currently applied methods measurement should be sufficiently accurate.
3. RESULTS AND DISCUSSION 3.1 Material homogeneity The data obtained for the chemical composition of 10 randomly sampled cubes and the results obtained on 8 separately analysed slices from one cube to are given in table I. The following elements were analysed: Ba, Br, Ca, Cd, CI, K, Li, Mo, Na, Pb, S, Si, Sr and Zn. In the table the average and standard deviation and relative standard deviation (RSD) for both datasets are given. The data indicate that the relative standard deviation is less than 6% for several elements both in the individual cubes as well as within one cube. The high relative standard deviation for Mo is entirely due to the limited analytical sensitivity for this element. For Sr
67
and Zn in the individual cubes a RSD slightly higher than 6 YOis observed. For Ba and Pb in the slices of one cube a RSD of respectively 7 and 9% is observed. This higher variability is not due to statistically rejectable outliers. Based on an analysis of variance of the two datasets it can be concluded that Na, Pb and Si do not fUllfill the nulhypothesis at 95% confidence. In the case of Si this can be attributed to the very low RSD in the slices from one block. If the same more realistic RSD as for the separate blocks is applied the condition is hllfilled. In case of Na the difference is just significant, but not critical. In case of Pb the difference is significant, which may be attributed to known heterogeneity of MSWI ash samples. The occurrence of minute metallic particles may cause such effects [ 141. The results of the analysis of individual cubes from different batches as well as the analysis of subsamples taken from one cube indicate that the cubes are sufficiently homogeneous to Table I. Verification of homogeneity of test specimen based on chemical composition between cubes and within one cube. 10 Cubes
Br Ba Ca Cd K Li Mo Na Pb S Si Sr Zn
One Cube
Mean
SD
RSD
55.5 244.0 130325 28.3 5544 17.1 7.5 4581 692 6890 27998 1 324 1826
2.43 4.38 58.7 13.58 5.57 257 6676 5.12 132646 29.9 1.23 4.34 5.09 6109 282 0.85 4.97 17.9 2.28 30.37 5.6 188 4.11 4871 5.83 763 40 4.46 7291 307 3.65 284946 10207 6.69 330 22 6.38 1902 117
Mean
SD
RSD
3.81 2.24 18 7.06 5.20 690 1 5.02 1S O 5.26 321 0.92 5.11 1.26 22.52 302 6.20 70 9.13 5.34 389 1.67 4750 16 4.75 112 5.87
F
P
1.18 1.78 1.07 1.49 1 29 1.16 3.25 2 58 2.98 1.60 4.62 1.91 1.09
0.42 0.19 0.45 0.26 0.34 0.40 0.05 0.07 0.05 0.23 0.02 0.18 0.46
Fcrit 3.31 2.30 2.30 2.30 2.30 2.30 3.31 2.30 2.30 2.30 3.3 1 3.31 3.31
warrant the leaching intercomparison studies. From other work the role of total composition of elements has been shown to be of limited importance for the leaching behaviour of constituents and Pb in particular, which would imply that the somewhat larger variability of Pb may not be reflected in the leaching results [ 151. 3.2 Availability In table I1 the data obtained for the availability test on fine ground materials is given. In addition to the availabilities for individual elements in mdkg used as driving force in the calculation of the effective diffusion coefficients for the elements, the acid consumption in the test is reported. The two step extractions required about 5.95 mequiv of acid/g to control the pH at pH=7 and about 3 mequiv of acid/g for pH control at pH=4.
68
Table 11. Availability (mglkg) and acid consumption (mmoVg) Ba 1 102.9 2 94.6 3 111.6
Na 1 3797 2 3828 3 4005
Br
Ca
Cd
CI
K
Mo
Li
50 51
114200 116400 122500
22.9 14.6
8450 9164
5382 4874 5148
2.2 1.6 2.4
11.4 11.6
Pb
S
Si
Sr
Zn
Acid
71 72.7 61.5
6630 6850 7020
5315 4483
270 272
1249 1178 1304
7.1 6.5 6.2
The sum is 113 mequiv of acid. The three series of availability tests carried out in May 1993(1), September 1993 (2) and January 1994 (3) point at a significant reduction in acid consumption, which is tentatively attributed to carbonation. The plastic wrappings, in which the samples are stored, are not sufficiently airtight to prevent uptake of carbondioxide. Relative to the composition the availability for leaching amounts to about 34 % for Ba, 90 % for Br, Ca and CI, 100 YOfor K and S, 75 % for Cd, 30 % for Mo, 62 YOfor Li, 80% for Na, 10 % for Pb, 2 % for Si, 70 % for Sr and 60% for Zn. The reproducibility of the availability test is good for elements that can be analysed with sufficient analytical accuracy at the concentration levels encountered in the availability test. 3.3 Standard leachate The composition of the standard
standard deviations as indicated in table 111. In general the relative standard deviation is close to or within the desired analytical precision of 5 %, except for V and Mo with respectively 13 and 32 % RSD. This is caused by the analytical limitations of the methods used at the concentration levels in the leachate. The
Figure 1. Na in Standard Leachate
h
9
69 67 6 5 . . 63
N
_ . _ - - - _2 _ - - - . * - . :, : * ; . : * ! ;
3 -
61
3.= z
579 ; 5 55 53 ~
51;
' t :
, -
;: ~
8
'
t
- - - - - - - - -
4 9 - " " ' ~ " " ~ " " ' ~ ~
69 3.4 Tank leaching data
TableIII. Average composition and standard deviation of the standard leachate (mg/l).
3.4.I Effective dijjuusion coefjcierit Mean Stdev RSD N As was shown in several studies [17,18,19] the leaching of constituents Ba 0.0476 0.003 6.3 from stabilized waste is not related to 4 Ca 227 7.0 3.1 the chemical composition. So in 4 Cd 0.835 0.042 5.0 addition to establishing the homo10 0.542 0.010 1.8 cu 10 geneity in terms of chemical comF 5.65 0.39 6.8 8 position the reproducibility of leaching 0.0055 0.0018 32.0 Mo 10 of 10 randomly selected samples was 55.8 2.1 3.8 12 Na tested to ensure a usehl inter0.275 0.0136 5.0 Pb comparison. To assess the difference 4 52.8 S 12 1 . 5 2.8 in sample size on the release three 0.0062 0.0008 13.4 V 10 cubes of lOxlOxl0 cm were tested in 28.2 Zn parallel. Since the release is related to 1.2 4.4 12 the surface area, the diffusion coefficients should be the same, which has been tested statistically. the results are Figure 2. pDe versus given in table IV. The agreement relative analytical SD. between the effective diffusion coefficients, expressed as pDe = - log 18 I 1 I D, , between the smaller and larger cubes is generally good with the exception of S, Ca and to a lesser extent CI. The release of Ca from the larger cubes appears to be slightly higher than from the smaller cubes. With an increase in pD, the standard deviation generally goes up as well due to the fact that the concentration levels associated with higher pD, 's are 12 also lower and therefor closer to detection limits. In figure 2 this is 11 illustrated. The elements Na, K , Br and to a lesser extent CI can be 10' ' ' ' ' ' ' ' ' regarded as inert i.e. non-reactive with 0 10 20 30 40 50 60 70 80 90 100 the productmatrix and as such can be Rel. analytical SD (%) used to calculate the tortuosity or physical restrictivity [7,15]. Their DD,'~are verv close. Constituents such as Cd and Pb show high pD, 's (= low mobility) caused by a strong interaction with the productmatrix. The difference between the pD for inert constituents and reactive components can be expressed in a chemical retention value. This value is a knction of pH, redox condition, complexing constituents in the pore solution and sorption reactions on the walls of the pores.
i
1;
1
'
1
-
70
Table IV. Effective diffusion coefficients, standard deviations and relative standard deviations for 10 randomly selected 4 x 4 ~ 4cm cubes and 3 cubes of lOxlOxl0 cm. 10 Cubes Mean
CI Ba Ca Cd Br K Li Mo Na Pb S Si Sr Zn
11.69 13.43 13.80 16.09 11.36 11.14 12.22 13.66 11.24 15.51 15.06 14.85 13.10 17.32
3 Cubes
SD 0.04 0.08 0.04 0.06 0.03 0.02 0.04 0.08 0.02 0.13 0.04 0.05 0.05 0.26
RSD 7.98 16.26 9.46 13.64 7.58 4.66 8.64 17.14 5.45 25.33 9.46 11.06 9.87 9.87
Mean 11.57 13.33 13.62 15.78 11.21 11.08 12.17 13.99 11.18 15.21 14.61 14.24 12.93 16.38
SD
RSD
F
0.06 13.76 5.28 0.03 6.90 3.46 0.06 12.73 4.27 0.03 1.15 7.11 0.02 3.50 2.28 0.02 3.50 1.30 0.01 2.64 8.43 0.11 22.88 2.61 0.02 3.94 1.46 0.21 37.82 9.03 0.06 12.21 13.35 0.04 8 13 8.97 0.03 7.54 1.28 0.07 7.54 1.28
P
Fcrit
0.03 0.24 0.05 0.36 0.34 0.51 0.11 0.3 1 0.47 0.01 0.00 0.01 0.33 0.33
3.01 19.38 3.01 3.01 19.38 19.38 19.38 19.38 19.38 3.01 3.01 3.01 3.01 3.01
Of these pH is a very prominent Figure 3. Mean pDoand 95% release controling parameter as will be shown later. The relative confidence interval for Na. standard deviation ranges from 3 to 37 %. In the latter case the measurements were at or close to the detection limits of the analytical methods employed. In judging these numbers it should be remembered that these values are not directly measured, but derived from other measurements adding to the overall uncertainty of the final answer. In addition, the release derived from this leach parameter is proportional to the square root of the effective dihsion coefficient, which implies that the error is not propagated linearly in the final answer. The standard deviation in the release data as derived from the effective diffusion coefficient, the availability and the size of the specimen is discussed below. 3.4.2 Intercomparison of lank leach data.
At the time of writing of this paper only part of the dataprocessing was carried out. Therefor only data for Na are given here. For the pD, of Na the results of 18 participants in the intercomparison have been treated statistically using I S 0 5725[ 161. The repeatability and the
71
reproducibility standard deviation of the pD, for Na amount to resp. 0.071 and 0.095 (pD = 3.47 .10 - I 2 m2/s). In figure 3 the individual data units) with a mean value of 11.46 (D=,N~ obtained by the participants are given with the 95 % confidence interval. 3.4.3 Tortuosity or physical restriction Based on the mobility data for inert constituents (here Na) the tortuosity can be calculated. The results are given in table V. The tortuosity data for the small and the larger cubes are only slightly different. The tortuosity for the samples aged for another 4 month show tortuosities significantly higher than the earlier data. This is attributed to the hrther curing of the product. The samples
Table V. Tortuosity data
Series
Period
T
SD
10cubes
May93
212
18
3 cubes
May93
187
12
Aged
Sept93
379
30
exposed to leaching under a carbondioxide Air Sept 93 100 4 flush fall slightly below the aged data. It is surprising to note that the exposure to air c02 Sept93 328 40 during leaching apparently leads to a reduction of the tortuosity. This effect is not only noted for Na but also for K and Br. The reason for this observation is at present unclear
3.4.4 Release estiniates The final answer of a leaching test on monolithic specimen is a release expressed in mglm’ . The uncertainty in release estimates based on the calculated effective difision coefficients is given in table VI. The reproducibility in terms of release is satifactory for the intercomparison. The cumulative release patterns for Ca, S, Mo and Zn under different experimental conditions - after aging, exposure to air and to carbondioxide during leaching - are given in figure 4. The difference between the early measurements (1 month cure) and measurements after 5 month curing, is an overall decrease in release due to ongoing hydration of cement resulting in a denser pore structure. The release of Mo is not significantly influenced by the different conditions of leaching. In the case of Zn the release is directly related to the pH dependence of leaching. This aspect will be addressed in more detail below. Under the low pH conditions imposed by COz all of the Zn available for leaching can be leached in a relatively short period.
3.4.5 Leachingproperties as afrinction of aging and exposure to air and C02 The question has been raised to what extent the leaching properties of cement-based specimen of the type studied in this work will change appreciably as a result of aging (or rather hrther curing) Another aspect of leaching solid specimen has been the issue of leaching in a closed vessel, exposure to the air or exposure to carbondioxide In the latter two cases carbonation of the specimen occurs, which may alter the leaching behaviour of the specimen To assess this aspect, which is accompanied by a drop in pH of the leachant, pH static experiments were carried out in the pH range from 4 to 12 5 The difference in leachability between the cubes
t
100000 i
A '
100
'
" " " "
'
" " " "
10
1
"
100
200
50'
'
' " " -
'
' " " -
'
"
500
Total
_
Available ~
~
_
~
~
~
10
9
a e 9 0.1
[
0.05'
1
A
A
BI '
'
" " " "
10
0.1'
'
" " " "
100
500
1
'
'
" ' -
'
" " " "
10
100
"
500
Time (hours) Figure 4. Release of Ca, S, Mo and Zn from cement-stabilized MSWI fly ash measured in a closed vessel after l(square) and 5 month (plus) curing, exposed to the air (triangle) and CO after 5 month (dot) curing.
cured for about 4 month and the first series of measurements after 1 month curing in a closed vessel is reflected in a higher pD, for Na, K, CI and Br, which points at a increase in physical restriction (tortuosity) due to ongoing hydration resulting in a denser structure of the cementitious matrix. A final series of measurements after about 9 month curing points at a stabilization of the curing process, because about the same pD, for Na is obtained as in September 1993. The same trend is observed for CI, which also shows a stabilization of the pD, after 4 to 5 month of curing (figure 5 ) . The mobility of Ba and Ca is more significantly
73
reduced, which is most likely related to the behaviour of sulfate, as it behaves quite different in the aged series. In stead of diffusion controlled release sulfate shows a marked surface wash-off effect in the aged samples. This is probably the result of carbonation, which decomposes the monosulfate phase in the cement matrix. The cubes were stored wrapped in plastic, which also may have led to some surface carbonation. In the test with forced exposure to air (air bubbling through the solution) the mobility of Ba and Ca decrease hrther (pD. higher). Zn mobility is also decreased. In the test with forced exposure to carbondioxide (COZ bubbling through the solution) the mobility of Ca is significantly increased and that of Zn is more than 4 orders of magnitude higher. This effect is caused by the leachability of Zn as a function of pH (figure 6) as obtained from a pH static leaching test on finely ground cube samples. This figure confirms that from a leachate pH around 10 to a leachate pH between 5 and 6 about 4 orders of magnitude difference in leachability exists. From other studies incite (Zn(0H)Z ) was found to be the solubility controlling phase[ 121. Here the same mineral phase appears to be solubility controlling. In case of Mo the difference between the test conditions is limited, which is in accordance with the pH static data shown in figure 6. With the forced carbondioxide exposure a slightly higher mobility is noted compared to the data obtained in a closed vessel and with air exposure. The pH stat data agree qualitatively with other studies [20]. The forced exposure to carbondioxide, which is supposed to represent accelerated carbonation and aging, leads to unrealistic release data for z" some metals as sorption reactions are usually too slow to follow the rapid change brought about by excessive CO2 exposure. The pD, values and Na release data obtained by 13 participants are given in figure 5 and compared with the data obtained in the verification study. In figure 5 the pD. values for Na are plotted as a function of time showing the change in tortuosity with time due to the ongoing curing process and the agreement between these data and the results
Table VI. Release (mglm2), standard deviation and relative standard deviation for cubes after 28 days of curing. 10 Cubes ( 4 x 4 ~ 4cm)
Mean
Stdev
RSD ~
Br Ba Ca Cd CI K Li
409 1 I0 79956 1.17 98707 80858 49.60 2.01 51047 6.93 1087 1122 426 14.90
MO
Na Pb S Si Sr Zn
10 2.39 7.64 6.95 3541 4.43 0.09 8.11 3076 3.12 1723 2.13 2.01 4.05 0.16 7.94 1335 2.62 0.93 13.40 47 4.34 61 5.46 20 4.63 3.00 20.12
Figure 5. Change in tortuosity upon curing 11.90 9
11.70
11.50
11.30
*
A
11.10
0
2
4
6
8
Curing time (month)
10
74 500
0.5
.
100
h
d
OD
E
v
10
4 : : A
-8 0.1 % 3 +l
I
I
A
*
t
1
0
e
0.1
e 0.01 A
0.01
0.w1
Mo ~
0.005 3
4
"
5
'
6
'
7
'
8
'
9
'
1011
'
0.m1
1213
3
4
5
6
7
8
9 1 0 1 1 1 2 1 3
PH Figure 6. Leachability of Mo and Zn from crushed cement-stahilized MSWI fly ash as measured in a pH controlled test.
Table VI. Effective diffusion coefficients derived from measurements on different sized cubes, aged products and products exposed to air and COz (pD, = - log D, , D, in m2/s). Blocks( 10) 4x4x4cm May-93
CI Ba
Ca Br K MO Na Pb S Zn
11.69 13.43 13.80 11.36 11.14 13.66 11.24 15.51 15.06 17.32
Blocks(3) 10x10x1Ocm May-93
Aged blocks Sep-93
Air pH=7 Sep-93
c02 pH=5 Sep-93
11.73 14.80 14.54 11.37 11.35 13.48 11.49 15.28 15.03 17.99
11.66 16.60 16.44 10.45 10.99 13.57 10.91 15.23 14.80 18.18
12.19 13.59 13.12 10.99 11.36 13.21 11.43 15.24 14.19 13.16
11.57 13.33 13.62 11.21 11.08 13.99 11.18 15.21 14.61 16.38
obtained by participating laboratories in Europe 3.5 Potential uses of the tank leaching test This type of test is a usehl characterization test to identify the leaching controlling parameters and conditions for a given material or class of materials[7, 15, 17 - 191. Then shorter procedures should be developed as compliance tests to assess basically the same parameters
75
however with less accuracy due to the limited number of data points used to quantify the parameters. Options for developing such shorter procedures should aim at minimizing initial wash-off effects, ensure timely wetting of the entire product, make use of the square-root of time dependence of diffusion controlled release and optimize the liquid to surface area ratio in the test to facilitate chemical analysis. Based on the type of leach parameters derived from this test predictions of release at longer time scales than those corresponding with the actual test duration can be made. Based on the pH static information supplied other leaching conditions in terms of pH can be largely explained and most likely modelled after some firther studies into this issue. In addition, the leach parameters obtained can be used for management purposes to improve product quality and performance[7, 211. 4. CONCLUSIONS The tank leaching test studied in the framework of this intercomparison is a good characterisation method for the leaching behaviour of monolithic materials as it provides knowledge on release controlling parameters and allows prediction of release at longer time scales by the leaching parameters derived form the test results. The homogeneity of the specimen prepared for the intercomparison prove to be sufficient between batches as well as within one specimen. The leaching data obtained in one laboratory on randomly selected specimen prove to be sufficiently reproducible in terms of release for most elements studied. In a few cases the concentrations to be measured are close to the analytical detection limit leading to a higher relative standard deviation for those elements. Based on the mobility data for Na and CI , which reflect the tortuosity of the matrix, it can be concluded that the product has changed in leaching properties in the first few months. It appears to be stabilizing in its properties after about half a year. Upon aging the cement-based product is changing in its leaching properties. This is an aspect that can not be avoided for any type of cement-based product. By measuring the release parameters in one laboratory in three time intervals covering the period of analysis by the other participants the results can be placed in the proper perspective. Leaching under different pH conditions was studied, which results in differences in release, which can be largely explained by the results of pH stat experiments carried out on crushed material. Based on this longer test shorter procedures with a close correlation with this test can be derived, which may largely provide the same basic information. Standardization of such a short compliance method will start in the framework of CEN TC 292. It is recommended that the characterization method is also standardized by CEN TC 292.
Acknowledgenienl This work is carried out in the framework of the EEC Measurements and Testing programme (BCR) 93/94. Participants are: University of Aberdeen, GB (Prof.Dr. F.P. Glasser); POLDENINSAVALOR, FR (Dr. J. Mehu): RIVM, NL ( Drs. G.A. Rood); ENEA ,IT( Drs. S. Balzamo): VKI, DK (Dr. 0. Hjelmar); K K , DE (Dr. J. Vehlow ); WRc, GB (N. Blakey, K.Lewin); SGI, SE (J. Hartlen); VTT, FI (M. Wahlstrom); VITO, BE ( Ir. B. Leathem); Institute Quimic de Sarria, ES (Dr. J. Obiols): Junta de Residus, ES (Mrs H. Sala); INASMET, ES (G. Ortiz); Umwelt Bundesamt, DE (P. Henschel); UNINOVA, PT (Dr. A. Steiger Gargao); ISDS , IT (L. Musmeci); WTC, Canada (J. Stegemann).
76 5. REFERENCES 1. Waste Materials in Construction: Proceedings of the International Conference on Environmental Implications of Construction with Waste Materials,Eds. J.J.J.M. Goumans, H.A. van der Sloot, Th.G. Aalbers, Elsevier Science Publishers, Amsterdam, 1991. 2. H.A. van der Sloot. Leaching behavior of waste and stabilized waste materials; characterization for environmentalassessment purposes. Waste Management and Research, 8, 1990,215-228. 3. RILEM CPC-I 1.3. Methods for the determinationof apparent density and vacuum porosity.1984. 4. NEN 7341 (formerly NVN 2508). Determination of leaching characteristics of inorganic components from granular (wastes) materials. "I, Delft. 1993. 5. NEN 7345 (formerly Draft NVN5432). Determination of the release of inorganic constituents Delft. 1993. from construction materials and stabilized waste products. "I, 6. ANS. 16.1 Measurement of the leachability of solidified low-level radioactive wastes by a short-term test procedure. American Nuclear Society, Illinois 60525 USA 1986. 7. G.J de Groot and H.A van der Sloot. Proc. Sec. Int. Symp. Stabilizatiodsolidification of Hazardous, Radioactive and Mixed wastes. Williamsburg, Virginia, May, 29 to June 1, 1990. 8. DIN 38414 S4: Geman standard procedure for water, wastewater and sediment testing - group S (sludge and sediment); determination of leachability (S4). Institut fiir Normung, Berlin, 1984. 9. CEN TC 292 document: Proposed leaching test for granular solid waste. H.A van der Sloot, 0. Hjelmar, Th.G. Aalbers, M. Wahlstrom and A,-M. Fallman, February, 1993. 10 Bericht zum Entwurf fur eine technische Verordenung iiber Abfalle (TVA), 1988. Departement Federal de I'Interieur. Switzerland. 11 Toxicity Characteristic Leaching Procedure(TCLP). Federal Register Vol 5 1 No 114, Friday, June 13, 1986, 21685-21693 (proposed rules). Federal Register, Vol No 261, March 29, 1990 (final version). EPA Toxicity Test Procedure (EP-tox), Appendix 11, Federal register, Vol45(98), 1980, 33127 - 33128. Govemment Printing Ofice, Washington D.C. 12 R.N.J.Comans, H.A.van der Sloot, P.Bonouvrie. Proc. Municipal Waste Combustion. VIP 32. Air & Waste Management Association Pittsburg, Pennsylvania. 1993. 667 -679. 13 J. Mehu, Y.Perrodin, B. Sarrazin and J. Veron. Reference I . page293 - 300. 14 C.W. Versluijs, I.H. Anthonissen and E.A.Valcntijn. Mammcet '85. Report 738504008. RIVM, June 1990. 15 S.E.Sawell, A.J.Chandler, T.T.Eighniy, J.Hartlen, O.Hjelmar, D.Kosson, H.A. van der Sloot, J.Vehlow. The International Ash Working Group: Treatise on MSW Incinerator Residues. This conference Special Session. 16 IS0 5725. Accuracy (trueness and precision) of measurement methods and results. part 2. International Organization for Standardization, 1990. 17 P.L. C6te. Thesis : Contaminant leaching from cement-based waste forms under acidic conditions. MacMaster University, Hamilton, Canada, 1986. 18 D.S.Kosson, T.T.Kosson, H.A. van der Sloot.,"USEPA Program for Evaluation of Treatment and Utilization of Municipal Waste Combustor Rcsidues", Cooperative agreement CR 8 18178-01O.USEPA/RREL, Cincinnatti, September 1993. 19 M. Hinsenfeld. Reference 1: page 33 1-340. 20 G.J de Groot, H.A van der Sloot and J. Wijkstra. In: ASTM STP 1033, P.L. Cote and T.M. Gilliam, Eds, ASTM, Philadelphia, 1989, pp 170 - 183. 21 H.A van der Sloot, G.L. van der Wegen and E. Vega. Beoordeling van immobilisaten. Een voorstel voor criteria en testmethoden. CUR report 93-6. Civieltechnisch Centrum Uitvcering Research en Regelgeving. Gouda. 1993.
EnvironmentalAspects of Construction with Waste Materials J J J M Goumans, H A . van der SIoot and 7b.G. Aalbers (Editors) el994 Elsevier Science B. V. All rights reserved.
77
Immobilisation Potential of Cementious Materials F.P. Glasser Department of Chemistry, University of Aberdeen, Old Aberdeen, Scotland. Abstract The immobilisation potential of lime and Portland cement matrices is two-fold. They afford physical immobilisation by converting liquids, sludges and particulates to solids. They also afford a chemical immobilisation potential, sorbing and precipitating otherwise soluble species. Both the physical and chemical potentials are difficult to quantify. The chemical immobilisation potential is perhaps easiest to measure and model by laboratory simulations. Progress in this area, of characterizing mechanisms of insolubilisation, is described using the Cr-Mo-U triad as examples. Remaining problems necessary to predict the performance of cement-conditioned wastes are discussed. 1. INTRODUCTlON Lime and Portland cement are relatively inexpensive manufactured products having consistent properties. Cement and Ca(OH)2 may also be mixed with each other and with a variety of other reactive wastes including slags, coal-combustion fly ash etc., to form solid matrices. These have been used successfidly to immobilize a range of waste materials. Conner reviews commercial technologies (1). Physical containment is most important for mobile liquids, dusts, sludges and solutions. Because cements are tolerant of wet wastes and can, within limits, neutralize acids and tolerate salts, wastes can frequently be dispersed throughout a cementitious matrix. The matrix affords both physical resistance to leaching and a definite chemical immobilization potential. Characterizing these potentials requires special methodologies. The porositypermeability relations of cementitious matrices are not well understood; most reports have been restricted to formulations intended for load-bearing constructional applications. Most of the intrinsic porosity of well-made cement matrices is less than Ipm effective radius, so the matrix has only limited permeability. However, extrinsic porosity is often introduced at grain-paste boundaries and the amount of such porosity increases rapidly with increased water content (2). This, in turn, affects permeation properties. The mix water used in formulating cementitious matrices is partially combined in the cementitious solids. Normally an excess of water, beyond that required for complete hydration of the cement, is needed to achieve the necesary plasticity or fluidity for The chemically mixing: also, to achieve economic loadings of wet wastes. uncombined water is held in pores, from whence it can be extracted by pore fluid expression (3). Analyses of the expressed pore fluid provides a unique method of determining how much of the added waste species remains soluble. The pore fluid comprises the most leachable portion, and its analysis provides an immediate quantification of the source term for leaching. The solids of cement have high surface area and provide potential for sorption and a source of reactive species for precipitation and hence exert a strong modifjmg influence on pore water composition.
78 2. PHYSICAL IMMOBILISATION Several methods are available to determine pore structures in hardened cement paste. Mercury intrusion can be used to determine pore sizes in the range 0.005-5pm (approximately), with neutron scattering used for finer pores. However the bulk diffusional and leaching properties are determined by the larger pores, within the range of mercury intrusion, as well as by pore interconnectivity. A special type of porosity arises at the interface between particles and cement paste. Fig. 1 illustrates schematically the different ranges of porosity. Interface porosity characteristically arises when grains are physically unequal in size. The standard for comparison is that of the cement, lime, fly ash etc. particles, typically 5-50pm, with larger particles, especially >50-IOOpm, giving rise to intefacial porosity. The larger pores become partly filled with crystals, notably Ca(OH)2 and ettringite, a hydrated calcium sulfoaluminate. The remaining space is occcupied by a permeating aqueous phase. The picture is representative of the state of hydration achieved within a few weeks or
-
20 % Intruded Volume
Slag Cement blend, mcist cured at 18°C
/d
10
0.1 Pore entry diameter
Fig. 1. Mercury intrusion porosimetry scan of a typical slag-cement blend cured for 30d and 2 years. Pore entry diameter in pm. Slow hydration of slag converts much of the open porosity to closed porosity and the pore size distribution shifts to finer pores.
0.01
Fig. 2. Microstructure of a cement containing particulate material. The main part of the Figure is on a micrometer scale. A large particle occupies the lower lefthand corner. Porous regions, ( I ) , exist in partly hydrated paste, but a process zone of enhanced porosity occurs, (2), in the vicinity of the particle. The inset shows on a nanometer scale the structure of the paste. A large part of the intrinsic nanoporosity, (3), is associated with the gel constituent of the paste.
79 months. The presence of residual cement or blending agent indicates that the system still has the capacity for formation of more hydrate. Some of this additional hydrate goes into blocking pores, although interfacial porosity, shown as region 2 in Fig. I , Fig. 2, showing mercury intrusion data for a typical remains little affected. constructional cement, indicates how the porosity continues to decrease with cure. It should be recalled that mercury intrusion only measures uccrssihle porosity: upon continued hydration, pore blocking contributes to the decrease in accessible porosity. However the actual permeation properties achieved may represent a compromise between achieving good waste loadings and realizing low permeability. The permeation properties of cements intended for waste conditioning have received less study. Table 1 summarizes some of the anticipated similarities and differences between constructional mixes and conditioned wastes. A particular concern is that high waste loadings will degrade the microstructure which, in turn, will adversely affect the permeation properties. The intrinsic permeation properties achieved may thus represent a compromise between achieving good waste loadings and realizing low permeability. The formulations which are used in waste treatment do not generally require high strengths. Nevertheless, most cementitious matrices acceptable for conditioning will gain strength and become susceptible to cracking, which increases the effective surface area available for leaching. Cracking may arise from purely mechanical causes, e.g settlement, shrinkage, but it may also arise as a consequence of inherent dimensional instability, from chemical reactions with waste components or those in the disposal environment. To act as a conduit or channel for leachants, cracks have to be relatively large, greater than 1-2pm. Therefore, if cracking cannot be prevented, it is better to have a dense network of microcracks rather than a lower density of relatively wider cracks. However, at present, little theoretical or practical guidance be given on the prevention or occurrence of cracking and, where cracking does occur, on crack density.
Table 1 Physical Properties of Some Cement Formulations Property
Waste Treatment
Constructional Materials
water:solid ratio
as high as possible to minimise cost
generally as low as practicable
intrinsic porosity
apt to be high
low, except in air-entrained materials
interfacial porosity
variable, depends on waste
always present in normal mortars and concretes
microstructure
influenced by presence of soluble component
relatively constant (see Fig. 1).
permeability
highly variable
'ordinary' materials; 10-8 to 10-10 m / s ; 'special' materials, 10-10 to 10-12 m / s
80
3. CHEMICAL FIXATION 3.1 General principles Cements provide a strongly alkaline internal environment. Ca(OH)2 and calcium silicate hydrogel (shorthand, C-S-H) are available to buffer the pH to about 12.4 at 18°C. In the short term, any alkali in the cement or waste tends to raise the pH above 12.4 because counterions - other than OH- - are relatively insoluble: chloride and nitrate are the main exceptions. Hence the pH of a lime- or cement-conditioned matrix is likely to be within the range 12.4-14 and is well-buffered by the cement solids.
10
-
F
1
:
Pb
Zn
Cd
~~
0.001 -
pH range in alkaline cements I
I
I+
1
Fig. 3 . pH control of the solubilities of a few selected heavy metals. The examples shown are for amphoteric elements. The characteristic internal pH of Ca(OH)2, Portland and blended Portland cements is shown Many metals are relatively insoluble in alkaline aqueous environments. However, there are many examples of amphoteric elements, so-called because they are soluble at low and high pHs. In strongly basic solution they form soluble anionic complexes, e.g. (At(OH)4-: tin and chromium (111) are also examples). Fig. 3, taken after reference (4) is a guide to these relations. The thermodynamics of metal speciation are well known and it might at first sight seem that the chemical conditioning action of cement would not be too effective for amphoteric species. However, these considerations are too simplistic because they fail to include the often specific and very strong interactions between cement components and waste species. These interactions depress solubilities, perhaps by orders of magnitude. However, the conditioning action arises from a number of mechanisms; examples will be given . The cement components vary in composition and crystallinity. Some, like the gel binding phase - a calcium silicate hydrate - are largely non-crystalline but have a very high specific surface, leading to a sorptive potential. The more crystalline hydrates favour crystallochemical substitution and, in the case of layer structures, intercalation. When a waste stream is mixed with water, a complex series of reactions ensue. Assuming for simplicity an initially soluble species, the general sequence is as shown in Fig. 4.
81
I initial Waste Concentration :',
.= ._ n %
Pcn
-
I
initially
';Relatively ,-
Amorphous':, Precipitates.!,, Sorption, e t c j :'
Crys:allization.
Fig. 4. Decline in solubility of waste species in a cement matrix as a hnction of time. Not all species undergo all the reactions shown, but the declie in solubility with time is characteristically encountered The first stage of reaction, usually achieved within the first few minutes of mixing, is the precipitation of relatively insoluble precipitates of oxides, hydrous oxides or precipitates containing other readily-soluble anions hrnished by cement, e.g. sulfate, which precipitates barium. These precipitates are generally amorphous, so are difficult to detect directly. As the time scale of reaction is extended to weeks or These include (i) crystallization of months, a series of slower reactions occur. previously-amorphous precipitates with concomitant reduction in solubility, (ii) reaction between waste species and cement components, leading to fixation by ion exchange, sorption, etc., as a result of which waste components are bound in dilute form into matrix components and finally, (iii) reaction of initially formed precipitates with cement components to yield new phases. For example, a cement component A may react with waste component X giving a compound AmXn or AmXn zH20. Reactions in this latter category are driven by the approach to equilibrium. Since AG is thereby decreased, and AG -RTCnK where K has the dimensions of a solubility product, the overall result is a decrease in solubility. Of course, not all three classes of reaction will necessarily occur for any particular species. These considerations highlight the need for mechanistic studies. These can be pursued at different levels. Table 2 outlines briefly some of these. Single speciessingle cement component studies provide the most satisfactory way of isolating for hrther study the insolubilizing reactions. But the data obtained from simple systems may be too naive: for example, the insolubility of BaSOq has been noted. Cement systems also contain much OH and Ba(OH)2 is relatively soluble. In strongly alkaline environments will B a s 0 4 redissolve? In order to answer questions of this sort, it is almost always necessary to move to stage 2, in which the possiblity of more complex interactions can be investigated. Once this has been done, it may be possible to extrapolate the results into the future. Finally, level 3 studies are necessary to ensure
82
Table 2. Methods of Studying Fixation of Wastes in Cement
Simulate Conditions
Remarks and Notes
Selected single waste component and cement component
Results relatively easy to control, interpret and extrapolate but neglect complex interactions. Not readily accepted into action programmes, e.g. compliance with legislation.
Selected single waste component in "real", chemically complex cements and blends
More difficult to interpret results in a fundamental sense, but likely to include complex interactions and provide acceptable basis for decision making.
'Real,' mixed waste streams with chemically-complex cements and blends
Fundamental mechanisms almost impossible to interpret but results necessary to demonstrate compliance. Difficulty remains of extrapolating present performance into the future.
compliance with standards, although they do not themselves provide a mechanistic basis for understanding the basis of containment or of extrapolating into the future the performance of cemented waste forms. In general, therefore, a combination of approaches is necessary to ensure that a broadly correct perspective towards research and practice is maintained. Experimental techniques have been greatly improved whereby specific immobilization mechanisms can be identified and, if needed, isolated for firther study. Computer-based computational routines will enable fiture performance to be assessed, given site-specific information and an adequate database. 3.2 Case studies: chromium Chromium is widespread in industrial and domestic waste streams. Its toxicity is mainly associated with the upper, Cr(VI), oxidation state. The other common state, Cr(III), is known to be better immobilized in cement than Cr(V1). However, the source of the immobilisation potential is uncertain: Kindness et al. ( 5 ) have reviewed the literature and determined the source of the immobilisation potential. Cr(VI) substitutes in part for sulfate in ettringite, 3CaS04 At203 32H20. Its substitution for sulfate in other phases, e.g. in the AFm type phases, is less. However even in ettringite, where substitution is most favourable, relatively high aqueous CrO 2concentrations are required to sustain significant replacement of SO42- by CrO4 . Thus the potential for chromate immobilisation is poor. However, Cr(V1) is readily reduced to Cr(II1) by contact with metals, e.g. with Fe, or by ferrous salts. Many natural disposal environments are also likely to be reducing in nature; the reducing conditions are generated by the presence of organic matter, and bioactivity which produces CH4, C02, etc. Therefore Cr(VI), if present in waste, can be reduced to Cr(II1) and, in the majority of disposal situations, it is unlikely to reoxidise with the result that the target species for immobilisation is Cr(II1). In simulate experiments, Cr(OH)3 is rapidly precipitated in alkaline solutions. Cr(II1) is amphoteric, so precipitation may be incomplete. Moreover, the initial
4-
83 precipitate is often amorphous, or nearly so, and could be expected to have anomalously high solubility relative to crystalline Cr(OH)3. Experiment shows that the amorphous precipitate crystallizes with a few days at -18°C. But Cr(OH)3 is unstable in the cement environment. It gradually reacts with the aluminate phases, where it substitutes for A t in octahedral sites. The open, layered structures of AFm phases are kinetically most accessible to these exchanges. Kindness, et al. ( 5 ) have reported the synthesis of CaO-Cr203-H20 phases which are structural analogues of the aluminates. The denser hydrogarnet structure, based on 3Ca0 At203 6H20, is kinetically relatively inert, but synthetic experiments disclose it to be a very effective host for Cr(II1). Synthesis of various Cr-containing host phases and measurement of their corresponding Cr solubilities has shown that they are very effective hosts for the immobilisation of chromium. Since the solid solutions require time to form, pore fluid Cr(II1) concentrations tend to decrease with cure duration until an equilibrium level is reached. Theory and experiment are in good agreement that the stable solubility threshold in pore fluid is about 0.2-0.5ppm Cr(II1). Thus the potential for resolubilization, which might be expected from the amphoteric character of Cr(III), is suppressed at normal cement pH's by other factors. Further work does require to be done on the long-term behaviour of cemented Cr containing wastes. In the disposal environment ground water components, including carbonate, chloride and sulfate, may react with the aluminate phases. Comprehensive documentation about the long-term fate of Cr awaits further study of these reactions. 3.3 Case studies; molybdenum Compared to Cr, the chemistry of Mo in cements is less complex. It has only one common oxidation state, Mo(V1). With oxygen, it forms very stable M 0 0 4 ~ groups. The effective ionic size of M 0 4 groups increases in the order SO4 < CrO4 < MoO4, so that substitution of SO4 by Moo4 in calcium aluminosulfates, e.g. in ettringite, is even less favourable than for CrO4. Instead, Mo precipitates as CaMoO4, a phase isostructural with the naturally-occumng mineral powellite. Anhydrous phases of the structurally related zircon-fergusonite-scheelitefamily, to which powellite belongs, are often observed to precipitate directly from aqueous solutions across a broad range of pH's. The high lattice energies of these phases stabilize the anhydrous form; apparently, hydration energies are insufficient to form hydrates. Kindness et al. (6) have synthesized CaMoO4 and determined its solubility. Values are shown in Table 3, and the synthetic models predict quite successfully the Mo content of pore fluids made by spiking "real" cements with Mo.
Table 3. Solubility of Mo in Cement Matrices (water:ordinary Portland Cement 0.4,18"C) Aqueous Concentration after (davs) 5 10 15
20 30
Concentration. porn Mo in mix water
500 200 100 75 65 60
1ooo 500 250 125 80 60
2ooo 1000
400 200 100 60
84 3.4. Case Studies: uranium Stabilisation and immobilisation of uranium-containing
mine tailings and process residues arising from a range of sources, including phosphate processing residues, present problems. Most of the mineral residues are, however, compatible with cementitious formulations. Relatively little data exist on immobilisation mechanisms. Moroni has characterised some of the reaction products (7). Ca(OH)2 reacts readily with U(VI) solutions: the nature of the products obtained depends on the Ca/U ratio of the system. At high U loadings, becquerelite CaO 6UO3 l l H 2 0 , forms. However, such high loadings are unlikely to be encountered in practice and in any event, becquerelite appears not to be compatible with Ca(OH)2 or C-S-H gel. Nevertheless, becquerelite is much less soluble than schoepite, U 0 3 2H20, a normal weathering product of uranium under oxidising conditions. At 5 5 T , the U solubili of a becquerelite precipitate had, after repeated dispersion, decreased to between 1 0 8and 10-6M(.IC-1by 400d. Examination of the system CaO-U03-Si02-H20 is still far from complete. Present indications are that the phases likely to be formed in real cement compositions are uranophane, CaO *2UO3 2Si02 6H20, and a phase structurally related to CaU04. This synthetic phase is apparently a hydrous version of probable formula CaU04x(OH)2x. Formation of these phases, together with sorption of anionic U(V1) species on C-S-H gel, appears likely to limit U solubilities to 10-7 to 10-8hUt. Several other incompletely characterised phases occur in synthetic mixtures, but their occurrence is unlikely to affect the broad picture: at very low U concentrations, <10-8WC, sorption on C-S-H limits U solubility. At higher loadings, the buffering capacity of the cement system furnishes Ca, Si and OH necessary to form solubility-limiting phases 4. GENERAL DISCUSSION It is not generally possible to characterise the immobilisation potential of cement systems in short term experiments: weeks or months are required for reaction of cementitious components with water, as well as for reactions between waste components and cementitious material to reach a state of near completion. Given the complexity of the waste-cement-water system, it is generally not possible to conclude the equilibrium state of the system: indeed, the principal bonding component of cementitious matrices, C-S-H gel, is itself probably thermodynamically metastable. However, a steady state is frequently achieved within a few months or years, and the solubilities of waste components in entrapped water, as determined by pore fluid expression from 'real' cements, gives order-of-magnitude correspondence with solubilities determined by laboratory experiments made on simplified systems. Systems chosen for models must not be oversimplified. Simulate experiments and calculations must at least take into account likely interactions between waste species and cement components. The pore water analysis provides a good reality check between the two approaches. These are illustrated in Fig. 5. Case I corresponds to no interaction between cement and waste components. Portland cement and caseium are an example of this type of behaviour: a solubility limit for Cs is obtained only at very high concentrations. Case I1 is like case I, but the cement components show some sorption characteristically at low concentrations. Many cement phases exhibit sorption - either true chemisorption, or limited ion exchange, or both - which depresses pore water concentrations of soluble waste species. Cs in fly ash or slag cement blends is an example. Case 111 illustrates a situation where a solubility-limiting precipitate forms at low concentrations
85 of the waste species; of the examples described, Mo best fits this case. Case IV illustrates more complex interactions, where low concentrations condition sorption as with U(V1) or Cr(1II). Above some critical concentration a definite plateau occurs conditioned by formation of a solubility-limiting phase. The plateau is shown as horizontal; if, as when Cr(II1) substitutes for At?, solid solution dominates, this portion of the curve may have a definite slope or curvature. The task for the experimentalist is, therefore, to determine the nature and strengths of the interactions and calibrate the response of cement systems to different waste loadings. This requires a determination of the solubility-limiting mechanisms. As has
Ioy cciiiceiilialioii waste s p c i e s
-*
Fig. 5 Selected examples showing solubility patterns of a hypothetical waste species in pore fluid as a knction of waste loading. See text for discussion. been shown, it is not generally practicable to deduce these mechanisms without recourse to experiment. Once the chemistry has been done, the physical aspects of diffusion can be analysed. Exchange of pore fluid with groundwater by a percolation process is likely to condition releases; such exchanges are often site specific and depend crucially on the matrix formulation, surface area and mass as well as the hydrological aspects of the disposal site. Of course, alteration of the cemented waste by the action of aggressive and corrosive groundwater, and the transporting action of other complexing species released from the waste need to be assessed to complete a safety case. 5. CONCLUSIONS
It is generally possible to determine the mechanism whereby waste species become fixed in cement. for Cr(II1) Mo(VI) and U(VI), solubilities across broad ranges of concentration tend to be limited by formation of complex salts involving one or more of the cement components. These substances may require considerable time to
86
develop, but the ultimate reaction products and their solubilities can be predicted with confidence by using a mixture of laboratory studies and examination of 'real' cements loaded with the waste species. One important chemical aspect remains, which is to model the impact of cement deterioration on release. In this way, a knowledge base is being created which enables a source term to be defined leaving the way clear to determining the percolation characteristics of particular wasteforms at specific sites and the resulting impact on leachability. 1.
2.
3. 4. 5. 6. 7.
REFERENCES J.R. Conner, Chemical Fixation and Solidification of Hazardous Wastses, Van Nostrand-Reinhold, New York, 1990. T.C. Powers and T.L. Brownyard, 'Studies of the Physical Properties of Hardened Portland Cement' Bulletin 22, Portland Cement Assoc., Chicago, pp 992, 1948. P. Longuet, L. Burglen and A. Zelwer, Rev. des Materiaux de Construcion et de Travaux Publics. No. 676, 35-41, 1973. J.R. Conner in 'Chemistry and Microstructure of Solidified Waste Forms' R.D. Spence (ed.) Lewis Publishers, 1993. A. Kindness, A. Macias and F.P. Glasser. Waste Management (in press). A. Kindness, E.E. Lachoswki, A.K. Minocha and F.P. Glasser. Waste Management (submitted). L.P. Moroni 'Immobilisation of U(W) in Cement' MSc Thesis, University of Aberdeen, 1993.
Environmental Aspects of Construction with Waste Materials JJJM Goumans. H A van der Sloot and lI8.G. Aalbers (Editors) 01994 Elsevier Science B.V. All rights reserved.
87
Coal Fly-Ash Leaching Behaviour and Solubility Controlling Solids R. Garavaglia - CISE Tecnologie Innovative - Segrate Milano, Italy P. Caramuscio - ENEL Centro Ricerche Residui - Brindisi, Italy
ABSTRACT Experiments have been Carried out with large-scale (1 cubic meter) lysimeter test cells filled with different coal fly-ashes. The leaching behaviour of up to 35 elements was investigated. Lysimeters were irrigated with distilled water under controlledconditions; leachate collection system and storage tank were kept under inert gas to avoid reaction with atmospheric gases. Integrated leachate samples have been collected on a weekly basis and analyzed for complete chemical characterization. Examples of the mobilization trends are presented. Measured concentration values were used to calculate elemental speciation in solution with the geochemical code MINTEQA2. Our results support the hypothesis that geochemical reactions, which can be modelled and predicted, control the leaching of elements from coal fly-ashes. A number of solubility controlling minerals are suggested for both major and trace elements, in the pH range from 8 to 12 (Al, Ba, Ca, Cr, Cu, Fe, Mg,Mn, Ni, Pb, S, Si, Sr, and Zn). No solubility controlling solid pbases were found for As, Mo, Sb, V e U. From these results, predictions could be denved on the leachate elemental concentration as a function of pH. Generally, the predicted values are well correlated to the experimental data.
Introdktion The use of fossil fuels for electric power production has increased throughout the world in the past decade. As a result, huge amount of combustion solid waste, mainly coal fly-ash and bottom ash, are produced. At present, coal-fired power plants account for more than 15% of total Italian electric power generation, and produce about 1.5 Mton/year ashes. Though coal fly-ash is a potentially marketable secondary material, for many different reasons (last but not least Italian law enforcements) substantial amounts are currently disposed of in landfills. Both the fly-ash and the bottom-ash are enriched in major and trace elements which have the potential to adversely affect ground water quality if released into the subsoil in sufficient amount. These elements include arsenic, boron, chromium, molybdenum, selenium, sulphur, and vanadium. There is a need to: 1) accurately assess the mobilization of elements that results from the weathering of these wastes; 2) understand those chemical processes in the ashlwater system that dictate the concentrations of these elements in the leachate. The final goal of our research effort is to develop a predictive capability for short- and longterm leaching behaviour of fossil fuel solid wastes. In the past, most studies on the effects of combustion waste disposal on ground water quality have been focused on characterizing the chemical composition of the waste and collect empirical measurements of elemental extraction rates in short-term leaching tests(' to 'I.
88
Rai and others authod7 to 17) have shown that an approach based on thermochemicaldata and principles can be applied in predicting the upper limits of elemental concentration in leachates generated from these wastes. If a solubility controlling compound that has a rapid dissolutionprecipitation kinetics is present in the waste, or if a secondary phase forms that dissolves and precipitates quite rapidly, then the solid phase will be in equilibrium. In this case, the concentrations of the element in the liquid phase can be predicted on the basis of the precipitation-dissolution reactions defined by thermochemical data. This approach has been particularly successful when applied to coal fly-ash and municipal solid waste incinerator residues and has been demonstrated('*) to have general validity also for different waste materials. We are currently carrying out experiments with large-scale (1 cubic meter) lysimeter test cells filled with different coal fly-ashes. The leaching behaviour of up to 35 elements has been investigated. Measured concentration values were used to calculate elemental speciation in solution with the geochemical code MINTEQA2. Our results support the hypothesis that geochemical reactions, which can be modelled and predicted, control the leaching of elements from coal fly-ashes. From these results, predictions could be derived on the leachate elemental concentration as a function of pH. Generally, the predicted values are well correlated to the experimental data.
Ash Samples Two coal ashes were investigated: a slightly alkaline fly ash from representativecoal combustion in an Italian coal-fired 1) power plant; the choice was based on a comprehensive chemical and physical characterization of fly-ashes generated at different plants from different coals. a strongly alkaline fly-ash, derived from Flue Gas Desulphurisation experiments by 2) means of direct injection of limestone (calcium carbonate) into the burner; this results in sorption of sulphur oxide by calcium sulphatehlphite formation and decomposition of excess lime to give calcium oxide. Upon leaching, the calcium oxide tubs into calcium hydroxide, which confers high pH values to the leachate. This quite unusual fly-ash was selected in order to investigate solubility-controlling solids over a pH range spanning up to high alkaline values. Throughout this paper, those samples will be referred to as Fly Ash and Alkaline Fly Ash, respectively. The elemental composition of both samples is reported in table 1; results of quantitative mineralogical characterization are listed in table 2. Particle size distribution and radioisotopes content measurement, together with standard batch leaching procedures were also performed.
89 Table 1:
total concentration of major and trace elements in the fly ash samples used in this study.
Total concentration (mg/kg)
Table 2:
I
Ash
I
Ba
I
10001
1800
Cd
I
0.2
I
0.2
Fe
1
50300
1-
I
Ash
I
Si
I
233000
I
181000
T1
I
2.0
I
0.6
Alkaline Fly Ash
Alkaline Fly Ash
3oooO
mineralogical composition of the fly ash samples used in this study. (N.D. = not detected).
Mineralogical composition (% w/w) Fly Ash
Alkaline Fly Ash
8.4
5.0
14.4
15.3
CaSO,
N.D.
2.4
Lime
CaO
N.D.
9.2
Calcite
CaCO,
N.D.
2.3
Quartz
SiO,
Mullite
2Si0,
Anhydrite
Amorphous
+ 3A1,0,
I
77
I
65
90
Lysimeter cells The lysimeter cells are cylindrical vessels made of PVC, about 1 m3 available volume. Only inert materials (plastics and washed quartz sand) were allowed to contact the residues into the cells. Leachate was collected at the conical bottom of the vessels and stored in polyethylene tanks. In order to avoid any reaction with atmospheric gases, leachate collection system and storage tank were continuously flushed with nitrogen. The lysimeters were covered so to avoid uncontrolled rainwater to enter them, but to allow free evaporation. The cells were irrigated with distilled water through a sprayer system; 25 litres were delivered once a week, thus simulating a 25 mm rainfall. The dry ashes were mixed with an amount of water (previously determined by Proctor compaction tests), then layered over a quartz sand bottom layer, compacted and topped with more quartz sand. Hydraulic properties of the ash beds were characterized through porosity and permeabilit measurements. Edometricpermeabilities(according to ASTM D2435) turned out to be 8 10-gYand 5 lo4; total porosities resulted in 53% and 61 %, for the compacted Fly Ash and for the Alkaline Fly Ash, respectively. The latter showed a marked water reactivity, both during compaction and subsequent irrigation; moreover, it shrank so that the first few leachate samples percolated through fissuration. After that the compacted material changed and showed hydraulic permeability through porosity. The experiment on Fly Ash started in late 1991, and it has been carried on until November 1993; test on Alkaline Fly Ash begun in summer 1992 and it is still under way. During harsh winter periods, the test had to be stopped due to freezing. Leachate collection and analysis Leachate integrated samples have been collected on a weekly basis and analyzed for complete chemical characterization. Leachate was removed from the storage tank by a peristaltic pump, on-line filtered and measured for quality parameters such as pH, conductivity and redox potential. Aliquot were preserved and stored for the different analytical tasks. A variety of instrumental techniques have been used: titrimetry (for alkalinity determination); Ion Chromatography(anionsdetermination);Flame Atomic Emission, ElectrothermalAtomization Atomic Absorption, Inductively Coupled Plasma Optical Emission and Inductively Coupled Plasma Mass Spectrometry (elemental content); radionuclides were also monitored by Gamma-Ray or Alpha Spectrometry. Thanks to the multielemental capability of ICP-MS instrumentation, the concentration of up to 35 elements was routinely determined; many other elements were also monitored, their concentrations being lower than the analytical detection limit. Geochemical modelling The geochemical code MINTEQA2(”) was used to calculate the element speciation in the leachate solutions. Some implementation were done to the standard thermodynamic database of the MINTEQA2 code; two were most significant: a) the addition of the component M o o t - together with equilibrium constants for aqueous species and solids as published by complex Krupka et al.(20), and b) the modification of the stability constant for the CU(OH)~” from the original value of log K = - 13.7 to the value log K = - 16.2, as discussed by Comans et al.(17) and reported by others(21). Forecasting on element concentrationwas obtained by assuming equilibrium between leachate and potential solubility controlling minerals in the ash. These minerals were selected on the basis of their saturation indexes in prior MINTEQA2 run and the fitting of free ion activities (calculated by means of speciation runs) with theoretically expected values (from thermodynamic solubility constants).
91
Results and discusswn Leachate general characteristics Lysimetry on Fly Ash was stopped at a WS ratio of 1.32, corresponding to 2.5 pore volume; the test on Alkaline Fly Ash is still under way and it has reached a L/S ratio of 0.73 (1.19 pore volume). Figure 1 shows the leachates pH values vs. pore volume. Fly Ash gave an almost constant value between 8.5 and 9.0 during the first year of investigation, the pH then steadily arose up to 10.8 and finally dropped to about 9.2. Alkaline Fly Ash, as expected, gave a more strongly alkaline leachate, with a maximum pH value of 12.9. Both leachates had reducing properties, more pronounced for Alkaline Fly Ash, as shown by redox potential (figure 2). Leachates showed typical high sulphate content ( > 2000 mg/l), with high concentration of alkaline metals (Na up to 400 mg/l and K about 200 mg/l), and Ca (about 500 mg/l). Chloride and Mg content were also significant (with maxima of more than 50 mg/l), while B, bromide, Li, Mo, and Si were between 1 and 20 mg/l range. All other elements were at trace levels (less than 1 mg/l).
Figure 1: leachates pH values vs. pore volume
Figure 2: lachates redox potential (referred to Ag'IAgCI) vs. pore volume
Elements leaching behaviour in Fly Ash Respect to their mobilization trend, elements leached from Fly Ash can be grouped into four broad categories: 1) having an almost constant concentration throughout the investigated U S range; they are B, Ba, Ca, Co, S (sulphate), Si, Sr, W, and Zn.This behaviour is typified in figure 3 and 4 by B and Sr. showing an initial release followed by a more or less steep decrease to lower 2) concentration; these include Br, C1, Cr, Fe, K, Li, Mg, Mn, Mo, Na, and U. Examples are given in figure 5 and 6 for Mn and Mo. exhibiting delayed release, with concentration "peaks" in later stage of the leaching 3) process, as Al, Cs, Rb, Re, Se, Te, and V. Figures 7 and 8 ilkstrate this behaviour in the case of Ni and V. 4) having an increase in concentration still at the later time of the process, as shown in figure 9 for As; this group includes As, Ga, and Sb. A fifth group might include those elements whose concentration was always found to be less than the instrumental detection limit; the most significant being Cd, Th and Ti.
92 .-
I
I
O I
0 0.2 0.4 0.8 Od
1
1.2 1.4 1.8 1.0 2
r l l l l l , l l r T l l
0 0.2 0.4 0.0 0.8 1
2.2 2 4 2.8
pow vobne
1.2 1.4 1.8 ld
2
2.2 2 4 2.8
pore volum
Figure 3: B leaching behaviour vs. pore volume
Figure 4 Sr leaching behaviour vs. pore volume
.:'
,i"i-----, ,
,
,
pom volume
Figure 5: Mn leaching behaviour vs. pore volume
0 C U 0 4 0 8 Od
1
,
,
~
1.2 1 4 1 8 1.8 2 22 Z4 2 8
pare volume
Figure 7 Ni leaching behaviour vs. poxe volume
0 0
02 04 08 00
1
1 2 1 4 1.8 V.0
2
22 24 28
pore volume
Figure 6 Mo leaching behaviour vs. pore volume
0 0.2 0.4 0.8 0.8 1 1.2 1.4 1.8 1.8 2 22 2.4 1.8
pow volume
Figure 8: V leaching behaviour vs. pore volume
NOTE:instrumental detectionlimit was indicated when significant respect to measured concentration
0 0.2 0.4 0.11 Od
1
1.2 1.4 1.8 1.0 2 2 2 2 4 2.8
pore volume
Figure 10 As leaching behaviour vs. pore volume
93
Aqueous speciation and identification of solubility-controllingsolids The concentrations in the liquid phase measured in some selected leachate samples from the early stage of the lysimetry precess of both ashes were subjected to geochemical speciation using MINTEQA2 code. The calculated free ion activities and the mineral phases saturation indexes were the obtained most relevant information derived. Aluminium, The free ion activities of A13+ in speciated solution fit well with theoretical values calculated for Al(OH)3 amorphous and crystalline form (gibbsite), as shown in figure 10. It seems that Al(OH), amorphous up to pH 8.5, and gibbsite at higher pH values, are the solubility controlling minerals for Al. Those secondary solid phases dictate the upper permissible A1 concentration into leachates through a reci itation-dissolution mechanism. Similar conclusions were obtained by other authors(5p'*13* 4i17). Based on this model hypothesis, equilibrium total A1 concentration were calculated over a pH range from 7 to 13. These predicted values are well correlated to the experimental data from the lysimetry tests (figure 11; the downward segment of the solid line between pH 8.5 and 9 in figure is intended just for graphical reasons to indicate the switching over the two minerals; it should be conveniently represented by a shaded area).
P P
+
-16 -
I
5-103.2JoO.O1
7
8
s
i
I
o.rn14
561 10
11
12
IS
7
0
I
PH
10
11
11
1s
PH
Figure 10: A13+ activity vs. pH calculated from experimental data (dots) and predicted (solid lines)
Figure 11: total dissolved Al in leachates and prediction assuming equilibrium with am. Al(OH)3 and gibbsite (solid line)
Arsenic, MINTEQA2 speciation runs indicated As to be present as As(II1 ; this is consistent with results from Turner(' 1. None of the As solid phases included into MINTEQA2 data base was found to be a potential solubility-controlling solid. The leachate concentration of this element might be 0.1 controlled either by precipitation/dissolution reaction, adsorption/desorption or dissolution rate of the fly ash matrix. In a previous o Po u) a, t m i a o i u ) i m t w m o m z ( o column leaching investigation on the same Fly us Ash we had evidence of a much delayed As Figure 12: total As concentration vs. L/S ratio in release at very high W S ratio (figure 12). This suggest adsorption as a candidate controlling Fly Ash column leaching mechanism.
4
(ID
94
Barium, Ba++ ion activities in leachates did not perfectly agreed with that deriving from barite &SO4 solubility (figure 13). In Alkaline Fly Ash leachates, measured concentration were in many cases lower than the instrumental detection limit. On the other hand, no other potential solubility controlling minerals showed to fit the data. A coprecipitated (Ba, Sr)S04 solid was suggested by Ainsworth and Rai(') and others('2i 13* 14); our data on calculated activities did not support this hypothesis. Anyway, total solution concentration based on barite equilibrium gave reasonable prediction, as shown in figure 14. 1
0.1
2 -I0 -11
o.mc
i! 7
0.01
I
0.0001
8
9
10
11
te
8
9
PH
I1
10
*2
13
PH
Figure 13: B a + + activity vs. pH calculated from experimental data (dots) and predicted (solid line)
Figure 1 4 total dissolved Ba concentration in leachates and prediction assuming equilibrium with barite (solid line)
Calcium and Sulphur, Ca leaching behaviour in Fly Ash was satisfactorily explained with gypsum CaS04'2H20dissolutionlprecipitationreaction. Alkaline Fly Ash leachates had lower sulphate content and resulted grossly undersaturated; at pH greater than 11, wairakite CaAl2SipOI2*2H,O theoretical solubility curve (assuming gibbsite equilibrium, and log[H4Si04] = -4.5 from experimental figures) gave a good fit, as figure 15 shows; as it was already explained, the segment between pH 10,5 and 11 is merely conventional. The modelled total concentration confirmed agreement with experimental data (figure 16). As a consequence, also sulphate concentration could adequately be predicted on the same hypothesis (not shown).
.
1.000-
2
im-
t
10-
1-
I
-8
I
I
0
10
11
12
13
PH
Figure 15: Ca++ activity vs. pH calcualted from experimental data (dots) and predicted (solid lines)
0.1
I
! 7
0
0
10
I1
I2
1s
PH
Figure 16: total dissolved Ca in leachates and prediction assuming equilibrium with gypsum and wairakite (solid line)
95
Chromium, Due to the reducing properties of the leachates, Cr was calculated to be in present as Cr(III). Calculated c?+ ion activities are in good agreement with theoretical values for amorphous and crystalline Cr(OH)3, depending on pH (figure 17). Predicted total Cr concentration in leachates nicely match the experimental values, as shown in figure 18. Once more, the segment of the solid line between pH 8.5 and 9 is intended just for graphical reasons to indicate the switching over the two minerals; it should be conveniently represented by a shaded area. im
-5
10-
1-
90-
56,
7
om01 8
@
10
11
12
I
8
7
13
rtMI :: 0 10
PH
I
I1
I2
1s
PH
Figure 17: C?+ activity vs pH calculated from experimental data (dots) and predicted (solid line)
Figure 18: total dissolved Cr m leachates and prediction assurmng equilibrium with cristalline and amorphous Cr(0H)j (solid line)
Copper, Cu equilibrium is suitably explained with tenorite CuO dissolution reaction over the investigated pH range; this is consistent with results presented by F r ~ c h t e r ( ' ~ ~ .' ~Man , ' ~Y) of the analytical data fell below the instrumental detection limit. d
10
d1
-+
-10-
2
3 -P-10
-
-18
-
0 01
o mi D.L. 0 om1
-20 7
7
\
01
+ -12-
8
D
10
11
12
13
PH
Figure 19: Cu++ activity vs. pH calculated from experimental data (dots) and predicted (solid line)
8
-0
lo
11
12
1s
PH
Figure 20: total dissolved Cu in leachates and prediction assuming equilibrium with tenorite (solid line)
Carbonate, The total gas amount available for exchange with percolating waters is limited by diffusion through the porous media. Thus, leachate solutions resulted not to be in equilibrium with atmospheric CO,, as figure 21 shows. The calculated logs of C03= activity showed a linear relationship with pH values that might be roughly approximated as follows:
96
Equation (1) fits the experimental data and was used to model the carbonate activity in prediction of total concentration for all elements controlled by carbonate solid phases. In figure 21 both the line from equilibrium with atmospheric gases (solid) and the line from equation 1 (dotted) are plotted.
,
,?-
d7
o
D
10
ii
12
is
PH
Figure 21:calculated C03-- activities vs. pH
lrt~~~Due to the low redox potential of leachates solutions, speciation calculation indicated Fe to be present mainly as Fe(I1). Siderite FeCO,, a possible solubility controlling solid, resulted over 5 order of magnitude undersaturated. On the other hand, ferrohydrite Fe(OH), solubility could explain the calculated logs of F e + + + activity (figure 22). This is also consistent with findings of Fru~hter('~? 14). Total iron concentration are adequately predicted by this hypothesis, as figure 23 shows for an average Eh value of -100 mV.
4J! 7
a
I 0
10
11
12
19
PH
Figure 22: Fe3+ activity vs. pH calculated from experimental data (dots) and predicted (solid line)
o.m1
! 7
o
D
10
11
ia
I
ia
PH
Figure 23: total dissolved Fe in leachates and prediction assuming equilibrium with ferrohydrite (solid line) at Eh=-lOO mV
Magnesium, In the pH range from 8 to 10, which is typical of fly-ash leachates, Mg leaching behaviour appears to be controlled by dolomite precipitation-dissolution.At higher pH brucite Mg(OH)2 seems to control solubility, as reported in figure 24. The dotted line in figure 24 represents Mg equilibrium at carbonate activities derived from empirical relation (equation l), and assuming Ca ion activities corresponding to gypsum solubility. Measured Mg in leachate fits nicely with predicted total dissolved Mg concentration, as shown in figure 25 (the segment between pH 9 and 10 is meant only for graphical reasons). Brucite control at high pH values was also found by corn an^('^) for the leaching of MSW incinerator residues.
91
-10 42
! 7
I 8
D
10
11
12
0.m1
J
I
7
13
8
0
PH
10
11
12
1s
PH
Figure 24: Mg++ activity vs. pH calculated from experimental data (dots) and predicted (lines)
Figure 25: total dissolved Mg in leachates and prediction assuming equilibrium with dolomite and brucite (solid h e )
Manganese and Nickel, Once the empirical correlation (equation 1) of carbonate vs. pH was assumed, the pH dependence of calculated Mn+ and Ni+ activities closely resembled that of carbonate mineral phases: rhodochrosite MnC03 and NiC03.6H20, respectively. Other possible solubility controlling solids (hydroxides) would have given a defferent slope, equal to -2pH. Thermochemic data for NiC03.6H,0 are not included into the MINTEQA2 database, and were taken from Kr~pka(~').As figure 26 and 27 show, modelled solutions are to some degree undersaturated respect to the proposed minerals. Carbonates are known to form mixed solid phases, whose solubility is lower than the pure one. As a consequence, predicted total dissolved concentration based on MnC03 and NiC0,.6H20 solubility were one order of magnitude overestimated (not illustrated); nevertheless, they can be used to set the upper limit of possible leachate concentration. +
0
+
I
I
I -14 7
1 8
0
10
2 11
12
i 13
PH Figure 26: Mn++ activity vs. pH calculated from experimental values dots) and predicted (dotted line)
-16
-
7
8
1 0
10
4 11
12
L1 s
PH
Figure 27: N i + + activity vs. pH calculated from experimental data (dots) and predicted (dotted line)
Molvbdenum. Mo forms very mobile oxyanions and gave quite high concentration levels in the early stage of the Fly Ash leaching process (figure 6). The most insoluble Mo compound are thought to be powellite CaMo04, PbMo04, and SrMo04; none could explain the Moo4-ion activities as calculated by MINTEQA2. Powellite was already suggested as a controlling solids in the leaching of MSW incinerator residues(17) and in hot-water extract of many fly ashes('), but was excluded for coal fly-ash leaching(13714).
98
LggL Pb leaching behaviour was explained on the basis of Pb(OH), solubility (figure 28). Experimental data were satisfactorily fitted by predicted total dissolved concentration, plotted in figure 29. 1
I
d-
-
d-
+ -10 -
g.1, 8.1. -10 -
-11 -
m ! 7
I)
D
10
11
12
7
13
1
0
PH
10
11
12
13
PH
Figure 28: Pb++ activity vs. pH calculated from
Figure 29: total dissolved Pb concentration in
experimental data (dots) and predicted (solid line)
leachates and prediction assuming equilibrium with Pb(OH)2 (solid line)
Antimony, MINTEQA2 calculation showed the species Sb(OH), to be dominant. The only Sb solid that could undergo a precipitation-dissolutionreaction resulted SbO,. The equilibrium between Sb components into solution (as defined in MINTEQA2) and SbO, depends on both pH and Eh values. Calculated activities in leachates do not support SbO, as a candidate solubility controlling solid. Silicon, Si concentration resulted consistent with quartz solubility at pH lower than 10, and with wairakite CaAl2Si40,,~2H,O at pH value up to 12 (figure 30); wairakite theoretical solubility was calculated assuming gibbsite and gypsum equilibrium for A1 and Ca, respectively. Predicted leachate concentration vs. experimental data are shown in figure 31. At very high pH value, prediction are grossly overestimated, pointing out that some other control reaction has to take place.
0.1
7
8
0
10
11
12
PH
I
4 7
8
0
10
11
12
13
PH
Figure 30: HqSiOq activity vs. pH calculated from
Figure 31: total dissolved Si concentration in
experimental data (dots) and predicted (solid lines)
leachates and prediction assuming equilibrium with quartz and wairakite (solid line)
99
0 -2-.
l.m,
.--- ..
im -
-.
Y( 10-
- 4
+ &
B
--. -,*
- - - .. --. ~
-eb*
--.
4-
.lmndal.
-10 -
.--
e
1
I0.1 0.01 -
-12 7
om1
Vanadium, Calculated activities for V02+ ion vs. pH lied on a straight line that did not fit 4 the theoretical values derived neither from Ca2V207 nor from Pb2V207. We were not able to identify any other candidate solubility controlling solid. V leaching behaviour showed increasing concentration still at the end of the lysimeter test (figure 8). Previous column 1leaching experiments on the same Fly Ash O I I confirmed delayed long-term release with 0 50 im 150 2m 260 much higher concentration at L/S ratio > 20 us (figure 34). Figure 3 4 total V concentrationvs. LIS ratio in Fly Ash column leaching
Uranium, Modelled U4+ ion activities had the same slope vs. pH than the theoretical data pertinent to uraninite UOz, but resulted up to four order of magnitude lower; the latter being the most stable U solid phase in MINTEQA2 data-base. Therefore no control by a mineral phase could be proposed.
Zinc. The plot of Zn'
calculated activities vs. pH is reported in figure 35 with solubility data for different solid phases: zincite ZnO, orthosilicate Zn2Si04, and metasilicate Zn2si0, (assuming quartz equilibrium), and hydrozincite ZnS(OH),(CO,), (assuming log[CO,--l = 14 - pH). The closest fit refers to Zn2Si04. The slope of modelled data points seems to match that of hydrozincite, with two order of magnitude undersaturation. As it was the case for Ni and Mn, also for Zn mixed carbonate phases are possible, whose solubility is much less than the theoretical one. Anyway, prediction of concentration levels based on Zn orthosilicate, reported in figure 36, gave quite reasonable estimate. +
100
I
a! 7
I
8
10
11
1P
1.3
PH
Figure 35: Zn++ activity vs. pH calculated from experimental date (dots) and predicted (lines)
7
I
e
10
11
1P
1s
PH
Figure 36: total dissolved Zn in leachate and prediction assuming equilibrium with orthosilicate
Our results from large-scale lysimetry experiments support the hypothesis that geochemical reactions, which can be modelled and predicted, control the leaching of elements from coal fly-ashes. A number of solubility controlling minerals are suggested for both major and trace elements, in the pH range from 8 to 12, namely: A1 concentration in leachate seems to be consistent with Al(OH)?- equilibrium, in its amorphous or crystalline form, depending on the pH; Ba mobilization appears to be controlled by barite &SO4 solubility; Ca is controlled by gypsum CaS04.2H20 at pH lower than 10.5 and by wairakite CaA12Si4012~2H20 at higher pH; Cr is at equilibrium with amorphousor crystalline Cr(OH),, depending on pH value; Cu is controlled by tenorite CuO dissolution; Fe is at equilibrium with Fe(OH),; Mg seems to be controlled by dolomite CaMg(CO,), at pH lower than 11, and then by brucite Mg(OH)2; Mn seems to be controlled by rhodochrosite MnC0, dissolution; Ni appears to be at equilibrium with the carbonate NiC0,.6H20; Pb is controlled by its hydroxide Pb(OH)2; Si concentration is consistent with quartz solubility at pH lower than 10 and with wairakite CaAI2Si4Ol2~2H2O at higher pH value; Sr Seems to be at equilibrium with celestite SrSO, up to pH = 10 and then with strontianite SrCO,; Zn solubility seemed to match that of Zn2Si04. For those elements forming oxyanions, As, Mo, Sb, V, and U, no solubility controlling solid phases have been found. The leachate concentrationof these elements might be controlled by an adsorption-desorption mechanism. From these results, predictions could be derived on the leachate elemental concentration as a function of pH. Generally, the predicted values are well correlated to the experimental data from coal fly-ash lysimetry.
101
References
(9)
Page, A.L.; Elseewi, A.A.; Straughan, I.R.. 1979. Physical and Chemical Properties of Fly Ash from Coal-Fired Power Plants with Referece to Environmental Impacts. Residues Rev. 7 1:83. Adriano, D.C.; Page, A.L.; Chang, A.C.; Straughan, I.. 1980. Utilization and Disposal of Fly Ash and Other Coal Residues in Terrestrial Ecosystems: A Review. J. Environ. Qual. 9:333. Summers, K.V.; Rupp, G.L.; Gherini, S.A.. 1983. Physical-ChemicalCharacteristic of Utility Solid Wastes. EPRI EA-3236. Electric Power Research Institute, Palo Alto, CA . Suloway, J.J.; Skelly, T.M.; Roy, W.R. ; Dickerson, D.R.; Schuller, R.M.; Griffin, R.A.. 1983. chemical and Toxicological Properties of Coal Fly Ash. Environ. Geology Notes 105. Illinois Dep. of Energy and Nat. Resour., Champaign, IL. Ainsworth, C.C.; Rai, D.. 1987. Chemical Characterization of Fossil Fuel Combustion Wastes. EPRI EA-5321. Electric Power Research Institute, Palo Alto, CA. Wu, E.J.; Chen, K.Y.. 1987. Chemical Form and Leachability of Inorganic Trace Elements in Coal Ash. EPRI EA-5115. Electric Power Research Institute, Palo Alto, CA. Rai, D.; Ainsworth, C.C.; Eary, L.E.; Mattigod, S.V.; Jackson, D.R.. 1987. Inorganic and Organic Constituents in Fossil Fuel Combustion Residues. EPRI EA5176. Electric Power Research Institute, Palo Alto, CA. Rai, D. ; Eary, L.E.; Mattigod, S.V., Ainsworth C.C.; Zachara J.M.. 1987. Leaching Jkhaviour of Fossil Fuel Wastes: Mineralogy and Geochemistry of Calcium. p. 3-15 in G.J. McCarthy et al. (ed.) Fly Ash and Coal Conversion By-products Characterization, Utilization,and Disposal III. Materials Research Society Symposia Proceedings. Vol. 86. Materials Res. Soc., Pittsburgh, PA. Rai, D.; Mattigod, S.V. ;Eary, L.E.; Ainsworth, C.C.. 1988. Fundamental Approach for Predicting Pore-Water Composition in Fossil Fuel Combustion Wastes. p. 317. In G.J. McCarthy and F.P. Glasser (ed.) Fly Ash and Coal Conversion By-products Characterization, Utilization, and Disposal. Materials Research Society, Meterials Research Society Symposia Proceedings Vol. 113. Materials Res. Soc., Pittsburg, PA. Roy, W.R.; Griffin, R.A.. 1984. Illinois Basin Coal Fly Ashes. 2. Equilibria Relationship and Qualitative Modelling of Ash-Water Reactions. Environ. Sci. Technol. 18:739. Mattigod, S.V.. 1983. Chemical Composition of Aqueous Extracts of Fly Ash: Ionic Speciation as a Controlling Factor. Environ. Technol. Lett. 4:485. Fruchter,J.S.; Rai, D.; Zachara, J.M.. 1990. Identification of Solubility-Controlling Solid Phases in a Large Fly Ash Field Lysimeter. In Proceedings: Environmental Research Conference on Groundwater Quality and Waste Disposal. EPRI EN-6749. Electric Power Research Institute, Palo Alto, CA. Fruchter, J.S. ; Rai, D.; Zachara, J.M.; Schmidt, R.L.. 1988. Leachate Chemistty at the Montour Fly Ash Test Cell. EPRI EA-5922. Electric Power Research Institute, Palo Alto, CA. Fruchter, J.S.; Rai, D.; Zachara, J.M.. 1990. Identification of Solubility-Controlling Solid Phases in a Large Fly Ash Field Lysimeter. Environ. Sci. Technol. 24: 1173.
102 (15)
Mattigod, S.V.; Rai, D.; Eary, L.E.; Ainsworth, C.C.. 1990. Geochemical Factors Controlling the Mobilization of Inorganic Constituents fron Fossil Fuel Combustion Residues: I. Review of the Major Elements. J. Environ. Qual. 19: 188. Eary, L.E.; Rai, D.; Mattigod, S.V.;Ainsworth, C.C.. 1990. Geochemical Factors Controlling the Mobilization of Inorganic Constituents from Fossil Fuel Combustion Residues: 11. Review of the Minor Elements. J . Environ. Qual. 19:202. Comans, R.N.J.; van der Sloot, H.A.; Bonouvrie, P.A.. 1993. Geochemical Reaction Controlling the Solubility of Major and Trace Elements During Leaching of Municipal Solid Waste Incinerator Residues. In J. Kilgroe (ed.)Proceeding 1993 Municipal Waste Combustion Conference. March 30 - April 2, Williamsburg, Virginia.AWMA, Pittsburg, PA. van der Sloot, H.A.. 1991. Systematic Leaching Fkhaviour of Trace Elements from Construction Materials and Waste Materials. p. 341-352. In J.J. Goumans, H.A. van der Sloot, Th. G. Aalbers (4.)Waste Materials in Construction. Elsevier Science Pubblishers, Amsterdam, NL. Allison, J.D.; Brown, D.S.; Novo-Gradac, K.J.. 1991. MZNTEQA2IPRODEFA2,A Geochemical Assessment Model for Environmental @stems: Version 3.0 User's Manual. EPAl60013-917021. U.S. Environmental Protection Agency, Athens, GA. Krupka, K.M.; Erikson, R.L.; Mattigod, S.V.; Schranmke, J.A.; Cowan, C.E.. 1988. Thermochemical Data Used by the FASTCHEM Package. EPRI EA-5872. Electric Power Research Institute, Palo Alto, CA. Morel, F.M.M.. 1983. Principles ofAquatic Chemistry. John Wiley and Sons, New York, NY. Turner, R.R.. 1981. Oxidation State of Arsenic in Coal Ash Leachate. Environ. Sci. Technol. 15: 1062.
Environmental Aspects of Construction wilh Waste Materials JJJ.M. Goumans, H A . van der SIoot and l3.G. Aalbers (Editors) @I994 Elsevier Science B, K AN rights reserved.
103
MODELLING CA-SOLUBILITY IN MSWI BOTTOM ASH LEACHATES Rob N.J. Comans and Jeannet A. Meima Netherlands Energy Research Foundation (ECN) P.O. Box 1 , 1755 ZG Petten, The Netherlands
Abshsct This study focuses on modelling the leaching of the major element calcium from fresh MSWI bottom ash leachates. It is emphasized that Ca-minerals in bottom ash mainly control leachate pH and hence, exert a strong influence on the leaching of contaminant metals and oxyanions from these waste materials. The geochemical speciation code MINTEQA2 was used to model the dissolved Ca concentration in leachates collected from carefully controlled batch laboratory experiments with fresh bottom ash covering a range of liquidkolid ratios and pH. and gypsum (CaS04*2H,0) control It is shown that ettringite (Ca,AI,(S04),(OH),,~26H~O) Ca-leaching from fresh bottom ash at pH 10-12, and pH < 10, respectively. Ca-availability at low pH appears, therefore, to be controlled by the dissolution of gypsum. It is hypothesized that the coexistence of the minerals ettringite, gypsum, and gibbsite (AI(OH),) controls the pH of fresh bottom ash at a value of approximately 10.2. The influence of the Ca-chemistry of leachates from alkaline waste materials, such as MSWI bottom ash, on contaminant leaching is discussed.
1.
INTRODUCI'ION
Bottom ash from municipal solid waste incinerators (MSWI) is produced worldwide in ever-increasing quantities and increasingly finds its destiny in construction. The relatively high concentrations of trace metals in comparison to natural solids such as soils and rocks [ I ] causes, however, environmental concern and may limit, depending on legislation, the use of bottom ash as a construction material. In a previous study [3], we have shown that geochemical reactions control the leaching of both major and trace elements from MSWI residues. A number of possible solubility controlling minerals and complexation processes in solution have been suggested that can to a large extent explain the observed leaching behaviour as a function of pH. Knowledge of these elementary processes enables us to modeVpredict the leaching of contaminants from bottom ash under the specific conditions of the environments in which it is applied. Moreover, once we know the mechanisms leading to the critical release of contaminants, we can apply geochemical knowledge to improve the
104
quality of bottom ash by interfering in these mechanisms. Many investigators have studied the leaching behaviour of bottom ash, focusing on the potential contaminants. If however, we want to reveal the underlying geochemical mechanisms controlling the leaching of contaminants, we need to understand the overall chemical changes in the ash when it contacts water. Bottom ash consists mainly of glasses and minerals, composed of geochemically abundant elements [1,2]. The dissolution of these phases and the reprecipitation of dissolved elements in new secondary solids affect not only the major ion chemistry but control also the concentration of trace elements. The leaching of these potential contaminants is strongly related to major chemical factors such as pH, redox potential and ionic strength, which in turn are related to the chemistry of the abundant elements. In the present study we focus on the Ca-chemistry of bottom ash and attempt to complete the modelling of the Ca-solubility in fresh bottom ash, using recently published thermodynamic data. We emphasize the importance of this element in particular, because Caminerals constitute for the major part the alkalinity of alkaline waste materials, such as MSWI bottom ash [2], coal fly ash [4], and steel slag [5]. The dissolution and reprecipitation of Caminerals mainly control, therefore, leachate pH in both the fresh materials and during weathering. We believe that the ability to model leachate pH is mandatory for the prediction of contaminant leaching and, as we will discuss below, that the quality of bottom ash, and probably other alkaline waste materials, can already be improved significantly by modifying pH, through interference in the Ca-chemistry.
2.
MATERIALS AND METHODS
2.1. Bottom ash samples The MSWI bottom ash samples, AVI-1 and AVI-2, have been described in a previous study [3], and were selected from two representative Dutch incinerators. The samples (1 5-20 kg each) were ground in a jaw crusher until all material, except for < 500 g of metal scrap, passed through a 2 mm sieve. Total concentrations of the major elements in the two bottom ash samples are listed in Table 1. Table 1 . Total concentration of major elements in the bottom ash samples of this study. Total concentration ( g k g ) AVI-1
AVI-2
AVI-1
AVl-2
Na
19.4
18.8
Al
39.8
41.9
K
10.1
10.4
Fe
56.7
62.4
Ca
94.6
86.8
Mn
0.87
0.81
Mg
12.8
11.2
CI
5.17
4.60
22 1.5
246.1
Si
S,,
11.1
8.13
105
2.2. pH-stat leaching experiments Leaching experiments were performed in a pH-stat system, since the solubility of many (trace) elements is known to be strongly dependent on pH. The system allowed the pH to be monitored and adjusted to setpoint in 8 reactors simultaneously in a range between pH 4 and 13. Bottom ash suspensions in water were prepared at liquid/solid (L/S) ratios of 2, 5 and 10 by adding 75, 30 or 15 g of solid material to 150 mL of nanopure H,O in 250 mL Teflon (PFA) reactors. The reactors were kept open to the atmosphere at 25+1 OC and were stirred continuously during the 24-hour equilibration period, using a Teflon coated magnetic stirring bar. The pH was adjusted with 1 M analytical grade HNO, or NaOH. After equilibration, the suspensions were filtered through 0.45 pm membrane filters. Major elements in the solutions were analyzed using a combination of inductively coupled plasma-atomic emission spectrometry (ICP-AES), flame atomic absorption spectrometry (AAS), and ion selective electrodes (ISE). Trace element analyses and results are reported elsewhere [3].
3.
RESULTS AND DISCUSSION
3.1. Geochemical modelling The geochemical speciation model MINTEQA2 (version 3.1 1) was used to calculate the element speciation in the solutions obtained from bottom ash leaching at different pH values in a range from 4 to 13. In addition, element concentrations were modelled/predicted by assuming equilibrium between the leachates (at L/S=5) and potential solubility-controlling minerals in the bottom ash solids. These minerals were selected on the basis of their saturation indexes in prior MINTEQA2 runs and/or on their likeliness to be present or formed under the experimental conditions. The thermodynamic data from the standard MINTEQA2 (version 3.11) database were used except that the mineral ettringite (Ca6AI2(SO,),(OH),,.26H,O) was added using the solubility product recently published by Atkins et al. [6]. The model predictions are presented as total element concentrations, rather than free ion activities, in the leachate solutions at each pH. This approach enables presentation of the results together with the analytical leaching data in a graph of log-concentration versus pH which maintains the characteristic shape and concentration levels of general leaching curves.
3.2. Caleaching Calcium is the major cation that is released from the bottom ash samples below pH 10 [3]. Total dissolved calcium concentrations in the leachates are shown in Figure 1 as a function of pH. Dissolved Ca-concentrations are very low at strongly alkaline pH (> 12), increase steeply over two orders of magnitude between pH 12 and 10, and show a gradual further increase when the pH decreases towards a value of 4. The highest Ca-concentrations at each liquid/solid (L/S) ratio approach the "available" concentration, measured using the Dutch availability test [7]. Below pH 10, the concentrations increase also with decreasing L/S ratio, in a nearly proportional manner, whereas Ca-leaching at strongly alkaline pH is
106 independent of L/S ratio.
3.3. Modelling Belevi et al. [8] have suggested that the major reactions in bottom ash involving Ca are:
CaO(s)
+
H,O
#
Ca2+ +
20H
(hydrolysis of lime in unquenched bottom ash)
(solubility of portlandite in hydrolysedkpenched bottom ash). The authors postulate that hydrolysis of lime starts immediately after the quench tank and progresses during storage of the fresh bottom ash. The bottom ash samples used for the present study have been quenched and, hence, are likely to contain portlandite. Figure 1 shows the Ca-concentrations modelled with MINTEQA2 for the conditions of our experiments, assuming equilibrium with portlandite. It is evident that, at the "natural" pH of the samples (i.e. pH 10.3-10.4), the leachates are not in equilibrium with portlandite. Only at pH 1 3 (with added base in the pH-stat) do the Ca-concentrations approach the solubility line of portlandite. Belevi et al. [8] and Zevenbergen & Comans [2] have also calculated that the leachates of their quenched bottom ash samples are undersaturated with respect to this mineral. In our previous modelling study of MSWI bottom ash [3], we have not been able to model Ca-solubility between pH 10 and 12. The Ca-mineral ettringite (Ca6AI,(SO,),(OH),,~26H,O)has been shown to exist in alkaline waste materials containing sufficient Ca, Al, and SO, [9].Ettringite is a well known mineral that plays an important role in cement chemistry [e.g. 101. A solubility product for this mineral was recently published by Atkins et al. [ 6 ]and was added to the MINTEQA2 database for the purpose of this study. Caconcentrations modelled assuming equilibrium with ettringite are included in Fig. I , The concentrations measured in the leachates between pH 10 and 12 follow the same slope and are similar to the values modelled on the basis of ettringite solubility. W e postulate, therefore, that it is ettringite that controls Ca-leaching between pH 10 and 12 and, hence, at the "natural" pH of the bottom ash samples used for this study. Very recently, ettringite has indeed been identified by XRD analysis of a fresh bottom ash sample from the same incinerator as sample AVI-1 [2].
107
100000
AVI4,US=lO \
C
g m
AVI-1, U S 5
10000 A
AVI-I, U S = 2
0
AVIP.US=lO
A
AVIZ,US=2
_ _ _ _ _ ETTRlNGlTE CI\LCITE
1
2
4
6
8
10
12
14
GYPSUM PORTLANDITE
PH Figuie 1 Dissolved Ca-concentrations in leachates from the two bottom ash samples, as a function of pH and liquidkolid ratio, and MINTEQA2 predictions assuming equilibrium with different Ca-minerals.
In addition to ettringite, two other minerals containing Ca, Al or SO, may coexist in equilibrium with these bottom ash leachates at a pH of about 10: gypsum (CaSO4*2H,O) and gibbsite (AI(OH),) [3]. It is interesting to note that MINTEQAZ calculations using these three minerals as coexisting phases, in equilibrium with the leachates, predict a leachate pH of 10.14. This value is very close to the "natural" pH of bottom ash samples AVI-I and AVI-2 (10.2-10.3). In a very recent review of leaching data of about 400 MSWI bottom ash samples, it is shown that more than half of the samples have a "natural" pH of about 10 [ I l l . We hypothesize that this pH value of fresh bottom ash results from the coexistence of ettringite, gypsum, and gibbsite in the bottom ash matrix. When the pH of the bottom ash samples is lowered to values of below pH 10, the leachates become undersaturated with respect to ettringite, and Ca-concentrations increase only slightly with decreasing pH. At L/S values of 10 Lkg, the solutions are calculated to be in equilibrium with gypsum (CaS0,*2H20). At lower L/S ratios, the leachates appear to be oversaturated with respect to this mineral. Gypsum has frequently been identified by XRD analysis of fossil fuel combustion residues [9]and MSWI bottom ash [ 1,2]. As gypsum seems to control Ca-leaching at low pH, this mineral may be the phase that controls Ca-availability when measured using the Dutch availability test [see e.g. 71.
3.4. Cahnrtion piwesses
The leaching experiments shown in Fig. 1 were performed in experimental vessels which were open to the atmosphere. Hence, carbonation of the bottom ash suspensions is likely to
108
occur, especially at high pH. When CO, is absorbed by the leachates, calcium carbonate is likely to precipitate according to the reaction:
Ca2+
+
CO,(g)
+
H,O
*
CaCO,(s)
+
2H’
Fig. 1 includes a line indicating the Ca-concentrations modelled assuming equilibrium with calcite (CaCO,) and independently measured total inorganic carbonate concentrations [3]. It is obvious that Ca-concentrations in the 24-hour leachates are not in equilibrium with calcite. However, fresh bottom ash from the same incinerator as sample AVI-I, exposed to the atmosphere for 100 days in an aqueous environment, releases Ca in concentrations close to those in equilibrium with calcite and the partial pressure of CO, in the atmosphere (P = 10-3,5atm.)[2]. The pH in those experiments had decreased to 8.0, which is close to the pH of calcite in equilibrium with the atmosphere (8.4). At this point it is important to emphasize that, for reliable modelling of Ca-solubility, it is mandatory to measure total dissolved carbonate. These measurements are not only needed to enable modelling the solubility of carbonate minerals, but also because dissolved carbonate strongly affects Ca-speciation at alkaline pH, because of the stability of the dissolved CaCO: complex (see also discussion in [2]).
3.5. Implications for contaminant leaching The Ca-chemistry of bottom ash can exert a strong influence on the leaching of potential contaminants. We have previously shown that the leaching of particularly the heavy metals Cd, Cu, Pb, Zn is probably controlled by (hydr)oxide or carbonate minerals [3]. The solubility and, hence, the leaching of those phases is strongly dependent on leachate pH. The solubilityminimum for these heavy metals lies between pH 8 and 9. The carbonation process, which changes the pH of the fresh bottom ash (controlled by ettringite/gypsum/gibbsite) from 10.2 to the value controlled by calcite in equilibrium with the atmosphere, i.e. 8.4, may, therefore, reduce heavy metal leaching. A further decrease of pH, e.g. by prolonged contact with acid rain or groundwater may, however, lead to a strong increase in heavy metal leaching. The acid neutralizing capacity of Ca-minerals in bottom ash constitutes, therefore, a very important factor controlling the period of time during which the system can maintain a pH of 8. Calcium minerals may also limit contaminant leaching through direct binding of the elements. Cadmium, and to a lesser extent other heavy metals, possess a strong affinity for the surface of calcite [I21 and may be sorbed on this mineral by coprecipitation or solidsolution formation during the carbonation of bottom ash. Ettringite has been shown to have a strong affinity for oxyanions of As, Se [I31 and possibly Sb and Mo. We are currently investigating these and other binding processes of contaminants to single solid phases in MSWI bottom ash.
109 4.
CONCLUSIONS
The geochemical speciation code MINTEQA2 has proven to be very useful in modelling the dissolved Ca concentration in leachates collected from carefully controlled batch laboratory experiments with fresh bottom ash, covering a range of liquidkolid ratios and pH. and gypsum (CaS04*2Hz0) We have shown that ettringite (Ca6AI,(S04),(OH),,-26H,0) control Ca-leaching from fresh bottom ash at pH 10-12, and pH < 10, respectively. It is hypothesized that the coexistence of the minerals ettringite, gypsum, and gibbsite (AI(OH),) in fresh bottom ash controls the pH at a value of about 10.2. In contact with the atmosphere, absorption of CO, by the alkaline leachates will lead to carbonation of bottom ash and the formation of calcite. In equilibrium with the atmosphere, this mineral controls the pH at approximately 8.4. In view of the solubility-minimum of the (hydr)oxide and carbonate phases that are likely to control heavy metal leaching, the carbonation process may have a favourable effect on the leaching of these contaminants. Laboratory experiments with single solid (Ca-)phases, such as ettringite and calcite, are needed to investigate their potential to bind metallic and non-metallic contaminants and limit leaching from bottom ash.
5,
REFERENCES
Kirby, C.S. A geochemical analysis of municipal solid waste ash. Ph.D. thesis, Department of Geological Sciences, Virginia Polytechnic Institute and State University (1 993). 2. Zevenbergen, C. & Comans, R.N.J. Geochemical factors controlling the mobilization of major elements during weathering of MSWI bottom ash. These proceedings. 3. Comans, R.N.J., van der Sloot, H.A. & Bonouvrie, P.A. Geochemical reactions controlling the solubility of major and trace elements during leaching of municipal solid waste incinerator residues. In: Kilgroe, J. (ed.) Proceedings 1993 Municipal Waste Combustion Conference, Williamsburg, VA. Air and Waste Management Association, Pittsburg, PA, 1993, pp. 667-679. 4. Schramke, J.A. Neutralization of alkaline coal fly ash leachates by CO,(g). Applied Geochemistry, 7 (1992) 48 1-492. 5 . Comans, R.N.J., Van Der Sloot, H.A., Hoede, D. & Bonouvrie, P. Chemical processes at the redox/pH interface during the application of steel slag in the aquatic environment. In: Goumans, J.J.J.M., van der Sloot, H.A. & Aalbers, Th.G. (eds.) Waste Materials in Construction. Elsevier, Amsterdam, 1991, pp. 243-254. 6. Atkins, M., Macphee, D., Kindness, A. & Glasser, F.P. Solubility properties of ternary and quaternary compounds in the CaO-AI,O,-SO,-H,O system. Cement and Concrete Research, 21 (1991) 991-998. 7. Comans, R.N.J., Van Der Sloot, H.A. & Bonouvrie, P.A. Speciatie van contaminanten tijdens uitloging van AVI-bodemas (in Dutch with English abstract). ECN-C--93-090 (1993). 8. Belevi, H., Stampfli, D.M. & Baccini, P. Chemical behaviour of municipal solid waste incinerator bottom ash in monofills. Waste Materials & Research, 10 (1992) 153-167. 1.
110 9.
10.
11.
12.
13.
Mattigod, S.V., Rai, D., Eary, L.E. & Ainsworth, C.C. Geochemical factors controlling the mobilization of inorganic constituents from fossil fuel combustion residues: I. Review of the major elements. Journal of Environmental Quality, 19 (1990) 188-201. Odler, I. and Abdul-Maula, S. Possibilities of quantitative determination of the AFt (ettringite) and AFm (monosulphate) phases in hydrated cement pastes. Cement and Concrete Research, 14 (1984) 133-141. Chandler, A.J., Eighmy, T.T., Hartlen, J., Hjelmar, O., Kosson, D.S., Sawell, S.E., Van Der Sloot, H.A. & Velow, J. Treatise on Municipal Solid Waste Incinerator Residues (in preparation, 1994). Comans, R.N.J. & Middelburg, J.J. Sorption of trace metals on calcite: applicability of the surface precipitation model. Geochim. Cosmochim. Acta, 51 (1987) 2587-2591, van der Hoek, E.E., Bonouvrie, P.A. & Comans, R.N.J. Sorption of As and Se on mineral components of fly ash: relevance for leaching processes. Applied Geochem. (1994, in press).
Environmental Aspects of Conshuction with Waste Materials JJJM Goumans, H A . van der Sloot and l3.G. Aalbers (Editors) @1994Elsevier Science B.V. AN rights reserved.
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Particle Petrogenesis and Speciation of Elements in MSW Incineration Bottom Ashes T.T. Eighmf, J.D. Eusden, Jr.", K Marsellab,J. Hoganb, D. Domingo", J.E. Krzanowski', and D. Stampfli' "Environmental Research Group, Department of Civil Engineering, A1 15 Kingsbury Hall, University of New Hampshire, Durham, N.H. 03824, U.S.A. bGeology Department, Bates College, Lewiston, Maine 04240, U.S.A. 'Mechanical Engineering Department, Kingsbury Hall, University of New Hampshire, Durham, N.H. 03824, U.S.A. Abstract The speciation of elements in municipal solids waste incineration bottom ash is important with respect to its impact on leaching behavior and to its treatment for utilization or disposal. We used a variety of techniques to identify the speciation of major, minor and trace elements in both intact bottom ash particles as well as bottom ash powders. Petrography and scanning electron microscopy/x-ray microanalysis (SEM/XRM) were used to classify intact particles and identify ash particle petrogenic sequences. Two distinct features were seen. Particles are comprised of about 15% of materials present in the MSW waste feed to the incinerator. The remaining portion of the particle (85%) is melt structure. A typical particle contains waste glass (10%) waste soil minerals such as pyroxenes, SiO, (quartz), and feldspars (2%), waste metals and metal alloys (2%). and waste organics (1%). Particles are also comprised of slag or melt products, derived from the MSW feed material, that include opaque glass (25%), isotropic glass (20%). schlieren (lo%), and spinel-group minerals (magnetite, hercynite, chromite) (10%) and melilite group minerals such as C%Al,SiO, (gehlenite) and MgCazSi,07 (akermanite) (20%) which precipitated out of the melt as it cooled. The paragenic sequence is similar to that described for melelite-bearing, igneous rock systems. The system can best be described petrogenically using the CaO-MgO-Al,O,-SiO,-N~O-FeO(CMASNF) system. The melt structure was formed at about 1,200"C. Thermodynamically incompatible phases are present in the ash, making it reactive to aging (oxidation, hydrolysis), weathering, and diagenesis. Increasing the silicon content of the ash could result in the formation of more geochemically stable phases. The residue was ground into powders less than 300 p n in size. Magnetic and density separations were performed to segregate powders for further analysis. The residue is comprised of approximate equal fractions of magnetic, high density; nonmagnetic, low density; and nonmagnetic, high density material. Isodynamic separation of the nonmagnetic fraction was also effective in separating minerals. SEM/XRM of powders fractions in thin section was particularly
112
effective in identifying major minerals in identifiable mineral structures as well as minerals associated with "hot spots" of minor and trace elements. These minerals include many pyroxenes, quartz, feldspars, and melilite-group minerals as well as many spinels. Lead appears to largely be incorporated in complex silicate melt structures. X-ray powder diffraction (XRPD) confirmed the presence of minerals seen by petrography and SEM/XRM. X-ray photoelectron spectroscopy (XPS) of powder surfaces also documented the presence of many of these minerals. A number of oxides and carbonates were also seen with XPS, reflecting the role of O,(g) and CO,(g) in altering the speciation of the particle exterior surface. X P S is particularly well suited for identifying phases associated with leaching at this surface. Solid phases controlling leaching, as determined with the geochemical thermodynamic code MINTEQA2, are not always the same as ones observed with the above mentioned methods. The role of mineral respeciation and diagenesis in controlling leaching is highlighted. The use of such models in predicting leaching behavior is discussed. 1.0 INTRODUCTION
Municipal solid waste incineration is a viable management strategy for treating combustible municipal solid waste that cannot be recycled. Organic material is oxidized, The volume of material is reduced while exothermic energy is recovered. Less volatile inorganic contaminants in the waste feed remain in the bottom ash while more volatile inorganic contaminants are captured in air pollution control devices. The residues from municipal solid waste incineration have been generally characterized with respect to composition and to speciation of some elements (Eighmy et al., 1993). Methods frequently employed involve petrography (Project EKESA, 1992; Vehlow et al., 1992), X-ray powder diffraction (DiPietro et al., 1990; Gardner, 1991; Kirby & Rimstidt, 1993; Ontiveros, 1988; Stampfli, 1992), Fourier transform infrared spectroscopy (Henry et al., 1983), scanning electron microscopy/)
113
speciation. Finally, the ability to understand, model, and predict longer-term leaching behavior requires fundamental knowledge about speciation. The approach taken by our group to characterize speciation of elements in municipal solid waste incineration bottom ash is shown in Figure 1. Elemental composition is determined by combinations of neutron activation analysis (NAA) and HCl/HNO,/HF total digestion coupled with inductively-coupled argon plasma (ICAP) atomic emission spectroscopy. Bulk sample mineralogy is determined by Xray powder diffraction (XRPD) and petrography; the petrography encompassing polarized light microscopy, reflectance microscopy, and scanning electron microscopy/X-ray microanalysis (SEM/XRM) of larger, intact particles. Petrography can also be used to examine particle petrogenesis and mineral paragenesis; the process by which particles form and minerals coalesce from glass phases as the particles cool. Characterization of the near-surface and surface environment is accomplished in particle size-reduced powders with SEM/XRM analysis and X-ray photoelectron spectroscopy (XPS). These methods can provide semi-quantitative compositional information and element associations by dot mapping. Element valence states, bonding environments and speciation are also provided from XF'S. X P S is particularly valuable as a surface analytical technique because its information can be coupled to leaching behavior (Eighmy et al., 1993). All of these methods are used to provide an integrated synopsis about speciation; particularly as it relates to leaching behavior. We employ two leaching procedures and the geochemical thermodynamic model MINTEQA2 to model solid phase control of leaching and compare the controlling solids to the ones observed by the other analytical methods. A fractionation scheme has been developed to separate wastes as a function of magnetic properties and particle density (Figure 2). This is necessary to help concentrate elements present at low concentration in the residue so as to allow for better chance at detection with the above mentioned methods. 300 pm minus particles are separated with a hand magnet and with 1,1,2,2-tetrabromoethane.This produces magnetic and non-magnetic, high and low density fractions. Isodynamic separation is also conducted on the non-magnetic fraction; producing O.SA magnetic, 1.OA magnetic, and 1.5A non-magnetic fractions on bottom ash specimens. The purpose of this study was to examine mineral paragenesis in municipal solid waste incineration bottom ash as well as the speciation of major, minor and trace elements in these residues. This study is one part of a much larger speciation study (Eighmy et al., 1994). The results presented here suggest that the bottom ash is comprised largely of soil grains, waste glass, and waste metals that have partially melted and become components of silicon-poor pyromaphic opaque glass, isotropic glass, schlieren, spinels and melilite group minerals. A number of spectroscopies confirm the presence of these major minerals as well as many trace minerals. Leaching behavior is only partially controlled by these solid phases; indicating the importance of respeciation and diagenesis in the leaching processes.
114 Petrography and SEM/XRM on Parllcles
SEMlXRM on Tola1 Composlllon
Powder Fracllons
XPS on
XRPD on Fracttons
Leachablilty
Figure 1.
Geochemical Modeling of Leachlng Behavlor
Approach used to characterize element speciation in municipal solid waste incineration bottom ash.
*
Powders
Denrlty Separation
Dendty Separation
114
d x
Nan-Magnetlc Non-Magnellc Low Density Hlgh Density
-31.0 Amps
Figure 2.
Magnetic Low Density
+ x
0.5 Amps Nan-Magnetic
Nan-Magnellc
1.5 Amps Magnetlc
0.5 Amps Magnetic
1.0 Amps Magnetlc
1.5 Amps Non-Magnetlc
Sample fractionation scheme.
Magnellc High Density
115
2.0 METHODS
Incinerator Description. The residues collected in this study were from a larger U.S. EPA-sponsored study on the solidification/stabilization treatment of incineration residues (Kosson et al., 1993). The facility is a modem mass bum system with a nominal capacity of 2,000 tonnes per day. The waste feed is primarily household with some commercial and non-hazardous industrial waste. The facility has three parallel trains consisting of a charging chute, primary combustion chamber and moving grates, boiler and economizer, lime slurry spray drier scrubber, and fabric filters. Boiler and economizer ash is collected and added to the bottom ash process stream prior to quenching. The bottom ash also contains grate siftings. As part of a larger EPA project (Kosson et al., 1993), bottom ash was collected on September 14-16, 1989. Full stream cuts from the vibratory conveyor were used to collect residue. The material was screened using a 5.0 cm screen. About 4,000 Kg of screened material was collected. The bottom ash was air-dried, particle-size reduced to pass 1.27 cm screens. A large baffled tank was used to homogenize the material. A dried 20 Kg subsample was sent to the University of New Hampshire for this project. The 20 Kg subsample was cone and quartered to produce a 3 Kg working sample. Processing. The 3 Kg working sample was ground using Proctor hammers and a Weber Brothers & White (Hamilton, Mich.) Model S500 pulverizer (rotary hammer mill). It was possible to grind greater than 60% of the original dry sample to produce a powder that would pass a 300 pm stainless steel sieve. Powders were stored under vacuum desiccation. Fractionation. Magnetic separation and density separation (1,1,2,2tetrabromoethane, TBE, 2.95 g/cm3 at 20°C) were used to produce four powder fractions:magnetic, high density (MHD); magnetic, low density (MLD); non-magnetic high density (NMHD); and non-magnetic low density (NMLD). Additionally, isodynamic separation using a Frantz Isodynamic Separator (S.G. Frantz Co., Trenton, N.J.) with settings of 25" forward tilt and 15" side tilt was used to produce 0.5 amp magnetic ( O S A M ) , 1.0 amp magnetic (LOAM), and 1.5 amp non-magnetic fractions (1.5 ANM). Mass fractions were determined for all the derived fractions. Thin Sections. Intact bottom ash particles in two size ranges (0.4 - 1.27 cm, 0.1-0.4 cm) as well as the powder fractions were made into petrographic thin sections using, a low viscosity epoxy (Epoxy Tech, Billerica, Mass.) and a regimen of high vacuum infiltration and pressure impregnation prior to curing. Thin sections were cut and lap polished with diamond paste prior to mounting. Total Composition. Total elemental composition (mass of element/dry weight of residue) was conducted using either neutron activation analysis (NAA) or total digestion/inductively coupled argon plasma (ICAP) atomic emission spectroscopy. Petrography on Intact Particles. Petrographic analysis, mineral identification, and particle classification of the 0.4 - 1.27 cm and 0.1 - 0.4 cm bottom ash particles in thin section was done using an Olympus BH-2 polarizing light microscope. Field of view magnification ranged from 2Ox to 4OOx. Microfiche copies of the thin sections, magnified to 200%, were used as maps to facilitate the above work. Standard
116
petrographic techniques were used to determine relative percentages of the following: isotropic, clear glass; opaque, black, metallic glass; epoxy-filled vesicles; unfilled vesicles; minerals; microstructural textures; and other miscellaneous ash products. Modal abundances of these features were estimated by comparing the unknown at a certain field of view to published charts for estimating percentage composition of rocks and sediments (Terry and Chilingar, 1955). Minerals were identified using the following standard optical methods and characteristics: pleochroism, birefringence, optic sign, mineral shape or habit, cleavage, relief, color, and extinction position. SEM/XRM on Intact Particles. To better constrain the mineral and ash chemistry, scanning electron microscopy and energy dispersive x-ray microanalysis (SEM/XRM) was done on the thin sections. The thin sections were carbon coated using a Denton Vacuum Desk-1 carbon evaporation unit. A JEOL JSM-6100 Scanning Electron Microscope equipped with a Kevex Super-Dry detector and a Kevex Delta Class IV analyzer was used for the SEM/XRM analyses. SEM/XRM on Powder Fractions. A Link AN-10,OOO energy dispersive XRM workstation was used in conjunction with a Pentafet atmosphere-thin window detector (lithium-drifted silicon crystal) and an AMR 1OOO SEM. XRPD on Powder Fractions. A Rigaku-Geigefflex goniometer was used for XRPD. A copper tube (44 kV, 34 mA, 1500 W) was employed as an X-ray source. A divergence slit of 1". a scattering slit of lo, a receiving slit (crystal) of 0.8", a receiving slit (monochromator) of 0.6" produced the best peak definition and lowest background. Powders were run in triplicate; double sided tape was used to hold the samples. A scan rate of 0.5" 28 per minute was used. Scans were conducted from 6.00 to 9.00". Most peak intensities were measured at 1,OOO to 10,OOO cps. Tungsten internal standards were employed. The goniometer output was directed to a Spectraphysics integrator (SP 4270) to precisely determine peak location and area. A PC-based search match program, MICRO-ID, was used to identify possible crystalline phases in the residues. The search procedure involved element exclusion based on NAA data, a 0.5" 28 error window, three lines of match, a minimum relative intensity of 1, and calculation of figure of merit based on weightings of 90% for peak location and 10% for peak intensity. Searches were based on PDF sets 1-41. XRPD analyses were also conducted on the leached 300 pm-minus unfractionated material from the total availability leaching test (see below). XPS on Powder Fractions. A Perkin Elmer Physical Electronics Division 5100 hybrid X P S was used to conduct quantitation and chemical shift analysis work. Samples were either packed into tiny, planchette-mounted cups or pressed onto the sticky-side of copper tape. An Art ion beam was used to etch the particle surfaces; this was done to remove adventitious carbon and oxygen. Sputter etching was usually for 3 minutes with a 4 kV beam voltage to the ion gun. A procedure was developed to initially scan the sample under low resolution from 1,100 to 0 eV, then high resolution scan the carbon 1s and silicon 2p3/2 windows for energy referencing purposes. The sample was then etched and rescanned. Finally, high resolution scans of the windows of elements of interest was done for a 2 hour total acquisition time. Adventitious carbon was used for static charge referencing to conduct charge
117
corrections. The NIST X P S database (version 1.0) was used to identify speciation from binding energy chemical shifts of principal photoelectrons. A tolerance of 0.2 eV was used. X P S analyses were also conducted on the leached 300 pm-minus unfractionated material from the total availability leaching test (see below). Leaching Tests and Leaching Modelling. Two leaching procedures were used to examine leaching behavior and solid phase control of leaching. The first procedure, based on the Dutch total availability leaching procedure [NVN 25081, was used to quantify the element mass fraction available for leaching. A sequential extraction is used; the first is conducted at pH 7 for 3h at a liquid to solid (L/S) ratio of 100 (L/Kg). 8.0g of <300 pm unfractionated bottom ash powder is added to 800 ml of Milli-Q@type 11 water and stirred in an open beaker. pH is controlled using 6N HNO,. At the end of the three hours, the leachant is filtered (0.45 pm polycarbonate filter) and the residue and filter are leached again at L/S 100 at pH 4.0 for 4h. 6N HNO, is again used to control pH. After filtration, the pH 7.0 and pH 4.0 leachants are combined and analyzed for dissolved constituents using ion chromatography (IC) or ICAP. The leached residues are dried, stored under vacumn and saved for XRPD and X P S . The pH 7.0 extraction solubilizes oxyanions. The pH 4.0 extraction solubilize cations. Both fractions, when compared to total composition, estimate the fraction that is environmentally available for leaching in a geologic timeframe. The second procedure is a pH-dependent procedure designed to examine leaching over a large pH range. A pH range of 2 to 14 with 2 pH unit increments is used. Each extraction is done at an L/S of 6.0. 65g of <300 um unfractionated bottom ash powder is placed in 500 mL plastic bottles. Milli-Q water and either HNO,, NaOH, or KOH are used to set pH values to prescribed levels. The bottles are placed on rotary tumblers and agitated for 48 h. After extraction, the leachants were filtered (0.45 pm polycarbonate filter) and quantified using IC or ICP. The geochemical thermodynamic model MINTEQA2 (Allison et al., 1990) was used to examine dissolution behavior and solid phase control of leaching as a function of pH. 3.0 RESULTS AND DISCUSSION Total Composition. The major constituents (> 10,000 mg/Kg) in the ash are Si, Fe, Ca, Al, Na, and C1. Although not qualified by NAA or total digestion/ICAP, oxygen is the most predominant element (ca. 400,000 mg/Kg) in the residue. Minor constituents (l,OOO-lO,OOOmg/Kg) include Mg, K, Ti, Zn, Cr, Pb, Mn, and Ni. Trace constituents (< 1,OOO mg/Kg) include up to 31 elements. These concentrations are typical for MSW bottom ash (Kirby and Rimstidt, 1993; Kosson et al., 1993). Fractionation. Figure 3 provides information on the mass distribution of the powdered bottom ash in the various fractions. The principal fractions from magnetic and density separation are MHD (26.25 wt %), NMHD (8.86 wt %), and NMLD (62.11 wt %). The principal fractions from isodynamic separation are magnetic (28.11 wt %), 0.5 AM (38.94 wt %), 1.0 AM (9.21 wt %), and 1.5 ANM (20.02 wt %).
118 USBA ferrous < 2.97 glcm’ 1.681
Deonty .%pimuon
-
,
USBA magnetic
USBA ferrous >2.97 gtcrn’ 26.25%
28.11% DeDsity Separation
USBA nonferrous <2.97 glcm’
Density Scpuuion
Mapodic sBp.rscion
USBA nonmagnetic
62.11%
71.14%
USBA nonferrous >2.97 gtcm’
tknsity Sepmtion
8.86%
magnetic at 0.5 amps
9.21% USBA magnetic at 1.5 amps
Magnetic Scpuuion
1.84%
I Figure 3.
I
USBA
Fractionation Data for the Bottom Ash from a U.S.Mass Bum Combustor.
I
119
Particle Classification. The municipal waste incinerator bottom ash intact particles were classified into two major groups: non-combusted waste products from the MSW feed and melt products. Non-combusted waste products consist of four principal types: waste glass; waste minerals: waste metals; and waste organics. These represent fragments of bottles, glassware, metals, cans, soils, etc. that did not or could not combust at the temperatures (and time period) they were exposed to in the incinerator. Melt products are defined as products resulting from the melting and partial melting of the waste and the minerals that crystallized from the melt. Five different types of melt products were recognized: opaque, metallic glass; isotropic silicate glass: schlieren: spinel-group minerals: and melilite-group minerals. In general 85% of the bottom ash is composed of melt products and 15% noncombusted waste products. Figure 4 shows the estimated modal abundances of the principal types of the non-combusted waste products and melt products. A common feature of all particles are vesicles or gas bubbles, presumably formed during boiling of the melt products. Vesicles comprise between 10 - 25% of the total volume of the particles, therefore providing for a significant measure of particle porosity.
Men Produd Melilile Group Mineral 20.0%
Men Produd Spinel Group Minerals 10.0%
Schlloren 10.0%
W u l a Mlnwala 20% Was10 Meld; 20% W a l e Organla 1.0% lrolroplc G l a u 20.0%
Opaque Glass 25.0%
Figure 4.
Pie graph showing percentages of non-combusted waste products and melt products in municipal waste bottom ash.
SEM/XRM Chemistry of Melilite-Group and Spinel-Group Minerals in Intact Particles. In order to better characterize the melt products, XRM spectra were acquired and representative mineral formula determined for the melilite-group and
120
spinel-group minerals. Since mineral identification using transmitted light techniques was quite successful, the XRM analyses simply confirmed the mineral identifications and provided chemical information useful for chemically subdividng the various mineral types within the melilite and spinel groups. Melilite group minerals included two general types of chemistry: gehlenite (CagU,SiO,), and akermanite (Ca, (Mg, Fe2+)Si,0,). Sodium is common in small amounts in gehlenite and is incompatible with iron and magnesium. When magnesium and iron are detected in akermanite, sodium is absent. The results of the melilite group chemical analysis agree with the description of naturally occurring melilite-group mineral chemistry described by a host of authors and summarized succinctly by Deer et al. (1992). Spinel-group minerals included three general types of chemistry: spinel-hercynite ((Mg,Fe)Al,O,), magnetite (Fe,O,), and chromite (FeCr,O,). A majority of the spectra have appreciable silicon. This is attributed to: 1) interactions of the electron beam with not only the spinel-group mineral but also the matrix silicate glass surrounding the mineral, and 2) minute inclusions of matrix silicate glass within the crystal structure due to rapid nucleation. The results of the spinel-group chemical analysis agree with the description of naturally occurring spinel-group mineral chemistry described by Deer et al. (1992) and a more exhaustive treatise by Rumble (1976). Related Natural and Experimental Ingenous Systems. Based on the known conditions of incineration and the melt products described for this study, it is believed that the incinerator paragenetic sequence is similar to that described for melilitebearing, igneous rock systems observed in both naturally occurring and experimentally produced associations (McBirney, 1993; Yoder, 1979; Soulard et al., 1992). With the exception of FeO, the CaO-MgO-Al,O,-SiO,-N~O (CMASN) experimental system described by Soulard et al. (1992) is quite similar in composition to the melt products discussed in this study. Furthermore, the pressure (1 bar) and temperatures of the experimental system are close to those within the incinerator. This system also provides a series of phase relationships many of which are consistent with the petrographic observations of the melt products. It more adequately describes the crystalline mineral phases present than, for instance, the CaO-Al,O,SiO, (CAS) system used by Lichtesteiger and Zeltner (Project EKESA, 1992) to estimate incinerator temperatures. The temperatures of the experimental runs for the study by Soulard et al. (1992) range from 1,238" to 1,143"C. Because of the presence of iron-bearing melt products, a better system to describe the incineration paragenesis for this study would be the CaO-MgO-Al,O,-SiO,-Na,OFeO (CMASNF) system. Though no such natural or experimental igneous system was discovered in the literature (it may well exist), the additional iron component would, based on thermodynamic arguments, presumably lower the melting temperature of the melt products. Therefore, using the values of Soulard et al. (1992) as a maximum, it is estimated that the melt reached temperatures no greater than approximately 1,200"C.
121
The melting and melt products within the incinerator are probably in a state of local equilibrium and the waste products are analogous to igneous xenoliths; blocks of foreign rock within in a magma. They are preserved due to the lack of complete melting and therefore assimilation into the melt. The common association of melilte and quartz in other ash residue studies (Vehlow et al., 1992; Project ESKESA, 1992) has also been mentioned as an indicator of complex heterogeneous, disequilibrium combustion and melting. It is indeed not surprising that quartz and waste glass survive incineration as their silica content make them highly refractory. Disequilibrium is also indicated by the mere presence of the opaque, metallic glass and the isotropic silicate glass. Glass is inherently at disequilibrium or in a metastable state since it is an amorphous solid lacking the arrangement of atoms into the crystallograhic state (Carmichael, 1979). Glass is formed through a two step process where a stable liquid is supercooled to finally produce a glass. The equilibrium temperature of crystallization of a liquid is higher than the glass transformation temperature. For example, Carmichael (1979) shows that the temperature of equilibrium crystallization of feldspathic liquids (producing sodic and potassium feldspars) is between 1,ooo" and 1,500"C, whereas supercooling of this liquid to produce glass occurs between 750" and 900°C. One can speculate that by varying the chemical components of the melt products and/or the external controls within the incinerator (pressure and/or temperature) a significantly different ash residue would result. For example, the melt products could be made silica-oversaturated by either completely incorporating the available waste glass, waste quartz, and waste feldspars present in the MSW feed, or by adding additional SiO, to the waste stream. In either case, the resulting mineralogy would shift to more typical silica-oversaturated melt products similar to naturally and experimentally studied granitic systems. These rocks typically have a mineralogy of quartz, plagioclase feldspar, and potassium feldspar. This might be environmentally advantageous as granitic rocks are less likely to react strongly with the atmosphere and surface or groundwaters. The tendency for alkaline, mafic, silica-undersaturated rocks (such as the incinerator residue described here) to chemically alter after crystallization is much greater and accounts for its propensity to weather and undergo diagenic reactions. SEM/XRM on Powder Fractions. Table 1 summarizes the results from SEM/XRM evaluation of powder fractions. The principal mineral phases identified with petrographic analyses of intact particles were also identified with SEM/XRM analysis of powder fractions. This minerals include ackermanite (MgC+Si,O,), gehlenite (C+Al,Si,O,), magnetite (Fe,O,), and hercynite (FeA1204). SEM/XRM analyses of powder fractions also revealed many waste soil minerals, metal alloys, and spinels. Lead was present in complex silicate melts and as lead oxide (PbO). Like petrography, SEM/XRM revealed that elements with reduced valencies are present in the material. The data also suggest some success in recovering minerals according to their magnetic properties and density. Some caution is needed in the absolute interpretation of the results because of problems associated with beam diameter and particle size as well as the confounding effects of particle location with depth in the
122
thin section. Nevertheless, SEM/XRM was particularly useful in probing mineral structures and "hot spots". Hot spots were usually regions where elements present in low concentration in the residue (e.g. Cr,Pb, Zn, etc.) were revealed with dot maps. The particle region containing the element of interest could then be probed at higher magnification. This allowed estimation of mineral formulas for minerals present at low concentration in the ash which might not be detectable by other methods. Table 1 SEM/XRM-Based Mineralogy of Ash Fractions
Mu2 MgCa#i,O (Akermanite) (Ca,Na)(Fe'5 +,Mn,Zn)Si,O, (Aegirine) Fe,O, (Magnetite) FeA1204(Hercynite) FeO (Wustite) ZnFe204(Franklinite) FeS (Pyrite) NaAlSi,O, (Albite) Fe,TiO, (Ulvaspinel) -Si,O, (spinels)
LQAM CaMg(CO,), (Dolomite) C5s(P04), (Whitlockite) SiO, (Quartz) NaA1Si30, (Albite) CaSI,O, NaAlSi,O, (Sodian Anorthite)
NMHD FeO (Wustite) MgCO, (Magnesite) Al,O, (Corundum) SiO, (Quartz) C5s(P04), (Whitlakite) Fe,04 (Magnetite) ZnO (Zincite) MgCaSi,O, TiO, (Rutile) ZnC1, (Corrunite) CaMgSi,O, (Diopside) PbO MgCa$i,O, (Ackermanite) Fe,Ni (Taenite) cu alloy Al alloy Lead Silicates
NMLD M@2°4
BaSO, (Barite) CaSO, (Anhydrite) CaHPO, Ca&,Si,O, (Gehlenite) SiO, (quartz) C (graphitic carbon)
LuuW NaA1Si30, (Sodian Anorthite) C%(PO,), (Whitlockite) CaMg(CO,), (Dolomite) AlO
XRPD on Powder Fractions. Table 2 summarizes the results from XRPD. The principal minerals identified with XRPD were quartz (SiO,), halite (NaCl), calcite (CaCO,), hematite (Fe,O,), taenite (Fe, Ni), zincite (ZnO), nantokite (CuCl),
123
Table 2 XRPD-Based Mineralogy of Ash Fractions NMHD
NMLD
Mm2
A1203(Corundum) ZnO (Zincite) Mg3A12si3012 CuCl (Nantokite) FeS (Galena) Sb,Sr
CaCO, (Calcite) SiO, (Quartz)
SiO, (Quartz) NiTiO, Fe203(Hematite) (Fe, Ni) (Taenite) Ca(Mg,AJ)(Si,Al),O, (Aluminian Diopside) C (Graphite) CaMg(SiO,), (Diopside)
NaCl (Halite) SiO, (Quartz) (Ca,Na)(Si,Al)& (Fe,Ni) (Taenite) K,HCr,AsO,,
CaCO, (Calcite) C (Graphite) NaCl (Halite) SiO, (Quartz) Ba,Sb, C%Na4(P04F)602 A&P,S6
SiO, (Quartz) CaCO, (Calcite) KNaPb8(P04)6 Rb,Fe(SO,),O
Other Ca,SiO, 0.3H20 MgSiO, NaAlSi,O, (Albite) NaAlSiO, (Nepheline) CuMgSi,O,j MgMnSi,O, (Donpeacorite) CaS04-2H,0 (Gypsum) BaSO K2CaMg(SO4)3
Pb2Si04 Pb,SiO, Pb,Si,O, Pb5Si3Oll KPb8NdF2(P04)6 Zn NaCuPO,
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corundum (A1203), graphite (C), and diopside (CaMg (Si03)2). Also present were albite (NaAlSi,O,), nepheline (NaAlSiO,), and gypsum (CaSO,ZH,O). Many minerals from the pyroxene class were seen as were calcium and magnesium silicates. Interestingly, a number of lead silicates were observed. Many of these minerals were seen with petrographic methods and SEM/XRM on intact particles. Elemental Zn was also seen. A few minerals were identified that suggest the residues were not formed under completely oxidizing conditions. Such minerals contain reduced elements and include nantokite (CuCl), galena (FeS), and graphite (C). It is possible to have geochemically incompatible phases in bottom ash particles given the mineral paragenesis that occurs. The minerals identified here are similar to ones observed by Kirby and Rimstidt (1993) and Sthpfli (1993) in their extensive XRPD analyses of bottom ashes from MSW incineration. Triplicate XRPD analyses of unfractionated, leached bottom ash powders subjected to the total availability leaching test found the following minerals: SiO, (quartz), Fe,Ni (taenite), C (graphitic carbon), MgSO,QH,O, CaMgSi,O, (diopside), Ca(Mg,Fe,Al)(Al,Si),06 (augite), and Na4A12Si,0,. XPS on Powder Fractions. The speciation of elements in the various powder fractions as determined by XPS are shown in Table 3. XPS is a promising technique for determining element speciation. This is particularly true because of its ability to interrogate the near surface environment (e.g. lo&. This surface layer is the first layer to dissolve, sorb or desorb during leaching. While the NIST database is comprehensive, it does not contain extensive mineralogies for elements of interest (e.g. Pb, Cr, Zn, Si) that are typically found in geologic specimens or combustion residues. The most prevalent minerals identified by X P S were quartz (SiO,), calcite (CaCO,), mullite (Al,SiO,), halite (NaCl), gibbsite (Al(OH),), corundum (A1203), zincite (ZnO), hematite (Fe203),and anhydrite (CaSO,). These species were also seen in the leached ash. The method also routinely identified other oxides such as CaO and FeO. These phases, not routinely seen by petrography, SEM/XRM or XRPD, most likely represent surface oxide layers on calcium-and iron-bearing minerals. It is important to note that magnetite (Fe,O,), a bulk mineral of some predominance that was seen with petrography, SEM/XRM, and XRPD, was not seen with XPS. It is possible that at the mineral surface, under oxidizing conditions, the more oxidized hematite (Fe,O,) forms. XPS is a surface sensitive technique and would tend to identify surface species rather than bulk (interior) species. The method also identified a number of carbon species including halogenated hydrocarbons, hydrocarbons, carbides, and calcite (CaCO,). Though not seen with other methods, calcite may be an amorphous surface layer on particles. Graphite, seen with SEM/XRM and XRPD, was not seen with XPS.
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Table 3 XPS-Based Mineralogy of Ash Fractions
M m SiO, (Quartz) A12Si05(Mullite) A12Si,0,*2H,0 CaO NaAlSi,O, (Albite) Fez03
QLW SiO, (Quartz) Al(OH), (Gibbsite) Al,O, (Corundum) A12Si05(Mullite) NaAlSi,O, (Albite) CaO NaCl (Halite) CaSO, (Anhydrite) Fe203(Hematite) ZnSO, ZnCl,
NMHD SiO, (Quartz) Al(OH), (Gibbsite) CaCO, (Calcite) Fe,O, (Hematite) NaCl (Halite) Ca,Si,09 Al,O, (Corundum) CaO ZnC1, ZnSO, NaAlSi,O, (Analcite) Carbides
NMLD SO, (Quartz) A1203(Corundum) NaCl (Halite) NaAlSi,O, (Albite) CaSO, (Anhydrite) ZnSO, ZnC1, Fe,O, (Hematite) Carbides
1.5 ANM SiO, (Quartz) Al,O, (Corundum) Al,Si,O7*2H,O NaAlSi,O, (Albite) CaO NaCl
Leaching Behavior. The data from the total availability leaching test are shown in Figure 5. Generally, about 11.4% of mass of bottom ash was lost during the leaching test. Principal constituents that readily dissolve are Ca, Cd, C1, Cu,K,Mg, Na, Pb, and Zn. The data from the pH-dependent leaching tests are shown in Figures 6-11. Figure 6 presents data for Na, K and Cl. These elements show more or less pH-independent dissolution across the entire pH range. Figure 7 shows data for Ca, SO,, Al and Si. Each of these elements shows pH-dependent dissolution phenomena and solid phase control. A clear minimum is seen for Al. Figure 8 shows the leaching data for Cd, Pb, Zn and Cu. All of these elements exhibit pH dependent leaching. Figure 9 shows pH dependent leaching of Fe, Mn,Sr and Mg. Figure 10 shows similar data
126
Bottom Ash Fraction Available for Leaching 140 120 c C
$100 Q,
a
ai 380 .m
2
6 60 .-c 0 2
LL
40
20
0
Figure 5.
Al Ca Cd CI Cr Cu Fe K Mg Mn Na Ni Pb Si Zn
Bottom Ash Element Fractions Available for Leaching.
127
10000
2
-.-F 1000 c
v K versus pH
Na versus pH
" i TIFOOO Y
0
c
E g
c
100
0
c
0
0
10
0 2 4 6 8 101214
0 2 4 6 8 101214
PH
PH
CI versus pH 10000
7
c
.-0 w L
1000
i
L
c
8 c
0
0
100
0 2 4 6 8 101214
PH
Figure 6.
pH Dependent Leaching Behavior for Na,K and C1.
128
SO4 versus pH
Ca versus pH $0000
I
10000
2
c
c
8 c
1000
0
0
100
.-c 0 c
-
E
U
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1
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\
1000
-
-
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8 c
0
0
I
100
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I
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I
0 2 4 6 8 101214
PH
Al versus pH
10000.0
10000.00 s l 1000.00 E 100.00 C .2 10.00
\
E ti
c 8
1.00 0.10
8
0.01
1000.0 c .G CI
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*-
0.00
Figure 7.
Si versus pH
I
I
I
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.‘ I
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I SiO, I (A,Pt
100.0 10.0 1.0 0.1
0 2 4 6 8 101214
0 2 4 6 8 101214
PH
PH
pH Dependent Leaching Behavior for Ca, SO,, Al and Si. The horizontal lines inside the graphs are detection limits. The dotted lines show solubility curves for the indicated controlling solid phase.
129
Cd versus pH 10.000
gE
Pb versus pH 1000.00
2
E 100.00
1.000
Y
.-c5. 10.00
Y
e
c
c
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0
0
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0
--...Anglesite
0.01
0 2 4 6 8 101214
0 2 4 6 8 101214
PH
PH
Zn versus pH
Cu versus pH
10.00
1 .oo ~
0.10
0 2 4 6 8 101214 PH
Figure 8.
, ,
h..;. ,
0.01 0 2 4 6 8 101214 PH
pH Dependent Leaching Behavior for Cd, Pb, Zn and Cu. The horizontal lines inside the graphs are detection limits. The dotted lines show solubility curves for the indicated controlling solid phase.
130
Fe versus pH
Mn versus pH 1000.00
Ferrihydrite
7100.00 E
Y Y
.-g 10.00 U
gc 8 c 0
0
0.01
0 2 4 6 8 101214
PH
PH
Sr versus pH
Mg versus pH 10000.0
2 1000.0
i
F
Y
Y
s
c
s
0.10
0 2 4 6 8101214
100.0
0 .= e c. C 8 C
1.00
100.0 .c
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gc 8 c
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Figure 9.
I
I
I
I
I
I
00
l~
~~
10.0 1.0
-
*
- -Magnesite
0.1
0 2 4 6 8 101214
0 2 4 6 8 101214
PH
PH
pH Dependent Leaching Behavior for Fe, Mn,Sr, and Si. The horizontal lines inside the graphs are detection limits. The dotted lines show solubility curves for the indicated controlling solid phase.
131
-
Cr versus pH
100.00
100.00 i 0, t u c 0 '= 10.00
b 4
0)
E !
10.00
C
0 .-
5
1.00
e
c
+
c
C
a,
g 0
B versus pH
a,
0.10
u
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0
0
0.01
1.oo
0 2 4 6 8 101214
PH
Ni versus pH
Co versus pH
-1 00.00
10.00
2
0)
F
&E10.00
Y
C
C
0 .c
E
1.00
z s
0.10
c
8 101214
PH
\
0 .-c
0 2 4 6
1.00
e
+
8 0.10
C
C
s
0.01 0 2 4 6
8 101214
PH
Figure 10.
0.01
-
0 2 4 6 8101214
PH
pH Dependent Leaching Behavior for Cr, B, Ni and Co. The horizontal lines inside the graphs are detection limits. The dotted lines show solubility curves for the indicated controlling solid phase.
132
Ba versus pH
Mo versus pH
-i
1000.003 -
10.0
F
Y
g
1
..._.Barite 4
F
2 100.00 F c
.-0
1.0
5 + c 8 c
g!
c)
c
8 C 0
0 0
0.1
1.00 0.10 1 0.01 °
.
O
O
0 2 4 6 8 101214
0 2 4 6 8101214
PH
PH
V versus pH
-i
10.000
F 1.000
Y
c
0
-= 0.100 ! 4-
c
g 0.010
s
0.001
0 2 4 6 8 101214
PH
Figure 11.
pH Dependent Leaching Behavior for Mo, Ba and V. The horizontal lines inside the graphs are detection limits. The dotted lines show solubility curves for the indicated controlling solid phases.
~
133
for Cr, B, Ni and Co. A minimum is seen for Cr. Figure 11 shows data for Mo, Ba, and V. The nature of phases controlling solubility is discussed below. Geochemical Modeling. Only one approach was used to model leaching behavior in the bottom ash. The data from the pH dependent tests was used. The approach assumes that equilibrium-based solid phase control of leachate chemistry occurs over a wide pH range. This solid phase control approach was viewed as an appropriate use of MINTEQA2 for the type of leaching test that was used (low L/S, solid phase in equilibrium with the solution, solid phase dominates). At each pH, all leachate constituents were inputted into the model (including acid or base needed to adjust the pH). Each condition was run on MINTEQA2 without allowing solids to precipitate. Minerals with either positive or negative saturation indices close to zero were identified. A select group of 13 minerals was then inputted into the files at each pH as infinite solids (i.e. solids that will not precipitate or dissolve but which will control its aqueous phase constituents in the presence of all other solid and aqueous phase reactions and conditions). The output of the modeling was then plotted, where appropriate, in Figures 6-11. The premise here is that a solid with a saturation index close to zero is the most likely solid controlling the composition of the leachate at that pH. By reintroducing the solids into the model as infinite solids, the specified aqueous phase concentration of a component species should be close to the observed values and the solubility plot should be close to or mimic the behavior of the actual leaching data. As shown in Figure 6, no solids were identified as controlling leaching behavior for Na, K and C1. These elements demonstrate pH-independent dissolution and most likely represent the dissolution of NaCl, ZnCl,, and unidentified potassium salts such as KCl. As shown in Figure 7, Ca and SO, solubility apper to be controlled by anhydrite (CaSO,). A slightly better fit was seen for anhydrite over gypsum (CaS0,*2H20). The fit was most appropriate between pH 4 and 10. Anhydrite also controls leaching of Ca and SO, in ESP ash (Theis et al., 1993; Eighmy et al., 1994). Comans et al. (1993) found gypsum to control leaching in their MSW bottom ash. An excellent match was found for Al solubility. Gibbsite [Al(OH),] is the controlling phase for this bottom ash, ESP ash (Eighmy et al., 1994) and Dutch bottom ash (Comans et al., 1993). No clear controlling solid could be identified for Si. Amorphous SiOl was used as it most closely resembles the type of silicate minerals seen in the residue; however, the match is not a good one. Comans et al. (1993) identified wairakite (CaAI,Si4O,,*2H2O) as a controlling solid for their ash; this mineral did not provide a good match in our residue and it was not identified in the residues with the various speciation methods that were employed. As shown in Figure 8, Cd solubility was found to be controlled by otavite (CdCO,); particularly at higher pH values. A corrected p q P of 12.24 was used (Zachara et al., 1992). This solid also controls Cd solubility in ESP ash (Eighmy et al., 1994) and in Dutch bottom ash (Comans et al; 1993). Pb solubility appeas to be controlled by anglesite (PbSO,), although the data is sparse. This solid also controls Pb leaching in ESP ash (Theis et al., 1993; Eighmy et al., 1994). Comans et al.
134
(1993) found Pb(OH), to control in their bottom ash. Zn solubility is controlled by ZnSiO,; this was also seen for various ashes by Gardner (1991). Cu solubility is controlled by Cu(OH),. As shown in Figure 9, Fe solubility is controlled by ferrihydrite (Fe(OH),). This solid also controls in ESP ash (Eighmy et al., 1994) and in Dutch bottom ash (Comans et al., 1993). Sr solubility could not be modeled adequately. Celestite (SrSO,) gave the closest match; this solid does control Sr leaching in ESP ash (Eighmy et al., 1994). Mg leaching appears to be controlled by magnesite (MgCO,); particularly at higher pH values. No suitable matches were found for Mn. As shown in Figure 10, no suitable matches were found for Cr, B, Ni, and Co. B is not in the MINTEQA2 data base as a component. The solids present in the data base for Cr, Ni and Co are limited. As shown in Figure 11, Ba solubility is controlled by barite (BaSO,). V solubility may be controlled by Pb,V,O, although the match is not that good. Mo is not in the MINTEQA2 data base as a component. Some of the data presented here suggest that XPS is the preferred method to verify the presence of solids that control pH-dependent leaching behavior (i.e. Al(OH),, CaSO,). Data also suggest that some soluble solids dissolve and new solids form under aqueous conditions that ultimately control leaching (i.e. BaSO,, MgCO,, Fe(OH),, Cu(OH),). Such diagenic transformations indicate that respeciation occurs during leaching and that simple carbonate and hydroxide solids form and therefore control leaching of their constituent cations. The fact that much of the pH-dependent leaching behavior can be described by geochemical thermodynamic models means that the use of such models to predict long term leaching behavior, leaching after treatment, or leaching during specific disposals scenarios is possible. Such efforts will need to be fine-tuned, however, by expanding the thermodynamic database in the models as well as incorporating sorption phenomena. 4.0 CONCLUSIONS Bottom ash particles from MSW combustion is principally comprised of a melt structure (85%) and an MSW uncombusted or unmelted waste feed component (15%). The melt structure, formed from waste feed components, is comprised of opaque glass, isotropic glass and schlieren where spinel-group minerals (Fe,O,, FeAl,O,, CrFe,O,) and melilite-group minerals (Ca&SiO,, MgCa,Si,O,) precipitated out in the melt as it cooled. The waste feed components contain soilderived SiO,, pyroxenes, feldspars and uncombusted organic carbon. An analogous paragenic sequence for the formation of melt is a melilite-bearing igneous rock described by the Ca0-Mg0-A120,-Si0,-N~0-Fe0(CMASNF) system. The melt structure was formed at about 1,200"C. Thermodynamically incompatible phases are present in the ash, making it reactive to weathering, aging (oxidation, hydrolysis), and diagenesis. Many species in the residue were defined by SEM/XRM, XRPD, and X P S evaluations of powder fractions. Species were documented for 0, Si, Fe, Ca, Al,
135
Na, C1, Mg, Ti, Zn, Cr, Pb, and Cu. Some of these species are present at trace levels ( c 1,OOO mg/Kg). All the methods provided comparable results. X P S , however, identified mineral species at the particle surface most involved in leaching behavior. The residues exhibit pH-dependent leaching behavior. Solid phases controlling leaching sometimes differ from those seen in the unleached residues; suggesting dissolution and reprecipitation reactions are occurring that control pH dependent leaching. Overall, bottom ash is not readily leachable. 5.0 ACKNOWLEDGEMENTS
This research was supported by the U.S. EPA Risk Reduction Engineering Laboratory (CR 818157-Ol), U.S. EPA Region 1 (X1001525-01),Chemical Waste Management Inc., and Environment Canada. We thank Trish Erickson, Jim Ryan, Carlton Wiles, Juan Perez, Jesse Conner, Bill Holm, David Hay, Gordon Owen, Steven Sawell and John Martin for their assistance in this effort. 6.0 REFERENCES
Allison, J.D., Brown, D.S. & Novo-Gradac, K.J. (1990) MINTEQA2/PRODEFA2,A Geochemical Assessment Model for Environmental Systems: Version 3.0 User's Manual. Environmental Research Laboratory, U.S. EPA, Athens, Georgia. Carmichael, I.S.E. (1979) Glass and the glassy rocks, p. 233-242. In (H.S. Yoder, ed.) The Evolution of the Igneous Rock: Fiftieth Anniversary Perspective. Princeton University Press, Princeton, New Jersey. Comans, R.N.J., van der Sloot, H.A. & Bonouvrie, P. (1993) Geochemical Reactions Controlling the Solubility of Major and Trace Elements During Leaching of Municipal Solid Waste Incinerator Residues, p. 667-679. In (J. Kilgore, ed.) Proceedings 1993 International Conference on Municipal Waste Combustion, AWMA, Pittsburgh, Penn. Deer, W., Howie, R. & Zussman, J. (1992) An Introduction to the Rock Forming Minerals. Longman Scientific and Technical, Essex, England, 696 p. DiPietro, J.V., Collins, M.R., Guay, M. & Eighmy, T.T. (1990) Evaluation of pH and oxidation-reduction potential on leachability of municipal solid waste incineration residues. p. 2B:21-33. In Proceedings of the International Conference on Municipal Warte Combustion, April 11-14, Hollywood, Florida, AWMA, Pittsburgh, Penn. Eighmy, T.T., Domingo, D., Stampfli, D., Krzanowski, J., & Eusden, J.D. (1993) The Speciation of Elements in MSW Combustion Residues, p. 457-470. In (J. Kilgroe, ed.) Proceedings of the 1993 International Conference on Municipal Waste Combustion Conference; AWMA, Pittsburgh, Penn. Eighmy, T.T., Krzanowski, J., Domingo, D., Sthpfli, D., Eusden, D., Marsella, K, Killeen, K, Guardenier, H. & Hogan, J. (1994) The Nature of Lead., Cadmium and
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Other Elements in Incineration Residues and Their Stabilued products. U.S. EPA Final Report (in press). Gardner, K.H. (1991) Characterirarion of Leachates From Municipal Incinerator Ash Maten&. Master's Thesis, Clarkson University, p. 112. Henry, W.M., Barbour, R.L., Jakobsen, R.J. & Schumacher, P.M. (1983) Inorganic Compound Ident$cation of Fly Ash Emissiom From Municipal Incinerators, EPA600/53-83-095, U.S. EPA, Cincinnati, Ohio. Kirby, C.S. & Rimstidt, J.D. (1993) Mineralogy and surface chemistry of municipal solid waste ash. Environ. Sci TechnoL, 27, 652-660. Kosson, D.S., Kosson, T.T. & van der Sloot, H. (1993) Evaluation of Solid$cation/Stabiluation Treatment Proceses for Municipal Waste Combustion Residues. USEPA/RREL Final Report. McBirney, A.R. (1993) Igneous Petrology. Jones and Bartlett Publishers, Boston, 508
P.
Ontiveros, J.-L. (1988) A Comparison of the Composition and Properties of Municipal Solid Waste Incinerator Ashes Based on Incinerator Configuration and Operation. Ph.D. Dissertation, Rutgers University, 219 p. Pliiss, A. & Ferrell, R.E. Jr. (1991) Characteristics of lead and other heavy metals in fly ash from municipal waste incinerators. Haz. Waste Haz. Mad, 8, 275-282. Project EKESA. (1992) Emksiomabschatnmgfur kehrictschlacke: Aufrtraggebergemeinschafi f u r das Projekt EKESA, Eidgenossische Anstalt fur Wasserversorgung, Abwasserreinigung, und Gewasserschutz (EAWAG), Dubendorf, Switzerland, 159 p. Rumble, D. (1976) Oxide Miner&: Mineralogical Society of America Short Course Notes, vol. 3, Mineralogical Society of America, Washington, D. C.. Soulard, H., Provost, A, & Boivin, P. (1992) Ca0-Mg0-Al,0,-Si02-N~0(CMASN) at 1 bar from low to high N%O contents: Topology of an analogue for alkaline basic rocks: Chem. Geol. 96,459-477. Stiimpfli, D. (1992) Final Report: Cements and Bottom Ash Chemistry (ClaAC) - A Study Wth X-RayPowder Difiaction. Environmental Research Group Final Report, University of New Hampshire, 69 p. Terry, R.D., & Chilingar, G.V. (1955) Summary of "Concerning some additional aids in studying sedimentary formations" by M.S. Shvetsov. J. Sed Pet. 25:229-234. Theis, T.L., Iyer, R. & Gardner, KH. (1993) Dynamic evaluation of municipal waste combustion ash leachate, p. 680-689. In (J. Kilgroe, ed.) Proceedings I993 International Confeence on Municipal Waste Combustion, AWMA, Pittsburgh, Penn. Vehlow, J., Schneider, J, & PfrangStotgG. (1992) Restoffe - Charakterisierung, behandlung verwertung. In: Furshung und Entwicklung in Kemforschungszentrum Karlsnrche zur Hausmillverbrennung: Symposium 25 jahre LIT 5 jahre TAULRA: Kerforschungszentrum Karlsruhe GMBH, Karlsruhe. Yoder, H.S. (1979) Melilite-bearing rocks and related lamprophyres, p. 391-412. In (Yoder, H. S., ed.) The Evolution of the Igneous Rock: Fifseth Anniversaty Perspectives, Princeton University Press, Princeton, New Jersey.
Environmental Aspects of Constnrction with Waste Materials J.J.J.M. Goumans, H A . van der SIoot and l k G . Aalbers (Editors) el994 Elsevier Science B.V. AN rights resewed.
137
An Approach to the Assessment of the Environmental Impacts of Marine Applications of Municipal Solid Waste Combustion Residues 0. Hjelmar, E.Aa. Hansen, K.J. Andersen, J.B. Andersen and E. Bjsrnestad VKI, Water Quality Institute, Agern All6 11, DK-2970 Hsrsholm, Denmark
Abstract A comprehensive approach to the assessment of the environmental impacts of marine land reclamation with municipal solid waste incinerator (MSWI) residues is described. Based on a study of the potential environmental impacts of marine foreshore land reclamation of MSWI bottom and fly ash in Bermuda, a methodology including sampling and characterization of bottom ash, fly ash and combined ash, performance of laboratory leaching tests with ocean water and rainwater on granular and solidified ash, studies of acute and chronic toxicity of ash leachates to local marine organisms, performance of baseline studies, development of acceptance criteria, modelling of the release of potential contaminants resulting from various disposal options at two locations, modelling of the dilution of released contaminants in the surrounding sea and assessment of the resulting environmental impact, is presented. 1.
INTRODUCTION
Municipal solid waste incinerator (MSWI) bottom and fly ashes have been shown by numerous studies [e.g. 1,2,3,4,5] to produce leachates which initially are characterized by high concentrations of inorganic salts and moderate to low concentrations of trace elements. As the leaching progresses with time, the contents of both salts and trace elements in the leachate may decrease to very low levels. For well burnt-out ash, the content of organic, biodegradable matter in the leachate is generally low. In practice, leaching of incinerator residues may occur over a very long period of time. This suggests the employment of a controlled contaminant release strategy for bulk application
or disposal of incinerator ash. Such a strategy would imply that the ash should be monofilled and that the application or disposal site should not be equipped with a leachate containment system. Instead it should be designed carefully to reduce the rate of transfer of contaminants out of the site to an environmentally acceptable level, and the site should be located in an environment which is not very sensitive to the anticipated impact. Proper design and proper selection of a site location therefore play crucial roles in the protection of the environment and should always be based on a preceeding assessment of the resulting environmental impacts. The rate of transfer of Contaminants out of the site may be reduced by restricting the flow of water through the site or by retarding or reducing the leaching of contaminants. Stabilization of granular ash with cement will often accomplish both of these objctives simultaneously. A
138 geologically stable, sloping top cover of low permeability combined with a surface drainage system can minimize the infiltration of rainfall into a site. The dominating components of MSWI ash leachate, inorganic salts, are totally compatible with seawater, and a location of a MSWI ash bulk application or disposal site at or in the sea would eliminate any potential groundwater pollution problems caused by leaching of salts. The efforts should then be concentrated on reducing the rate of release of trace elementslheavy metals (and possibly organic compounds) to an acceptable low level. Although the high salt content of the seawater itself may initially enhance the leaching of certain trace elements from the ash through the formation of chloride complexes, the sea offers a very stable, well buffered chemical environment with a constant pH of 8.1 which is generally favourable in terms of low trace element solubility and which is a safeguard against any substantial long term changes of the leaching conditions. During the period 1989 - 1991, the Water Quality Institute (VKI)in conjunction with Bermuda Biological Station for Research Inc. (BBSR) and the Danish Hydraulic Institute (DHI) performed a comprehensive study with the objective of assessing and minimizing the environmental impacts of disposal of MSWI bottom ash and fly ash by marine land reclamation in Bermuda [6,7,8]. The study which was carried out for Ministry of Works and Engineering, Bermuda, was initiated in order to enable the prevention of unacceptable impacts on Bermuda’s vulnerable marine environment from disposallapplication of an annual amount of approximately 16000 tonnes of combined bottom and fly ash (without acid gas cleaning residues) expected from a planned refuse incinerator designed to treat 50,000 tonnes of garbage per year. Due to the scarcity of land and the mid-ocean location, marine land reclamation of the MSWI ash was an attractive option in Bermuda, provided unacceptable emvironmetal impacts could be avoided. This paper presents an overview of the approach, the methodology and the results of the study. 2.
APPROACH
The approach of the study is outlined in figure I . Initially, a few selected disposal options were defined in terms of location (Tynes Bay and Castle Harbour, see figure 2), physical design, geotechnical and hydrogeological properties, rate and mode of disposal of ash, etc. The options or scenarios included disposal of both unsolidified and cement stabilized incinerator ash. A number of laboratory leaching tests were then designed and conducted on MSWI ash samples in order to estimate the rate of leaching of contaminants from the ash under the conditions corresponding to each disposal option, and to evaluate the effect of various efforts to minimize the rate of leaching, particularly cement stabilization of the ash. Ash samples from waste-toenergy facilities were procured from the USA (Saugus, Massachusetts) and Europe (Jersey, Channel Islands). The leaching studies which included the performance of column, batch and tank leaching tests with rainwater and seawater were set up to simulate local conditions as closely as possible. The interpretation of some of the leaching results were supported by chemical equilibrium calculations using a computer model. Several ecotoxicological tests (bioassays) were performed in the laboratory to determine the acute and chronic toxicity towards sensitive local marine lifeforms of the most contaminated leachate from the leaching tests. The dilutions needed to render the leachate non-toxic were calculated
139
Descriptionof disposal options
I
and rwtew of existing leaching data
L
I
Choke of disposal option@)
Performance of baseline study
Land reclamation and development of monitoring procedures
Is impact
Fiiure 1 Outline of the approach of the study.
and used together with chemical composition data for the development of recommended acceptance criteria. Baseline field studies were performed at both site locations to describe the present condition, the species composition and diversity of the communities of flora and fauna at the main nearshore marine habitats (rocky shores, coral reefs, sediments, seagrass beds). Bioaccumulation of selected trace elements by transplanted mussels kept in cages was investigated in preparation of future biomonitoring of the disposal site. Data were collected on the background levels of trace elements in water, sediment, and selected biological species. Samples of leachate from an existing dump site at the Civil Air Terminal at Castle Harbour were collected and analyzed.
140
Bermuda had no guidelines which define specific acceptance criteria for the environmental impact of a disposal or land reclamation site. A set of suggested acceptance criteria in terms of maximum emissions of selected trace elements has therefore been developed as part of the project work. The rate of emission of selected contaminants was then estimated for each disposal option over a period of 10 years of disposal and 20 years following that by combining the experimentally determined leaching data with a mathematical description of the rate of disposal of the ash and the flow of water through the sites. The estimated rate of emission of dissolved contaminants was used as input to a dynamic mathematical simulation model which predicted the dilution of the solutes in the marine waters surrounding the two site locations. A number of field measurements were conducted at the two sites to provide the necessary information on current and flushing conditions. Finally, an environmental impact assessment was performed for each of the disposal options considered. The calculated concentrations of leachate/contaminants in the marine waters surrounding the sites were compared to the results of the bioassays, the baseline study, available information on the toxicity of specific contaminants, and the suggested acceptance criteria. The safety margins and uncertainties involved in the entire assessment procedure was evaluated, and the disposal options were ranked in order of increasing environmental impact andlor uncertainty.
3.
DISPOSAL OPTIONS
Two locations have been considered for ash disposal. The primary site for consideration is at the Civil Air Terminal at Castle Harbour. The secondary site is at Tynes Bay in Devonshire. Both sites are shown in figure 2. Castle Harbour is the second largest of the inshore basins of the Bermuda platform, covering some 10.5 km2. The morphology, water exchange, sediments and biota have been significantly impacted by dredging, reclamation and causeway constructions in the past. The proposed disposal site is located adjacent to an existing scrap metals dump at the Civil Air Terminal. The bottom sediment in the disposal area is typically a fine silty mud, and the average water depth is 12 m. It is anticipated that the ash disposal will continue to an approximate height of 5 m above sea level. Tynes Bay is located on the north shore of Bermuda in Devonshire which is also where the incinerator will be located. The bay is open to the North Lagoon, which extends about 11 kilometers to the north and east and has a maximum depth of 20 m. The shoreline is rocky, and the seabed is sandy. The average water depth in the proposed disposal area is 5 m. As in Castle Harbour, ash disposal to a height of approximately 5 m above sea level is anticipated. Three different foreshore land reclamation disposal scenarios have been considered at each location:
141
Disposal by placing cured 1 m3 blocks of cement stabilized combined ash directly into the sea from the coastline. Above sea level, the concretelash mixture may be poured directly onto the surface of the site as it is raised. Disposal by backtipping of untreated combined ash into sequences of 15 m x 15 m compartmentshasins made of sheetpiles which have been rammed into the seabed. Combined solution. In the sheetpile compartments, cement stabilized combined ash blocks are placed below sea level. Unstabilized combined ash is placed above sea level.
Figure 2 Map of Bermuda showing the two potential site locations.
4.
SAMPLING AM) CHARACTERIZATION OF MSWI ASH
Since no incinerator ash was available in Bermuda, samples of bottom ash (BA), fly ash (FA) and combined ash (CA) were procured from two MSWIs which were chosen on the basis of their similarity to the future situation in Bermuda regarding waste feed composition, combustion and gas cleaning technology, and ash management. The Martin facility in Jersey was chosen as representing a resort island and the Wheelabrator facility in Saugus, Massachusetts was chosen as representing the strong North American influence on the consumer products and, consequently, the waste composition in Bermuda. The Jersey facility (Martin design) had a capacity of 240 tonnes of refuse per day and the Saugus facility (Von Roll design) had a capacity of 1500 tonnes of refuse per day. As will the in-
142 cinerator in Bermuda, both facilities were equipped with electrostatic precipitators for collection of fly ash, and neither of them had acid gas scrubbers at the time of sampling. At the Jersey facility, the samples were combined from a large number of subsamples collected over a period of one week. At the Saugus facility, the ash samples were collected over a period of less than one day (bottom ash and fly ash) and two days (combined ash). At the Jersey facility, the bottom ash was collected dry before it was quenched whereas the bottom ash from the Saugus facility and the combined ashes were collected after quenching. Both fly ashes were collected dry. The dry bottom ash from Jersey was crushed with a hammer. Material larger than 45 mm and ferromagnetic material were removed from the bottom and combined ashes which were then reduced to manageable sample sizes (100 - 120 kg) by means of a riffle or by coning and quartering. Representative BA and CA samples were crushed to < 2 mm in a jaw crusher and, after further subdivision by a riffle, ground to < 0.090 mm in a tungsten carbide mortar grinder. The ground BA and CA and the FA were subjected to chemical analysis. The results of the chemical analysis which are presented in table 1 indicate that the elemental compositions of the residues from Jersey and Saugus, Mass. are quite similar to each other and not very different from the compositions of similar residues from other incinerators, although the lead concentrations in the FA from both Jersey and Saugus are unusually high (about twice the level found in most fly ashes). The zinc concentrations in both FA, BA and CA from Jersey and Saugus are also relatively high. The Jersey FA reacts almost neutrally with water whereas the Saugus FA creates a higher pH upon contact with water. The Jersey BA, in contrast, has a more alkaline reaction with water than the Saugus BA, probably because the Jersey BA was sampled dry and the Saugus BA after quenching. The two combined ashes are quite similar to each other, both in pH and alkalinity. The content of organics is described in general terms as loss on ignition at 550 "C (LOI) and total organic carbon (TOC). The FA and BA samples from each facility were in addition analyzed for content of polychlorinated dibenzo-p-dioxins (PCDD) and dibenzofurans (PCDF). It appears from the TOC and LO1 results that the residues from Saugus, Mass. have a somewhat higher content of uncombusted material than the residues from Jersey. This is in accordance with the visual appearance of the residues.The concentration of PCDD and PCDF is very low in the bottom ash and somewhat higher in the fly ash from both facilities. The concentration levels found in the fly ash samples are lower than or within the lower end of the ranges reported by others. 5.
LABORATORY LEACHING TESTS ON MSWI ASH
5.1 Granular ash Samples of granular bottom ash (BA), combined ash (CA) and fly ash (FA) were subjected to column and batch leaching tests with ocean water and artificial rain (demineralized water acidified to pH = 4 with HNO,). 10 kg samples of BA and CA and 6 kg samples of FA from each facility were placed in plexiglass columns of diameter = 0.15 m and height = 0.80 m and leached with seawater and artificial rain, respectively, in upflow at a flow rate corresponding to 15-20 mm124 hours. Leachate fractions corresponding to the liquid/solid ratios (LIS) = 0.00.2, 0.2-0.5.0.5-1.0, 1.0-2.0. and 2.0-5.0 I/kg were collected from each of the 12 columns and
143 Table 1 Analyses of MSWI ash. All
ELEMENT
element analyses
UNIT
JERSEY F A C l L I T Y FA
Fe
ca K Mg Na S P Ti Mn Ag As Be Cd co Cr cu Hg Mo Ni Pb
se Sn
Sr V
Zn pH (1% s l u r r y ) ANC eqv/kg
LO1 TOC
are total analyses.
7.6 2.8 27 8.2
g/kg g/kg
FA: F l y ash BA: Bottom ash CA: C d i n e d ash
ANC: LOI:
TOC:
BA
SAUGUS FACILITY CA
FA
220 52 1.7 140 a5 12 14 22 3.5 4.2 7.2 1.7 4.1 29 2700 < 5 26 310 4800 0.26 6.3 190 2100 1.2 260 200 a4 6200
23 32 390 4800 0.28 6.4 150 5200 2.2 320 100 a6 9100
170 68 45 22 74 23 13 29 22 5.1 12 0.90 5a 140 1300 370 29 340 920 7.0 21 92 15000 7.8 1900 100 80 41000
11.1 3.5 5.9 5.5
9.9 3.1 19 7.8
10.2 2.5 45 8.0
220 53 6.3 120 a2 13 13 25 4.6 3.7 6.8 1.5 8.3 41
iaoo
BA
CA
210 52 1.2 150 69 9.2 9.4 27 6.1 2.9 7.0 1.6 7.3 38 1500 12 33 300 3400 0.22 14 190 3900 0.63 300 200 51 5300
210 56 1.1 120 76 9.8 10 25 8.9 2.7 6.7 1.3 9.2 26 1000 23 24 480 3100 0.20 4.4 160 3100 0.48 280 200 53 5500
9.6 2.0 26 13
3;o
9.7 35 17
Acid n e u t r a l i z a t i o n c a p a c i t y ( a l k a l i n i t y ) Loss on i g n i t i o n T o t a l organic carbon
Table 2 Results of PCDD and PCDF analysis of ash samples. Units: pgikg.
PARAMETER
JERSEY F A C I L I T Y FA
140 4a 0.10 0.27 1.5
PCOD ” PCDF 2,3,7,a-~c~o 2.3.7,a-TCDF TCDD-EPV (Eadon)
’’
1) 2) I’
SAUGUS F A C I L I T Y BA
FA
1.5 - 1.9 0.44 0.64 < 0.03 < 0.007 0.014 - 0.046
-
115 69 0.21 0.52 2.5
S u n o f TCOD, PnCDD, HxCDD, HpCOD and OCDD S u n o f TCOF, PnCDF, HXCDF, HpCDF and OCDF
BA
-
9.0 10 3.4 - 4.5 < 0.07 < 0.05 0.14 0.22
-
144
subjected to chemical analysis. When the column experiments were completed, two consecutive batch leaching tests, each at LIS = 10 IIkg, were performed on well mixed material from each column with the appropriate leaching medium. The two filtered fractions (0.45 pm filter) were combined to represent LIS = 5.0-25 for each leaching test. Samples of BA and CA were subjected to additional batch leaching tests with seawater at LIS = 25, 50, 75, 100, and 200 Ilkg. The collected leachate fractions were analyzed for pH, conductivity, TDS, alkalinity, sulfate, chloride, total-N, total-P, Na, K, Ca, Mg, As, Cd, Cr, Cu, Hg, Mn, Ni, Pb, Se, Sn, Zn, and non-volatile organic carbon (NVOC). Some of the results are presented in terms of total leached quantities in tables 3, 4 and 5 . Table 3
Total amounts of some trace elements and NVOC leached in column and batch leaching tests with artificial rainwater (L/S = 0 - 25 I k g ) . Units: mg/kg of ash.
PARAMETER
JERSEY F A C I L I T Y FA
AS Cd Cr
cu HQ Pb Se Zn NVOC
0.12 0.05 < 0.03 < 0.12 0.05
0.7 0.40 < 0.3 104
SAUGUS F A C I L I T Y
BA
CA
< 0.01 < 0.013 < 0.02
< 0.06 < 0.02
0.42 0.094 0.05 0.063 < 0.3 490
0.61 0.11 0.21 0.12 < 0.3 440
0.038
FA
EA
0.073
0.018
< 0.06
< 0.014
< 0.03 < 0.02
< 0.02
0.084 1.1 0.18 < 0.3 170
CA
0.53 0.032 0.32 < 0.03 0.3 230
0.047 0.043 < 0.03
0.35 < 0.03
0.26
< 0.005 < 0.3
300
Table 4 Total amounts of some trace elements and NVOC leached in column and batch leaching tests with ocean water (LIS = 0 -25 I k g ) . Units: mgkg of ash.
I 11
FA As Cd Cr
cu Hg Pb
se Zn NVOC
SAUCUS F A C I L I T Y
JERSEY F A C I L I T Y
PARAMETER
0.081 2.9 < 0.02 < 0.02
0.07 15 0.42 0.8 110
BA
0.024 0.21 < 0.02
1.4 0.053 0.93 0.095 0.6 470
CA
0.060 0.71 < 0.02
0.96 0.099 5.7 0.11 2.8 220
FA
0.072 0.93
BA
0.028 1.3
< 0.02 < 0.03
< 0.02
0.08
< 0.03
18 0.24 2.6 67
2.7 0.40 < 0.03
6.6 310
CA
0.028 0.88 < 0.02 0.88 < 0.03
0.33 < 0.03 5.7 270
145 Table 5 Accumulated amounts of trace elements leached in sequential batch leaching tests on CA and BA from both facilities with seawater at L/S = 2 x 100 lkg.
RESIDUE
ACCUMULATED AHWNTS LEACHED (mg/kg)
ACCUM.
L/S
As
Cd
cu
Ni
Pb
Zn
CA Jersey CA Jersey
100 200
0.7 1.2
4.1 7.2
4.1 6.1
< 0.2 < 0.4
13 22
3.3 8.0
CA Saugus CA s a u g u s
100
0.5
200
< 0.7
3.4 5.4
2.8 3.9
< 0.2 < 0.4
11 20
3.9 7.7
BA Jersey BA Jersey
too
< 0.2 < 0.4
0.10 0.35
1.5 2.4
< 0.2 < 0.4
0.096 1.9
3.3
200
BA Saugus BA Saugus
100 200
< 0.2 < 0.4
2.6 3.7
2.6 4.8
0.4 < 0.6
4.3 7.1
7.1 12
-
The composition of the leachate is dominated by inorganic salt ions such as Cl-, SO:, Na+, K + , and Caz+, but these contaminants are of no environmental concern in a marine disposal situation. The concentration level of trace elements varies between the different tests but is generally moderate to low. Most parameters in most of the tests (but not all) appear in the highest concentrations in the first few fractions of leachate collected. Cd, Cu, Pb, and Zn are leached more extensively by seawater than by rainwater, possibly due to chloride complexation, whereas the leaching of As, C r (which is below the limit of detection), Hg, and Se is approximately the same in seawater and rainwater. The data show quite clearly that Pb, As, and Se are leached in greater amounts from FA than from BA, and that more Cu and NVOC are leached from BA than from FA. The trace elements leached in the largest amounts are Pb, Zn, Cd, and Cu. The relatively large amounts of NVOC leached could also give rise to some concern and emphasize the importance of good burn-out of the bottom ash. Improved burn-out could probably reduce the amount NVOC leached by more than 75 %. A special set of leaching tests was carried out in order to determine whether PCDDs and PCDFs could be leached from the incinerator residues with seawater. No detectable amounts of these compounds were leached, and it may be concluded that polychlorinated dioxins and furans are unlikely to be problematic in relation to leaching of incinerator residues. Any PCDDs and PCDFs present can be expected to be strongly associated with particulate matter which should be prevented from escaping.
Cement stabilized ash Combined ash from the Jersey facility was stabilized with sulfate resistant low alkali cement and allowed to cure for 7 days at 40 "C and 100 percent relative humidity before being subjected to dynamic tank leaching with seawater in order to determine the rate of release of key Contaminants from the monolithic waste forms. 5.2
146 The ash which had been crushed to 10 mm was stabilized after some of the metal pieces which had been removed earlier were returned to the ash (approximately 2 percent of the total weight). The metal was reintroduced into the ash in order to simulate an ash quality which Bermuda may produce. Two mixtures were prepared: one containing 92.6 percent (w/w) of ash and 7.4 percent (w/w) of cement and one containing 87 percent (w/w) of ash and 13 percent (w/w) of cement, both containing 12 percent of water in addition to the ash and cement. The mixtures were cast in rigid plastic molds with diameters = 10.5 cm, settled by vibration, and cured. Cube shaped specimens (approximately 6 cm x 6 cm x 6 cm) were cut out of the cured test materials and used in tank leaching experiments. Dry densities of 1.80 and 1.90 g/cm3 were obtained with 7.4 and 13 percent of cement, respectively. Due to lack of material, the compressive strength could only be determined using non-standard cylinders of height = 5 cm and diameter = 10.5 cm. For these samples the unconfined compressive strength after 7 days was found to increase from 2.5 MPa to 6.0 MPa when the cement content increased from 7.4 to 13 percent. The cut out cubes were subjected to tank leaching with ocean water at 20 "C. In the tank leaching experiments, each of the monolithic waste forms was suspended in a closed seawaterfilled polypropylene tank by a nylon string. At certain time intervals (increments of 0.7, 1 , 3, 6, 11, 18, 92, and 47 days), the water (0.9 litres) was replaced by new seawater and analyzed for pH and trace elements (As, Cd, Cu, Pb, and Zn) after filtration through a 0.45 p n filter. This allows the calculation of the flux and accumulated leached amounts of the trace elements as a function of time. The initial rates of release were highest for Cu and Pb, an order of magnitude smaller for As and Cd and below the detection limit for Zn. After approximately 6 months, the rates of release had decreased by a factor of 100 to 1000 and were below the detection limit for As and Pb. The Cu flux data were comparable to Cu fluxes determined by others in similar tests on blocks of solidified combined MSWI ash containing 18 percent (wlw) cement. The leaching of organics from the stabilized ash was not measured but may safely be assumed to be very low in comparison with the leaching of organics from granular ash. 6.
TOXICITY STUDIES
The toxic effects of the first few fractions of leachate from the column leaching experiments on FA, BA, and CA from each of the two facilities with seawater were studied both on local species, on one of the most dominant species of the zooplankton, the marine copepod Acarfia fonsa, and on a standardized marine test organism, the rotifer Brachionus plicarilis. Suitable local organisms were selected for their relevance to the potential incinerator ash land reclamation sites as well as their niche in the Bermuda ecosystem. As this project was the first of its kind in Bermuda, and was the first to utilize these organisms, exhaustive tests were carried out and various experimental techniques had to be evaluated. The effects of ash leachate on the early stages of the purple urchin Lyfechinus variegafus and the limpet Siphonaria alfernafawere studied. These organisms live along the rocky shores of Bermuda and occur in close proximity to the Tynes Bay site. Both organisms showed a similar overall response to the various ash leachates with a general trend of decreased toxicity with increasing fraction number and increased dilution. Overall, the leachates produced by the Jersey
147
bottom and mixed ash proved to be the most toxic. The leachate samples from the Massachusetts ash proved to be markedly less toxic. The copper and lead contents in the Jersey leachate correlated well with toxicity. This was in sharp contrast to the results obtained at the VKI and in Bermuda using the rotifer Brachionus plicarilis. The FA leachate from both incinerators gave results similar to each other. In this case the toxicity did not correlate to the copper and lead concentrations. A study on the effects of ash leachate on primary productivity of the symbiotic algae from the sea anemone (Aiprasiapallida) was carried out. Primary productivity is the basis of the world's food chains and the importance of this type of algae in Bermuda made it a good candidate for testing as similar algae are major organic producers on coral reefs. The results were generated using 14C-labellingtechniques on algae exposed to various concentrations and types of toxicants. Jersey BA leachate had the most pronounced effect with fractions 1-3 inhibiting productivity to some extent but only at highest leachate concentrations. The combined ash leachate had the opposite effect as fractions 1 and 2 stimulated primary productivity. Fraction 1 of the FA leachate inhibited photosynthesis but only at extremely high concentrations. The Massachusetts ash leachate were even less toxic than the Jersey ash. Tests on the effects of the incinerator ash leachates on the marine rotifer Brachionus plicarilis were carried out as this is a "standardised" test for toxicity. The rotifers were hatched from cryptobiotic cysts and exposed for 24 hours to a dilution series of the first two fractions of leachate from BA, FA and CA from both facilities. No effects were seen after the standard 24 hours nor were effects seen after the test was extended for 48 hours. The copper levels were higher in the leachates than levels of copper that would normally cause an effect. It may be that the toxicity of the metal to the organisms in this case were modified by other constituents leached from the ash. The high NVOC (480 mgll) measured in the Jersey bottom ash leachate may have an important influence on the bioavailability of copper in seawater. The chronic effects of the incinerator ash leachates on the anemone Aiprusiapallida were studied over a 120 day period. Asexual bud reproduction was the most sensitive test found in all of the studies of local species where Jersey CA leachate toxicity effects were shown at concentration of leachate at 1 mlll with no significant effect below 0.25 mM. Massachusetts CA leachate showed an effect at concentrations greater that 0.25 mVI. Other parameters such as survival, growth, biochemical measurements and metal accumulation showed no effects at the concentrations tested. The acute toxic effects of the ash leachates on adults of the marine copepod Acarria fonsa were studied over periods of 24 and 48 hours. Acarria ronsu is not native to Bermuda's waters but it is a well referenced test organism. The highest mortality were found for the FA leachates, particularly the initial leachate from Jersey FA (LCIO = 5 - 20 ml/l), whereas the BA and CA leachates, especially from the Saugus facility, were less toxic (LCIO = 23 - > 100 mlll). The chronic toxic effects of the initial leachate from Jersey FA and BA on Acarria fonsa were studied as the effect on the growth (19 days) and subsequent egg production (96 hours exposure) of fertile females of Acartia tonsa (life-cycle test). In the egg production test, EC20 values of approximately 0.5 ml/l for both the FA and BA leachates were found. The corresponding EC50 values were 0.5 - 1.0 mlll for FA leachate and > 10 ml/l for BA leachate.
148
Although they are not directly comparable, the concentration levels below which no chronic effects can be expected from the most toxic leachate fractions are of the same order of magnitude (0.25 - 1.0 mlll) both for Aiptusiapallida and Acartia tonsa. In a number of the toxicity tests on local organisms, the FA leachates were found to be less toxic than the BA leachates, whereas the FA leachates were more toxic to Acarria ronsa than the BA leachates. These apparent discrepancies may have been caused by different copper complexing capacities of the seawater types used for dilution in the various tests. 7.
BASELINE STUDIES
Various studies were carried out to evaluate present environmental conditions around the two proposed disposal areas (Tynes Bay and the Civil Air Terminal at Castle Harbour). The studies have included hydrographic, chemical and biological studies of the area as well as a "groundwater" contamination survey at an existing metals dump at the Airport. The hydrographic studies have provided physical oceanographic information on the two sites for incorporation into the physical dispersion model. The hydrologic regimes of the two areas have been described using data collected from two current surveys using Anderaa RCM 7 current meters. Data for current speed, direction, tidal oscillation, temperature and salinity have been gathered at 10 minute intervals for 1-2 weeks at a time. Generally, the currents in Castle Harbour are very weak, while the current speeds at Tynes Bay are somewhat faster. The biological measurements have consisted of benthic surveys and "mussel watch" programmes to determine the baseline environmental quality of the areas. As in the toxicity tests a locally important representative, the turkey wing mussel Arca zebra, was employed. Chemical data on water and sediments have provided baseline information on the state of the environment at the present time. Macrofaunal studies: sub-tidal surveys were carried out at eight sites in each area for the macrofauna (larger organisms such as worms, clams etc.) existing on and in the sediments. The results indicate that Tynes Bay and Castle Harbour contain different macrofaunal communities. The differences arise primarily from environmental conditions, and reflect the differing sediment characteristic of both areas. There is a significant seasonal change in species abundance (numbers of organisms) at Castle Harbour between winter and summer. There were no significant changes at Tynes Bay. There was a 30% turnover in species composition (types of organisms) at both sites which may indicate a seasonal cycle of species occurrence. Conditions should be continually assessed in the monitoring programme in order to determine future changes. Macrofloral studies: phototransects were collected for the large algae and seagrasses at all eight of the sampling sites at Tynes Bay during the winter of 1989 and the summer of 1990. Results indicate a great deal of variability between the two sampling periods. This variation may be due to seasonal differences but may also be due to continual changes within these populations. Further surveys should be carried out over time to establish firm baselines so that the an-
149
thropogenic effects can be distinguished from natural variations. The macroalgae Penicillus sp. in particular may be a good indicator organism for the determination of thermal effects in the Tynes Bay area. Meiofaunal studies: were also carried out in both areas. Meiofauna (microscopic organisms that live in sand) were collected and sorted, but were archived for later use in the monitoring programme. Rocky shore assessments were carried out at Tynes Bay as intertidal communities are exposed to contaminants from the island as well as the sea. These surveys consisted of quadrate measurements of limpets, barnacles and mussels on the rocky area of Tynes Bay. Surveys were carried out on a seasonal basis. In Bermuda there are periodic, naturally occurring die-backs of these communities which should be continually monitored so that the correct cause of these changes can be assessed. Baseline chemical studies have been carried out at eight field stations at each site. Water samples have been taken for pH, salinity and the trace metals. Concentrations of constituents in the water samples taken were of the same order as a BBSR study conducted in 1984. Sediment samples at both sites were analyzed for grain size distribution, and trace metal content. Generally, the Castle Harbour sediments are finer than those from Tynes Bay, which would be expected for this lower energy environment. These fine, silty sediments contain higher concentrations of trace metals. The closer to the airport metal dump, the higher are the concentrations of copper, zinc, lead, iron and cadmium. For the most part the concentrations decrease with increasing distance from the source. Bioaccumulation studies of trace metals by the mussel, Arcu zebra have been carried out during two periods, winter and summer, at both sites. The mussels were analyzed after one and two months' deployment in the field and were used as "sentinel" organisms. In contrast to the findings for the sediments, there were no significant gradients of trace elements between the mussels placed at different distances from the metals dump at Castle Harbour, although the highest contents of Cd, Cu, Fe, Pb, and Zn were found in mussels adjacent to and SW of the dump, particularly during the winter period. However, the concentration differences were small, and there were no gradients for Cr and Ni. The general level of trace elements in the mussels were comparable at the two sites. The use of mussels as sentinel organisms should be considered as part of the monitoring phase. In order to investigate the contamination potential of the existing metals dump at the Civil Air Terminal at Castle Harbour, three "groundwater" sampling wells were installed in the site. Chemical analysis of samples of the groundwater or leachate showed a minor contamination with organics. The concentration levels of cadmium, copper, lead, and nickel were approximately one order of magnitude higher in the water inside the site than in the surrounding Castle Harbour. Chromium concentrations were of the same order of magnitude inside and outside the site.
150
8.
ACCEPTANCE CRITERIA
Since Bermuda has no guidelines which define specific acceptance criteria for the environmental impact of a disposal site, the study has included the development of a set of proposed acceptance criteria. The overall purpose of developing acceptance criteria for the emission of contaminants into a water body is to protect the ecosystem (living pelagic or benthic organisms and the sediment) as well as humans from bioaccumulation and toxic effects. Only impacts on organisms in the water body are considered in this study. Various methods of defining and calculating acceptance criteria for single substances in the water body, including the methods developed by the U.S. EPA [9] and the OECD [lo], have been evaluated. The acceptance criteria suggested for Bermuda are based on modified U.S. EPA Water Quality Criteria for saltwater (all the modifications represent increased restrictions). It is suggested to use these single substances criteria in conjunction with another set of criteria based on the actually observed toxicities of the most toxic fractions of ash leachate tested in this study on local marine organisms. The suggested single substance acceptance criteria are expressed as two sets of maximum resulting ambient water concentrations of relevant trace elements (As, Cd, Cr, Cu, Pb, Hg, Ni, and Se) at Tynes Bay and Castle Harbour. One of the sets of criteria should ensure that no acute toxic effects occur, the other that no chronic toxic effects occur. It is suggested that the maximum calculated mean concentration of each trace element in the mixing zone should not cause any chronic effects, and the maximum concentration in the leachate at the point where it penetrates the bottom sediment should not cause acute effects. The mixing zone is suggested as the minimum compartment size used in the dilution model (250 m x 250 m). Using the results of the dilution modelling, the acceptance criteria based on chronic effects may be used to calculate the site specific rates of release of contaminants from the disposal sites at either location which result in ambient water concentrations not exceeding the criteria concentrations. The rates of release corresponding to the acceptance criteria in the mixing zones at each site are called acceptable loads (AL). Levels of e.g. lo%, 50% or 100% of the calculated acceptable loads may be chosen. At the 10% level, the calculated acceptable rates of release of Cd, Cu, Hg, and Pb are 0.07,0.34, 0.02 and 4.7 kglday, respectively, for Tynes Bay and approximately half of that at Castle Harbour, thus reflecting the less favourable dilution conditions at Castle Harbour. There may be reasons other than just specific toxic effects to reduce the release of Pb to less than the figure shown above (e.g. considerations concerning accumulation in sediments and biomass and subsequent possible effects on human health as well as the ecosystem). The other type of suggested acceptance criteria which is based on the observed toxic effects of the initial ash leachates indicate that minimum resulting leachate concentrations of 10 - 23 ml/l (corresponding to a dilution of 44 - 100 times) are required to avoid acute effects and 0.16 - 0.4 mlll (corresponding to a dilution of 2500 - 6300 times) are required to avoid chronic effects. These criteria relate to the initial leachate from granular ash. The leachate from cement stabilized ash will be much less concentrated and much less toxic at any time.
151
9.
MODELLING OF EMISSIONS
Based on the results of the ash leaching tests as well as meteorological, hydrological, and geological data and a number of assumptions and preconditions concerning the design and layout of the sites, the emission of trace elements (Pb, Cd, and Cu for granular ash and As, Cd, Cu, Pb, and Zn for cement stabilized ash) have been modelled for six disposal scenarios: Disposal of 1 m3 blocks of CA stabilized with 13 % cement directly into the sea at Tynes Bay (Scenario 1) and Castle Harbour (Scenario 4). Disposal of granular CA by backtipping into 15 m x 15 m compartments contained by sheetpiles rammed into the seabed at Tynes Bay (Scenario 2) and Castle Harbour (Scenario 5). Combined solution. In the sheetpile compartments, blocks of CA stabilized with 13% cement are placed below sea level and granular CA is placed above sea level at Tynes Bay (Scenario 3) and Castle Harbour (Scenario 6). In order to obtain an adequate description of the rate of release of Cd and Pb from the granular ash placed below sea level by backtipping, it was necessary to supplement the experimental data with calculations using the hydrogeochemical equilibrium and speciation model MINTEQ2A.
For each scenario, the rates of emission of the above mentioned trace elements were then calculated, first during a period of 10 years of disposal of an annual amount of 16,000 tonnes of combined ash, and subsequently for 20 - 30 years following that. The maximum daily emissions of each trace elements were calculated for each scenario and compared (see table 6). The lowest maximum daily emissions will occur from Scenarios 1 and 4 (disposal of cement stabilized ash blocks): 0.01 glday of Cd, 4.9 g/day of Cu, and 0.54 g/day of Pb at both locations. The highest maximum daily emissions will occur from Scenarios 2 and 5 (disposal of granular ash): 6.2 glday of Cd, 42 g/day of Cu, and 23 g/day of Pb at Tynes Bay and slightly less at Castle Harbour. The maximum daily emissions of contaminants from the combined solution (Scenarios 3 and 6) are in between those. The model predicts that in most cases the maximum daily emissions will occur at the time of maximum ash deposition, i.e. at the end of the 10 year disposal period (in a few cases later than that). After this, the emissions keep decreasing with time to a low level. This is due to the fact that the rate of release of most contaminants from the ash decreases as the leaching process progresses but it also reflects the method and rate of disposal chosen. 10. MODELLING OF DILUTION Mathematical hydrodynamic models (two-dimensional) have been set up to describe the current conditions in the marine areas at the two designated potential disposal sites. Using the current data collected, simulations have been performed with a transport-dispersion model coupled to a hydrodynamic model in order to investigate the spreading of dissolved contaminant emissions from the potential sites in the surrounding waters. For the purpose of calculations, the model uses a 250 m by 250 m grid. Mean excess concentration fields for the simulated periods have been computed and mapped for each location.
152 Table 6 Calculated maximum daily emissions of selected trace elements.
Scenario 1 Scenario 2 Scenario 3
1
Castle Harbour
I
Scenario 4 Scenario 5 Scenario 6
0.16
0.16
0.01 6.2 0.1
4.9 42 6.9
0.54 23 1.6
2.8
0.01 5.4 0.07
4.9 31 6.7
0.54 19 1.3
2.8
1.8
2.4
For the Tynes Bay location a reasonable agreement between measured and calculated current velocities was obtained. Two situations were considered: The first was dominated by moderate northerly wind and the second by moderate westerly wind. In the first case the contaminants were found to spread primarily towards the northeast along the coast. In the second situation the contaminants would spread towards the northwest, the plume keeping clear of the coast. In both cases a dilution of 10 times that in the source compartment closest to the site were reached at a distance of approximately 3500 m from the site. Generally, current velocities are low in Tynes Bay due to the sheltered and shallow nature of the area. Thus horizontal mixing is slow despite slightly favourable wind conditions.
For the Castle Harbour location, the recorded current speeds are extremely low, and no correlation to wind data is apparent. Consequently, the model calculations failed to reproduce the recorded currents. The modelled current velocities are very low, though, and indicate that the emitted contaminants will spread to all of Castle Harbour and most of Ferry Reach at concentrations ranging down to one tenth of that in the source compartment of the model. Half the amount of contaminants discharged will be flushed to the North Lagoon via Ferry Reach. The model calculations - however inadequate - along with other evidence indicate that horizontal dispersion at Castle Harbour is slower than in Tynes Bay. Recordings and modelling results indicate that the wind may induce vertical circulations which is the cause of the partial failure of the attempt to reproduce the conditions in Castle Harbour using a two-dimensional model. In terms of dilution potential Tynes Bay is the better of the two disposal sites considered. The hydraulic measurements indicate that the dilution potentials at neither of the two locations, Castle Harbour and Tynes Bay, are particularly favourable. However, a poor dilution potential does not necessarily disqualify the sites, it merely means that in order to comply with the suggested acceptance criteria, the emissions of contaminants will have to be lower than they would have to be if the dilution had been better. The initial dilution of the discharged contaminants in the source compartments adjacent to either site becomes important. The volume of seawater in this Compartment is 340,000 m3 at Tynes Bay and 755,000 m3 at Castle Harbour.
153 11.
IMPACT ASSESSMENT
The calculated emission scenarios indicate that disposal of cement stabilized combined ash may reduce the maximum emissions (glday) of key contarninants by a factor of 9 (Cu), 43 (Pb), and 620 (Cd) at Tynes Bay and 6 (Cu),35 (Pb), and 540 (Cd) at Castle Harbour as compared to disposal of granular combined ash. If the environmental impacts of the various disposal options are to be evaluated using the acceptable load (AL) values developed in this study, it must be decided which level of security should be used in relation to the water quality criteria for single elements based on chronic effects (i.e. 10%AL, 50% AL or 100% AL). These acceptable loads are expressed as emission values (kglday or g/day) and apply to a specified volume of water (250 m x 250 m x the average depth), corresponding to a mixing zone (the smallest volume of resolution considered in the dilution model). The final decision of which set of criteria should apply is a matter for local consideration. In the following, only the most restrictive of the above mentioned emission criteria (10% AL) are considered. In table 7, the calculated average excess concentrations of Cd, Cu, and Pb (and for the cement block scenarios 1 and 4 also As and Zn) in the water in the innermost compartment (the source compartment covering 250 m x 250 m, closest to the site) resulting from the maximum daily emissions are calculated and compared to the existing average background concentrations (data exist only for Cd, Cu and Pb) and the acceptance criteria (AC). None of the predicted average excess concentrations of trace elements for any of the scenarios at Tynes Bay or Castle Harbour exceed 5 % of the existing average background concentrations (most of them are much smaller), and none of the existing average background concentrations exceed 30 % of the acceptance criteria level. A comparison of the predicted maximum daily emissions of selected trace elements with the acceptable load criteria based on chronic effects of single elements (AL and 10% AL) is presented in table 8. The results indicate that the maximum emissions of trace elements for all of the scenarios will be at least 5 times lower than the most restrictive of the criteria (10% AL). The emission of Cd in scenario 2, Tynes Bay, and scenario 5, Castle Harbour, come closest to the criteria values (10% and 20 %, respectively, of the 10% AL val
It can be calculated that dilutions of 40-100 times and 2700-6300 times of the strongest leachates produced in the laboratory tests are necessary to avoid acute and chronic toxic effects, respectively. Using the dilution model, it can be calculated that the average dilution of the leachate produced in scenarios 2 and 5 (the most critical scenarios) will be approximately 160,000 times (Tynes Bay) and 84,000 (Castle Harbour) within an area of 250 m x 250 m adjacent to the disposal site. No chronic effects should therefore occur outside this area even if the contaminants are not totally evenly distributed. Due to the chosen grid size (250 m by 250 m) used in the mathematical modelling of the dilutions at Tynes Bay and Castle Harbour, only the average resulting contaminant concentrations inside the volume of water corresponding to the source compartment adjacent to the disposal sites can be estimated.
154
Table 7 Calculated resulting maximum excess concentrations of selected trace elements in the innermost compartment of the model (250 m x 250 m) compared to existing average background concentrations at both sites and to acceptance criteria (AC) in terms of trace element concentrations in the water bodies.
SCENARIO
ELEMENT
AVERAGE BACKGROUND CONC.
Pull
CALCULATED EXCESS CONC. P8/ I
ACCEPTANCE CRITERIA (AC)
1(9/1
T m e s Bay
0.000018 0.000001 0.00057 0.000063 0.00033
5 0.1 1 .o 5.6
0.2 < 0.07
0.00073 0.0049 0.0027
0.1 1 .o 5.6
0.03 0.2 < 0.07
0.00001 1 0.00081 0.00019
0.1 1 .o 5.6
0.03 0.2 < 0.07
0.000041 0.000002 0.0013 0.00014 0.00072
5 0.1 1 .o 5.6
1 1 1 1 1
As Cd cu Pb 2n
Scenario 2 scenario 2 Scenario 2
Cd
0.03
cu
Scenario Scenario Scenario Scenario Scenario
Scenario 3 scenario 3 Scenario 3
Pb Cd
cu Pb
0.03 0 .2 < 0.07
C a s t l e Harbour
4 4 4 4 4
AS
Scenario 5 Scenario 5 Scenario 5
Cd
0.03
cu
0.2 < 0.07
0.0014 0.0079 0.0049
0.1 1 .o 5.6
scenario 6 Scenario 6 Scenario 6
Cd
0.03 0.2 < 0.07
0.00001 7 0.0017 0.00033
0.1 1.0 5.6
Scenario Scenario Scenario Scenario scenario
Cd
cu Pb 2n
Pb
cu Pb
Therefore, possible areas or spots within that compartment with concentrations exceeding the acute and chronic toxicity levels cannot be defined. Consequently, with the mathematical model used it is not possible to predict whether or not the emissions may cause an acute or chronic toxicity effect in a limited area (very) close to the disposal site (e.g. within a distance of a few meters from the site). This will depend very much on the disposal scenario chosen (disposal of cement stabilized material causing the least effects) and on the actual design of the disposal site which may have great influence on the initial dilution. In conclusion it may therefore be stated that, given the uncertainties described and the imperfections of the techniques employed, all of the three disposal options at each of the two site locations considered appear to comply with the proposed acceptable load (10% AL) outside a 250 m by 250 m compartment adjacent to the sites. The size of this compartment is determined by the resolution of the model and it is most likely that the area within which there is a risk of exceeding the acceptance criteria in reality will be much smaller. In fact, the data indicate that for the stabilized ash scenarios (1 and 4), the body of water constituting the innermost compartment could probably be reduced to less than 2 % of its present size and still be in compliance with the suggested criteria. For the granular ash scenarios (2 and 5), a similar
reduction to 10 and 20 % of the present size of the innermost compartment appears possible without the 10 % AL being exceeded. Table 8 Calculated emissions of selected trace elements compared to acceptable loads (AL) for each scenario.
ELEMENT
SCENARIO
PREDICTED MAX I MUM EM1SSlON (PME) gfday
ACCEPTABLE LOAD (AL)
SUGGESTED
PME/lO% A t
LIMIT 10% AL
g f day
gf day
%
Tvnes Bay
1 1 1 1 1
AS
Cd cu Pb Zn
0.16 0.01 4.9 0.54 2.8
600 6800 47000
60 680 4700
0.018 0.72 0.011
Scenario 2 Scenario 2 Scenario 2
Cd cu Pb
6.2 42 23
600 6800 47000
60 680 4700
10 6.2 0.49
Scenario 3 Scenario 3 Scenario 3
Cd
0.1 6.9 1.6
600 6800 47000
60 680 4700
0.17 1 .o 0.034
0.16 0.01 4.9 0.54 2.8
270 3100 22000
27 310 2200
1.6 0.025
Scenario Scenario Scenario Scenario Scenario
cu Pb
Castle Harbour
4 4 4 4 4
AS
Scenario 5 Scenario 5 Scenario 5
Cd
5.4
cu Pb
31 19
270 3100 22000
27 310 2200
20 9.9 0.88
Scenario 6 Scenario 6 Scenario 6
Cd cu Pb
0.07 6.7 1.3
270 3100 22000
27 310 2200
0.26 2.1 0.060
Scenario Scenario Scenario Scenario Scenario
Cd cu Pb Zn
Acceptable load (AL):
0.040
Calculated emission which u i l l b r i n g the average concentration o f t r a c e elements i n the uater i n the innermost compartment o f the model (250 m x 250 m) t o the acceptance Level.
10% AL:
10% o f the above mentioned emission (suggested c r i t e r i a ) .
PME/lOX AL:
The Pr edicted naximm m i s s i o n o f t r a c e elements r e l a t i v e t o the suggested acceptable Load value, 10% A t , expressed i n percentages. A value above 100 % indicates t h a t the chosen c r i t e r i a value has been exceeded.
The suggested acceptance criteria are based on toxicity data and are aimed primarily at the protection of fish, shellfish and wildlife in the water body. Other considerations may lead to more restrictive criteria e.g. concerning the emission of Pb. This should not be prohibitive in relation to the ash disposal since the predicted emissions of Pb in no case exceed 1 % of the most restrictive criteria (10%AL). The proposed supplementary requirements that ash leachate having a composition corresponding to the strongest concentrations produced in the laboratory leaching tests should be diluted approximately 100 times to avoid acute toxic effect and
156
approximately 6300 times to avoid chronic effects are based on the results of the bioassays performed on local marine organisms. The criteria and requirements have been conservatively set with respect to toxicity (and existing background concentrations). They do not, however, directly consider potential effects related to long term fate and accumulation of contaminants. The monitoring programme mentioned below should include activities addressing this issue. From an environmental point of view, the most favourable disposal option considered is scenario 1 (disposal of cement stabilized combined ash blocks at Tynes Bay) followed by scenario 4 (disposal of cement stabilized combined ash blocks at Castle Harbour). The least favourable solutions are scenarios 2 and 5 (disposal of granular combined ash by backtipping at either site). It should be noted that a substantial amount of organic material (NVOC) was leached from the granular ash. This could cause some concern and could probably be mitigated considerably in the case of Bermuda if care is taken to ensure a good burnout in the incinerator. Mercury which is particularly undesirable in the environment is leached from the granular ash in minor but measurable quantities, and objects containing mercury (e.g. small batteries and thermometers) should be prevented from entering the feed stream to the incinerator to the greatest extent possible. It should also be noted that there is an existing metals dump at one of the potential locations of the future incinerator ash disposal sites, the Civil Air Terminal at Castle Harbour. This study has shown that the "groundwater" or leachate within the dumpsite is slightly contaminated with organic compounds and trace elementsheavy metals. The bioaccumulation studies indicate that the dumpsite is already having a slight impact on the accumulation of trace elements, particularly lead in mussels deployed in Castle Harbour near the site. The predictions are based on the assumption that good engineering practice prevents the escape of particulate matter from the disposal site. The calculations pertaining to the cement stabilized ash disposal scenarios are based on the assumption that the civil engineering capability to produce mechanically durable 1 m3 cement stabilized ash blocks is available.
12. SAFETY MARGINS AND UNCERTAINTIES The approach of this study has been to attempt to simulate and predict the release and subsequent environmental impact of contaminants from various MSWI ash disposal scenarios. The basic idea is that whereas the release of contaminants cannot be totally prevented, the release may indeed be controlled through the application of appropriate precautions (e.g. cement stabilization) and kept at a rate which is acceptable to the environment. Since environmental acceptability is not a universally well defined concept, the study has included a discussion of acceptability criteria. An actual set of acceptance criteria for Bermuda has been proposed, and the disposal scenarios considered have been evaluated in relation to these criteria. As a consequence of the choice of methodology the conclusions of the study are based on a series of assumptions, choices, theoretical considerations, model calculations, and results of practical tests and field investigations. This adds an element of uncertainty to the study and the results. Since the overall objective has been to achieve an environmentally acceptable and sustainable solution, it has been attempted to adopt a conservative approach in most cases where
157
a choice has been necessary. This means that safety margins have been included at several points of the study. In order to facilitate an evaluation of the conclusions and results, this chapter presents an overview of the major choices made, the major points of uncertainty of the study, and the major safety margins included. In an attempt to further illustrate and somehow quantify the degree of safetyhncertainty associated with the conclusions of the study, the key operationsldecisions have been identified and evaluated for each of the 6 scenarios considered. Each of these key steps of the project has been assigned a safetyhncertainty factor. A factor of 1 means that the operation or decisiodestimation in question is considered realistic and is not expected to contribute any significant degree of uncertainty (nor is it expected to provide any additional safety margin) to the study. A factor higher than unity designates that the operation in question is believed to be on the safe side, i.e. the number is indicative of a certain margin of safety. The size of the factor represents a (rough) quantitative evaluation of the relative degree of exaggeration or conservatism applied at this point of the study. A factor smaller than unity indicates that the operation in question may be associated with a certain degree of uncertainty. The larger the uncertainty is believed to be, the smaller will the factor be. In order to obtain an overall evaluation of the safety marginhncertainty of the study, all the individual factors, including the result of the environmental impact assessment, i.e. the ratio between the suggested acceptable load and the calculated emissions, are multiplied with each other for each scenario. Table 9 summarizes all the individual safetyhncertainty factors and presents the resulting products for each scenario. This exercise exhibits a strong element of subjectivity but it does allow at least a comparison of the degree of safety or the safety margin associated with each of the scenarios. The higher the resulting product, the "safer" are the conclusions in question (from an environmental point of view). If the results are evaluated in absolute terms, a product of 1 indicates a result which is realistic and satisfactory - provided all important (factors) have been accounted for. It is therefore more comfortable to have a resulting product higher than 1 since this would provide some insurance against errors and uncertainty factors which might have been overestimated or overlooked. A resulting product < 1 may be acceptable - it is just associated with a degree of uncertainty which is higher, the smaller the product is. A resulting product of unity or lower could possibly be increased by looking more closely at (and possibly by changing) one or more of the individual steps involved in the procedure and thereby perhaps reduce the uncertainty and increase the size of the factor(s). Using the resulting overall safety factors calculated in table 9, the study of the various scenarios may be ranked in order of decreasing safety or increasing uncertainty: Rank
Option
1
Scenario 1 Scenario 4 Scenario 3 Scenario 6 Scenario 2 Scenario 5
2 3 4
5 6
Resulting safety factor (stabilized ash, Tynes Bay) (stabilized ash, Castle Harbour) (combined solution, Tynes Bay) (combined solution, Castle Harbour) (granular ash, Tynes Bay) (granular ash, Castle Harbour)
325 150 10 4.8
2 1
Table 9
Overview of individual safety/uncertainty factors and resulting products for each diposal scenario.
Safetv or risk factor for each d m s a l scenario Tynes Bay Castle Harbour 1 2 3 4 5 6
KEY OPERATIONS/DECISIONSOF THE STUDY A. Using Jersey & Saugus ash (instead of future Bermuda ash)
1
1
1
1
1
1
B. Selecting only key contaminants for leachate analysis and subsequent evaluation
0.5
0.5
0.5
0.5
0.5
0.5
C. Using accelerated laboratory data instead of actually observed full scale data (not available)
0.5
0.1
0.1
0.5
0.1
0.1
D. Using only the most contaminated initial leachate fmm the column tests for the bioassays
100
20
20
100
20
20
E. Selecting specific organisms and specific test methods for the bioassays
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
0.1
G. Lowering the suggested acceptable load to 10 96 of the calculated value
10
10
10
10
10
10
H. Risk of increased contaminant release due to cracking of stabilized blocks
F. Using modified US EPA Water Quality Criteria which only relate to toxicity to organisms living in the water body (but using the criteria for chronic rather than acute toxicity)
0.1
1
0.1
0.1
1
0.1
1.
Safety margin due to the use of constant flux in the calculation of the emission from blocks
10
1
10
10
1
10
I.
Using maximum daily contaminant emissions rather than averaged values in the impact assessment
1
2
1
1
2
1
K. Modelling of dilution
1
1
1
1
1
1
L. The result of the environmental impact assessment (suggested acceptable load/calculated emissions)
130
10
100
60
5
47
A-L The overall safety marein (the Droduct of all safetvlrisk factors) rvnes Bay Scenario 1: Scenario 2: Scenario 3:
Disposal of blocks of cement stabilized combined ash Disposal of untreated (granular) combined ash Disposal of stabilized ash blocks below sea level and untreated ash above sea level
4 R
Castle Harbour Scenario 4: Disposal of blocks of cement stabilized combined ash Scenario 5 : Disposal of untreated (granular) combined ash Scenario 6: Disposal of stabilized ash blocks below sea level and untreated ash above sea level
159
The above ranking is identical to the ranking according to the results of the environmental impact assessment alone (step L).
13. CONCLUSIONS A methodology for assessment of the potential environmental impact of marine land reclamation with MSWI residues has been developed and demonstrated in a study in Bermuda. For various land reclamation scenarios the rate of release of potentially harmful components into the surrounding sea has been estimated by combining the results of laboratory leaching tests on granular and cement stabilized MSWI ash with hydraulic models describing the flow of water through a site. Baseline field studies have been performed to describe the ambient marine environment near the proposed land reclamation sites. Bioaccumulation studies and ecotoxicological tests have been performed on local marine organisms. A set of acceptance criteria for discharge of trace elements into the sea has been developed and proposed for Bermuda. The dilution of the released trace elements in the marine waters surrounding the sites has been predicted using a dynamic mathematical simulation model, and the environmental impact of each scenario has been assessed. The estimated environmental impact of land reclamation with granular combined ash was not surprisingly substantially higher than the estimated impact of land reclamation with cement stabilized combined ash.
Even though the assessment procedure described in this paper is based on a substantial amount of practical experience and dedicated experimental work and modelling, it is still associated with a significant degree of uncertainty. In order to reduce this uncertainty and improve the assessment ability, it is important to monitor new land reclamation sites closely and to review the assessment procedure when actual full scale data become available.
14.
ACKNOWLEDGEMENT
The work described in this paper was funded by the Ministry of Works and Engineering, Hamilton, Bermuda. The views expressed are those of the authors and does not necessarily express the views of the Ministry.
REFERENCES Hjelmar, 0. (1987): Leachate from incinerator ash disposal sites. In: Proceedings of the International Conference on Municipal Waste Incineration, NITEP, Montreal, Canada, October
1-2,1987. Hjelmar, 0. (1989):Characterization of Leachate from Landfilled MSWI Ash. In: Proceedings of the International Conference on Municipal Waste Combustion, Hollywood, Florida, April 1 1-14,
1989. Hjelmar, 0. (1991):Field Studies of Leachates from Landfilled Combustion Residues", Presented at WASCON '91,Environmental Implications of Construction with Waste Materials, Maastricht, The Netherlands, November 10-14,1991.
Hjelmar 0. (1992): Municipal Solid Waste Incinerator Flue Gas Cleaning in Denmark: Residue Properties and Residue Management Options, in Proceedings of ISWA Specialized Conference on Incineration and Biological Waste Treatment, Amsterdam, The Netherlands, September 1-3, 1992. Hjelmar, 0. (1993): Leachate Management at Incinerator Residual Landfills. In: Proceedings of the 9th International Conference: Incineration: The Issues; Cork, Ireland, November 4-5, 1993 (in press). Water Quality Institute: Assessment of the Environmental Impact of Incinerator Ash Disposal in Bermuda. Final Report. Prepared for Ministry of Works and Engineering, Hamilton, Bermuda. Hersholm, Denmark, 1993. Bermuda Biological station for Research, Inc.: Marine Environmental studies to Determine the Impact of the Mass Burn incinerator Proposed for Tynes Bay, Bermuda. Report prepared for Ministry of Works and Engineering, Hamilton, Bermuda. Ferry Reach, Bermuda, 1991. Danish Hydraulic Institute: Bermuda, Ash Disposal. Hydraulic Studies. Hersholm, Denmark, 1989. Jickells, T.D and A.H. Knap (1984):The distribution and Geochemistry of some Trace Metals in the Bermuda Coastal Environment. Estuarine, Coastal and Shelf Science (18), pp 245-262. US-EPA (1984): Estimating "Concern Levels" for Concentrations of Chemical Substances in the Environment. Environmental Effects Branch. Health and Environmental Review Division. Norton, S., M. McVey, J. Colt, J. Durda and R. Hegner (1988): Review of Ecological Risk Assessment Methods. US-EPA, Oftice of Policy, Planning and Evaluation. EPA/230- 10-88-041.
Environmental Aspects of Conshrction with Waste Materials JJ-IM. Goumans, H A . van der SIoot and Th.G. Aalbers (Editors) 01994 Elsevier Science B.V. All rights reserved.
161
Quality Assessment of Granular Combustion Residues by a Standard Column Test: Prediction versus Reality M. Janssen-JurkoviEovaa, G.G. Hollmanb, M.M. Nass' and R.D. Schuilingb "KEMA Environmental Services, KEMA, P.O. Box 9035, 6800 ET Amhem, The Netherlands bDepartmentof Geochemistry, Institute of Earth Sciences, P.O. Box 80.021, 3508 TA Utrecht, The Netherlands 'Department of Soil Science and Plant Nutrition, Wageningen Agricultural University, P.O. Box 8005, 6700 EC Wageningen, The Netherlands
Abstract The laboratory leachability tests introduced by the Dutch Standardization Institute (NNI) as draft standards NEN 7341 and 7343 for the analysis of the leaching behaviour of granular materials and/or fly ashes, will play an important role in the near future. Particularly the column test will play a decisive role in the Decree on Building Materials currently in preparation, concerning whether or not to permit the application of these materials in constructional works. In view of the role this test will play, it is important to determine the degree of correspondence between leaching as 'simulated' in the column test and leaching occurring in reality. However, the principles underlying the present column test are not included in the introductory text to draft standard NEN 7343. This lack of information prevents us from understanding exactly which practical conditions are being simulated in this test. The results presented here show that there is a great discrepancy between the leaching of elements in the column test and in a natural environment. Hence, it is necessary first to translate the results obtained in the column test into realistic field conditions. This article will discuss one option by which this could be accomplished. The results also demonstrate that the current availability test (NEN 7341) is not adequate to determine the maximum amount of elements available from a certain material for leaching.
162
1. INTRODUCTION
The Decree on Building Materials currently in preparation requires that all primary and secondary building materials be provided with a quality statement before admitting them to the building market. For most secondary building materials this means that their leaching characteristics should be determined according NEN 7343. The column test described in NEN 7343 is based on the idea that this test simulates leaching of granular materials over a medium-range period of about 30 years, while the cascade test simulates leaching of these materials over a period of several hundred years. It should be possible by means of the so-called availability test (NEN 7341) to determine under natural conditions the maximum amount of elements available from a building material for leaching in the very long term under extreme conditions. The development of standard leachability tests has taken over ten years of efforts from many institutes. These efforts have resulted in the development of tests that are highly reproducible. In view of the ‘simulation function’ that was assigned to the column test in a later stage and the role this test will have to play in the decision whether or not to admit e.g. secondary granular materials to the building market, it is however of great importance to determine: - whether this test is really adequate to simulate leaching over a certain period of time, and moreover - how the test conditions selected affect the intended simulation, if one is to establish whether the simulating (timescale) function ascribed to this test has in fact a realistic or merely an arbitrary value. This article will discuss a suitable test method to determine this. Furthermore, this report will present fairly recent findings resulting from a comparison of leaching values obtained in standard leachability tests with those obtained under field conditions of pulverized fly ash from a power station.
2.TESTlNG OF STANDARD LEACHABILITY TESTS ASCRIBED TO THESE TESTS
FOR THE VALUE
2.1 Full Standard Leachability Test The full standard leachability test comprises a column test, a cascade test and an availability test. In the column test demineralized water (pH4) passes in upward direction through a column containing fly ash. The percolate is collected in seven fractions: after L/S (i.e. Liquid/Solid) 0.1; 0.2; 0.3; 0.5;1.0;5.0;and 10.0. Based on the sum of the fractions an estimate is made of the total amount of elements leached from the fly ash tested in its natural environment over a medium-range period of about 30
163
years. In the cascade test the material is extracted five times at an L/S of 20 in acidified demineralized water (pH4) for a maximum period of 23 hours. The test serves to predict long-term leaching over a period of several hundred years. In the availability test the fly ash to be tested is extracted in demineralized water (L/S lOO),first at a constant pH of 7 and the second time at a constant pH of 4, each time for a period of three hours. The availability test serves to determine the maximum fraction of an element available for leaching in the natural environment.
2.2 The lest Procedure Table 1 lists the total amounts of elements in the fly ash used in the test procedure. Figure 1 represents the procedure for testing the influence of the standard leachability tests on element immobilization and/or mobilization in the granular materials subjected to these tests. This figure demonstrates that the fly ash was subjected first to the standard column test and subsequently to the cascade test.
Table 1 Total amounts of elements in the flv ash sample examined. Amounts of elements in mg/kg As
Mo
Se
27.7
9.3
16.1
Ba 2,166
Fe 61,693
v 332
The availability of elements was determined at the start of the test, after the column test and after the cascade test. Element immobilization in the column test was calculated according to the following equation:
in which ALl = amount immobilized during the column test co = available before column test Abefore = leaching during column test, and L o = availability after column test. Aafter w
164
Q fly ash
availability test
column test
availability test
el-El
cascade test
availability test
Figure 1. Outline for research on the influence of standard leaching tests on the (im)rnobilization of elements.
To determine the influence of the cascade test on the material already subjected to the column test, equation (1) was adapted as follows: AL =
kfter w
L a - Aafter ca
(2)
in which amount immobilized during the cascade test available after column test (i.e. before cascade test) leaching during cascade test, and = availability after cascade test.
Alca
= = =
Aafler
L, Aafter cB
The percentage of element immobilization during the column test was calculated using:
%Aim
=
k x p e c t e d after w
- hetermined after w
-%elore
Wl100
(3)
165
To calculate the immobilization percentage during the cascade test, equation (3) was adapted as follows:
%Aim =
Aexpected afler ca
- betermined afler ca
(4)
-Aafler cn/100 in which Aexmcted afler ca
=
betermined afler w
(5)
-
Findings of the Test Procedure Table 2 contains the data obtained by the procedure represented in Figure 1.
2.3
Table 2 Available and/or leached amounts of elements in the standard leachabilitv tests. type of test in consecutive order
leached and/or available in pg/kg As
Mo
Se
Ba
Fe
v
availability in material at start of test
18,350 10,600 10,200 column test
10
60,700
1,350
5,020
590
9,750
200
76,840 250
"availability after column test" -expected"
18,340
5,580
9,610
50,950
1,150
76,590
-determined
1,720
5,160
7,220 164,340
2,640
19,260
cascade test
420
1,240
6,030
42,160
8,450
6,590
"availability after cascade test" -expected*)
1,300
3,920
1,190 122,180
0
12,670
-determined
7,370
3,440
5,040 260,200
9,760
30,440
as might be expected on the basis of the leaching values found in the column test *) on the basis of kfler cn - L, Table 3 shows the influence of the standard column and cascade tests on element immobilization and/or mobilization calculated with equations (1)and (2). Table 4 gives the percentage values of the amounts of immobilized elements.
166
Table 3 Element immobilization (-) and/or mobilization (+) during the standard leachability tests in pg/kg. tvDe of test
As
Mo
Se
Ba
Fe
V ~~~
’)
column
- 16,620
cascade”
+ 6.070
- 420 - 480
- 2,390
+ 113,390 + 1,490
+ 3,850 +
138,020
- 57,330
+ 9.760 + 17.770
in respect of the column test
Table 4 Element immobilization (-) and/or mobilization (+) during the standard leachability tests in %. type of test column cascade‘’ ’)
As
Mo
Se
- 90
-4
- 23
+ 353
-9
+ 53
+
Ba
Fe
V
187
+ 110 + 370
- 75
+ 84
+ 92
in respect of the column test
Discussion of the Findings Obtained from the Test Procedures 2.4 On the basis of the leaching tolerated by the Decree on Building Materials for Soil Protection, the application of a (secondary) building material as examined in the present study would not be allowed (without any restrictions). The maximum permissible leaching values for Mo and Se would be exceeded. It is the question though, also with a view to the significance of such an observation, if this application restriction (or rejection) would indeed be justified. In other words, a decision on this matter requires that the time-simulating function of the standard leachability tests, and especially that of the column test, should be motivated. Element Immobilization during the Column Test and in Practice Table 3 shows that immobilization of As, V and, to a lesser degree, Se already occurred during the 3-week column test. This means that availability of these elements after the column test is lower than might be expected on the basis of the standard leachability tests. For As as much as 90% of the total available amount is immobilized in the column test and for V this is 75%. The element Se appears to have been immobilized for 23% only, while Mo has hardly been immobilized at all (Table 4). From Figure 2,which represents a comparison between leaching of Mo in a field lysimeter and in the column test [I], it can be derived that immobilization of Mo under field conditions already occurs within a short period. Figure 2 shows that
167
exposure to natural weathering conditions results in a decrease of the cumulative amount of Mo leached in the column test and the field lysimeter; i.e., the decrease in amounts of Mo leached in column tests, carried out at increasing L/S ratio's in the field lysimeter, is greater than the amount of Mo leached during this time in the lysimeter : natural weathering leads to immobilization of a part of the Mo present. Between L/S 2.8 and L/S 4.1 the amount of Mo immobilized according to figure 2 is 0.66 mg/kg. Compared to a fresh alkaline fly ash of similar type, the results indicate an immobilization for Mo of 1.45 mg/kg after 8 years of natural weathering.
0 ' 0
1
I
I
I
2
4
6
I
8 Liquid/Solid
I
1
I
10
12
14
Lysimeter Column t=O Column t=4 Column t=6 Column t=8
I
Figure 2. Leaching of Mo from fly ash under field conditions and from this fly ash in a column test after 4, 6, and 8 years of weathering. The fact that immobilization of the elements stated occurs, under field conditions in fly ash types similar to that used in the examination, and in even higher degrees, is illustrated in the Figures 3 and 4. These figures give concentrations of various elements as a function of the depth in a fly ash deposit where activities were stopped as long as 20 years ago. Figure 3 shows that immobilization of As and Se occurs at the same level as that of the macro-elements Fe, Mg and Ca, while the elements Ba, Mo and V correspond to the levels of the macro-elements Al and Si (Figure 4).
168 Concentration (rng/kg)
Concentration (rng/kg)
200.
a5 150. u)
poo50. O!
I
4
I
10
2.01+
3.0
1 .o
-
0.0’
0 Depth (m)
Figure 3. Concentration of Fe, Mg, Ca, As and Se in the solid phase as a function of depth in a 20 year old fly ash deposit.
.
.
.
.
.
.
. 0,5 1.0 1,5 2,O 2,5 3,O
I
Depth (m)
Figure 4. Concentration of Si, Al, Ba, V and Mo in the solid phase as a function of depth in a 20 year old fly ash deposit.
169
Coprecipitation of trace elements with macro-elements in the environment is a phenomenon known from geology. As an example can be mentioned the high metal content in iron concretions in the Dutch soil [2].The fact that there is a correspondence between leaching of As and of Fe can be corroborated also by findings from our own research [3]. The Influence of the Cascade Test on Leaching The cascade test mainly has a mobilizing effect on the trace elements examined. For instance, the immobilization of As, Se and V that occurred in the column test would be decreased by the mobilizing effect of the cascade test. The mobilization of Fe and Ba that started in the column test would continue reinforced. On the basis of the elements examined it can be stated in general that in a cascade test more leaching occurs than in a column test. Determination of Availability for Leaching As stated earlier, the availability test according to NEN 7341 is assumed to yield the (very) long term emission (in mg/kg) that might occur under field conditions. This means that the availability established at the start of the test i.e. "before the column test" should have the highest value. Leaching in the column and/or cascade test should subsequently reduce the availability, depending on the extent of leaching. Table 2 demonstrates, however, that element availability after the column and/or cascade test does not decrease proportionately to the extent of leaching (e.g. for As, Se and v). The most remarkable finding is the increase of the initially determined availability due to the mobilizing effect of the column test for the elements Ba (factor 3.2) and Fe (factor 2.3). Availability of these elements continued to increase due to the influence of the cascade test: for Fe, which had already been leached for more than the total available amount, nearly 10 mg/kg still appeared to be available. Another remarkable fact is that of a given element a (far) greater amount can be leached in the column and cascade test than was initially leachable according to the availability test (Fe, factor 6.4), whereas the exact opposite occurs in nature.
3. RELATION OF THE L/S RATIO TO A TIMESCALE
The text accompanying draft standard NEN 7343 provides no information on the background of this test. This lack of information prevents us from establishing which practical conditions are simulated in this test. By way of example, this study compares the leaching values obtained in the standard column test and those obtained for a similar fly ash under (semi-)practical conditions (lysimeters with an effective layer thickness of 0.9 m, 1 L/S = 1.7 years). The values in Tables 5 and 6 show that (based on knowledge available so far) no timescale for practical conditions can be assigned to the L/S value of the standard column tests. Thus the amount of elements leached in the column test will not be
170
equal to that leached over a period of 30 years under field conditions. For some elements the amount leached will be smaller, for other ones larger. Previous results [3] indicate that most of the anions, those of Mo being an exception, are less leachable in the column than under field conditions, while for cations this depends strongly on the element itself. This stresses the importance of finding an explanation for the contradictions observed. Table 5 Cumulative leaching of major elements from alkaline fly ash in a column test (mg/kg).
K
Mg
Na
20 0.5 0.25 12.8 0.10 624 0.01 1.0 0.5 12.8 0.17 1235 0.02 35 2.1 1 12.8 0.24 2259 0.03 51 2 12.9 0.32 3925 0.04 100 4.2 5 12.7 0.79 6024 0.17 151 10.4 10 12.7 8.78 7157 0.66 175 20.8
0.14 0.27 0.41 0.59 1.02 1.45
124 220 298 478 596 630
Time L/S days kg/kg
pH
Al
Ca
Fe
P 0.17 0.30 0.43 0.69 0.84 1.00
Si
S
215 215 364 364 630 630 1188 1188 1677 1677 1733 1733
Table 6 Cumulative leaching of major elements from alkaline fly ash in a field lysimeter (mg/kg)'. Time L/S days kg/kg
pH
Al
211 0.429 475 0.782 1202 2.176 1279 2.437 1296 2.542 1489 2.673 2000 3.352 2070 3.457 2269 3.633 2413 3.842 2672 4.039 2798 4.143
10.1 10.9 8.5 7.5 10.1 10.4 8.3 9.9 7.8 7.6 7.5 8.6
15 19 25 25 26 26 27 27 28 28 28 29
~________
*
Ca
Fe
Na
K
-
-
4 13 131 163 1 76 188 295 312 337 376 412 426
144 195 307 322 326 332 353 358 368 379 389 393
0.23 0.27 0.29 0.32 0.32 0.32 0.32 0.32 0.33 0.33 0.33 0.33 ____
0.01 0.02 0.13 0.13 0.14 0.14 0.16 0.17 0.17 0.17 0.26 0.27 ~_______
156 186 234 239 241 243 253 256 262 267 272 274
P
Si
0.08 0.08 0.28 0.28 0.28 0.28 0.28 0.28 0.28 0.28
3.9 6.3 9.6 10.1 10.4 10.7 10.8 11.0 11.4 11.7 NA 12.0 NA 12.3 __
= Results obtained from the same type of fly ash as presented in Table 5.
S NA NA NA NA NA NA
> 29
>49 NA NA
> 74
> 88
171
4. POSSIBLE EXPLANATIONS FOR THE DISCREPANCY BETWEEN LEACHING IN THE STANDARD COLUMN TEST AND UNDER FIELD CONDITIONS
The factor time was pointed out as the principal cause of the discrepancy between leaching in the column test and under field conditions. This factor determines the degree in which the influence of the various other factors in a laboratory test or in a fly ash layer is expressed. The duration of a standard leachability test is far too short for certain immobilization processes that are important under field conditions. In the course of decades or even centuries the natural conditions will exert their influence on fly ash leaching by the continuation of such processes as humification, weathering and the formation of secondary minerals. This will result in a modification of qualities such as the pH, Eh, specific surface and the adsorption complex. The findings of this study demonstrated that for fresh fly ashes the parameters specific to fly ash ageing, such as specific surface (0.7 m’/g), pore surface (4.5 m’/g) and cation exchange capacity, CEC (1.4 meq/100 g) did not change during the three-week column test (L/S lo), whereas for fly ash after four years of leaching in lysimeters (L/S approx. 2.7) these values changed significantly, to 19.6 m’/m, 15.2 m’/g and 6 meq/100 g for specific surface, pore surface and CEC, respectively. SEM photographs show that after the column test the fly ash particles are still smooth, while those in the lysimeters are covered with an irregular crust formed by formation of new minerals and amorphous phases (Figure 5).
Figure 5. Left-hand side: ‘fresh’ fly ash; Middle: the same fly ash after the column test (L/S 10); Right-hand side: the same fly ash after a residence time of four years in the lysimeters (L/S 2.7).
172
Due to the high concentrations of elements in solution at the start of the weathering process and the resulting high supersaturation, the precipitation reactions were initially so fast that there was no time to develop clearly marked phases of a certain crystalline structure. Instead, an amorphous crust was formed with the combined stoichiometry of these minerals. The entire process has been documented in a qualitative weathering model as represented in Figure 6. According to this model fly ash weathering will occur in various phases. Phase 1 At the first contact between fly ash and water the oxides (especially CaO) and salts of high solubility that are present at the surface of the fly ash particles, dissolve. Moreover, there is a rapid exchange between ions in solution and ions at or near the fly ash surface. This phase is accompanied by a strong rise of the pH up to values exceeding 11. Phase 2 Due to the high pH value and hydratation, the glass phase starts to dissolve. Cations are detached from the glass network and dissolve. Mg, Al and Si precipitate on the surface of the fly ash particles, where they form amorphous crusts. The pH value decreases gradually. As the fly ash particles are divided into an exterior reactive and an interior, less reactive glass phase [4], the dissolution rate will decrease in time. Phase 3 It still is unclear if and how the crusts formed continue to grow afterwards. It is possible that in-situ conversion into the Mg-AI-Si phase, of which the crusts consist, takes place at the interface with the glass matrix. A probability is that the crusts provide a certain degree of protection against solution of the glass matrix and that the importance of diffusion processes increases. Phase 4 The eventual development of the crusts formed in the course of the ageing process is determined largely by the pH trend. In the pH range from 7 to 9-10 a restructuring takes place whereby O-Si-O-AI-OH chains are formed, materials from which clay minerals such as smectite and kaolinite can be formed. At a pH value exceeding 9 or 10 the formation of clay minerals is prevented by complexation of silica in solution, in which case zeolites are more likely formed. The precise development of these reactions during the ageing process is as yet unknown. Chemical analyses and analyses using a secondary electron microscope (SEM) on samples obtained from experiments carried out at 110, 150 and 200°C to speed up the weathering process, appear to confirm that the formation of zeolites is related to the crusts enclosing the fly ash particles. Results suggest that the formation of zeolites takes place at the interface between solution and crusts, with calcium, potassium, sodium and silicon polymers being taken up from solution and the crusts supplying aluminium, silicon and nucleation sites. When the pH value has become low enough due to the influence of weathering reactions, the formation of zeolites can be replaced by the formation of smectites. Analyses by X-ray diffraction, differential thermal and thermogravimetric analyses and infrared spectroscopy appear to indicate that the crusts indeed possess a structure bearing similarities to that of a clay mineral [5].
173
Phase 1 : Original fresh fly ash: solubilisation of salts and oxides soluble salts and oxides,
Mn*
CaO + H,O = Ca2* + 2 0 H -
H 2 0 , Na+ Phase 2: dissolution of glass phase, formation of crusts
Mg + ySi +zAl = Mg-Al-Si-phase
Phase 3:diffusion, continuing formation of crusts
Phase 4: transformation of crusts a) Ca2+, Na+, K', H,SiO,
pH < 9 a 10: formation of smectites
b) Ca2+,Na', K*, Si,0,(OH),4.
pH > 9 a 10: formation of zeolites
Figure 6. Model of weathering for alkaline fly ash.
174
Fly ashes after seven years of natural weathering in lysimeters show that phases 1 and 2 take place under natural weathering conditions. Whether phase 4 will
eventually be reached depends on the activities of the cations in solution reaching sufficiently high levels. This in turn depends on such factors as permeability and thickness of the fly ash deposit. However, the possibility of the formation of smectites and zeolites is confirmed by saturation indices calculated from the composition of pore water and percolates sampled from various fly ash deposits exposed to natural weathering. Table 7 gives an overview of the minerals calculated to become supersaturated and a comparison with those actually observed to be supersaturated. Table 7 Minerals predicted to be formed, according to calculated saturation indices [6] and data from the literature [7-91, and minerals observed in weathered fly ash [lo-111. Predicted Mineral name boehmite gibbsite Al-hydroxide (amorphous)
Approximate composition
AlOOH AI(W, AI(OH), AIOHS04 Fe-hydroxide (amorphous) Fe(OH), Fe-oxyhydroxide (am.) FeOOH hematite Fez03 magnesiumferrite MgFe,O, calcite CaCO, magnesite MGO, an hydrite CaSO, gypsum CaS0,'2H,O ettringite Cakl2(so,),(OH),,*2sH,O monosulphate Ca,AI,SO,(OH),,'GH,O portlandite Ca(OH), br uc it e MWHh whitlockite CWO,), apatite Ca,(PO,),OH quartz SiO, amorphous silica SiO, Mg-Ca-Na-K-Al-Si phase (amorphous) Al-Si phase (amorphous) proto-imogolite Ca-aluminate Ca-silicate Ca-zeolite CaAI,Si,O,,'yH,O phillipsite NaKAI,S&O,,'ZH,O feldspar (Na.K)AISi,O, halloysite AI,Si,O,(OH), MgAl-phyllosilicate smectite chlorite MgA~,Si,O,,(OH), illite 16,Mg,,kl,,Si,,O,,(OH), prehnite Ca@W,O,,(OH), kerolite Mg,Si,O,(OH), sepiolite MS,SbOtdOH),'WO mica KAI,Si,O,,(OH),
Field
Observed
Column Literature
This Study
Literature
t
t
+
t t
t
+ t
t
t
t t t
t
+
t
+
t
t
t
+
+ t
t
t
t
t
t
t
+
+
+
t
t t
+ t t t t
+
+
+
+
t
t
+
+ + t
+
t
+
+
t
t
+ t t
t
t t t
t
+ +
+
t
+ +
175
5. DEVELOPING A PREDICTIVE MODEL
On the basis of the results obtained it was attempted to develop a predictive model for (long-term) leaching of fly ash and other combustion residues. For the development of this model the approach represented in Figure 7 [9] will be adopted. This approach is based on an estimate of the composition of percolate from the residues, using the following input data: - total element concentrations in the residue and leachability of the element in question; - the different types of solid phase and their solubility; - the kinetics of solution and precipitation reactions; - adsorption and desorption reactions; - pH value, redox potential, ionic strength; - hydrology.
1 FLY ASH
Fly Ash Characteristics Total Chemical Composition Primary and Secondary Minerals and Gasses Extractabllity/Leachabilily(Eh, pH)
-
REACTIONS
THERMOCHEMICAL DATA
Precipitation Dissolution Reactions
Y
Adsorption Desorption Reactions
ADSORPTION
LEACHATE COMPOSTION
Figure 7 . Diagram for predicting the leaching behaviour of combustion residues [9].
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The modelling results obtained so far [6] demonstrate that the weathering processes in a fly ash deposit probably lead to immobilization of a significant part of the trace elements present. This is partly due to coprecipitation with and adsorption to aluminium and iron (hydr)oxides, smectites and possibly even zeolites. However, there is as yet insufficient information available on a number of input data to enable the development of a predictive model for leaching. This is especially the case for the kinetics of dissolution and formation reactions and for the development of adsorption and desorption reactions. Various literature sources assert that the rate of mineral formation or dissolution in fly ash is so high that it is fair to assume an approach based on thermochemical principles [E, 12, 13, 141. This would mean that the upper limit of the element concentrations in the percolate might be predicted on the basis of data on the solubility of certain solid phases. Our research, however, indicates that a number of major reactions are highly influenced by the factor time and are therefore kinetically determined [6]: - after the relatively rapid dissolution of the exterior reactive glass which reprecipitates into an amorphous crust, the remaining glass, which forms the bulk of the fly ash, dissolves only gradually. Another aspect involved here is that particles are protected by the amorphous crusts. - The amorphous phases formed, such as the crusts, will in time convert into crystalline compounds. - The saturation indices calculated indicate that some minerals (e.g. apatite and nontronite) are extremely supersaturated. Nevertheless, these minerals are not found in the weathered fly ash. This is due either to unknown complexation reactions or to the fact that crystallization of these phases requires more time and is therefore kinetically determined. The kinetics of the weathering processes have so far been insufficiently subjected to scientific analysis. However, these changes over time mean that the adsorption characteristics will also change over time. Desorption and adsorption processes at the surface of solid phases (e.g. iron (hydr)oxides) are highly important in regulating the concentrations of trace elements such as As, Cr, Se and Mo in solution. To be able to predict the leaching of these ecol
