Municipal solid waste incinerator residues

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Author: A.J. Chandler | T.T. Eighmy | O. Hjelmar | D.S. Kosson | S.E. Sawell | J. Vehlow | H.A. van der Sloot | J. Hartlén

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MUNICIPAL SOLIDWASTE INCINERATORRESIDUES

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Studies in Environmental Science 67

MUNICIPAL SOLID WASTE INCINERATOR RESIDUES The INTERNATIONAL c o m p r i s e d of

ASH WORKING GROUP,

(in alphabetical order):

A. John Chandler

David S. Kosson

T. Taylor Eighmy

Steven E. Sawell

A.J. Chandlerand Associates Ltd., Willowdale, Ontario, Canada University of New Hampshire, Ourham, New Hampshire, U.S.A.

Jan Hartl6n

Rutgers, The State University of New Jersey, New Brunswick, New Jersey U.S.A. Compass Environmental, Burlington, Ontario, Canada

Hans A. van der Sleet

Swedisch GeotechnicalInstitute, LinkEping, Sweden

Netherlands EnergyResearchFoundation, Petten, The Netherlands

Ole Hjelmar

JiJrgen Vehlow

VKI WaterQuality Institute, Hersholm, Denmark

1997 ELSEVIER Amsterdam

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Lausanne

Forschungszentrum Karlsruhe GmbH, Institute of TechnicalChemistry, Karlsruhe, Germany

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NewYork-

Oxford -

Shannon

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Tokyo

ELSEVIER SCIENCE B.V. Sara Burgerhartstraat 25 P.O. Box 211, 1000 AE Amsterdam, The Netherlands

ISBN 0-444-82563-0 © 1997 ELSEVIER SCIENCE B.V. All rights reserved. No part of this publication may be reproduced, stored in a retrieval system or transmitted in any form or by any means, electronic, mechanical, photocopying, recording or otherwise, without the prior written permission of the publisher, Elsevier Science B.V., Copyright & Permissions Department, P.O. Box 521, 1000 AM Amsterdam, The Netherlands. Special regulations for readers in the U.S.A. - This publication has been registered with the Copyright Clearance Center Inc. (CCC), 222 Rosewood Drive Danvers, Ma 01923. Information can be obtained from the CCC about conditions under which photocopies of parts of this publication may be made in the U.S.A. All other copyright questions, including photocopying outside of the U.S.A., should be referred to the 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

PREFACE

The International Ash Working Group (IAWG) was established in 1989 to conduct an in-depth review of the existing scientific data and develop a state-of-knowledge treatise on MSW incinerator residue characterisation, disposal, treatment and utilisation. The topics of operator and worker health and safety, and health risk assessment were beyond the scope of this project, and therefore have not been addressed. Members of the IAWG had been involved in various research and development programs concerning MSW incineration residues for several years prior to establishing the IAWG. The IAWG has met regularly since its inception to discuss aspects of residue characterisation and management, as well as offering a forum for other researchers to provide their perspectives on the issues. The project soon grew beyond the original scope, due in part to the need to examine the ever increasing volume of published research data which became available in the early 1990's. In addition, the IAWG project was designated as an Activity under the International Energy Agency's (lEA) Bioenergy Agreement Task Xl - Conversion of MSW to Energy 1991 - 1994. This final treatise and the Summary Report represent the culmination of the IAWG efforts over the period from February 1990 through July 1996. The input of information from colleagues, along with other information available from the literature and personal contacts, was used to formulate the conclusions and recommendations summarised in this document. The results of this effort have been presented in extended seminars, in conjunction with both the WASCON '94 Conference (June 1994) in Europe and with the Municipal Waste Combustion Conference (April 1995) in North America. In addition, the IAWG co-sponsored and participated in the "Seminar on Cycle and Stabilisation Technologies of MSW Incineration Residues" along with the Japan Waste Research Foundation (March 1996) in Japan. Currently, the IAWG continues to operate as a sub-group of Thermal Conversion Activity under the IEA's Bioenergy Agreement Task XlV - Energy Recovery from Municipal Solid Waste.


vii

AUTHORS A. John Chandler A. J. Chandler and Associates, Ltd. Willowdale, Ontario Canada T. Taylor Eighrny University of New Hampshire Durham, New Hampshire United States of America Jan Hartldn Swedish Geotechnical Institute Linkoping Sweden Ole Hjelmar Danish Water Quality Institute H~rsholm Denmark David S. Kosson Rutgers, The State University of New Jersey New Brunswick, New Jersey United States of America Steven E. Sawell Compass Environmental Burlington, Ontario Canada Hans van der Sloot Netherlands Energy Research Foundation Petten The Netherlands Jtirgen Vehlow Forschungszentrum Karlsruhe GmbH Institute of Technical Chemistry Germany

... VIII

THE INTERNATIONAL ASH WORKING GROUP A. John Chandler A. J. Chandler and Associates, Ltd. Willowdale, Ontario Canada

Shin-ichi Saki (Since 1994) Environment Preservation Centre Kyoto University Japan

T. Taylor Eighmy University of New Hampshire Durham, New Hampshire United States of America

Steven E. Sawell Compass Environmental Burlington, Ontario Canada

Jan Hartl6n Swedish Geotechnical Institute Linkoping Sweden

Hans van der Sloot Netherlands Energy Research Foundation Petten The Netherlands

Ole Hjelmar Danish Water Quality Institute H~rsholm Denmark David S. Kosson Rutgers, The State University of New Jersey New Brunswick, New Jersey United States of America

JQrgen Vehlow Forschungszentrum Karlsruhe GmbH Institute of Technical Chemistry Germany

DISCLAIMER This report was prepared by the International Ash Working Group (IAWG). The work was sponsored by the agencies listed herein, who are not necessarily in agreement with the opinions expressed by the IAWG. Neither the sponsoring agencies (including its members), nor the IAWG, nor any other person acting on their behalf makes any warranty, express or implied, or assumes any legal responsibility for the accuracy of any information or for the completeness or usefulness of any apparatus, product or process disclosed, or accept liability for the use, or damages resulting from the use, thereof. Neither do they represent that their use would not infringe upon privately owned rights. The IAWG also does not, and never intended to, discuss or make recommendations with regard to health and safety issues concerning facility operators or workers. Furthermore, the sponsoring agencies and the IAWG hereby disclaim ANY AND ALL WARRANTIES, EXPRESSED OR IMPLIED, INCLUDING THE WARRANTIES OF MERCHANTABILITY AND FITNESS FOR A PARTICULAR PURPOSE, WHETHER ARISING BY LAW, CUSTOM, OR CONDUCT WITH RESPECT TO ANY OF THE INFORMATION CONTAINED IN THIS REPORT. In no event shall the sponsoring agencies or the IAWG be liable for incidental or consequential damages because of the use of any information contained in this report. Any reference in this report to any specific commercial product, process or service by tradename, trademark, manufacturer or otherwise does not necessarily constitute or imply its endorsement or recommendation by the IAWG and the sponsoring agencies or any of its members.

SPONSORING AGENCIES The IAWG is grateful for the financial and technical contributions made to this project by the following agencies/organisations/companies:

Major Sponsors Asea Brown Boveri (Switzerland) Danish Ministry of Energy Energy, Mines and Resources Canada Environment Canada European Commission Forschungszentrum Karlsruhe (Germany) International Energy Agency International Lead Zinc Research Organization Integrated Waste Services Association (USA) Japan Waste Research Foundation LAB (France) Management Office for Energy and the Environment (Netherlands) National Institute of Public Health and Environmental Protection (Netherlands) Swedish National Board for Industrial & Technical Development Takuma Co., Ltd. (Japan) United Kingdom Department of Environment United States Environmental Protection Agency Wheelabrator Environmental Systems (USA)

Minor Sponsors American Society of Mechanical Engineers Greater Vancouver Regional District (Canada) Kubota Corporation (Japan) Northeast Waste Management Officials Association (USA) New Jersey Department of Environmental Protection (USA) Waste Processing Association (Netherlands (WAV))

TECHNICAL CONTRIBUTORS The IAWG gratefully acknowledges the technical contributions made during the course of this project by: T. Aalbers - RIVM, Netherlands M. Adams - VROM, Netherlands I. H. Anthonissen - RIVM, Netherlands J. A t w a t e r - University of British Columbia,Canada J. Barniske - Umweltbundesamt, Germany J. B e r r y - Wheelabrator Environmental Systems Ltd., USA S. B i n n e r - V~lund, Denmark R. B o e h m - PBI, Netherlands H. Borrmann - Forschungszentrum Karlsruhe, Germany J. P. B o r n - VVAV, Netherlands R. Braam - PBI, Netherlands S. Burnley - Energy Technology Support Unit, United Kingdom D. C h a m b a z - BUWAL, Switzerland A. Chamberland - Tiru Inc. (formerly with Montenay Inc.), Canada W. Chesner- Chesner Engineering, P.C., USA B. Christensen - Environment Canada S. C o o k - Bermuda Biological Station R. C o m a n s - ECN, Netherlands S. Dalager- dk TEKNIK, Denmark A. Damborg - Danish Water Quality Institute C. Dent - AEA Technology, United Kingdom A. M. F~llman - Swedish Geotechnical Institute A. Finkelstein - Environment Canada J. Fraser - Wastewater Technology Centre, Canada M. G a l l o - Rutgers University, USA D. Goetz - University of Hamburg, Germany J. G r o n o w - United Kingdom Department of Environment T. Guest - Montenay Inc., Canada L. Gullbrand - Swedish National Board for Industrial and Technical Development G. Hansen - United States Environmental Protection Agency D. Hay - Environment Canada S. Hetherington - Compass Environmental Inc., Canada F. Hoffman - Rutgers University, USA G. Hoffmann - Umweltbundesamt, Germany R. H u i t r i c - LA County Dept. of Sanitation, USA L. Johansson - Swedish Geotechnical Institute B. J o h n k e - Umweltbundesamt, Germany T. Kimura - Kubota Corporation, Japan J. Kiser- Integrated Waste Services Association, USA R. Klicius - Environment Canada O. Knizik - Greater Vancouver Regional District, Canada M. K n o c h e - LAB, France K. Knox - Knox Associates, United Kingdom

T. Kosson - Rutgers University, USA H. K r u i j d e n b e r g - NOVEM, Netherlands P. Leenders- (formerly with VEABRIN - Netherlands) G. L u e r s - Corning Glass Ltd., USA T. Lundgren - Terratema AB, Sweden D. Mitchell - AEA Technology (formerly with Warren Spring Laboratory), United Kingdom K. Oberg- Swedish Environmental Protection Agency G. Owen - Environment Canada J. Pappain - Peel Resource Recovery Inc., Canada J. Pearson - AEA Technology, United Kingdom A. Petsonk- Swedish Environmental Protection Agency B. Putnam - International Lead Zinc Research Organization G. Rigo - Rigo & Rigo Associates, Inc., USA J. Robert- Energy, Mines & Resources, Canada F. Roethel - University of New York at Stoney Brook, USA H. Roffman - AWD Technologies, USA S. Sakai - Kyoto University, Japan M. Sheil - New Jersey Dept. of Environmental Protection & Energy, USA B. Simmons - California Board of Health, USA D. St~mpfli - formerly with EAWAG, Switzerland J. Stegemann - Wastewater Technology Centre, Canada L. Stieglitz - Forschungszentrum Karlsruhe, Germany M. Stringer - Greater Vancouver Regional District, Canada H. T e j i m a - Takuma Co., Ltd., Japan T. Theis - Clarkson University, USA B. T i m m - Swedish Environmental Protection Agency J. Tsuji - formerly with Environmental Toxicology International Inc., USA A. van Santen - Energy Technology Support Unit, United Kingdom J. F. Vicard - LAB, France J. Vogel - Heidelberger Zement, Germany H. V o g g - Forschungszentrum Karlsruhe, Germany S. Waring - AEA Technology, United Kingdom C. Wiles - National Renewable Energy Laboratory, USA M. Winka - New Jersey Dept. of Environmental Protection & Energy, USA J. W i t t w e r - Environment Canada D. Wexell -Corning, Inc., USA W. Wormgoor - TNO, Netherlands The IAWG also wishes to thank all of the other people not mentioned here, who in their own way assisted us in this endeavour. A special "thank you" to S t e p h e n H e t h e r i n g t o n for his patient efforts in revising and reformatting this document.


xiii

TABLE OF CONTENTS PREFACE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

V

CHAPTER 1 - INTRODUCTION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1.1 A BRIEF H I S T O R I C A L E X C U R S U S ........................... 1.2 THE D E V E L O P M E N T OF W A S T E I N C I N E R A T I O N ................ 1.3 O B J E C T I V E OF THIS TREATISE . . . . . . . . . . . . . . . . . . . . . . . . . . . . . REFERENCES ...............................................

1 1 2 12 13

CHAPTER 2 - MUNICIPAL SOLID WASTE 2.0 INTRODUCTION ......................................... 2.1 DEFINITION OF M U N I C I P A L SOLID W A S T E . . . . . . . . . . . . . . . . . . . . 2.2 C O M P O S I T I O N OF M U N I C I P A L SOLID W A S T E . . . . . . . . . . . . . . . . . . 2.3 QUANTITY AND MANAGEMENT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.1 Canada . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.2 Denmark . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.3 France . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.4 Germany . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.5 Japan . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.6 The Netherlands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.7 Sweden . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.8 Switzerland . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.9 United Kingdom . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.3.10 United States . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.4 CHEMICAL CONSTITUENTS ................................ REFERENCES ...............................................

15 15 17 21 22 25 26 28 31 32 35 37 37 39 41 51

CHAPTER 3 - MUNICIPAL SOLID WASTE INCINERATION TECHNOLOGIES . . . . . . . . . 3.1 FUEL RECEIPT A N D H A N D L I N G 59 3.2 AVAILABLE COMBUSTION ALTERNATIVES .................... 3.2.1 Mass Burning Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . European Type Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . Grates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Furnace Design . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Operating Philosophy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Modular Incineration Systems . . . . . . . . . . . . . . . . . . . . . . . . Other Mass Burn Variants . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2.2 Refuse Derived Fuel Systems . . . . . . . . . . . . . . . . . . . . . . . . Semi-Suspension Burning Systems . . . . . . . . . . . . . . . . . . . . Stoker Fired Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

59 61 62 62 65 70 73 76 77 79 82 85

xiv Fluidised Bed Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . HEAT RECOVERY SYSTEMS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . IN-PLANT RESIDUE M A N A G E M E N T . . . . . . . . . . . . . . . . . . . . . . . . . . 3.4.1 Bottom Ash . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.4.2 Grate Siftings . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.4.3 Heat Transfer System Ash . . . . . . . . . . . . . . . . . . . . . . . . . . . REFERENCES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

3.3 3.4

85 87 89 90 92 94 95

CHAPTER 4 - AIR EMISSION CONTROL STRATEGIES . . . . . . . . . . . . . . . . . . . . . . 4.0 INTRODUCTION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1 COMBUSTION CONTROL . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1.1 Theory . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Compensation for Fuel Variability . . . . . . . . . . . . . . . . . . . . . . Factors Controlling the Chemical Reaction Rate . . . . . . . . . . . 4.2 POST-COMBUSTION C O N T R O L . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2.1 Unit Processes For Air Pollution Control . . . . . . . . . . . . . . . . Particulate Matter Control Systems . . . . . . . . . . . . . . . . . . . . Electrostatic Precipitators . . . . . . . . . . . . . . . . . . . . . . . . . . . Fabric Filter (Baghouses) . . . . . . . . . . . . . . . . . . . . . . . . . . . Gaseous Controls . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Wet Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Dry Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Metals Control in Dry Systems . . . . . . . . . . . . . . . . . . . . . . . Mercury Control with Activated Carbon . . . . . . . . . . . . . . . . . NOx Removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3 TYPICAL APC INSTALLATIONS . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3.1 Hogdalen, Sweden . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3.2 Munich South, Germany . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3.3 Warren County, New Jersey, USA . . . . . . . . . . . . . . . . . . . . 4.3.4 Zirndorf, Germany . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3.5 Vestforbr~nding, Copenhagen . . . . . . . . . . . . . . . . . . . . . . . 4.3.6 Lausanne, Switzerland . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3.7 Bremerhaven, Germany . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3.8 Stuttgart, Germany . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . REFERENCES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

97 97 97 97 98 98 101 103 103 104 106 109 109 111 112 113 115 118 118 118 122 122 125 125 128 128 131

CHAPTER 5 - REGULATION OF MSW INCINERATORS . . . . . . . . . . . . . . . . . . . . . 5.1 EXISTING MSW INCINERATOR OPERATING GUIDELINES . . . . . . . . 5.1.1 Furnace Temperature and Residence Time . . . . . . . . . . . . . . 5.1.2 Combustion Efficiency and Carbon Monoxide . . . . . . . . . . . . 5.1.3 APC Temperatures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.1.4 Other Aspects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2 AIR EMISSION STANDARDS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.2.1 Chronological Changes in Emission Standards . . . . . . . . . . . 5.2.2 Emissions of Combustion Products and Acid Gases . . . . . . .

135 137 137 139 139 140 140 141 144

XV

H y d r o g e n Chloride . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Particulate Matter . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . S u l p h u r Dioxide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . O x i d e s of Nitrogen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Carbon M o n o x i d e . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Total H y d r o c a r b o n s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . H y d r o g e n Fluoride . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Trace Metals Emission S t a n d a r d s . . . . . . . . . . . . . . . . . . . . 5.2.3 5.2.4 Trace O r g a n i c Emission S t a n d a r d s . . . . . . . . . . . . . . . . . . . . 5.3 CURRENT ASH AND RESIDUE DISPOSAL PRACTICES .......... 5.3.1 Disposal of Bottom Ash . . . . . . . . . . . . . . . . . . . . . . . . . . . . Canada ....................................... Denmark ...................................... France . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . G e r m a n y and Switzerland . . . . . . . . . . . . . . . . . . . . . . . . . . Netherlands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sweden ....................................... United K i n g d o m . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . United States . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Japan . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3.2 Disposal of Fly Ash and A P C Residues . . . . . . . . . . . . . . . . Canada ....................................... D e n m a r k & the Netherlands . . . . . . . . . . . . . . . . . . . . . . . . . France . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Germany ...................................... Netherlands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sweden ....................................... Japan . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5.3.3 Utilisation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Canada ....................................... Denmark ...................................... France . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Netherlands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Germany ...................................... Sweden ....................................... Japan . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . United States . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . REFERENCES ..............................................

144 144 144 145 145 145 145 145

CHAPTER 6 - ISSUES RELATED TO INCINERATOR ASH SAMPLING . . . . . . . . . . . . 6.0 INTRODUCTION ........................................ 6.1 THE C O N C E P T OF THE REPRESENTATIVE SAMPLE ........... 6.1.1 Waste T y p e . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.1.2 T y p e of Incinerator/APC S y s t e m . . . . . . . . . . . . . . . . . . . . . 6.1.3 Residue S t r e a m s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.2 O B J E C T I V E S OF M A T E R I A L S A M P L I N G P R O G R A M S ...........

167 167 167 168 169 170 171

147 149 149 149 149 150 150 150 152 152 152 153 153 153 153 154 154 155 155 155 155 156 156 156 157 157 158 161 161 161

xvi 6.3 6.4

AVAILABLE SAMPLING PROTOCOLS . . . . . . . . . . . . . . . . . . . . . . . . SAMPLING CONSIDERATIONS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.4.1 Increment Collection Classification . . . . . . . . . . . . . . . . . . . . 6.4.2 Bias . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.4.3 Precision . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Number of Increments in Composite Sample . . . . . . . . . . . . . 6.4.4 Size of Increments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.4.5 Collection Procedures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.4.6 Sampling Streams Other Than Bottom Ash . . . . . . . . . . . . . . Grate Siftings and Heat Recovery Ash . . . . . . . . . . . . . . . . . APC Residues . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Storage Piles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sampling from Trucks or Containers . . . . . . . . . . . . . . . . . . . 6.4.7 Sample Preparation Concerns . . . . . . . . . . . . . . . . . . . . . . . Sample Size Reduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . Preservation of Samples . . . . . . . . . . . . . . . . . . . . . . . . . . . Sample Containers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sample Storage . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Laboratory Sample Preparation . . . . . . . . . . . . . . . . . . . . . . Laboratory Sample Subdivision . . . . . . . . . . . . . . . . . . . . . . Drying . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Size Reduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Balance of Materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . SAMPLE COLLECTION R E C O M M E N D A T I O N S . . . . . . . . . . . . . . . . . 6.5 6.5.1 Generic Bottom Ash Testing Protocol . . . . . . . . . . . . . . . . . . 6.5.2 Generic Boiler Ash Sampling Protocol . . . . . . . . . . . . . . . . . 6.5.3 Generic APC Residue Sampling Protocol . . . . . . . . . . . . . . . 6.5.4 Documentation of Sampling and Preparation Procedures . . . . E X A M P L E S OF SAMPLING STRATEGIES . . . . . . . . . . . . . . . . . . . . . . 6.6 6.6.1 Bottom Ash . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Regulatory Testing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Research Testing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.6.2 Grate Siftings . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.6.3 Boiler/Economiser Ash . . . . . . . . . . . . . . . . . . . . . . . . . . . . Regulatory Sampling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Research Sampling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 6.6.4 Air Pollution Control System Residues . . . . . . . . . . . . . . . . . Regulatory Testing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Research Testing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . REFERENCES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

174 175 175 177 178 178 179 180 181 181 182 182 183 183 183 185 185 186 186 186 187 187 188 188 188 191 192 193 194 194 194 196 197 198 198 199 199 199 200 200

CHAPTER 7 - CHARACTERISATION METHODOLOGIES . . . . . . . . . . . . . . . . . . . . . 7.1 PHYSICAL TESTING . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.1.1 Visual Observation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

203 203 203

xvii

7.2

7.3

Method . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bottom Ash . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Fly Ash and APC Residue . . . . . . . . . . . . . . . . . . . . . . . . . . Particle Size Distribution . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.1.2 Test Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Dry Sieve Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Fine Particle Analyses Methods . . . . . . . . . . . . . . . . . . . . . . 7.1.3 Density . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bulk Density . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Specific Gravity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Laboratory Testing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Field Testing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.1.4 Absorption Test . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Test Method . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.1.5 Water Content . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Test Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.1.6 Proctor Compaction Test . . . . . . . . . . . . . . . . . . . . . . . . . . . Standard Proctor . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Modified Proctor . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.1.7 Strength and Strength Development . . . . . . . . . . . . . . . . . . . 7.1.8 Bearing Capacity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.1.9 Durability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Soundness Tests . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . LA Abrasion Test . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Freeze-Thaw Tests . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.1.10 Permeability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Test Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . CHEMICAL COMPOSITION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2.1 Sample Preparation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Size Reduction Techniques . . . . . . . . . . . . . . . . . . . . . . . . . 7.2.2 Inorganic Analyses . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Digestion Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Specialty Methods for Specific Elements . . . . . . . . . . . . . . . . 7.2.3 Analytical Measurement . . . . . . . . . . . . . . . . . . . . . . . . . . . . Destructive Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Non-Destructive Analytical Methods . . . . . . . . . . . . . . . . . . . 7.2.4 Loss on Ignition . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7.2.5 Total Carbon, Carbonate, Sulphur and A m m o n i a . . . . . . . . . . 7.2.6 Acid Neutralisation Capacity . . . . . . . . . . . . . . . . . . . . . . . . . 7.2.7 Organic Analyses . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sample Preservation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Extraction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . CHEMICAL SPECIATION M E T H O D S . . . . . . . . . . . . . . . . . . . . . . . . . 7.3.1 Separatory Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . . .

203 203 205 205 205 206 206 206 207 207 207 208 208 208 208 209 209 210 210 211 212 212 213 213 214 214 214 221 221 221 223 223 225 226 226 229 232 234 235 236 236 237 237 238 238

xviii Sample Drying . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Particle Size Reduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . Magnetic Separation Techniques . . . . . . . . . . . . . . . . . . . . . Density Separation Techniques . . . . . . . . . . . . . . . . . . . . . . Selective Phase Dissolution Methods . . . . . . . . . . . . . . . . . . 7.3.2 Impregnation, Thin-Sections, and Thin-Foil Methods . . . . . . . 7.3.3 Analytical Methods for Solid Phase Chemical Speciation . . . . Transmitted Light Microscopy . . . . . . . . . . . . . . . . . . . . . . . Scanning Electron Microscopy ...................... Petrography (Morphology) . . . . . . . . . . . . . . . . . . . . . . . . . . Scanning Tunnelling Microscopy . . . . . . . . . . . . . . . . . . . . . X-Ray Powder Diffraction . . . . . . . . . . . . . . . . . . . . . . . . . . Petrography (Mineralogy) . . . . . . . . . . . . . . . . . . . . . . . . . . . Scanning Electron Microscopy/X-Ray Microprobe Analysis . . Scanning-Transmission Electron Microscopy/X-Ray Microprobe Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . Auger Electron Spectroscopy ....................... X-Ray Fluorescence Spectroscopy . . . . . . . . . . . . . . . . . . . X-Ray Photoelectron Spectroscopy . . . . . . . . . . . . . . . . . . . Secondary Ion Mass Spectroscopy . . . . . . . . . . . . . . . . . . . Electron Energy Loss Spectroscopy . . . . . . . . . . . . . . . . . . . X-Ray Adsorption Spectroscopy and Extended X-Ray Adsorption Fine Structure . . . . . . . . . . . . . . . . . . . . . . . Nuclear Magnetic Resonance . . . . . . . . . . . . . . . . . . . . . . . . Infrared Spectroscopy and Raman Spectroscopy . . . . . . . . . REFERENCES . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

................. MECHANISMS CONTROLLING THE FATE OF ELEMENTS . . . . . . . . 8.1.1 Fundamentals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.1.2 Processes in the Combustion Chamber . . . . . . . . . . . . . . . . Physical Parameters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemical Reactions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sintering and Related Processes . . . . . . . . . . . . . . . . . . . . . Physicochemical Transformations . . . . . . . . . . . . . . . . . . . . . 8.1.3 Mechanisms in the Boiler . . . . . . . . . . . . . . . . . . . . . . . . . . . Condensation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Corrosion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.1.4 Mechanisms in the Dust Removal System . . . . . . . . . . . . . . 8.1.5 Mechanisms in the Air Pollution Control System . . . . . . . . . . MASS STREAMS IN A MUNICIPAL SOLID WASTE INCINERATOR . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . LITHOPHILIC ELEMENTS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.3.1 Fundamentals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

CHAPTER 8 - FATE OF ELEMENTS DURING INCINERATION

8.1

8.2 8.3

240 240 241 242 242 243 245 247 247 247 249 249 250 250 250 251 251 251 252 252 253 253 254 254

263 264 264 265 265 266 272 277 280 280 281 284 285 285 287 287

xix 8.3.2 8.3.3 8.3.4

8.4

Alkali M e t a l s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Earth-Alkali M e t a l s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Heavy Metals . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Iron . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chromium ...................................... Nickel . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Copper ........................................ VOLATILE ELEMENTS ................................... 8.4.1 Halogens ...................................... Chlorine ...................................... Fluorine . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . B r o m i n e and Iodine ..............................

288 290 291 292 294 295 295 296 296 297 298 300

Sulphur ....................................... Nitrogen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.4.2 Volatile M e t a l s . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mercury ....................................... Cadmium ...................................... Zinc . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Lead . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Arsenic ........................................ Antimony ...................................... O t h e r Volatile E l e m e n t s . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.5 CARBON AND SELECTED CARBON COMPOUNDS ............. 8.5.1 Total C a r b o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.5.2 P o l y c h l o r i n a t e d D i b e n z o - p - D i o x i n s and -Furans . . . . . . . . . . . 8.5.3 Polychlorinated Biphenyls .......................... 8.5.4 Polychlorinated Benzenes .......................... 8.5.5 Polychlorinated Phenols ........................... 8.5.6 Brominated Hydrocarbons .......................... 8.5.7 Polycyclic A r o m a t i c H y d r o c a r b o n s . . . . . . . . . . . . . . . . . . . . REFERENCES ..............................................

300 302 303 304 305 307 308 310 311 312 312 312 314 320 322 323 324 324 326

CHAPTER 9 - BOTTOM ASH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.1 PHYSICAL CHARACTERISTICS OF BOTTOM ASH .............. 9.1.1 Gross C o m p o s i t i o n . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Reject Fraction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Visual Classification .............................. Water Content .................................. Ferrous C o n t e n t . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Loss on Ignition . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . D i s s o l v a b l e Solids C o n t e n t . . . . . . . . . . . . . . . . . . . . . . . . . . 9.1.2 Gravimetric Characteristics ......................... Specific Gravity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Absorption .....................................

339 342 342 342 343 345 346 347 351 351 353 353

XX

Unit Weight . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Gradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Percent Fines . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.1.4 Durability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Soundness . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Abrasion Resistance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.1.5 Geotechnical Properties . . . . . . . . . . . . . . . . . . . . . . . . . . . . Proctor Compaction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Field Compaction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . California Bearing Ratio (CBR) . . . . . . . . . . . . . . . . . . . . . . . Permeability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.1.6 Influence of Combustor Type and Operation on Physical 9.1.7 Characteristics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.1.8 Influence of Aging on Bottom Ash Physical Characteristics .. PARTICLE MORPHOLOGY, MINERALOGY, AND ALKALINITY OF BOTTOM ASH . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.1 Morphology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.2 Mineralogy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.3 Alkalinity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.2.4 Influence of Combustor Type and Operation on Bottom Ash Surface Area, Mineralogy and Alkalinity . . . . . . . . . . . . . . 9.2.5 Influence of Aging on Bottom Ash Surface Area, Mineralogy and Alkalinity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . INORGANIC CHARACTERISTICS OF BOTTOM ASH . . . . . . . . . . . . . 9.3.1 Elements Present in Bottom Ash . . . . . . . . . . . . . . . . . . . . . 9.3.2 Major Matrix Elements (> 10,000 mg/kg): O,Si,Fe,Ca,AI,Na,K,C . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.3 Minor Matrix Elements (1,000 to 10,000 mg/kg): Mg, Ti, CI, Mn, Ba . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.4 Other Minor Elements (1,000 to 10,000 mg/kg): Zn, Cu, Pb, Cr . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 9.3.5 Other Trace Elements Including Oxyanionic Elements (720 5 2 >480- 240 - 500 > 100 - 50 pm material, >99% of the 5 pm material and 98% of the 1 pm and smaller material. Unlike fabric filters, emissions from ESPs can vary depending on the concentration of particulate matter in the flue gas stream. ESPs can be designed for temperatures as high as 375~ although more commonly they operate at about 200~ Some volatile metals will escape capture in ESPs operating at elevated temperatures because they will not condense and form particles that can be removed. Without an appreciable temperature drop volatile organic and metallic species will flow through the system. This is particularly the case for mercury as shown in Environment Canada's NITEP studies (Environment Canada, 1986). In addition, since trace metallic and organic species tend to condense on surfaces and the ESP has only limited fine particle control, the removal effectiveness for species condensing on the surfaces of the fine particles is reduced. Furthermore, ESPs operating at elevated temperatures have been shown by Vogg et a1.(1990) to be likely to have enhanced PCDD/PCDF in their residue streams. This finding suggests that de novo synthesis of these compounds occurs at the operating temperatures found in some units. On the other hand, ESPs operating at low temperatures (..:. . . :';;=

.. ;~.. ~.,. .,..~~.; .-

Reaction

additives

Mixing

From

boiler

Cona~tioning

9-

Plu'lddd,~l,~e

Flue g a s

pretreatment

CDAS reactor. The flue gas is first cooled with water to suitable temperature, then mixed with dry sorption agent.

(courtesy Fl~kt Review, 1990)

122 control of mercury, which causes the mercury to condense out in the form of mercuric sulphide. The back end of this system is equipped with an SCR reactor for NOx control. To operate the SCR system, the gas passes through a heat exchanger and burner to raise the gas temperature to the appropriate operating temperature. The NH3 is added and the gas passes through the SCR reactor prior to being discharged to the stack. No dust control is installed after the SCR reactor and ammonia salts will not be present in the plant residues. The fabric filter dust from the APC system is recirculated similar to the Hogdalen application until such time as it transfers to the waste silo. From here the material is wetted only to prevent dusting during transport to the disposal site.

4.3.3 Warren County, New Jersey, USA Typical of many facilities in both the US and Europe, the Warren County facility utilises a spray dry absorber (SDA) for acid gas and trace compound removal, Figure 4.8 (Jorgensen et al., 1991). The system was supplied by Environmental Elements Corporation. Lime slurry is introduced into the SDA where it neutralises the acid gases and serves as nucleating sites for the condensation of trace materials. The particulate matters in the gas stream are then removed in a fabric filter. Typical of many U.S. installations, the APC residues are combined with the bottom ash before being transported to the disposal site.

4.3.4 Zirndorf, Germany The schematic for this plant's APC system, Figure 4.9, illustrates the use of a wet scrubber system in a retrofit situation (Beckert and Jungmann, 1992). The retrofit was performed by ABB W+E Umwelttechnik in combination with ABB Fl~kt-Umwelttechnik. The existing ESP and HCI scrubber were upgraded with the addition of a wet SO2 scrubber and an absorption stage. The residue from the absorption stage is returned to the incinerator where any organics are destroyed and residual mercury is liberated to be trapped in the organic sulphides in the HCI wet scrubber. A heat exchanger removes heat from the gases before they pass to the HCI scrubber and transfer the heat to the gas leaving the scrubber system. The HCI scrubber uses water as the scrubbing media in a contact tower. A bleed from this system maintains the pH of the solution at an appropriate level. Organic sulphides are added to the scrubber liquid to improve the separation of trace metals, particularly mercury. The sump of the scrubber is the eventual sink for all trace metals in the system with the bleed from this circuit combined with that from the SO2 scrubbing circuit. In the wastewater treatment system the pH is adjusted with a lime slurry and forms gypsum and organic sulphides which precipitate and are filtered from the effluent.

Figure 4.8 Process Flow Sheet -Warren County

I Boiler Ash

Mixer

Waste Product Waste to Disposal

The absorption stage is returned to the incinerator where any organics are destroyed and residual mercury is liberated to be trapped in the organic sulphides in the HCI wet scrubber

Jorgensen et al., 1991

A

h,

W

Figure 4.9 Process Flow Sheet - Zirndorf, Germany

Beckert and Jungrnann, 1992

125 The SO2 scrubber is a packed tower system. Sodium hydroxide in softened water is used as the scrubbing liquid. A bleed from this system maintains the salt content of the scrubber liquid at an optimal level. The gases are reheated before passing to a compact reactor where fresh lime and powdered activated carbon are added to the gas stream to remove trace organics and residual trace metals such as mercury. The sorbents are then captured in a fabric filter where the filter cake serves to polish the acid gas removal process and remove trace organics and trace metals. The filter cake is recirculated in the system and then eventually disposed by introducing it into the high temperature zone of the furnace where trace organics are destroyed. The only residue from this system is the filter cake from the wastewater treatment plant.

4.3.5 Vestforbra~nding, Copenhagen The APC system at this plant represents one of the latest wet scrubber concepts, Figure 4.10. The system was designed by G5taverken of Sweden. The existing hot side ESP has been maintained and is followed by an economiser, gas/gas heat exchanger, quench system, and HCI and SO2 scrubbers. The latter scrubber is designed to operate at 50~ to generate hot water and recapture energy while condensing much of the moisture out of the stack gases. The gases are reheated before exiting to the stack. Wastewater treatment in this system includes limestone coarse pH adjustment, lime slurry fine pH adjustment along with organic sulphide to bind trace metals, and polymer addition to aid settling and separation of the suspended particulate matter. The sludge from the system is mixed with the ESP residue and landfilled.

4.3.6 Lausanne, Switzerland The LAB system installed at Lausanne is one of a range of systems manufactured by this company. The systems are applied with different levels of equipment to achieve the emissions limits required. Figure 4.11 shows a typical installation of the EDV 7000 variant. In this case the scrubbers are open vessels without packing. Proprietary nozzles are used to generate a high density water curtain in the vessel that neutralises the gases. The gases then pass to electrofiltering modules where the particles are charged. Particles are removed from the gas stream by a water spray. The system is completed by a water droplet removal device based upon centrifugal force. The wastewater treatment system is similar to those seen in other plants outlined above, and produces a sludge requiring disposal.

Figure 4.10 Process Flow Sheet - Vestforbranding, Copenhagen

(courtesy Gotaverken Miljo AB)

127 Figure 4.11 Process Flow Sheet - Typical LAB EDV 7000

(courtesy LAB S.A.)

128

4.3.7 Bremerhaven, Germany Detailed flow sheets for two German installations, Figures 4.12 and 4.13, show variations in APC strategy (Lange, 1992). Figure 4.12 illustrates the system used at Bremerhaven. The use of SNCR NOx control through the injection of ammonia into the furnace is similar to that shown in other facilities. The unique part of this installation is the steam stripping of the scrubbing solution from the wet scrubber to recover ammonia which is then reused in the NOx control system. The quantities illustrate the low ammonia level in the flue gas. An ESP is used to remove the particulate matter from the flue gas leaving the furnace and clean it before it enters the scrubber. A wastewater treatment facility is noted, but no details of the system are available.

4.3.8 Stuttgart, Germany This facility utilises a spray dryer/absorber followed by an ESP to treat the scrubber solution from the APC system, Figure 4.13, after Lange (1992). The SCR NOx control system on the plant is similar to the installation shown for Zirndorf.

Figure 4.12 Process Flow Sheet - Bremerhaven, Germany

temperature

:

30

OC

to waste water

Flow sheet of the waste gas purification plant with the SNR process (municipal waste incineration plant Bremerhaven) L

(Lange, 1992)

Figure 4.13 Process Flow Sheet - Stuttgart, Germany

heat exchanger

blower

waste gas heating

-

SCR reaktor

. Flow sheet of the waste gas purification plant with the SCR process I municipal waste incineration plant in Stuttgart )

(Lange, 1992)

131 REFERENCES

ASME. Retrofit of Waste-to-Energy Facilities Equipped with Electrostatic Precipitators. An ASME Research Report prepared by H.G. Rigo and A.J. Chandler under the direction of the ASME Research Committee on Industrial and Municipal Waste. CRTD Vol. 39. 1996. AWMA. Air Pollution EnQineerin.q Manual/Air & Waste Mana.qement Association. Edited by A.J. Buonicore and W. Davis. Van Nostrand Reinhold, New York, NY., 1992. Beckert, P. and G. Jungmann. "Retrofit of the Flue Gas Cleaning System of the Waste to Energy Plant Zirndorf". A paper presented at the 7th IRC conference Berlin, Nov. 1992. In Wa.ste Management International, Karl J. Thome-Kozmiensky (ed.), Berlin: EF-VerI. fur Energie- und Umwelttechnik, ISBN 3-924511-64-0, 1992. Environment Canada. "The National Incinerator Testing & Evaluation Program (NITEP) Air Pollution Control Technology". Environment Canada Report EPS 3/UP/2, September 1986. Fl~kt. "Cleaning Flue Gases in Energy from Waste Plants". A sales document from Fh~kt, Sweden, 1991. Guest, T.L. and O. Knizek. "Mercury Control at Burnaby's Municipal Waste Incinerator". Proceedings of the 84th..Annual AWMA Meetin& Vancouver, B.C. Paper 91-103.30, June, 1991. Guest, T.L., 1993. Mercury Control in Canada. Proceedings of the 86th Annual AWMA Meeting. Paper 93-WP-109. 01. Denver, Colorado. June. Hartenstein, Hans-Ulrich and Anthony Licata. "The Application of a Low Temperature Selective Catalytic Reduction System for Municipal & Hazardous Waste Combustors". Proceedings of the 17th Biennial Waste Processing Conference. ASME. 1996. Heath, Patrick B., 1995. Design and Installation of Powdered Activated Carbon Storage and Injection Systems for Municipal Solid Waste Incinerators. Proceedings of the 88th Annual AWMA Meeting. Paper 95-RP-147B.01. San Antonio, Texas. June. Herrlander, B. "Recent developments in de-NOx technology". An article in Fl~kt Review No. 73, Stockholm, Sweden, February 1990. Hofmann J.E. et al. "NOx Control for Municipal Waste Combustors". A&WMA Annual Meeting, Pittsburgh, PA, 1990.

132 Hurst, B.E. and C.M. White. 'q'hermal DeNOx: A Commercial Selective Non-Catalytic NOx Reduction Process for Waste to Energy Applications". A paper presented at The ASME 12'h Biennial National Waste Processing Conference, Denver, Colorado, June, 1986. Jorgensen, C. et al. "Two and a Half Years Operating Experience at the Warren County Energy Resource Recovery Facility". A paper presented at the 2nd International Conference on Municipal Waste Combustion, Tampa, Florida. Published by A&WMA Association VIP-19 Municipal Waste Combustion, Pittsburgh, PA, April, 1991. Lange, M. "Legal Regulations and Technical Measures Limit the Emissions of Waste Incinerators in the Federal Republic of Germany". A paper in A Selection of Recent Publications (Volume 3) prepared by Umweltbundesamt, Berlin, Germany, 1992. Licata, A., M. Babu, and L-P Nethe, 1994. An Economic Alternative to Controlling Acid Gases Mercury, and Dioxin form MWC's. Proceedings of the 87th Annual AWMA Meeting. Paper94-MP-17.06. Cincinnati, Ohio. June. LIRPB. Lon.q Island ReQional Pla.nnin.q Board, The Potential for Beneficial Use of W_aste-to-Ener.c]v Facility Ash. Seven volume report. Hauppauge, NY, 1992. McDonald B.L., G.R. Fields and M.D. McDaniel. "Selective Non-Catalytic Reduction (SNCR) Performance on Three California Waste-to-Energy Facilities". A paper presented at the 2nd International Conference on Municipal Waste Combustion, Tampa, Florida. Published by A&WMA Association VIP-19 Municipal Waste Combustion, Pittsburgh, PA, April 1991. Miller, J.A. and G.A. Fisk. "Combustion Chemistry". C&ENAug. 31, 1987, pp. 2248. Rigo, H.G., 1993. How Good are Today's Mercury Test Methods and Controls? A paper presented at the Ash VI Conference, Arlington, Va. November. Seeker, W.R., G.C. England and R. Lyon. "Advanced Pollution Control in Municipal Waste Combustors Using Natural Gas". A paper presented at the 2nd International Conference on Municipal Waste Combustion, Tampa, Florida. Published by A&WMA Association VIP-19 Municipal Waste Combustion, Pittsburgh, PA, April 1991. Seeker, W.R., W.S. Lanier and M.P. Heap. "Municipal Waste Combustor Study: Combustion Control of Organic Emissions". A Report prepared for the U.S. EPA by Energy and Environmental Research Corporation EPA/530-SW-87-021C. Irvine, CA, 1987.

133 Sierhuis W.M. and J.G.P. Born. November, 1994. PCDD/F Emissions Related to the Operating Conditions of the Flue Gas Cleaning System of MWI- Amsterdam. A paper from Dioxin '94 Kyoto, Japan, published in Organohalogen Compounds Volume 19. U.S. EPA. "Air Pollution Engineering Manual/U.S. Environmental Protection Agency" AP-40. Prepared by the Los Angeles Air Pollution Control District, 1973. U.S. Environmental Protection Agency, 1995BID. Municipal Waste Combustion: Background Information Document for Promulgated Standards and Guidelines -- Public Comments and Responses. Emission Standards Division, U.S. Environmental Protection Agency Office of Air and Radiation, Office of Air Quality Planning and Standards Research Triangle Park, North Carolina 27711. October. Vogg H., A. Merz, L. Stieglitz, F. Albert and G. Blattner. "Zur Rolle des Elektrofilters bei der Dioxin-Bildung in Abfallverbrennungsanlagen", Abfall.wirtsch.aft Journal, 2, p.529, 1990. Zmuda, J.T. and P.V. Smith. "Retrofit Acid Gas Emission Controls for Municipal Waste Incineration: An Application of Dry Sorbent Injection". A paper presented at the 2nd International Conference on Municipal Waste Combustion, Tampa, Florida. Published by A&WMA Association VIP-19 Municipal Waste Combustion, Pittsburgh, PA, April 1991.


135

C H A P T E R 5 - REGULATION OF MSW INCINERATORS

During incineration, most organic-based materials are destroyed by complete oxidation to carbon dioxide and water vapour but inorganics remain substantially untouched. Inorganics either partition within the various residue streams or are entrained into the flue gas stream. As noted in the previous chapter, considerable work has been undertaken in the past 10 to 15 years to reduce the level of emissions to the atmosphere through the stack. This has resulted in the application of new APC technologies incorporating both new operating philosophies and new equipment. These changes have influenced the characteristics of the residue streams. For instance, the use of lime to remove acid gases has increased the mass of the fine residue generated in the systems which, in turn, has increased the alkalinity of this material, thereby influencing the potential solubility of trace metals in the disposal site. APC technologies based on wet scrubbers generate less solid residues, but they generate a wastewater stream requiring treatment. To address the regulatory concerns, various jurisdictions have developed specific MSW incineration regulations governing air emissions, residue disposal, and wastewater treatment. This chapter reviews these regulations. Regulations represent an evolving set of limits on the operation of MSW incinerators. Newer limits are more stringent and comprehensive in response to requirements that are aimed at controlling an increasing number of substances (contaminants). Each jurisdiction has adopted standards they consider appropriate for their circumstances. Typically, both the toxicity and persistence of contaminants in the environment are considered during the evaluation process. Examples of the contaminants that could be considered are shown in Table 5.1. This list was derived from the Canadian MSW incinerator operating guidelines (CCME, 1989). The contaminants listed cover the broad range from combustion related gas emissions to trace metallic and organic compounds. As noted in an earlier chapter, changes to the operational characteristics of a facility can influence atmospheric emissions and consequently influence the characteristics of the residues. Regulations governing the design and operation of MSW incineration facilities to minimise air emissions address both: ~ operational factors such as combustion conditions; and, ~ the quantity of materials released through the stack. Emission standards exist in many jurisdictions but they are not presented on a consistent basis. Both temperature and diluent concentration can vary between jurisdictions. The values in this report have been corrected to 11% 02 and reference conditions of dry gas at 25~ and 101.3 kPa pressure. (Note: The comparative diluent and temperature correction used in the United States is 7% oxygen and 68~ Thus, while the Ontario and US emission concentrations are equivalent, when comparing the two standards with their normal diluent basis, the Ontario values would appear to be

136 T a b l e 5.1 List of Air E m i s s i o n C o n t a m i n a n t s Acid Gases 9 9

Hydrogenchloride Oxides of nitrogen

9 9

..

Hydrogenfluoride Oxides of sulphur

Trace Metals *Cd - C a d m i u m Fe - Iron *Be - Beryllium *Pb - L e a d Mo - Molybdenum *Cr - C h r o m i u m Ca - C a l c i u m *Ni - Nickel

*V - V a n a d i u m Si - Silicon AI - A l u m i n u m Ti - T i t a n i u m Mg - M a g n e s i u m B - Boron Ba - Barium P - Phosphorous

K *Hg Na *As *Zn *Sb Mn Bi

- Potassium - Mercury - Sodium - Arsenic - Zinc - Antimony - Manganese - Bismuth

Co *Se *Cu Te Ag Sn

- Cobalt - Selenium - Copper - Tellurium - Silver - Tin

*metals selected, bY the committee as most important for health and the environment. Organics Polychlorinated dibenzo-p-dioxin (PCDD) homologues TCDD PeCDF HxCDF HpCDD OCDD

tetra penta hexa hepta octa

Polychlorinated dibenzo-furan (PCDF) homologues TCDF PeCDF HxCDF HpCDF OCDF

tetra penta hexa hepta octa

C h l o r o b e n z e n e s (CB) CI-2 benzenes CI-3 benzenes CI-4 benzenes CI-5 benzenes CI-6 benzene Polychlorinated B i p h e n y l s (PCB) C h l o r o p h e n o l s (CP) CI-2 phenols CI-3 phenols CI-4 phenols CI-5 phenol

Polyaromatic Hydrocarbons (PAH) Acenaphthylene Acenaphthene Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Chrysene Benzanthracene Benzo(e)pyrene Benzo(a)pyrene Benzo(b,k)fluoranthene Perylene o-Phenylenepyrene Dibenzo(a,h)anthracene Benzo(g,h,i)perylene

Other Particulate matter Opacity After C C M E , 1989

Carbon monoxide Total hydrocarbons

Oxygen or Carbon dioxide

137 about 70% of the U.S. levels.) Wet standard levels are converted to dry, assuming the average moisture level will be 20%. Residue treatment and disposal regulations exist in many jurisdictions. The scope of these can vary from requirements for disposal under controlled conditions to quality standards such as the level of carbon in the material being disposed. Some jurisdictions have developed regulations governing the utilisation of residues. Both these issues are discussed later in this chapter. Wastewater treatment aspects of the disposal of wet APC residues is not discussed because, for the most part, these regulatory requirements do not specifically address wastewater discharges from MSW incineration facilities.

5.1 EXISTING MSW INCINERATOR OPERATING GUIDELINES Operational guidelines have been developed to assist regulators in standardising the design parameters of incinerator facilities. A summary of typical conditions is provided in Table 5.2. The intent of these standards is to ensure good combustion conditions and minimal organic emissions. The design guidelines are based upon theoretical calculations of the combustion process. In some cases, the applicability of these calculations is influenced by new trends in technology. In addition, some regulations stipulate the objectives for incineration. For example, German regulations mandate the recovery of energy from MSW incinerators; no systems are used solely to incinerate waste.

5.1.1 Furnace Temperature and Residence Time The operating temperature concept used in North America was developed from hazardous waste incinerator test data and some now suggest that such standards may be inappropriate for MSW systems. Because the reactions needed to take the combustion process to completion occur very quickly, if the temperature is sufficient and there is enough oxygen, these restraints on temperature and residence time may be artificial and can limit the development of new technologies. Thus, while there is a regulatory move to relax the prescriptive design aspects for MSW incinerator regulations, staff charged with the responsibility of making decisions on applications still find the guidance of these standards helpful. When considering temperature values in the table it should be remembered that some of the apparent differences result from conversion between various temperature units. The value 982~ represents 1800~ a common value for the United States whereas, Canada and the European countries specify temperatures in ~ and choose 1000~ as a suitable number. Both numbers are within the accuracy obtainable from conventional measurement systems.

Table 5.2 Incinerator Operating Conditions Jurisdiction

Temperature ?c)

Residence Time (Seconds)

Minimum Excess Oxygen

CO Level (mglRm3@ 11% 0,) Averaging Times

1 hour

I 4 hours

U.S. 1991

see note California Guidelines

982

1

300 (none specified)

Pennsylvania BAT Criteria

982

1

300 (8 hour)

New Jersey Guidelines

815

1

6

1000

1

6lRDF 3

Peel, Ontario Perml

1000

1

6

B.C. Guideline

1000

1

1000

11

CCME Guidelines

Ontario A7 Guidelines 1995

1

1850

Eurowan Communitv Guidelines Germany

1

850

12

1

I

I

I

1

90% of readings el50 (24 hours)

850

2

6

Denmark

850

2

6

1

I

850

12

France

850

2

7

Belgium

800

1

6

Italy

950

2

Norway

800

1.5

Sweden (6 tph) Japanese Guideline 1990

850

850

1 800

see note

I

APC Operating Temperature (DC)

17 mar. test temD.

6IRDF 3

United Kingdom

Netherlands

I 1

24 hours

loo

I

100

620 (10 minute) 825 (1 minute)

loo

2

6 6

12

16

Notes: Under carbon monoxide the US Regulations distinguish between incinerator types. Under 4 hour averaging times: 40 (35 ppmdv) refers to modular units; 80 (70 ppmdv) to mass burn waterwall or refractory wall, and, fluidised bed units; and, 120 (105 ppmdv) to RDF mixed with pulverised coal units. Standards for both new and existing facilities are identical. Under the 24 hour averaging time category: existing units are limited to 80(7O ppmdv) for mass bum rotary refractory units; 160 (140 ppmdv) for RDF spreader stokers with or without coal and 200 (175 ppmdv) for mass bum rotary waterwall systems. For for new plants the 24 hour average for RDF systems except Mass Bum Rotary Waterwall are limited to 120 (105 ppmdv) for CO with the MBWW Rotary units having a limit of 80 (70 ppmdv).

139 Residence time specifications vary from the stringent 1 second residence time at 1000~ specified in Ontario Guideline A-7, to a more flexible approach such as that favored by some European countries. The Ontario guideline provides a definition of the zone where 1000~ could be expected to occur. Europeans require a longer time at lower temperatures to allow some flexibility in design. Germany has gone one step further in the latest documents, (17 BImSchV; November, 1990). That regulation states that if testing data collected at the specified conditions is the same as that collected at lower temperatures or shorter residence times, the facility can be operated at the lower level.

5.1.2 Combustion Efficiency and Carbon Monoxide Carbon monoxide (CO) operating levels are used as a surrogate for good combustion conditions. They also are used to calculate various combustion efficiency factors. In some jurisdictions combustion efficiency may be defined as the ratio of CO to the sum of CO and CO2; in others it may be the ratio of CO to CO2. Because CO can be measured directly, the trend is towards setting a standard for CO levels at the boiler exit. CO standards are contained in Table 4.3. Averaging time variations are evident in the CO standards. The Danish standards are for 10-minutes and 1-minute respectively. The latest German standards specify three values based upon differing averaging times for CO. The 24-hour or daily mean must be 50 mg/m3; with an hourly mean of 100 mg/m3; and, 90% of all readings in the 24 hour period being less than 150 mg/m 3. The German values are not considered emission values but rather used as an operational parameter. Where additional values are shown under the 4-hour category in Table 5.2, they are explained in the footnote. Notable by its absence is the lack of a carbon monoxide standard in the new Ontario Guideline A-7. It should be recognized that under the operating procedures in the province, each new facility will have a specific Certificate of Approval that will specify operating parameters of this nature. These numbers can be tailored for different combustion technologies thereby reducing the complexity of the new guideline.

5.1.3 APC Temperatures Performance of the air pollution control systems has been found to be correlated with temperature at the inlet to the APC (NITEP, 1986). Some jurisdictions are recommending operating temperature restrictions on these systems. The aim is to lower the temperature to increase the trace contaminant and acid gas removal efficiency while maintaining temperatures high enough to minimize corrosion in the system and blinding of the fabric filters. At low temperatures the hygroscopic nature of the reaction products of the sorbent injection systems can lead to formation of 'mud-

140 like' material that coats the bags and turns to a solid non-porous mass. Canada included a maximum inlet operating temperature recommendation in the CCME (1989) document. The 1995 US regulations specify that the facility cannot be operated with an inlet temperature to the last particulate control device in the APC that is any higher than 17~ more than the 4-hour block average during the most recent successful dioxin test period.

5.1.4 Other Aspects There are other design related parameters such as capacity, (throughput) and auxiliary fossil fuel fired burner capacities included in some regulations. During operation, incinerators are subject to upset if the loading rate is too variable or too far from the design point. Limitations on feed rate are included in some standards including the 1995 U.S. standards which limit loading to no more than 10% above that used during the last test series. This is to prevent excessive particulate entrainment and potential trace organic formation downstream of the combustion chamber (US EPA, 1995BID). The German, Dutch and English regulations and others require that the system be equipped with auxiliary burners to enable the furnace to reach operating temperature before MSW is added, and to maintain combustion efficiency in the event of a drop in fuel level in the system. Several jurisdictions have added staff/operator training requirements to regulations. The new US EPA rules require ASME certification of senior staff; British Columbia MOE's proposal requires that all staff be trained to a level acceptable to the Ministry. Owners of MSW incineration facilities attempt to hire staff who have previous experience at similar facilities. Lead operators generally have such experience and others on staff are promoted to positions of increased responsibility if they have appropriate qualifications and sufficient time in the facility.

5.2 AIR EMISSION STANDARDS The operating emissions from a facility are related to both the type of APC system installed and the nature of the MSW received at the facility. Generally, the greater the efficiency of the APC system, the lower the emissions, although the presence of commercial or industrial type waste in the fuel stream may raise trace contaminant levels in the emissions. Regulators set performance standards based upon their desire to minimise emissions; however, these standards vary from jurisdiction to jurisdiction based upon the perceived requirements in that area and their regulatory interpretation of the issues discussed earlier. This section reviews existing regulations for several groups of emissions:

141 9conventional combustion products and acid gases 9trace metals 9trace organics. Regulatory limits can be based upon either the absolute emission number or a removal efficiency determined by the ratio of emissions to input to the APC device. Other variations in the standard setting process include the use of various averaging protocols based either on the time of sampling or the number of samples taken. While the following summarises some of these issues, no attempt has been made to address the differences in sampling times, rather the data are normalised to standard temperature and pressure and a standard basis diluent concentration of 11% 02. 5.2.1 Chronological Changes in Emission Standards

As noted in the introduction, newer standards tend to be more stringent. Table 5.3 illustrates the trends for emission standards of conventional pollutants in those countries where the data are easily traced. Regulations in the United States have had numerous changes since 1989. They are extremely complex, given that they were written to address different technologies and sizes of facilities, but all exhibit the same decreasing trends in emissions. Table 5.3 Chronology of Municipal Solid Waste Incinerator Emissions Limits - Combustion products and Acid Gases (Values expressed as mg/Rm 3 @ 11% 02) ..... Jurisdiction (Country/State/Prov.)

Hydrogen Chloride

France

1982 1986

155

Nethedands

1984

1080

Switzerland

Hydrogen Fluoride

Sulphur Dioxide

Oxides of Nitrogen

120

23

690

290

Particulate Matter

Carbon Monoxide

Hydrocarbons (as CH4)

200

1530

11

78

1900

120

1989

9

1

37

65

5

46

1986

28

4

460

460

46

92

9

1991

18

2

46

74

9

46

18

1986

46

2

90

460

32

92

18

1990 mean 24 hour

9

1

46

184

9

46

9 (as carbononly)

Denmark

1986

83

2

240

33

83

1991

60

2

276

37

92

1994 Proposed

50 10

2 1

240 50

30 10

100 50 (daily)

Germany

Sweden

* = as total carbon for old plants

200

20 * 10 (totalcarbon)

142 Four major standards are currently enforced: the Canadian (CCME, 1989); European Economic Community (EEC, 1989); German (17 BImSchV, 1990); and the U.S. (EPA, 1995). A summary of these standards and those from other jurisdictions is presented in Table 5.4. As noted above, the US EPA standards have evolved since the late 1980's and those in the table represent the latest edition. These regulations are currently being challenged in the U.S. court system and may be subject to revision. The EEC Directive sets out minimum standards for MSW incinerators in all countries of the EEC. All new facilities, as of December 1, 1990, are required to meet the standards, whereas existing large facilities have until December 1996 to comply. Interim standards on facilities smaller than 6 tonnes per hour applied in December 1995, forcing local standards to adjust by 1996. The Germans have added other requirements to the EEC Directive noting that existing installations needed to comply by March 1994, and absolute compliance must occur by December 1996. The rule allows local jurisdictions some discretion in requiring tighter controls where necessary to protect the environment, while still allowing flexibility in operating conditions where it can be demonstrated that the alternative operating conditions do not have a detrimental impact on the quality of air emissions. The newest U.S. regulations were passed into law in late 1995. They apply across the country and set a minimum performance standard in much the same manner as the European Directive. These standards require compliance at existing facilities by 2000, but states have the ability to accelerate compliance by including these standards in state regulations. The current court challenge relates to the distinction between different sizes of facilities included in the standards and the absolute limits that will apply to some of the existing smaller facilities. The Canadian CCME guideline was developed by a joint federal/provincial committee and was meant to act as a basis for provincial regulations for new MSW incinerators in Canada. The new Ontario guideline applies only to new facilities built after December, 1995. The guidelines do not apply to existing facilities in Ontario unless they undergo modification or expansion. Currently these facilities have specific operating permits which contain marginally higher emission standards. As mentioned earlier, these specific permits allow more stringent standards to be set for specific facilities if necessary. As is the situation in Canada, all countries tend to view national standards as minimum operating levels. Local jurisdictions can apply more stringent standards. These are reflected in facility specific operating permits and thus there appear to be a plethora of standards in some countries when, in fact, the operating limits at a particular facility reflects the requirements of the local area.

Table 5.4 Municipal Solid Waste Incinerator Emissions Limits - Combustion Products and Acid Gases (mglRm3@ 11% 0,) Jurisdiction (Country/State/Prov.) European Economic Community 1991 Italy 1991 United Kingdom 1992 (new plants) Belgium 1991 Netherlands 1989 Sweden 1986 Norway 1992 Switzerland 1991 Austria 1989 Germany 1990 mean 24 hour Germany 1990 1R hour max Denmark 1991 mean 24 hour U.S.A. NSPS 1995 New Facilities Existina . ... ... . .Facilities - -...- ..>35 .. tod .r - 8 . . 35 tpd Existing >35 & 225 tpd Canada CCME Guidelines 1988

Note:

I

0.20"

Italy

Switzerland 1986

9

Trace Metals by Cate.qory

Ontario Guideline A-7 1995 0.014 Cd 0.14 Pb 0.057 Hq Generally, Hg and Cd are in Class I but Sweden has Hg only and the old German and British Columbia standards include TI in Class I. Class II has As and Ni in the EC; Class III for the EC is Pb, Cr, Mn and Cu; in the Netherlands & Switzerland Pb and Zn; elsewhere the class contains Pb and Cr. * the French regulations adopted the EC Directive but tightened the cadmium and mercury emission levels. ** represents total for each compound Hg and Cd. *** the German standard combines As, Co, Cr, Ni, V, Sn, Sb, Pb, Cu, and Mn. .... the U.K. standard combines As, Cr, Ni, Sn, Pb, Cu, and Mn. NA no standards for Category III. 1" Individual for each plant. 1"1" Total of Pb, Sb, Cr, Cu, Mn, V, Sn, As, Co, Ni, Se, Te should not exceed 1.0 mg/Nm 3 (max hourly average)

147 The regulation of trace metal emissions from incinerators is a new development in North America. The U.S. NSPS (1995) regulations are based upon the Maximum Achievable Control Technology, i.e., those demonstrated by the five best plants in operation, [MACT] criteria and these values have been adopted by the province of Ontario, Canada. Previously, Ontario had applied the Point of Impingement dispersion calculation procedure to determine acceptable stack emissions. The province of British Columbia adopted the TA Luft (1986) values in setting permit conditions for the Burnaby facility. Previously, in the U.So, NESHAP standards exist for lead, mercury and beryllium, along with PSD limits for lead with some state and local standards also existing (Jordan, 1987).

5.2.4 Trace Organic Emission Standards Standards for organics are listed on Table 5.6. These are more difficult to develop than inorganic standards because there are differences in toxicity within a family of organic compounds. For example, there are 210 congeners of polychlorinated dibenzo-pdioxins (PCDDs) and polychlorinated dibenzofurans (PCDFs) divided into eight homologues for each compound. Since each congener has a wide range of toxicity, scientists developed a scaling factor (the Toxic Equivalency Factor, TEF) for the 17 congeners that are considered to be the most toxic. A Toxic Equivalent (TE) value is then calculated by:

mE=~ (XixTEF,) i=1 ,n

where

X~ = concentration of a specific congener; and, TEFi = the toxic equivalency factor for that congener.

This cumulative value can then be used to estimate the potential total toxicity relative to a single congener, namely 2,3,7,8-TCDD, which is considered the most toxic congener of the PCDD/PCDF group. The most common scheme for applying PCDD/PCDF toxicity factors is shown in Table 5.7. Not all legislation has adopted the ITEQ approach. The Eadon equivalency factors used in some European legislation result in a value that is approximately twice the ITEQ value for the same sample. This implies that a value of 0.1 Eadon is equivalent to 0.05 ITEQ. With that in mind, the Danish value is close to the CCME Canadian level. Furthermore, while not considered particularly significant, there are subtle difference in the sampling and analytical methods used for these compounds in Europe and strict comparisons to North American numbers may lead to invalid conclusions. The German value results from measurements expressed on a wet basis and for a considerably longer averaging time than that conventionally used in North America. The net effect of this is that the German value equates to approximately 0.3 ng ITEQ/Rm 3 @ 11% oxygen. On the basis of toxic equivalents, the Swedish standard would appear to be the lowest at 0.05 ng/Rm 3.

148 Table 5.6 Municipal Solid Waste Incinerator PCDD/PCDF Emissions Limits Jurisdiction (Country/State/Province) PCDD/PCDF (ng/Rm 3) (Toxic Equivalents) Netherlands 1993

0.1

Sweden 1986 (Recommended)

0.1 Eadon

Germany 1990

0.1

U.S.A. 1995

9 (total) 88 (total) 21 (total) exceptESPequippedfacilities42 (total)

New >35 tpd Existing >35 & 225 tpd Canada CCME Guidelines 1988 British Columbia 1991 Ontario Guideline A-7 1995 Japan 1990

0.5 0.5 0.14

0.55

Table 5.7 Toxicity Equivalency Factors (TEFs) for Specific PCDD and PCDF Congeners Positions Equivalency Homologue Chlorinated Factor Dioxins TCDD 2,3,7,8 1 PeCDD 1,2,3,7,8 0.5 HxCDD 1,2,3,4,7,8 0.1 1,2,3,6,7,8 0.1 1,2,3,7,8,9 0.1 HpCDD 1,2,3,4,6,7,8 0.01 OCDD 1,2,3,4,6,7,8,9 0.001 Furans TCDF 2,3,7,8 0.1 PeCDF 1,2,3,7,8 0.01 2,3,4,7,8 0.5 HxCDF 1,2,3,4,7,8 0.1 1,2,3,7,8,9 0.1 1,2,3,6,7,8 0.1 2,3,4,6,7,8 0.1 HpCDF 1,2,3,4,6,7,8 0.01 1,2,3,4,7,8,9 0.01 OCDF 1,2,3,4,6,7,8,9 0.001 Note: * When only homologue test data are available, then the most conservative (largest) equivalency factor should be applied. NATO CCMS, 1988

149 The U.S. EPA 1995 standard for PCDD/F does not involve the use of toxic equivalents; rather compliance with the standard specified must be determined through the use of a specific sampling and analytical method that quantifies the amount of each of the 2,3,7,8 substituted congeners (isomers). These are totalled and the total is compared to the standard. This effectively uses the same congeners that are included in the ITEQ scheme, but applies no weighting to the various levels of these material. 5.3 CURRENT ASH AND RESIDUE DISPOSAL PRACTICES The bulk of the residue generated at an MSW incinerator consists of grate ash. In Canada, most European countries and Japan, bottom ash is handled separately from the other residue streams, whereas the current trend in the United States is to combine all the residue streams and dispose of this material in dedicated landfills. In most jurisdictions, bottom ash is disposed by landfilling and regulations governing this activity have been developed. However, utilisation of the material as a substitute lightweight aggregate has emerged as a viable option to landfill disposal and experience has led to various regulations being developed specifically for utilisation practices. Since the regulations and management practices for ash are evolving in all countries, the current trends and approaches are described below. 5.3.1 Disposal of Bottom Ash Canada In Canada, ash is currently handled as two separate streams, bottom ash and fly ash/APC residue. The CCME guidelines (1989) recommended ash be handled in this manner to prevent contamination of the bulk of the material by the high trace metals concentrations in the fly ash. Furthermore, the guidelines suggested that bottom ash could be disposed in a conventional municipal landfill, which is the current procedure in British Columbia, Ontario and Quebec, although the bottom ash from the Burnaby, B.C., facility is used for road construction purposes within a landfill. Denmark Since 1984, Denmark has utilised a very large portion of the bottom ash generated at its incineration facilities. In 1993/94, Denmark utilised between 400,000 to 450,000 tonnes of processed bottom ash which represents almost 100% of the total amount generated (Hjelmar, 1994). The ash is processed by screening and removal of ferrous materials to generate an upgraded product. Previously, approximately 26% of the total amount produced was disposed in dedicated monofills. The term monofill refers to segregation of ash from other waste materials, including MSW. A further 36% was used for fill or land reclamation purposes (Ludvigsen, 1991). The impetus for wide spread use of ash stems from the imposition of a State tax on disposal which was initiated in 1987 (Hjelmar and Ludvigsen, 1993).

150

France

France has a landfill regulation (Law on Waste Disposal, 1975, revised 1992) which suggests that landfilling is the last resort after all recyclable uses have been made of the material. This encourages incineration of waste after recycling and utilisation of residues where appropriate. Any material landfilled must contain less than 5% organic matter and the TOC of the leachate is limited. The act stipulates categories of materials according to its solubility. A material can be recycled if the solubility is less than 3%. There are two classes of landfills: Class 1, Hazardous Waste with a solubility greater than 5% and less than 10%; and Class 2 where the material has a solubility less than 5%. All landfills must be lined. Bottom ash is the only material that is currently considered appropriate for re-use, however, given the limits on recycled materials, most residues will require some treatment before they can be used or they will require disposal.

Germany and Switzerland In Germany and Switzerland, landfill disposal of materials requires that the residues meet a loss on ignition criteria (a measure of the unburned material in the ash) of less than 10% and contain less than 10% soluble salts. Furthermore, leachate from the residue must meet criteria for various trace metals based on elution with distilled water. In 1993, the German Bundesministerium f0r Umwelt issued a new directive on landfills used for both MSW and incinerator residue disposal. This legislation defines two classes of landfills based on the total organic carbon, loss on ignition at 550~ leachate quality as defined by DEV $4, and solubility. The new directive promotes utilisation as the preferred option for bottom ash. If there are no available markets for utilisation, disposal should be at a Type 1 landfill. After simple in-plant treatment, bottom ashes from properly operated incinerators will be able to meet the criteria (Schneider et al., 1994). These are summarised in Table 5.8. The objective of these standards is to reduce the reactivity of materials being placed in landfills.

Netherlands In the Netherlands, the Soil Protection Act of 1987 provides statutory authority for protection of the environment through limiting the pollution of soil by anthropogenic activities. Under the act, construction materials are regulated to prevent contamination from industrial residues that may be used in construction. The regulations limit releases to the environment to a small percentage of the existing level of that contaminant in the first metre of underlying soil. The effect of these regulations is to emphasise the efforts of controlling releases of contaminants found in very low concentrations in the Dutch environment. This is a distinct difference from many other jurisdictions where release

151 Table 5.8 Summary of German Landfill Requirements Standard Not To Be Exceeded Type1 ... Type 2

Classification Criteria For Landfills TOC (%)

1"

3**

LOI (%)

3*

5**

pH

5.5- 13

5.5- 13

Conductivity (uS/cm)

10,000

50,000

TOC (eluate, mg/L)

20

100

Phenols

0.2

50

Solubility (%)

...... 3

.

6

Leachate Quality (mg/L) (from DEV $4 Procedure) As

0.2

0.5

Cu

1

5

Hg

0.005

0.02

Zn

2

5

Cd

0.05

0.1

Cr§

0.05

0.1

Ni

0.2

1.0

F

5

25

Pb

0.2

1

NH4

4

200

CN

0.1

O.5

AOX * = for new incinerators TA Siedlungsabfall, 1993

0.3

1.5 ** = for old incinerators

152 limits are based largely on the toxicity of the contaminants. Under the Soil Protection Act, a separate regulation deals with the disposal of wastes (Regulation for Disposal). In the Netherlands, a large portion (>80%) of the bottom ash produced is utilised in embankment and roadbase applications. Ferrous rejects are recycled.

Sweden It is estimated that Sweden produces 400,000 tonnes of bottom ash and 60,000 tonnes of fly ash and APC residues annually (F~llman and Hartldn, 1992). This quantity fills 250,000 m3 of dedicated monofill space in recently approved disposal sites. Each site has its own permit requirements which were approved by the Environmental Franchise Board. Furthermore, monofills that are used for both bottom ash and APC residues must dispose of these streams in separate cells. Current recommendations suggest that leachate be collected for the initial filling period and after this time infiltration should be kept below 50 mm/year by the use of proper soil covers. The Swedish regulators are currently monitoring disposal requirements developing in the rest of Europe with a view to amending their standards. Regardless of the standards imposed, local citizens are afforded an opportunity to review and comment on any landfill development plans during the approval stages. Efforts to develop a suggested re-use criteria are also under way in Sweden as discussed in the next section. United Kingdom In the United Kingdom, no special provisions exist for the disposal of ash from MSW incinerators, although the issue is under review. All ash generated in society, be it from residential, commercial or industrial establishments, is classified as a "controlled waste". Controlled waste must be disposed at approved licensed facilities that can handle the material. Licensing requirements reflect the need to preserve the environment and ensure neither the water resources nor public health are endangered by the disposal practice. The current practice is to co-dispose with MSW or to use the material as cover in older landfill sites. These sites are under reducing conditions and the theory is that they present a more stable environment for the containment of trace metals. The regulations governing ash disposal are expected to change when the new Air Emissions Regulations force facilities to install new APC systems in 1996. United States Up until the mid 1980s, most MSW incineration residues in the U.S. were disposed in co-disposal situations with MSW. Regulations for disposal varied by state and local situation, and considerable debate and confusion existed about the status of these materials with respect to the RCRA Subtitle C (hazardous waste) testing and management requirements. This was brought about by an exemption for "household waste" from the provisions of Subtitle C. However, in the Spring of 1994, the U.S. Supreme Court ruled that MSW incinerator ash was no longer exempt from testing using the Toxicity Characteristic Leaching Procedure (TCLP). Thus ash (combined or

153 separated bottom ash and APC residues) which passes the criteria associated with the TCLP can be landfilled or monofilled, however, ash which fails the criteria must be disposed as a hazardous waste. This involves disposal in a secure landfill with provisions including a series of liners and leachate collection and treatment facilities which are more stringent than the design criteria for Subtitle C landfills. Most new facilities are using monofills for combined residue disposal, but where space is limited, interest in utilisation is increasing. Although the majority of the facilities combine the residue streams, a small number segregate the bottom and fly ash/APC residue streams to facilitate treatment of the fly ash/APC residue. Furthermore, regulations regarding ash management still vary widely from State to State. For example, New York State requires semi-annual testing for ash and are developing a procedure to handle this material as a special waste (e.g., Bill 10780, State of New York), whereas the State of Florida has permitted the use of ash in artificial marine reef construction projects. Moreover, although some States actively discourage the practice of co-disposal with MSW, other States endorse the practice.

Japan In Japan, the Waste Disposal and Public Cleaning Law, which addresses all aspects of waste disposal, was thoroughly amended in 1991. Under that law, incinerated ash is classified as either bottom ash or fly ash. Bottom ash is treated as normal domestic waste and disposed directly into sanitary landfill sites. 5.3.2 Disposal of Fly Ash and APC Residues The options for handling and disposing of the finer ash streams from incinerators are more limited. Most jurisdictions treat the material as a hazardous waste.

Canada In Canada, the fine ash material must be handled as a hazardous waste. The disposal options include transfer to a hazardous waste disposal facility or treatment of the residues prior to disposal. Various treatment alternatives from disposal in secure landfills to solidification are being evaluated, but there are few regulations in place to evaluate the efficacy of a treatment process. The exception is in British Columbia, where the treated ash must pass a battery of laboratory tests prior to disposal in a conventional landfill. The testing protocol includes evaluating the treated residue using chemical, engineering, durability and leaching tests (Government of British Columbia, 1992). Denmark & the Netherlands In Denmark, APC residues from the dry or semi-dry processes and fly ash are currently classified as hazardous wastes and are disposed in dedicated monofills with leachate

154 collection systems and bottom liners, and often with impermeable cover layers. Wet scrubber sludges are generally monofilled alone or are mixed with fly ash residues. All of these measures are only considered temporary solutions until suitable treatment systems are made available. At sites in Denmark and the Netherlands, APC residues are stored in polyethylene bags in landfills that have leachate collection systems and bottom liners. Generally, APC residues in the Netherlands are sent to a hazardous waste landfill site, although nearly 40% of the ESP residues from Dutch facilities are currently utilised as a very small percentage filler in asphaltic concrete mixes, but this practice is waning. This residue stream is segregated from all other streams for this purpose.

France

The 1991 French law on MSW incineration adopted the EEC directive on air emissions but has tighter mercury and cadmium standards. This has resulted in an increased use of wet APC systems, and hence, more sludge from these systems. The changes in regulations have fostered increased study into ways of modifying residues to meet the disposal criteria mentioned above. Immobilisation of contaminants by solidifying with hydraulic binders is being practised in some areas, and four organisations are currently exploring vitrification alternatives. One manufacturer is utilising the wet scrubber system to modify the residue to meet the criteria (Knoche, 1992).

Germany

In Germany, the APC system has to be designed in a way to minimise the production of harmful residues (Bundesministerium, 1993). Heat recovery system ash is separated from dry/semi-dry scrubber residues in some facilities. The fly ash and APC residues are classified as a hazardous waste and requires disposal in either approved landfills or preferably in underground disposal sites such ash old salt mines or in special cells of municipal waste disposal sites. Mehlenweg (1990) estimates that 5% of the total fly ash/APC residue stream (210,000-240,000 tonnes/year)is deposited in underground sites, less than 1% is re-used and the balance goes to surface storage. To minimise the release of dust from surface stored materials, it is packaged in large bags or moistened. German regulations allow boiler/economiser and filter ashes to be modified to reduce the need for controlled disposal, however, few methods have been developed to the commercial scale. Heat recovery and filter ashes along with APC residues contain high quantities of water soluble salts and in Germany they are required to be disposed of in hazardous waste landfills. Limited work has been done to explore the options for treatment/re-use of these materials, but these processes have not progressed beyond pilot scale (Juritsch, 1990; Kurzinger, 1990).

155

Netherlands At the present time, APC residues and fly ash are considered hazardous wastes and are generally managed in a similar manner to that used in Denmark. Sweden In Sweden, APC residues are disposed separately from bottom ash. The Environmental Franchise Board is responsible for setting the requirements for these disposal sites. It has been found that the Swedish infiltration limit of 50ram/year is not suitable for limiting releases from APC residues and fly ash. A new practice is to stabilise these materials before disposal. This is done at one facility in Stockholm by adding 40% low calcium cement to the residue stream. This increases the volume of the stream but further retards the infiltration rate into the material. Other options for the disposal of APC residues and fly ash are currently being examined, including slurry deposition to achieve better compaction, advanced immediate compaction to reduce permeability and the use of plastic covers during deposition to reduce infiltration. Japan In Japan, fly ash and APC residues are treated as a domestic waste under special control. Before disposal, they have to be tested via a leaching procedure and compared to waste disposal standards. In order to treat fly ash, the Ministry of Health and Welfare specified four treatments: 9Melting and solidification 9Solidification by cement 9Stabilisation using chemical agents 9Extraction with acid or other solvent After all standards have been passed, the treated fly ash could be disposed directly into sanitary landfill sites with other domestic wastes.

5.3.3 Utilisation Two fundamental concerns with utilisation applications are that 1) the physical properties of the material are appropriate for the intended application (i.e., bearing capacity, compaction, etc.), and 2) the application does not lead to environmental degradation. The latter situation relates mainly to the leaching of metals and salts from the ash, since the potential loading of ash within a fill application may pose a potential problem. In Europe, the materials are used as a civil engineering material, largely as base and sub-base for roadways. Each country has considered the environmental implications of these uses and developed guidelines for implementation. While the subject of utilisation is discussed in more detail later in the document, a brief discussion of existing regulations governing the use of residues is included here.

156 Canada As mentioned previously, although no major efforts have been devoted to utilisation in Canada, the Greater Vancouver Regional District has evaluated bottom ash for utilisation applications, and currently uses the material for construction of roadways within a landfill site. The bottom ash undergoes ferrous removal prior to leaving the Burnaby facility, but no other processing is done other than compaction during placement.

One of the major impediments to bottom ash utilisation is that there has been little economic incentive to divert materials from landfill. Should sufficient regulatory criteria be put in place to allow the use of bottom ash as a lightweight aggregate, it is likely that the practice would be considered for ash from some of the major facilities. Denmark Although part of the bottom ash stream from incinerator facilities in Denmark has been used in sub-bases for roads, bicycle paths and parking lots since 1974, the first Danish utilisation requirements were not developed until 1983 (Statutory Order No. 568 of Dec. 6, 1983). Moreover, these requirements only applied to the use of small to moderate amounts of ash. Large scale applications (>30,000 tonnes or 5 m thickness) are regulated under the Environmental Protection Act (Disposal and Discharge Permit section). Additional guidelines for road sub-bases were developed in 1989 by the Danish Highway Department (Pihl et al., 1989 in Hjelmar, 1990). The Statutory Order is currently being reviewed. Ferrous material is removed from the ash by screening and then magnetic separation to generate an upgraded material for recycling purposes. The portion of the bottom ash stream which cannot be used is disposed in dedicated monofills.

A Danish testing protocol has been developed to determine the suitability of ash for utilisation based on chemical parameters (Hjelmar, 1990). The conditions include a pH >9 for a 1% slurry of the material, alkalinity of >1.5 eqv/kg, metals levels as determined from a HNO3 leach of Pb 99%). The first one applies a thermal treatment in a rotary kiln at about 400~ under oxygen deficient conditions, then copper salts are added as catalysts. A full scale facility is in operation (Schetter et al., 1990). A second process, the 3R Process, utilises the combustion chamber of the incinerator itself to decompose PCDD/PCDF in extracted and compacted filter ashes (Merz et al., 1989). All vitrification processes proposed for APC residues may also be appropriate for treatment of PCDD/PCDF as well.

320 Figure 8.27 Incinerator

Concentrations (TE) and Partitioning of PCDD/PCDF in a Modern

Based on the information given above, a modern MSW incinerator which is well operated and equipped with adequate APC devices is capable of meeting the most stringent emission regulations for PCDD/PCDF. In some European countries, this limit is 0.1 ng(TE)/Rm 3, however, even lower limits have been achieved during recent pilot plant studies ( 4.75 mm. "ASTM C127, C128.

Figure 9.7 Bottom Ash Absorption as a Function of Time 30

Absorption,

%

14

Absorption,

F i n e Ash

H!I

25

20

%

..... Hourly Average

Overall Average

C o a r s e Ash

tjIIttttLIIjtt tII tt ttittittittitF1 ..... Hourly Average

0

Overall Average

I

1

[

I

t

I

2

4

6

8

lO

12

.~ampllng

_ ~ L

14

[

_._1~__

16

18

20

Day

0

I

I

1

I

I

I

1

[

I

2

4

6

8

10

12

14

16

18

Sampling

Day

20

The vertical bars are the 95 % confidence intervals After Eighmy et al., 1992

Unit Weight

Table 9.9 provides information on bottom ash unit weight from a number of facilities in the United States and Canada. Typical mean values range from 955 to 1,420 kg/m 3. Typical minimum and maximum values are 732 and 1,510 kg/m 3 for all measurements, respectively. There is reasonably good agreement about bottom ash unit weight values for each of the facilities that were evaluated. These unit weight values show that bottom ash is a lightweight aggregate.

357 Table 9.9 Bottom Ash Unit Weight Country Facility Canada GVRD

Unit Weight, kg/m 3 Reference Min Max Mean Median n 1,3701,5101,420 4 WASTE Program, 1993

Southwest Brooklyn, Nu a,b 732 1,229 1,054 62Chesner et al., 1988 Concord, NH ",c 1,039 1,234 1,157 1,159 20 Eighmy et al., 1992 Dry Scrubber I a'b 1,090 1,183 1,152 4 LIRPB, 1992a Dry Scrubber 2 a'b 955 956 955 2 Dry Scrubber 3a,b 1,102 1,378 1,215 5 Dry Scrubber 3",b 1,150 1,312 1,234 5 c Fraction less than 19 mm. a ASTM C29. b Fraction less than 50.8 mm.

United States

Figure 9.8 shows bottom ash unit weight as a function of time at the Concord, New Hampshire facility (Eighmy et al., 1992). As can be shown in the figure, bottom ash unit weight was relatively constant over the 1.5 year sampling period.

9.1.3 Gradation Figure 9.9 shows a typical bottom ash grain size distribution. The distribution is classified as well-graded, meaning that there is equal abundance of coarse and fine material. Such uniform gradation is important to the compactability of bottom ash and the potential to utilise bottom ash as an aggregate substitute. Table 9.10 provides information on bottom ash effective size and uniformity coefficients. Effective size is determined by calculating the grain diameter where 10% of the material is passing. Coefficient of uniformity is the ratio of the grain diameter in millimetres corresponding to 60% passing by weight to the effective size of that material. As can be seen in Table 9.10 the mean effective size of bottom ash is 0.293 mm. Minimum and maximum values for all measurements are 0.127 and 0.508 mm. The mean uniformity coefficient was 21.6 with minimum and maximum values of 12.8 and 38.0. The data that are shown in Table 9.10 are from only one facility in the United States. The effective size and uniformity coefficients indicate that bottom ash is a wellgraded gravelly sand. Table 9.10 Bottom Ash Effectiv e Size and Uniformity Coefficients Country Facility Effective Size, mm I UniformityCoefficient Reference Min Max Mean Median n I Min Max Mean Median n USA Concord, ... NHa 0.1270.5080.2930.25472112,82138.02!.:6820.6872 Eighmy et al., 1992 " Fraction less than 19 mm.

358 Figure 9.8 Bottom Ash Unit Weight as a Function of Time 1300

Unit

Weight,

kg/m3

1200

1100

1000

..... 95~0 C. I.

Average

900 0

1

i

i

t

i

L

i

2

4

6

8

10

12

14

Sampling

_.!

i~

16

18

Day

After Eighmy et al, 1992 Figure 9.9 Bottom Ash Grain Size Distribution 100

Passing,

Percent

0 0.01

i

J

iJ~lllJ

%

i

....

i

Jl~Jlll

0.1 Particle

After Eighmy et al., 1992

|

,

,Jl

1 Size. mm

100

359 Figure 9.10 shows how bottom ash effective size and uniformity coefficients vary as a function of time for bottom ash samples obtained from the Concord facility. The data indicate that both effective size and uniformity coefficient were reasonably constant over the 1.5 year sampling period. Figure 9.10 Bottom Ash Effective Size and Uniformity Coefficient as a Function of Time 0.6

Effective Size, mm

40

Uniformity

Coefficient

35

0.5

30

0.4

,-.. li

i. , ,:,'I

25

0.3

20

. i!/!

0.2 0.1 ..... Hourly Average

0

..... Hourly Average

Overall Average

1

i

1

i

i

i

i

i

i

2

4

6

8

10

12

14

16

18

Sampling

Day

0

20

0

Overall Average

i

t

i

i

i

i

i

i

i

2

4

6

8

10

12

14

16

18

Sampling

Day

After Eighmy et al., 1992 Percent Fines The concentration of fine material in bottom ash is an important consideration when bottom ash is to be used as an aggregate substitute. The percent fines can frequently create problems because that fraction is highly absorptive for water, asphaltic cement and Portland cement. Frequently, high fine contents create a material that has a tendency towards freeze-thaw susceptibility and durability failure. The values for percent fines shown in Table 9.11 are from facilities in Germany, the Netherlands and the United States. The fraction denoting fines can be different in Europe compared to the United States. In Europe a fine fraction is usually denoted as that material passing a 63 pm mesh sieve. In the United States, a fine fraction is denoted as that material passing a 75 pm mesh sieve. Nevertheless, mean values for fines range from about 1.9 to 7.4%. Minimum and maximum values for all measurements are 1.0 and 10.1% respectively. There is good agreement between the values seen from the different facilities in the different countries.

360 Table 9.11 Bottom Ash Percent Fines Country Facility Min Max a Germany A 2 7 Bla,b 3 8 B2,.c 2 7 C" 2 6 Netherlands AVI 1' 6.2 9.5 AVI 2" 3.7 7.2 AVI 3" 4.8 9.6 AVI 4' 4.3 10.1 United Southwest Brooklyn, Nud,e 1.0 3.2 States Concord,NH~'~ 2.17 6.57 Dry Scrubber 3d,~ 2 3 Dry Scrubber 3d,~ 1 3 a Finesdenoted as fraction passing 63 pm. d b Unit 1 e c Unit 2

Reference Fines, % Mean Median n 4 5 Vehlow, 1992 5 4 5 4 4 4 7.3 29 TAUW, 1988 5.9 26 7.4 26 7.3 26 62 Chesneret al., 1988 1.9 72 Eighmy et al., 1992 3 . 9 6 3.92 10 LIRPB, 1992a 2 2 10 Finesdenoted as fraction passing 75 pm. ASTMC136 (dry sieving technique).

As bottom ash exhibits some friability, the production of fines during processing operations may occur. The levels of fines in bottom ash are somewhat problematic with regards to the utilisation of bottom ash in civil engineering construction applications because fines increase water capillarity and promote frost susceptibility. This means that under certain utilisation scenarios this fine fraction may need to be removed from the bottom ash process stream. At certain ash processing facilities in Europe, fines are removed by trommeling processes using agricultural trommels. Figure 9.11 shows how percent fines vary as a function of time at the Concord facility (Eighmy et al., 1992). The data indicate that the fine fraction was relatively uniform over the 1.5 year sampling period. The variations seen over time were as great as the variations observed within four consecutive hourly sampling events.

9.1.4 Durability The assessment of the durability of bottom ash is an important characterisation when considering the utilisation of bottom ash. Frequently, bottom ash is considered as an aggregate substitute and it is important to characterise how durable bottom ash is in comparison to natural aggregates.

Soundness The sodium or magnesium sulphate soundness test is generally the accepted method for aggregate soundness testing. Data are provided in Table 9.12 on percent losses

361 observed to bottom ash samples subjected to soundness testing. The data are from facilities in the United States. Soundness testing usually involves evaluation of both fine (< 4.75 mm) and coarse (> 4.75 mm) fractions. The data shown in Table 9.12 indicate that for the fine fraction, mean percent losses ranged from 1.6 to 11.91. The coarse fraction mean values were 2.6 and 2.9. As with sorption, the fine fraction is more susceptible to expansive fragmentation compared to the coarse fraction. This is because bottom ash fine material is more porous than the coarse material. There is a wide variation in values for the fine fraction seen between facilities within the United States. It is not clear as to why this variation exists. Figure 9.11 Bottom Ash Percent Fines as a Function of Time Percent

Passing,

%

..... H o u r l y

0

--

0

i

2

Average

Overall A v e r a g e

t.

P

~

~

L

_l

J

4

6

8

10

12

14

16

Sampling

Day

18

20

The vertical bars are the 95% confidence intervals After Eighmy et al., 1992 Abrasion Resistance The Los Angeles abrasion test measures the ability of an aggregate material to maintain its physical integrity under defined abrasive conditions. The test is conducted on two different size fractions, a coarse and a fine fraction, termed "B" and "C", respectively. The test is considered to be highly aggressive with respect to evaluating lightweight porous aggregate materials. Table 9.12 provides data on LA abrasion resistances for bottom ash B and C fractions from facilities from the United States. Typical percent losses observed for both fractions are around 40 to 45%. These values are considered to be high, however they are typical for porous lightweight aggregate materials.

362 Table 9.12 Bottom Ash Durability Country Facility

United States

Concord, NHc Dry Scrubber I d Dry Scrubber 3 d

Soundness % Lossa Fine Coarse Min Max Mean Median n Min Max Mean Median 2.63 10.38 14.32 11.91 11.48 4 2.51 2.76 2.63 . . . . 1.7 3.4 2.7 4 2.9 1.1 2.4 1.6 5 1.6 4.0

Country Facility n

" b

Concord, NH c Dry Scrubber 14 Dry Scrubber 24 Dry Scrubber34 Dry Scrubber34 ASTM C88. ASTM C131.

5

LA Abrasion, % Loss b C Fraction

B Fraction United States

n 4

Min

Max Mean Median

4 46.4 48.2 - 57.8 58.5

47.3 58.2

. . . . 5 57.6 61.0 - 46.0 61.8

. 59.6 54.9 c d

47.3 -

n

Min Max Mean Median

n Reference

2 4

42.6 44.2 43.4 40.4 40.7 40.6

43.4 -

2 E i g h m y e t al., 1992 2 LIRPB, 1992a

5

43.1 46.1 44.6 44.9 48.3 46.6

-

2 5

6

41.3 47.5 45.0

-

6

Original fraction less than 19 mm. Original fraction less than 50.8 mm.

9.1.5 Geotechnical Properties Many of the utilisation scenarios envisioned for bottom ash involve the use of bottom ash as an aggregate substitute subjected to compaction. Despite the potential problems of durability with bottom ash, bottom ash is a highly compactible material that upon compaction has high levels of E-modulus and strength. The fine fraction in bottom ash, the retentive capacity of bottom ash for holding water and the porosity of bottom ash mean that careful attention must be given to bottom ash water content prior to compaction.

Proctor Compaction The compactibility of a granular material is frequently assessed in the laboratory using Proctor compaction testing. Figure 9.12 shows a typical Proctor compaction curve with an optimum moisture content at maximum dry density. Table 9.13 provides data on bottom ash Proctor moisture optimums as well as maximum Proctor densities. Data are provided from facilities from the Netherlands, Sweden and the United States. As can be seen in the table, mean Proctor densities range from 1530 to 1739 kg/m 3. The minimum and maximum values that are observed are 1242 to 1838 kg/m 3 for all measurements. Good agreement is seen between facilities from the different countries. The geotechnical moisture content, at which maximum compaction occurs, tends to range from about 9.6% to 20%, with typical mean values of 13 to 16%. These moisture optimums are similar to those seen for gravelly sands in their ability to allow compaction to occur. There is good agreement between facilities from different countries with regard to moisture optimums that are observed.

363 Figure 9.12 Bottom Ash Proctor Compaction Curve Dry Density,

1.9

1000 k g / m 3

Typical Ash Zero Air Voids

1.6

--

8

I

I

,I

12

16

20

Moisture


Content, %

Table 9.13 Bottom Ash Proctor Moisture and Proctor Density Compaction Country

Facility

Proctor Density (kg/m 3) Min

Max

Proctor Moisture (% W C )

Mean Median n Min Max Mean Median

Reference n

Netherlands AVI 1"

1,513 1,665

1,602

-

29 11.9 16.5

13.3

-

AVI 2"

1,543 1,630

1,573

-

26 10.9 16.0

13.0

-

16

AVl 3'

1,445 1,620

1,530

-

26 10.6 18.7

14.2

-

26

20 9.6 16.5 1,825 1 1,748 18 12.0 16.0 1,345 1 2 15.0 17.0 5 14.3 14.8 5 20.8 21.7

12.9 15.5

..... -

20 1 Hartl6n & Ro~lbeck, 1989

15.4 -

16.0 12.8

18 Eighmy et al., 1992 1 LIRPB, 1992a

16.0 14.6

-

21.3 -

16.6

Sweden United States

AVl 4" 1,475 1,630 1,530 Maim6 ~ ......... Concord, NH c 1,619'""'1,838 ""'1,739 Dry Scrubber 1d Dry Scrubber 2 d 1,242 1,298

1,271

Dry Scrubber 3 d 1,500 1,588

1,545

Dry Scrubber 3 d 1,550 1,580

1,566

Mass burn d

Standard Proctor. Standard Proctor.

-

-

-

c d

2,463

1

-

-

29 TAUW, 1988

2 5 5 1 Kosson et al., 1992

ASTM D1557 (Modified Proctor). ASTM D1557 (Modified Proctor).

The E-modulus can be evaluated for bottom ash materials as a function of Proctor compaction. Work by Hartl~n and Elander (1986) shows very high E-modulus values for ashes compacted in the fresh form and aged. The E-modulus values that are observed indicate that bottom ash is a strong aggregate material when it is in a compacted state.

364

Field Compaction Field compaction can be an aggressive, energetic process. Frequently, the compaction methods that are used in the field can break down particles in bottom ashes. Figure 9.13 shows how bottom ash particle size distributions become finer after field compaction efforts. The data, provided by Hartl~n and Rogbeck (1989), show that most full scale field compactors will fracture and fragment bottom ash and change the grain size distribution. At this time it is not clear if the degree of grain size redistribution is problematic with respect to civil engineering structural fill applications. Many bottom ash utilisation studies have shown that bottom ash can be successfully used as a compacted aggregate material in road sub-bases or in wind barriers and embankments. However, the fines content may require control as this influences frost susceptibility in cold climate applications. Figure 9.13 Bottom Ash Particle Size Distribution after Field Compaction Siltl

"

I

sand

Gravel

.... I

Silt

-,.., r r i._ C~.

100

0.06

0.2

0.6

2

90

6

20

./;y

/M

80

70 V e-

/

50

*~9

so

9

70

0

10

~

0 0.063

1

).2

0.6

4

Mesh Opening, mm

2

6

20

60

/,Y 7

30

S

20 10

0.5

Gravel

/,#"

*-~11 0.25

I

,/

-~

20

San~

1 0 0 0.06

~

0 r

~>

I

Grain Size d. in mm

Grain Size d. in mm

r

_

1

16

4

16

63

Mesh Opening, mm

2 months old

1 year old

The shaded area represents the original grain size distribution The solid line (A) shows the new grain size After Hartldn and Rogbeck, 1991 California Bearing Ratio (CBR) The CBR is a determination of the strength and stability of a compacted material. Values greater than 100% for CBR are seen in bottom ash. Table 9.14 provides information on CBR values at 0.1 inches and 0.2 inches for bottom ash samples obtained from facilities in either the Netherlands or the United States. Mean values seen at 0.1 inches range from 51.8 to 79.7% with minimum values at 22 and maximum values of 112.5 for all measurements. The CBR at 0.2 inches is higher. Typical mean

365 values are 39.0 to 154.5%, with minimum and maximum values ranging from 32.0 to 167.3% for all measurements. There is not good agreement between the data from different countries, but it is not clear why. Table 9.14 Bottom Ash Penetration Resistance Country

Facility

CBR @ 0.1 Inches a

CBR @ 0.2 Inches a

Reference

Min

Max

Mean Median n

Min

Max

Mean Median n

Netherlands AVI 1

24.0

65.0

52.0

-

29

32.0

76.0

62.4

-

AVI 2

38.0

62.0

51.8

-

26

52.0

74.0

61.2

-

26

AVI3

22.0

42.0

31.7

-

26 28.0

50.0

39.0

-

26

46.0

34.1

42.1

AVI4

27.0

United

Concord, NH

63.0 112.5

States

Dry Scrubber 1

.

.

.

.

1

Dry S c r u b b e r 2

.

.

.

.

2

121.0 158.7 139.9

-

2

Dry Scrubber 3

.

.

.

.

5 122.7 167.3 154.5

-

5

Dry Scrubber 3

.

.

.

.

5

-

5

79.7

80.0

20

34.0

58.0

20

92.0

136.5 110.2

29 T A U W , 1988

-

-

38.7

-

126.0

20 107.5 2 0 E i g h m y e t a l , 1 9 9 2 146.0

90.1

1 LIRPB, 1992a

Figure 9.14 shows how CBR varies as a function of time at the Concord facility. The data show that CBR exhibits some variability as a function of time. Figure 9.14 Bottom Ash CBR as a Function of Time CBR, %

CBR

a t 2.54

mm Penetration

CBR '

9O

130 1 120

a t 5.08 m m P e n e t r a t i o n

@~

f" ...........

.

ll0t. -

--

70

i

100 I

Average 50

i

~ _ L

0

2

4

6

L

8

i

..... 95% C. I.

10

L___

12

S a m p l i n g Day


~I___~L 14 16

i 18

80

70

I

0

Average

..... 05% C I

L

t

i

I

I

I

2

4

6

8

10

12

q9a m p l i n g

Day

14

16

18

,

366

9.1.6 Permeability The permeability of bottom ash, or its ability to transmit water via percolation, is an important component with regards to characterising the hydraulic regime to which bottom ash can be subjected. Because bottom ash is a well-graded material and can be compacted to high densities, it is expected that under compactive efforts the permeability of bottom ash will be quite low. Frequently, an assessment of the permeability of bottom ash is needed to model leaching of bottom ash, to model water balances of water moving through bottom ash and to assess the ability of bottom ash to freely drain. Utilising permeability testing apparatus, permeabilities that have been observed in bottom ash are usually in the low 106 cm per second range. Table 9.15 shows some bottom ash permeabilities obtained in studies conducted in both Denmark and Sweden. At maximum density, it appears that bottom ash permeability can range from about 0.2 to 10.0 x 106 cm/s. Such permeabilities are considered to be relatively low for wellgraded materials and reflect the presence of fine material which increases the tortuosity within bottom ash. Such low permeability values suggest that bottom ash may be subject to some infiltration but could also create some surface runoff. Table 9.15 Bottom Ash Permeability Permeability 106 cm/s

Reference

-

3.5-4.4

Geoteknisk Institute, 1992

Malm5

0.2-10.0

Hartl~n & Elander, 1986

Country

Facility

Denmark Sweden

9.1.7 Influence of Combustor Type and Operation on Physical Characteristics There has not been a great deal of study conducted on the influence of combustor type and combustor operation on the physical properties of bottom ash. The comprehensive N ITEP program was the only large scale study that has been conducted to date that has looked at the influence of poor combustor operation and combustor type on ash characteristics. The only data that is available at this time is information on the loss on ignition content for a variety of facilities operated under both good and bad conditions. As shown in Figure 9.15, two-stage systems operated either under good or bad conditions produce bottom ash with significantly higher LOI values than mass burn or RDF systems.

367 Figure 9.15 Influence of Combustor Type on Bottom Ash LOI LOI

30-

r... o ~

20-

tO O) O) 0 _J

10-

0 -

[; 2_Stage

I I

Mass_Burn

!

I

RDF

9.1.8 Influence of Aging on Bottom Ash Physical Characteristics There have been some studies conducted in Sweden, Germany and the U.S. on the influence of aging on certain physical characteristics of bottom ash. In Sweden, it has been shown by Hartl~n and Rogbeck (1989) that the E-modulus of bottom ash will increase as bottom ash ages over time when compacted at optimum moisture under Proctor compaction testing. This increase in strength is attributable to the formation of mineralogical phases that increase particle interlocking within the bottom ash. Additional studies on aging in Sweden from bottom ashes at the Malmo facility have shown that when ash is aged for almost a year, the Proctor compaction characteristics are much better than when ash is freshly collected (Hartl~n and Elander, 1986). Again, aging is thought to increase the formation of certain mineralogical phases that increase the durability of the residue and interlocking characteristics of the particles in the residue. Studies in Sweden have also evaluated the influence of aging on the gross composition gradation of bottom ashes generated from the Malmo facility. There do not appear to be any significant differences between the nonmagnetic fraction, the glass fraction, the ceramic material, stone material and organic material in either fresh or aged fractions (Hartl~n and Lundgren, 1992).

368 Studies have been conducted in Germany to look at the influence of aging on a number of civil engineering properties. The data, presented by Vehlow (1992), show the influence of aging on leachable solids in bottom ash as bottom ash ages. The concentration of leachable solids in bottom ash decreases during aging. This is particularly true for facilities A and C (see Table 9.1). Facility B did not exhibit the same trends. The susceptibility to freeze/thaw fracturing has also been evaluated for aged materials in German facilities and the data suggests that the susceptibility to freeze/thaw erosion decreases with aging. This is particularly true again for facilities A and C. The data from facility B does not support this observation. Also evaluated in the German study was the raw density of aged material compared to fresh material at the three facilities. In all cases the raw density tended to increase with aging. This phenomenon is not presently understood.

9.2 PARTICLE MORPHOLOGY, MINERALOGY, AND ALKALINITY OF BOTTOM ASH Particle morphology, mineralogy and alkalinity of bottom ash play important roles in both the physical and chemical characteristics of bottom ash. The particle morphology of bottom ash is an important component in its physical characteristics and performance because of the angular nature of bottom ash particles. Bottom ash also tends to be a rough-textured material and this is an important property with regards to its physical performance. The mineralogy of bottom ash is thought to be important to understanding the leaching behaviour of bottom ash; however, the mineralogy also plays an important role in the compactibility and strength development of bottom ash as it ages with time. Finally, buffer capacity is an important component for both physical and chemical performance because of the role of the carbonate buffer system in bottom ash and the influence that has on strength development, particle aging and leaching.

9.2.1 Morphology Figure 9.16 provides some scanning electron microscopy micrographs and petrographic thin section micrographs of bottom ash particles. As can be seen in the SEM micrographs, bottom ash is an angular material. Slag-like material can be seen; the slag material is porous and contains vesicles. The petrographic thin section of bottom ash clearly shows that bottom ash possesses a high degree of internal porosity that is connected to the exterior of the particles. This vesicle porosity provides a large surface area for chemical reactions to take place and for leaching phenomenon to occur. As shown in Table 9.16, a number of researchers have looked at the specific surface area of bottom ash using BET absorption isotherms. Typically, bottom ashes have surface areas of 3 to 46 m2/g dry weight of bottom ash. These are considered to be very high degrees of surface area for a granular material. For instance, traditional

369 soils have surface areas that are orders of magnitude less than that. Mercury porosimetry is also used to look at internal pore diameters as can be seen in Table 9.16. A number of the pores in the material are considered to be quite small in nature. Values of less than a tenth of a micron are seen. The data also suggest that combustor type has an influence on the type of surface area and pore diameter that is found in bottom ash. The reader should refer to Chapters 12 and 13 for further discussion on the role of surface area in chemical leaching phenomenae. Figure 9.16 SEM Micrographs (a,b) and Petrographic Thin Section Micrographs (c,d) of MSW Bottom Ash

370 Table 9.16 Bottom Ash Surface Area BET Country Facility Surface Area m2/g

Pore Diameter" IJm

Reference

RDF b 4.605 Gardner, 1991 PRF b 3.286 0.0947 Gardner, 1991 RDF b 9.469 0.0786 Gardner, 1991 RK-MB b 28.184 2.117 Gardner, 1991 Unknown 9.4-46.3 Theis & Gardner, 1990 MB b 3.2 0.0342 Kosson et al., 1992 Mercury porosimetry. RDF=Refuse-Derived Fuel, PRF=Processed Refuse-Derived Fuel, RK=Rotary Kiln, MB=Mass Burn

United States

In comparing the petrographic thin sections of bottom ash with the surface area and porosimetry data provided in Table 9.16, it is clear that the internal porosity that is connected to the exterior accounts for a great deal of the surface area measured by nitrogen BET absorption isotherms. Compared to most aggregates, bottom ash is considered to be a light-weight porous aggregate with more angularity and more surface roughness and textures than many traditional aggregates.

9.2.2 Mineralogy The mineralogical characteristics of bottom ash play a very important role in the leaching behaviour of bottom ash. It is estimated that bottom ash contains numerous mineral phases. Such diversity complicates our understanding of the leaching behaviour of bottom ash and more research needs to be conducted on the role of mineralogy in bottom ash aging and strength development. Nevertheless, there are at least four studies that have been conducted that have examined the mineralogy of bottom ash. The studies have been conducted by St~mpfli (1992), Vehlow et al. (1992), Kirby and Rimstidt (1993) and Eighmy et al. (1994). All four of these studies have used rigorous procedures employing x-ray powder diffraction (XRPD) and other methods as precise procedures for estimating the nature of the mineral phases. St~mpfli (1992) has examined bottom ash for the presence of those mineral phases that are associated with strength development as bottom ash ages. St~mpfli used XRPD to determine the mineralogy. Minerals identified include SiO2, CaCO3, Fe304, Fe203, Fe, FeO, Ca~,I(OH)7.6.5 I--120,Na2Si20~, and CaSO4. Others are shown in Table 9.17.

371 Table 9.17 Mineral Phases in Bottom Ash (in relative order of decreasing abundance) St~impfli (1992)b Vehlow et al., (1992)o Kirby and Rimstidt Eighmyet al. (1993)d (1994)e SiO2 CaCO3 Fe304

Fe304 SiO2 (Ca, Na)2(AI,Mg)(Si,AI)207

SiO2 CaSO4 o 2 H 2 0 3(AI203)~ TiO2

Fe203 Feo

CaCO3 KAISi308

Fe203 FeO

FeAI204 SiO2

FeO

NaAISi308 CaAI2Si208

CaSO4 KCI

CaG(PO4)2 Fe203

FeCr204 Ca(Mg,Fe)Si206 Fe2SiO4 Cr203 Fe203 CaMgSiO4

NaCI

CaSO4 CaO AI(OH)3 NaCI ZnCI2 NaAISi308

Ca2AI(OH)T.6.5H2 O Na2Si2Os CaSO4 (Ca, Na)(Al,Si)2Si8 NaAISi308

AI203 Ca(OH)2 CaSO4

b Basedon XRPD c Basedon petrographyand XRPD

Ca2AI2SiOT MgCa2Si207 Fe304

AI2SiOs TiO2 Based on XRPD Basedon petrography,XRPD,XPS, SEM/XRM

Vehlow et al. (1992) have conducted extensive characterisations of bottom ashes from three facilities in Germany. They used XRPD and petrography. Data are presented in Table 9.17. The principle phases found in ash from the German facilities are glass, magnetite, quartz, melilite and feldspar. A number of other minor phases were also identified. Agreement was seen in the relative presence of major phases amongst the three facilities. Vehlow et al. (1992) also looked at aging effects. Kirby and Rimstidt (1993) studied bottom ashes containing small quantities of fly ash. They used XRPD as well. Principle minerals include (% abundance) Fe203 (3.7%), CaCO3 (3.5%), NaCI (0.5%), SiO2 (2.3%), magnetic spinel (3.5%), TiO2 (1.1%), and CaSO4 o2H20 (1.8%). The majority of the non-LOI mass of the sample was amorphous glass and minerals present below the detection limit for XRPD. Table 9.17 provides further information. Eighmy et al. (1992) examined the characteristics of bottom ash using XRPD, petrography, SEM/XRM and surface microanalytical techniques for samples from a U.S.

372 facility. The bottom ashes were ground and separated using magnetic and density gradient separation procedures. Table 9.17 provides a summary of the data. Many of the phases found in the U.S. bottom ashes are similar to the ones identified by the other studies.

9.2.3 Alkalinity The buffer capacity of bottom ash is an important component in the leaching characteristics of bottom ashes. The acid neutralising capacity of the residue is a measure of how many milliequivalents of nitric acid are required to reduce the pH of one gram of residue to a value of 4.3. The endpoint of the titration can vary. Some researchers use a value of 7.0, while others use the more traditional carbonate alkalinity endpoint. To put this measure into perspective, one gram of residue would need to be leached with 45 litres of acidic precipitation to reduce the pH from 12.0 to 7.0. Table 9.18 provides some information on the acid neutralising capacity as well as the initial pH or the inherent pH of bottom ashes for samples collected from Canadian and U.S. facilities. Typically, bottom ash has an initial pH ranging from 10.5 to about 12.2. This is in part due to the presence of calcium hydroxides produced from CaO hydrolysis in the bottom ash. The acid neutralising capacity of bottom ash ranges from about 1.2 to 4.1 milliequivalents per gram. This means that bottom ash is reasonably well-buffered. Such buffering capacity indicates that bottom ash can moderately resist changes in pH. Table 9.18 Bottom Ash pH and Acid Neutralising Capacity Country

Facility

Canada

LVH SWARU QUC

Initial pH

.

ANC, meq/g Reference

10.20

3.05

Sawell et al., 1989b

-

4.11

Sawell et al., 1989a

11.39

2.15

Sawell and Constable, 1988

United States Concord, NH 10.5-12.2

1.2-3.0

Eighmy et al., 1992

Figure 9.17 provides a typical titration curve for bottom ash. The data indicate that there are a number of locations in the titration curve where a slight degree of buffering takes place. These buffers tend to occur at a pH of around 10, 8 and 5. Such locations for buffers are attributable to the carbonate system. Figure 9.18 shows the change of botom ash acid neutralising capacity as a function of time for bottom ashes collected from the Concord facility (Eighmy et al. 1992). The acid neutralising capacity is relatively variable over time.

373 Figure 9.17 Bottom Ash Titration Curve I0.000 8.000

Ave A N C - 2.1 m e q / g

6.000 "i-

4.000 2.000 0.000

. . . .

I

i

J

,

1.0

0.0

9

i

. . . .

!

2.0

. . . .

!

3.0

. . . .

I

4.0

. . . .

5.0

!

. . . .

!

6.0

. . . .

i

7.0

. . . .

8.0

m e q / g dry bottom esh After Eighmy et al., 1992 Figure 9.18 Bottom Ash ANC as a Function of Time 4.5 0

First Hour 9 Second Hour Z~ Third Hour A Fourth Hour 9 Daily Composite i l l

4.0 ol O" Q)

3.5 9

E 3.0

0 z 10,000 mg/kg), some are present as minor constituents (>1,000 but u} C1.

~ooo

E o

0

9

2s

&

7's

~;o

i

~3o

Days

Kullberg et al., 1989 11.4 PARTICLE MORPHOLOGY AND MINERALOGY

During the WASTE Program (1993), qualitative analysis was conducted on some of the dry APC system fabric filter residue samples using scanning electron microscopy (SEM) to determine the morphology of the particles. The particles were categorised into the 5 types listed in Table 11.4. In addition to the identification exercise, the relative percentage of the different particle types was determined optically (Table 11.4). The most common particle types in the dry scrubber residue were the polycrystalline and opaque irregular shaped particles, whereas the overwhelming majority of the fabric filter residue were polycrystalline in structure. The dry scrubber residue characteristics were very similar to those of heat recovery system ash (see Chapter 10), whereas the morphology of the fabric filter residue was dominated by the powdered lime particles encrusted with flue gas condensation/reaction products.

451 Table 11.4 Summary of Morphological Characteristics and Estimated Relative Percentages of Particle Types in Dry APC System Residues Morphology Estimated Relative % Size Range Fused Spheres Crystalline Polycrystalline Opaque Char

DS

FF

DS

FF

15 10 30 30 15

5 230 150 230

Bulk

Particle Size (microns)

10-

6,-

O"

.)141

4260

61110

111150

151230

>230

Bulk

Particle Size (microns)

Stuart, 1993 LOI values ranging between 5.9 - 10% (w/w) have been recorded for dry/semi-dry APC system residue from two-stage incinerators (Environment Canada, 1993). In addition, older RDF semi-suspension incinerators were found to generate dry/semi-dry APC residues containing higher LOI values (11% w/w) compared to values ranging between 4.1 - 7.9 % (w/w) for similar residues from a modern RDF incinerator (Environment Canada, 1993). 11.7 CHEMICAL CHARACTERISTICS

Knowledge of the chemical characteristics, i.e. pH and acid neutralisation capacity and the chemical composition of the APC system residues and their dependency on various variables, is necessary for an understanding of the behaviour of these residues during handling, treatment, disposal and/or utilisation. In order to provide an overview of the chemical characteristics of the various types of APC system residues, a data base has been compiled using data from incinerators in Canada (6 facilities), Denmark (7 facilities), Germany (4 facilities), Jersey, Channel Islands (1), the Netherlands (6 facilities), Sweden (7 facilities) and the USA (8 facilities). In all, information has been collected on APC system residue chemical characteristics from 39 incineration facilities. Some background on the origin of the data is given in Tables 11.6, 11.7, 11.8 and 11.9.

455 Table 11.6 Origin of Data on Composition of Fly Ash (FA) from Mass Burning (BM) Incinerators Country Type of Incinerator Sampling Y e a r Reference Residue . . . . Point Sampled ..... Canada

FA FA

Denmark

FA FA FA FA

Germany

MB, Quebec City . MB, Quebec City

ESP ESP

1991 1986

MB, Vestforbraending MB, Amagerforbraending MB, Kolding II MB= Fasan

ESP ESP ESP ESP

1985+92 1985 1992 1992

FA FA FA FA

MB, Bamberg MB, G6ppingen MB, GSppingen MB, Oberhausen

ESP ESP ESP ESP

1982 1982 1984 1987

Schneideret al., 1983 Schneideret al., 1983 Schneider,1986 . Vehlow, 1988

Jersey, Channel Is.

FA

MB, Bellozane

ESP

1988

Hjelmar et al., 1993

The Netherlands

FA FA FA

MB, Amsterdam MB, Rotterdam MB, den Haa~l . . . .

ESP ESP ESP

1985186 Versluijset al., 1990 1985/86 Versluijset al., 1990 ! 985/86 Versluiiset al., 1990

Sweden

FA FA FA

MB, GRAAB, G6teborg ESP MB, Uppsala ESP . MB, Avesta.... ESP.

MB, Glen Cove, LI FA MB, Saugus, Mass. -FA ESP = Electrostatic Precipitator USA

ESP ESP

Eighmy,1992 ....Sawell &.Constable, 1988 Hjelmar,1987 & 1993 Hjelmar,1987 Hjelmar,1993 Hielmar, 1993 .

1988 1988 1988

SGI data base, 1993 SGI data base, 1993 SGI data base, 1993

1987 1989

LIRPB, 1993 Hjelmaret al., 1993

The data have been divided into groups representing ESP fly ash from mass burn incinerators (see box plots Table 11.6), dry and semi-dry APC system residues from mass burn incinerators (Table 11.7), wet scrubber APC system residues from mass burn incinerators (Table 11.8) and dry/semi-dry APC system residues from two-stage incinerators and RDF-fed incinerators (Table 11.9). Box plots depict all of the data that is available in the data base for each type of residue. The central box for the elements in each plot extends from the first quartile to the third quartile, with a horizontal line across the box to indicate the median value. The first quartile denotes the twenty-fifth percentile, the median denotes the fiftieth percentile and the third quartile denotes the seventy-fifth percentile. The height of the box equals the interquartile range. Lines are sometimes drawn out from the quartiles to adjacent values, defined as those data points less than 1.5 times the interquartile range beyond the first or third quartiles. Values farther than 1.5 times the interquartile range from the box are considered outliers and are denoted by individual circles in each of the plots. The width of each box in a plot is proportional to the number of observations it represents.

456 Table 11.7 Origin of Data on the Composition of Dry(DP) and Semi-dry (SDP) APC System Residues from Mass Burn (MB) Incinerators Country

R e s i d u e Incinerator

Sampling Year Point Sampled

Reference

Canada

DP + FA DP + FA DP + FA DP + FA SP + FA

MB, GVRD MB, GVRD MB, GVRD MB, Flakt/QUC MB, Flakt/QUC

FF FF FF FF FF

1991 1988 1991 1986 1986

Eighmy, 1992 Sawell et al., 1990 WASTE Program, 1993 Sawell et al., 1987 Sawell et al., 1987

Denmark

DP + FA DP + FA SP + FA SP + FA

MB, Nordforb. MB, REFA MB, Amagerforbraending MB, KARA

FF ESP FF ESP

1989 1989 1989/92 1989

Hjelmar, Hjelmar, Hjelmar, Hjelmar,

Germany

SP + FA

MB, DiJsseldorf

ESP

1982

Schneider et al., 1983

Netherlands

DP (-FA) MB, Incinerator 1

Sweden

DP + FA SP + FA SP + FA

MB, Ht)gdalen MB, Sysav MB, Linkt~ping MB, Karlstad

USA

DP + FA DP + FA SP + FA SP + FA SP+FA

MB, "Dry Scrubber 1" MB, "Dry Scrubber 2" MB, 3x750 tpd MB, "Dry Scrubber 3" MB, 500tpd

(-FA) + FA ESP

:

9 9

Without fly ash (precollected) Including the fly ash Electrostatic precipitator(s)

1992 1992 1992 & 1993 1992

van der Sloot, 1992 FF CY

FF FF FF CY

1985188 1988 1988 1988

SGI database, SGI database, SGI database, SGI database,

1988 1988/89 1989 1988/89

L I R P B ,1993 LIRPB, 1993 Kossonet al., 1993 LIRPB, 1993 Stuart, 1993

:

9

1993 1993 1993 1993

Fabric filter Cyclone system

Table 11.8 Origin of Data on the Composition of Wet Scrubber Residues from Mass Burning (MB) Incinerators Country

R e s i d u e Incinerator

The WP (-FA) MB, Incinerator2 Netherlands WP (-FA) MB, Incinerator 3 Sweden

Sampling Year Reference Point Sampled WWT WWT

1990/91 van der Sloot, 1992 van der Sloot, 1992

WP (-FA) MB, GRAAB, G0teborg WWT 1988 SGI database, 1993 WP (-FA) MB, GRAAB, GOteborg VVW3" 1989 Hjelmar, 1994 WP + FA MB, GRAAB, G0teborg WWT 1989 Hjelmar, 1994 WP (-FA) MB, Uppsala WWT 1988 SGI database, 1993 WP: Wet scrubber process residue + FA: Mixed with the precollected fly ash WWT: Wastewater treatment system (-FA): Not mixed with the precollected fly ash

457 Table 11.9 Origin of Data on the Composition of Dry APC System Residues from Two-Stage Mass Burn (MB) Incinerators and an RDF-fed Semi-Suspensi0n Incinerator Country Residue Incinerator

Sampling Point

Year Sampled

Reference

Canada

CT/FF CT/FF CT/FF

1992 1988 1988

PRRI, 1992 Sawell et al., 1990 Sawell et al., 1989b

USA CT/FF SS RDF

DP + FA DP + F DP + FA

MB, 2-stage, Inc. A MB, 2-stage, Inc. B MB, 2-stage, LVH

DP + FA SS, RDF, Inc. 1 CT/FF 1990 Sawell et al., 1991 Combined residue from conditioningtower and fabric filter Semi-suspension 9 Refuse-derived fuel

9

The use of box plots allows the reader to compare the data bases that are available for each type of residue. Several of the data sets have outliers that denote a skewness in the data. Such skewness means that the data is not normally distributed. As can be seen from Tables 11.8 and 11.9, data are available on fly ash and dry/semidry APC system residues from six or seven countries, covering both Europe and North America. In contrast, information on wet scrubber system residues is only available from Europe (Table 11.8) and information on APC system residues for two-stage and RDF-fed systems is only available from North America (Table 11.9). Only a few data sets are available on the characteristics of dry/semi-dry APC residues from systems with pre-collection of fly ash. Studies have shown that the content of most trace elements in the acid gas cleaning residues is very low and primarily originating from the lime or the process water if a very effective fabric filter precollection system for fly ash is employed upstream of the acid gas cleaning system (Vehlow, 1993). 11.7.1 pH and Acid Neutralisation Capacity The pH resulting from contact of an APC system residue with water is the most important factor controlling the solubility of various heavy metals and trace elements from these residues. Since the pH may change with time due to interaction with the surroundings (acid/base reactions and transport), the buffering capacity of a residue plays an important role in maintaining a certain pH level, or determines the rate of change of pH with time. Most APC system residues are highly alkaline (pH 11-12.5) and the buffering capacity is generally equal to the alkalinity. The pH of a residue is normally measured in a distilled water suspension at L/S ratios between 20 and 100. Based on a contact time of 0.5 hrs at 100:1 L/S ratio, pH ranges

458 of 7.0- 11.3 for ESP fly ashes, 12.1 - 12.5 for dry and semi-dry APC system residues, and 10.5 for wet APC system sludge mixed with fly ash have been recorded (Hjelmar, 1993). All these residues were from mass burn incinerators. Figure 11.6 shows pH ranges measured in leachates generated with water at L/S = 20 from ESP fly ash and dry/semi-dry APC system residues from mass burn incinerators, and dry/semi-dry APC system residues from a two-stage and a semi-suspension (RDF) incinerator (Sawell & Constable, 1988; Sawell et al., 1991; WASTE Program, 1993). All the dry/semi-dry APC residues are seen to be highly alkaline, whereas the ESP fly ash ranged from slightly acidic to moderately alkaline. Practically all dry/semi-dry APC system residues are highly alkaline due to the presence of excess lime. Normally, ESP fly ash also contains enough alkaline material (e.g. oxides and carbonates of calcium, sodium and potassium carried over from the boiler) to create an alkaline pH in equilibrium with water. The alkaline core of the ESP ash particles is, however, often coated by sorbed acidic condensation products. The relative amount of acidic condensation products to alkaline material depends on the conditions and temperatures in the ESP. In some cases an ESP ash generates an initially acidic pH when contacted with water, after which the pH gradually increases to an alkaline level (Hjelmar, 1987; Sawell & Constable, 1988). In other cases, the amount of alkaline material is insufficient to neutralise the acidic condensation products, resulting in a neutral or slightly acidic pH (as in Figure 11.6). Figure 11.6 Boxplots of pH at L/S = 20 for ESP Fly Ash and Various Dry APC System Residues

pH T

I,,Z

12-

I

8

tO-

8 -

A = 2-Stage, Boiler B = 2-Stage, Economiser

(Bridle & Sawell, 1986; Sawell et al., 1989; Sawell & Constable, 1990)

C = Mass Bum, Boiler D = Mass Bum, Economiser

(Waste Program, 1993) (Sawell & Constable, 1988)

E = RDF, Economiser

(Sawell et al., 1991)

6 A

B

C

[ "--T'-

I

I

!

, D

E

459 In column leaching tests with dry and semi-dry APC system residues, an initial depression of pH (9.7 - 10.8) in the first leachate has been observed, even under equilibrium conditions with high amounts of excess lime (Hjelmar, 1992). This was probably caused by the very unusual solubility conditions imposed by the high ionic strength ( >10 moles/litre or up to 500 g dissolved salts/litre). At higher L/S ratios and lower ionic strengths, the pH reached a level of 12.1 - 12.5. The alkalinity or the acid neutralisation capacity (ANC) of the residues may be measured either as a full titration curve (see Chapter 7), or as the amount of acid equivalents used to titrate the residue to a fixed endpoint, usually pH = 7 or 4.4. Some typical ANC titration curves for APC system residues are presented in Figure 11.7. Hjelmar (1992 and 1993) has reported ANC values (titration to pH = 7) for ESP fly ashes of 2.5 - 3.5 eqv/kg, for dry and semi-dry APC system residues of 5.6 - 12 eqv/kg, and for a wet APC system sludge mixed with fly ash of 4.5. This corresponds with the values shown in Figure 11.7. The ANC of the dry and semi-dry APC system residues reflect the amount of excess lime present which again reflects the operating conditions (i.e. the stoichiometric ratio of lime addition or whether or not the residue is recycled). The influence of recycling on the ANC of dry scrubber residues from the scrubber reactor and the fabric filters (the bulk of the material) is illustrated in Figure 11.8. The data indicate that the concentration of acid gas reaction products increases with increased exposure to the flue gas stream, thereby reducing the ANC. Figure 11.7 ANC Titration Curves for an ESP Fly Ash (FA) and Two Semi-Dry APC System Residues (SD) from Fabric Filters 14

11.2

FA from ESP

8.4 -r-

(3.

SD(1) from FF

5.6

o

SD(2) from FF 2.8

.

5

Acid Added Stuart, 1993; Hjelmar, 1994

(eqvlkg

10

dry residue)

15

460 Figure 11.8 Comparison of ANC Values for Dry APC System Residues with and without Lime Recycle 21 18

. . . . . . . . . . . . . . Fabric Filter (No R e c y c l e ) o...

9

,~

9 . "

~

e-

o

~ Z
91% capture) at temperatures below 150 - 160~ At temperatures over 200~ mercury removal efficiency was negligible, if not nonexistent. However, the use of activated carbon or sodium sulphide has been demonstrated as effective control reagents for mercury (Guest and Knizek, 1991). Sodium sulphide injection results in effective formation of stable HgS. Consequently, mercury (as well as sulphate) concentrations are increased in the residues. Conversely, it has been speculated that mercury sorbed onto activated carbon is not as stable, since it is susceptible to reduction by carbon. Although this has not been confirmed, further study into the phenomenon should be conducted to ensure that there is no substantial release of elemental mercury into the atmosphere through reduction, or methylation reactions.

Figure 11.12 Trace Element Concentrations in Fly Ash (FA), Semi-DrylDry APC Products with Fly Ash (SPIDP) and Wet Scrubber Products without Fly Ash (WP-FA)

Antinony, c d

.

Cnroniun, c r

Y

a

1000

Vanadium, V

iOOD

400 200

S i l v e r . ~g 120 1

ioo

C O M l t . CO

1

472 T a b l e 11.16 T r a c e Elements ( pH=p~,the surface would adsorb cations. Table 13.15 depicts some PH=pcfor various adsorbent mineral phases.

561 Sorption phenomena are usually modelled using empirical, experimentally determined sorption constants based on solute activity or using mechanistic explanations of the electrostatic interactions that occur at the particle surface. Both approaches can be modelled in the modelling programs described in Chapter 15. Table 13.15 Estimates of pHD=e for Various Minerals Mineral

pHDzc

y-AI203

8.5

Anatase (TiO2)

5.8

Birnessite (5-MnO2)

2.2

Calcite (CaCO3)

9.5

Corundum (a-AI203)

9.1

Goethite (a-FeOOH)

7.3

Hematite (a-Fe203)

8.5

Magnetite (a-Fe304)

6.6

Rutile (TiO2)

5.8

Quartz (a-SiO2) 2.9 Adapted from Davis and Kent, 1990 with permission from the Mineralogical Society of America

13.5.2 Activity-Based Sorption Models The first activity-based sorption model is based on the distribution coefficient, Kd (Allison et al., 1990). Using the convention of M as a metal and SOH as a hydroxylated surface sorption site, consider the following reaction depicting m sorbing to the surface site: SOH + M -

SOH" M

(13.75)

The ratio of sorbed metal concentration to the total analytical metal concentration in solution at equilibrium is the distribution coefficient. It is analogous to an equilibrium constant: Kd = [SOH" M] [M]

(13.76)

562 It is more appropriate to examine the Kd relationship based on the activity of participating aqueous metal solute: K da c t _ -

By convention [SOH

{SOH" M}

(13.77)

{i}

9 M] = {SOH ~ M} and equation (13.77) becomes: act_

Kd

-

[SOH" M] u [M]

(13.78)

Many Kd"= values are tabulated in the sorption review document prepared by Rai and Zachara (1984). The second approach that employs activity in an empirical way to depict sorption is the Langmuir adsorption model (Allison et al., 1990). Again, using the SOH and M terminology, consider the following reaction: SOH + M ,= S O H - M

(13.79)

where at equilibrium K Lact

_ -

{SOH

9M }

(13.80)

{M} {SOH}

Conventionally, the Langmuir constant is derived experimentally using various quantities of SOH and M. To place ~"= into the more familiar context of the Langmuir adsorption isotherm, a mass balance on surface sites is needed (Allison et al., 1990): [SOH]T = [SOH

9 M] + [SOH]

(13.81)

Combining equations (13.80) and (13.81) produces:

[SOH

9 M] =

K act

L [SOH]T Ya[M] act

1 + "L I~"

Yi[ M]

(13.82)

The only difference between Kda~ and ~a~t is that the Langmuir equation assumes a finite concentration of SOH. Values for ~"~ can be found in Rai and Zacchara (1984).

563 The third activity-based empirical sorption model is the Freundlich model (Allison et al., 1990). Again using the SOH and M terminology, consider the following reaction: SOH + l/nM = SOH. M

(13.83)

where at equilibrium Kfact =

{SOH 9M} {M} TM {SOH}

(13.84)

Imposing the convention that {SOH ~ M} = [SOH ~ M], equation (13.84) becomes: [SOH

9M] = Kfact {MM+}TM

(13.85)

The 1In term is a mass action stoichiometric coefficient related to M. Kf"ct is similar to Kd~ if n=l. Kfact differs from ~act in its implicit assumptions about an unlimited supply of unreacted surface sites at equilibrium. Kf"~tvalues can be found in Rai and Zacchara (1984). The fourth empirical, activity-based sorption model is the ion exchange model (Allison et al, 1990). Again using the SOH terminology, but denoting the exchangeable metal as M and the sorbing metal as M2, consider the following SOH- M 1 - M 1 § M 2 ,, S O H - M 2

(13.86)

where at equilibrium, Kex =

{M1}{SOH'M 2} {M2}{SOH'M 1}

(13.87)

K.x differs from the other three constants by assuming a substitution reaction occurs at SOH. K,x values are found in Rai and Zacchara (1984). There are no conventions as to the applicability of these four empirical models (Allison et al., 1990). When modelling sorption processes, all four models can be tested (provided appropriate constants are available). The literature does suggest that mechanistic models based on electrostatic considerations do a better job at modelling sorption (Allison et al., 1990); frequently parameters for these models are not available (as discussed below).

564

13.5.3 Electrostatic Surface Complexation Models There have been a number of attempts to model adsorption to mineral surfaces while taking into account electrostatic interactions between charged surfaces and solutes. The models explain how both cations and anions adsorb as a function of pH, adsorbent site density and ionic strength (Westall and Hohl, 1980). Models such as the constance capacitance model (CCM) and the diffuse double layer model (DLM) were developed (Davis and Kent, 1990). A third model, the triple layer model (TLM), was developed to have multiple adsorption planes on the mineral surface to allow for outer sphere as well as inner sphere complexes to form (Leckie, 1988). The TLM is viewed as most applicable and is presented here. The TLM is schematically depicted in Figure 13.18 (Leckie, 1988). As described by Evans (1989), tightly bound inner sphere complexes reside in the inner or surface plane, ~. (or cx plane). Outer sphere complexes reside in the adjacent plane, )~ (or 13 plane). Noncomplexed species reside in the diffuse layer, A,~(or d plane). To maintain electroneutrality, the charge density, o, must equal the intrinsic charge density of the mineral such that:

(13.88)

Oint + Ois + Oos + 0 d = 0

Figure 13.18 Depiction of the Triple Layer Model % ~r=

o'p

I I I

Od

I I I

NO~

§

I

-r" ?~ 0 0

(I)

~:" ~:;!

:Z~

O~b*.... I

OebOH I

O

I I

I

I

....

I

....

,Np

....

,"9;

....

l~b 2.

~L~! a

....

Pt~OH§

, |

a~

NO~

I~a§ N

NO 3

pb 2+

I

i"L;: Ol'

~'#

Na* NO~

NO~ I

~--~-.~ O H +

~

NO~

Na* NO~ NO 3

N% pb 2+

NO3

CI_ ,, J

13

immobile layer

d diffuse

layer

Reprinted with permission from Leckie, 1988. Copyright Lewis Publishers, an imprint of CRC Press, Boca Raton, Florida. 9

565

Hayes (1987) has depicted the types of binding mechanisms that can occur in both the inner and outer spheres (Figure 13.19). The figure shows various inner sphere and outer sphere complexation reactions that can occur. Figure 13.19 Schematic Depiction of Coordinative Surface Complexes and Ion Pairs at Oxide Surfaces Metal

Oxygen

Other Examples

C~c,-b ~

",v..,,,>Water '~"" Molecules

C~a~1b

t3

I-. ~-. No;. ClO,-

Na* K + 2+ 2. , , Ca , Mg

uter-Sphere Complexes

c~,L co;k

"O"-~>__O_

pb.

Cu

f Divalent Transition Metal Ions

Monodentate

Divalent Transition Metal Ions

Bidentate / Inner-Sphere

"O~~

U~ap~exe s

F

,,O-'- H P

OH

Mononudear

SeO~;,~02

Binuclear

From Hayes, 1987 with permission of the author

566 Consider the following reaction for the monovalent metal ion M § (Allison et al., 1990): SOH

+

+

H s + M s ,- ( S O . M )

-

(13.89)

where SOH and Ms+ are the same surface binding site and sorbing metal and Hs§ is the sorbed proton that must deprotonate from SOH to allow for formation of the sorbed complex SO 9M. By convention,

and

{H~} = {H *} [e-~~

(13.90)

{ms} = {m *} [e -YI~FIRT]

(13.91)

+

where e "~F/RT is the Boltzmann factor for either the a or 13 planes depicted in Figure 13.19. Equations (13.86) through (13.88) can be written as an equilibrium expression (Allison et al., 1990): K = {SO

9M} {H *} [e-~~

(13.92)

{SOH} {M *} [e-~I3F/RT]

Other forms of equation (13.89) can be provided for a hydrolysis reaction of the type (Allison et al., 1990): SOH + M 2+ § H20 - 2H~ - S O . MOH

(13.93)

producing, K = {SO

9MOH} {H +}2 [e-I~~

2

{SOH} {M 2+} {H20 } [e-~I]F/RT]2

(13.94)

A similar approach can be taken for a sorbing monovalent anion (Allison et al., 1990): +

SOH + A s- + H s ,, SOH 2 A 9

(13.95)

producing, {SOH 2 A} 9 [e-~IBF/RT] K

._.

{SOH} {A-} {H *} [e-I~~

(13.96)

Finally, a similar approach can be taken for a sorbing divalent anion (Allison et al., 1990):

567 SOH + A 2- + H s =* SOH 2 A 9 t

(13.97)

producing K

=

{SOH 2 A-} 9 [e-e"F/RT]2 {SOH} {A 2-} {H +} [e-~oF/Rm]

(13.98)

The geochemical model MINTEQA2 has the capability to model sorption using estimated parameters for the TLM as well as the CCM and DLM. The TLM model has been widely used to model adsorption of cations and anions (Davis and Kent, 1990). For adsorption to oxides, Davis and Leckie (1978, 1980), Balistrieri and Murray (1982), Hsi and Langmuir (1985), Catts and Langmuir (1986), LaFlamme and Murray (1987), Zachara et al. (1987), Hunter et al. (1988), and Zachara et al. (1987) have successfully used the model. It has also been applied to non-hydrous oxide minerals like quartz, titanium and clay (Shuman, 1986). 13.5.4 Adsorption Data Davis and Kent (1990) have compiled some interesting data as to how cations and anions adsorb to mineral surfaces. As shown in Figure 13.20, there is a narrow pH range where a cation or anion goes from near zero adsorption to high levels of adsorption. This adsorption edge occurs at pHads, the pH where significant sorption occurs, which is close to the Prize:. Cations adsorb at high pH by forming inner sphere complexes with the deprotonated hydroxyl functional groups. Anions adsorb at low pH when forming inner sphere complexes with protonated functional groups. 13.6 A UNIFIED APPROACH TO LEACHING The information compiled in this chapter allows us to assemble an approach to characterising the leaching process. Understanding fundamental leaching behaviour of a residue such as ash requires the consideration of many factors (Figure 13.21). The speciation of the elements in the solid phase plays a fundamental role in controlling the nature of the leachate. This can then be related to that fraction of an element that is available for leaching. Particle morphology, porosity, and diffusion pathlength are also critical in assessing the role of diffusion in controlling reactions. Attempts to characterise those leaching processes that are kinetically based or thermodynamically based is an additional approach that is needed. The thermodynamic mechanisms can be modelled with geochemical codes. Additional data can be gleaned by conducting leaching studies to assess the effects

568 of pH, Eh and ligands on dissolution phenomena. The processes surrounding sorption cannot be ignored; these studies can be compared to the various sorption models that are presently available. Finally, the role of the L/S ratio and time must be considered. Figure 13.20 Adsorption Edges for Various Cations and Anions

From Parks, 1990 with permission of the Mineralogical Society of America Now that these concepts have been explained, they need to be applied. The next chapter sets out various types of leaching tests which can be performed to determine certain characteristics of ash under specific sets of leaching conditions.

569 Figure 13.21 Schematic of Fundamental Leaching Behaviour

Available

Fraction

Particle

Chemical Speciation

Therm odynamics

Morphology

FUNDAMENTAL LEACH IN G BEHAVIOUR

Kinetics

Influence of LS, Time

Influence of pH, pE & Ligands Sorption

570 REFERENCES

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572 Fruchter, J.S., D. Rai and J.M. Zachara. Identification of Solubility-Controlling Solid Phases in a Large Fly Ash Field Lysimeter. Environ. Sci. Technol. 24, pp. 1173-1179, 1990. Furuichi, R., N. Sato and G. Okamoto. Reactivity of Hydrous Ferric Oxide Containing Metallic Cations..Chimia 23, pp. 455-463, 1969. Garrels, R.M. and C.L. Christ. Solutions, Minerals, and Equilibria. Harper and Row, New York, 1965. Giovanoli, R. J.L. Schnoor, L. Sigg, W. Stumm and J. Zobrist. Chemical Weathering of Crystalline Rocks in the Catchment Area of Acidic Ticino Lakes, Switzerland. Clays Clay Min 36, pp. 521-529, 1989. Grandstaff, D.E. Some Kinetics of Bronzite Orthopyroxene Dissolution. Geochim. Cosmochim. Acta 41, pp. 1097-1103, 1977. Guggenheim, E.A. Applications of Statistical Mechanics, Clarendon Press, New York, 1966. Harvie, C.E., N. Mr and J.H. Weare. "The Prediction of Mineral Solubilities in Natural Waters: The Na-K-Mg-Ca-H-CI-SO4-OH-HCO3-CO2-H20 System to High Ionic Strengths at 25~ ''. Geochim. Cosmochim. Acta 48, pp. 735-751, 1984. Harvie, C.E. and J.H. Weare. The Prediction of Mineral Solubilities in Natural Waters: the Na-K-Mg-Ca-CI-SO4-H20 System from Zero to High Concentration at 25~ Geo.chim. Cosmochim. Acta 44, p. 981-987, 1980. Hayes, K.F. Equilibria, Spectroscopic and Kinetic Studies of Ion Adsorption at the Oxide Aqueous Interface. Ph.D. Dissertation, Stanford University, Palo Alto, CA, 1987. Helgeson, H.C. Thermodynamics of Hydrothermal Systems at Elevated Temperatures and Pressures. Am. J. Sci. 267, pp. 729-804, 1969. Hem, J.D. Equilibrium Chemistry of Iron in Groundwater. In Principles and Applications of Water Chemistry Edited by S.D. Faust and J.V. Hunter. John Wiley and Sons, New York, p. 625, 1967. Hering, J.G. and W. Stumm. Oxidative and Reductive Dissolution of Minerals. In Mineral-Water Interface Geochemistry Edited by M.F. Hochella, Jr. and A.F. White. Mineralogical Society of America, Washington, D.C., p. 427, 1990. Hsi, C.D. and D. Langmuir. Adsorption of Uranyl onto Ferric Oxyhydroxides: Application of the Surface Complexation Binding Site Models. Geochim. Cosmochim. Acta 49, pp. 1931-1941, 1985.

573 Kim, H.-T. and W.J. Frederick Jr. Evaluation of Pitzer Ion Internaction Parameters of Aqueous Electrolytes at 25~ Single Salt Parameters. J. Chem. EnQ. Data 33, pp. 177-184, 1988. Kluge, G., H. Saalfeld and W. Dannecker. Untersuchun,qen des LanQzeitverhaltens von M011verbrennun.qsschlacken beim Einsatz im Strassenbau. Unweltforschungsplan des Bundesministers des Innern, Forschungsbericht Nr. 103 03 006. Berlin, 1980. Krupka, K.M., R.L. Erikson, S.V. Mattigod, J.A. Schramke and C.E. Cowan. Thermochemical Data Used by the FASTCHEM Packa.qe. EPRI EA-5872, EPRI, Palo Alto, CA, 1988. Kummert, R. and W. Stumm. The Surface Complexation of Organic Acids on Hydrous u J. Colloid Interface Sci. 75, pp. 373-385, 1980. LaFlamme, B.D. and J.W. Murray. Solid/Solution Interaction: The Effect of Carbonate Alkalinity on Adsorbed Thorium. Geochim. Cosmochim. Acta 51, pp. 243-250, 1987. Leckie, J.O. Coordination Chemistry at the Solid/Solution Interface. In Metal Speciation: Theory, Analysis and Application Edited by J.R. Kramer and H.E. Allen. Lewis Publishing, Chelsea, MI, 41, 1988. Lindsay, W.J. Chemical Equilibria in Soils. J. Wiley and Sons, New York, 1979. Millero, F.J. and D.R. Schreiber. Use of Ion Pairing Model to Estimate Activity Coefficients of the Ionic Components of Natural Waters. Am. J. Sci. 282, pp. 15081540, 1982. Naumov, G.B., B.N. Ryzhenko and Khodakovsky. Handbook of Thermodynamic Data. US Geological Survey WRD-74-001. NTIS-PB-226 722/AS, Washington, D.C., 1974. Nordstrom, D.K. and J.L. Munoz. Geochemical Thermodynamics. Blackwell Scientific Publications, Palo Alto, CA, 1986. Nriagu, J.O. Lead Orthophosphates. IV. Formation and Stability in the Environment. Geochim. Cosmochim. Acta 38, pp. 887-898, 1974. Oberste-Padtberg, R. and K. Schweden. Zur Freisetzung von Wasserstoff aus M6rteln mit MVA-Reststoffen. Wasser Luft Boden 34, pp. 61-62, 1990. Pankow, J.F. Aquatic Chemistry Concepts. Lewis Publishers, Chelsea, MI,, 1991. Pankow, J.F. and J.J Morgan. Kinetics for the Aquatic Environment. I. Environ. Sci. Technol 15, pp. 1155-1164, 1981 a.

574 Pankow, J.F. and J.J. Morgan. Kinetics for the Aquatic Environment. II. Environ. Sci. Technol. 15, pp. 1306-1313, 1981b. Parker, V.B., D.D. Wagman and W.H. Evans. Selected Values of Chemical Thermodynamic Properties. Tables for the Alkaline Earth Elements (Elements 92 through 97 in the Standard Order of Arrangement). U.S. Nationa.I Bureau of Standards Technical Note 270-6, Gaithersburg, MD, 1971. Parks, G.A. Surface Energy and Adsorption at Mineral/Water Interfaces: an Introduction. In Mineral-Water Interface Geochemistry Edited by M.F. Hochella Jr. and A.F. White. Mineralogical Society of American, Washington, D.C., p. 133, 1990. Pitzer, K.S. Thermodynamics of Electrolytes. I. Theoretical Basis and General Equations. J. Phy..s. Chem. 77, pp. 268-277, 1973. Pitzer, K.S. Theory Ion Interaction Approach. In Activity Coefficients in Electrolyte Solutions. Edited by R. Pytkowicz. CRC Press, Boca Raton, FL, p. 157, 1979. Pitzer, K.S. Characteristics of Very Concentrated Aqueous Solutions. In Chemistry and Geochemistry of Solutions at Hi~..h Tempe.ratures and pressures. Edited by D.T. Rickard and F.E. Wickman. Pergamon, Oxford, p. 249, 1981. Pitzer, K.S. and L. Brewer. Thermodyn.amics, 2nd Edition, McGraw-Hill, New York, 1961. Pitzer, K.S. and J.J. Kim. Thermodynamics of Electrolytes. IV. Activity and Osmotic Coefficients for Mixed Electrolytes. J. Am. Chem. Soc_ 96, pp. 5701-5707, 1974. Pitzer, K.S. and G. Mayorga. Thermodynamics of Electrolytes. I1. Activity and Osmotic Coefficients for Strong Electrolytes with One or Both Ions Univalent. J.. Phys. Chem.. 77, pp. 2300-2308, 1973. Pourbaix, M. Atlas of Electrochemical Equilibria. Pergammon Press, Oxford, 1966. Pytkowicz, R.M. Equilibria, Nonequilibria, and Natural Waters, Vol. 1 and 2, John Wiley and Sons, New York, 1983. Rai, D. Inor.qanic and Organic Constituents in .Fossil Fuel Combustion Residues. Volume I A Critical Review. EPRI EA-5176, EPRI, Palo Alto, CA, 1987. Rai, D. and J.M. Zachara. Chemical Attenuation Rates, Coefficients and Constants in Leachate Mi,qration, Volume 1" A Critical Review.. EPRI EA-3356, EPRI, Palo Alto, CA, 1984.

575 Robie, R.A., B.S. Hemingway and J.R. Fisher. Thermodynamic Properties of Minerals and Related Substances at 298.1K and 1' Bar Pressure and at Hi.qher Temperatures. Geological Survey Bulletin No. 1452, U.S. Government Printing Office, Washington, D.C., 1978. Robie, R.A., B.S. Hemingway and J.R. Fisher. Thermodynamic Properties of Minerals and Related Substances at 298.15~ and 1 bar Pressure and at Higher Temperatures. Geolo.(]ical Survey Bulletin 1452. U.S. Government Printing Office, Washington, D.C., 1979. Rossotti, F. The Determination of Stability Constants.. McGraw-Hill Co., Inc., New York, 1981. Roy, W.R. and R.A. Griffin. Illinois Basin Coal Fly Ashes. 2. Equilibria Relationships and Qualitative Modelling of Ash-Water Reactions. Environ. Sci. Technol. 18, pp. 739742, 1984. Rubin, J. Transport of Reacting Solutes in Porous Media Relations Between Mathematical Nature of Problem Formulation and Chemical Nature of Reactions. Water Resources Res. 19, pp. 1231-1252, 1983. Scatchard, G. The Excess Free Energy and Related Properties of Solutions Containing Electrolytes. J. Am. Chem. Soc 90, pp. 3124-3127, 1968. Schindler, P.W. and W. Stumm. The Surface Chemistry of Oxides, Hydroxides and Oxide Minerals. In Aquatic Surface Chemistry. Edited by W. Stumm. John Wiley and Sons, New York, p. 83, 1987. Schnoor, J.L. Kinetics of Chemical Weathering: A Comparison of Laboratory and Field Weathering Rates. In Aquatic Chemical Kinetics" Reaction Rates of Processes in Natural Waters Edited by W. Stumm. John Wiley and Sons, New York, p. 475, 1990. Schnoor, J.L. and W. Stumm. The Role of Chemical Weathering in the Neutralization of Acidic Deposition. Schweiz Z. HYdrol. 48, pp. 171-193, 1986. Schott, J. and J.-C. Petit. New Evidence for the Mechanisms of Dissolution of Silicate Minerals. In Aquatic Surface Chemistry Edited by W. Stumm. John Wiley and Sons, New York, p. 293, 1987. Schott, J., R.A. Berner and E.L. Sj0berg. Mechanism of Pyroxene and Amphibole Weathering. I. Experimental Studies of Iron-Free Minerals. Geochmi. Cosmoch.im. Acta 45, pp. 2123-2135, 1981.

576 Schumm, R.H., D.D. Wagman, S.M. Bailey, W.H. Evans and V.B. Parker. Selected Values of Chemical Thermodynamic Properties. Tables for the Lanthanide (Rare Earth) Elements (Elements 62 through 76) in the Standard Order of Arrangement. U.S. National Bureau of Standards Technical Note 270-7. U.S. Government Printing Office, Washington, D.C., 1973. Schwertmann, U. and R.M. Taylor. Iron oxides. In Minerals in Soil Environments Edited by J.B. Dixon and S.B. Weed. Soil Science Society of America, Madison, WI, p. 145, 1977. Sigg, L. Surface Chemical Aspects of the Distribution and Fate of Metal Ions in Lakes. In Aquatic Surface Chemistry Edited by W. Stumm. John Wiley and Sons, New York, p. 319, 1987. Sigg, L. and W. Stumm. The Interactions of Anions and Weak Acids with the Hydrous Goethite (a-FeOOH) surface. Colloids Surf. 2, pp. 101-117, 1981. Sill~n, L.G. and A.E. Martell. Stability Constants of Metal-Ion Complexes. Supplement No. 1, Chemical Society, London, 1971. Smith, R.M. and A.E. Martell..Critical Stability Constants. Plenum Press, New York, 1976. Sposito, G. The Thermodynamics of Soil Solutions, Clarendon Press, Oxford, 1981. Sposito, G. Chemical Models of Inorganic Pollutants in Soils. CRC Crit. Rev. Environ. Control 15, pp. 1-24, 1984. Sposito, G..The Chemistry of Soils, Oxford University Press, N.Y., 1989. St~mpfli, D. Final Report: Cement and Bottom Ash Chemistry (CABAC). ERG Report. UNH, Durham, NH, 1992. St~mpfli, D., H. Belevi, R. Fontanive, and P. Baccini. Reactions of Bottom Ashes from .MSW Incinerators and Construction.Waste Samples with Water. EAWAG Project 3335, EAWAG, Dubendorf, Switzerland, 1990. Stumm, W. and G. Furrer. The Dissolution of Oxides and Aluminum Silicates; Examples of Surface Coordination-Controlled Kinetics. In Aquatic Surface Chemistry, Edited by W. Stumm. John Wiley and Sons, New York, p. 197, 1987. Stumm, W. and J.J. Morgan. Aqu.atic Chemistry.. John Wiley and Sons, New York, 1981.

577 Stumm, W. and E. Weiland. Dissolution of Oxide and Silicate Minerals: Rates Depend on Surface Speciation. In Aquatic Chemical KineticsLReaction Rates of Processes in. Natural Waters Edited by W. Stumm. John Wiley and Sons, New York, p. 367, 1990. van der Wegen, G. Orienterend Onderzoe.k. Naar O.orzaak Binding in een Monster AVISlakken. Rapportnumme, 91149, Intron, the Netherlands, 1991. Verink, E.D. Simplified Procedure for Constructing Pourbaix Diagrams. J. Educational Modules Materials Sci. EnQineer 1, pp. 535-560, 1979. Wagman, D.D., W.H. Evans, V.B. Parker, I. Halow, S.M. Bailey and R.H. Schumm. Selected Values of Chemical Thermodynamic Properties. Tables for the First ThirtyFour Elements in the Standard Order of Arrangement. _U.S. National Bureau of Standards Technical Note 270-3 U.S. Government Printing Office, Washington, D.C., 1968. Wagman, D.D., W.H. Evans, V.B. Parker, I. Halow, S.M. Bailey and R.H. Schumm Selected Values of Chemical Thermodynamic Properties. Tables for Elements 35 through 53 in the Standard Order of Arrangement. U.._S.National Bureau of Standards. Technical Note 270-4 U.S. Government Printing Office, Washington, D.C., 1969. Wagman, D.D., W.H. Evans, V.B. Parker, I. Halow, S.M. Bailey, R.Ho Schumm and K.L. Chumey. Selected Values of Chemical Thermodynamic Properties. Tables for Elements 54 through 61 in the Standard Order of Arrangement. U~S. National Bureau of. Standards Technical Note 270-5 U.S. Government Printing Office, Washington, D.C., 1971. Wagman, D.D., W.H. Evans, V.B. Parker, R.H. Schumm and R.L. Nuttall. Selected Values of Chemical Thermodynamic Properties. Compounds of Uranium, Protactinium, Thorium, Actinium, and the Alkali metals. U..S. N_ati0nal Bureau of Standards Technical. Note 270-8 U.S. Government Printing Office, Washington, D.C., 1981. Wagman, D.D., W.H. Evans, V.B. Parker, R.H. Schumm, I. Halow, S.M. Bailey, K.L. Churney and R.L. Nuttall. The NBS tables of Chemical Thermodynamic Properties. Selected Values for Inorganic and C1 and C2 Organic Substances in SI units. J. Phy. Chem. Ref. Data 11 (Supplement No. 2), pp. 1-392, 1982. Warren, C.J. and M.J. Dudas. Weathering Processes in Relation to Leachate Properties of Alkaline Fly Ash. J. Environ. Qual. 13, pp. 530-538, 1984. Warren, C.J. and M.J. Dudas. Formation of Secondary Minerals in Artificially Weathered Fly Ash. J. EnYiron. Qua!. 14, pp. 405-410, 1985. Warren, C.J. and M.J. Dudas. Mobilization and At..tenuati0n of Trac.e Elements in Artificially Weathered Fly .Ash. EPRI EA-4747, EPRI, Palo Alto, CA, 1986.

578 Wieland, E., B. Wehrli and W. Stumm. The Coordination Chemistry of Weathering: III a Generatlization on the Dissolution Rates of Minerals. Geochem. Cosmochim. Acta 52:1969-1981 Westall, J.C. and H. Hohl. A Comparison of Electrostatic Models for the Oxide/Solution Interface. Adv. Coll. Inter. Sci. 12, pp. 265-294, 1980. Whitfield, M. An Improved Specific Interaction Model for Seawater at 25~ Atmosphere Total Pressure. _Marine Chem. 3, pp. 197-213, 1975a.

and 1

Whitefield, M. The Extension of Chemical Models for Sea Water to Include Trace Components at 25~ and 1 Atmosphere Pressure. Geochim. Cosmochim. Acta 39, pp. 1545-1557, 1975b. Wollast, R. and L. Chou. Kinetic Study of the Dissolution of Albite with Continuous Flow-Through Fluidized Bed Reactor. In T.he Chemis..try of Weatherin.Q, Edited by J.l. Drever, NATO AI Series C 149, pp. 75-96, 1985. Zachara, J.M., D.C. Girvin, R.C. Schmidt and C.T. Resch. Chromate Adsorption on Amorphous Iron Oxyhydroxide in the Presence of Major Groundwater Ions. Environ. Sci. Technol. 21, pp. 589-594, 1987. Zachara, J.M. and G.P. Streile. Use of Batch and Cplumn..Methodolo.qies to Assess Utility Waste Leachin.q and Subsurface Chemical Attenuation. EPRI EN-7313, EPRI, Palo Alto, CA, 1991. Zevenbergen, C., J.P. Bradley, T. van der Wood, R.S. Brown, L.P. van Reeuwijk, and R.D. Schuiling. Weathering as a Process to Control the Release of Toxic Constituents from MSW Bottom Ash. In Geoconfine 93 Edited by M. Arnoud, M. Borres and B. C6me. A.A. Balkema, Rotterdam, p. 591, 1993.

579

CHAPTER 14- LEACHING TESTS 14.0 LEACHING TESTS Now that the concepts behind leaching phenomena have been introduced, discussing leaching tests is appropriate. Invariably, these tests are involved in the regulation of residues, as well as in the interpretation of leaching phenomena. Careful consideration should be given to the specific "tools" that are selected to characterise ash. Clearly, it is preferable to use a number of tools, rather than a single tool, for work in both science and regulation. At the end of this section, a unified theory of leaching is presented. This will move away from the strict use of concentration data and toward normalisation of leaching data to release, fractions leached and fluxes from residues. The rationale for using this approach to develop models and management scenarios will be discussed. Generalised and detailed reviews of leaching tests are found in Jackson et al. (1984), F,~llman (1990), Environment Canada (1990), Zachara and Streile (1991) and van der Sloot et al. (1991; 1993). The information from Environment Canada (1990) provides much of the basis for the following discussion. 14.1 PURPOSE OF LEACHING TESTS In general, a leaching test involves contacting a solid material with a leachant to determine which components in the solid will dissolve in the leachant and create a leaching solution or leachate. To investigate the various processes governing the extent and rate of leaching, endless variations can be introduced by changing test variables, such as leachant composition, method of contact, liquid-to-solid (L/S) ratio, contact time and system control (pH, pE (or Eh), temperature). Leaching tests have a wide range of objectives, the most common of which are presented in Table 14.1. Leaching tests are typically used to provide information about the constituent concentration or the constituent release from a waste material under reference test conditions, or under conditions that more closely approximate the actual disposal site. This information may subsequently be used in mathematical models to predict long term leaching.

14.1.1 Classification of Leaching Tests For the purposes of this discussion, leaching tests have been separated into two broad categories on the basis of whether or not the leachant is renewed: 1) extraction tests (no leachant renewal), and 2) dynamic tests (leachant renewal).

580 Table 14.1 .Leaching Test Objectives Objective

Description

Identification of leachable constituents

Determine which constituents of a waste are subject to dissolution upon contact with a liquid.

Classification of hazardous wastes

Compare wastes against performance criteria for classification of wastes as hazardous or nonhazardous.

Evaluation of process modifications

Determine if modifications to a wastegenerating process result in a less leachable waste.

Comparison of waste treatment methods

Determine whether a given waste treatment method/process results in superior containment of contaminants.

Quality control in waste treatment

Verify the efficiency of a treatment process using a simple pass/fail criterion.

Design of leachate treatment systems

Obtain a typical leachate to perform treatability experiments.

Field concentration estimates

Express leaching over time (e.g. to be used as a source term in groundwater modelling).

Parameter quantification for modelling

Quantify partition coefficients and kinetic parameters to be used in transport modelling.

Risk assessment

Estimate potential impact of waste disposal on the environment.

The concept of leachant renewal is based on modifying the leaching system to promote solution control of leaching rather than solid phase control.

Extraction Tests Extraction tests include all tests in which a specific quantity of leachant is contacted with a specific quantity of waste for a certain length of time, without leachant renewal. (This definition does not include analytical extractions or digestion procedures which are used to measure the total constituent concentration in an ash sample). The leachate is separated from the solid and analysed either at various times during the

581 test, or, as in most extraction tests, at the end of the test. The analysis of leachates generated at various times can help determine the kinetics of the leaching process or if equilibrium has been attained. The underlying assumption in this type of test is that an equilibrium condition is achieved by the end of the extraction test (i.e. the concentrations of solutes in the leachate become constant). In this no-flow system, an equilibrium condition occurs when there is no net transfer of components from the solid phase to the leaching solution, or vice versa. Sampling in an extraction test over time to derive kinetic information or to monitor the attainment of equilibrium is difficult since it must be done without modifying the residueleachant interactions, which are a function of factors such as the L/S ratio and gaseous exchanges. This can be accomplished in three ways: nondestructive sampling and analysis of parameters such as pH, conductivity or specific ions removing small volumes (aliquots) that are negligible when compared with the total volume preparing as many parallel extraction tests as data points required and performing destructive analyses. Extraction tests can be further divided into four subcategories: agitated extraction tests non-agitated extraction tests sequential chemical extraction tests concentration buildup tests

Agitated Extraction Tests

Agitated extraction tests (Figure 14.1) are performed to reach steady-state conditions as quickly as possible. They measure the chemical properties of a waste-leachant system, as opposed to rate-limiting mass transfer mechanisms. Agitation ensures a homogeneous mixture, promotes contact between the solid and the leachant and reduces boundary layer thicknesses. Sample particle size reduction is often performed to increase the surface area to volume ratio of the solid to enhance liquid/solid phase contact and to eliminate mass-transfer limitations. Generally, this reduces the duration of the test by reducing the time required to reach a pseudo-equilibrium condition in the leachate. This procedure may also have the effect of overestimating the short-term release of constituents. A steady-state leaching environment can also be attained in a column apparatus by recirculating the collected leachate back into the column.

582 Figure 14.1 Agitation Extraction Test Crushed Solid Waste

Monolithic Solid Waste

o00 (:]DO ~

C]

OOO E]

Agitation

After Environment Canada, 1990

Non-agitated Extraction Tests A non-agitated extraction test is performed to study the physical mechanisms that are rate-limiting in leaching. The underlying assumption behind a non-agitated extraction test is that the physical integrity of the solid matrix and mass transfer constraints (both internally within the sample and externally in the boundary layer) affect the amount of contaminants that are leached during the test. Two types of non-agitated tests are illustrated in Figure 14.2. They can be performed on large particle-sized residue samples, concrete-type or monolithic samples. The disadvantage of running a non-agitated test is that a much longer contact period may be required to reach equilibrium conditions than is required in an agitated test. The advantage of this type of test is that rate-limiting mechanisms of leaching due to the physical integrity of the solid matrix are taken into account. These tests are presented in further detail in Chapter 20.

583 Figure 14.2 Static Leach Test

A) Static test with monolithic solid waste

A) Statictest with nonmonoUthic solid waste


Sequential Chemical Extraction Tests A sequential chemical extraction test is composed of a battery of non-agitated extraction tests (Figure 14.3). It involves performing sequential elutions of aliquots of a sample with different leachants (i.e. A, B, C, D and E in Figure 14.3), which are increasingly more aggressive in terms of chemical attack toward the residue (Figure 14.3a). One type of method assumes that each successive leachant also extracts the sum of contaminants extracted by all preceding leachants. The other type of method is conducted by subjecting the same aliquot of sample to each leachant (Figure 14.3b). The amount extracted in each elution is associated with a certain chemical form or mineral phase in the solid phase. The Sequential Chemical Extraction Procedure, originally compiled by Tessier et al. (1979), was adapted to sewage sludge incinerator ash by Fraser and Lure (1983), and then further modified for MSW incinerator ash (WTC, 1990). The test has been used in different studies (Wadge and Hutton, 1987) (Tessier Method); Environment Canada, 1993 (modified)), however, results by Khebohian and Bauer (1987) and discussion by Nirel and Morel (1990) (on the Tessier method) have shown that resorption and reprecipitation reactions can dramatically alter the mass fractions that are obtained in the different extractions. This limitation has been recognised and the latest studies using the modified method have basedmuch of their interpretation on the operationally defined extractions (e.g., peroxide extractable) rather than the implied chemical species (Environment Canada, 1993). Consequently, it appears that although the method is not appropriate for determining the chemical species, relating the operationally defined extractions to exposures under different leaching conditions (e.g., fraction available for leaching under acidic leaching conditions versus severe reducing conditions) is an appropriate set of interpretations.

584 Figure 14.3 Sequential Chemical Extraction Tests

Leachant a)

A

C

D

E

C

D

E

Wffh different waste samples

Leachant A b)

B

B

With the same waste sample and liquid/solid separation between elutions

After Environment Canada, 1990 Concentration Buildup Tests

In a concentration buildup test, an extraction is achieved at a very low cumulative L/S ratio. Aliquots of samples are successively contacted with the same leachant (Figure 14.4). The contact of leachate with fresh solid material can be considered as a model for an elemental volume of water flowing through a large body of residue and approaching saturation with respect to specific mineral phases. The purpose of this test is not to collect kinetic information, but to characterise a leachate saturated with soluble residue constituents. In some cases, this may simulate the actual pore water composition of a granular material in column leach tests or in outdoor disposal or utilisation scenarios. Dynamic Tests

Dynamic tests include all tests in which the leachant is continuously or intermittently renewed to maintain a driving force for leaching that is solution-controlled. The intermittent tests may be conducted by alternating leaching periods with dry periods to study the effects of desiccation or unsaturated flow conditions. Dynamic tests provide information about the kinetics of solid phase dissolution and contaminant flux. Information is generated as a function of time, and attempts are often made to preserve the residue's physical integrity. These two factors lend this category of leaching tests to the investigation of more complex mechanisms of leaching.

585

Dynamic tests can be further divided into subcategories according to how the interface between the waste and the leachant is defined. Tests in which individual waste particles are used to define the interface are called serial batch tests. The tests in which a characteristic dimension of the waste, (such as the external geometric surface area or the geometric surface area perpendicular to flow) is used to define the interface include flow-around tests and flow-through or column tests. Figure 14.4 Concentration Buildup Tests

1

2

N Discard

II

Agitation


Serial Batch Tests A serial batch test is conducted using a granular or crushed sample which is mixed with leachant at a given US ratio for a specified period of time (Figure 14.5). The leachate is then separated from the solids and replaced with fresh leachant until the desired number of leaching periods have been completed. The waste/leachant mixture is normally agitated to promote contact. Kinetic information regarding contaminant dissolution is obtained using the concentrations measured in the leachate from each of the leaching periods. Data from serial batch tests can be used to construct an extraction profile to infer the temporal release of leachable constituents. Flow-Around Tests In flow-around tests, a sample of residue is placed in the leaching vessel and the flow of fresh leachant around the residue provides the driving force to maintain leaching. The L/S ratio is modified to express the volume of leachant divided by the surface area

586 of the solid sample. Samples are usually monolithic, although non-monolithic or crushed residue may be used if it is confined in some manner. Agitation is generally not performed. Leachant flow is either continuous (Figure 14.6a), in which case it is sampled and analysed periodically, or it is intermittently renewed (Figure 14.6b). The latter method is generally simpler from an experimental point of view, but the renewal frequency must be sufficient to prevent a buildup of contaminants at the residue/leachant interface, which may inhibit further leaching by reducing the diffusional gradient. Figure 14.5 Serial Batch Tests

m

m

m

AgitaUon After Environment Canada, 1990

Flow-Through Tests In a flow-through or column test, an open container is packed with a porous solid and leachant is passed through, either continuously or intermittently. The effluent is sampled periodically and analysed for the parameters of interest. The results are used to examine contaminant removal in which the primary transport mechanism is advection. There are two basic types of flow-through tests characterised primarily by the shape and size of the container. The first type is a column test which is performed using a small cylindrical container (Figure 14.7a). The second is a lysimeter test which is conducted in a large rectangular or cylindrical container (Figure 14.7b). In general, the size of the sample used in a flow-through test tends to be large to minimise the effects of sample heterogeneity and wall channelling effects. The depth of waste in either type of test varies according to the individual experiment.

587 Figure 14.6 Flow-Around Tests leachant

@

leachant

a) continuous leachate renewal

9

9

9

b) intermittent leachate renewal

After Environment Canada, 1990 Figure 14.7 Flow-Through Tests leachant (downflow)

y_,

Y

leachant

leachant (upflow)

a) Columns After Environment Canada, 1990

I leachate b) Lysimeter

588 Columns may be operated either in an upflow or downflow mode, whereas lysimeters are always operated in a downflow mode. Flow through the solid depends upon its hydraulic conductivity, as well as the hydraulic gradient, and varies with the individual test. Mini-columns may be used to achieve a relatively rapid breakthrough of leached species. Since head losses may be large and a rapid breakthrough is desired, the leachant is usually delivered under pressure and at a constant flow rate. The advantages of minicolumns include: L/S ratios that are similar to those of real waste-leachant systems a known and easily varied average fluid velocity negligible axial dispersion or spreading of the solute a simple estimation of both equilibrium and kinetic coefficients automation permitting the rapid output of data. These tests are not applicable when large volumes of leachate are needed for a variety of analytical tests. Care should also be taken when conducting flow-through tests to avoid unnatural channelling of water and clogging by fine material or biological growth. In lysimeter tests, channelling cannot be avoided. It is a factor that occurs in the field, and its influence should be modelled in the laboratory, although quantifying it is difficult. Biodegradation of organics can also be a problem in columns, although in some cases experiments are intentionally set up to measure the effects of biological activity. Flowthrough tests can also be modified to examine other site-specific influences, such as vegetation on the surface of the container, or layered media, such as ash and geological material.

14.1.2 Leaching Test Variables

There are several experimental variables which are common to all extraction and dynamic tests. These variables need to be considered when designing a leaching test for specific purposes.

Sample Preparation

Depending on the nature of the waste and the test to be performed, the sample may require one of the following preparatory steps: 9 9 9 9 9 9 9 9

liquid/solid separation sub-sampling particle-size reduction surface washing compaction preservation curing aging

589 Liquid/solid separation may be performed on residues containing a free liquid phase. The leaching test is conducted only on the solid portion of the sample. The free liquid phase constitutes the initial leachate, which may be analysed separately to estimate the pore-water concentration or it can be included with the final leachate for analysis. Liquid/solid separation can be accomplished by various methods, including settling and decanting, centrifugation or pressure filtration through filter media of various types. Sub-sampling is generally required when several different tests or replicates are to be performed on the same sample. Waste samples should be thoroughly mixed before sub-sampling is performed. (See Chapter 6). Particle-size reduction is required for most extraction tests. The goal is to reduce the time required to reach steady-state conditions by increasing the contact surface area between the solid and the leachant. However, care should be taken to prevent the loss of volatile compounds in the solid if they are of interest. Particle-size reduction is usually carried out by grinding (e.g. mortar and pestle, centrifugal grinder or hammer mill). These issues are also discussed in more detail in Chapter 6 and 7. Surface washing may be performed prior to testing small monolithic samples in flowaround tests. The surface is washed to remove small detachable particles and readily soluble salts by quickly dipping the sample in an aqueous solution. Compaction or remolding is often required for flow-through tests. Reproducibility and field simulation considerations require that samples be compacted to a pre-specified density using methods such as vibration, proctor compaction or modified proctor compaction. Sample preservation is performed to avoid biological activity. This is a greater problem in tests of long duration, such as column tests. Various chemical treatments are available, such as the addition of sodium azide, however, none offer complete efficacy. Curing may be performed on samples that have been transformed into a solidified mass using various chemical additives, such as Portland cement. It allows the waste sample to gain physical and engineering properties, i.e. high unconfined compressive strength and low permeability, that are considered to be important in reducing leachability. Curing can be used to achieve a variety of chemical reactions within the waste, although this term usually refers to cement hydration reactions. Aging may be promoted on any type of waste sample to account for the physical, chemical and biological alterations that a waste might undergo in the field. Chapter 13 discusses classes of aging reactions that can occur in residues.

Leachant Composition

The release of contaminants from a waste in any leaching test may be strongly influenced by the initial leachant composition, especially at high L/S ratios, or with the

590 use of an aggressive solution. Chemical properties of the leachant that influence contaminant mobilisation are indicated in Table 14.2. Examples of three types of commonly used leachants, i.e. water, site liquid and chemical solution, are identified in Table 14.3. Several advantages and disadvantages of these leachants are outlined in Table 14.4. Table 14.2 Important Factors in Leachant Composition Factor Release Mechanism Affected Dissolution/precipitation of metals, speciation of inorganic species Adsorption/desorption of solutes Oxidation/reduction of inorganic species Ionic exchange of metals, speciation chemistry and solubility products

pH Eh, redox potential Ionic strength Chelating and complexing agents Buffering capacity After Environment Canada, 1990

Metal solubility All above properties

Table 14.3 Commonly Used Leachants Type of Leachant Water

Common Uses

Nonaggressive, baseline medium without buffering capacity Site liquid (real or Simulates site-specific synthetic) leaching conditions Chemical solution Examines metal speciation and organic compound binding After Environment Canada, 1990

Examples Distilled, deionised and tap water Rainwater, groundwater, surface water, landfill leachant, seawater Strong chemical solution (acidic, basic, reducing, oxidising, complexing, solvent,etc.)

591 Table 14.4 Advantages.and D isadvantacjes of Commonly Used Leachants Leachant

Advantages

Disadvantages

Pure water

Reliable, simple standard

Lack of background composition may result in dissolution of common ions

Site liquid

Best field case model

Requires characterisation (to obtain leaching results by subtraction)

Several synthetic liquids available

Results not comparable with other leaching studies Labour intensive (sampling and preservation)

Chemical solution Allows for the study of waste chemistry After Environment Canada, 1990

Aggressive, difficult to relate data to field conditions

Method of Contact Since a leaching test is primarily a system to study the transfer of contaminants from a residue to a liquid, it is important to consider the aspects of the test conditions that promote mass transfer, such as agitation, and to consider the effect of mass exchange with other components of the system, primarily the leaching vessel and the atmosphere.

Agitation of the leachant-solid slurry generally hastens reaching equilibrium conditions by maintaining maximum contaminant concentration gradients at the leachant-solid particle interface. Different methods can be used to agitate the waste, including shaking (wrist action or reciprocation) stirring (magnetic or paddle) tumbling gas bubbling. In static or non-agitated tests, the leachant-solid interface is usually the geometrical surface area of the solid form. There is usually no provision for mixing because diffusion of leached constituents within the leachate is assumed to be much faster than the rate of release by mechanisms such as dissolution from the surface or diffusion from within ash particles. Ensuring that the leachate is well mixed before sampling is important, however.

592 It may be important to identify and quantify exchanges of chemical species other than between the solid and the leachant. Exchanges between the leachant and the leaching vessel are always undesirable, whereas exchanges with the atmosphere depend in large part upon the objectives of the test, such as leaching with carbonic acid. To minimise exchanges with the leaching vessel, glass or stainless steel should be used for organic contaminants and plastic for inorganic contaminants. If the cost is not prohibitive, polytetrafluoroethylene is considered to be acceptable for both. For the purpose of verifying the mass of constituents adsorbed to the container wall, the emptied leaching vessel can be extracted with a strong solvent. The test system may be either open or closed to the atmosphere. The choice depends on the specific leaching problem being examined. For example, a closed system provides a better simulation of the saturated groundwater environment, whereas an open system models problems like a storage pile and unsaturated disposal environments more accurately. An open system facilitates sampling, leachant renewal and periodic or continuous adjustment of the pH or redox potential. However, a system that is open to the atmosphere allows for the loss of volatile compounds, including water and organics, and the introduction of CO2 and 02 from the air. Losses due to evaporation may have to be accounted for in an open system. Although volatile organics are generally not a concern with incinerator ash, there are several apparatus configurations that will prevent volatile contaminants from escaping. If there is no headspace in the leaching vessel, volatiles will remain in either the solid phase or the leachate. If there is a headspace, volatiles will be partitioned in the gas phase. Analysis of the headspace allows for an evaluation of this loss. Even for experiments carried out in closed containers and under controlled conditions, penetration of gases through plastic container walls can have a significant effect, especially over long durations. This is seen in reaction vessels kept at a low redox potential when oxygen diffuses into the vessel.

Liquid-to-Solid Ratio

The L/S ratio is the ratio of the amount of leachant in contact with the residue to the amount of waste being leached. Although this definition appears straightforward at first glance, it can become confusing because of the many ways in which the two variables in the ratio have been defined. The L/S ratio has been expressed as: volume of leachant/mass of solid mass of leachant/mass of solid and volume of leachant/surface area of solid (for monolithic material).

593 Furthermore, when using the first two expressions, the mass of solid being leached can be calculated on a wet weight or a dry weight basis. Another problem arises because of the various ways that the volume or mass of leachant can be calculated, depending on whether or not the liquid phase of the solid is included in the total leachant volume. The preferred way to report L/S is the mass of leachant to the dry mass of solid. Figure 14.8 illustrates how these various ways of defining the amounts of waste and leachant can give different L/S ratios for the same system. The three fractions shown in Figure 14.8 include the amount of leachant added, the liquid phase associated with the solid, and the solid phase. Figure 14.8 Liquid and Solid Fractions of the Waste Leachant System ,,,

A

Added Leachant

B Liquid Phase of Waste

C

Solid Phase of Waste

After Environment Canada, 1990 If the residue is dry, then the L/S ratio is simply A/C. If the residue is wet, there are three ways to define the L/S ratio: 1) the residue is the sum of the liquid and solid phases, i.e. L/S = N(B+C). 2) the leachant is the sum of the amount of leachant added plus the liquid phase associated with the solid, i.e. L/S = (A+B)/C 3) the liquid phase associated with the solid is excluded from the calculations, and residue is the solid phase only, i.e. L/S = A/C.

594 Leachate concentrations of highly soluble species (e.g. sodium, potassium) are generally inversely proportional to the L/S ratio of all of the species which have been removed from the solid. However, if the release of a species is limited by solubility, the final concentration is independent of the L/S ratio and simply equals the maximum solubility concentration. In general, the leachate concentration will be controlled by a number of competing factors, namely, the amount of contaminant available, solubility and kinetically controlled chemical reactions. Thus, the relationship between the L/S ratio and concentration is complex, and different for each species of interest. Selection of an appropriate US ratio depends on the objectives of the leaching test, the solubility of species of interest and analytical constraints. The ratio should be low enough to avoid dilution of contaminants to less than analytical detection limits. However, the ratio also must be high enough to prevent solubility constraints from limiting the amount of contaminants that can be leached from the waste. The selected ratio should be somewhere between these two limitations. Practical values for the L/S ratio range from 0.1 to 100:1. To place these values into perspective, most landfilled residues are exposed to L/S values of less than 3:1 during the operational life (10 to 20 years) of a disposal facility. After closure, the type and integrity of the cap will influence further increases in the L/S.

Contact Time

The total amount of time that a leachant is in contact with a solid sample before the attainment of equilibrium will influence the amount of contaminant released. In extraction tests, the contact time is equivalent to the duration of the test, whereas in dynamic tests, it is a function of the flow rate, or the number of elutions, in addition to the test duration. The contact time for extraction tests should allow equilibrium conditions to be reached for the contaminants of interest. This is generally in the order of hours to days for samples that have undergone particle-size reduction. For concrete-based or monolithic samples, it can be in the order of weeks to months. The contact time for dynamic tests should be sufficient to allow for observation of the processes of interest. Diffusion processes may be quantified within a few weeks, although several months may be required to study slow chemical reactions.

Temperature

Temperature affects the results of extraction and dynamic tests. Both the van't Hoff relationship, which applies to thermodynamic equilibrium constants and solubility products, and the Arrhenius relationship, which applies to kinetic processes such as adsorption and diffusion, indicate that properties or mechanisms relevant to leaching vary exponentially with temperature.

595 For convenience, most leaching tests are performed at room temperature. Higher temperatures may be used to accelerate the rate of leaching (although this may also change the properties of the waste) or to simulate the effects of biological activity in a landfill or the self-heating from exothermic reactions.

Leachate Separation

Leachates are commonly separated from agitated non-monolithic wastes by filtration using a 0.45 pm membrane filter (a convention used to define soluble species). However, very small colloid particles can pass through a 0.45 pm filter. A smaller pore size filter (0.2 pm) should be used if these particles are to be removed. The use of the smaller filter size should be reported with the data. Glass fibre filters are chosen when hydrophobic, low solubility organic molecules are expected in the leachate since they may have a high affinity for filters composed of an organic polymer. Membrane filters, such as cellulose acetate, should be used for metal species in place of glass. The same care used to select a leaching vessel should be applied when selecting the filter material. Filtering the leachate from non-agitated monolithic samples may not be necessary if the method of contact generates only dissolved species. This should be verified before sampling.

14.1.3 Compilation of Leaching Tests The leaching tests presented in Environment Canada (1990), Fallman (1990), and van der Sloot et al. (1991, 1993) serve as the basis for the compilation of various leaching tests presented here. The reader can refer to these references for precise details of each method. Table 14.5 summarises the various agitated batch extraction tests that are used for regulatory purposes or for research into leaching characteristics of waste. All the methods specify the type of leaching vessel to be used, the type of sample preparation that is required, the amount of sample that is needed, the type of leachant to be employed, the L/S ratio that is used, the type of agitation that is required and the duration of the test. Most methods also specify the type of filtration that is to be employed to allow for quantification of total dissolved constituents in the leachate. Table 14.6 specifies two non-agitated extraction tests that are commonly used to examine sequential dissolution of mineral phases in a solid or the fundamental dissolution and effective diffusion parameters of a solid dissolving under static conditions.

Table 14.5 Agitation Leach Tests Test Name and Proponent

Status of Leaching Development Vessel

Sample Preparation

Sample Mass

Leachant

US Ratio

-EP f o x U.S. EPA Method 1310

Standard regulatory method (1980)

Unspecified

Non-monolithic waste; phase separation Monolithic waste; particle-size reduction

100 g

Deionised water 0.5 N acetic acid (max. 2.0 meq H+lg solid)

20:l

-LEP MOE (Ontario)


Wide mouth, 1250 mL cylindrical bottle

Phase separation by 0.45 pm membrane filter

50 g of d?, sol~ds

Distilled water Acetic acid (2.0 meq H+lg dry solids)

20.1

End over end (10 rPm)

24 hours

0.45 pm filtration

-TCLP U.S. EPA Method 1311


Any material compatible with waste, zero head-space extractor (ZHE) for volatiles

Cuttinglcrushing and grinding Solidlliquid phase separation No structural integrity

100 g (25 g for ZHE)

Buffered acetic acid 1) pH = 4.93 2) pH = 2.88

20:l

End over end (30 rpm)

18 hours

0.6 to 0.8 pm borosilicate glass fibre filter combines liquid phase with extract

-Q.R.S.Q. MOE (Quebec)


>1 L bottle

No phase separation 100 g dry Inorganics: buffered Grinding solids acetic acid (0.82 No structural integrity 50 g for meq H+lg dry volatiles solids) Organics: distilled water

10:l

End over end (10 to 20 rPm)

24 hours

30 min decantation, 0.45 pm filtration

-WET California


Polyethylene or glass container

Milling, 0.45 pm filtration

50 g

0.2 M sodium citrate at pH 5.0

10:l

Table shaker Rotary Extractor

48 hours

Centrifugation 0.45 pm filtration

-X31-210 French Leach Test AFNOR (France)

Proposed standard for waste (Dec 1992)

Straight wall, 2 L bottle

Remove free liquid Reduce particle size to Vss ~o q " utidt

1

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=

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for

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column leaching curve from the minicolumn experiment for c o m p o n e n t [i], the volume of ash in the column, and the flow rate.

632 15.2.2 Modelling Solid Phase Dissolution Theis et al. (1992) used their approach to model minicolumn leaching experiments for a number of MSW incineration residues. Results for a soluble constituent (potassium) from an ESP ash are discussed further. The data are shown in Figure 15.10. Potassium in the ESP ash is readily soluble, it initially appears in high concentration in the minicolumn leachate and declines precipitously over time without any reprecipitation. The column was operated over 30,000 pore volumes during the one hundred-hour run. The model simulated the observed data quite well. Figure 15.10 Potassium Leaching and Modelling from Minicolumn Experiments

o

POTASSIUM CONC. VS TIME ak c = De

1.27E-3 t

1.18E-3

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After Theis et al., 1992 with permission of the Air and Waste Management Association 15.2.3 Modelling Solid Phase Reprecipitation and Solubility Control The data for lead are shown in Figure 15.11. Lead in the ESP ash is less soluble than potassium. It appears in an initially high concentration in the minicolumn leachate and declines over time. The model simulates soluble lead in the leachate quite accurately.

633 However, in the case of lead, at two different positions in time and space, two solids are predicted to precipitate out and control lead dissolution. The first solid to appear is anglesite (PbSO4) early in the column run. The second, lead hydroxide [Pb(OH2)], controls later in the run. In support of these observations, earlier pC-pH plots of lead data from the column showed aqueous phase lead to be in equilibrium first with anglesite and then with lead hydroxide. Figure 15.11 Lead Leaching and Modeling from Minicolumn Experiments

LEAD CONC. VS Ti M E

-3.5

-4~ -4.5~"

m
11. Several test methods used for regulatory purposes are consistent with the general pH dependent leaching curve indicating the solubility control for Pb is very significant. Zinc (Figure 16.3h) The leaching behaviour of Zn as a function of pH is more consistent than any of the other elements, indicating of a high degree of solubility control. In the pH range below 6, the leachability reflects the amount of Zn available for leaching with the maximum leachability reached between pH 4 and 5. As in the case of Pb, the Zn leaching is characterised by a sharp increase in leachability at high pH due to formation of anionic hydroxide complexes.

16.3.2 APC Residues The release of constituents from APC residues as a function of pH is crucial for the understanding of the chemistry behind the observed release from APC residues. Figure 16.4 presents the solubility of B, Ba, Cd, Cr, Cu, Hg, Mo, Pb and Zn from APC residues as a function of pH. The data were obtained from (Eighmy, 1993; van der Sloot et al, 1992; Kosson et al., 1993; Environment Canada, 1993, Hjelmar, 1993). Characteristic solubility controlled release is indicated although insufficient data was available to develop unified leaching curves.

Boron The leachability of boron from filter ash (without scrubber residue) is consistent with that of bottom ash (Eighmy et el, 1993). At pH greater than 10 the solubility decreases.

654

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Barium

The leachability of Ba does not show any significant dependence on pH except a slight decrease around pH = 12.

Cadmium

The solubility of Cd from APC residues follows a similar pattern to Cd in bottom ash, except that the solubility slope is shifted to higher pH caused by complexation with the high chloride content in these residues.

Chromium

Cr solubility measured with the ANC procedure (Kosson et al., 1993) shows a consistent pattern which would reflect Cr III leachability at pH < 4 and a chromate leachability plateau from pH 4 to 10. At pH greater than 10, the concentration of chromate also decreases.

Copper

The leachability data for Cu show a large amount of scatter. The pH stat data for ESP ash are fairly consistent indicating very low solubility of Cu corresponding to tenorite solubility. The irregular behaviour of Cu in APC residues is not well explained.

Mercury

The leachability of Hg indicates a pattern similar to many other metals but only limited data are available.

Molybdenum

The leachability of Mo is scattered. No distinctive trend in its behaviour has been observed.

Lead

The pH dependent leaching behaviour of Pb is consistent with only a few outliers. It is striking to note that the leachability of Pb at pH around 12 exceeds the leachability of Pb at pH = 4. This can result in significant release because the pH of many APC residues is greater than 11. Pb is considered one of the elements of concern in APC residues because results from regulatory testing exceed many current regulatory limits.

Zinc

The leachability of Zn is quite consistent considering the widely different origin of the

658 residues included in the figure. This may point at the presence of one important solubility controlling phase. 16.4 GEOCHEMICAL MODELLING OF LEACHING EQUILIBRIA In Chapter 15, several studies on modelling of leaching behaviour from incinerator residues have been discussed (Di Pietro et al., 1987; van der Sloot et al., 1987; Eighmy et al., 1993, Kirby and Rimstedt, 1993; Comans et al., 1993). Two complementary approaches have been used: Identifying and quantifying the solid phases in ash and running MINTEQA2 through a solid phase approach (Eighmy et al., 1993); and, Using aquatic chemistry to identify the solubility controlling phases using MINTEQA2 based on the measured leachate composition (Comans et al., 1993) Different sophisticated analytical techniques (see Chapter 7.2.7) are needed for the first approach to carry out species identification and quantification. Eighmy at al (1993) indicated fairly good agreement between test data and modelling results following this approach (See Table 15.5 and 15.6). The leaching test used for the comparison was carried out at controlled pH=4 using a high L/S value. The modelling results for AI and Si are significantly below the measured concentration levels because according to the model quartz (SiO2), diaspore (AIO(OH)) and alunite (KAI3 (SO4)2(OH)6) precipitated out. The second approach presented covers a wider range of pH conditions. Elemental concentrations have been predicted by assuming equilibrium between the leachates (at L/S=5) and potential solubility-controlling minerals in bottom ash. These minerals were selected on the basis of their saturation indexes in prior MINTEQA2 runs and their likeliness to be present or formed under the experimental conditions (Chapter 9). The thermodynamic data from the standard MINTEQA2 (version 3.11, Allison et al., 1991 ) database were used, unless indicated otherwise. As molybdenum was not present in the database, it was added as the component MoO42 together with equilibrium constants for aqueous species and solids reviewed by Rai and Zachara (1984). The model predictions are presented as total element concentrations, rather than free ion activities, in the leachate solutions at each pH. This enables presentation of model results together with the analytical leaching data in a graph of log-concentration as a function of pH, which maintains the characteristic shape and concentration levels of the unified pH leaching curves.

659 16.4.1 Bottom Ash Figure 16.5 presents the total dissolved concentrations of major and trace elements in leachates as a function of pH, as well as MINTEQA2 predictions assuming equilibrium with different mineral phases. In the following paragraphs, leaching curves and possible solubility controlling processes are discussed for each element. Calcium and Sulphate Ca and SO4 are the major dissolved components leached from the bottom ash samples, generally reaching concentrations of 3 g/L or greater. Since the solubility of Ca in fossil fuel combustion residues may be controlled by calcite (CaCO3), portlandite (Ca(OH)2) or gypsum (CaSO4*2H20) (Rai, 1987), the same mineral phases have been considered for MSW incinerator bottom ash. MINTEQA2 calculations indicate that the leachates are not in equilibrium with atmospheric CO2 (350 ppm) and that measured concentrations of Ca and CO3 are not in equilibrium with calcite. The steepness of the portlandite solubility curve indicates that the leachates are not in equilibrium with this mineral either. The Ca and SO4 concentrations between pH 4 and 10 do not depend very strongly on pH and follow the solubility curve for gypsum. This relationship is confirmed by plotting the data in a log Ca2*-activity versus log SO42"-activity diagram. Gypsum is a soluble mineral and does allow high concentrations of both Ca and SO4 to be leached from bottom ash. However, provided enough time, carbonation of the alkaline ash and the formation of calcite will further limit Ca leaching in the near-neutral pH range. Calcium concentrations at pH greater than 10 start decreasing at a lower pH and less sharply than predicted by portlandite solubility. The mineral phase controlling Ca solubility at strongly alkaline pH remains as yet unknown. Ettringite, which has been observed in field applications of compacted bottom ash, may play a role in the solubility of Ca at high pH. Stability data for ettringite is needed to be able to verify this possibility. Magnesium Mg concentrations are essentially independent of pH between pH 4 and 7. Normally magnesium also is largely controlled by carbonate minerals. The system is not in equilibrium with atmospheric CO2 as indicated for Ca already. At higher pH values, concentrations decrease sharply. MINTEQA2 calculations indicate this decrease to be consistent with the solubility line of brucite (Mg(OH)2), a mineral which has often been shown to control magnesium solubility. The data point at pH > 13 deviates due to the very high ionic strength of this particular sample. Silicon Dissolved silicon decreases between pH 4 and 10 and increases again at strongly alkaline pH. This leaching pattern is not in agreement with either amorphous or crystalline (quartz) SiO2, as calculated with MINTEQA2. Fruchter et al. (1990) have

660 Figure 16.5 Dissolved Metals in Bottom Ash Leachates as a Function of pH Compared with MINTEQA2 Predictions Assuming Equilibrium with Different Mineral Phases 10000

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9

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2

4

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661 Figure 16.5 Continued

10-

1000 t

100

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|

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(g) "~'

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0,001

,

t

t

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t

t

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662 suggested wairakite (CaAI2Si4012*2H20) as the solubility controlling mineral for Si in coal fly ash leachates. Si data from the bottom ash leachates between pH 4 and 10 agree remarkably well with the solubility pattern of wairakite and strongly suggest this mineral phase controls silicon leaching. The data at higher pH deviate from wairakite solubility and suggest that another mineral controls dissolved Si under strongly alkaline conditions. The agreement between measured and calculated data appears to be better than the solid phase approach for this matrix component.

Aluminum AI leaching from bottom ash is strongly pH-dependent and results in the characteristic V-shaped Iog-concentration/pH curve which is typical for Al-(hydr)oxide solubility. Model calculations indicate that the leaching data closely follow the solubility curves of gibbsite (crystalline AI(OH)3) and amorphous AI(OH)3. The data below pH 7 are more similar to those calculated in equilibrium with amorphous AI(OH)3, whereas gibbsite seems to be the solubility controlling mineral at higher pH. Similar observations have been reported for coal fly ash (Fruchter et al., 1990). The formation of aluminosilicates may control the solubility of AI in the high pH ranges. Ettringite has been observed in field application of compacted bottom ash (INTRON, 1991 )(see also Chapter 9) and may also play a role in AI solubility control at high pH. Kirby and Rimstedt (1993) also suggest amorphous aluminum hydroxide as the most likely solubility controlling phase. Iron Fe is leached at relatively high concentrations at low pH, decreases strongly toward neutral pH values and remains essentially pH-independent at neutral to alkaline pH. Amorphous iron hydroxide, or ferrihydrite (Fe(OH)3), is the most obvious solubility controlling mineral for dissolved iron. The measured data at acid to neutral pH follow the calculated solubility line for ferrihydrite, but are up to two orders of magnitude higher in concentration. Similar observations have been attributed to the presence of colloidal iron or inaccuracy of published solubility data for ferrihydrite (Fruchter et al., 1990). The pH-independence at neutral to alkaline pH is not consistent with the calculated ferrihydrite solubility. A second mineral may be controlling Fe leaching in this pH range. Manganese Dissolved Mn also decreases strongly with pH, and reaches minimum values at approximately pH 10. A slight increase in concentration occurs at higher pH values. Although the solutions are not in equilibrium with pyrocroite (Mn(OH)2), the shape of the curves may suggest another less soluble Mn-(hydr)oxide to control Mn leaching. Recent work suggests manganite (MnO(OH)) as the solubility controlling phase. In view of the difference in concentration at the different LS values studied, Mn is probably not solubility controlled at pH below 8. Mn(hyrdr)oxide appears to control the solubility at pH >12.

663

Sodium, Potassium, Bromide and Chloride

The alkali metals Na and K, as well as Br and CI, are very soluble and are leached in high concentrations from the bottom ash in a pH-independent manner. No solubility controlling solids were found for these elements.

Cadmium

Cd is leached in high concentrations at low pH and decreases strongly toward pH values of 8-9. At higher pH, Cd concentrations are near the ICP detection limit of 1 IJg/L shows the stability lines for ottavite (CdCO3) and amorphous-Cd(OH)2. Ottavite stability was calculated using independent measurement of total carbonate (as was mentioned above, the leachates were not in equilibrium with atmospheric CO2). Cadmium concentrations in the leachates between pH 6 and 9-10 are close to values predicted on the basis of equilibrium with ottavite. Cadmium is also known to have a very high affinity for the surface of calcium carbonate (Comans and Middelburg, 1987). Although there is no evidence from solubility calculations that dissolved Ca and CO3 were in equilibrium with calcite, some calcite might nevertheless have been formed considering the high Ca concentrations, pH, and contact with the atmosphere. Co-precipitation or solid-solution formation with calcite may limit Cd concentrations to lower levels than would be predicted by the solubility of ottavite (Comans and Middelburg, 1990). It is uncertain at present what role this sorption process plays in controlling Cd solubility during bottom ash leaching and whether it may explain some of the cadmium concentrations below the solubility of ottavite. All leachates were at least two orders of magnitude under saturated with respect to amorphous Cd(OH) 2, which rules out the relevance of this mineral in systems containing sufficient carbonate. The role of carbonation and calcite formation during aging of bottom ash appears to be a significant factor for the retention of Cd in the ash matrix. Previous removal of chlorides, thus avoiding the formation of soluble Cd-complexes, will enhance this process. Accelerated carbonation may prove a means to improve ash characteristics.

Copper

Leaching of fresh bottom ash generally leads to high initial concentrations of copper (Figure 16.6). Cu concentrations decrease from pH 4 to 6 but remain constant at higher pH at levels (approx. 1 mg/L), depending on the L/S ratio. Tenorite (CuO) has been suggested as the controlling phase for Cu leaching from coal fly ash (Fruchter et al., 1990) and may also be relevant for bottom ash. However, the leachates are systematically over saturated with respect to this mineral, except at low pH where concentrations are lower than predicted by tenorite solubility. Substantial amounts of dissolved organic matter can often be released from the uncombusted fraction in the bottom ash. It is speculated that dissolved organic material may increase copper solubility because of the high affinity of this metal for organic species. The relevance of this process was investigated by giving an ash sample an 8-hour heat treatment at 550 ~ prior to the leaching experiments. This

F~gure16 6 Leachab~lltyof Copper from Bottom Ash

(b)

(a)

Demineralized water

-

-

B A - ~ & z ~, ~ 5% H, flushing q,

\

3

2

4

5

8

7

6

9

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1 0 1 1 1 1 1 3 1 4

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.

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r

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1

(dl

a

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s

l

9

1 0 1 1 1 2 1 3 1 4

pH

\ \\

0

01:

001

/ \\

j

CU

2

3

'

0

\

d

4

BA heated 1 0 5 5 0 S--, ~ ~ "

4

L

L

5

-

Prcah BA

Reducing

1 cond~tlons

0

'

6

L

7

-

8

pH

d-

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j

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U

abc-

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10

d

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3-

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L

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//

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.

9 1 0 1 1 1 2 1 3 1 4

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cornp~ex.tion

01

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,

3

,,

4

/

/

\\ \

/ >

5

6

7

8

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9

1 0 1 1 1 2 1 3 1 4

/ 8

pH

Unified pH curve Redox potential in batch experiments under alr exposure and in a close container purged with 5% hydrogen Cu leaching from fresh bottom ash, same BA after addition of a mixture of organic acids, same BA after heating to 550°C to remove organlc matter, and Cu leachlng data point for a rotary kiln facility proposed factors and their pH dependence for leaching pf Cu from bottom ash under ambient conditions

665 procedure was assumed to destroy organo-copper complexes by removing any organic material left in the ash sample. As Figure 16.6c indicates, the leaching of Cu is reduced by the temperature treatment, especially between pH 6 and 11. At low pH, copper complexation by organic species is reduced by protonation of the acid functional groups on the organic material, whereas at high pH the formation of stable hydroxo-copper complexes may out compete Cu complexation by organic species. The treated samples followed the tenorite solubility pattern, but are under saturated with respect to this mineral. The copper solubility as predicted by MINTEQA2 calculations is, however, strongly affected by the stability of the aqueous Cu(OH) complex. Log K for the reaction Cu2§ + 2H20 - Cu(OH) + 2H § equals to -13.7 in the MINTEQA2 database and is more than two orders of magnitude higher than the value of -16.2 which has been reported by others (Rai and Zachara, 1984) as also been shown to be consistent with biological uptake of dissolved copper (Blust et al., 1991). If the value of-16.2 is used, the solubility of tenorite is calculated to be much lower, with leachates below pH 7 being under saturated, and above pH 7 over saturated with respect to this mineral. In summary, the solubility of Cu during bottom ash leaching is complicated by organic complexation and disagreement as to the stability of the aqueous Cu(OH) complex. In view of the above, the deviation of the heat treated ash sample, within an order of magnitude, from the Cu concentrations calculated in equilibrium with tenorite does not rule out this mineral as a possible solubility-controlling solid. Another approach that was taken was to extract the same bottom ash sample with a mixture of dissolved organic compounds. In Figure 16.6b, the results of the experiment are given indicating an increase in the release by about one order of magnitude after leaching with a mixture of acetic acid, propionic acid, butyric acid and valeric acid, which are present in leachates from domestic waste disposal. It appears that the addition of organic complexants increases the leachability of Cu, particularly in the pH range 9 - 10.5. It is interesting to note that a similar, but less pronounced, peak is observed in the unified pH plot for Cu at this same pH interval (Figure 16.6a). In Figure 16.6d, the role of different factors -inorganic copper (tenorite), organic acids, DOC controlling Cu leachability is schematically presented. The influence of reducing conditions on Cu leachability was addressed by using a recently developed redox stat (Comans, 1993) based on equilibration with H2 gas. The relation between Eh and pH under normal atmospheric conditions and forced reducing conditions is given in Figure 16.6b, whereas the consequences for the Cu leachability are presented in Figure 16.6c. A greater than one order of magnitude difference in leachability in the neutral to moderately alkaline pH range is observed. These results indicate that both changes in redox and removal of organic matter can result in decreased Cu release.

666

Molybdenum Mo is known to form very mobile oxy-anions, and is mainly present as MoO42 above pH 5. This element was measured only in the leachates from one installation (urban), with concentrations increasing strongly with pH between 4 and 7, and remaining virtually constant at higher pH. Of the Mo solids reviewed by Rai and Zachara (1984), which were added to the MINTEQA2 database for the purpose of this study, PbMoO4 and CaMoO4 are the most likely phases to control Mo solubility in the leachates. MINTEQA2 calculations showed the leachates not to be in equilibrium with PbMoO4. The Mo concentrations predicted from solubility control by CaMoO4 are shown in Figure 16.5h and could explain the leaching data between pH 7 and 10. The relationship is confirmed by plotting the data in a log Ca2*-activity versus log MoO42-activity diagram. CaMoO4 may, therefore, control Mo leaching in this pH range. This mineral has also been suggested to control Mo concentrations in high-Ca waters (Hem, 1985), but was calculated not to control Mo leaching from coal fly ash (Fruchter et al., 1991). Lead The Pb leaching patterns for the two bottom ash samples are clearly different from each other. While one bottom ash sample (urban) shows a strong pH dependency, with concentrations decreasing over three orders of magnitude from pH 4 to 10, dissolved Pb leached from the second sample (rural) remains high and is relatively pH-independent (Figure 16.7c). Equilibrium with Pb(OH)2 could explain the first ash sample data above pH 9 or 10, but the mechanisms behind the high Pb leaching from the second ash sample are unclear. Most dissolved Pb data from the first ash sample leachates between pH 7 and 10 are close to the ICP detection limit and do not allow conclusions as to the solubility controlling solid. Modelling studies on fly ash by Gardner (1991) indicate solubility control by PbSO4 in the lower pH ranges and Pb(OH)2 solubility above pH=7. Whether Pb solubility in bottom ash is partly controlled by sulphate remains to be proven. Zinc Similar to the other heavy metals discussed above, Zn is leached in high concentrations at low pH and decreases strongly (four orders of magnitude) with pH values up to 10 (Figure 16.7d). Complexation with hydroxide in solution increases dissolved Zn at strongly alkaline pH. The leaching pattern follows the solubility line for zincite (ZnO), with concentrations remaining down to 1 order of magnitude lower. A second possible solubility controlling solid is ZnSiO3. The solubility of ZnSiO3 in Figure 16.7b shows a pH-dependency similar to the leaching data and limits, except at very high pH, dissolved Zn to 1-3 orders of magnitude lower concentrations. Both zincite (at intermediate pH) and ZnSiO3 (pH 4-6 and pH > 12) may contribute in controlling Zn leaching from bottom ash. Gardner (1991) calculates Zn leaching from fly ash to be limited by the solubilities of Zn(OH)2 and ZnSiO3. Zincite is, however, less soluble than Zn(OH)2 and gives a better prediction of the data. It appears that the solubility control in bottom ash and fly ash is not significantly different for Zn.

Figure 16.7 Thermodynamic Activities from Column Experiments on MSWl Fly Ash -2 -4 A

+

N

D

f5 M 0 H

-6

(a)

;/ 0 0

-8

-2

-

on

-

0 0

0

-10 -

0

a ,

-I2 1

(b)

0

2

3

4

5

6

1 2 6

-

" " 8

'

" " 10

12

14

16

668

16.4.2 Modelling of APC Residue Leachability The modelling of APC residue leachability requires the application of corrections for the activity coefficient using Pitzer equations (Pitzer et al., 1973 & 1974) because the Davies equation and the Debye-H(3ckel equation are not valid at the ionic strengths present in leachates from APC residues. However, the general trend as observed for bottom ash is valid for APC residues. Extending the modelling efforts from bottom ash to APC residues opens a new area of research. The work of Gardner (1991) of modelling fly ash leaching from small columns and batch extractions is very much in accord with observations on bottom ash as described in the previous section. Figure 16.7 presents the solubility control of Pb and Zn from fly ash as a plot of element concentration against counter ion. This type of plot illustrates the boundaries of solubility control by specific mineral phases. It is clear that in different pH ranges, different minerals may control the leachability. It is, however, unlikely that many phases will control leachability at the same time and to the same extent. This implies that with the identification of the most relevant mineral phases (usually 2 or 3), the leachate composition should be adequately predicted. The task is now to sort out the most relevant phases in the relevant pH regions.

16.4.3 Application of Geochemical Modelling Results In combination with XRD and other surface techniques as discussed in Chapter 7, specific mineral phases can be identified. The modelling of the solution properties does not provide direct evidence for the existence of specific minerals in the bottom ash matrix, but the results of such studies support the hypothesis that geochemical reactions control the leaching of both major and trace elements from incinerator 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 processes can be used to: Predict the long-term behaviour of incinerator residues in the environment Improve the interpretation and further development of regulatory leaching tests, and Chemically modify the ash and/or its environment in utilisation or disposal, to minimise contaminant leaching Further work is needed to clarify the behaviour of some of the elements discussed above. In a number of cases, the degree of agreement between model predictions and actual measurements is very promising. General leaching behaviour and the impact of waste feed on incinerator operating conditions can be identified based on modelling efforts. In Figure 16.8 some of these

669 Figure 16.8 Identification of pH Domains with Solubility Control and Availability Control for Cd, Cu, Mo, Mg, Pb and Zn in Bottom Ash 1 O0

1000

Variation with the availability for leaching

I \

\\

,

i= d

e0 0

\

\ k

0.1

\

0.01

\ /

.........

0

' .... 4

2

1000

100

8 o l u b l l l t y control by Inorganic cormtltuents

' .... 8

' .... ' .... ' .... 8 10 12 14

0.01

0

....

' .... "'"' 2 4

e

....

a v a i l a b i l i t y for leaching

O/0

%

/

,N ~ / / /

/

' .... ' .... ' .... 8 10 12 14

Region of aolublllty control

-O-

9

10

~

C)

9 I

~0 0.1

go

1

0

0

0

0.01

0.1

Mg 0.01

\ \

//

,' /

Vadation with the availability for leaching

Region Of solubility COntrol by bruclla

0

c o

o.,

~~//

Variation due to the T

o

"4', ~ 4,', I "4,

"

Cd

0.00001

vE

X\

\00\

",,,, o Ic~

0.0001

Region of solubility control by DOC 0

\

Xx

0.001

OI

\

\

"

0 0

\

X

/

10

\;~

0

\

Reglo~ Of solublllty control by chloi'Ido and omrboflite

/-I_ %.,U

~

10o O

p ~

OD

Vm'iatJoe with the availability for leaching

4

Ms

~o

I

J

z L'- T - -

T -- T--

T--'T~'Z

5

8

7

9

11

8

10

12

1:)

0.001

........ 0 2

10000

1000

100

Region of solubility control

Region of solubility r

100

0.1

go

0.01

Pb

0.001

,,,,I 0

.... 2

a,,,,j 4

.... 6

pH

! .... 10

0

:I

1

o.1

i .... 12

9

lO

o/

! .... 8

14

4~?o

1000

~O I I

12

Variation wllh the availability for leaching

~1 Varlmtlon with the I availability for leaching

O

' ......................... 4 6 8 10

14

0,01

0

9169

Zn

......... 2

' ......... 4 6

t I I I

' .... ' .... ' .... 8 10 12 14

pH

670 trends are presented, which indicate regions where leachate concentrations are controlled by availability and regions controlled by solubility as defined by geochemical reactions. Such general trends appear to be relevant for other types of residues as well. 16.5 RELEASE RATES OF ELEMENTS The observed release of a specific element over a time interval is a result of the element's availability, solubility and rate of mass transfer from the solid to the liquid phase. The rate of mass transfer between phases is most frequently controlled by diffusion through the porous solid phase, geometric dimensions of the solid phase, and the mode of liquid solid contact. In general, the following cases can be used to describe the release of elements and inorganic species from incinerator residues: Localised equilibrium is attained between the solid phase and the contacting leachate, resulting in a leachate which is saturated with respect to the element of interest, i.e., solubility is controlling release. This case typically occurs for inorganic species, except alkali metals and halogens (e.g., Na, K, CI), with percolation of infiltration through residues in the field and during column leaching tests or sequential batch extractions of small particles in the laboratory. For this case, cumulative release with respect to time reflects the product of the liquid to solid ratio (LS) with the elemental solubility at the specified conditions (e.g., pH and Eh). .

,

Equilibrium is locally attained between the solid phase and the contacting leachate, but the resulting leachate is not saturated with respect to the element of interest because of limited availability or flow channelling. This case typically occurs for alkali metals and halogens with percolation of infiltration through residues in the field and during column leaching tests or sequential batch extractions of small particles in the laboratory. For this case, greater than 50% of the availability is released at an LS = 1. Release is controlled by diffusion through the solid phase and the contacting leachate is not in equilibrium with the solid. This case typically occurs for most elements and species when flow is around either a monolithic material (e.g., solidified/stabilised residues) or a granular material compacted to low permeability such that is behaves as a monolith (e.g., residues used as a road base and overlain by an asphalt layer). For this case, cumulative release is a function of effective diffusivity, availability and release time interval.

The following sections present observed release rates for elements as a function of L/S to address Cases 1 and 2, above, and diffusion controlled release to address Case 3.

671 16.5.1 Release As a Function of Liquid to Solid Ratio

The leaching of granular residues is in many cases dictated by percolation, which can be represented by column experiments. This is valid both for utilisation and for disposal, unless measures have been taken to minimise infiltration drastically. In that case infiltration may be reduced to such an extent that diffusion becomes the rate controlling transport mechanism. It also is important to establish the relationship between column and batch testing results because while column tests more closely reflect field scenarios, batch tests are more efficient to carry out in the laboratory. A comparison of leaching data from both column and batch tests has been carried out for a wide variety of waste materials (WAV, 1988, 1992). The two approaches were agreement except for the cases where the initial release and depletion of one species subsequently resulted in altered release for a second species (van der Sloot et al., 1993). In addition, some long term processes, such as those caused by biological activity can be observed in column tests carried out at low flow rates over extended time periods, but cannot be addressed in short batch tests. Test results generally are expressed in mg/kg to allow a direct indication of release. Presentation of column leaching results in the form of cumulative release as a function of L/S may be used to discern between different types of release behaviour (Figure 16.9): Type I-

Rapid release of highly soluble species due to under saturation of the constituents in the leachate even at low L/S. This results in rapid wash out of these species in a percolation dominated system with release of the available quantity within L/S less than 1 or 2. The slope of the cumulative release as a function of L/S typically is greater than or equal to 1 at L/S less 1 followed by a slope of approximately 0, indicating depletion. Elements that exhibit this type of behaviour include alkaline metals and halogens.

Type II-

Release controlled by solubility in the aqueous phase which most often is a strong function of pH. Cumulative release as a function of L/S is approximately linear with the slope dictated by the elements solubility. Elements which exhibit this type of behaviour include Pb and Zn.

Type III-

Delayed release due to retention in the matrix by a second species controlling solubility which is depleted after a limited time interval. This behaviour is characterised by a transition from linear release at a lesser slope to a greater slope. An example of this behaviour is the release of sulphate which initially may be limited by barium until depletion of that element occurs.

Type IV-

Enhanced initial release due to the presence of a complexing agent which increases the solubility of the element of interest. This behaviour

672

Figure 16.9 Types of Release Identified from Column Tests or Sequential Batch Tests

Type I

[c]

tEj

~Jf---

Availability

..........................................................

o

2

LS

2

10

[c]

~

Type II

J

I

10

[El

T

/

/ 2

~,

2

LS

Type III

[c]

1.8

10

[El

/

j

2

L$

Type IV

1 2

LS

IO

'

2

10

[c]

......

10

[El

/]

2

LS

/

673 is characterised by a transition from linear release at a greater slope to a lesser slope. Unlike for Type I release, depletion does not occur. Examples of this behaviour are the initially increased release of copper in the presence of organic acids, and the initially increased release of cadmium from APC residues in the presence of high chloride concentrations.

Bottom Ash In several studies (Versluijs et al., 1990; Hjelmar, 1992; F~llman, 1992; Eighmy, 1992, van der Sloot et al., 1991) column leaching experiments have been carried out using bottom ash. The results of these studies, presented in Figure 16.10, can be used to compare release from ash generated in Canada, Denmark, Germany, The Netherlands, Sweden, and the United States. The release of the various constituents has been presented as a function of L/S with ranges for total content and availability indicated for comparison. Results for bottom ash from batch extraction tests with distilled water at L/S=20 are provided along with the column results in Figure 16.10. Cumulative release of Cd, CI, Cu, Mo, Na, Ni and sulphate observed for column tests at L/S=10 and release observed in batch tests at L/S=20 were similar, indicating that the leachable fraction of these elements was released at US less than 10. Solubility controlled release was observed for Ba, Ca and Pb between L/S=10 and 20 based on a linear increase in cumulative release from the column to the batch data.

Antimony

The release of antimony from bottom ash ranged from 0.0005 to 0.5 mg/kg. In contrast, the availability of Sb is approximately 5 to 10 mg/kg, which implies that Sb is largely retained in the ash matrix. Release patterns were very consistent and indicate slow dissolution with increasing L/S. The wide range in Sb release most likely was related to variability in the waste feed composition.

Cadmium

The release pattern for Cd indicated a small initial release of a highly mobile species followed by a slow increase with increasing L/S. The release patterns generally had a slope of 0 after L/S=1, which indicated that the soluble Cd-species was depleted. Possible species responsible for this release behaviour are anionic Cd chloride complexes or organically complexed Cd. The cumulative release at US=I 0 ranged from 0.005 to 0.05 mg/kg compared to availability which was between 0.5 and 5 mg/kg, indicating that Cd release was limited and significant retention in the matrix occurred.

674 Figure 16.10 Release (mg/kg dry ash) of Selected Metals as a Function of the Liquid to Solid Ratio 1oo

~ ~"

10000

As

1000 I

lO

Total

1

Availability

Total

~= 10010I

]~a

Availability ....I

0,1

--~ 0,01 n" 0,001

~

0,1 0,Ol

0,0001 =,

9

0,001

100

LS (llkg) 100

1 10 LS (llkg)

CI Total

A,,eilability

...........................................

j l P I l - n u-u-ll

~ ] - , * i B - - o

0,1

"==t

0,01 0,001

100

1OOOO

Cd

Total .1 Availability

0,1

E 1000

D-i~-a

.............................................

0,0001 0,00001

........

0,1

IOO0

.....

,,,|

Total

lOO "~

,

1 10 LS (llkg)

1

........

00

,

100

100

1 10 LS (llkg)

0,1 0000

Cr

1000 ~

Total

100 l

Avaiiabllity

,I

Availability

Cu

I

0,1 ~

0,01 0,001 0,0001

0,1 ~ 0,01

0,1

1 10 L8 (=lkg)

........ 100

0,0Ol

i

o,1

........

~

1 Le

........

(I/kg)

:

lO

........

lOO

675

Figure 16.10 Continued 100

10o0

Mo Total

10

t

100

Total

10_ u

9

9

Availability. . . . . ]

9

Availability

~ 0,1 0,01

Ni

_

E n ~

i

,

,

,

.....

,

.

,

,

u . , , l

1

0,1

LS (llkg)

,

_

10

0,001

i

. . . . . .

lO0

............ 1 10 100 LS (llkg)

0,1

100000

100o0 lOOO I

Total

lOO l

I

~b

S04

Total = Availability

.........................................

A~ailability

~, 10000 r

i

0,1

-~ iooo m.'

O,Ol 0,0Ol 0,0001

,

,

o,1

,

9 ......

1 lO LS (l/kg)

1000 Availability

10

Total

I

lO0 lO

...:,

0,1 B

0,01

B

!1 II Ii

0,001

,,

0,1

~1111 9

............................. 1 10 100 LS (llkg)

1000

"'=l

1

0,1

0O00

Sb

100

0,0001

100

:

lOO

~

II

Zn

Availability

~

o,1

I

f,~

D-I~--D

n-=---~

0,01 ........

i

1 10 LS (llkg)

i

,

i

. . . . ,i

100

0,001

o,1

1 lO LS (I/Vo)

lOO

676

Chromium

The release pattern for Or indicated depletion of a soluble species after L/S=1 (slope = 0). The fraction initially released may be a chromate species, reflecting the high mobility of Cr (VI) in comparison with Cr(lll). A steady increase in release with increasing L/S (slope = 1), indicating slow dissolution, was noted for only one case. Cumulative release at US=10 ranged from 0.005 to 0.5 mg/kg, in contrast to availability between 2 and 10 mg/kg.

Copper

The release of copper was very consistent among different sources of ash, as indicated by the parallel pattern of the release curves. The release of Cu was indicative of a highly soluble fraction that was washed out within L/S-1 (approximately 2 pore volumes). The concurrent release of a dissolved organic species which complex Cu has been postulated as an explanation for this initial release. The cumulative release at L/S=10 from many different sources was within a relatively narrow range of 0.3 to 10 mg/kg, which is a small fraction of the availability which varied between 50 and 200 mg/kg.

Molybdenum

Two distinct data groups can be distinguished for Mo. The lower release curve was related to input from typical domestic sources, while it appears that the relatively high releases of Mo observed for two facilities resulted from industrial contributions of Mo rich waste streams which were included with the incinerator feed. Cumulative release at L/S=10 ranged between 0.2 and 10 mg/kg, which was approximately equal to Mo availability. These results indicate that Mo is quite mobile and almost all leachable Mo can be depleted from bottom ash in a relatively short time span. The relatively high initial concentrations of Mo in leachates may be of concern in some jurisdictions.

Lead

The release pattern for Pb was reasonably consistent for most cases, with increasing cumulative release with increasing L/S (slope = 1)indicating solubility controlled release. In a few cases, however, depletion appeared to have occurred. This type of release behaviour may be explained by a decrease in pH during the leaching process, which resulted in decreased Pb solubility. Cumulative lead release was considerably variable between data sets. At L/S=10, the cumulative release varied between 0.005 and 10 mg/kg. The sensitivity of Pb solubility to pH was the primary reason for the wide range in release observed. However, the cumulative release at L/S=10 was several orders of magnitude less than the availability, which indicated considerable retention of Pb with the ash matrix.

677

Nickel

The release of Ni was similar to that of Cu. Both Cu and Ni have been demonstrated to form strong complexes with dissolved organic matter, which is hypothesised to have been responsible for initial release of both elements. The cumulative release at L/S=10 varied between 0.02 and 0.5 mg/kg, with only two cases having significantly less release. Ni availability was several orders greater than the observed cumulative release, indicating significant retention in the ash matrix.

Sulphate

The release of sulphate approaches the maximum leachability at L/S=100, which indicates that sulphate can be depleted from the bottom ash matrix. The quantity of soluble sulphate present in the ash is such that it becomes a constituent of major concern in relation to potential effects on local ground water quality.

Zinc

The release patterns for Zn were fairly consistent for a given data set, indicating very slow dissolution from slightly soluble phases. However, considerable variability existed between data sets. Cumulative release of Zn at L/S=10 varied from 0.01 to 3 mg/kg. The primary reason for the variability has been attributed to variable pH, because of the sensitivity of Zn solubility to small changes in pH. This was similar to the behaviour observed for Pb. However, the cumulative release at US=10 was a small fraction of the availability, which ranged from 50 to 500 mg/kg.

Alkali Metals and Halogens (e.g. Na, K, CI, Br)

Alkali metals and halogens, such as Na, K, CI and Br, were completely leached from the bottom ash at L/S=1-2. There was no retention of these elements in the matrix, which implies that most equilibrium batch extraction tests can be used to assess the potential release. The quantity of these elements present coupled with their rapid release should be carefully considered during the development of management practices for bottom ash. The column leaching test results discussed above demonstrate consistent release patterns reflecting depletion, dissolution or in some cases a delayed release due to changes in chemical conditions with time of leaching. For a specific element such as Pb or Zn, cumulative release may vary over several orders of magnitude depending on the origin of the particular bottom ash sample evaluated. This observed variation in cumulative release results primarily from variation in ash alkalinity which controls leachate pH and solubility of elements. This effect is substantiated through the presentation of measured concentrations of elements in leachates from column tests as a function of pH in Figure 16.11. The unified pH-solubility curves also are presented for reference. In general, the pH of leachate from column tests with bottom ash does

678

Figure 16.11 Release (mg/kg dry ash) of Cu, Pb and Zn as a Function of pH to Indicate Solubility Controlled Release in Column Leaching Test (n) as Compared to the Unified pH Curves (e) 100 -

Cu _0

o

o

O0

-= o ~ , O~ o

~o

1

o

.

.

.

.

t

10

.

11

oo oo 0 0

o

o B o O []

[] o

o

0,1

o _ oo u

,.,

o

.

.

.

I

,

'

'

I

13

12

pH

o

1000O o

o

~ ~/

~

10

o

o

1000

o

o

oO

oO. o w ... .~fi_ ~_~~oO~O o-

o

o

o 11

12

13

pH

14

100000 Zn ,-, ..... E o

10000

1000

o

O

._ E

8

10

o

n

100

J~4b@@~n -OOnLl

~ 0

0

11

0

DO

0

n

0

O00D'~D

12

pH

13

14

679 not vary over a wide range. Pb and Zn data from the column experiments are consistent with the unified pH curve, which illustrates the solubility controlled release as a function of pH for bottom ash. Measured concentrations and trends as a function of pH from column and batch tests are consistent. APC Residues Column leaching experiments have been carried out using APC residues in several studies. The release from ESP ash has been reported in two studies carried out in The Netherlands (Versluijs et al., 1990; Born, 1993). Additional studies have examined release from dry and semi-dry scrubber systems in Denmark (Hjelmar, 1991) and the United States (Theis and Gardner, 1991).

The results of these studies, presented in Figure 16.12, can be used to compare release from the different types of APC residues. The release of the various constituents has been presented as a function of L/S with ranges for total content and availability indicated. Results for ESP ash from batch extraction tests with distilled water at L/S=20 also are provided along with the column results for comparison. A typical feature of results from column leaching of APC residues is an initially moderately alkali pH at very low L/S ratios, followed by increasing pH with increasing L/S ratios. It is hypothesised that this behaviour results from the formation of a shell of acid gas reaction products surrounding a carbonate shell, which in turn surrounds a hydroxide interior of the APC residue particle. The carbonate shell subsequently dissolves during leaching with increasing L/S. This effect is greater for dry and semi-dry scrubber residues because of the injection of lime into the flue gas. The resulting changes in leachate pH also effect the release of elements for which solubility is a strong function of pH. From the leaching curves as a function of L/S it is clear that the cumulative release is constant or increases only slightly after an L/S=1 (As and Cr from two installations are the primary exceptions). This implies that the release within one L/S constitutes the amount available for leaching at the pH controlled by the residue and represents soluble chemical species that are washed out of the system with minimal retention. This mode of release is prominent for many constituents in ESP ash as evidenced by the release-L/S profiles. The variability in leachability of ESP residues within and between installations is substantial. It appears to be related more to the input to the incinerator and facility operation than in the case of bottom ash. Initial leachate concentrations and release of individual elements from APC residues during column tests typically are in agreement with the solubility as a function of pH derived from batch testing. Initial solubilities may increase relative to those observed for bottom ash because of the high ionic strength and chloride concentration of the solution. However for APC

680 Figure 16.12 pH and Release (mg/kg dry ash) of Selected Metals as a Function of the Liquid to Solid Ratio pFl

As 0,1

== 11

j

0,01

" 0,001

0,0001 0,01

0,1

1

10

100

.......

0,01

I

.......

0,1

LS 1000

........

I

10

1o0000O

Cd

100

I

1 LS

I

.......

100

CI

=

{,o

e == 1o000o

==

~

o,1 0,01

o

, 0,01

0000 0,1

1

10

100

. . . . . . . ; . . . . . . ; ........ = ....... 0,01

0,1

1 LS

LS 10

10

,oI

100

Cr

e .=

100

Cu

1.1-I-I

m-D--I

. . . . . . . . ',

....... I

....... i

0,1

1

==

.. 0,1 "

0,1

-t: [

0,01

................................. 0,01

0,1

1 LS

f 10

100

0,011

I

0,01

LS ( l l k g )

10

........ ; 100

681 Figure 16.12 Continued

100

100

Ni

Mo

10

lO

J

v

j

-

)

1

0,1

........ I

0,1

0,01

........ I

........ :

1

10

0,1

........

:

100

O,Ol

........

0,01

a

........

i

0,1

1000

10000 -~

m

........

i

100

Zn

Pb

1000

v

I

10

LS

LS

E

........

1

lOO

100

9 m

tr

lO ram=

r

1 0,01

........

I

0,1

........

:

1 LS

........

I

10

........

I

100

1 0,01

........ : 0,1

I

m

m m

........ : 1 LS

|

m

........ I 10

........ I 100

682 residues, availability is an important parameter controlling the release level with depletion occurring for many elements at L/S less than 1. As a consequence of these observations, column tests are not very functional for the evaluation of APC residue leachability. Availability and pH dependent solubility are much more a practical tools to assess APC properties. 16.5.2 Diffusion Controlled Release

Assessment of constituent release most frequently assumes that the predominant mode of leaching will be by percolation of water through the porous solid matrix. However, in several circumstances the release may not reflect percolation and solid-liquid equilibrium, but rather be controlled by diffusion through the porous solid. This situation may occur when a fine-grained material is placed in a surrounding matrix with a higher permeability, thus creating a preferential flow around, instead of through, the material (Environment Canada, 1990). A similar situation will be encountered when the material is compacted during placement to form a less permeable matrix, or when infiltration of water is minimised through use of low permeability barriers such as compacted clay covers. Evaluation of release under these circumstances requires measurement of diffusion controlled release fluxes. Measurement of diffusion controlled release fluxes for monolithic solids is achieved by refreshing the leachant in contact with a well-defined surface geometry at regular intervals. An analogous test for compacted granular material has been developed which maintains an undisturbed surface through use of a thin layer of glass beads placed on top of compacted granular material in an inert mold. This test was first applied in the framework of a USEPA study on stabilisation/solidification of incinerator residues (Kosson et al., 1993). After an initial delay in the release caused by the layer of glass beads, the release profile generated permits the calculation of tortuosity and chemical retention within the matrix (refer to Chapter 21). The intrinsic leaching properties obtained from this type of measurement allows prediction of release at longer time scales than the actual testing period. In addition, it provides an estimate of diffusive contribution to transport in a low flow column experiment. Table 16.5 presents the measured values of tortuosity and pD for several elements in untreated bottom ash, APC residue and combined ash (Kosson et al., 1993). Estimation of the pD values are based on using the availability of each element in the specific matrix as the driving force for diffusion. This approach results in a more accurate estimate of the release parameters than use of the total concentration as the driving force. Compaction of bottom ash and combined ash resulted in a reduction in permeability, which was reflected in the pD values for mobile species, such as Na, K, and CI. The difference in physical retention between bottom ash and combined ash compared to APC residue was substantial. The latter behaved as a loose powder without physical

683 Table 16.5 Diffusivities (-Iog(m2/s)) Measured using the Compacted Granular Leach Test for Residues Element

Bottom ash pD

Tortuosity

Std. dev.

24

Combined Ash

APC Residue

pD

Std. dev.

pD

13.6

0.2

As AI

14.06

0.09

13.63

0.19

Ba

12.3

0.08

12.04

0.08

Br

10.05

0.17

9.91

0.12

Cd

>15

Ca

12.68

0.06

12.75

CI

10.5

0.16

Cr

11.73

*

Cu

>14.8

Fe

14.2

Pb

16.17

>15

15.2

0.6

0.09

10.3

0.15

10.52

0.21

9.3

0.1

10.35

*

11.6

0.15

14.57

0.17

14.2

0.2

0.19

15.2

0.21

11.2

0.2

0.09

16.3

0.37

14.2; 11.8; 12.9

Li

11.93

*

11.69

*

Mg

14.66

0.72

15.1

1.06

Ni

>13

11.02

*

NO3

11.37

0.36

10.36

0.60

K

10.12

0.07

10.18

0.08

Si

14.49

0.12

13.48

0.06

Na

10.24

0.08

10.26

0.09

Sr

11.5

0.07

11.45

0.09

SO4

15.62

*

15.27

0.18

16.01

Zn

15.71 data. Kosson et al., 1993

Std. dev.

25

11.2

0.2

9.0

0.1

9.0

0.1

0.17

13.6

0.4

0.35

15.9

0.3

* single

restriction; only chemical retention was due to the chemical environment (high pH). The mobility data for combined ash and bottom ash are very similar, which is largely related to the fact that the pore water conditions (e.g. pH) are very similar. The low retention values for K, CI, Br, nitrate and Li are in agreement with the high release of these constituents observed in column experiments. The high retention values for

684 several other elements also are in agreement with release observed from column and batch tests. The difference between APC residue and bottom ash is significant for Ca, F e, Pb and minor for sulphate and Zn. In the case of Pb the difference is largely attributed to the higher pH in the APC residue versus the bottom ash. The large variability in Pb leachability from APC residue reflects the sensitivity to minor changes in pH. The further development and application of this procedure needs to be explored for predicting release from incinerator residues.

16.6 RESIDUE LEACHING IN THE CONTEXT OF REGULATORY LEACHING TESTS AND WASTE FROM OTHER SOURCES 16.6.1 Regulatory Tests and pH Dependent Leaching Regulatory leaching tests in most jurisdictions are batch tests which control the leaching conditions through definition of the leachant composition, L/S ratio, sample preparation (e.g., particle size reduction) and equilibration period. The final pH of the leachate is controlled to a large extent by the acid neutralisation capacity of the waste being tested. Measured leachate concentrations of elements and species are then compared to fixed regulatory thresholds. Figure 16.13 illustrates the impact of alternative presentations on the interpretation of results. Clearly, pH is the most important independent variable. However, most regulatory data that is obtained is focused on a narrow pH range dictated by the waste being tested, which limits the utility of results for defining characteristic release behaviour. This pH range may not reflect actual disposal or utilisation conditions, and may be in a range for specific elements where small changes in end point pH cause large changes in observed leachate concentration. Thus, the result is often a testing scenario that differentiates between the alkalinity of a waste, not the potential for release of specific elements and environmental impact. Figure 16.13 clarifies this effect through schematic presentation of the results of several regulatory leaching tests for Pb. The unified pH-solubility curve and the concentration range and threshold values for each test are presented. Deviations from a general leachability pattern can be related to changes in solubility controlling properties either imposed by the test, such as the Swiss TVA or the California WET test, or by changes in ash characteristics due to changes in input or operation conditions of the incinerator facility. None of the regulatory tests consider quantity of total soluble salts present in the material being evaluated.

German DIN 38414 (1984) and French AFNOR X-31-210 (1988) These tests are similar in that they are based on the conditions dictated by the material to be tested. The final leachate pH can vary widely depending on the acidic or alkaline properties of the material being tested. The pH range commonly observed for incinerator residues extends from 9 to 12. In this region, the solubility of some elements changes rapidly with minor changes in pH, especially metals which exhibit amphoteric behaviour. Consideration of this phenomenon should be given when interpreting test results, thus ensuring that the associated test conditions are relevant to actual management practices. For example, leachate pH from bottom ash disposed

685 Figure 16.13 A Comparison of Different Methods of Presentation of Regulatory Leaching Test Results for Lead

Regulatory limit

..........

200

.

10

.

.

.

.

.

.

.

.

.

e

~., . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

0

~

0 0

.

.

.

. Aft..}

A

J

g~

Regulatory limit

1

oo /

~

_

|

13

~

0 ......

0

I

I

I0

20

0.3

.

30

4

l

i

!

,

,

,

l

,

w

6

7

8

9

10

11

12

pH

LS

50 f .

. . . . . . . . . . Regulatory [ limit e 9

[ 0

~o e~

"

~

4

.......... 2'. . . . . . . . . . . . . . . . . . •. . . . . . . . . . . . . . . . . . . . . . . . .

10 L

v

t I

--v-+"

1

. . . .

.~

d

0

1 /.

I

9

10

OA~ . . . .

9

Ot 11

pH

I'\ I \

0

9 9 . _ %" o 0

ISS DIN

ro

0.1 ,TCLP

o

oo ~

+ I 12

_ 13

0.01

I

4

S

.I

1

6

7

.

1

1

I

l

I

8

9

10

11

12

pH

13

686 in landfills is rarely greater than 10 and test conditions manipulated higher than pH 10 will require different interpretation.

Japanese Leaching Test

The Japanese Leaching Test is based on extraction with water for 6 hours at L/S=20, however, the short duration of the test may not allow sufficient time for several elements to reach equilibrium. Testing results should resemble those obtained by the DIN and AFNOR procedures, although there is currently a paucity of leaching data on incinerator residues.

Swiss TVA (1988)

The Swiss test method generally produces results which are consistent with the pH dependent leaching curves discussed previously. The acceleration of carbonation provided by bubbling pure carbon dioxide through the leachant produces some side effects that are not representative of field conditions. This test is very sensitive to minor changes in pH because testing is carried out in the pH region where solubility changes quickly with pH. The test has also been shown to produce an exceptionally high release of oxy-anions. This is attributed to diminished retention capacity of the ash by the conversion of lime and basic calcium silicate phases to calcite, via additional carbon dioxide. The rate of retention of oxy-anions in other mineral phases is not as fast as the rate of release resulting from carbon dioxide injection, since these re-mineralisation reactions can be relatively slow. Slow sorption of oxy-anionic species, such as arsenate, selenite and molybclate, on hydrated ferric hydroxide phases is likely to occur in the field, but is too slow to be of significance during the laboratory leach test. This aspect needs to be considered for the relevance of the TVA test for oxy-anions.

USA, Califomia WET Test

The California WET test is a modification of the TCLP test which uses dilute citric acid as the leachant instead of acetic acid. This modification results in much greater complexation of metals than in TCLP. Results of the WET test indicate availability when evaluated on a release basis. Insufficient data is available to assess the comparability of WET test data for anionic species.

US EP Toxicity Test (1980), the Toxicity Characteristic Leaching Procedure (TCLP) (1990) and the Regulation 309 (now 347) Leach Procedure

The intended final pH of 5 of the EP Toxicity and TCLP tests is in a pH region where major changes in metal solubility can result with minor variations in pH. The final pH of the leachate from both tests is highly dependent on the acid neutralisation capacity of the waste, and incinerator residues typically require much more than the maximum 2 meq of acid per gram of ash addition to bring the pH down to 5. However, variation in buffering capacity leads to inherent variability in the test results.

687

Alternative Approach Since many of these tests expose ash to conditions which typically would not prevail in most disposal scenarios, an alternative approach is suggested for the regulatory evaluation of incinerator residues. The leaching properties of these materials are now well characterised. Testing of the materials should be focused on evaluating whether or not the sample being tested exhibits the same characteristic leaching behaviour for elements or species of concern. This would allow for implementation of management strategies which are most effective for the general material class and rejection of materials with characteristics beyond acceptable variability limits. An example of this approach would be to specify acceptable ranges of alkalinity and leachability in relevant pH domains, in particular at neutral pH. This would be valid for disposal and for utilisation. However, the criteria for acceptance would be different. 16.6.2 Systematic Leaching Behaviour Among Different Incinerator Residues Streams and Other Wastes Figure 16.14 presents a comparison of the release (mg/kg) of several constituents from different residue streams. It is striking to note that the curves show many similarities in spite the substantial differences in composition and origin. The solubility control in the pH region 6 to 11 is very pronounced for the metals. The primary difference observed between the residues is the total availability, which is generally higher for the combined ash than for bottom ash. For the APC residues, the availability can be one to two orders of magnitude greater than for bottom ash. The grate siftings typically indicate approximately an order of magnitude greater availability for Cu and Pb compared to bottom ash. The behaviour of Ba is very consistent for all incinerator residues. The solubility of Ba was independent of pH for most residues except from boiler ash, which decreases with increasing pH. The key to this behaviour is the influence of sulphate, which was not reported, because BaSO4 is the solubility controlling phase in most cases. The most significant difference between incinerator residues was noted for Cd, where the greater chloride concentrations in the APC residues results in increased Cd leachability at alkaline pH. For Cu, low content of organic matter in certain residues (e.g., ESP ash, boiler ash) tends to lead to low leachability of Cu. However, LOI is not a proper measure for the possible increased Cu leachability, because char alone does not affect Cu leachability (see section 16.2.3.1). Cr is more soluble in the boiler ash and SD-FF residue than the grate siftings and the D-FF residue. The reason for this is unclear but may be related to Cr speciation (e.g., Cr.6 is more soluble at neutral and alkaline pH than Cr*3). The leachability of Zn is the most consistent for all residues. Even the availability does not differ more than a factor of three to five. Table 16.6 provides a comparison of the total content, availability, and leachable fraction of several elements in the different residue streams. The availability is

688

Figure 16.14 Comparison of the Leaching of Selected Metals as a Function of pH from Bottom Ash, Combined Ash, ESP Filter Ash, Fabric Filter Ash, Grate Siftings and Boiler Ash 100 -

A Dry-FF= Grate sil'dngs o Boiler ash

10-

9Bottom ash o8o

~ o

t~t

E

r

I=

8

0,1

B

0,01

0,001

....

0

I ....

~ ....

I ....

I

2

4

6

8

....

I . . . .

I

10

....

12

I

14

pH 100 & Dry-FF a Grate siftings 9Boiler ash

10

9 AA,,

~

Ok

4>

=_.

~o

~

ee

A

9Bottom ash

E

G

0,1

O O

0,01

Ba

0,001 2

4

6

8 pH

10

12

14

689

Figure 16.14 Continued 1000 & Semi-dry-FF o Combined ash

106

+ESPash 9D n/-FF [] Grate siQngs

10

9Boiler ash J

o

9 0

*

8

o

0 (~) 0

0,0.%

~8 O8c~#A

(~U

0,01

....

a

I ....

I ....

4

9

+

O

2

9

o

9O

0,001

9

040

0,1

9Bottom ash

9

+ 9

9

+

O

I ....

6

I

pH

....

.~

+e

I ....

8

I

10

....

12

I

14

10000,00 9Bottom ash A Semi-dry-FF

1000,00

n Combined ash + ESP ash

100,00

00

9 00

Dry-FF [] Grate si~ngs

E (J

§176

10,00

9Boiler ash

+

O

cJ 1,00

O

O O ~

[] &

9

+

A

0,10

0,01

Pb

....

9

I ....

I ....

2

4

I ....

6

I ....

8 pH

+

+

I ....

10

I ....

12

1

14

690

Figure 16.14 Continued 100,00 I

A Semi-dry-FF & Dry-FF , Boiler ash + ESP ash o Combined ash

+ A , ~ A'~ d~A&

+

10,00

9

O

1,00 -~

9 & ~o t eSO,e~ + oo~ o" ~o ~

%

E

r

0,10

O

O

I

%0 * ~,,~,,

0,01

Cd

0,00

9 ~176 I

0,00 0

2

4

6

pH

8

10

12

14

100,00 9Bottom ash A Semi-dry-FF o Combined ash ,, Dry-FF o Grate sifdngs o Boiler ash

51,o~ o o

10,00

--~ 1,00 E

~

c

0,10

0

~%

9149

n

Cr

0,01

~ &

O

A

r-t r-k

l.Detection

0,00

. . . .

0

I ....

I ....

I ....

t

2

4

6

8

..................................

pH

~t,

r'n

limit {

....

I ....

10

I ....

12

I

14

691 Figure 16.14 C o n t i n u e d

10000,00 9Bottom ash 1000,00

qb Oo~

9 _=i_A&.~

9Semi-dry-FF

z~

oE~h

o Combined ash o Grate siltings

100,00 e

Ot

E

'-:'. U c O

& Dry-FF

0 9

e Boiler ash

0 9 o,t-

+ ESP ash

o

10,00

&

,&

&

cJ

00

1,00

+

[]

Zn

0,10

9

9

O

9~p

+ogL

O

0,01

....

0

I

2

....

i

4

....

I ....

6

r

....

8 pH

II I ....

10

I ....

12

I

14

Table 16.6 Total Content and Availability of Several Elements for Each MSWl Residue Stream Availability is presented on the basis of release per mass of ash, release per mass of MSWl incinerated, and fraction of total content Stream

As Min

Availabilitv (malMo M S ~ Bottom ash 0.09 Grate siflings Boiler ash Filter ash APC resiaue Total available mass per element

B Max

Min

Ba

Ca

Max

Min

Max

Min

Cd Max

Min

Max

21000 0.15 1050

1000 .

n

O .-1 t~

#.

100 -

109 1 0.01

I

I

0.1

1

Waste Loading Table 20.1 Ranges of Tortuosity Values Measured for Untreated Incinerator Residues, Products Containincj Residues and Reference Materials Material Unconsolidated granular Waste 1 Compacted Bottom Ash 2 Compacted Fly Ash 2 Stabilised Bottom Ash 2 Stabilised Fly Ash 2 Stabilised Fly Ash 6 Stabilised Fly Ash 3 Pavement Blocks Containing Bottom Ash 4 Pavement Block Reference 4 Asphalt Concrete Containing Bottom Ash 4 Asphalt Concrete Reference 4 Asphalt Concrete Containing Fly Ash 5 Asphalt Concrete Reference s 1. ECN Work, Various Wastes 2. ECN Study 3. BCR intercomparison

Tortuosity 2-4 20 - 25 5-10 1-6 25 - 35 200 - 210 7-19 35 1900 - 3300 24000 450 - 1400 > 3000 4. Mammoet 5. UNH Whitehead 6. Mehu/vdSloot Record

874 with decreased aqueous solubility of the element of interest either as a function of pore solution pH or the presence of a solubility limiting species. Elements or species that are highly soluble (e.g., chloride, sodium) have relatively little or negligible chemical retention. Table 20.2 provides a comparison of chemical retention values observed for several elements in untreated and treated residues and reference materials. Note that relatively high chemical retention values are observed for untreated residues and low values are observed for asphalt matrices. Table 20.2 Ranges of Chemical Retention Values Measured for Untreated Incinerator Residues, Products Containing Residues and Reference Materials Chemical Retention Values

Material

a b c d e f g h i j k I

Cd

Cu

Pb

Zn

CI

n.a 5 x102 2 x102 1.6 x 108 3 x 108 40,000 >30,000 (220) 150-600 1 n.d.

32,000 8x104 7x102

3.1 x 106 1,300-9x104 2x102-2x106 13,800 70,000 - 98,000 13,900 1.5 x 106 (820) 1,300- 3,100 1 n.a.

4 x 105 1.5x102-4x106 3x106-3x107

2.9 1.2-1.5 1-5 2.5 1.5 - 2 2 92 (110) 2- 3 1 1.4- 2

n.d.

>2 x 104 (780) 17- 26 1 n.a. n.a.

a. Compacted Bottom Ash b. Compacted Fly Ash c. Stabilised Bottom Ash d. Stabilised Fly Ash e. Stabilised Fly Ash f. Stabilised Fly Ash n.d. - not detected

20.5.3

n.a.

790,000 7 x 102 (17,800) 100- 200 1 n.d. n.d.

n.d.

g. Pavement Blocks Containing Bottom Ash h. Pavement Block Reference i. Asphalt Concrete Containing Bottom Ash j. Asphalt Concrete Reference k. Asphalt Concrete Containing Fly Ash I. Asphalt Concrete Reference n.a.- Not analysed

Solubility

Cadmium solubility as a function of pH is presented in Figure 20.18 for untreated APC residue from a semi-dry scrubber/fabric filter unit and the same APC residue SIS treated with Portland cement and treated with soluble phosphate (Kosson et al., 1993). It can be seen that the effect of treatment on solubility was minimal. Similar information is presented in Figure 20.19 for chromium in untreated and treated APC residue and bottom ash, and for lead in Figure 20.20. The characteristic solubility curves for chromium were the same for untreated bottom ash, cement stabilised bottom ash and

875 Figure 20.18 The Effect of SIS Processes on the Solubility of Cadmium for Bottom Ash and APC Residues Bottom ash, untreated 100

-

--

APC residue, untreated

. . . . .

100 - ~

_ .... odla o dido

o%c~~ o ~

Oo

dB~

a)~

o

o

,,.,..

%

b)

o

0.1

0.1

Cr 0.01

0

o P"

2

4

6

pH

8

10

12

0.01

14

Bottom ash, phosphate treatment

0

2

4

6

8

:_:_

10

pH

!

12

14

APC residue, phosphate treatment

100

100

10

~Oo~ v

% o oo

o

% 0.1

0.01

0

2

4

6

pH

o%

0,1

oo

8

10

12

0.01

14

Bottom ash, Portland cement treatment

0

2

4

6

8

pH

12

14

APC residue, Portland cement treatment

lOO

100 OO

Q:~o

o

o

f)~ o.l.J O.Ol

10

~ c

e

~

.

.

pH

0.1 ~

~

~

0.01 -~ 0

o

c~ 2

........ 4

~

~. 6

,., 8

__ 10

12

.... 14

876

Figure 20.19 The Effect of SIS Processes on the Solubility of Chromium for Bottom Ash and APC Residues APC residue, untreated

Bottom ash, untreated 10

a>

oo

1 r

0.01

.

r

.

,

2

eo,~O

g

r

,

0

%

b) g

Cr

0.1

o Cr ]

O

q~ 8 "_ --_

.

-

4

.

6

.

pH

o

o

.

.

8

12

O.1

0.01

.

10

14

~176

. . . . . . . 2 4 6 8 pH

0

10

12

APC residue, phosphate treatment

Bottom ash, phosphate treatment

oo

o

c)

..

o

I

d)

oz

0

0.1

o.oi

.

0

.

2

.

.

4

o o

1

g

Ib

~3

...

.

.

.

.

6

.

pH

8

.

.

10

.

12

o.oi

14

Bottom ash, Portland cement treatment

c,~176

0.1

0

2

4

~ 8 6

pH

8

1

0.1

='-

_

_

.

2

.

.

4

.

.

.

6

.

.

pH

8

.

10

f) g o

m,~

.

12

_

i4

{oc;) o

c~

0

14

10

o

0.01

12

APC residue, Portland cement treatment

10 ~

.e) ~

10

Q~o o o d=~ O.1

0.01

% ~-

0

. . . . . . . . . . . . 2 4 6 8 10 12

pH

14

14

877 Figure 20.20 The Effect of SIS Processes on the Solubility of Lead for Bottom Ash and APC Residues

APC Residue, Untreated

Bottom Ash, Untreated 1,000

.

.

.

.

.

1,000

10o

100

o

0

~o

a)

o

0.1

0

2

4

6

o

oo

8

pH

% E

9

(~:)

10

12

b)

oo

9

10

JD

0.1

14

o

%

'~ 1

Bottom Ash, Phosphate Treatment

o:

Pb

0

o ,..,.,

2

4

6

pH

,=

8

10

12

14

APC Residue, Phosphate Treatment

-

1,000

6;~OOo

..-.,.

1 fO00

-

,

,

to O

100

c) ~ .....

o

10

10

0

.o 13.

%o

8o ~~176 o

tt

ooo

g~

Pb

0.1

0

2

. . . . . .

4

6

pH

8

. . . . .

10

Bottom Ash,. Portland Cement Treatment 1,000

~

.....

,

.o

l~.

10 1 0.1 1 0

o oO

wb 2

o

,r

4

0~

~ 6

pH

8

10

12 14

...........

2

4

6

pH

8

10

12

I

14

1,(X)O

[oPb)

o

0

APC Residue, Portland Cement Treatment

......

lO0

e) ~

0.1

12 14

100

f) ~

JO 0.

r

10 o o 1

0.1

Pb

0

"

" 2

. . . . . 4 6

-

pH

. . . . 8 10

"

12

14

878 cement stabilised APC residues. These curves exhibited a regime of intermediate solubility between pH 5 and 10, and sharply increasing solubility with pH decreasing below 5. The solubility between pH 5 and 10 may be indicative of chromate (Cr+6). The curves for untreated APC residue, APC residue treated with soluble phosphate and bottom ash treated with soluble phosphate indicate increasing solubility only at pH less than 5. This suggests that treatment with phosphate may have facilitated conversion of chromium to less soluble speciation, while cement-based treatment of APC residue facilitated the formation of more soluble chromium species. For lead, treatment of APC residue resulted in decreased solubility between pH 5 and 11 while treatment of bottom ash had limited effect. The solubility of an element or species of interest also can significantly effect the testing protocols required for estimating diffusion controlled release. A critical assumption in the tank leaching protocols is that the leachant remains dilute with respect to species of interest during each leaching interval. Figures 20.21 and 20.22 present the measured leachate concentrations for copper, cadmium, lead and zinc from several studies. The unified pH solubility curves for bottom ash (see Chapter 16) are provided for comparison. For the case of copper, all observed leachate concentrations were significantly less than the unified solubility curve. However, for most data sets, the observed leachate concentrations of Cd, Pb and Zn are below or equal to the solubility curve. In a few cases, data above the solubility curve have been measured. In these cases, the dilute solution criteria have not been met and the release data have to be interpreted with caution. Future testing for these cases should be carried out either with increased liquid to solid surface area ratios or with shorter leaching intervals. If release is controlled by solubility, other means of assessing long term release are needed. The use of release assuming an effective diffusion coefficient based on the highest pDe measured during testing is a worst case assessment of release. For the purpose of evaluating acceptability of release, this may already be adequate in view of the limit values concerned.

20.6 INTEGRATED INTERPRETATION OF pD,, AND AVAILABILITY Reductions in constituent release from a treated material can be achieved either through reducing the fraction of that constituent available for release, through reduction of the rate of release (increased pDe), or through modification of both critical parameters. In addition, the PDe can be modified either through physical or chemical effects of treatment. A useful mechanism for evaluating the combined effects of both availability and pD~ is needed. Estimation of diffusion controlled release from specific geometries and conditions over prolonged time periods can be accomplished through use of intrinsic leaching parameters and application specific geometries and exposure conditions. An important advantage to the use of intrinsic leaching parameters derived from the availability and diffusion release test data is the ability for prediction of release under conditions other

879

Figure 20.21 A Comparison of Tank Leaching Concentrations for Various Products as a Function of Solution pH and the Unified pH-Solubility Leaching Curves for Cu and Cd

Cu 10

9 r o

A

0.1 U C

o

US FA P3 US BA

9

O

O

G)

9

o.ol

e@

9

9

u

US BA

9

o.ool 0.0001

=

'

6

',

'

7

', .'

8

',

,

9

I

,

I

,

t

,

I-

,

P1

US FA P2 9

9

P4

US BA P9

A

...I

Unified pH MSWI BA

NL ASPHA 2% BA

I

10 11 12 13 14 pH

Cd

0.1 E c o

== u r o

0.01

9 ,&

0.001

9

0.0001

0.00001 6

7

8

9

10 11 12 13 14 pH

Unified pH MSWI BA France FA NL ASPHa 2% FA

880 Figure 20.22 A Comparison of Tank Leaching Concentrations for Various Products as a Function of Solution pH and the Unified pH-Solubility Leaching Curves for Pb and Zn

Pb A

NL asph Fa 9

0.1

0.01

0.001

R--

NL concrete BA

o

Unified pH MSWl BA

US BA P1

9

US FA P1

9

US FA P2

9

US FA P3

9

US BA P4

m

0.0001

'

5

I

'

I

6

'

7

I

'

8

I

'

9

:

"

~

"~.-;

"

;

"

;

10 11 12 13 14

pH

US BA P9

9

DK FA

9

France FA

Zn 10 A

NL asph Fa 9

ol

E t-

u c o

-o

Unified pH MSWl BA

US BA P1

9

US FA P1

0.1

o

0.01 O

0.001 9 4NIb

0.0001

9

~

6

"

',

7

"

:

8

"

I

9

'

I

10

pH

'

I

11

9 '

I

12

'

I

13

'

I

14

NL beton BA

a

9

US FA P2

9

US FA P3

9

US BA P4 US BA P9

881 than those studied in the laboratory. Many factors affect the translation of laboratory results to prediction of field behaviour. Field environmental conditions that are important include residue aging, contact with infiltration and precipitation frequency, temperature cycles, direct abrasion or erosion and the specific application scenario. Thus, estimates of field releases must be carefully derived. However, simplified models can be used to indicate relative releases and provide order of magnitude or limit case assessments. The availability and pDe leach parameters can be used to predict the release of contaminants during a given time period for a variety of application geometries. A 3-dimensional diffusion model enables one to take actual dimensions into account, so differences in leaching from a product with a cubic versus a flat rectangular shape can be described. With the 3-D model, release from only one side of the material also can be modelled. A 3-dimensional model is based on the analytical solution of the linear diffusion from a parallelepiped, which initially is at a constant concentration, to an infinite region outside with a constant surface concentration (Crank, 1989; de Groot, 1993). The diffusion profile is calculated in all three dimensions according to the equation: C0

13.3 ,=0 m.o n.o

XGOS

(2/+1)(2_m'-~2n+11

cos (21+1)nx 2a

(2m+ 1)nYcos (2n+ 1)nz exp(-tal, m,n) 2b 2c

where:

(20.26) a,,m,n=-~t 2~.--~1)2+(2% +1

+ 2n+1 2t

'C"

(20,27)

Integration of the constituent flux across the surface boundary with respect to time results in an expression for calculation of cumulative release (Crank, 1989):

Mt=fo

-n---~- p

~176 ~0 ~o (2/+1)2(2--~+1i2(2n+1)2

Application of Equation 20.10 permits estimation of the cumulative release of a constituent as a function of time. The cumulative release, expressed as fraction of the total leachable quantity (Rma,), can be calculated using the 3-D model for different product configurations and bulk applications based on the effective diffusion coefficient

882 measured during laboratory testing or in the field under well-defined boundary conditions. The usual boundary condition applied for field translation is that the surface concentration of the leaching constituent is effectively zero. The cumulative release, expressed as fraction of the total leachable quantity (Rm~), can be calculated using the 3-D model for different product configurations and bulk applications based on the effective diffusion coefficient measured under well-defined boundary conditions. The relative release from standard sizes with dimensions of 10 x 10 x 10 cm and 15 x 15 x 45 cm were calculated as a function of time for different effective diffusion coefficients ranging from PDe=9 to pDe=15 (PDe = -log De with De in [m2/s]). The cumulative release-time curves are provided in Figure 20.23. Between blocks of increasing size, the difference is largely a shift of the cumulative release for a given pD e to a longer timescale. It takes longer to reach the maximum leachable quantity, but the leachable quantity may ultimately be reached unless the chemistry or other release controlling factors change. A significant shift in the cumulative release curve is apparent for the roadbase simulation. For a base of 15 cm thickness, 50% of the highly mobile components (PDe=9) will be leached from the slab in less than approximately one year, assuming permanent contact with water. A 45 cm thick slab will reach the same level of relative release in about 6 years. It is important to consider that translation of lab data to field conditions further involves several factors such as corrections for the ambient temperature and degree of contact with water. Estimating release during utilisation must consider adjustments to the pD measured in the laboratory to reflect anticipated field conditions (Kosson et al., 1995). Temperature, the fraction of time the surface is wetted, and the degree of water saturation are important considerations. While the diffusion coefficient is a function of the diffusivity of the constituent of interest in water, tortuosity, and chemical retention, only diffusivity (Do) is significantly a function of temperature. The temperature dependence of diffusivity in dilute ionic solutions can be considered to be proportional to the absolute temperature over limited temperature ranges [20.28], e.g.,

DT2-

OTIT2 71

(20.29)

The above relationship assumes that the viscosity of the pore water (leachate) does not change significantly over the temperature range of interest. Alternatively, the effect of temperature on D has been correlated for release from cement stabilised products containing waste materials according to [20.29]:

PD1-PD2=0"71 '-~1-

(20.30)

The cumulative release for a series of wet dry cycles can be approximated based on Equation 20.24 by.:

883 Figure 20.23 Estimation of Fractional Cumulative Release as a Function of Time and Effective Diffusion Coefficient (pDe)

MONOLITHIC Block

.f

= o n-

~ E

= (3

0.80

~

~

.-

--

MATERIALS Block

cm

-~ ~-

"" /I / ..."~:.1= ",o-Q~/ ,," i11/ ..."'/

1.00

~= n" E

lOx10x10

CONSTRUCTION

. ..

~: ~r" n"

0.80

0.60

-/

0.20

,'

/

.."

../

r

.-"

-"

o.oo ~

10-2

/

.

'

~

10-1

/ ,a,.,a,.:

/

_ /'

'

~.- - ~ . . . . . . . . .

100

101

102

~

0.40

E

0.20

= (3

/

/

,

~.~ . ~ 10-1

10 .2

Time [years]

DIFFUSION

COMPACTED

/

/

0.60

/

0.40

// ~:>=,o'

0.20

/ /

1 )'2

//

/

10 "1

'

/

//

'

/

,//

/ / ~>=11// f /

100

/

/

~

/

I

..

)

9

/

."

."

."

....'"" .

/

."

/~=,..=

/ / /,,>/,,

,

100

.~.~..~.~. 101

102

103

101

Time [years]

/

/

/

//

/

/

//

/

/

/

/

=

~>,.12 .

/

102

cm

/

."" =0,.,3

...'"

45

1.00

," "

.'"

LAYER

Roadbase

f f---:: ~:>-,e /

/

GRANULAR

15 cm

1.00

0.80

/

//~=11/--I..-"

Time [years]

CONTROLLED

Roadbase

/

/

I

/r / .'" / /

0.00

03

,

//

,'

/

e=

/

/

/ /

e

,

cm

1.00

0.60 0.40

15x15x45

I

103

E n" n" ,,._. r e) r _.e rr

0.80

//,"' /

0.60

~>

0.40

E (~

0.20

0.00

10 .2

~=111

11//

10-1

100

101

Time [years]

102

103

884

Mr,,,1d

(20.31 )

or -o-,

=

=MtF,,~d

(20.32)

Note that the calculation of IVItassumed a continuously water saturated material. Figure 20.24 illustrates the effects of temperature and wet/dry cycles on diffusion controlled release from a 45 cm thick roadbase (Kosson, 1995). The cumulative effect of these conditions can be significant over long time intervals. Calculation of cumulative release curves for a variety of geometries and applications is impractical on a routine basis. A simple one-dimensional diffusion model, assuming a constant source, can be used as an approximation. The advantage of this approach is that cumulative release is only a function of availability, pDe and the exposed geometric surface area. The one-dimensional model is independent of application specific geometry. This approach is valid as long as the concentration in the material has not decreased substantially, avoiding species depletion. If depletion does occur over the time period of interest, the one-dimensional model will over predict release, providing a conservative estimate for decision making. However, the amount of a specific element released will not exceed the availability of that element for all cases. The one-dimensional model is based on the "semi-infinite slab" solution of the diffusion equation provided by Crank (1989):

M,=2pCo (-~) ''2 where: Mt Co D t r

= = = = =

(20.33)

cumulative release [mg/m 2] availability [mg/kg] effective diffusion coefficient [m/s 2] time [s] density [kg/m 3]

This is the same basis which was used to estimate the pD from tank leaching data. The initial and boundary conditions for this solution are (i) the initial constituent concentration is uniformly distributed in the matrix; (ii) the exposed surface for leaching has a liquid concentration which is maintained essentially at zero; and (iii) depletion does not occur. Figure 20.25 provides a comparison of cumulative release estimated using the three-dimensional model and the one-dimensional model applied to a 10 x 10 x 10 cm block and a 45 cm thick roadbase (one surface exposed) for two different

885 values of pDe. The effect of depletion on release is indicated by the greater predicted release by the one-dimensional model. However, in all cases the release estimated by the one-dimensional model will be equal to, or greater than the release predicted by the three-dimensional model. This confirms that the one-dimensional model provides a conservative estimate of release, provided that no significant changes in chemistry occur. Figure 20.24 The Effect of Intermittent Wetting and Temperature on Diffusion Controlled Release of Product Constituents

Road base 0.45 m

1.20

X

1.00

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

pD= 10, Sat., 20 ~

E r~

0.80

(D (/) (1)

0.60

pD= 10, Sat., 10~

-

_

pD= 10, Wet/dry, 10 ~

x

(D

rY .O >

0.40

l:::

0.20

~

/

/

o 0.00 -"- " ~ 0.1

--I--I-1

r'l'r'l--

"-

1

T

~

~ ~

~i

I

i

,

10

Time (years)

i

,~

,~

I

100

i

i

~ ~,

,~

1000

886 Figure 20.25 A Comparison of Fractional Cumulative Release for Products Estimated Based on the "Infinite Slab" Approximation and Accounting for the Product Geometry and Depletion Block 10x10x10 cm

1.20 . . . . . .

I

x

E E

rv" n,'

v

!

1.00 ................ ~.................

0.80

iIIll ,

pD=10 / /

I

(D (D n" ~

E

fO

0.60

iJ /

O 0. 0.20

~

j'

i

pD=14

, I

" .t

0.00 0.01

~ 1

0.1

' 100

10

1000

Time (years)

1.20 X

E E

n,, n,"

v I

1.00

Road base 45 cm

..................................................

0.80

~-.......

I

(D

(D n"

0.60 0.40

E O

./

0.20 0.00 0.001

0.01

0.1

1

10

Time (years)

/ 100

1000

887 Interpretation of the availability and pD information can be further simplified if a fixed time interval is defined over which to consider release. For example, a 100 year interval may be considered as the useful life for roadbase or other construction applications. Use of the one-dimensional model permits the development of charts which provide estimated cumulative release per unit area as a function of only availability and pD. These charts are referred to in this text as "cumulative release nomographs". Examples of cumulative release nomographs are provided for several elements in Figures 20.26 through 20.28. Availability and pD are presented on the x-axis and y-axis, respectively. Diagonal lines indicate lines of constant cumulative release (e.g., 10, 100, 1000, etc. mg/m 2 over a 100 year interval). Lines of constant cumulative release decrease in value from lower right to upper left of each figure. Data plotted above and to the left of a line of constant cumulative release provide less release than the indicated value; data below and to the right provide greater release than the indicated value. Cumulative release nomographs provide a straightforward method of interpretation for evaluation of data from laboratory testing of diffusion controlled release. An acceptable cumulative release can be defined for specific applications or locations based on evaluation of the potential impacts of that release. Data for applications which result in release less than the defined limit would be considered acceptable. Thus, diffusion controlled release information can be readily incorporated into a regulatory framework. Consideration of the cumulative release nomographs presented in Figures 20.26 through 20.28 also provide information about the effectiveness of various treatment processes and products for incinerator residues. Observations are summarised in the paragraphs that follow. Cadmium (Figure 20.26) Cumulative release for bottom ash incorporated into cement pavement blocks was between 0.1 - 1.0 mg/m 2. SIS treated bottom ash and fly ash incorporated in asphalt had release between 1.0 - 10 mg/m2. Untreated APC and SIS treated residue typically had cumulative release > 10 mg/m 2.

Copper (Figure. 20.26) Cumulative release for fly ash in asphalt, bottom ash in paving blocks and residue from the 3R process was between 10 and 100 mg/m 2. Untreated and SIS treated APC residue cumulative release was between 100 and 1000 mg/m 2. Untreated and SIS treated bottom ash and combined ash generally had cumulative release in excess of 1000 mg/m 2. Lead (Figure 20.27) SiS treated bottom ash, bottom ash in cement pavement blocks and fly ash in asphalt generally had cumulative release less than 100 mg/m 2. Untreated bottom ash,

888 Figure 20.26 Availability-pDe Plot of Release Parameters for Cadmium and Copper from Tank Leaching Tests on Untreated Incinerator Residues and Several Product Materials Containing Residues

1~l

/o.,

"F

/.

15

/".I:.

l"r ~

9

14

/.

/

l~ / ,

9

9

13 0.01

/,

,

e e

~

I !

9

0.1

i

10

100

1000

Availability (mg/kg) 16

~.~10

/,

15

,,,-,

14

,~. , , ~

§

j

s.

13

/100

..

./v

_p......

q) a a-12

11 10

[ j/..V , ooo /,oooo .

1

.

.

.

.

.

.

.

10

.

.

.

.

"

9

.. .. .. .. .. .. .. .. .. .. .

100

9

I

.

Ou cu

.

.

.

1000

Availability (mg/kg) Diagonal lines represent constant cumulative release (mg/m 2) estimated for 100 yrs using the "infinite slab" approximation

889 Figure 20.27 Availability-pDe Plot of Release Parameters for Lead and Zinc from Tank Leaching Tests on Untreated Residues and Several Product Materials Containing Residues

19/

,~1o /

18tPb

j ~oo s O " "~'-'.kd

17

1~

15

o.

9

13

~

|

I

12

~

11 10 9

1000 1

.

10000

10

-100000 1 O0

Availability 18

1000

10000

(mg/kg)

Zn / ' ~ y,/'~176 / '~176176 / /...'o / /

17 cI e~

/'

16

/

"/

/.;.. "J. ~

/

o0

15

14

/.

......

13 10

/,

77"-

100

.....

Z

1000 Availability

,o. ooo . . . . . . . . . . . . . . 10000

100000

(mg/kg)

Diagonal lines represent constant cumulative release (mg/m 2) estimated for 100 yrs using the "infinite slab" approximation

890 untreated combined ash, SIS treated combined, SIS treated fly ash, and fly ash in asphalt had cumulative release between 100 and 1000 mg/m 2. Untreated APC residue and SIS treated APC residue generally had cumulative release in excess of 10,000 mg/m 2. Phosphate treated APC was the exception to this with cumulative release between 100 and 1000 mg/m 2. Zinc (Figure 20.27) Bottom ash in cement pavement blocks had cumulative release between 10 and 100 mg/m 2. SIS bottom ash, SIS combined ash and fly ash in asphalt had cumulative release between 100 and 1000 mg/m2. Untreated bottom ash, combined ash and APC residue, and SIS treated APC residue generally had cumulative release in excess of 1000 mg/m 2. Sodium (Figure 20.28) Bottom ash in cement pavement blocks had cumulative release of approximately 10 mg/m2. Bottom ash in asphalt had release between 100 and 1000 mg/m2. All untreated residues and other treatment conditions had release in excess of 1000 mg/m 2. Several processes resulted in cumulative release in excess of 10,000 mg/m 2. It is important to recognise that because the PDe for sodium is less than 10.5 for most cases, the majority of the cumulative release will occur during the first ten years of application with subsequent depletion (see Figure 20.24). Chloride (Figure 20.28) Bottom ash in asphalt, fly ash in asphalt and bottom ash in cement paving blocks all had cumulative release of approximately 100 mg/m2 or less. Untreated and SIS treated bottom ash and combined ash generally had cumulative release between approximately 10,000 and 100,000 mg/m 2. SIS treated APC residue and fly ash had cumulative release in excess of 100,000 mg/m 2. As with sodium, the majority of the release will occur during the first few years of application. Figure 20.29 summarises the affects of various treatment processes on diffusion controlled release from incinerator residues. In general, incorporation of residues into asphalt decreases release through increased physical retention. Therefore, the process can be effective for both highly soluble salts and trace metals. Treatment of residues by SIS processes decreases release through chemical respeciation and modification of the matrix alkalinity. Therefore, SIS processes may be effective for reducing release of trace metals but are generally ineffective for reducing release of highly soluble salts, although release may be delayed. SIS processes also may be ineffective when highly soluble salt Ioadings are so great, as in the case of APC residues, that a highly porous matrix remains after depletion of the salts. Availability of all of the residue constituents may be reduced by any of the above processes by having limited residue loading in the final product (dilution effects).

891 Figure 20.28 Availability-pDe Plot of Release Parameters for Sodium and Chloride from Tank Leaching Tests on Untreated Residues and Several Product Materials Containing Residues 13

12

11

l::l

10

9

8

J,.oo:. .....

7 1 oo

"..a ........ ' 1 o00o

10oo Availability

12

11

/

'

'

" 1 ooooo

(mg/kg)

100

100

9 ,~

s Ill

10

/-

Ei

7 100

/

T/". ~,4

,.oooo..l ,,!oo7o

...... I

.............

1000

10000

Availability

100000

1000000

(mg/kg)

Diagonal lines represent constant cumulative release (mg/m2) estimated for 100 yrs using the "infinite slab" approximation

892 Figure 20.29 A Schematic Illustration of Treatment Process and Environmental Conditions Effects on the Availability-Effective Diffusion Coefficient Relationship

18

Poresealing Tortuosity increase by cementation Low effective watercontact Lower temperature

17 16

Decreasing availabilih/ by vitrification

15

(b C~ O_

14 13 12

Dilution with additives Matrix incorporation by recrystallisation

n

11 10

m

m

I I0

Desorption by remineralisation Less retention due to acidification I I O0 1000

Availability (mg/kg)

10000

893 REFERENCES

American Nuclear Society Standard Committee Working Group ANS 16.1. "American National Standard Measurements of the Leachability of Solidified Low-Level Radioactive Wastes by a Short-Term Procedure" American Nuclear Society, La Grange Park, IL, 1986. Barna, R., P. Moszkowicz, J. Mehu, H.A. van der Sloot. "Waste Solidification Modeling and Simulation of Sodium Chloride Release from Leached Concrete", Int. Symp.Waste_ ManaQement 1992. Praag, pp. 65- 68, September 1992. Batchelor B. and K. Wu. "Effects of Equilibrium Chemistry on Leaching of Contaminants from Solidified/Stabilised Wastes", In Chemistry and Microstructure of Solidified Waste Forms. Edited by R.D. Spence. Lewis Publications, Boca Raton, FL, 1993. Crank, J. The Mathematics of Diffusion, Oxford University Press, New York, 2nd edn., 1989. de Groot, G.J., D. Hoede. Verfiinin.q van de beschrijvin.q van de .uitlo,q.in~ van resstofprodukten, ECN-C-93-085. ECN publication. 1993. (3 D model) de Groot, G.J., H.A. van der Sloot and J. Wijkstra. "Leaching Characteristics of Hazardous Elements from Coal Fly Ash as a Function of the Acidity of the Contact Solution and the Liquid/Solid Ratio", In Environmental aspects of stabilization and solidification of hazardous and radioactive wastes, ASTM STP 1033, P.L. Cbte and T.M. Gilliam, Eds., American Society for Testing and Materials, Philadelphia, 1989, pp 170-183. de Groot, G.J., H.A. van der Sloot, P. Bonouvrie and J. Wijkstra. "Karakterisering van het uitlooggedrag van intakte produkten", Mammoet deelrapport 09, March 1990. Eighmy, T.T., D. Crimi, S. Hasan, X. Zhang, D.I. Gress. "The Influence of Void Change, Cracking and Bitumen Aging on Diffuional Leaching Behavior of Pavement Monoliths Constructed with MSW Combustion Bottom Ash", In: Proc. of the 74th Transportation Research Board Meeting, Washington, D.C., Jan. 1995. Hinsenveld, M. "Towards a New Approach in Modelling Leaching Behaviour", Waste Materials and Construction. Edited by J.J.J.M. Goumons, H.A. van der Sloot and Th.G. Aalbers, Elsevier, Amsterdam, 1992. Kosson,D.S., T.T. Kosson and H. van der S l o o t . Evaluation of Solidification/Stabilization Treatment Processes for Municipal Waste Combustioq Residues, NTIS PB93-229 870/AS, 1993. Milieutechnisch onderzoek AVI slakken toepassing Rijksweg 15. Rijkswaterstaat, Dienst Weg en Waterbouwkunde. Report D 0421-71-001. July 1992.

894 NEN 7341. Leaching characteristics of soil-, construction materials and wastes Leaching tests - Determination of the availability of inorganic constituents for leaching from construction materials and waste materials. NNI (Dutch Standardization Institute, Delft), 1994. NEN 7345. Leaching characteristics of soil-, construction materials and wastes Leaching tests - Determination of the release of inorganic constituents from construction materials, monolithic wastes and stabilized wastes. NNI (Dutch Standardization Institute, Delft), 1994. van der Sloot, H.A., and D. Hoede. AVl-bodemas als aanvulmateriaal. ECN-c-91-044, 1991. van der Sloot, H.A., G.J. de Groot and J. Wijkstra. "Leaching Characteristics of Construction Materials and Stabilization Products Containing Waste Materials", In: Environmental.Aspects of Stabilization and Solidification of .Hazardous and Radioactive Wa.stes, ASTM STP 1033, P.L. Cote and T.M. Gilliam, Eds., American Society for Testing and Materials, Philadelphia, 1989. van der Sloot, H.A., G.J. de Groot, J. Wijkstra, and P. Leenders. "Leaching Characteristics of Incinerator Residues and Potential for Modification of Leaching",in Proceedin.qs International Confer.ence on Municipal Waste Combustion.., Hollywood, 1989. van der Sloot, H.A. "Systematic Leaching Behaviour of Trace Elements from Construction Materials and Waste Materials", in Waste Materials in Construction~ Proceedin(:]s of the .International Conference on Environmental. Implications of Construction with Waste Materials, Elsevier, Amsterdam, 1991. Whitehead, I.E. "An Enviromental Evaluation of Bottom Ash Substitution in Pavement Materials", Master Thesis, University of New Hampshire, Durham, New Hampshire, 1992.

895

C H A P T E R 21 - U T I L I S A T I O N 21.1

INTRODUCTION

Utilisation of MSW incinerator residues is being conducted or considered for a variety of applications in many countries. Interest in utilisation principally is motivated by the potential for extending existing ash landfill capacity, and thus reduces disposal costs, and in some regions the substitution for natural aggregates. The relative importance of each of these factors varies considerably from country to country and between regions within a country. Primary applications include use as: an aggregate substitute in paving applications, including as compacted base, or in bituminous pavement, an aggregate in terrestrial Portland cement applications, including cement block and prefabricated or field erected forms an aggregate substitute in Portland cement-based marine applications such as artificial reefs and shoreline protection 9

daily cover for municipal waste landfills, or

9

granular fill material for embankments.

Bottom ash is the primary material being utilised or considered for utilisation in construction applications. However, there are some notable exceptions, for example, a small percentage of fly ash has been used as a fine aggregate filler in asphalt in The Netherlands and the use of combined ash has been considered in the United States. Almost all of these applications, except use as daily landfill cover, would involve some degree of ash processing, either physical and/or chemical. For example, most applications would require screening of ash to achieve a desired particle size gradation or would result in ash encapsulation in another matrix. Utilisation through recovery of chemical constituents (e.g., CaCI2) and recycling of ferrous and nonferrous metal is discussed in Chapter 17 (separation processes). A typical pavement consists of the following layers or a subset combinations of layers depending on design (listed from the top driving surface down): a shim/levelling course, a wearing~surfacecourse, a binder course, a base course, a sub-base course, a compacted subgrade, and a natural subgrade. The shim/levelling course is placed on the surface to level ruts and depression and typically consists of a fine grain sand. The wearing~surface course is the top 1 to 5 cm and the binder course is below the binder course. The binder course serves as the bottom portion of the roadbed if

896 needed. Otherwise, it will be placed between the wearing course and the base course. The base course is normally the lower portion of the pavement. However, a sub-base may be required and is place directly below the base. The pavement is built from the bottom to the top on a subgrade that has been prepared by compaction. The entire roadbed is placed on natural subgrade or fill. Applications in the marine environment include shoreline protection and artificial reefs. These involve the use of the residues mixed with Portland cement to form concrete structures. Shoreline protection is the process of creating physical resistance to disruptions, such as storm events, natural erosion, and boat wakes. Examples of shoreline protection are bulkheads, sea walls, breakwaters, jetties, and piers. Artificial reefs are constructed to provide structures for the growth of marine organisms and attraction of fish while additionally serving as shoreline protection. All of the above applications require a final product which has a high degree of physical durability. 21.2 CURRENT AND PLANNED PROJECTS

The following sections provide a summary of current and planned utilisation projects and testing programs in several countries at the time of writing. 21.2.1 Canada

Because of the availability of landfill and the modest production of incinerator ash, there is little incentive, either economic or environmental, to pursue ash utilisation applications in Canada. The only example of ash utilisation is the use of bottom ash from the Burnaby Incinerator facility to construct access roads within a landfill. As utilisation becomes more wide spread in other countries, it is expected that the practice will be considered more seriously in Canada. 21.2.2 Denmark

Danish incinerator facilities produce approximately 420,000 tons of bottom ash annually (including grate siftings and boiler ash which in most cases are mixed with the bottom ash) the overwhelming majority of which is utilised. Since 1974, screened and sorted bottom ash has been utilised in Denmark for civil engineering purposes, particularly as subbase material at parking lots, bicycling paths, and paved/unpaved residential and major roads (Hjelmar, 1992). As a subbase material, the bottom ash is usually substituted for the diminishing supplies of natural gravel in various parts of Denmark. Thus, the incentives for utilisation include both natural resource conservation and economic benefit. A substantial portion of the economic benefit is derived from the avoided costs of landfilling, which is typically $150 (US) per tonne.

897 21.2.3 Germany In 1991, the 48 German incinerators in operation produced 2.56 million tonnes (Tg) of bottom ash. About one half of the annual production was utilised (1.2 Tg) and almost 0.2 Tg of scrap ferrous was recovered for recycling. The remaining amount (about 1.2 Tg) was landfilled (Johnke, 1993). The bottom ash is sieved and ferrous metal is removed prior to utilisation. The primary uses reported were as aggregate for compacted roadbase and embankments, primarily in demonstration projects. The extent of utilisation varied considerably from state to state. There remains strong opposition against the use of bottom ash even though positive results have been obtained from demonstration projects. Following an inquiry of 176 municipalities with >10,000 inhabitants, only 6 make use of bottom ash on a regular basis, with another 11 making use of it on a limited basis. Air pollution control residues, including fly ash and scrubber residues, are undergoing pilot-scale and provisional evaluation for use in the coal mining industry as filling and sealing materials for excavation cavities, and as aggregate substitute in grouts (Plate, 1992). The hard coal mines of Ruhrkohle AG (RAG) consume approximately 1.5 million tonnes of grout per year and the potential capacity to incorporate APC residues in grouts is approximately 20,000 tonnes/yr. Approximately 50,000 tonnes of APC residue has been used for a pilot test of mine filling and sealing operations. Scrubber residues also are being considered for use in alinite cement (Oberste-Padtberg, 1992). In this case, scrubber residue is used as a substitute for lime as a raw material. Scrubber residues are pelletised with other raw materials and about 20% water, and then treated in a rotary kiln to form an alinite cement clinker which is subsequently ground into cement. 21.2.4 The Netherlands Approximately 600,000 tons of bottom ash and 80,000 tons of fly ash are produced annually (1988) in The Netherlands (Born, 1994). Ferrous scrap, representing about 70,000 tonnes/yr, is separated magnetically at all facilities and recycled in the steel industry. The government policy is to achieve utilisation of more than 80% of incinerator residues. In practice, approximately 95% of the bottom ash is currently utilised (Born, 1994). The principal motivation for utilisation in the Netherlands is the shortage of suitable natural aggregate and the lack of available landfill space. The primary use of bottom ash is in the following applications: road base material for roads and industrial sites material for embankments, noise and wind barriers aggregate in concrete and concrete products, and aggregate in asphalt concrete

898 A total of more than 2 million tons of bottom ash has been used in the listed applications. In addition, 30-40% of the fly ash produced since 1984 has been used as a fine aggregate filler in asphalt concrete. The following is a summary description of several major ash utilisation projects: Caland Wind barrier- This project carried out in 1985 used more than 650,000 tons of bottom ash in an embankment with a length of 700m and a height of 15m. The ashes are covered with a primary cover layer of 0.5m compacted clay with a sand drainage layer (0.5m) and top soil (1.0m) overlaying the clay layer. The slope of the compacted ash is 1 2 and 1:2.5. The mean compaction factor was 97.5% with a wet density of 1840 kg/m 3. Groundwater quality is monitored on both sides of the embankment. Highway A-15 Rotterdam - Approximately 400,000 tons of bottom ash was used in an embankment for this major roadway construction. Ash is covered with a compacted sand-bentonite mixture with a minimum thickness of 20 cm to reduce water infiltration. Road Base Material - Several projects with bottom ash as road base material have been carried out in Rotterdam and North Holland. In Rotterdam, primarily a mixture of ash, crushed rubble aggregate and additive (50%-50%-10%) has been used. The base thickness is 25-30 cm. Concrete Paving Blocks - A project in Keilehaven carried out in 1984 used more than 300,000 concrete paving blocks in which the coarse aggregate was replaced up to 40% with the 5-8 mm fraction from bottom ash. After five years of traffic, it was concluded that there was no difference in physical properties between standard concrete paving blocks and those using bottom ash as the coarse aggregate replacement (Leenders, 1988). Laboratory investigations of environmental properties of concrete paving blocks with 20% replacement of the coarse aggregate by bottom ash indicated no significant difference from conventional paving blocks (KEMA, 1986).

Hartel Canal Pilot Project - A pilot project with asphalt containing bottom ash was carried out in 1987 along the banks of the Hartel Canal. A length of approximately 50 m was coated with about 100 tonnes of asphalt containing 30% bottom ash. It was found that the mixture temperature was required to be 40~ higher and the bitumen content 3% higher than traditional material. 21.2.5 Sweden

Current utilisation of incinerator residues in Sweden is very limited due to the regulations for licensing the use of any residual products in a specified manner (under the Environmental Protection Act). This also applies to controlled tipping.

899 Uncertainty with regard to whether or not a license is obligatory under the Environment Protection Act for the utilisation of residues has resulted in essentially no full-scale utilisation of incinerator residues through 1993. However this uncertainty is supposed to be resolved within a short time frame (e.g., 2 years) with the overall intent to increase utilisation of waste materials. New regulations are likely to require quality control based on environmental parameters. It is unclear if incinerator bottom ash will be deemed acceptable since these parameters have not been established. The legal, environmental and engineering aspects of utilising incinerator bottom ash have been studied in a comprehensive project (Lundgren & Hartl~n, 1991 ). The results based on field and laboratory tests indicated that sorted bottom ash can be used as embankment fill, base course material in low traffic roads and under light buildings and floor structures. Recommendations included that the use of bottom ash should be restricted to applications where the ash is covered by a low permeability material, such as asphalt, and deposited away from the groundwater table or ground water catchment areas. It was also recommended that the thickness of a deposit should be limited to 3.0 m and the ash placed well above the ground water table until further experience is gained. 21.2.6 United States

In the United States, several factors have influenced the possible uses for incinerator residues in construction applications. The shortage of existing landfill space and the difficulty of securing new sites has created a situation in which either disposal fees are costly, disposal space must be sought in distant locations, or disposal will not be possible. Hence, recycling and reuse of residual wastes has been suggested as the preferred management option. A detailed summary of planned and ongoing demonstration projects utilising municipal solid waste incineration residues in the U.S. has been prepared by Hoffman (1993). The following paragraphs and Table 21.1 highlight summary information for project parameters. Type and Distribution of Application: Approximately half of the 23 identified projects utilising residues are concrete applications. Of those 11 projects, 5 involve using residues in concrete blocks for buildings, 4 marine application, and 2 are used for landfill functions. Six of the remaining 12 projects are asphalt road paving applications. The most common initially proposed was in the road wearing surface. Project plans have been substantially modified in several cases, resulting in more road sub-surface demonstrations. Other projects include utilisation of loose MSW residue aggregate as the base layer underlying a paved parking lot, commercial scale substitution of fly ash for a raw material in cement production and fill material for inactive salt mines. Plans for the remaining three projects do not contain specific identification of use.

900 Table 21.1 Summary of Incinerator Ash Utilisation Projects in the United States Project MSWl Facility Ash Type Type .Asphaltic Applications Mass Burn/DS-FF Combined ennepin County, Minnesota Pavement emonstration Hillsborough County Department of Solid Waste Municipal Incinerator Ash Reuse, Research Development and Demonstration Project, Florida

Mass Burn/ESP

Combined

McKaynite Demonstration, Acline Street, Florida

Mass Bum/ESP

Combined

McKaynite Demonstration, Ruskin, Florida

Mass Bum/ESP Mass Bum/DS-FF

Bottom Bottom

New Hampshire Bottom Ash Paving Project

Mass Burn/DS-FF

Bottom, Combined

NYSERDA- Phase Ila, New Jersey

Mass Burn/DS-FF

Bottom, Combined

Concrete/Cement Ash Management Building, OH (Montgomery County)

Mass Burn/ESP

Bottom

Center for Innovative Technology, VA

Mass Burn/DS-FF

Bottom, Combined

Commerce Refuse-to-Energy Ash Treatment and Reuse, Los Angeles County, CA.

Mass Bum/DS-FF

Combined

Fly Ash Stabilisation Building, OH (Montgomery County) Islip, Blydenburgh Landfill, Long Island N.Y.

Mass Bum/ESP

Bottom

Mass Bum/DS-FF

Combined

Pinellas County, Florida Artificial Reef

Mass Bum/DS-FF

Scrubber, Bottom

Residential Foundation, New York

RDF/ESP

Bottom

SEMASS Administration Building Project, MA

RDF/DS/ESP

Bottom

SUNY Artificial Reef Demonstrations

Mass Burn/ESP Mass Burn/DS-FF

Combined Combined

SUNY Boathouse Demonstration

Mass Burn/ESP

Bottom, Combined

Other Commercial Grade Cement Production (Tacoma, Washington) (shale replacement)

RDF/FF

Fly Ash

City of Albany, NY, Parking Lot Demonstration (loose aggregate)

RDF/ESP

Bottom

Hawaii: Field Tests of Use as Landfill Cover

RDF/ESP

Bottom

Metropolitan Washington, D.C. Demonstration

Mass Bum/DS-FF

Bottom, Combined

NYSERDA - Phase II B - New York City

Mass Burn/DS-FF

Bottom, Combined

RDF ESP

Refuse Derived Fuel Electrostatic Precipitator

DS FF

Dry Scrubber Fabric Filter

901

Geographic Distribution: The majority of projects cluster in two areas - Northeastern United States (New England, N.Y., and N.J.), and Florida. Additional locations include Los Angeles County, California and Honolulu, Hawaii. Factors influencing this distribution are lack of landfill space, availability of natural aggregate, and vendor and government interest in utilisation. Type of Incineration Facility: All but five of the sources of incinerator residue for the demonstration projects are mass burn facilities. Of those five, four employ refuse derived fuel (RDF) in a conventional boiler. The fifth also uses RDF, but co-combusts it with coal and wood waste in a fluidised bed burner. Slightly more than half of the incineration facilities supplying residue for the projects are equipped with a scrubber and fabric filter combination. One of this group is equipped with an additional lime injection system in the furnace, and one is fitted with a de-NOx treatment system. Eight are equipped with ESPs only.

Incinerator Residue Fraction: Six projects propose using only bottom ash, and one uses only fly ash or APC residues. The others either use bottom and combined ash in the experimental design or are debating which residue streams to employ. The trend is toward bottom ash use, rather than combined ash. Incinerator Residue Processing: In each project, the incinerator residue is processed to some degree before it is substituted for natural materials in asphalt and concrete media. Pre-combustion processing is employed in RDF: large non-combustible items and ferrous metals are removed, and waste is shredded. Post combustion processing frequently consists of ferrous and nonferrous metal removal and particle size control. In addition, stockpiling (aging) is done to improve the engineering performance of the residue. Because virtually all of the projects substitute incinerator residue for natural aggregate, efforts are made to supply the replacement material in a form as similar as possible to the natural material. Some companies have patented their aggregation processed and have applied trade name to their products (e.g., Ardellite, Boiler Aggregate, McKaynite, Permabase Plus, Rolite, etc.). A range of 5% to 90% substitution for natural aggregate in asphalt and concrete applications exists among the projects.

21.3 CURRENT REGULATORY FRAMEWORK The regulatory framework for utilisation of incinerator residues is evolving due to the current debate. Therefore, the following descriptions of regulatory frameworks are intended only to provide a summary of different approaches. Only Denmark has a regulatory framework which is not under revision.

902 21.3.1 Denmark

The utilisation of granular incinerator ash for civil engineering purposes in Denmark has been regulated since 1983 by rules issued by the Danish Ministry of the Environment (Statutory order No, 568 of Dec.6, 1983). These rules apply to the use of small and moderate quantities of incinerator ash for specified purposes. Large scale applications of incinerator ash involving more than 30,000 tons of ash and ash applied in layers thicker than 5 m are regulated under the Disposal and Discharge Permit Act (Section 5 of the Environmental Protection Act). In 1989, these regulations were supplemented by a set of technical guidelines for the utilisation of bottom ash as a subbase material, issued by the Danish Highway Department (Phil et al., 1989). The principles of the rules and guidelines regulating incinerator ash utilisation in Denmark are illustrated in the diagram shown in Figure 21.1. In principle, both bottom ash and fly ash or combined ash may be utilised. In practice, however, all fly ash and virtually all combined ash will fail to meet the conditions set on the heavy metal content of the ash. Therefore, the ability to collect the bottom ash separately from the fly ash at the incinerator is mandatory to ash utilisation. There are chemical composition requirements for each use. Each portion of incinerator ash, with a maximum 5,000 tons, intended for utilisation must be sampled. The sample, which may be collected on-line (e.g., from a conveyor belt) or from a stockpile, must be a composite of at least 50 sub-samples of 2 kg each. If the ash has not been screened prior to sampling, the composite sample is passed through a 45 mm screen to remove large objects. In order to facilitate the subsequent crushing, ferromagnetic material, pieces of nonmagnetic metals, and pieces of unburnt material (paper, fabric, etc.) may be removed from the screened and air-dried ash sample, which is then reduced to 5 kg by means of a riffle sampler. After crushing to 9.0 > 1.5 eqv/kg DW < 3,000 mg/kg DW

Cadmium

< 10 mg/kg DW

Mercury

< 0.5 mg/kg DW The distance to drinking water wells must be 20 m or more The ash must be placed above the highest groundwater table The maximum average thickness of the ash layer is 1 m, and the thickness of the ash layer must not exceed 2 m.

If the ash is to be used in unpaved single-lane roads and unpaved squares (maximum surface area of 2000 m2), the following additional requirements must be met: The distance to drinking water wells must be 20 m or more The thickness of the ash layer must not exceed 0.30 m. If a utilisation project of this nature involves less than 100 tons of ash of approved quality, it may proceed without any permit as long as the above conditions are met. If an ash utilisation project involves more than 100 tons but less than 30,000 tons of ash, a detailed description of the project must be submitted to the local authorities (county and municipality) in advance for approval. The applicant must then wait for 4 weeks. If he receives no (negative) reply within the 4 week period, the project is approved as submitted. Each county council may refuse the project if it is in conflict with environmental protection considerations or may ask the applicant to change the project or to provide an environmental impact assessment before resubmitting the project. Large scale applications of more than 30,000 tons of ash will usually covered by the legislation on disposal (Figure 21.1). When the ash is used a sub-base in road construction, it must comply with the following additional performance related conditions set by the Danish Highway Department: The bottom ash must not be mixed with fly ash (or any other materials); The bottom ash must be quenched immediately, and it must be stored for at least 1 month prior to utilisation; and, The bottom ash must be screened to maximum particle size of 50 mm, contain less than 9 percent (w/w) of fines below 0.075 mm, the loss on ignition (at 1000~ must be less than 10 percent (w/w), and the content of water must be between 17 and 25 percent.

905

21.3.2 Germany Guidelines for road construction have been developed (Hoesel, 1986), including options for the utilisation of bottom ash (grate ash only or grate ash combined with boiler ash) in the road surface (with or without binder), and use in road base and fill in areas such as parking lots, promenades, noise protection walls, etc. These guidelines serve as the basis for subsequent regulations. The raw ash has to be stored a minimum of 3 months to reduce water content (initially about 30%) and allow swelling to occur. Sieving and ferrous removal prior to utilisation is also required. The following properties are specified for bottom ash: grain size < 32 mm; splintering during freeze-thaw testing between 0.5-8.5 wt.%; proctor density of 1.5-1.9 Mg/m 3 at 11-18% moisture content; LOI < 5%; and pH of 8 - 12 in water. Bottom ash use should be at least 1 m above the groundwater table. In water quality protection areas, additional requirements recommended were: pure metals < 5%; unburnt material < 0.5%; LOI < 5%; particle size < 0.063 mm < 7%; soluble matter < 2%; leaching parameters based on DEV $4 test to include pH, conductivity, CI, sulphate, EOX, TOC, Pb, Cr, Cd, Cu, Ni, Zn. Later guidelines prohibited the use of ash in water quality protection areas. Monthly monitoring of ash quality is also recommended. Individual German states have issued regulations for ash utilisation. For example, in Hassia, bottom ash has to be pretreated according to the guidelines presented by Hoesel (1986) along with additional specifications (Hessisches, 1988). Ash must be aged for more than 2 months and have an LOI < 2%. Using the German standard leaching test (DEV $4), solubility measured as solid residue of evaporation must be < 1%, and limits for ions are (mg/I) NH4 0.4, CI 250, S04 600, F 3, Pb 0.1, Cd 0.004, Cr 0.04, Cu 0.5, Ni 0.04, Zn 0.5, and Hg 0.001. The moisture content, pH and conductivity must be recorded. Every 2 years a PCDD/PCDF analysis is required but no limit is provided.

21.3.3 The Netherlands The Netherlands currently has the most extensive framework proposed for utilisation of waste materials including incinerator residues. Management of these residues is regulated under the general framework established for solid wastes including dredge spoils, construction debris and other industrial and combustion residues (Eikelboom, 1992). The philosophical basis for the regulatory framework includes lifecycle management, (ii) marginal environmental burdening and (iii) user acceptance. The goal of lifecycle management is to maintain or modify the physical and environmental properties of residues to achieve the highest quality practical for recycling as granular construction materials, as many successive times as possible. The goal of marginal environmental burdening is to establish incremental increases in ambient soil and water

906 contaminant concentrations below which environmental impacts are negligible or acceptable. The goal of user acceptance is to allow routine residue utilisation in environmentally acceptable applications with public and product user confidence. These goals have resulted in the development of a detailed set of regulations and supporting research. The key aspects of the regulatory framework are: 9

Classification of waste substances and building materials, Establishment of target values for soils (including groundwater) and surface water, Establishment of standardised leaching tests, composition requirements and evaluation protocols for building materials, and

9

Certification of residues for use

Classification of waste substances and building materials is based on whether the materials are granular or monolithic, the use of additional emission controls (e.g., covers, liners, etc.), and the degree of contact with water. The target values for soils and groundwater are presented in Table 21.3. They were established based on a survey of soil and groundwater quality within the country. These target values then were used to determine acceptable marginal burdening levels for contaminants released from residues during use (Table 21.4). Marginal burdening is defined as "an increase of 1% in the level of pollution in relation to the target values in 1 meter of soil over 100 years." Soil composition was assumed to include 10% humus and 25% lutite. The assumed soil composition was used to estimate the distribution contaminants between assimilation by the soil and transport to groundwater. This approach indicated that marginal burdening of soil was also protective of groundwater. Contaminant release limits for building materials incorporating wastes are based on standardised column leach test for granular materials (NEN 7343) and a monolith leach test (NEN 7345) for molded construction materials (Aalbers, 1992). Detailed structured assumptions and extrapolations based on the physical structure of the material, application and leaching mechanisms have been developed to permit development of limits for laboratory tests. Utilisation of monolithic and granular wastes is classified based on contact with water and whether or not additional barriers (e.g., liners and covers) are employed: Type A -

Submerged or always in contact with water

Type B -

Primary contact with water is from precipitation (estimated contact with water 14% of time)

907 Table 21.3 Target Values for Soil and Groundwater Quality in The Netherlands Substance Ground Groundwater Surface Water mg/kg

pg/I

IJg/I

Cr

100

1

5

Co

20

20

NA

Ni

35

15

9

Cu

36

15

3

Zn

140

65

9

As

29

10

5

Mo

10

5

NA

Cd

0,8

0,4

0,05

Sb

(2.6)

NA

NA

Se

(1)

NA

(10)

Sn

20

10

NA

Ba

200

50

(200)

Hg

0,3

0,05

0,02

Pb

85

15

4

V

(68)

NA

NA

F

500

500

1500

CN-complex

5

10

NA

CN-free

1

5

(50)

(500)

150,000

100,000

20

300

8,000

S O4 Br

CI (200) 100,000 200,000 values from "Beleidstandpunt Over De Notitie Milbowa" (0.5-0.2-1992 with the exception of the values, which are in brackets). the values of chloride, fluoride, bromide and sulphate in surface water are limit values. NA = not available Sb, Se, and V from [13] Aalbers, 1992 -

908 Table 21.4 Maxirnum Acceptable Marginal Burdening Levels for Contaminants Released from Residues During Use in The Netherlands Substance

Ground

Groundwater

max. accept, immersions mg/m2 per 100 year

max. accept, immersion mg/m2 per 1 year

As

400

Ba

3000

Cd

10

Co

300

Cr-tot

1500

Cu

500

Hg

4

Mo

150

Ni

500

Pb

1000

Sb

35

Se

15

Sn

300

V

950

Zn

2000

Br

300

CI F

30000 7000

SO4

45000

CN-tot

70

CN-free

15

909 V1 building materials-

application without additional emission controls

V2 building materials-

application with additional emission controls (e.g., covers, liners, etc.)

The relationship between the maximum allowable release and the release observed during the standardised monolith leaching test is described by: where:

Im,x(J yr)= Em,• Imax(J yr)=maximum acceptable emission into the ground in a period of J year (mg/m2); Em~(64d)=maximum acceptable emission out of a material determined with the tank leaching test in 64 days (mg/m2); f~xt=extrapolation factor for the extrapolation from Em,~(64d) to Em~(J yr); fbev=Correction factor for wetting period; f~=correction factor for the difference between the laboratory temperature and the temperature in the field; f~so=isolation factor for V2 building materials (this factor is 1 for V1 materials).

The criteria for determining if diffusion is the controlling release mechanism is based on the slope of the cumulative release curve from the monolith leach test: Slope > 0.6-Emission is not diffusion controlled (more rapid) and the column leach test should be applied 0.35 < Slope < 0.6-Emission is diffusion controlled; Slope 12-1ow mobility 10.5 < pDe < 12-intermediate mobility pD~ < 10.5-high mobility

910 If the pDe is less than 10 for a species, depletion of that species can be anticipated to occur during the 100 yr assumed use interval. The release is estimated based on one dimensional diffusion from a flat plate. The quantity available for release is estimated based on the availability leach test (NEN 7340). The effects of weathering on the rate of diffusion are assumed to decrease the release rate. This implies physical durability and the occurrence of weathering mechanisms, such as carbonate uptake, which occur over the 100 year assumed use interval. The correction to account for this effect was to assume an increase in the PDe of 0.01/yr. The cumulative effects of the above derivation are summed in an overall extrapolation factor or multiplier to translate release from 64 day tank leaching results to a 100 yr release interval (Tables 21.5 and 21.6). Table 21.5 Extrapolation Factors for Determining Release in the Field from Laboratory Leaching Results for Cases where Depletion is Anticipated (The Netherlands) Layer thickness (m):

0.3

0.5

0.7

1.0

2.0

10

fuit

f,~

f,,t

fuit

fu,t

8 2

2

3

5

10

24

9 5

8

11

16

23

24

10 15

21

23

24

24

24

pD~ f,,

f,, : E(100yr)/E(64d)

Table 21.6 Extrapolation Factors for Determining Release in the Field from Laboratory Leaching Results with Geometric Considerations, Depletion and Assumed Effects due to Weathering Included (The Netherlands) Layer thickness (m): pO~

0.3

0.5

0.7

1.0

2.0

10

f,~

f,~t

f,~t

f,,~

.f,,~t

f,,~t

8

2

2

3

5

10

15

9

5

8

11

15

15

15

15

15

15

15

15

15

>10 fe~t= E(100yr)/E(64d)

Chloride and sulphate are considered exceptions to the above extrapolation factors because the maximum acceptable emission is defined to be that which would be released over a period of one year. This recognises the limited natural attenuation which occurs for these species and the potential direct impact on groundwater resources. The extrapolation factor, fe~, for chloride and sulphate is 2.4. The other

911 factors for translating release from tank leaching results to field conditions are to correct for differences in the frequency of wetting, temperature and isolation (e.g., barriers such as liners and covers). The correction factor, fb,,v, for always immersed conditions is 1.0 while for non-immersed applications it is 0.14. The factor of 0.14 is based on an average precipitation occurrence, and hence surface wetting, of 14% of the time in The Netherlands. While this is used as a multiplier to emission, a more appropriate approach would be to modify the release time interval using (0.14) ~ The correction factor for temperature, from,was based on an Arrhenius approach to diffusion kinetics and a mean laboratory and field temperatures of 20 C and 10 C, respectively. This resulted in ft~ equal to 0.7. The correction factor for isolation, fiso, was assumed to be 0.14 when addition barriers were employed and 1.0 when the material was directly exposed. Inversion of the above derivation permits estimation of acceptable laboratory results based on defined field conditions as Emax(64d) = Imax(100yr)/(fext*ft~m*fbev*fi,o) A similar laboratory to field extrapolation approach was applied for granular materials. In this case, the principal mechanism of contaminant release is assumed to be by percolation through the granular material. This is in contrast to when a granular material compacted in place results in a low permeability layer which may be treated as a monolith or diffusion controlled leaching. Examples of typical limit values for leaching test results on monolithic and granular materials are presented in Table 21.7. Since 1995, the limit value on composition $1 has been withdrawn following discussions between regulators and industry. 21.3.4 Sweden

In Sweden, both the utilisation and disposal of residues are treated in the same manner under The Environmental Protection Act (Hartl~n, 1989 and 1991; F~llman, 1992). Thus, individual regulatory reviews including local and regional authorities are required for each specific application (Figure 21.2). The general philosophy in evaluating utilisation applications is that the specific utilisation scenario should result in (i) improvement of general environmental conditions, and (ii) have less environmental impact than disposal. Examples of uses that may satisfy these criteria are use as cover in a municipal waste landfill, in road paving applications where the residue is covered with an asphalt layer, or where natural aggregate is in limited supply. To date, very little utilisation has occurred, and utilisation regulations are under review. 21.3.5 United States

The United States currently does not have national standards for the utilisation of residues. Requirements for the U.S. Environmental Protection Agency to develop

912 criteria for ash utilisation are being considered in pending legislation. In the absence of national guidelines, several states have developed applicable regulatory requirements. Florida and New York are the two states that have the most extensive requirements.

Table 21.7 Maximum Acceptable Limits for Leaching Test Results on Granular (mg/kg) and Monolithic (mg/m2) Materials (proposed, The Netherlands) Leaching

standard values oBB granular materials in mcj/kg

Substance

U1

U2

$1

U1

U2

$1

0.30

3.0

375

25

125

750

As

standard values oBB products in mg/m 2

Ba

4.0

40

7500

350

1750

15000

Cd

0.010

0.10

10

0.70

3.5

20

Co

0.20

2.0

250

15

75

500

Cr

1.0

10

1250

90

450

2500

Cu

0.35

4.0

375

30

150

750

Hg

0.010

0.050

5

0.30

1.5

10

20

250

Mo

0.050

0.50

125

4.0

Ni

0.35

4.0

250

30

150

500

Pb

0.80

8.0

1250

75

375

2500

Sb

0.030

0.30

50

2.5

13

100

Se

0.020

0.20

50

1.8

9.0

100

Sn

0.20

2.0

250

20

100

500

V

0.70

7.0

1250

60

300

2500

Zn

1.4

14

1250

125

625

2500

Br

0.20

2.0

500

20

100

1000

600

5000

5000

2250

11250

1000000

CN- complex

0.050

0.50

125

4.5

23

250

CN-free

0.010

0.10

25

0.90

4.5

50

5.0

50

4500

CI

F

SO 4 750 10000 25000 A = A-type application" B = B-type application

440

2200

9000

15000

45000

40000

913 Figure 21.2 Regulatory Reviews Required for Specific Applications of Ash Utilisation in Sweden APPLICANT

FRANCHISE BOARD

PLANNING I

COUNTY COUNCIL

LOCAL AUTHORITY

ENVIRONMENTAL PROTECTION BOARD

OTHER PARTIES

INFORMATIONAND PRELIMINARYDISCUSSIONS

APPLICATION TREATMENT

COMPLETION

CIRCUI.A,TION FOR COMMENT

, I

PUBLIC HEARING DECISION, TERMS OF CONCESSION

INSPECTION L OPERATION

r

io

I INSPECTION '~ PROGRAMME

,._sPECT,om

INSPECTION PROGRAMME INSPECTION

INSPECTION

Florida requires an ash management plan as part of the operating permit for an incinerator facility. These plans must be updated and reviewed at least once every five years (Florida reg. 17-702.400), and must address the methods, equipment and structures needed to control dispersion of ash during handling, processing, storage, loading, transportation, unloading and disposal. The plan must consider potential pathways for human and environmental exposure including inhalation and direct contact (human exposure) and migration to soil, groundwater and surface water (environmental exposure). Recycling of ash (utilisation) is explicitly discussed in the regulations (Florida reg. 17-702.600). The generator of ash, at least monthly must describe the chemical and physical properties of the ash which is to be recycled. Prior to the recycling of ash, the process and use of the ash must be shown not to cause discharges of pollutants to the environment. In addition, in order to utilise ash, the following steps must be completed: describe chemical and physical properties of the finished product line; identify the quantity of ash used in the product; identify the quantity of product to be marketed or used; demonstrate the process will physically or chemically change the ash residue so that any leachates produced after processing will not cause a violation of surface or ground water standards;

914 demonstrate ash or products will not endanger human health or the environment; performance standards need to be established as well as operational criteria. New York also requires the development of an ash management plan for each incinerator facility (New York reg. section 360-3.5). Ash generation, handling, storage, testing, transportation, treatment, and disposal or beneficial use plan must be included. Ash utilisation is regulated by a beneficial use petition. The party who desires to beneficially use ash must petition to utilise the ash. There is no permit involved directly with ash utilisation. As a result of no permit being required, there are no public hearings required. The party petitioning for ash to be used as an ingredient or as a substitute for a raw material must: demonstrate that the resulting material is not a waste requiring disposal; have a known market or disposition; 9

not accumulate the material speculatively; have contractual arrangements with a second person for use as an ingredient and this person has to have the equipment to do so; chemically and physically characterise the ash; identify the quantity and quality to be marketed; describe the proposed method of application or use, available markets and marketing agreements; demonstrate that the intended use will not adversely affect the public health, safety, welfare and environment; provide a description of each product mixture, if the use of the ash includes the mixing with different types of materials.

21.4 TECHNICAL REQUIREMENTS The three major categories of applications include use as a lightweight aggregate either for road base construction, as a fill for embankments, or as an amendment for Portland cement concrete or bituminous asphalt. Most physical utilisation criteria are based on standard engineering tests. Specific physical requirements will vary based on the type of application (e.g., asphalt pavement, cement concrete, structural fill, etc.) and local construction regulations. Table 21.8 provides a summary of the physical testing

915 requirements that may be required and typical acceptable values. Many of the tests traditionally specified for construction may not be directly applicable to bottom ash testing. However, they may be necessary for market acceptance. Most frequently, bottom ash can be blended with other aggregates to achieve specific design criteria.

Table 21.8 Physical Criteria and Property Ranges for Utilisation of Bottom Ash Requirement

Asphalt Pavement

Concrete

Particle Size Distribution

Specific 9 to location & application design Uniformity 9 coefficient (d60/dl0) can be specified total 9 content of fines (1 year (inert waste). Corresponding calculations for surface water show that they have little barrier effect.

953

22.4.3 Caps and Top Covers The capping of a landfill serves the dual purposes of modifying the infiltration of precipitation into the landfill and isolating the waste from the surroundings. The final capping of landfills with low permeability materials, such as clays, can greatly reduce the rate of leachate production. It is general practice to complete landfill sites with a layer of compacted clay or an artificial membrane, followed by soil and/or subsoil to support vegetation. In some cases (e.g. for MSW landfills) a gas drainage layer must be placed underneath the cap. Sloping surfaces, surface drainage systems and capillary barriers may be used to help reduce the infiltration of precipitation into the landfill if this is the part of the disposal strategy. It should be noted that synthetic cap materials, although impermeable in principle, may not be durable over a longer period of time. Furthermore, unlike most clay caps, a synthetic cap may (over a certain period of time) entirely eliminate infiltration, but it does not allow controlled infiltration at a predetermined rate which may be called for by the disposal strategy (e.g. leachate containment and collection in conjunction with waste mineralisation processes within the landfill or controlled contaminant release). A European review of clay cap performance (Knox, 1991) produced the following conclusions: Percolation through clay caps may range from 0 to 200 mm/annum depending on its quality and on materials and drainage arrangements above the clay layer. In many parts of Europe this represents a large reduction compared to effective rainfall. The performance of the cap is determined as much by what is put on top of it, as by the quality of the barrier layer itself. Percolation through the cap is extremely dependent on the efficiency of lateral drainage above it. For a given amount of effective rainfall, percolation is minimised when lateral drainage is maximised. To achieve a low percentage percolation, the ratio of hydraulic conductivity in the cap to that in the soil or drainage layer above it should be no greater than 104. The distance between field drains should be no greater than 20 m. Desiccation cracking of caps may lead to a large increase in percolation. To prevent desiccation of a cap placing sufficient soil or other material on top of it is necessary. Typically, at least 0.9 m of soil or subsoil may be needed under northern European conditions. In more arid climates greater depths may be needed to counteract desiccation.

22.4.4 Geotechnical Stability The geotechnical stability of the landfilled material (e.g. bearing capacity and slope stability) must in general be sufficiently high to accommodate trucks and various types

954 of waste moving and compacting equipment. In the longer term, it must be ensured that settlements, particularly differential settlements, do not hamper or interfere with the intended performance of liners, caps and drainage systems.

22.4.5 Abatement of Noise, Odour and Fugitive Dust Problems Proper precautions must be taken to minimise noise, odour and fugitive dust problems during the operation of a landfill. Keeping the residues moist or contained are among the measures used to avoid fugitive dust problems at residue landfills. In traditional landfilling, it is common practice to apply a daily cover of topsoil to the landfilled waste in order to minimise problems with odour, dust and vermin. Depending on the quality of the soil used, this may have the undesirable effect of creating a number of hydraulically isolated waste cells within the landfill.

22.4.6 Monitoring of Leachate Quantity and Quality Both the quantity and the quality of the leachate generated at any landfill equipped with leachate collection systems should be monitored during the period of operation and during the postclosure period until the waste in the landfill has reached final storage quality. Climatic data (e.g. precipitation and temperature) should also be monitored. The main objectives of monitoring the quantity of leachate produced at a landfill are 1): to evaluate through water balance calculations whether the leachate collection system is functioning as intended; 2): to check the efficiency of any systems intended to modify the rate of production of leachate and to obtain information on the actual rate of production of leachate at the landfill in question, thus enabling proper planning of leachate management; and 3): to provide information on the accumulated amount of leachate produced at a given landfill and hence allow an evaluation of the attained degree of leaching of the waste. The main objectives of monitoring the quality of leachate produced at a landfill are 1): to provide a basis for the selection of indicator parameters for the monitoring of groundwater (and surface water) monitoring; 2): to ensure that the leachate quality complies with the criteria in relation to on-site management and direct discharge, onsite treatment or treatment at a municipal wastewater treatment plant; and 3): to provide information on the progress of the waste stabilisation processes occurring within the landfill and, eventually, to provide background for an assessment of whether or not the waste has reached final storage quality. The leachate sampling frequency should be higher during the period of landfill operation than during the subsequent postclosure period. The analytical programme will depend on the type of waste present in the landfill and possibly also on the perceived environmental risks. A programme may consist of a routine analytical

-~

955 programme with relatively few parameters carried out on leachate samples collected relatively frequently and a more comprehensive extended analytical programme which is carried out on samples collected at longer time intervals. For instance, for leachates from incinerator residues, the simple analytical programme could comprise the determination of pH, conductivity, total dissolved solids (TDS), nonvolatile organic carbon (NVOC), chloride, sulphate, alkalinity and temperature. In addition to this, the extended programme could include measurement of adsorbable organic halogen (AOX), sulphide, total N, ammonia N, NOx N, Na, K, Ca, Mg, Fe, Pb, Cu, Cr, Hg, Ni, Zn and possibly As and Mo. The sampling point should be chosen carefully and the sampling procedure should allow for correct measurement of sensitive leachate quality parameters such as pH, redox potential, sulphide content, etc.

22.4.7 Monitoring of Groundwater and Surface Water Quality The main objectives of groundwater quality monitoring at landfills are 1): to provide information on background levels and natural variations of downstream groundwater quality prior to operation of a landfill; 2): to detect at the earliest possible time any unintended leakage of leachate from a lined landfill into the aquifer; and 3): to ensure that the anticipated impact on the aquifer is not exceeded if a controlled contaminant release strategy is applied to a landfill. The implementation of a groundwater monitoring programme which is highly dependent on local conditions includes the performance of geological and hydrogeological surveys to determine how many monitoring wells are needed, where they should be located and how they should be designed to ensure maximum likelihood of detecting any leachate plume at an early stage. It also includes the installation of such monitoring wells and implementation of proper groundwater sampling procedures. A detailed discussion of these subjects is, however, beyond the scope of this study. Several books and manuals are available on these issues. An overview is provided by Christensen et al. (1992). The analytical requirements for a groundwater monitoring programme should be based upon information on the compOsition of the leachate from the landfill in question and the background quality of the groundwater. The monitoring parameters should generally be chosen among components which are present in the leachate at significant concentrations, which are relatively mobile in the aquifer, and/or which are present in the groundwater at low background concentrations. The latter condition facilitates the detection of a leachate plume due to a large relative concentration contrast between contaminated and uncontaminated groundwater. In order to establish background levels, and the natural or seasonal variations of the groundwater to be monitored it is proposed that the samples from the monitoring wells are analysed 4 times a year for 2 years prior to commencement of landfill operation. The analytical programme carried out during this period should correspond to the

956 extended programme for leachate described in Section 22.4.6. When this has been accomplished, both the sampling frequency and the analytical programme should change. The sampling frequency should be high enough to ensure that a leachate plume will be detected before it has migrated significantly past a monitoring well. The sampling frequency should also be low enough to ensure that the water that is drawn into the well from the aquifer has been replaced between sampling events by water flowing from upstream. The sampling frequency will thus depend on the rate of flow of the groundwater which must be determined for each individual landfill site. It is recommended, however, that a groundwater sampling frequency of not less than once a year is adopted in any case. An analytical programme could consist of a basic set of parameters (pH, conductivity, TDS, NVOC, chloride, sulphate, ammonia, Na, K and Ca) which for each individual landfill may be supplemented with further parameters. The basic parameters are designed to provide an early warning for components which are present in most incinerator residue leachates. The supplementary parameters must be selected for each individual landfill based on knowledge of the actual waste accepted, its leaching properties, the composition of the leachate and the mobility and background values of various leachate components in the aquifer. Hjelmar et al. (1988) report that very early signs of a plume of bottom and fly ash leachate in an aquifer were observed as continuously rising concentrations of chloride and calcium (due to ion exchange between Na and K in the leachate and Ca in the soil) and other mobile salts in groundwater from a downstream monitoring well. The monitoring of the groundwater quality should continue during the postclosure period, possible at a reduced sampling frequency, until the waste in the landfill has reached final storage quality. The main objectives of surface water quality monitoring at landfills are 1): to provide information on background levels and natural variation of the quality of surface waters to which leachate is directly or indirectly discharged; and 2): to detect any unintended impact of leachate on a surface water system and to ensure that the anticipated impact on the surface water system is not exceeded if a controlled contaminant release strategy is applied to a landfill. It should be noted that monitoring of surface water systems generally is more difficult and much less likely to produce useful results than monitoring of groundwater. Surface water systems are usually much more diverse and dependent on local conditions than aquifers. There may be none or several surface water bodies downstream of a landfill and within the same catchment area or close enough to be affected by spillages of leachate. With the exception of stagnant waters, most surface waters systems generally have much higher dilution potentials than aquifers do. This means that leachate which may be discharged directly, may be removed rapidly or diluted by the

957 surface water to the extent that direct detection by chemical analysis is difficult or impossible. In cases where surface water quality monitoring is found desirable, the design of the chemical analytical programme may be based on the same principles and include similar ranges of analytical parameters as for groundwater while taking into account the nature of the surface water body in question. Analytical parameters such as chloride, sulphate, Na, K, Ca and Mg are, of course, irrelevant as contamination indicators in marine surface water bodies. Due to the low concentration levels of contaminants which must be expected for affected surface water systems, the monitoring programme may in certain cases alternatively be based on eutrophication indicators (e.g. chlorophyll-a) or on biological monitoring (in marine waters, e.g. placement and subsequent analysis of mussels). 22.4.8 Leachate Treatment

The most common leachate disposal route for landfills with active leachate removal systems is to sewer and subsequently to a biological wastewater treatment plant without any pretreatment. Biological treatment only affects the content of biodegradable organic contaminants (and perhaps ammonia) in the leachate. Leachates with high organic Ioadings may require aerobic biological pretreatment to remove organics and ammonia prior to discharge to sewer or surface water. Incinerator residue leachates, however, normally have relatively low contents of organics, but may have high contents of inorganic salts (e.g. CI, SO42, Na§ K§ Ca 2§ and variable contents of trace elements/heavy metals. Such leachates may be treated/pretreated by pH and/or redox potential adjustment, precipitation (e.g. with TMT)/filtration and/or adsorption on activated carbon to remove dissolved trace metals, suspended solids and non-degradable organics prior to discharge to sewer or surface water bodies. Preconcentration or removal of salts may be accomplished by reverse osmosis or evaporation, although this may not be advisable for this type of leachate from an economic or a life cycle perspective (Hjelmar et al., 1995). In most cases, biological wastewater treatment will have little effect on incinerator residue leachates dominated by inorganic constituents, but discharge to sewer may be convenient. If such leachates have very low contents of trace elements and organics, they are compatible with seawater and may be discharged directly into marine surface water bodies without pretreatment. Leachates from landfilling of poorly combusted residues may have high contents of organic material and may require (and benefit from) biological wastewater treatment. 22.5 DISPOSAL PRACTICES

Data have been compiled for a number of countries regarding present and future disposal practices for incinerator residues. In almost all jurisdictions, regulations have

958 been or are presently undergoing modifications. In Europe, for instance, a number of countries will have to modify regulations to come into compliance with an anticipated European Union (EU) waste disposal directive (European Union, 1995) and other EU waste management legislation. In the US, the Supreme Court ruling in May 1994 which lifted the exemption of MSW incinerator ash from being tested as a potentially hazardous waste may give rise to changes in some states (Supreme Court of the United States, 1994). Consequently, the situation is somewhat dynamic, and since the collected information may be outdated fairly rapidly, only a brief, general outline of the current disposal practices in some European and North American countries is presented in this section. Disposal practices for MSW incinerator residues vary widely across the world. Substantial variations in disposal practices are also found within countries consisting of federations of states or provinces such as US, Canada and Germany. Table 22.8 summarises the disposal strategies corresponding to the landfilling practices and/or policies of some countries. Table 22.8 MSW Incinerator Residue Disposal Practices in Various Countries Disposal strategy Bottom ash APC residues Combined ash Total Germany Canada US containment/dry Denmark (d) tomb The Netherlands (d) .. Germany Leachate Denmark Denmark (d) US containment and France France (a) collection Germany Germany (wet Sweden scrubber) Switzerland Sweden (a) The Netherlands Switzerland(a) Controlled Sweden Sweden (a) contaminant Denmark (b,c) release Unrestricted Canada contaminant release (a): Residue treatment or stabilisation required (b) Past practice (c) Planned future practice (pending legislation) (d) Temporary practice (awaiting improved treatment and disposal technology) Monofill disposal represents the most common method of bottom ash disposal in Europe and Canada, although co-disposal of incinerator residues with other wastes,

959 including MSW, does occur. The strategies employed for disposal of bottom ash cover the entire range from total containment over leachate containment and collection and controlled contaminant release to unrestricted contaminant release. At present, however, the prevailing disposal strategy for bottom ash is containment with some type of leachate collection. In Canada and several European countries, APC system residues are listed as hazardous or special wastes, and disposal at highly engineered monofills with extensive environmental protection systems corresponding to total containment or leachate containment and collection is normally required, often in conjunction with treatment/stabilisation of the residues. In France, for instance, stabilisation of APC residues is required prior to disposal at hazardous waste landfills, and in Germany most of the residues from dry/semi-dry APC processes are placed in underground storage in old salt mines. At one disposal site in Sweden, residue from the dry APC process is stabilised with approximately 30% (w/w) of a special type of cement; the material (37% residue, 16% cement and 47% water) is poured onto the site as a slurry in 2.5 m deep layers (cells) and allowed to cure. The median concentration of chloride in the leachate which consists primarily of surface runoff water is approximately 20,000 mg/I and the median concentration of lead is approximately 0.06 mg/I over a period from 2 to 6 years after the commencement of disposal (Sundberg and Tuutti, 1994). In the US, most of the incinerator residues generated are managed as combined ash. The combined ash is frequently disposed in landfills with relatively stringent design standards for leachate containment and collection, or total containment strategies. Most regulations evolve from the State level, provided Federal criteria are met. As of 1991,25 states required monofill disposal, 5 states allowed for monocells inside MSW landfills, and 16 states allowed codisposal of incinerator ash and raw MSW. According to the above mentioned Supreme Court decision, all MSW incinerator ash in the US must now be tested using the TCLP test (see Chapters 14 and 16). If the ash passes the requirements, it may be disposed in a Subtitle D landfill for nonhazardous waste. Conversely, if it fails, it must be placed in Subtitle C landfills for hazardous waste (as specified in the Resource and Recovery Act). The ash may be treated and combined prior to testing.

22.6

DISPOSAL RECOMMENDATIONS FOR INCINERATOR RESIDUES

Based on the data presented on incinerator residue leaching characteristics and the discussion of disposal strategy, a number of conclusions may be drawn concerning the feasibility of various disposal and leachate management options. The disposal strategies applicable to mineral wastes such as incinerator residues and organic waste types are, as previously discussed, very different and generally incompatible, both in the short and long-term. Co-disposal of incinerator residues and raw MSW is therefore generally not advisable. Different types of incinerator residues

960 may also exhibit significant differences in behaviour when landfilled, and separate management and disposal of, e.g. bottom ash and APC residues, is therefore recommended, since the opportunity of applying different disposal strategies to different types of incinerator residues when appropriate is lost if bottom ash and APC system residues are combined. This recommendation differs from the current practice in the US, where disposal of combined ash is frequently managed using very stringent landfill design standards (e.g. in double lined monofills). While this practice generally is environmentally protective, at least in a short-term perspective, it does require perpetual maintenance and can be costly, especially since the bulk of the combined ash (bottom ash) may not require the same level of perpetual care as a combined ash or APC residue fill. Overall, the optimal disposal strategy for the various incinerator residues may be considerably different from the traditional disposal strategies applied to raw MSW. The most feasible disposal options for bottom ash, APC residues and combined ash are briefly discussed below and summarised in Table 22.9. Table 22.9 Summary of Incinerator Residue Options and Recommendations Disposal strategy option

Bottom ash

APC..residue Combined ash

Total containment/dry No Possibly No tomb (e.g. salt mines) Leachate containment and Yes (a) Yes (a) Yes (a) collection Controlled contaminant Yes (b) Maybe (b) Maybe (c) release Unrestricted contaminant No (c) No (c) No (c) release (a) If requirements for controlled contaminant release are not met (e.g. as a first stage of disposal). (b) If requirements are met. May require prior or in-situ treatment of the residues or may be second stage of disposal. (c) Only after final storage quality criteria are met. 22.6.1 Bottom Ash

With proper siting (e.g. close to the sea or in an area without vulnerable aquifers), a disposal strategy based on controlled contaminant release seems appropriate and should be pursued for landfilling of bottom ash. Pretreatment (e.g. washing or stabilisation) or an initial disposal stage entailing containment, collection and treatment of leachate may in several cases be required (Belevi et al., 1992). The possibilities for controlling the geochemical and biogeochemical conditions within a bottom ash landfill

961 through ash quality requirements, ash treatment and landfill design should be investigated further. The rate of leachate production may be controlled partly through the design of the landfill. The construction of any bottom ash disposal site based on a controlled contaminant release strategy must be preceded by a thorough environmental impact assessment which ensures that the rate of release of contaminants into the surrounding environment will not exceed an acceptable limit, neither in the short nor long-term. If, for some reason, a solution requiring containment and collection of leachate is chosen, either temporarily or indefinitely, it becomes necessary to manage and dispose of the leachate. Leachate from bottom ash is generally accepted at wastewater treatment plants as long as it does not constitute a major proportion of the total input to the facility. It has in some cases been necessary to reduce the pH of the leachate (by addition of sulphuric acid) and/or to elevate the redox potential from a reducing level to an oxidised level (e.g. by addition of hydrogen peroxide) prior to treatment at a wastewater treatment plant. In most cases no pretreatment has been necessary. Dilution is practically the only beneficial effect of biological wastewater treatment on bottom ash leachate containing mostly inorganic salts and little or no organic degradable matter.

22.6.2 APC Residues (Fly Ash and Acid Gas Scrubbing Residues) A sustainable disposal solution for the APC residues, particularly fly ash and residues from the dry/semi-dry acid gas scrubbing processes, must eventually be based on a controlled contaminant release strategy and will almost certainly require extensive pretreatment of the residues. A two-stage treatment process involving removal and possibly recovery of the soluble salts (washing/extraction) followed by stabilisation, vitrification or fixation of the remnant may be appropriate for this purpose (Hjelmar, 1992). Considerable efforts are currently being spent on the development of such processes. In the meantime, disposal of APC residues must generally be based on less sustainable strategies involving total containment/entombment or containment and collection of leachate. The same requirements concerning proper siting and design of a landfill and performance of an environmental impact assessment as mentioned above for bottom ash apply to a controlled contaminant release disposal strategy for APC residues. The leachate produced at APC residue disposal sites based on containment and collection of leachate generally has a high concentration of inorganic salts and in some cases also relatively high concentrations of trace elements, particularly Pb and Cd. Such leachate is often accepted at municipal wastewater treatment plants without prior treatment, provided it does not constitute a major proportion of the total input to the plant. As mentioned for bottom ash leachate, the only beneficial effect of such a treatment is dilution. Leachates with a high content of heavy metals may have an adverse effect on the sludge from a biological treatment plant. If necessary, the

962 concentration of several trace elements in the leachate (e.g. Cd and Pb) may be reduced substantially by subjecting the leachate to pretreatment including adjustment of pH and sedimentation/flocculation with TMT. This treatment may be relatively expensive if large amounts of leachate are produced. Removal of the inorganic salts from the leachate (e.g. by evaporation) is not economically feasible or environmentally desirable under most circumstances. 22.6.3 Combined Ash

Although separate management and disposal of the different residue streams are believed to be technically and economically advantageous, both in the short and longterm, combined ash is still generated in the US. In principle, the disposal requirements for combined ash are similar to those described for APC residues. The proportion of residue requiring relatively stringent environmental protection measures when landfilled is increased substantially by the mixing of APC residue and bottom ash which also precludes utilisation and renders pretreatment or in-situ treatment of the residue more difficult and less efficient than it would be for separate ash streams. REFERENCES

Andersen, L. & J. Boll. "Leaching of APC System Residues from MSW Incinerators, Pilot Scale Experiments", In: Udvas.kning fra sla.q~er, lord o.q afraid, ATV-K0miteen vedrc~rende ,qrundva.ndsforurening, Lyngby, Danmark, pp. 85-106 (in Danish), 1994. Belevi, H. and P. Baccini. "Long-Term Behaviour of Municipal Solid Waste Landfills", Waste Management & Research, 7:43-56, 1989. Belevi, H., D.M. Stampfli and P. Baccini. "Chemical Behaviour of Municipal Solid Waste Incinerator Bottom Ash in Monofills", Waste ManaQement & Research, 10153167, 1992. Cambotti, R.K. & H.K. Roffman, "Municipal Waste Combustion Ash and Leachate Characterization. Monofill- Fifth Year Study, Woodburn Monofill, Woodburn, Oregon", Report prepared by AWD Technologies, Pittsburgh, Pennsylvania, 1993. Christensen T.H., P. Kjeldsen and J.L.C. Jansen. "Groundwater Control Monitoring at Sanitary Landfills." In (Christense.n, Cossu and Ste.Qemann, eds.): Landfillin.q of.Waste: Leachate. Elsevier, London, 1992. Comans, R.N.J., H.A. van der Sloot and P.A. Bonouvrie. "Geochemical Reactions Controlling the Solubility of Major and Trace Elements During Leaching of Municipal Solid Waste Incinerator Residues", in.. Proceedings of....th.e.. 1993 International Conference on Municipal Waste Combustion, March 30 - April 2, 1993, Williamsburg, Virginia, USA, 1993.

963 European Union. "Directive on the Landfilling of Waste", Brussels, Belgium, 1995. Hjelmar, O. "Leachate from Incinerator Ash Disposal Sites", in ProceedinQs of the International Workshop on Municipal Waste Incineration, Montreal, Canada, October 1-2, 1987. Hjelmar, O. "Characterization of Leachate from Landfilled MSWI Ash", in Proceedinqs of the International Conference on Municipal Waste Combustion, Hollywood, Florida, April 11-14, 1989. Hjelmar, O. "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, O. "Municipal Solid Waste Incinerator Flue Gas Cleaning in Denmark: Residue Properties and Residue Management Options", in Proceedings of ISWA Specialized Conferenc_~..on Incineration and Biolo.qical Waste Treatment, Amsterdam, The Netherlands, September 1-3, 1992. Hjelmar, O. "Leaching Properties of Fly Ash from MSW Incinerators", Report for the National Agency for Environmental Protection, VKI Water Quality Institute, H~rsholm, Denmark, 1993 Hjelmar, O. "Disposal Strategies for Municipal Solid Waste Incineration Residues", J_ Haz. Mats., 47, pp. 345-368, 1996. Hjelmar, O., K.J. Andersen, J.B. Andersen, E.A. Hansen, A. Damborg, E. Bj~rnestad, A.H. Knap, C.B. Cook, S.B. Cook, J.A.K. Simmons, R.J. Jones, A.E. Murray, M.J. Lintrup, H. Schr~der, F.J. Roethel. "Assessment of the Environmental Impact of Incinerator Ash Disposal in Bermuda", Final Report, Prepared for Ministry of Works & Engineering, Hamilton, Bermuda, by the Water Quality Institute, H~rsholm, Denmark, 1993. Hjelmar, O., E. Aa. Hansen and A. Rokkjaer. "Groundwater Contamination from an Incinerator Ash and Household Waste Codisposal Site", In UNESCO Wor.kshop on Impact of Waste Disposal on Groundwater and Surface Wa.ter, Copenhagen, Denmark, 1988. \\

Hjelmar O., L.M., Johannessen, K. Knox, H.-J. Ehrig, J. Flyvbjerg, P. Winther and T.H. Christensen "Management and Composition of Leachate from Landfills", Final Report to the Commission of the European Communities, DGXl A.4, Waste '92, Contract No. B4-3040/013665/92, 1995. Johannessen, L.M., O. Hjelmar and J. Riemer. "A New Approach to Landfilling of Waste in Denmark", in Proceedin.qs of Sardinia "93, IV International Landfill Symposium, S. Margherita di Pula, Italy, 11-15 October 1993, 1993.

964 Knox, K. "A Review of Water Balance Methods and Their Application to Landfill in the UK". Report prepared for the UK Department of the Environment, DOE report No. CWM 03/91, 1992 Knox, K. "Control of Landfill Leachate", in Proceedin.qs of .the 8th International Conference: Water: Supply and Quality, Cork, Ireland, 1992. Kosson, D.S. Personal communication, 1995. Kosson, D.S., H.A. van der Sloot and T.T. Eighmy. "An Approach for Estimation of Contaminant Release During Utilization and Disposal of Municipal Waste Combustion Residues", J. Haz. Mats. 47, 1996. Lyons, M.R. "The WES-PHix Ash Treatment Process", Wheelabrator Environmental Systems Inc., Hamilton, NH, USA, 1995. Sundberg, J. and K. Tuutti. "Solidification of APC System Residue from H(~gdalenverket", Final Report 1994, Terratema ab, Link(~ping, Sweden (in Swedish), 1994 Supreme Court of the United States. "Syllabus: City of Chicago et al. v. Environmental Defense Fund et al.", Certiorari to the United States Court of Appeals for the Seventh Circuit. No. 92-1639. Argued January 19, 1994 - Decided May 2, 1994. Reimann, D.O. "Abwasserbehandlung aus M011verbrennungsanlagen", M011und Abfall, 19 (1):1-7, 1987. Thygesen, N., F. Larsen and O. Hjelmar. Environmental Risk Screenin.q of Utilization and Disposal of MSWI Bottom .Ash, Miljc~projekter 203, National Agency for Environmental Protection, Copenhagen, Denmark, (in Danish), 1992. van der Sloot, H.A., R.N.J. Comans, T.T. Eighmy and D.S. Kosson. "Interpretation of MSWI Residue Leaching Data in Relation to Utilization & Disposal", in Proceedin.qs of the Internation_al Recyclin(:] Conference, Berlin, Germany, 1992.

965

INDEX

abrasion resistance 361 absorption 353 as a function of time 355 accuracy 167 acid neutralising capacity 372, 454 as a function of time 372 acid gas cleaning residue 444 leaching 745 acid-base reactions 491 actinides 400 active environmental protection systems 942 leachate 942 activity coefficient 508, 511,553 activity-based sorption model 561 adatoms 551 adsorption edge 567 advection 487 aggregate 895 substitute 339 aging 550, 588 reactions 556 agitation 581 extraction tests 581 air controlled 66 excess 62 injection 73 over-fire 63 starved 76 under-fire 63 air pollution control residues 441 alkali metals fate in combustion 288 alkalinity 368 ALS Process 749 aluminum metallic 271 amphoterism 540 anion 517

ANS 16.1 843, 844 antimony 673 partitioning 311 apparent specific gravity 353 arsenic 648 partitioning 310 artificial reefs 896 ash deposition mechanisms 419 Brownian forces & Eddy diffusion 421 diffusiophoresis 421 Fickian diffusion 421 gravitation 421 inertial impaction 421 interception impaction 421 thermophoresis 421 asphalt 895 auger electron spectroscopy 251 availability 854 control 639 test 640 barium 657 base salt 507 batch tests 483, 684 BET surface area 370 analyses 495 bias 177 bidentate 528 binders blast furnace slag 777 cement kiln dust 777 coal fly ash 777 lime kiln dust 777 boiler ash 496 boiler tubes fouling 88 boilers convection 88 economiser 88 hoppers 88 radiant 88

966 superheater 88 boron 653 bottom ash 495 bottom ash removal drag-chain 90 plate 90 ram 90 bottom ash composition 401 bottom liner 934 Boudouard reaction 267 boundary layer 491 bromine containing organic compounds 300 partitioning 300 Brownian motion 498 bulk analysis 44 density 495 environment 238 specific gravity 353 as a function of time 353 cadmium 650, 657, 663, 673 partitioning 306 calcium CaCI2 production 752 partitioning 291 California bearing ratio 364 WET test 684 Canada 22 caps 953 carbon C/H ratio 267 CO formation 313 monoxide 139 oxidation reaction 312 partitioning 312 cation 517 cation/anion balances 517 CCME 22 chelate 523 chemical aspects of leaching 487 binding 766 composition 41 characteristics 425, 454 acid neutralisation capacity 426 cadmium 433 CB 437

chemical composition 428 chromium 433 CP 437 lead 433 nickel 433 PAH 437 PCB 437 PCDD 435 PCDF 435 pH 426 solubility 426 zinc 433 evaluation tests 773 acid neutralisation capacity 775 availability 773 tank leaching 775 total metal concentration 773 water solubility 773 retention 854 weathering 550 chloride 663 chlorinated benzenes 406 phenols 406 chlorine electrochemical recovery 758 partitioning 297 chromium 650, 657, 676 cleaner technology 931 closed system 592 co-disposal 944 cold-crown melting 792 collection procedures 180 column test 483, 585, 670 combined ash 409, 931 leachate 940 combustion products 272 zone 267 compacted granular leaching test 682 compaction 588 properties 447 complexation 488 complexing agents 489 compressive strength 447 concentration buildup tests 581,584 concrete 898 condensation process 280 congruent dissolution 553 contact time 579, 594

967 containment, collection and treatment of leachate 945 contaminants cadmium 154 lead 147 mercury 147 PCDD/F 147 contamination 944 control combustion 97 emission 118 fuel variability 98 mercury 113 metals 112 particulate 111 trace organics 101 controlled contaminant release 950 copper 650, 657, 663, 676 influence on PCDD/PCDF formation 319 partitioning 295 corrosion chloride induced 283 sulphate induced 282 criteria 915 crystal radii 287 structure 509 crystalline phases 496 cumulative flux 490 release 599, 854 curing 588 Darcy's Law 498 Davies equation 513 de novo synthesis 106 Deacon Process 283 Debye-Heckel equation 513 demonstration projects 897, 899 Denmark 25 density 447 separation 242 depth of analysis 247 diagenesis 509, 556 differential thermal analysis 246 diffusion 498, 841 coefficient 501,843 dioxins 406 direct approach 44 disposal 931

practices 957 strategies 940, 947 dissociation constant 524 dissolution 488, 841 reaction 488, 531 dissolvable solids content 351 DOC 7O0 documentation 193 dot maps 250 dry scrubber residue 442, 496, 679 dry system residues 442 durability 360 dynamic tests 579, 584 dynamic multicomponent flow-through leaching model 607 earth-alkali elements concentrations in waste and residues 291 fate in combustion 297 economic benefit 419 edge 551 effective diffusion coefficient 843 size 357 electron energy loss spectroscopy 252 electrostatic precipitator ash 441 surface complexation model 564 elements 377 association 246 bonding 246 distribution by country 380 related to biogeochemical cycling 377 elemental composition 460 embankment 897 emission standards 17 BImSchV 142 CCME 142 EEC 142 Ontario A-7 139 U.S. EPA 142 emitted radiation 248 energy requirements 803 entombment 948 environmental impacts 932 fugitive dust emissions 932 leachate 932 EP Tox 596 equilibrium 487

968 constant 520 pH 488 systems 488 ettringite 764 EU Landfill Directive 952 evaporative cooling 111 exotic elements 400 extended Debye-Heckel equation 513 x-ray adsorption fine structure 253 external resistance 491 extracting agents 244 extraction tests 579, 580 fabric filter residue 442 factors influencing representativeness 168 residue streams 170 type of APC system 169 type of incinerator system 169 waste type 168 FASTCHEM 611 fate 946 ferrous content 346 Fick's second law 501 field compaction 364 investigations 933 fill material 895, 899 filtration 595 final storage quality 944 flow-around tests 585 flow-through tests 585, 586 flue gas stream 420 fluid flow 486, 498 velocity 487 fluorine partitioning 299 FLUWA Process 749 flux 490 fly ash 441 and acid gas scrubbing residue leachate 937 formation enthalpies 269 PCDD/PCDF 317 formation constant 524 fouling chemical reaction 421 condensation 421

corrosion 421 particulate 421 FOWL 634 France 26 French X31-210 leaching test 596 fuel fired melter 804 fugacity 508 fugitive dust 954 furans 406 furnace configuration centre flow 73 contra 70 parallel 73 fluidised bed 85 rotary kiln 77 walls refractory 63 water 63 fusion 791,792, 817 gas bubbling 591 gas-side fouling 420 generation rates 15 geochemical modelling 658 thermodynamic equilibrium models 609 geochemistry 507 geological barrier 952 geotechnical stability 953 German DIN leaching test 684 Germany 28 Gibbs free energy of formation 520 minimisation 608 Gibbs' fundamental equation 269 Gibbs-Helmholtz equation 269 glass fraction 367 leaching 799 glassy phase 496 gradation 357 grain size distribution 346 granular material 584 (testing) 843 grate ash 339, 496 feed rates 63 manufacturers 63

969 reciprocating 65 rocking 65 roller 66 siftings 65, 339, 496 travelling 66 grinding 222 gross composition 342 guidelines 97 design 97 operating 97 GCmtleberg equation 513

intrinsic properties 942 ion activity 511 activity product 534 exchange 746 pairs 515 iron partitioning 293 separation 739 isodynamic separation 241 isotopes 400

halogens characteristics 296 thermal dissociation of hydrogen halides 297 HCI recovery 755 heat transfer surface 420 heating value 41 heterogeneous reactions 507 historical excursus 1 homogeneous reactions 507 Horsefall cell-type incinerator 4 hydraulic conductivity 447, 499 hydrodynamic dispersion 503 hydrogen evolution 451 hydroxyl 507

Japan 31 Japanese leaching test 686 jaw crushing 240

impregnation 243 Incongruent dissolution 547 increment collection classification 175 ~ndirect approach 44 ~nfiltration 492 influence of aging 367 of combustor type and operation 374 infrared spectroscopy 254 injection systems duct 111 furnace 111 reactor 111 inner sphere complexes 523 inorganic characteristics 376 constituents 460 salts 957 intended lifetime 942 internal porosity 491 resistance 491

kinetic model 507 systems 488 kinks 551 L/S 933 laboratory-field translation 881 landfill disposal APC residue 149 fly ash 149 grate ash 149 landfill design 945 operation 945 siting 945 lanthanides 400 leachant 491,579 composition 579, 589 renewal 579 leachate collection systems 951 management 948 treatment 957 leaching 905 behaviour 637 modelling 490, 622 regime 493 scenarios 493 solution 492 system 485, 491 tests 489, 509, 510, 579 time frame 493 lead 653, 657, 666, 676, 711 partitioning 309

970 ligand 507 liquid-to-solid ratio 488, 579, 592 lithophilic elements 339 heavy metal concentrations in waste and residues 292 local equilibrium assumption 487 Los Angeles abrasion test 361 loss on ignition 347, 453 lysimeter 586 macro-encapsulation 763 magnetic separation 241 major elements 464 matrix elements 379 management aspects 899 Martin reverse-acting grate 8 mass burn incinerators European 62 modular 62 mass balance 813 burn incinerator 704 streams in a MSW incinerator 285 transfer 498 constraints 491 rate 492 maximum concentration levels 933 melt structure 369 mercury 657 chloride induced 304 Hg recovery from flue gas scrubbing solutions 751 porosimetry 495 metal cation 507 metallic components in fly ashes 272 separation from bottom ashes 739 metastability 536 micro-encapsulation 763 MINEQL 609 mineral predominance 509 waste 943 mineralogy 250, 368, 450 minicolumns 626 minor elements 469 matrix elements 383

MINTEQ 6O9 MINTEQA2 530 monitoring 954, 955 monofill 149, 934 monolithic material 592 (testing) 843 morphology 247, 450 MR-Process 746 MSW management tool 419 MSWl APC residues 679 boiler ash 687, 695 bottom ash 673 economiser ash 642 ESP ash 679 grate siftings 637 semi-dry scrubber residues 679 wet scrubber residues 638 multiphase heterogeneous system 491 municipal solid waste definition 15 near surface environment 238 NEN 7345 843, 844 Netherlands 33 neutral species 512 nickel 653, 677, 711 partitioning 295 NITEP 24 nitrogen oxide emission limits 302 oxide formation 302 rfate in combustion 302 non-agitated extraction tests 581-583 non-selective catalytic reduction (SNCR) Exxon's DeNOx 116 NOxOut 117 RAPENOx 117 NOx removal 115 nuclear magnetic resonance 253 NVOC 932 on-line cleaning 420 open system 592 organic additives 783 characteristics 406 constituents 473 other

971 minor elements 385 trace elements 388 outer sphere complex 522 oversaturation 534 oxidation-reduction potential 489 PAH 324 partial pressure 508, 543 particle loading 420 characteristics 486, 495 morphology 368, 424, 496 char 425 crystals 425 fused spheres 425 opaques 425 polycrystallines 425 size 642 distribution 422, 444 reduction 221,581 particulate control 111 cyclone 103 dry systems 111 electrostatic precipitators 111 fabric filter 106 settling 103 wet systems 109 matter 441 passive environmental protection systems 942 paving applications 899, 911 PCB 320, 700 PCDD/PCDF abatement strategies 317 characteristics 314 concentrations 317 influence of burnout 317 of chlorine 314 of copper 319 thermal destruction 319 toxic equivalents 314 Peclet number 504 penetration resistance 365 percent fines 359 percolation 485, 842 permeability 366, 842, 950 petrography 247 pH 372, 454, 486, 488

static test 724 pH-dependent leaching 599 phase transfer of SiO2 273 philosophy 941 PHREQPITZ 517 phthalates 474, 476 physical aspects of leaching 485 characteristics 342 evaluation tests 771 durability 772 grain size distribution 771 moisture content 771 permeability 772 Proctor compaction test 772 total carbon content 771 unconfined compressive strength 772 Pitzer equation 512 plasma arc melter 804 polishing 245 polyaromatic hydrocarbons 406 polychlorinated benzenes 322 biphenyls 320, 406, 474, 476 dibenzofurans 474 dibenzo-p-dioxins 473 phenols 323 polycyclic aromatic hydrocarbons 324, 474, 476 pore area 495 diameter 369, 486 porosity 486, 842 postclosure care 942 potassium partitioning 289 pozzolanic reactions 766 precipitation reaction 531 precision 178 predictive capability 607 preservation 588 pretreatment 957 processing 895, 901 RDF 79 waste 62 Proctor compaction test 362 density 362 moisture 362 tests 447

972 PRODEFA2 6O9 proton 507 proximate analysis 41 pyrolysis 265 Raman spectroscopy 254 RDF incinerators semi-suspension 82 stoker 82 reaction stoichiometry 488 REDEQL 609 redox 488 potential 725 reaction 508 regulations utilisation 137 regulatory issues 901 leach test 648 reject fraction 342 release 590, 860, 910 models 858 nomographs 887 representative sample 167 Reynolds number 499 road base 670, 897 salt mines 948, 959 sample drying 240 preparation concerns 183 preservation 185 size reduction 183 storage 186 sampling considerations 175 protocols 174 strategies 194 APC system residues 199 boiler/economiser ash 198 bottom ash 194 grate siftings 197 sanitary 943, 944 saturated solution approach 611 surface dry specific gravity 353 saturation index 534 scanning electron microscopy 247, 424 transmission electron microscopy 246

tunnelling microscopy 249 seasonal variations 18 secondary ion mass spectroscopy 252 selective non-catalytic reduction 115 phase dissolution 242 semi-dry process residues 442 separation processes 735 crystallisation 751 definition 735 distillation 755 electrochemical 755 evaporation 751 Ion Exchange 749 on-site 737 physico-chemical and chemical 741 sequential chemical extraction tests 581,583 serial batch tests 585 shaking 591 short-term effects 940 sintering 272, 791,830 site density 555 siting 951 skeletal density 496 slag 339 sodium NaCI production 751 partitioning 289 sulphate soundness test 360 SOLGASWATER 609 solid phase control 492 phase approach 611 solidification 764 SOLMINEQ 88 517 SOLTEQ 517, 610 solubility 874 control 544, 642 product 532 solution control 492 solvent 491 soot blowing 420 sorption 112, 488, 489 reactions 489 surface 112 soundness 360 specific gravity 353, 495 stabilisation 682, 764 stability constant 524

973 standard enthalpy 520 entropy 520 free energy change 520 static leach tests 591 stirring 591 stockpiling 918 Stoke's law 500 storage ash 63 waste 59 submerged electrode melter 804 sulphate 659, 677 attack 778 gypsum production 752 thermal stability 268 sulphation 426 sulphur partitioning 301 surface adsorption models 611 area to volume ratio 486 reactions 505 sorption reactions 508 wash off 841,847 washing 588 Sweden 35 Swiss fly ash treatment process 743 Switzerland 37 tank leaching tests 843 TCLP 596 TDS 724 technical requirements 914 temperature fuel bed 265 gas phase 265 zones on the grate 265 thermal treatment 803 thermodynamic equilibrium model 507 thin foils 245 section 243 time series of leachate quality measurements 936 time-dependent data 339 TMT (trimercaptotriazine) 442, 933 TMT#15 776 top atomic layer 238

covers 953 tortuosity 501,842 tortuous path length 491 total amounts of major and trace components leached 937 available concentration asymptote 604 containment 948 intrusion volume 496 toxic equivalency factor 314 equivalent 147 trace elements 377, 470 oxyanionic elements 377 transmitted light microscopy 247 triple layer model 564 tumbling 591 TVA leaching test 597 ultimate analysis 41 underground storage 948 undersaturation 534 unified approach 491 leaching 653 pH curves 648 uniformity coefficients 357 unit weight 356 United Kingdom 37 United States 39 unrestricted contaminant release 948 utilisation 638, 895 valency 246 Van't Hoff relationship 522 vapour pressure of metal compounds 272 vesicles 496 visual classification 343 vitrification 791,805 vitrified ash composition 801 leaching 801 volatile metals concentrations in waste and residues 303 fate in combustion 303 Volund Incinerator at Gentofte 9

974 Von Roll Borsigstrasse Incinerator at Hamburg 11 washing 741 air pollution control residues 743 bottom ashes 741 waste categories combustibles 18 fines 18 glass 18 metals 18 non-combustibles 18 paper and card 18 plastics 18 putrescibles 18 textiles 18 waste composition 17 management policy 931 WASTE Program 45 WATEQ3 609

water content 345 solubility 452 weathering 493, 764 erosion 764 freeze/thaw 764 wet/dry 764 WES-Phix process 779 wet process (WP) residue 442 scrubber residue 442 x-ray fluorescence 230 photoelectron spectroscopy 251 powder diffraction 249 XRPD 451 zero point of charge 555 zinc 653, 657, 666, 677, 711 partitioning 307

Studies in Environmental Science Other volumes in this series 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33

Atmospheric Pollution 1978 edited by M.M. Benarie Air Pollution Reference Measurement Methods and Systems edited by T. Schneider, H.W. de Koning and L.J. Brasser Biogeochemical Cycling of Mineral-Forming Elements edited by P.A. Trudinger and D.J. Swaine Potential Industrial Carcinogens and Mutagens by L. Fishbein Industrial Waste Management by S.E. Jorgensen Trade and Environment: A Theoretical Enquiry by H. Siebert, J. Eichberger, R. (3ronych and R. Pethig Field Worker Exposure during Pesticide Application edited by W.F. Tordoir and E.A.H. van Heemstra-Lequin Atmospheric Pollution1980 edited by M.M. Benarie Energetics and Technology of Biological Elimination of Wastes edited by (3. Milazzo Bioengineering,Thermal Physiology and Comfort edited by K. Cena and J.A. Clark Atmospheric Chemistry. Fundamental Aspects by E. Meszaros Water Supply and Health edited by H. van Lelyveld and B.C.J. Zoeternan Man under Vibration. Suffering and Protection edited by (3. Bianchi, K.V. Frolov and A. Oledzki Principles of Environmental Science and Technology by S.E. Jorgensen and I. Johnsen Disposal of Radioactive Wastes by Z. Dlouh~] Mankind and Energy edited by A. Blanc-Lapierre Quality of Groundwater edited by W. van Duijvenbooden, P. Glasbergen and H. van Lelyveld Education and Safe Handling in Pesticide Application edited by E.A.H. van Heemstra-Lequin and W.F. Tordoir Physicochemical Methods for Water and Wastewater Treatment edited by L. Pawlowski Atmospheric Pollution 1982 edited by M.M. Benarie Air Pollution by Nitrogen Oxides edited by T. Schneider and L. Grant Environmental Radioanalysis by H.A. Das, A. Faanhof and H.A. van der Sloot Chemistry for Protection of the Environment edited by L. Pawlowski, A.J. Verdier and W.J. Lacy Determination and Assessment of Pesticide Exposure edited by M. Siewierski The Biosphere: Problems and Solutions edited by T.N. Veziro~lu Chemical Events in the Atmosphere and their Impact on the Environment edited by (3.B. Marini-Bettblo Fluoride Research 1985 edited by H. Tsunoda and Ming-Ho Yu Algal Biofouling edited by L.V. Evans and K.D. Hoagland Chemistry for Protection of the Environment 1985 edited by L. Pawlowski, (3. Alaerts and W.J. Lacy Acidification and its Policy Implications edited by T. Schneider Teratogens: Chemicals which Cause Birth Defects edited by V. Kolb Meyers Pesticide Chemistry by (3. Matolcsy, M. Nadasy and Y. Andriska Principles of Environmental Science and Technology (second revised edition) by S.E. J~rgensen 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 60 61 62 63 64 65 66

Chemistry for Protection of the Environment 1987 edited by L. Pawlowski, E. Mentasti, W.J. Lacy and C. Sarzanini Atmospheric Ozone Research and its Policy Implications edited by T. Schneider, S.D. Lee, G.J.R. Wolters and L.D. Grant Valuation Methods and Policy Making in Environmental Economics edited by H. Folmer and E. van lerland Asbestos in Natural Environment by H. Schreier How to Conquer Air Pollution. A Japanese Experience edited by H. Nishimura Aquatic Bioenvironmental Studies: The Hanford Experience, 1944-1984 by C.D. Becker Radon in the Environment by M. Wilkening Evaluation of Environmental Data for Regulatory and Impact Assessment by S. Ramamoorthy and E. Baddaloo Environmental Biotechnology edited by A. Blazej and V. Privarova Applied Isotope Hydrogeology by F.J. Pearson Jr., W. Balderer, H.H. Loosli, B.E. Lehmann, A. Matter, Tj. Peters, H. Schmassmann and A. Gautschi Highway Pollution edited by R.S. Hamilton and R.M. Harrison Freight Transport and the Environment edited by M. Kroon, R. Smit and J.van Ham Acidification Research in The Netherlands edited by G.J. Heij and T. Schneider Handbook of Radioactive Contamination and Decontamination by J. Severa and J. B~.r Waste Materials in Construction edited by J.J.J.M. Goumans, H.A. van der Sloot and Th.G. Aalbers Statistical Methods in Water Resources by D.R. Helsel and R.M. Hirsch Acidification Research: Evaluation and Policy Applications edited by T.Schneider Biotechniques for Air Pollution Abatement and Odour Control Policies edited by A.J. Dragt and J. van Ham Environmental Science Theory. Concepts and Methods in a One-World, Problem-Oriented Paradigm by W.T. de Groot Chemistry and Biology of Water, Air and Soil. Environmental Aspects edited by J. T61gyessy The Removal of Nitrogen Compounds from Wastewater by B. Halling-S~rensen and S.E. Jorgensen Environmental Contamination edited by J.-P. Vernet The Reclamation of Former Coal Mines and Steelworks by I.G. Richards, J.P. Palmer and P.A. Barratt Natural Analogue Studies in the Geological Disposal of Radioactive Wastes by W. Miller, R. Alexander, N. Chapman, I. McKinley and J. Smellie Water and Peace in the Middle East edited by J. Isaac and H. Shuval Environmental Oriented Electrochemistry edited by C.A.C. Sequeira Environmental Aspects of Construction with Waste Materials edited by J.J.J.M. Goumans, H.A. van der Sloot and Th. G. Aalbers. Caracterization and Control of Odours and VOC in the Process Industries edited by S. Vigneron, J. Hermia, J. Chaouki Nordic Radioecology. The Transfer of Radionuclides through Nordic Ecosystems to Man edited by H. Dahlgaard Atmospheric Deposition in Relation to Acidification and Eutrophication by J.W. Erisman and G.P.J. Draaijers Acid Rain Research: do we have enough answers? edited by G.J. Heij and J.W. Erisman Climate Change Research: Evalution and Policy Implications (in two volumes) edited by S. Zwerver, R.S.A.R. van Rompaey, M.T.J. Kok and M.M. Berk Global Environmental Biotechnology edited by D.L. Wise

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