The Handbook of Environmental Chemistry Volume 3 Part A Edited by 0. Hutzinger

Pedro Áureo

Environmental Chemistry is a relatively young science. Interestin this subject, however, is growing very rapidly and, although no agreement has been reached as yet about the exact content and Iimits of this interdisciplinary discipline, there appears to be increasing interest in seeing environmental topics which are based on chemistry embodied in this subject. One of the first objectives ofEnvironmental Chemistry must be the study ofthe environment and of natural chemical processes which occur in the environment. A major purpose of this series on Environmental Chemistry, therefore, is to present a reasonably uniform view of various aspects of the chemistry of the environ- ment and chemical reactions occurring in the environment. The industrial activities of man have given a new dimension to Environ- mental Chemistry. Wehave now synthesized and described over five million chemical compounds and chemical industry produces about hundred and fifty million tons of synthetic chemieals annually. We ship billions of tons of oil per year and through mining operations and other geophysical modifications, large quantities of inorganic and organic materials are released from their natural deposits. Cities and metropolitan areas ofup to 15 million inhabitants produce large quantities ofwaste in relatively small and confined areas. Much of the chemical products and waste products of modern society are released into the environment either during production, storage, transport, use or ultimate disposal. These released materials participate in natural cycles and reactions and frequently Iead to interference and disturbance of natural systems.

The Handbook of Environmental Chemistry Volume 3 Part A Edited by 0. Hutzinger Anthropogenie Compounds With Contributions by R.Anliker, G.C.Butler, E.A.Clarke, U.Förstner, W Funke, C. Hyslop, G. Kaiser, C. Rappe, J. Russow, G. Tölg, M. Zander, V. Zitko With 61 Figures Springer-Verlag Berlin Heidelberg GmbH 1980 Professor Dr. Otto Hutzinger Laboratory of Environmental and Toxicological Chemistry University of Amsterdam, Nieuwe Achtergracht 166 Amsterdam, The Netherlands ISBN 978-3-662-15998-9 Library of Congress Cataloging in Publication Data Main entry under title: Anthropogenie compounds. (The Handbook of environmental chemistry; v. 3, pt. A-). Includes bibliographies and index. I. Pollution- Environmental aspects. 2. Pollution- Toxicology. 3. Environmental chemistry. I. Butler, Gordon Cecil, 1913-. Il. Series: Handbook of environmental chemistry; v. 3, pt. A-. QD31.H335 vol. 3, pt. A, etc. [QH545.Al] 80-16609 ISBN 978-3-662-15998-9 ISBN 978-3-540-38522-6 (eBook) DOI 10.1007/978-3-540-38522-6 This work is subject to copyright. All rights are reserved, whether the whole or part of the material is concerned, specifically those of translation, reprinting, re-use of illustrations, broadcasting, reproduction by photocopying machine or similar means, and storage in data banks. Under §54 of the German Copyright Law where copies are made for other than private use, a fee is payable to the publisher, the amount of the fee to be determined by agreement with the publisher. © by Springer-Verlag Berlin Heidelberg 1980 Originally published by Springer-Verlag Berlin Heidelberg New York in 1980 Softcoverreprint ofthe bardeover Istedition 1980 The use of registered names, trademarks, etc. in this publication does not imply, even in the absence of a specific statement, that such names are exempt from the relevant protective laws and regulations and therefore free for general use. 2152/3140-543210 Preface Environmental Chemistry is a relatively young science. Interestin this subject, however, is growing very rapidly and, although no agreement has been reached as yet about the exact content and Iimits of this interdisciplinary discipline, there appears to be increasing interest in seeing environmental topics which are based on chemistry embodied in this subject. One of the first objectives ofEnvironmental Chemistry must be the study ofthe environment and of natural chemical processes which occur in the environment. A major purpose of this series on Environmental Chemistry, therefore, is to present a reasonably uniform view of various aspects of the chemistry of the environment and chemical reactions occurring in the environment. The industrial activities of man have given a new dimension to Environmental Chemistry. Wehave now synthesized and described over five million chemical compounds and chemical industry produces about hundred and fifty million tons of synthetic chemieals annually. We ship billions of tons of oil per year and through mining operations and other geophysical modifications, large quantities of inorganic and organic materials are released from their natural deposits. Cities and metropolitan areas ofup to 15 million inhabitants produce large quantities ofwaste in relatively small and confined areas. Much of the chemical products and waste products of modern society are released into the environment either during production, storage, transport, use or ultimate disposal. These released materials participate in natural cycles and reactions and frequently Iead to interference and disturbance of natural systems. Environmental Chemistry is concerned with reactions in the environment. It is about distribution and equilibria between environmental compartments. It is about reactions, pathways, thermodynamics and kinetics. An important purpose of this Handbook is to aid understanding of the basic distribution and chemical reaction processes which occur in the environment. Laws regulating toxic substances in various contries are designed to assess and control risk of chemieals to man and his environment. Science can contribute in two areas to this assessment; firstly in the area oftoxicology and secondly in the area of chemical exposure. The available concentration ("environmental exposure concentration") depends on the fate of chemical compounds in the environment and thus their distribution and reaction behaviour in the environment. One very important contribution of Environmental VI Preface Chemistry to the above mentioned toxic substances laws is to develop laboratory test methods, or mathematical correlations and models, that predict the environmental fate of new chemical compounds. The third purpose of this Handbook is to help in the basic understanding and development of such test methods and models. The last explicit purpose of the Handbook is to present, in concise form, the most important properties relating to environmental chemistry and hazard assessment for the most important series of chemical compounds. At the moment three volumes of the Handbook are planned. Volume 1 deals with the natural environment and the biogeochemical cycles therein, including some background information such as energetics and ecology. Volume 2 is concerned with reactions and processess in the environment and deals with physical factors such as transport and adsorption, and chemical, photochemical and biochemical reactions in the environment, as weil as some aspects of pharmacokinetics and metabolism within organisms. Volume 3 deals with anthropogenic compounds, their chemical backgrounds, production methods and information about their use, their environmental behaviour, analytical methodology and some important aspects of their toxic effects. The material for volume 1, 2 and 3 was each more than could easily be fitted into a single volume, and for this reason, as weil as for the purpose of rapid publication of available manuscripts, all three volumes were divided in the parts A and B. Part A of ail three volumes is now being published and the second part of each of these volumes should appear about six months thereafter. Publisher and editor hope to keep materials ofthe volumes one to three up to date and to extend coverage in the subject areas by publishing further parts in the future. Plans also exist for volumes dealing with different subject matter such as analysis, chemical technology and toxicology, and readers are encouraged to offer suggestions and advice as to future editions of "The Handbook of Environmental Chemistry". Most chapters in the Handbook are written to a fairly advanced Ievel and should be of interest to the graduate student and practising scientist. I also hope that the subject matter treated will be of interest to people outside chemistry and to scientists in industry as weil as government and regulatory bodies. It would be very satisfying for me to see the books used as a basis for developing graduate courses in Environmental Chemistry. Due to the breadth of the subject matter, it was not easy to edit this Handbook. Specialists had to be found in quite different areas of science who were willing to contribute a chapter within the prescribed schedule. It is with great satisfaction that I thank ail 52 authors from 8 contries for their understanding and for devoting their time to this effort. Special thanks are due to Dr. F. Boschke of Springer for his advice and discussions throughout all stages of preparation of the Handbook. Mrs. A. Heinrich of Springer has significantly contributed to the technical development of the book through her conscientious and efficient work. Finaily I like to thank my family, students and coileagues for being so patient with me during several critical phases of preparation for the Handbook, and to some colleagues and the secretaries for technical help. Preface VII I consider it a privilege to see my chosen subject grow. My interest in Environmental Chemistry datesback to my early college days in Vienna. I received significant impulses during my postdoctoral period at the University of California and my interest slowly developed during my time with the National Research Council of Canada, before I could devote my full time to Environmental Chemistry, herein Amsterdam. I hope this Handbook may help deepen the interest of other scientists in this subject. Amsterdam, May 1980 0. Hutzinger Contents Mercury G. Kaiser and G. Tölg Historical Background . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Production-, Use-, Shipment-, andRelease Data . . . . . . . . . . . . . . . . . . Anthropogenie Discharged Mercury . . . . . . . . . . . . . . . . . . . . . . . . . Naturally Released Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemistry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Eiemental Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mercury Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Analytical Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Total Mercury Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Distinction Between Individual Mercury Compounds . . . . . . . . . . . Transport Behaviour in the Environment . . . . . . . . . . . . . . . . . . . . . . . . Transport into the Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . Natural Input . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Transport in the Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemical, Biochemical and Photochemical Reactions . . . . . . . . . . . . . . Conversion Between Inorganic Forms . . . . . . . . . . . . . . . . . . . . . . . Conversion Between Organic and Inorganic Forms . . . . . . . . . . . . . Conversion Between Organic Forms . . . . . . . . . . . . . . . . . . . . . . . . . Transalkylation Reaction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Metabolism . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Uptake ofinorganic Mercury . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Organic Mercury Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biotransformation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biodegradation - Decontamination of Polluted Areas . . . . . . . . . . . . . . Accumulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Persistence . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biological Effects and Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biological and Toxicological Effects . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1 3 3 8 8 8 10 12 12 16 17 17 17 19 23 23 24 25 25 25 25 26 26 28 29 38 39 39 43 X Contents Cadmium U. Förstner Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Production, Consumption, and Use . . . . . . . . . . . . . . . . . . . . . . . . . . . . Production . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Consumption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Use . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . General Chemistry, Mineralogy, Geochemistry, Aquatic Chemistry . . Chemistry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mineralogy . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Geochemistry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Aquatic Chemistry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Analytical Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sources, Pathways, and Reservoirs in the Environment . . . . . . . . . . . . . Sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Pathways . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Reservoirs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cycling of Cadffiium in Natural Systems . . . . . . . . . . . . . . . . . . . . . Chemical Reactions: Sorption and Release of Cd on Particulates . . . . . Leaching Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Remobilization Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biological Uptake and Accumulation of Cadmium in Organisms . . . . . Uptake in Plants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Uptake, Absorption, Storage, and Excretion in Animals . . . . . . . . . Food Chain Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Indicator Organisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Human Intake, Absorption, and Excretion ofCadmium . . . . . . . . . . . . Food Concentrations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Intake from Food, Water, and Air . . . . . . . . . . . . . . . . . . . . . . . . . . Absorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Body Distribution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Excretion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biological Half-Time in Rumans . . . . . . . . . . . . . . . . . . . . . . . . . . . . Toxicological Aspects of Cadmium Pollution . . . . . . . . . . . . . . . . . . . . . Toxic Effects on Aquatic Organisms . . . . . . . . . . . . . . . . . . . . . . . . . Toxic Effects on Rumans . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Regulations . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References ................................................. 59 60 60 61 62 64 64 64 65 66 68 69 70 71 74 80 81 81 83 85 85 86 88 90 91 91 93 94 95 95 96 96 96 98 99 101 Polycyclic Aromatic and Heteroaromatic Hydrocarbons M.Zander Origin and Formation ....................................... 109 Chemistry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 112 Nomenetature ........................................... 112 Contents Building Principles . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Relationships Between Topology, Stability, and Reactivity ofPAH Synthetic Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Analytical Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tansport Behaviour in the Environment . . . . . . . . . . . . . . . . . . . . . . . . Air . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Water . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemical and Photochemical Reactions . . . . . . . . . . . . . . . . . . . . . . . . Metabolism . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biodegradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . The Overall Environmental Fate ofPAH . . . . . . . . . . . . . . . . . . . . . . . Toxicology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . XI 112 114 116 118 119 119 120 120 122 125 125 126 128 Fluorocarbons J. Russow Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Production and U se . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemistry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Analytical Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Transport Behaviour in the Environment . . . . . . . . . . . . . . . . . . . . . . . Chemical and Photochemical Reactions . . . . . . . . . . . . . . . . . . . . . . . . Metabolism . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biodegradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Accumulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biological Effects and Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 133 134 136 137 137 142 145 145 145 145 146 Chlorinated Paraffins V. Zitko Production and Applications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemistry . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Determination . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chlorinated Paraffins in the Environment . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 149 151 153 154 156 Chloroaromatic Compounds Containing Oxygen C. Rappe Chlorophenols . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Production, Use, Contaminants . . . . . . . . . . . . . . . . . . . . . . . . . . . . Analytical Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Physical and Chemical Properties . . . . . . . . . . . . . . . . . . . . . . . . . . . 157 157 158 158 XII Contents Transport Behaviour . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemical and Photochemical Reactions . . . . . . . . . . . . . . . . . . . . . Metabolism and Biodegradation . . . . . . . . . . . . . . . . . . . . . . . . . . . Accumulation and Persistence . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biological Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Halogenated Dipheny1 Ethers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chlorinated Dibenzo-p-dioxins and Dibenzofurans . . . . . . . . . . . . . . . Chemical and Physical Data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Occurrence ofPCDDs and PCDFs in Industrial Chemieals . . . . . Formation ofPCDDs and PCDFs . . . . . . . . . . . . . . . . . . . . . . . . . Analytical Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Transport in the Environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemical and Photochemica1 Reactions . . . . . . . . . . . . . . . . . . . . . Metabolism and Biodegradation . . . . . . . . . . . . . . . . . . . . . . . . . . . Accumu1ation and Persistence . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Biological Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 159 159 159 159 160 160 161 161 163 165 169 170 17"1 171 174 176 176 Organic Dyes and Pigments E. A. Clarke and R. Antiker Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chemistry and U ses . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Production Data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Analytical Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ecological Aspects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Environmenta1 Assessment of Colorants . . . . . . . . . . . . . . . . . . . . . Elimination and Degradation Cycle . . . . . . . . . . . . . . . . . . . . . . . . Effluent Treatment Processes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Environmental Elimination Processes . . . . . . . . . . . . . . . . . . . . . . . Azo Dyestuffs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Triphenylmethane Dyestuffs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Xanthene Dyestuffs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Accumulation and Persistence . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Toxicological Aspects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Toxicity to Aquatic Organisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mammalian Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Legislation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 181 182 184 185 186 186 188 188 193 196 197 198 198 199 199 200 204 210 Inorganic Pigments W.Funke Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 217 Sources ofHazards in Using lnorganic Colorants . . . . . . . . . . . . . . . . 221 Production Process . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 221 Contents XIII Application . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Performance . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Removal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Welding . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Waste Disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Inorganic Colorants Based on Heavy Metals . . . . . . . . . . . . . . . . . . . . Lead Pigments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Chromate Pigments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Cadmium Pigments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Silica, Silicates and Asbestos . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Miscellaneous Inorganic Colorants . . . . . . . . . . . . . . . . . . . . . . . . . . . . Antimony . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Arsenic . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Barium ............................................... References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 221 221 222 222 222 223 223 224 225 227 227 227 228 228 228 Radioactive Substances G. C. Butler and C. Hyslop Glossary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Basic Concepts . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Radiation Doses and Units . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effects of Radiation and Dose-Effect Functions . . . . . . . . . . . . . . Dose Equivalent (H) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Committed Dose Equivalent (H 50) • • • • • • • • • • • • • • • • • • • • • • • • • • Dose-Equivalent Commitment (He) . . . . . . . . . . . . . . . . . . . . . . . . . Risk Estimates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Effective Dose Equivalent (HE) . . . . . . . . . . . . . . . . . . . . . . . . . . . . Collective Dose Equivalent . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Collective Dose Commitment (Si) . . . . . . . . . . . . . . . . . . . . . . . . . . Detriment and Dose Limits . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Transfer to Man . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Exposures ofNon-Human Biota . . . . . . . . . . . . . . . . . . . . . . . . . . . Selected Radionuclides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Tritium Oxide . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Krypton-85 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Strontium-90 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Iodine-131 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Caesium-137 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Radium-226 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Plutonium-239 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Subject Index 231 232 232 232 233 233 233 234 235 235 235 238 238 238 240 241 241 242 248 251 255 257 260 264 268 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 271 Volume 1, Part A: The Natural Environment and the Biogeochemical Cycles The Atmosphere. M. Schidlowski The Hydrosphere. J. Westalland W. Stumm Chemical Oceanography. P. J. Wangersky Chemical Aspects of Soil. E. A. Paul and P. M. Huang The Oxygen Cycle. J. C. G. Walker The Sulfur Cycle. A. J. B. Zehnder and S. H. Zimier The Phosphorus Cycle. J. Emsley Metal Cycles and Biological Methylation. P. J. Craig Natural Organohalogen Compounds. D. J. Faulkner Subject Index Volume 2, Part A: Reactions and Processes Transport and Transformation of Chemicals: A Perspective. G. L. Baughman and L. A. Burns Transport Processes in Air. J. W. Winchester Solubility, Partition Coefficients, Volatility and Evaporation Rates. D. Mackay Adsorption Processes in Soil. P. M. Huang Sedimentation Processes in the Sea. K. Kranck Chemical and Photo Oxidation. T. Mill Atmospheric Photochemistry. T. E. Graedel Photochemistry at Surfaces and Interphases. H. Par/ar Microbial Metabolism. D. T. Gibson Plant Uptake, Transport and Metabolism./. N. Morrison and A. S. Cohen Metabolism and Distribution by Aquatic Animals. V. Zitko Laboratory Microecosystems. A. R. Isensee Reaction Types in the Environment. C. M. M enzie Subject Index List of Contributors Dr. R. Anliker ETAD Clarastr. 4-6 CH - 4005 Basel 5 Switzerland G. Kaiser Max-Planck-Institut für Metallforschung D -7070 Schwäbisch Gmünd Federal Republic of Germany Dr. G. C. Butler Div. of Biological Seiences National Research Council ofCanada Ottawa, Canada KlA OR6 Prof. C. Rappe Dept. of Organic Chemistry University ofUmeä S - 901 87 Umeä Sweden Dr. E. A. Clarke ETAD Clarastr. 4-6 CH- 4005 BaselS Switzerland Dr. J. Russow HoechstAG D - 6230 Frankfurt/M. 80 Federal Republic of Germany Prof. U. Förstner Institut für Sedimentforschung Universität Heidelberg D -6900 Heidelberg Federal Republic of Germany Prof. W. Funke II. Institut für Technische Chemie Universität Stuttgart D - 7000 Stuttgart 80 Federal Republic of Germany Dr. Colleen Hyslop Div. of Biological Seiences National Research Council ofCanada Ottawa, Canada KlA OR6 Prof. G. Tölg Max-Planck-Institut für Metallforschung D - 7070 Schwäbisch Gmünd Federal Republic of Germany Prof. M. Zander Rütgerswerke AG D- 4620 Castrop-Rauxel Federal Republic of Germany Dr. V. Zitko Fisheries and Environmental Seiences Fisheries and Oceans Biological Station St. Andrews, N. B. Canada EOG 2XO Mercury G. Kaiser, G. Tölg Max-Planck-Institut für Metallforschung, Institut für Werkstoffwissenschaften, Laboratorium für Reinststoffe D-7070 Schwäbisch Gmünd, Federal Republic of Germany Historical Background The story ofmercury can be traced back to prehistoric times. A precise dating is, however, impossible because reliable written records are lacking [1]. The first evidences of the use of mercury originate from the ancient Chinese, who used the metal and its principal ore cinnabar as a medicine to prolang life [2] and cinnabar for the preparation of red ink [3]. Often the Hindus [4], the Egyptians [5, 6], the Hettities [7], and the Assyrians [8] were credited with the use ofmercury. Positive proofs for this assumption are, however, stilllacking [9, 10]. The metal is said to have been known very early in Persia [9, 11] but a chronological assignment is impossible [12]. The Phoenicians exploited cinnabar in Spain from the 8th century B.C. but there is no direct evidence of their involvement with the metal [14]. In the 5th century B.C. cinnabar was used as a pigment by the Greeks [13, 15] but Aristotle is reputed to be the first in Europe who mentioned the metal itself [16]. By the first century B.C. the preparation ofmercury by roasting cinnabar and distilling ofT the metal was weil known [17]. Roman writers describe for the firsttime the process of amalgamation [18] for the recovery of gold from garments [19]. The first recorded mention of an amalgamation process being worked on a large scale appeared in the 12th century in Egypt [9] and was technologically applied in Mexico and South-America to process silver in the 16th century [20]. About one and a half pounds of mercury were used to produce one pound of silver. The life expectance of the native miners was about 6 months [21 ]. Already in about 1567 Paracelsus described a therapy for mercurial diseases of miners [22]. Throughout the Middle Ages mercury was used as an intermediate to produce gold and silver from basemetals [23-26] and for the treatment of various diseases [5, 10, 27-29]. The endeavour to eure syphilis with Hg and its 2 G. Kaiser, G. Tölg compounds persisted till the 19th century [30] although the toxic nature of mercury was already reported by ancient authors [31] and the danger of mercury vapour had been demonstrated adequately in 1493 [22]. During the 18th century as chemistry slowly evolved from alchemy into a science, the physical and chemical properties of mercury were investigated [32-35], entraining a growing use ofthe metaland its compounds [36, 37]. The first anthropogenic release of mercury into the environment began with the industrial revolution. The steam engine was invented in 1705 entailing an increased consumption offossil fuels, andin 1892 a new technique for the production of chlorine and caustic soda by electrolysis using a mercury cathode. Moreover in 1900 organo-mercury compounds ("chlorophenolquecksilber") were introduced as fungicides to treat seed and from about 1950 as slimicides [21]. The quantities discharged into the environment remained unnoticed and were disregarded until serious hazards that occurred in the 1950s in Japan and Sweden were brought to light. 2400 I! Tjf I/HO T + 1 2200 1 1 1 1400 >. ~ 's 1200 I t 1000 I 800 I 600 I I I 400 200 I ll I I I I I II II I I I I I I I I I I I I I 1 I I I I I J j 11 T 1 I I I I I I I I I I tl tl d I l II II I I I I I I I I I I I I I I I I Fig.l. Production and consumption ofmercury in 1973 [68, 54] - - production;----- consumption; EEC: European Economic Community Mercury 3 Production-, Use-, Shipment-, and Release Data Occurrence. Mercury is commonly found in nature as the red sulfide (cinnabar) and in lesser amounts as the black sulfide (metacinnabar), the formula of which may be assumed as (Hg,Zn,Fe)(S,Se) [38:--42]. It is also found in a number of minerals in which it is not an essential constituent, and in which it substitutes for other elements [38, 41 ]. It is often combined with pyrite, quartz, calcite, dolomite, stignite and others. Major deposits of cinnabar are in Spain, Italy, Yugoslavia, USSR, USA, China, Mexico [29, 40-42]. The Hg content in ore ranges from 0.3-2% [39, 43-45]. About 80% ofthe world supply ofthe last years came from these countries [46] (Fig. 1). The deposits of mercury were formed when hydrothermal solutions from hot springs or volcanic activity penetrated unstable geological formations to replace porous sandstone or Iimestone formation with mineral solution containing mercury [39, 40]. Preparation. Mercury is still prepared, in principle, as described in the 16th century [47-49]. The mined ore is crushed, ground and concentrated by flotation before being roasted in a kiln [50, 51] at 500-600 oc in the presence of air, and sometimes with added iron and calcined lime to remove sulfur. The liberated mercury passes over with the combustion gases and is condensed in water-cooled condensers. The recovery is ~ 95%. Mechanical impurities can be removed by passing the metal through a perforated paper or leather. Contaminating heavy metals can be dissolved from mercury by pouring it in a thin jet through diluted nitric acid. The purity of the metal is ~ 99%. The metal is further purified by either threefold distillation or by electrolysis [52] and is commercially offered normally in 3 purity categories (I technically pure 99.995%, II chemically pure 99.999%, III analytically pure 99.9995%), but also 99.999995% [53] and 99.9999999% [110] qualities are available. Production. Quantities for the global production in 1973 are reported between 8,747 t [46] and 9,784 t [54]. It is assumed that the amount of the produced Hg kept nearly constant at about 10,000 t (Table 1) up to 1978. Consumption. Data in Table 1 and Fig. 1 indicate declining rather than increasing global use ofmercury. Precise data from eastern bloc countries are lacking. Most of the requirements of these countries are supplied by imports [54]. In 1973 the use of mercury in the USSR amounted to 1,800 t [55]. The production in Red China largely supplies its own requirements. Anthropogenie Discharged Mercury Specific Uses and Discharges. Table 1 shows the use of mercury by final use. Data up to 1976 are available only from the USA and the FRG [54, 56]. The table does not show the potential capacity to pollote the environment by the manufacture ofthe corresponding compounds by the chemical industry. An indication of the size of pollution problems posed by the use of mercury is given by the following examples valid for the USA: 4 G. Kaiser, G. Tölg Table 1. Consumption of mercury classified by use [54, 56] Consumption [t] Irrdustrial division EEC USA FRG (1973)a 1973 1975 1973 1975 b 1976 b 137 226 Electrolysis net cons. invesl 730 300 451 35 520 372 Electrotechnique and instruments 280 868 690 60.1 61.1 81.5 Paints 70 262 250 18.7 5.6 12.4 Catalysis 60 23 25 42 12 16 95 135 63 30 92 60 50.3 31 31.9 25 31.4 26 40 21 20 0.5 0.5 78.1 Agriculture Dental use Pharmaceutical products Labaratory products Others and stock Total 170 690 2,570 23 15 40 35 1,873 160 1,770 190 0.5 61.3 44.2 808 377.6 46.7 514.6 a In 1969 the total amount was 2,830 [84] If exportation and increase in stock are considered the consumption amounts to about 330 t [56] EEC: European Economic Community b 1. Mercury losses in industry are assumed to amount to 10,850 t over a period 1944-1959 [57]. 2. Up to 1974 recycled mercury accounts for less than 20% of the total consumption [58]. (About 500 t from 2,900 t total consumption.) 3. lt is estimated that in 1968 the chloralkali industry used 590 t only to maintain inventory [58]. Electrolysis. The chloralkali industry is usually the biggest consumer (Table 1) and has been one of the biggest polluters. In recent years this industry made every effort to reduce the emissions [59] as can be seen from the consumed and the emitted mercury quantities (Table 2), which have been evaluated in the FRG [56]. Although the mercury cell capacity might be replaced by the diaphragm process by which less mercury is released to the environment, there is a trend to the former on economical grounds. Electrical Apparatus and Control Instruments. Mercury finds widespread use in fluorescent, and discharge lamps, in industrial power rectifiers, and to a great extent in mercury-cell batteries. The most part thereof is assumed tobe lost, e.g., by breakage of thermometers [60], waste of fluorescent lamps 5 Mercury Table 2. Decrease of discharged mercury [g Hg/t Chl in the chloralkali electrolysis from 1972-1976a [56] Pathway of emission Year Waste water Outgoing air Different products Dump 1971 1972 1973 1974 1975 1976 (1.7) (1.7) (2.2) (2.3) (1.8) (2.05) 25 15 29 20 25 15 24 20 10 13 15 20 9 13 13 20 9 13 8 20 5 10 5 10 a Figures in parentheses indicate production of chlorirre in 106 [Uyr) [61-63] and batteries. In control instruments mercury metal is used in barometers, gauges, thermometers, lamp seals, electrical switches, etc. Recent developments aim at substituting mercury in dry batteries [64, 65]. Thermometers with an infra-red indication are in development [65]. Catalysis. Mercury chloride and sulfate are used for converting acetylene into vinyl chloride and acetate, (PVC, PVA production). The catalyst is regenerated and recycled [66]. In addition to this, mercury compounds are used for the conversion of acetylene to acetaldehyde and for the preparation of dye raw materials. Mercury in effiuents from factories converting acetylene to a variety of products has received particular attention in the Minamata incidence in Japan [67]. Paints. The fungicidal effect of some mercury compounds is taken advantage of in the production of protective paints. Mostly organic mercury compounds such as phenyl mercury acetate, oleate and dodecenylsuccinate are used. In 1969 about 3 x 106 t paints were produced in Europe [68], which is 40% ofthe world production [69]. This corresponds to about 5,000 t ofmercury which were painted onto surfaces [68]. The stability of the mercury compounds in the paints is quite 1ow. Photochemical breakdown and vaporization of both, mercury compounds and of the metal reduce the mercury content in paints quickly [70, 71]. Recently fungicidal compounds of zinc, copper and phenyl and sulphur derivatives have been testedas substitutes for mercury [72, 73]. Agriculture. Inorganic and organic (alkyl, alkoxy, aryl) mercury compounds have been used as seed dressing (potatoes, grains, flower bulbs, cotton, etc.) and as foliage sprays against plant diseases. These uses are dangerous because mercury compounds are brought in direct contact with the ambient environment and thus contaminate plants and birds [74]. Administrative laws have promoted replacement of mercurials by substitutes [68], whose efficacy is, however, smaller than that of mercurials. Amalgamation. In electrolytic processes mercury is used for the recovery of metals (Zn), furthermore as a reducing agent and for dental fillings. Today G. Kaiser, G. Tölg 6 amalgam residues are collected and recycled. They amount in FRG to 5 t annually [56]. Dentalamalgamsare today partly replaced by artificial products (acrylic-, epoxy resins) [75]. Pharmaceuticals. Mercury compounds are used for their antiseptic and preservative properties in soaps, cosmetics, antiseptic preparations. Some cases of mercury intoxication by absorption of mercury into human skin are known [76]. Most of the mercury thus used is lost to the environment via sewage and drain waters. Pulp and Paper. Organic mercury compounds (especially phenyl mercury acetate) have been used to prevent microorganisms (bacteria, fungi, algae etc.) from growing in pulp. In recent years official regulations have been issued to eliminate mercury from those papers which come into contact with food. None the less mercury is found in paper and board products because cellulose seems to concentrate mercury from contaminated caustic soda [77]. The pulp and paper industry has recently improved with respect to water pollution, but the air pollution via the incineration of the products remains. Artificial papers derived from polyolefines, may bring further improvement since substantially less slimicides are required in their production. In many countries the use of mercury in slimicides has been banned by governmental action. 14 t are estimated tobe released into the environment on a world-wide basis [78]. Other Uses. Smaller amounts ofHg are used in the production ofplastics, in the tanning industry, and as heat transfer agent [68]. Table 3. Global mined fossil fuels and ores and released mercury during burning and smelting processes in 1970 Mined quantity Mercury content [t] [j.lg/g] Assumed Released average mercury concentration [j.lg/g] [t] Crude oil 2-3-109 0.005-2 0.04 Bituminous, anthracitic coal 2.18·1Q9 0.012-33 [81, 84, 95, 96] Lignite 0.77·1Q9 0.036-0.056 [87] Raw material Coal (all types) 3·1o9 Naturalgas 0.6-1.35 ·109 Sulfide ores (Cu, Pb, Zn) Phosphorites Bauxite Minerals for cement preparation HQ9 0.18 0.3-1,000 1()2 Ref. [80, 81, 89, 95] 1 3·1()3 [80] 0.04 20 [80, 81] 1.5-20·1Q3a [80, 81] 2.5-3·1()2 [80] a Mercury produced in the smelting process is estimated to amount to 30,000 t [81] Mercury 7 Table 4. Comparisonofglobal, natural and anthropogenic mercury emission Pollution Discharged mercury [t] Ref. Volcanoes, geysers, weathering 0.5- 5·103 [81, 97] Degassing of crustal materials 25 -150·103 [98] Evaporation from ocean 23·1oJ [99] River, glacial ice runofT 3.8·1oJ [98] Nature: Man: 6 - 10·loJb [54, 95] Processing of minerals c 1.5- 20·103 b and ores Buming of fuels 0.1- 8·1oJ [80, 81] Mercury industrya [81, 88, 93-95] a Evaluated up to 1974 b Depending on the efficacy of recycling c Basedon data from 1970 Processing of Ores. The quantity of mercury discharged through stacks in the smelting process of sulfide ores (Cu, Pb, Zn) [79] is reported between 1,500 tfyr [80] and 30,000 tfyr [81] depending on the assumed Ievel ofmercury. Thus with zinc roaster gases of a Finish company 20 t were reported to be discharged annually [55]. In addition, emissions in the processing of phosphorites, bauxite, minerals of iron and manganese (müdes) have to be considered (Table 3). In the production of sulfuric acid from zinc ores mercury can be removed from roasting gases with a newly developed technique [82]. Fossil Fue/ Combustion. Although the mercury content of fossil fuels is small the burning oflarge quantities constitute an enormous pollution hazard [83-88]. The mercury Ievels in coal depend strongly on its origin [89]. In the [84] and 33 ~g/ [88, 90] were USA for instance Ievels between 0.012 ~g/ [87]. found. For Iignite mined in the FRG Ievels lie between 0.036--0.056 ~g/ Upon incineration ofthe fuels about 90% ofthe mercury is released into the atmosphere via the flue gas [84, 86, 91]. About 10% distribute in furnace bottom ash, precipitator ash and drainwater [87, 91]. In 1970 the global coal consumption was about 3 x 109 t [92]. This corresponds to arelease of mercury of 3000 t assuming an average content of 1 ~g/ [88] (Table 4). According to figures on record, the concentration of mercury in fossil fuel [81, 93, 94], and discharged mercury between 80 ranges between 0.005-33 ~g/ [81] and 1,800 tfyr [94] based on an annual consumption of2 x 109 t. G. Kaiser, G. Tölg 8 Natorally Released Mercury Natural polluters are volcanoes [100-103], geysers [104, 105], thermal fluids [100, 103, 106] and the earth crust itselfby weathering and erosion ofrocks. The latter process is held responsible for emissions from 500 [97] up to 150,000 tjyr [81, 98]. These quantities can be calculated both from the mercury concentration in the air and its precipitation by rainfall [98, 107]. The quantities that enter the oceans by river and ice cap runoff, 3.8 x 103 t [98], aresmall in comparison with the mercury stock of5-20 x 1Q1 tin the ocean [108] (Table 4). The immense reported range of discharged mercury quantities can be traced back partly to the different average Ievel of mercury in the respective matrix used for the calculation. Additionally, it is not known how much mercury is recycled by industrial processes today. It is assumed that about 50-80% ofthe global consumed mercury (about 104 t) is lost to the environment. The amount released by burning of fossil fuel andin smelting of metals and ores can not be estimated. At first view one ought to assume that there is no impact on the atmospheric and hydrospheric mercury burden by man, since the quantities released by natural processes [98, 107] are larger (Table 4). Analysis of glacial samples from Greenland indicate, however, a significant increase in mercury depostiton during the course oftime [98]. The reasons for this increase are believed tobe less due to industrial pollution than to those activities which result in greater exposure of the earth's crust through alteration of terrestrial surfaces thus allowing more mercury to enter the atmosphere [98]. Global reflections by these emissionsarenot expected but an impact on a local ecosystem can occur if industrially derived mercury is discharged uncontrolled into the environment. Chemistry Eiemental Mercury The chemical symbol Hg for mercury was derived both from the latin name Hydrargyrum, i.e., liquid silver, and "argentum vivum" meaning live or quick silver, or from the planet Mercury and the Roman God. Mercury can easily Table 5. Some physical properties of mercury Atomic weight Melting point Boiling point Density Vapour pressure Solubility in water Ohrnie resistance a Depends on purity 200.59 38.9 oc 357.3 oc 13.595 g/cm3 (0 °C) 0.189·10- 3 ffiffi (0 °C); 1.22·10-3 (20 °C); 2.8·10- 3 (30 °C) 6·10- 6 g/100 g (25 °C) 95.76·10-8 Q m (20 °C)a 9 Mercury be obtained in a pure state by heating ofmercuric oxide [109]. There are seven stable and eleven unstable known isotopes [110, 111]. The most useful ofthem being 203 Hg (half life:47 days, ß-, y-emitter) and 197 Hg (half life:65 h, y-emitter). Mercury is a glistening silvery metal. Some important properties for eiemental mercury are compiled in Table 5 [112] and for some mercury compounds in Table 6. The reactions of mercury with some common reactants are briefly compiled in Table 7. Removal from Contaminated Rooms. On account of the high vapour pressure mercury evaparates quickly into the air after being spilled. It can be removed from breathing air by sucking it through a filter consisting of different layers of CaC12 , Nal, and activated carbon, and from laboratory air by gassing the room with H 2S and by covering the floor and the benches with Table 6. Properties of some inorganic Hg(I) and Hg(II) compounds [112, 113] Hg(!) compounds HgzFz HgzC}z HgzBrz Hgzlz Hgz(N03)z·2 HzO HgzO (Hg)zS04 HgzS Hg(II) compounds HgFz HgClz HgBrz Hglz HgO (yellow, red) HgS (a) (ß) HgS04 HgSe Hg(N03)2·H20 Hg(N03)z·1/2 HzO Hg (Me)z Hg (Et)z Hg (Ph)2 MeHgCl EtHgCl PhHgCl PhHgAc Solubility [g/100 g water] Boiling (B), Sublimation (S) Decomposition (D) and Melting (M) point [OC] D 2·10- 4 4·10- 6 2·10-8 D i. 0.06 (25 °C) i. 570D 400 345 s 140 S, 290 D 70D lOOD D D 6.6 (20 °C) 0.62 (25 °C) 6·10-3 (25 °C) 5.3 ·10-3 (25 °C) 1·10-6 (18 °C) i. D i. s v.s. i. i. sl. s. s 645D 277M 241M 257M 350D 583.5 s 583.5 s D vcc.S 79M 79M 92.5 B 159 B 121.8 s 170M (S 193 M (S 271M 149M > 100) > 100) Key: i: insoluble; sl.s.: slightly soluble; v.s. very soluble; vcc.: vaccuum G. Kaiser, G. Tölg 10 H 2S-water. Iodine carbon has proved especially useful. Splashed or spilled mercury can easily be collected by taking it up with a capillary connected with a glass container and a pump. Table 7. Reaction of mercury with some common reactants [110] Reactant Conditions Reaction products Noblegases In discharge tubes HgAr, HgKr [114] Halogens At room temperature on excess of hal. Mercurous halide [115] Mercuric halide [116] Oxygen, air At about 350 "C, room temp. (u.v., electron bombardment HgO (Hg, 02 at temp. >350 "C Ozone S, Se, Te Dry hydrides HX (X = F, Cl) H2S, NH3, PH3, AsH3 etc. I Cl N02 Conc. H2S04 HN03 Ammonia solution Room temperature On heating ;;::: 200 "C > 200 "C Room temperature Room temperature Room temperature Room temperature In air HgO HgCh, Hgl2 Hg2(N02h/Hg2(N03h Hg I, Hg n, Sulfates Hg I, Hg ll, nitrites, nitrates Millons base Mercury Compounds A detailed description of mercury chemistry has been given [117, 118]. Here only some common inorganic and organic mercury compounds are cited. Inorganic Hg ( /) Compounds. Studies ofvarious equilibria support that in Hg(l) compounds two Hg atoms are associated to give Hg2 +ions. ~+ From the equilibrium constant one can infer [119, 120] that Hg(l) ions are moderately stable towards disproportionation in solution. In spite of this any reagent that reduces the activity of Hg(II) ions compared with that of Hg(l) ions will force the equilibrium to the right. Since many Hg(II) compounds are very insoluble, are slightly dissociated in solution or form stable complexes the number ofHg(l) compounds is limited. For instance, the addition ofOH-, s--, or alkylsulfides to a solution ofHg(I) salts gives Hg and HgO, HgS or Hg(II) complexes of the organic ligands. Apart from a few soluble salts such as nitrate, chlorate, and perchlorate most known Hg(l) compounds are sparingly soluble (Table 6). A detailed review on Hg complex formation ofHg(l) is available [121]. Mercury (//) Compounds. Hg(II) compounds with highly electronegative anions F-, N03, CI04) have ionic structures; they are dissociated and 11 Mercury hydrolysed in aqueous solution. Other halides, the mcide, and sulfide are covalent in nature. They are largely undissociated in water. Mercury forms a host of strong complexes with linear or tetrahedral coordination arrangements. Complexes where mercury is five-, or six-coordinate are less common [121]. Organomercury Compounds. A number of up-to-date reviews consider organomercury compounds in general [122-124], mercury alkyls [125], organometallic reaction mechanisms [126] and complex formation of the methyl mercury cation [127]. This contribution covers only some environmentally relevant compounds. The formation of monoalkyl mercurials from mercury and methyl iodide in the presence of sunlight was discovered in 1851 [128] and the dialkyls accidentally in 1858 [129] in an attempt to form methyl mercury cyanide by double decomposition. The extraordinary toxicity ofthese compounds caused fatal poisonings at that time [130]. More comprehensive studies on organomercurials have been resumed by 1900, when the important mercuration reaction which yielded relatively inoffensive aryl compounds was discovered by Dimroth [131 ]. Mercury acetate and benzene and its derivatives were found to react to give phenyl derivates. These compounds had then already been tested for their fungicidal effect. Mercury for Meta! Substitution (Transmetallation).lt is the most universally applied method in organometallic synthesis. General methods include: a) HgX2 and LiR or AlR3, b) HgX2 and RMgX, c) Hg or sodium amalgam and RX (X = halides, sulfates) giving rise to mono-, and dialkyl mercury compounds as in ---RHgX+MX RM+HgX2 2RM+HgX2 RHgX+R'M _ __,. RHgR.' +MX. Another possibility is a disproportionation reaction as in 2 EtHgl + 2 Nal----+ HgEt2 + Na 2 H~. The Grignard route is suitable for the synthesis of primary secondary and tertiary alkyl halides [132, 133]. Mercury for Hydrogen Substitution (Mercuration). In the mercuration reaction compounds such as Hg(II) acetate react readily with aliphatic [132 to 135] and aromatic [126, 136-138] compounds with replacement of -H by -HgOAc, e.g. CH2 (COR) 2 + Hg (0Ac)2 PhNH2 +Hg(0Ac) 2 CJI6 +Hg +oAc :;;;;===!!o= (RC0) 2 CH HgOAc + HOAc P·NH 2 C6 H4 HgOAc+HOAc ~ ~ CJI5Hg0Ac + [H+]. G. Kaiser, G. Tölg 12 Phenyl mercury acetate is widely used, and was introduced in the 1920's for seed dressing, and as a fungicide in the pulp industry. Addition Reaction (Oxymercuration and Related Reactions). Aliphatic organomercury compounds with selected groups containing oxygen or nitrogen can readily be prepared by the reaction of alkanes and to a lesser extent of alkenes with mercury II salts in the presence of appropriate nucleophiles [132, 133]. R1R2C = C R 3 ~ + HgX2 + R2YH.- R 1R2C(YRJ C(HgX) R~ + HX. Methoxymethyl mercury acetate is a representative which is still allowed for use as a seed dressing. Analytical Methods Many comprehensive reviews on the determination of total mercury and individual compounds exist [43, 139-148). Therefore, only a briefsummary is given here, considering also sources of systematic errors. The analysis can roughly be classified into two groups: 1. Procedures for total mercury (eiemental mercury, inorganic, organic mercury compounds) 2. Procedures which discriminate between the respective forms. Total Mercury Analysis In the determination of total mercury Ievels, mercury and its compounds are normally converted into mercuric ion. This presupposes a range of analytical operations which are associated with methodical and systematical errors causing the analytical result to be incorrect by orders of magnitude if environmentally relevant Ievels as low as pg/g aretobe determined. The sources of such errors are complex and are inherent in each step of the analytical procedure, but lie preferentially in the taking, preparation and decomposition of the sample and the separation of mercury from the matrix, unless appropriate precautions are met, e.g., cleaning and purifying all necessary tools and reagents and storing them under cleanroom conditions [149-151]. Taking and Preparation of the Sample. Systematic errors can be encountered in the following steps of the procedures: a) sampling, if the matrix is heterogeneous [152, 153], b) during storage by interactions of the traces of mercury with interfaces (adsorption, desorption) which is dependent on the working material and the matrix to be analysed, by volatilization from solids [163-166], and from liquids [167-172], e.g. caused by bacteria, by diffusion processes which may cause introduction of blanks and Iosses in mercury if the sample is storedin plastic containers [153, 173, 174], c) with disintegration and pulverizing [153], d) drying [153, 163, 175], ande) lyophilization [163, 176, 179] of the sample. In addition to this there is always the risk of contaminating the sample seriously by insufficiently cleaned tools and devices [153]. With gase- Mercury 13 ous samples mercury normally has to be preconcentrated directly or after passing a combustion unit. The absorbers have to ensure quantitative trapping (see section preconcentration). Decomposition of the Sample. In wet oxidation the most commonly applied reagents, are HF, HCl, HC103, HC10 4, HBr, HBr03, H 2S04, H 20 2, KMn0 4, K 2 Cr20 7, K 2S20 8, V20 5 and mixtures thereof. Normally open systems are used [180-187] some ofwhich are partly automated with respect to routine work [188, 189]. Some open systems may involve the risk ofintroducing blanks from outside and losing considerable amounts ofmercury by volatilization. Closed vessels with reflux systems and special receivers [190] evade this disadvantage but so-called pressure decompositions in single, [191-194] and multiple arrangements [195] using vessels with small surfaces and acids which can be obtained extremely pure, e.g. by subboiling point or isothermal distillation [196], are preferable ifng/g and lower levels aretobe determined. Volatilization and combustion techniques in open systems [93, 197, 199] may be associated with losses in mercury by incomplete trapping or incomplete decomposition of mercury compounds [198, 200]. Organic matrices can be ashed in closed systems under static conditions [201] or under dynamic conditions [202], using also HF- [203] or UHF- [200, 204] excited oxygen or an oxygen-hydrogen flame [205, 206]. In aqueous solutions organomercurials can be degraded, e.g. with u.v. rays [189, 207], thus avoiding introduction ofblanks by the decomposition agents. Separation of Mercury from the Matrix Volatilization. From inorganic solids e.g. metals, rocks and some soils mercury can be volatilized by heating the sample (900 oq in a stream of nitrogen, air or argon [93, 208, 209], from coal by combustion in oxygen [199], from nearly all organic solids by combustion in a HF- or UHF -induced oxygen plasma [200, 203], and from aqueous solutions and decomposition solutions after reduction of the ionic mercury to the eiemental form by aeration (cold vapour technique) [210, 211]. Some papers point at interferences encountered in this technique resulting in incomplete release ofmercury [153, 212, 215]. Miscellaneous Techniques. From solids a separation is possible by solvent extraction, e.g. with benzene in case of organomercurials [216-218], from aqueous solutions by precipitation [219], co-precipitation [220], precipitation exchange on thin layers, e.g. on ZnS which reacts with ionic mercury to form HgS [221], liquid-liquid extraction [222], e.g. with dithizone [223] and other complex forming agents [224--226], by Chromatographie methods, such as ion exchange [227-229], e.g., on Wofatite 1-150 [230, 231], thin layer [232, 223], paper [234--236], and gas chromatography [237-240], by electrophoresis on a macro scale [241], and by electrodeposition which allows separation yet in the lower ngfg and pg/g range [200, 242, 243]. Preconcentration. From gaseous samples mercury can be pre-concentrated by passing the gas stream through impinger flasks containing absorbing 14 G. Kaiser, G. Tölg solutions, e.g. permanganate sulfuric acid [244, 245], iodine potassium iodide, iodine monochloride [246] and others [247, 248]. Solid adsorbers, e.g. Cu, Ag, Au, Pt [249, 258}, Chromosorb W [253}, activated charcoal [252, 253, 255, 259], with KI- [260], CdS- [255], Au-impregnated or prepared filters [234, 261 ], glass fibers [262], glass wool [153] or glass beads [263} enable separation from a gaseous sample or from combustion gases and specific preconcentration. One should always keep in mind that in gaseous samples different forms of mercury may occur which additionally may be partiewate bound [254, 264]. For total mercury analysis the gaseous sample best is passed over catalysts, e.g. CuO at 900 oc [254] or Ag at 600 oc [200, 205] in combustion units to ensure degradation of mercury compounds and complete trapping. Tandem arrangements of the mentioned absorbers enable specific separation and preconcentration ofindividual mercury forms [261, 263, 265]. Methods of Determination Determination as a Meta/. Stock [266, 267] liberated mercury by heating the sample and trapped it in a cooled capillary. The diameter of the mercury dropletswas then measured under a microscope. This way he could determine the mercury in the J.Lg/g range with good accuracy. Spectrometric Methods Spectrophotometry ( Colorimetry). Dithizone is the most widely used reagent and covers a wide concentration range down to about 0.01 J.lg/g of mercury [268]. Possibilities of an elimination of interferences from other metals which are complexed by dithizone and the application of a multitude of other reagents are reviewed [43, 141, 143]. For instance, with Brillant Green as low as 1.7 ng ofmercury can be determined [269). Atomic Absorption (AAS). By far the most popular and widely used method of determining mercury is AAS. Flame [270, 271] and graphite tube atomisation can be applied. But better sensitivities are achieved if metallic mercury is carried by a gas stream into an absorption cell in the light path of the spectrometer where the absorption is measured at 253.7 nm or more sensitively at 184.9 nm ifthe system is purged with N 2 or a noble gas, or where the absorption line is splitted by a magnetic field (Zeeman effect) [272}. Numerous publications, which are extensively reviewed [273-276}, describe combinations of cold vapour techniques- involving a subsequent pre-concentration -, with AA spectrometry [211, 277-285] (see also section separation). Some procedures are partly [288-299] and some are totally automated [289, 286, 287] with respect to routine work. Arrangements where decomposition, separation, pre-concentration, and determination are connected closely together, reduce the risk of introducing blanks and losing mercury, e.g. by interchanges ofthe traces ofmercury with large surfaces [153, 300]. An example of a so called multi-stage procedure is shown in Fig. 2 where mercury can be 15 Mercury 11 1 Generotion vesset 2 Reduction sotution 3Delivery pump (0.2 ml/min) LDesiccont SAu-Absorber 6Heoting Coil 7PTFE - cell 8 Hollow cathode lamp 9 Spectn::ameter or photodiode I interference filter 10 Microwove cavity ( 3t, 111 11 Microwove generoter ArltOOmllmin)- Decomposition o) ocids ond oxidizing reogents bl gasphase I 0 2 .H2 102 ) dissolution of combustion residue Transfer of decomp. soluhon in genen::ation vessel Fig. 2. Determination of total mercury in organic matrices by flameless AAS (I) and OES-MIP (II) after decomposition and cold vapour technique monitored at 253.7 nm both by atomic absorption and emission with detection limits of0.5 and 0.05 ng respective1y. In this practically all decomposition methods can be app1ied. For routine ana1ysis of bio1ogica1 samp1es a semiautomated device (using a mixture of HC10 3 /HC10 4/HN0 3 for decomposition) has proved especially usefu1 [188]. Atomic Emission (OES). The use of microwave induced gas plasmas (MIP) main1y he1ium, and argon p1asmas as excitation sources in connection with different separation techniques (vo1atilization, e1ectrodeposition) enab1e mercury tobe determined down to the pg/g range [153, 200, 301-306]. Poorer sensitivities are achieved with inductive1y coup1ed high frequency p1asma(ICP) [307], radio frequency p1asma- [308], and arc- [309] excitation. Atomic Fluorescence. Simi1arly to AAS, AFS techniques have great1y improved, for instance by using e1ectrode1ess discharge tubes [310] or a separation step, e.g.cold vapour technique or amalgamation. This way detection 1imits of 3 ng (0.06 ng/g) can be obtained [311, 312]. X-ray Fluorescence ( XRF). Detection limits in the J.Lg-range can be achieved ifthe sample is directly applied [313, 314]. In connection with separation steps, e.g. cold vapour technique [315] and concentration steps, e.g. precipitation exchange [221] or ion exchange [316], even ng/g 1eve1s can be detected. Electroanalytical Methods. These include mainly potentiometry [317, 318] coulometry [319, 320] dc- and ac-polarography [321-324], amperometry 16 G. Kaiser, G. Tölg [325-329] anodic stripping- [330-335] and differential pulse anodic stripping voltammetry [336, 336 b] and anodic stripping chronopotentiometry [337]. With these methods mercury can be determined in the Jlg/g to the ng/g range. Neutron Activation Analysis (NAA). NAA enables sensitive determination of mercury and excludes the risk of introducing blanks provided that the sealing of the sample and of standards happen under blank controlled conditions in the comparative NAA [338, 339]. Detection Iimits down to 1 ng are reported if interferences are excluded by chemical separation of the activated mercury from the matrix [340-343]. Chromatographie Methods. Thin layer chromatography (TLC) [344], and paper chromatography [345] are used to determine mercurials directly in liquids, and in solids after decomposition, and more sensitively after extraction with, e.g. dithizone [268] but are mostly used as separation methods in combination with other more sensitive detection systems (see below). Gas chromatography (GC) allows moresensitive determination ofboth inorganic mercurials after transfer into suitable derivatives [346, 364] and organic mercurials directly in solutions [347] or aftersolvent extraction [216--218, 237, 239, 348, 349, 366]. A detailed review is given e1sewhere [146]. Miscellaneous Methods. These include radio-release [350], catalytic methods which take advantage of the capability of mercury to catalyse or inhibit reactions [351-353], and mass spectrometry to analyse for mercury insmall natural samples [354]. Distinction Between Individual Mercury Compounds The occurrence of various forms of mercury in the environment (eiemental mercury vapour, inorganic and organic mercurials), and the high toxicity of mercury vapour and some organomercurials necessitate distinction and determination ofthe individual species. From air the individual forms can specifically be absorbed or adsorbed (see section Separation). For liquids and decomposition solutions TLC, paper chromatography, and electrophoresis with use of various complexes, papers, coated slides, and developers have been successfully applied down to the upper ng/g range [237, 355-357]. GC is an efficient technique for separation of individual inorganic and organic mercurials andin connection with sensitive detectors, e.g., flame ionization (FID) and electron capture (ECD) a very sensitive determination method. Discrimination and determination of mercurials by Chromatographie methods are exhaustively studied and reviewed [146, 218, 349, 367]. These methods are often connected with other sensitive detection systems, e.g., high performance liquid chromatography (HPLC) with voltammetry [643] or GC with flameless AAS [358-360] or OES-MIP [361-363] or with mass spectrometry [254, 368]. Thus combining selectivity and high sensitivity to enable detection of individual mercury compounds down to the pg/g range. With the aid of differential pulse anodic stripping voltammetry mercury species complexed by organic ligands can be discriminated and sensitively determined [368 a]. Mercury 17 Transport Behaviour in the Environment Transport into the Environment Mercury is released into the environment mainly as particulate matter, eiemental vapour, HgC1 2 -vapour, inorganic mercurous and mercuric compounds, methyl mercury (II) compounds, dimethyl mercury, and phenyl mercury compounds [264]. Natural discharges occur almost always in relatively low concentrations and widely distributed. In contrast to this, manmade mercury enters the environment at only a few locations but in relatively large quantities, which are assumed to amount from a fraction [88, 94, 98, 370] to an equal order ofmagnitude ofthe natural burden [99, 371]. The following sources must be considered: Naturalsources volcanic activity, geysers and thermal fluids weathering of rocks degassing of the earth mantle transpiration and decay of vegetation emanation from the ocean Anthropogenie sources Natural Input Atmosphere. There are no data on emission from volcanoes, and geysers [100, 372, 373]. It is assumed however, that the amounts greatly exceed those released from deposits [94]. Weathering ofrocks does not significantly contribute to the atmospheric burden. Mercury sulfide impounded in rocks [374] is resistant to solubilization through weathering, and enters the geocycle mostly in form of mechanically degraded particulate matter. In this form it may, however, undergo chemical and microbial transformation to the eiemental form [143, 375]. When passing through the soil further transformation e.g. into organo mercurial with aid ofbacteria [164] enable mercury to reach the atmosphere. Decay and transpiration of land plants are another source for the atmospheric mercury burden [376]. An estimation of the global quantity released by transpiration of soil and vegetation amounts to 44,000 tjyr [108] without giving details about the exact origin of this quantity. A considerable but not yet determined amount of mercury is released from the surface of the oceans [95, 644]. One ofthe mechanisms which effect transition from the hydrosphere to the atmosphere is the bursting of gas bubbles [377-379]. The aerosols thus formed as weil as those lifted from the land surface, can be transported great distances and are distributed on a global scale if the particle size is small ( < 10 llm) [95, 380]. Hydrosphere. The flux of mercury from the continents to the oceans by river and ice cap runoff (3.8 x 103 tjyr) is much less than that from the continents to the atmosphere (2.5-15 x 104 tjyr). For themost part weathering ofrocks contributes to the river runoffin form offine particles [94, 99, 381]. 18 G. Kaiser, G. Tölg Oxidation of sulfide ores may result in mercurous and mercuric ions which are readily leached by rainfall and reach the oceans by runoffunder flow, and groundwater [382]. In certain areas mercury bearing deposits, thermal springs and mine drainage contribute significant amounts to streams. All these processes are affected by physical, chemical and microbial processes and geological conditions when water passes individual strata. An elucidation of the pathways, and an assignment of the released mercury to a definite source, is very difficult on account of their complex mechanisms. The transition of man-made mercury into the environment follows a similar pattem. The greater part is discharged directly via stacks and flues (industry, space heating, transport facilities, generation of energy) or with effiuents into the aqueous environment. ._B,iospher-:'1~ ...,; Pedospherel Rivers llndustry - - - - - - - - • 1 Soil I Effluents I Solution ... Hydrosphere Ocean Lake,River Sediment c ~ J \ ~ l1 Lithosphere Mining Rocks. Deposits Volcanic activity Fig. 3. Global mercury cycle Figure 3 shows a simplified model ofthe exchange ofmercury between the different compartments of the environment. A comparison of pre-man with present day cycles ofmercury [99, 108] show a global impact by man's activity Mercury 19 upon the environmental mercury burden. Such balances are, however, to be considered reservedly as a series of differing data exist depending on the assumed background and average concentrations ofHg in the corresponding matrices used for the calculations as e.g., weathering of rocks [94, 381-384] Hg stock in ocean [385, 386] and in the earth crust [41], atmospheric burden, natural output from the earth [88, 108]. Such calculations can only give approximate values with an uncertainty of about half an order of magnitude. Transport in the Environment Atmosphere. The individual forms ofmercury in the.atmosphere contribute to the overall mercury burden in the following way: mercury vapour 4%, Hg(II) halide 25%, monomethyl mercury 21%, dimethyl mercury 1%, particulate form 4% [264]. The existence and proportion of individual forms depend on many factors [375]. Statements on the percentage of mercury vapour and partiewate bound mercury are in conflict on occasion. They range from about 4% [263, 387, 388] up to 50% [376] for particulate bound mercury, which is susceptible to transportation or removal by impactation, and dry or wet precipitation [389, 390]. Varying proportions in the air near the ground reflect the irregular transport by winds. In meterological terms the horizontal dispersion is several orders ofmagnitude greater than the vertical [391] and therefore most fallout will occur near the place of emission [392]. This contrasts, however, with high mercury Ievels in Greenland ice [98], and studies which did not observe any noticeable reduction of gaseous mercury Ievels during rain storms. The decrease is due to an increase in ventilation [393]. Jet streams carry pollutants from the industrial areas of the northem hemisphere in a concentrated band around the globe. The pollutants are precipitated beneath them. Higher mercury levels of samples from below northerly jet stream paths and adjacent latitudes confirm this assumption [372]. The regional and global circulation depends on meterological factors, e.g. wind speed and direction, rainfall intensity, and atmospheric stability. Reviews comment on numerous Contradietory views on the effectiveness of atmospheric processes in the removal of mercury from the air [264], and give a model for the calculation of evaporation and recycling rates [371]. Some authors doubt the theory of a global circulation of mercury. There is no correlation between the long distance transport ofS02, N0 2 and mercury in air [394]. Hydrosphere. The transport of a trace elementinan aqueous medium is determined by several factors [395]. For mercury, the following factors are important: dissolution of ionic species and inorganic compounds [94, 396, 397], adsorption on and coprecipitation with solids, e.g. Fe20 3 [398], Iimonite [399] or clay [400, 401], incorporation in a crystalline structure [402], cationic exchange [399], formation of complexes with organic molecules, e.g. sulfur containing proteins and humic material etc., sorption and ingestion by viable 20 G. Kaiser, G. Tölg biota [108]. Pelargic organisms agg1omerate mercury bearing particles promoting Sedimentation. Thus mercury is removed by stream sediments and related fine grained materials within a distance of a few km after being introduced into streams [382, 403] depending on the composition of the aquatic medium, redox potential, pH, temperature, amount of the suspended sediment, minera1ogical-chemical nature of the sediment, presence of complexing agents and existence of aquatic biota. In rivers, sediments are transported with a speed between 4 and 80 kmfyr [404, 405]. Mercury in lakes will be deposited and covered with layers of other deposited materials at a rate between 5 mm/yr and less than 1 mmfyr depending on whether the Iake is eutrophic or oligotrophic [95]. When the mercury pollution source is eliminated, mercury will be slowly released from the bed sediment until a steady state condition is reached. The most important physical influences on the distribution of mercury are motions of any kind such as currents, waves, turbulent mixing processes which are extensively commented upon [380]. Some mathematical models were introduced which allow generat prediction of the behaviour of pollutants [406, 407] if chemical and physical behaviour in addition to runoff, topography, 1ocally induced currents for coasta1 areas [408] as well as action of bacteria, yeast and other microflora and the composition of the bottom sediment are known. Thus inorganic mercury which is high1y preconcentrated in bottom sediments [383, 409, 410] reaches the aquatic food chain [412, 413] via the methy1ating activity ofbacteria [164,411] and it is wide1y distributed in the biota (Fig. 4). Atmosphere - l chemieoll photochemicol reoctions (CH 3 )2 Hg~ l ~ Hg++. !f ~ Hg 0 ~CH - 3 Hg+- Bioto Distribution by food web Hydrosphere Sediment Fig. 4. Interconversion ofmercurials and their mobility in the aquatic environment [413, 646] 21 Mercury Soil. Mercury distribution in soils has a characteristic profile [414, 415] (see: accumulation of Hg in soil) and its mobility appears to be due to redox potential, pH [416], drainage, type of soil [417], and other factors [418-420]. Thus sulfur containing amino acids and proteins form very strong soluble complexes [441, 422]. Humic acidsform strong comp1exes of relatively low solubility [418, 423]. Investigations on selective extraction suggest that both metallic and ionic mercury are adsorbed in the form of a humate [424] since none of the common and stable mercury compounds including HgS were found [425, 426]. Thus, the leaching into deeper layers is small. The mobility of mercury in soil depends on many factors, e.g. reduction by chemical processes, microbes, plants, and other living organisms or biotransformation into volatile mercury compounds [166, 428-431]. Models describe the behaviour of mercury in soil and evaluate time constants between 36 and 3,600 yr [371, 427]. Aquatic Food Chain. Mercury, which arrives at the aquatic environment for the most part as inorganic and phenyl mercury [432], is quickly adsorbed by organic and inorganic particulates. This particulate matter is deposited in sediments, where, in turn, inorganic mercury may be transformed into methyl mercury. Phytoplankton (the main primary producers in the aquatic environment cf. [433]) as well as Zooplankton concentrate both inorganic and alkylated mercury [409, 434-436] and they thus enter the food chain efficiently (Fig. 5). Poilutor t dissolved substances ~ /~ ---t---\ ~ Bacteria ~ Phyfoplankton -small t Zooplankton / t 1sh ~ ....-------, Large fish lnsects-- Predator fish Higher plants- Herbivores/ Predator birds ....___ _ ___, suspended particles ~ Decomposers Fig. 5. Mercury cycle in the aquatic environment [645, 646] A significant transfer from coastal pollution sources into the open ocean marine biota, possibly occurs through the food web connecting inshore plankton - where the mercury concentration is relatively high - to higher trophic levelsrather than by direct transport through water [436] (see also accumulation in marine biota). Terrestrial Food Chain. Mercury may enter the terrestrial food chain by way of seed eating species. Comparative analyses on feathers of museum birds showed an increase in mercury concentration approximately at that time G. Kaiser, G. Tölg 22 when seed dressing with methyl mercury compounds started [437, 438]. In vivo studies on terrestrial fauna, e.g. predatory birds [439], game, singing birds, and rodents [440] with contaminated feeding confirmed the origin of increased Ievels in tissue and eggs. After the use of alkylmercurials for seed dressingwas banned, the mercury levels in wild-life decreased substantially [441]. Plantstake up small amounts of mercury in ionic complexed [442] and gaseous form through leaves [443], also from dry fallout [435], and via roots [442] (see also accumulation in plants). A transport from the leaves to the roots and into the fruits is more likely than the converse [444-450]. A repercussion on man from the uptake of mercury by plants from soil and the atmosphere is as yet unknown. Figure 6 displays a flow chart of the movements of mercury in the ecosystems. Accumulation in sea animals Drinking water Accumulation in terrestrial onim. Accumulation in plants I crops Exhalation Atmosphere Sea Surface water water 111 Absorption Drainage I River Soil Irrigation 0 01 .::.: u E111 Sediment Sewage stock I Flue gos Fossil fuel burning, Smelting process Depot lndustrial Domestic discharge Fig. 6. Flow chart ofindustrially derived mercury (modified from [609]) Agriculture 23 Mercury Chemical, Biochemical, and Photochemical Reactions At all times the disproportionation reaction is prominent in the consideration of reactions in respective media [413]. Which compound prevails, depends on its solubility and to which extent metallic mercury enters or leaves the system, and on external factors that affect the biosphere [451]. Conversion Between Inorganic Forms Hg2 +-+ HgS/HgSe: Wherever sulfide and selenide ions are present, mercury sulfide or selenide form owing to the great affinity ofmercury for sulfide sulfur (Ks = 1053) and selenide. The conditions under which HgS is stable in aqueous solutions can be evaluated by the Eh-pH diagram [94, 396]. HgS seems also to be stable under anaerobic conditions. In excess of sulfide ions the complex HgS~ is formed [452] depending on the pH [453]. Areaction which is believed to occur in soils but on which no exact information is available. HgS-+Hg2 +: Humic compounds (fulvic-, humic acid, humin) increase the solubility of HgS by complex formation [424, 426]. It seems likely that an enzymatic reaction [413, 454] oxidizes the sulfide to sulfite and sulfate releasing bivalent mercury ions, which then undergo further conversion. Hg2 +-+ Hg: The transformation from the cationic to the eiemental state can occur chemically under suitable reducingconditions, e.g. in the presence ofhumic acid [397] or by bacterial cultures (Pseudomonas), yeasts, and other microflora [455-457]. As a method for detoxification under strictly anaerobic conditions the reduction ofHg2 + to Hg0 becomes an important consideration [458]. Hgü-+ Hg2 +: The oxidation depends on the redox potential in a medium which can be calculated from the formula E = 850 + 30 log [Hgl+] !J. where a is an estimation of the strength of the binding between bivalent mercury and the available complex forming substance [459]. For mercury complexes with organic soil a has been calculated to be > 1021 [460]. This means that oxidation of metallic mercury to inorganic bivalent mercury takes place in an aquatic environment if an organic substance and oxygen are present [452], e.g. 24 G. Kaiser, G. Tölg Conversion Between Organic and Inorganic Forms +: Alkyloxyalkyl mercury compounds are unstable in acid media. At pH = 0 the half-life of the reaction is about I 0 min, in humic soil (pH = 5), three days [452]. 2 +: Organomercury compounds can be degraded ArHg+, R-Hg+~ chemically and biochemically and by the effect of u.v. radiation. In general, the stability of the compounds decreases as the carbon chain increases. Different papers describe the breakdown of organomercurials into inorganic mercury in aqueous solution by u.v. radiation from low pressure lamps [461] and with a special inversion radiator in the presence of oxidizing agents [207]. Hg 2 + ~ CH 3Hg+ /(CH 3) 2Hg: Mercuric ion can be abiotically methylated by, e.g. methylcobalamin {B 12-CH 3) [462, 463] or trimethylsilyl salts [464], and biotically by enzyme systems [357]. Two different pathways for methylation are reported. One in the presence of cell free extracts of a bacterium strain [465], B12CH 3 and ATP in anaerobic conditions where the methyl group is transferred in a nonenzymatic reaction to the mercury ion and B12CH3 is regenerated enzymatically. The other reaction is an enzymatic methylation of mercury bound to homocystein as observed, e.g. in cells ofNeurospora Crasia [357, 358]. Biological methylation has been observed in river, Iake and sea water [164, 466] in soils [467, 468], andin sediments where HgS also is methylated under aerobic conditions mostly in the top layers by the reaction of various strains [412]. The reaction begins with a chemical oxidation ofthe sulfide followed by biological methylation. CH3Hg+ and (CH 3) 2Hg are formed according to aerobic or anaerobic conditions respectively [459]. From soils a methylating substance can be extracted whose efficacy is dependent on temperature, mercury concentration, pH, and type of soil [430, 431]. Humic acid seems to affect the methylation [431]. Some microorganisms in the intestine of yellow sea tuna and in the slime ofthe fish have the capability to methylate mercury [469]. CH 3Hg+ ~ Hg0 : Microbial degradation occurs in river, sea, and Iake sediments [470]. Thus the environmental methyl mercury concentration is maintained at a minimum by the continuous cycle of breakdown and formation [471]. (CH3) 2Hg ~ Hg0 : In the atmosphere dimethyl mercury is photolysed by u.v. radiation whereby radicals might form [472]: R-O(CH 2 2 )nHg+~ (eH 3) 2H g u. V. CH 0 3 u. V. Hg 0 + C2H6. + CH 3 H g 0 ____,. Further decomposition of the monomethyl mercury radicalleads to metallic mercury and another methyl radical which can abstracthydrogen or recombine, to give rise to methane or ethane respectively [457]. Both organic and inorganic compounds in the atmosphere yield eiemental mercury in presence of sunlight [418]. Dust particles onto which mercury is adsorbed may act as activation sites for photochemical processes [376]. A direct photolytic cleav- Mercury 25 age of dimethyl mercury in the troposphere is unlikely, however, due to the influence of OH, O(lD), and 0 3 [473]. Conversion Between Organic Forms CH 3Hg+ ~ (CH 3) 2Hg: The formation of dimethyl mercury from bivalent mercury in the presence of vitamin B12 [457], e.g. in decomposing fish or in sediments runs over monomethyl mercury as an intermediate. (CH 3) 2Hg ~ CH3Hg+: Dirnethyl mercury is unstable at low pH values. During its breakdown monomethyl mercury is assumed to form as an end product or as an intermediate metabolite. Transalkylation Reaction CH3 Hg+ ~ CH 3Se3+ /(CH 3) 2Se2 +: A transfer of methyl groups from methyl mercury to selenium has been observed in in vivo studies. Dimethylselenide is released [474]. Selenium salts can also attack the Co-C bond in presence of thiols [475] resulting in a transfer of the methylgroups of the Hg-Co cycle to the selenium cycle. Moreover, anhydrous selenium salts react with methyl mercury to give dimethylselenide as a major product [476]. Metabolism Metabolism of mercurials has been investigated with labelled compounds and was found to correlate with a series of factors, e.g., type of compound (inorganic mercury differs from organomercurials, which themselves differ greatly from each other. Arylmercurials are rapidly metabolized whereas the metabolically stable alkylmercurials resist degradation into the inorganic form [477]), the species to which mercurials are administered, dose, mercury body burden and the toxic effects of the individual compounds. A detailed, excellent review has been given [478]. Mercury and its compounds can enter the organism via the lungs by inhalation, the gastrointestinal tract by ingestion, the skin and via the placenta into the fetus. Uptake of Inorganic Mercury Eiemental M ercury Vapour. The eiemental form penetrates the skin ofhumans [479], and animals [480], and on account of its sparing solubility in water, it penetrates on inhaling far down the bronchial tree to the alveoli [481]. In animals between 25 and 100% are retained [482-484], whereof a part was found tobe in the lung- with a half-life of 5-10 h [485], andin the blood [481 ]. A similar deposition mechanism is assumed to occur in humans as can be inferred from autopsies [486]. But less than 0.1% is absorbed by blood and organs ifmercury is administered to the gastrointestinal tract [487]. 26 G. Kaiser, G. Tölg Mercury Compounds. In general, aerosols of inorganic compounds are absorbed to a lesser extent than mercury vapour [488]. Deposition in the air ways and the lungs depends on particle size and density. Half-lives in the peripherallung tissue lie between a day and 1 yr [489]. Gastrointestial absorption of salts is governed by their solubility and may amount to 20% for mercuric acetate [490], but less than 2% and 8% for mercuric chloride for mice [491] and humans [492], respectively. Penetration of skin has been observed with mercuric oxide, ammoniated mercuric chloride [493, 494] with potassium mercuric iodide [495], mercuric chloride [496] and with various other compounds in pigs. Organic Mercury Compounds Alkylmercurials. Respiratory uptake of methyl mercury iodide, chloride and dicyandiamide were found on various animals [497, 498]. Within 45 s of exposure 50-80% ofthe offered dimethyl mercury was absorbed by mice [499]. Gastrointestial absorption has been studied with methyl mercury chloride on humans [500], and on mice [491] with ethyl mercury on cats [501] and with various alkyl mercury salts on rats [502]. Methyl mercury dicyandiamide is absorbed from water solution through the skin of guinea pigs [503]. Placental transferwas observed with methyl mercury salts in mice [504], and guinea pigs [505], and humans [506, 507] (see also toxicology). Aryl Mercury Compounds. Phenyl mercury acetate (aerosolfparticle size 0.6--1.2 Jlm) is absorbed by animals upon inhalation within 1 h [498], penetrates the skin of rats (25% within 24 h) [508] and of humans [509] and is better absorbed from the gastrointestinal tract than inorganic mercury as was found in experiments on various animals. Measurements of the excretion in the faeces indicate absorption rates between 10% [510] and 40% [490]. Only limited mercury levels were found in the foetus indicating a limited placental transfer [511, 512]. Biotransformation Inorganic Mercury. In contrast to in vitro studies on blood where eiemental mercury is quickly oxidized - no differences in distribution and toxicity between inhaled mercury vapour and absorbed mercuric salts, and a binding by haemoglobin solutionrather than by plasma were observed [513]- in vivo studies on various animals show a higher uptake by the brain which allows the conclusion that mercury in blood passes the lungs in eiemental form [514, 515], and is only slowly converted into ionic form by enzymes [516]. The reverse process can also occur [483, 514]. 67%-84% of the total blood mercury is found tobe in blood cells immediately after exposure to mercury vapour as opposed to 25-31% if mercury ions are injected intravenously to animals [515]. A large part of eiemental mercury is taken up by the erythrocytes where it may be dissolved in the lipid structure [477]. Mercury 27 Organic Mercury Compounds. Investigations with different methyl mercury salts on various animals showed no definite difference in metabolism [501, 517, 518]. Aftermonomethyl mercury administration mercury is mainly found in the blood cells. The extent depends on the species of the animal and on the dose administered [499]. Dirnethyl mercury administered to mice by inhalation or intravenous injection was found tobe in fat deposits [499]. Two kinds oftransformation of monomethyl mercury can be assumed. A metabolic transformation of the methyl groups in situ, or a breakage ofthe covalent bond between carbon and mercury. On the one hand the slow and even elimination of mercury after administration of monomethyl mercury to various animals indicates a rather high stability of the covalent bond. More than 90% of injected methyl mercury dicyandiamide was still found as organomercury after 6 weeks in liver, spieen, and blood, 75% in plasma and brain and 55% in kidney [519]. On the other hand there is evidence of a small breakage of the covalent bond in liver [499], andin the intestinal Iumen [422, 520]. 20-90% of dimethylmercury administered to mice are rapidly exhaled, the remainder was metabolized within 20 minutes after administration into methyl mercury ion and was detected mainly in liver and bronchi [521]. There is no difference in metabolism between different salts of methyl mercury in rats [501, 522]. The compounds are almost exclusively firmly bound to the haemoglobin in the red cells [522]. Eight days after administration in the organic form more than 94% were still detected in liver [523] and brain. Metabolism into inorganic mercury takes place in the organs mentioned with time but mainly in the kidney (34% after 8 d) [524]. Aryl Mercury Compounds. Investigations are mostly restricted to phenyl mercury. No measured differences in metabolism of the different salts have been established in animals [501, 504, 522]. High Ievels are found tobe in the blood- for the most part attached to blood cells- in liver andin kidney, not more than 20 and 10% respectively in the form of organic mercury [525]. In another study 85% of a subcutaneous administration dose appeared in the urine and about 5% in the breath within 4 days [526] which indicates relatively quick breakage of the mercury carbon bond probably after ortho-hydroxylation [527]. Alkoxyalkyl Mercury Compounds. The metabolism ofthistype of compound has mainly been investigated with methyloxyethyl mercury salts in animals, indicating a fairly rapid breakage ofthe carbon mercury bond [528]. Within 24 habout 50% of a singledosewas exhaled together with ethylene and carbon dioxide. The percentage of the organic mercury in the kidney decreased from 50% after a few hours to nearly zero after one day. About 10% ofthe mercury was excreted in the urine first in the organic then exclusively in the inorganic form [528, 529]. In conclusion we can summarize that alkyl mercury compounds (mainly methyl mercury) have the highest stability in the body. The highest Ievels are found tobe in the blood according to the declining order. 28 G. Kaiser, G. Tölg Alkylmercury > phenylmercury > inorganic mercury They are found also in tissue ofkidney, liver and brain. The distribution to the brain is very slow but mercury which is present there as methylmercury has a long half-life time. The excretion occurs mainly via faeces, via kidney into the urine, the hair and to a very small extent via the milk [500, 520, 530, 531]. The normal excretion with urine is about 10 Jlg/24h. Levels over 40 Jlg are assumed tobe due to an intoxication [532]. Biodegradation - Decontamination of Poiluted Areas Biological degradation is a natural process of decontamination and detoxication of polluted systems. In addition to this, measures have been proposed to restore areas locally polluted by man [43, 451, 533]. Biological Degradation. Microbes have the capability todegrade inorganic [534] and organic mercury compounds, as has been observed to occur in lake sediments [535], soil [536], sludge [537], andin model tests to study biodegradation ofmethyl mercury compounds [470]. A series of factors, e.g., type of microorganism, mineral salt-composition in the medium, supplementary nutriants, pH, temperature, and light have been shown to affect this process [451]. The bacterial strain Pseudomonas aeruginosa obtained from aquatic mediawas found to convert mercury ion to eiemental mercury [166, 538]. The strain K 62 from the genus Pseudomonas isolated from soil, which is capable of mercury uptake and conversion, was used to remove mercurials that were present in industrial waste waters. This strain shows a high resistance to both inorganic and organic mercurials which are loosely adsorbed onto the cell surface [534, 540]. The cell wall is then biologically stimulated to induce vaporization of mercury, a process which might be prompted by a gaseous substance secreted from the bacterial surface [541, 542] or the mercurial might be chemically transformed into a form, e.g., eiemental mercury, which is more volatile [543]. Furthermore, selected strains of bacteria even show a high degree of tolerance of mercury. For instance, Pseudomonas and Pseudomonas like bacteria exhibit growth inhibition at mercury concentrations in the percent range [541, 542]. The bacteriostatic activity of mercurials towards bacteria may be a result of a different type of chemical or biological binding. In a pertinent study it has been established that the mercurial is not deposited in the cell wall of the bacterium but is attached to the cytoplasma [544]. Removal of Contaminated Sediments by Dredging. This kind of decontamination of an aquatic ecosystem has been investigated by laboratory experiments [533], and practically exercised by dredging lakes [545, 546]. The dredged sediments can be deposited in settling ponds or they may be buried. This should, however, happen together with sand, silicates, or inert clays in order to bind mercury, thus avoiding recontamination by drainage water [533]. Mercury 29 Conversion of Mercury to Mercuric Sulfide. Techniques which have been proposed are a) covering a mercury sediment with FeS or FeS2, enabling formation of mercuric sulfide by exchange ofthe sulfide ion [533], b) change of the redox potential, which is to a very large extent determined by the concentration of dissolved oxygen in an aquatic medium [547]. The rate of biological conversion of mercury depends on it. Thus conversion of aerobic to anaerobic conditions, e.g., by adding oxygen consuming easily degraded organic substances such as glucose [533] or plants [547] favours the reduction of sulfate to sulfide and with it also the formation of mercuric sulfide. This is sparingly soluble and undergoes methylation only at very high concentrations. Simultaneously the redox potential in anaerobic environments can become so low that the oxidation of eiemental mercury which is necessary for biological methylation hardly occurs. Conversion of Mercury into Dirnethyl Mercury by Raising the pH. The process of biological methylation of mercury is determined by the pH value [548]. The composition ofmicroorganisms changes with pH levels. Higher pH favours those producing dimethyl mercury, which can evaporate while lower pH those forming monomethyl mercury [410, 549] which is more likely to accumulate in aquatic biota. Lower pH values adjusted with CaC03 yielded higher methyl mercury levels in fish [550]. Other Techniques. Proposed methods are: coverage of the bottom of mercury contaminated lakes with a plastic coating, amalgamation ofmercury with metals and the use of shrimps, crabs and clams to biologically extract mercury in aquatic media [551]. Most of these methods are only of theoretical value because they are either too costly, as large areas have normally tobe restored, or are associated with eco1ogica1 damage. Accumulation The extent to which mercury has been accumulated in the different ecosystems can be ascertained if the respective background concentrations are known. Atmosphere. Background concentrations vary between 0.001 and 50ng/m3 (264] depending on the extent ofurbanization. An average value of 1-2 ng/m3 is assumed (107]. Much higher values were measured over industrialized areas and mercury deposits {Tab1e 8). In air various forms ofmercury occur, which can be partly particulate bound [254, 264, 387]. Thus, e.g. in a speciation measurement, eiemental mercury (1-15 ngjm3) monomethyl, dimethyl, diethyl mercury (150-250 ng/m3), and particulate bound mercury (1-10 ngjm3) was found [254]. Mostly total mercury concentrations are given which depend on numerous factors, e.g., site and altitude [254, 375, 552] -lower Ievels are found at higher G. Kaiser, G. Tölg 30 Table 8. Mercury Ievels in air Description Concentration range Ref. [ngtm3] Global average Atlantic (1977) FRG (1977) Russia 1-10 0.4-20 2-37 < 10 USA (San Francisco Bay) 2-50 Summer 1-25 Winter 2-10 Chicago 150-550 Iudustrial areas 150-400 Urbanization Japan (non industr.) Air over deposits mines, geysers Air over agricult. area (fungicides) Volcanic exhalations Russia Hawai < 14 [89, 533, 261627] [627] [627] [674] [376] [387] [254, 533] [254, 388] [675] 30-1()6 [89, 533, 676] 1Q4 [675] 100-9,600 730-40,000 [89] [100] elevations; temperature and barometric pressure [552], sunlight, wind speed and direction [389, 553]. Daily but also diurnal differences have been established [256, 388]. Analyses of permanent ice sheet indicate an increase of the total mercury burden in course of time. The average concentration in ice for the period 800 B.C. up to 1952 was found tobe 60± 17 ng/kg as opposed to 125±52 ngfkg for the period 1952-1965 [98]. Hydrosphere. Part of the atmospheric mercury is washed out by rain. Levels in rain water lie between 0.005 and 0.48 J..Lg/1 [267, 417, 554]. The content of mercury in an aquatic environment (Table 9) depends on many factors. Mostly total mercury concentrations are given [555, 556]. A discrimination between inorganic and organic forms has been made for river and coastal sea water wherein organic mercury makes up about half ofthe total portion [557]. In ocean water a vertical distribution of mercury from about 0.1 J..Lg/1 at the surface to 0.15-0.27 J..Lg/1 at greater depths [558, 559] appears tobe due to the uptake of mercury by plankton and the subsequent conveyance to depths by marine biota [409]. For the applicability of surface waters for the drinking water supply nationalandinternational guidelines exist [560, 561] which are compared with mercury level of some German waters (Fig. 7). Sediments. Mercury which enters rivers, lakes and oceans for the most part ends up in the sediments [108, 380, 563, 564] (Table 10). There it is accumulated with a distinct increase towards the surface [563], which might be 31 Mercury Table 9. Mercury Ievels in aquatic media Description Rain water unpolluted Concentration Ref. range [J.lg/1] 0.02 -0.48 [2643 ' 267,417, 554] 0.25 [100] Surface water unpolluted 0.1 [555] Drainage water unpolluted 0.05 [684] Ground water unpolluted 0.01 -0.46 [403, 685] 1-1000 [43 3 ] 0.01 -0.2 [267, 403, 686] near Hg-deposits 0.5 -100 [43 3 ] Rhine (Wiesbaden) 0.03 -8.4 [556 3 ] Lakewater Ontario (Canada) 0.048 Lake Constance (FRG) 0.03 -0.38 [556 3 ] Sea water North sea (1934) 0.03 [267] (Belgium 1972) 0.03 -0.76 [556 3 ] Atlantic 0.001-1.6 [688, 687, 689] Greenland sea 0.016-0.364 [690] Pacific near shore 0.012-0.15 [559] Hot springs and minerat waters 0.01 -20 [43 3 ' 555] Oil field brines and saline waters 0.1 -230 [43 3 ' 555] near volcano near Hg-deposits Riverwater unpolluted "Reviews explained by the relatively high mobility of mercury in the interior of the anaerobic sediment and its continuous concentration in new deposits [380]. Profile analyses (Fig. 8) [565, 566] suggest that the preconcentration ensue from man's activity. Detailed compilations ofmercury in sediments are available [556, 564, 567]. Marine Biota. Plankton and zooplankton the firstlinks in the aquatic food chain take up and concentrate both inorganic and alkylated mercury com- 32 G. Kaiser, G. Tölg 5 E ·a; .J::. 4 c: c: ~ 0 -~ Cl> .J::. ~ a:: c: 0 - c: c: E äi c: CD E E i5 L.. Cl> ·a; .J::. a:: ·a; :6 CJ) c: 2 !!:! ~ c: ·a CJ) ~ ~ u :I "0 c: :.J :I c: Cl> 0 E 5 "0 Cl> Cl> .J::. .0 .0 <( s 0 0 c: ~ § 'ö 0 0 Cl> Cl> CJ) "0 0 CD 3 E c: u Cl> u 0 u ~ 2 T 0 :I: T c: T :t .0 (!) a:: T lL Fig. 7. Average Ievels of dissolved mercury in river and Iake water (FRG) and threshold Iimit values (TLV) for the applicability of waters as drinking water supply [561]. -:Maximum Ievels, a: TLV ofthe EEC [560], b: TLV ofthe FRG (1975 (562]), c: Internat. Standard (WHO) mercury concentration [ ,ug /g] 0 0.2 0.4 0.6 0.8 10 12 1960 0 0.2 0.4 0.6 0.8 10 12 1960 1940 1940 1920 1920 0 z 1900 ::::> 0 1880 a:: (!) 1860 u 1840 CD 1820 1800 ~ <( a 0 z 1900 -~ t Cl> ::::> 0 "0 20 -------30 40 ~ a:: 1880 15 ~5 10 E' ~ (!) ~ 1860 u 1840 CD <( b c: 15 20 24 Ci 26 Cl> "0 1820 ---~ <p ---;? )> 34 44 70 Fig. 8. Mercury in sediments oflake Ontario a [566] and Iake Windermere b [565] Mercury 33 Table 10. Mercury Ievels in river-, Iake-, and sea-sediments Description Concentration range Remarks Ref. 191h century [380] [267, 95] [385] [556] [556] [ftg/g] Background Ievel Unpolluted waters Ocean North Sea Lake ofüntario Wisconsin river and lakes Wisconsin river and lakes River sediment (Rhine, Koblenz) Swiss lakes Lake Sangchris (Illinois, USA) ~o.6 ~o.5 0.1 -1 0.01-5.7 0.35-1 Uplitted sedim. Fraction < 63 f.LID 0.4 -2.7 [680] 684 Vicinity of chloralkali industry [680] Fraction < 63 f.LID [681] 0.037 Before coal-fired power plant operation (1965) [54] Lake Sangchris (Illinois, USA) 0.049 Power plant in operation (mean 1968-1973) [54] Minamata Bay (1 apan) 2,010 Wet weight [682] 4.5 0.01-2.23 10 u c 0 u 1.0 Cl J: 0.5 0.1 +-..--.---r--r--r-.---...-r---r-.-...,...., 0 100 200 300 400 500 600 km Distonce from land Fig. 9. Concentration of mercury in plankton in relation to the distance from North American Coast [436] pounds by direct assimilation from the adjacent medium (Fig. 9) [409, 434 to 436]. Concentration factors up to 100,000 are reported [568]. Higher trophic Ievels feed upon these organisms thus forming a biological magnification from algae feeders (mercury concentration of0.001--0.18j.1.gjg) to predators such as pike, tune and shark (mercury concentration O.Ol-5.82j.l.gjg) [569]. In fish concentration factors of 5,000 up to 100,000 are reported [69, 568, 570] because they take up mercury by ingestion and from the adjacent water. 34 G. Kaiser, G. Tölg In pike caught, e.g., at various distances down stream from a paper mill up to 8 11g/g [571], andin rainbow trout exposed to methyl mercury (60 ngjg, 1 h a day) 17.41lg/g were found [568]. A comparison ofmuseum specimen and fish caught in a river with a chlora1kali plant clearly shows up an exponential increase ofmercury. Similar results yielded comparative studies on osprey and grebe [571] (Fig. 10). 15 ~ Osprey 0 Greot Crested Grebe Ql > ~ ()\ I 5 Probable natural Ievei 1840-1865 1865-1890 1890-1915 1915-1940 1940-1964 period Fig. 10. Mercury Ievels in feathers of osprey and great crested grebe [571] The highest concentrations in fish were found in the liver, kidney and muscle [568, 572, 573] (Table 11) but also in gills and skin depending on the water being contaminated with inorganic or organic mercurials [574]. Dirnethyl mercury and some monomethyl mercury compounds can directly be taken up by diffusion across the gills [575], while pike and trout are able to concentrate orally-ingested protein bound with methyl mercury in muscle tissue [576]. Mercury was not found to accumulate in tissues ofwater plants [572]. The mechanism of accumulation is not clear but seems to be a function of metabolic rate in individual fish, differences in selection offood objects as the fish matures, or the fish's epithelial surface area [572, 577]. Detailed studies review the accumulation ofthe mercury in water organisms regarding species, mercurial [556, 572, 578], exposure time, and distribution to different organs [556, 579]. Soil. The global average concentration of mercury in soil is estimated tobe somewhere between 50 [580] and 100 ng/g [95]. Figures range from 0.1 to 5 11g/g, depending on numerous parameters (see transport through soil). Mercury 35 Table 11. Total mercury concentration in some aquatic organisms Species Concentration range [llg/g] Plants Plankton Fish: pike pike rainbow trout perch Organs of pike: heart muscle liver kidney gill scales - X Remarks Ref. [llg/g] 0.03-0.64 0.1 -5 0.2 Ruhr (1970-1972) Depending on distance from shore [556] [436] 0.19-0.59 1.2 -8 2.8 0.02-0.08 0.57-1.9 0.44 The Netherlands (1970) Vicinity of paper mill FRG (1934) UK (1972) The Netherlands (1970) Sweden (1967) [556] [571] [267] [556] [556] [571] 0.03 0.85 1 0.78 0.64 0.3 0.1 Locally, close to strong polluters, such as chloralkali plants [581], coal-fired power plants [582] and deposits [583] the mercury levels can build up to as much as 10 Jlg/g and more. While in rocks mercury is distributed more or less homogeneously with depth, in soils it has its highest concentration in the upper 5 to 20 cm [175](Fig. lla, b). A profile analysis of a high bog (Fig. 11c) suggests this accumulated mercury tobe of anthropogenic origin. From that a background level of about 18 ng/g can be derived as opposed to about 250 ng/g at a depth of about 10 cm. On the immediate surface evaporation as a result of chemical and biological processes yield lower Ievels. The effects of mercury enriched soil on the terrestrial food chain are not yet known. Hg canc. [ ng/g I Hg canc. [ ng /g I 40 20 I I At / / / / 2 4 4 Bt a c 20 40 60 80 2 At ) -6 E Bt BC arg. material [ % I 60 80 100 Ah I I Ap 20 40 60 Bv c b :::>.8 6 - ~ 8 10 10 12 14 12 16 c 14 Fig. 11 a-c. Mercury distribution in soi1 [414], a: Arab1e, b: Forest, c: High bog, ----- o/oo Humus Hg canc. [ ng/g I 100 200 36 G. Kaiser, G. Tölg Terrestrial Plants and Fruits. Some plants take up mercury from soil depending on type of soil, plant, and form of mercury [100, 585-588]. For instance, mercurous or mercuric mercury chloride is taken up by the root system oflettuce, and carrot plants [435], pines and deciduous tree [589, 175] but there is little translocation into the aerial parts [586]. Mercury Ievels in soil ;;::: 1 mg/g reduce the yield of cultivated plants by 50% [585]. Grain, grown from dressed seed has up to two times the mercury content as crops from untreated seeds [590, 591]. In the application ofmethyl mercury [592], all parts ofthe plant contain methyl mercury [593]. These data are contradicted by other investigations which assume mercury not to be taken up by plants from a contaminated soil [444, 446, 448]. A translocation from leaves- after being sprayed or after uptake of airborne mercury [594]- into the root system [447], as well as into fruits [449, 450], e.g., Iimes [595] potatoe tubers [596, 597] the pulp of tomatoes, and into rice [598] is more likely to occur. A comparison ofmercury contents in food stuffs, from 1934 [267] with present-day data (Table 12) [69, 556, 578, 579, 585, 587, 614] Table 12. Mercury Ievels of some foodstuffs Concentration range [f.Jglg] Ref. Vegetables: fresh canned 0.001-0.05 n.d. -0.06 [586] [586] Fruit: apples 0.002-0.18 [556] Eggs: egg white total Hg methy!Hg 0.023 0.023 Meat: pork 0.003-0.5 Foodstuff [599] total Hg methyl Hg canned Meat Prepared food baby food sauces Flour Mushrooms Yellow Bolete Field Mushroom [556] [599] OX 0.074 0.068 0.01 n.d. n.d. [556] -0.02 -0.02 <0.005-0.1 [679] [556] 3.17 -8.77 3.21 -6.09 [556, 678, 599] [556, 678] Drinking water (Lake Constance) 0.01 -0.08 [556] Beer, Cider 0.01 -0.02 [556] 37 Mercury with consideration of mushrooms [599] and proportion of methyl mercury [600] does not show a significant increase unless mercury containing waste from industrial, agricultural or mining processes are discharged into local water systems or where plant, fruit and seed treatment have contaminated game, other wild life or food [601, 602]. Terrestrial Animals and Man. In general, terrestrial wild-life has lower natural contents of mercury than aquatic organisms [578]. There is a difference in distribution and accumulation in different organs [578] depending on the type ofmercurial [499, 511]. For instance, organs of pheasants from an area where seed was treated with methyl mercury [511] or which where continuously fed for 60 days with methyl mercury dressed seed [615] contained up to 10 and 40 times more mercury respectively, than normal birds. High mercury levels were found to occur in the liver, spieen, and kidney of a doe in the vicinity of a chloralkali plant [594]. Generally, in animals high concentrations were found to occur in muscles, kidney and liver but mainly in the kidney and especially in some parts ofthe brain with acute and prolonged eiemental mercury exposure [478, 578, 603]. For an evaluation of the risk of accumulation in different critical organs and different kinds of exposure a mathematical model for the kinetics of mercury exchange has been proposed [604]. An accumulation of mercury in Table 13. Mercury in human tissues [612] Tissue Concentration range [IJ.g/g] Blood totala plasmaa seruma Bone Brain Gastro-Intestinal Tract unspecified stomach Hair Heart Kidney Liver Lung Museie (Skeletal) Nails Ovary Pancreas Placenta Skin Spleen Urinea 0.005 -0.02 0.002 -0.01 0.012 0.45 0.005 -2.94 aiJ.g/ml 0.075 0.0083-2.27 1.25 -7.6 0.005 -0.15 0.0063-2.75 0.005 -3.7 0.01 -0.25 0.004 -0.71 0.07 0.2 0.05 0.06 0.003 -7.27 -2.14 -1.14 -0.12 -3.34 0.004 -1.5 4.3 -114 G. Kaiser, G. Tölg 38 man can easily be established by analyses of e.g., hair and blood of fish eaters [605-608). Normal blood is reported to contain between 5 and 20 ng/ml, but that of fish consumers up to 100 ng/ml. In hair the concentration factors are yet higher [441, 610]. In exceptional cases (Minamata patients) up to 249 Jlg/g [611] were found as opposed to about 5 Jlg/g in hair of persons with no occupational exposure [609]. Detailed compilations of mercury in human tissuesofnonexposed [612] (Table 13)andexposed persons[613] are available. Persistence Aquatic Environment. Polluted river, lake, and sea sediments are a serious hazard because the mercury being confined may remain active by methylation processes for some 100 yr [43] even if the source of pollution is eliminated [616). The persistence of methyl mercury is relatively high because it is metabolized very slowly. Retention times of one year up to 3 yr [286] depending on the species are reported (Table 14). The strong binding of Table 14. Methylmercury half-life times in fish [575] Species Half-life (days) Flounder Perch Pike Eel 400-700 500 500-700 900-1000 methyl mercury with fish can not even be disrupted by boiling or frying [77]. Biological half-lives of inorganic and phenyl mercury compounds generally are shorter than those ofmethyl mercury in all aquatic species [619]. Terrestrial Environment. Generally the persistence in animals depends on the mercury compound. For instance 90% ofphenyl mercury acetate administered intravenously, intramuscularly, and orally to rats, chicks, and dogs is metabolized into inorganic mercury and then excreted within 96 h [525] and 40-60 days after administration of a single dose [603]. Ethyl mercury similarly administered is detectable in kidney andin liver for at least 21 days [621). Methyl mercury, and eiemental mercury vapour taken up by the brain remain there for a long time [622]. In soil the persistence offungicides is many months [487, 578]. Compounds ofthis kind penetrate, the pulp oftomatoes, after application and persist for 2-3 weeks [623]. A review summarizes the distribution and retention of mercury in plants and fruits after severa1 treatments with fungicides [43]. Humans. The evaluation ofthe retention ofmercury in humans is mostly restricted to whole body measurements and depends on the type of mercurial Mercury 39 and dose administered. Long half-lives occur in the brain if humans are exposed to mercury vapour [77, 622]. Formethyl mercury a half-life of70-76 days is frequently stated [500, 531, 624--626]. Biological Effects and Toxicity Biological and Toxicological Effects The biological and toxicological activity of mercury which is reviewed in general [575, 630--633] and with special regard to inorganic [634] and organic mercury compounds [635], epidemiology [43, 68], Minamata disease [636], and genetic effects [637] depends on the form in which it is taken up, the route of entry into the body (skin, inhalation, ingestion), and on the extent to which mercury is absorbed. There is copious evidence to subdivide the individual forms into the following categories in declining order ofbiological and toxicological activity: Alkylmercury salts (methyl, ethyl) > mercury vapour > inorganic mercury-, phenyl- and methoxyethyl mercury salts. Aryl mercury is largely converted to the inorganic form and handled as such in the body. Short chain alkyl mercury compounds are more soluble in lipids than are those of mercury (II) or eiemental mercury. They are also about 100 times moresoluble in lipids than in water [638] enabling CH 3Hg+ to penetrate more readily into cells than inorganic forms. Lipotropy, affinity to SH-groups (thiols) [639, 640], and other biological interactions such as inhibition of enzyme systems [457, 633,641, 642], cause alkyl mercurials tobe 10--100 times more toxic than soluble inorganic forms. Inorganic protein-bound mercurials are absorbed to a low degree in the intestinal tract and injuries heal quickly if the exposure ceases. In contrast to this methyl mercury derivatives are almost totally absorbed causing irreversible lesions, implying genetic effects which cause both darnage of reproductive cells - inheritable darnage to following generations- and of the genetic material in the chromosomes of ordinary cells - disturbances of the nuclear material, which regulates cell function, thus giving rise to carcinogenesis and teratogenic darnage [637]. Breakage and abnormal chromosome division have been shown to occur in concentrations as low as 0.05 and 0.6 Jlg/g for phenyl and methyl mercury and methoxymethyl mercury respectively in experiments on plants [658, 659, 660], animals [658, 661], and on humans [662]. The dominant effect is on the spindie fiber mechanism which is responsible for the distribution of chromosomes into equal sets in the daughter cells. CH 3Hg+ partially inactivates this mechanism thus producing cells with erroneous distributions of single chromosomes (Mongolism is one ofthe congenital disorders, which depend on it [637]). The type of genetic darnage actually observed in humans (Japan, Iraq) indicates that the same mechanism is acting as in the animal experiments. Numerous investigations have been conducted to study, e.g., embryotoxic and teratogenic effects [653, 666, 664]. 40 G. Kaiser, G. Tölg Selenium has been found to protect against the toxicity of organic [677] and inorganic [677, 699, 700] mercurials by liberating dimethylselenide from the methyl mercury cycle [457]. Animals given a high dose of mercuric compounds lethal to controls survived when treated with selenium, and a decreased passage of mercury into foetuses and into milk occurred. A change of distribution and retention within the body has been observed [700]. Symptoms of Intoxication Inorganic Mercury. Inhaled mercury vapour injures the respiratory tract and the oral cavity, e.g. sore mouth, ulcerated gums, etc. arise (633], manifestedas coughing, bronchial inflamation, ehest pains, vomiting, excitement, tremors, irritability, diarrhoea and respiratory arrest (617, 620, 633]. Longer exposure may lead to death [486, 620]. Disturbances by dental fillings has as yet not been shown, however, in some cases, allergic reactions ofthe lips and the oral mucous membranes have been observed [691]. Ingestion of dissociated salts of bivalent mercury causes precipitation of proteins upon contact with the mucous membranes of the gastrointestinal tract and produces local pain, gastric pain, and vomiting. In acute poisoning organic changes arise, such as renal failure with all sequences, and inflamation of the oral cavity, which are both reviewed in detail (630]. Typical chronic poisoning, mostly caused by occupational exposure involves injury of the central nervous system [628], which takes effect in characteristic tremor of the hands and other parts ofthe body, erethism (628] a peculiar form ofpsychic disturbance, decreased productivity, increased fatique, loss of memory and self confidence [629], injury to the kidney [647], vascular symptomatology [648], idiosyncracy [649], and effects upon the skin which are particularly marked with mercuric chloride [650]. Organic Mercury Compounds. Organic mercurials are absorbed to the skin, by inhalation, and by ingestion (633]. Methyl mercury chloride discharged into the Minamata river initiated the first disease (Minamata disease) caused by environmental pollution (636, 651, 652, 692]. As a result over 100 persons were afflicted, causing 46 deaths and several cases of prenatal intoxication manifesting in characteristic symptoms, e.g. motor disturbances, mainly ataxia, mental symptoms, congenital malformations (see below) and cerebral palsy as a major effect [636, 653]. In the mothers concerned no serious symptoms occurred [654, 655] which is suggested tobe due to the relative ease of placental transfer of methyl mercury and its preferential concentration in the foetus [653, 656, 657]. Foetal erythrocytes contained 28% more mercury than those from the mothers [657]. The onset of tissue darnage can be correlated with the concentration ofmercury in the red blood cells (Fig. 12). Postnatal intoxication involves irritation of the mucous membranes of the respiratory tract, dermatitis, and eczema upon contanct with organomercurials [625]. In systemic intoxication the latent period ofweeks to months is a characteristic feature (666]. 41 Mercury lndividuals who have died from mercury poisoning Japanese with observed symptoms of poisoning from fish consumption ( Niigata) Swedish group in which chromosome breakage was observed Finnish people who consumed large omounts of fish and had no symptoms Swedes in polluted area who consumed large amounts of fish and had no symptoms Normal consumption - a segment of the Swedish population • lsolated case in which low Ievei found. 0 Ot2 0.4 0.6 0.8 1.0 t ''-------.---...3· 1.2 1.4 t Chromosome Fetal Darnage Overt Symptoms Fatal Darnage ( Estimate) May Occur Methyl Mercury[,...g/g l - Fig. 12. Relation ofmethyl mercury Ievels in blood to physical hazards [77, 441, 658] The classical picture contains three main symptoms: 1) Sensory disturbances in the distal parts of the extremities in the tongue and round the lips; 2) ataxia; 3) concentric constriction of the visual fields, hearing loss, symptoms from the anatomic and extrapyramidal nervous system, and mental disturbances [667]. Methyl mercury penetrates the blood brain barrier - more than mercury vapour [668]- and is distributed within the brain producing specific symptoms due to the destruction of the cells in the cerebellum and the visual and hearing centers [77]. Damaged functional nerve cells, in contrast to other types of cell, are not replaced by nerve cells produced by cell division. Their function is partly taken over by other existing nerve cells. Thus darnage may be cumulative. The latent period for a manifestation of lesions can be very long [613]. The onset oftissue darnage can be correlated with the concentration of mercury in the brain- assumed to be in the form of methyl mercury (Fig. 13). In the graph the practical daily intake of mercury with food [587], the acceptable daily intake (ADI) [68, 669, 670] and the practical residue limit for food set by the FAO and WHO [671] arealso stated. Maximum Allowable Concentrations (MAC-va/ues) of Mercury and its Compounds. The MAC-value, in Germany MAK (Maximale Arbeitsplatzkonzentration), is defined as: that average concentration in the air which causes no signs or symptoms of illness or physical impairment in all but hypersensitive workers during their working day (8h/5d a week) on a continu- G. Kaiser, G. Tölg 42 lntoke of methyl mercury [ mg/doy] 12 methyl mercury [ 1-Jg/g 1 in brain 12 Fatal ( Estimate for sensitive individuals) 10 Fish eaters (Niigato district) 0.75 0.6 Estimated individual intake (500g food0.5,ug/g) ADI 0.5 1 Estimated intake in USA Estimated intake in FRG 2- - - - - . For comparison practical residue Iimit in food (0.02 -0.05 .ug /g ) 5.0 Ouvert symptoms (sensitive individuals) l..O Eating of 2 fish ( 6.71-Jg/g) for 20 doys 3.0 Fetal damage 0.25 0.2 0.1 gg& 0.02 average Ievel in food (0 011-Jg/g) 3 Fig. 13. Calculated relationship between methyl mercury intake and Ievels of methyl mercury in brain tissue [490]. 1: Fora 75 kgman[669], therecommendedADI values lie between0.03 and 0.1 mgjday [670]. 2: Evaluated total mercury from per capita consumption withoutconsideration of beverages and fish [587]. 3: Without consideration offish [587]. Calculations ofbrain tissue Ievels based upon: Brain distribution of 15% of total body methyl mercury at 10-15%. Continuous exposure for 1 year. With an excretion rate of 1%/day oftotal body mercury the indicated Ievel will almest be reached ing basis, as judged by the most sensitive internationally accepted test [672]. The definition is by and large, comparable with the concept of Threshold Limit Values (TL V) in the USA [673] (Table 15). Up to now there are no MAC-values of individual mercurials. From occupational and suicidal intoxications as well as from animal experiments the quantities which exert adverse effects to human health can roughly be construed (Table 16). MAC-values should not directly be equalized with risk. 43 Mercury Many factors, e.g., route of entry into the body, penetration rate through the skin, and absorption rate play an important role. Sensitive persons may be affected at lower Ievels whereas others may tolerate much higher ones. Table 15. Maximum allowable concentrations of mercury and its inorganic and organic compounds MAC (mg/m3] Mercury/ Hg-compound Ref. Western countfies USSR Mercury vapour lnorganic comp. Organic comp. alkyl mercury salts 0.1 0.1 0.01 0.01 0.005 [672, 697] [672, 698] [672, 697] [698] Table 16. Quantities ofmercury and some ofits compounds at which no signs of intoxication in humans have been observed compared with some LDso values Mercury/ Mercury compound Route of up takea Eiemental Hg or i.v. Mercury vapour inhal. Hg2Cl2 inhal. ingest. HgCb Hg(CNh PhHg ac. (mouse) orally, i.p. [693, 696] [693] see MAC [694, 695] [669] or 0.1-0.2 0.5(LDso) [695, 696] or 0.2-1 (LDso) [695] n.d. For comparison: CH3HgCI (mouse) = 100-1000 <27 g Ref. <0.1-0.2 2-3 (LDso) Organic Hg-comp. • or Quantity [g] = i.p. LDso [mg/kg] body weight 14 i.p. or 8 26 intraperitoneally, i.v. [635] [635] = intravenously References 1. Li Chiao P'ing: The chemical arts of old China. Amer. Chem. Soc., Easton, Pa. 1948, p. 48 2. Leicester, H.M.: The Historical Background of Chemistry, John Wiley, Inc. New York 1961, p. 58 3. Britton, R.S.: Havard J. 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Man's Environment, Ottawa, Canada 1971, Feb. 15th 656. Snyder, R.D.: New England, J. Med. 284, 1014 (1971) 657. Tejning, S.: Rep. 68 02 20, Dept. Occupat. Med., Univ. Hospt., Lund, Sweden, p. 5 658. Löfroth, G.: Swedish Nat. Sei. Res. Council, Stockholm Sweden, 2nd Ed., Sept. 1970 659. Fiskesjo, G.: Hereditas 62, 314 (1969) 58 660. 661. 662. 663. 664. 665. 666. 667. 668. 669. 670. 671. 672. 673. 674. 675. 676. 677. 678. 679. 680. 681. 682. 683. 684. 685. 686. 687. 688. 689. 690. 691. 692. 693. 694. 695. 696. 697. 698. 699. 700. G. Kaiser, G. Tölg Ramel, C.: Nord. Hyg. Tisdkr. 50, 135 (1969) Ramel, C., Magnuson, J.: Hereditas 61,231 (1969) Skerfving, S., Nanson, K., Lindstein, J.: Arch. Environ. Health 21, 133 (1970) Ramel, C. in: The Mercury Problem Oikos (Suppl.) 9, 35 (1967) Löfroth, G.: Ecological Res. Comm. Bull. 4, Swedish Nat. Sei. Res. Council1969 Cohen, M.M.: Md. Med. J. 7, 236 (1958) Höök, 0., Lundgreen, K.-D., Swensson, A.: Acta Med. Scand. 150, 131 (1954) Tsubaki, T. in: Mercury in Man's Environment. Proc. Royal Soc. Can. Symp. 1971, 131, Feh., p. 15 Takahata, N. et al.: Folia, Psych. Neuro. Japan 24, 59 (1970) Morrison, A.B.: Canadium Approach to Acceptable Daily Intakes of Mercury in Foods, Spec. Symp. on Mercury in the Man's Environment, Ottawa, Canada, Feh. 1971 Berglund, F., Berlin, M.: in: Chemical Fallout (Eds.: Miller, M.W., Berg, G.G.) Charles C. Thomas Pub!., Springfield, Ill. 1969 FAO Working Party on: Pesticide Residues and WHO Expert Comm. on Pesticides Residues, WHO Techn. Rep. Ser. No. 370, FAO Agricult. Studies, No. 73, FAO, Rome 1967 Arch. Environm. Health 19, 891 (1969) Threshold Limit Values of Air-Borne Contaminants for 1968: Recommended a. Intended Values, Adapted at the 30th Ann. Meet. Amer. Conf. Gov. Ind. Hyg., St. Louis, 1968, May 13, (Eds.: Americ. Conf. ofGov. Ind. Hyg.) Cincinnati Sergeyev, Ye.: Proc. First All-Union Conf. Geochem. Meth. ofProspecting for Ore Deposits Moscow 1957 Fujimura, Y.: Japan J. Hyg. 18, 10 (1964) Neville, G.A.: Canad. Chem. Educ. 3 (l), 4 (1967) Ganther, H., Sunde, M.: J. Food Sei. 39, l (1974) Schelenz, R., Diehl, J.-F.: Z. Lebensmittel-Untersuch. Forsch. 154, 160 (1974) Krausehe K. in: [43], p. 51 Konrad, J.G. in: Environmental Mercury Contamination (Eds.: Hartung, R. and Dinman, B.D.) Ann Arbor Sei. Pub!., Inc., Ann Arbor, 1972, p. 52 Vernet, J.P., Thomas, R.L.: Eclogae, Geol. Helv. 65 (2), 307 (1972) Takeuchi, T. [42] Interlaboratory Comparative Analysis of the DFG: Trace Metals in Rhine Sediment, Bundesanstalt für Gewässerkunde, Koblenz, FRG, 1976 Wilander, L.: Grundforbattring21 (4), 151 (1968) Brune, D., Landström, D.: Radiochem. Acta 5, 228 (1966) Heide F., Lerz, H., Böhm, G.: Naturwissenschaften 44, 441 (1957) Aidin'yan, N. Kh., Ozerova, N.A., Gipp, S.K.: Akad. Nauk. SSR, Trudy Inst. Geol. Rudn. Mestorozhd. Petrogr., Mineral. Geokhim. 99, 5 (1963) Chester, R. et al.: Mar. Pol!. Bull. 4, 28 (1973) Leatherland, T.M. et al.: Nature 232, 112 (1971) Carr, R.A., Hoover, J.B., Wilkniss, P.E.: Deep-Sea Res. 19, 747 (1972) Turkeim, H.J.: Lancet 257, 1150 (1949) Kutsuna, M.: Minamata Rep. (Study Group of Minamata Disease) Kumamoto Univ., Japan 1968 Braun, W., Dönhardt, A.: Vergiftungsregister, Georg Thieme Verlag, Stuttgart 1970 Jordi, A.: Z. Unfallmed. Berufskrankh. 36, 136 (1943); Schweiz med. Wschr.1947, 621 Moeschlin, S.: Klinik u. Therapie d. Vergiftungen, 5th Ed., Georg Thieme, Stuttgart 1972 Bader, E.W.: Gewerbekrankheiten, 4 Ed., Urban & Schwarzenberg, München, Berlin 1954 Deutsche Forschungsgemeinschaft, Maximale Arbeitsplatzkonzentrationen 1979, Senatskommission z. Prüf. gesundheitsschädl. Arbeitsstoffe, Mitt. XIV, Harald Boldt, Boppard, FRG Korbakova, A.l.: Maximum Allowable Conc. Mercury in the USSR Working Paper, Stockholm Symp., Nov. 1968 (paper can be obtained from the Institute ofHyg. Karolinska, S-10401 Stockholm 60, Sweden) Parizek, J. et al. in: New Trace Elements in Nutrition (Eds.: Mertz, W., Cornatsev, W.E.) Macel Dekker Inc., N.Y. 1971 Kosta, L., Byrne, A.R., Zelenko, V.: Nature 254, 238 (1975) Cadmium U. Förstner Institut für Sedimentforschung der Universität Beideiberg D-6900 Heidelberg, Federal Republic of Germany Introduction Cadmium is regarded as one of the most toxic metals, although there is no rigid order of toxicity of trace metals in the environment [1, 2]. The acute toxicity of cadmium upon inhalation or ingestion was recognized long ago and its chronic toxic effect on workers exposed to dust and vapor has been known approximately since 1948 [3]. Pollution by cadmium in aquatic systems appears to be less widespread than that by mercury, but has nonetheless had hazardous effects on humans. During 1947 an unusual and painful disease was recorded as of a "rheumatic nature" in the case of 44 patients from villages on the banks of the Jintsu River, Toyama Prefecture, Japan. It became later known as the "itai-itai" disease (meaning "ouch-ouch") in accordance with the patients' shrieks resu1ting from painfu1 ske1eta1 deformities; it is estimated that approximately 100 deaths occurred due to the disease until the end of 1965 [4]. However, the cause ofitai-itai disease was completely unknown until 1961 when sufficient evidence led to the postulation that cadmium plays a role in its development [5]. Basedon the findings offurther government-supported studies, the Japanese Ministry ofHealth and Welfare declared in 1968: "The itai-itai disease is caused by chronic cadmium poisoning, on condition ofthe existence of such inducing factors as pregnancy, lactation, imbalance in internal secretion, aging, deficiency of calcium, etc." [6]. After the first alarming suspicion and initial diagnosis of the itai-itai disease, numerous detailed investigations have been carried out in many countries. Heavy cadmium pollution in aquatic systems (without indications of acute toxic effects on humans, however) has been recorded in the Hudson River Estuary, New York (nickel-cadmium battery factory), the Hitachi area near TokyofJapan (braun tube factory), from Palestine Lake, Indiana/USA (plating industry), Sörfjord/Norway and Derwent Estuary, Tasmania (smelting emissions), from the Neckar River/FRG (pigment factory), and the Ves- 60 U. Förstner dre River/Belgium. The accumulation of cadmium in agricultural soils and its increased uptake by plants is presently ofworld-wide concern. Although analytically valid long-term human balance studies are not available to enable precise estimates of the rate of cadmium absorption and excretion, it has been shown that cadmium originating in different pathways can constitute a health hazard to parts of the populations. Good examples of such pathways are food (e.g., crops grown on contaminated soil), air pollution and cigarette smoking. Summaries and discussions of the environmental impact and of dose-response relationships of cadmium have been given by Friberg et al. [6], by the Panel of Hazardous Trace Substances of the US National Institute of Environmental Health [7], by the Task Group on Environmental Health Effects of Cadmium, World Health Organization [8], by the Subcommittee on the Toxicology of Metals of the International Association of Occupational Health [9] and by the Working Group of Experts for Cadmium, prepared for the Commission ofthe European Communities [10]. Mining and smelting companies in various parts of the world established the Cadmium Association in the UK and the Cadmium Council in the US in 1976 to collect and disseminate information on cadmium to consumers, governments, and other concerned groups. These two organizations publish Cadmium Abstracts, a quarterly summary of current world Iiterature on the properties and uses of cadmium, its alloys, and compounds. The First International Cadmium Conference, organized jointly by the two bodies and the International Lead and Zinc Research Organization, was held in San Francisco in January 1977 [11], and the second conference in Cannes/France in February 1979. A handbook on the biogeochemistry of cadmium is presently edited by J. 0. Nriagu [12]. This work includes data on cadmium in natural waters, in the atmosphere, in soils, in biota, andin sediments, on occupational exposure to cadmium, metabolism and toxicity in organisms, human health (including immuno1ogica1, cellu1ar, teratogenic, mutagenic, and patho1ogica1 aspects). Production, Consumption and Use Production Cadmium' is basically recovered as a by-product from the smelting and refining of zinc concentrates at a rate of approximately 7-8lbs/t primary zinc. Nooresare mined and processed exclusively to obtain cadmium. Data for 1977 show that cadmium production in western industrial countrieswas approximately 14,000 t, 7% higher than in the year before. This rate is also 17% higher than that in 1975, the year when a distinct slump occurred in demand for cadmium in most of the western countries (Fig. 1). The majority of cadmium producing countries, particularly Japan and Western Europe, 1 According to Mirring Ann. Rev. 1978, p. 101-102 61 Cadmium Consumption "' c .9 ·~ 10000 ~ ==Asia+== =Australia= I 19551959 a; (avg) 5000 • ~ 19461950 (avg) 1960 1965 1970 1975 77 74 75 76 Fig. 1. Production and consumption of cadmium (metric tons). (Data from [7] and [13]) recorded significant increases in production during 1977; this, of course, reflects expansion of zinc output (Table 1). Against a background of a 2% increase in zinc production in 1977, the 7% increase in cadmium production raised the cadmium/zinc recovery ratio. This apparently means that stocks of semi-refined material were economically converted to saleable metal boosting somewhat smelter revenues [13]. An important source of cadmium is the processing of secondary raw sources (recycling materials). These are partially added to the cadmium production process with primary raw materials. Secondary raw materials are cadmium-containing end products which have become unuseable (such as nickel-cadmium batteries), cadmium-containing by-products that are subsequently reprocessed using special procedures (dust particles, cadmium-containing muds, etc.), and all types ofmaterials whose reprocessing has become economically feasible or necessary by legislative enactment. In the Federal Republic ofGermany in 1975 as much as 20% ofthe total processed cadmium came from recycled material. Consumption Consumption of cadmium in Western industrial countries rose at an average compound rate of about 3.5% between 1964 and 1974. In 1975, however, a slump in consumption of 34% occurred, but consumption rapidly picked up in 1976 and almost regained the previous levels ofa.round 14,000 t. In 1977 cadmium demand in some Countries was down again with the consumption in 62 U. Förstner Table 1. Majorproducersand consumers of cadmium (t) [13] Meta! consumption Meta! producfion 1974 1975 1976 1977a 1974 1975 1976 1,043 200 18 156 1.176 220 18 426 586 1,454 439 398 1,445 220 18 519 782 1,535 435 300 1,682 1,036 1,426 450 450 460 1,338 505 75 305 178 950 220 18 217 602 923 421 272 370 205 1,460 2,016 430 825 1,139 360 958 2,100 360 400 400 230 240 200 273 2,900 275 2,950 190 2,900 285 2,900 170 1,441 2,000 264 1,032 2,200 403 1,414 2,300 272 264 266 160 America Brazil Canada Mexico Peru USA 1,152 348 182 2,970 1,140 577 144 2,055 1,368 680 174 2,150 1,416 894 198 49 200 38 200 50 2,142 4,584 2,893 5,081 Asia India Japan Australia 59 3,084 759 53 2,741 543 34 2,566 648 36 2,794 608 160 968 165 120 444 96 1,181 174 Europe Belgium Bulgarlab East Germanyb Finland France West Germany Italy Netherlands Polandb Spain Sweden United Kingdom USSRb Africa Zaire 644 246 300 200 Sources: World Bureau of Meta! Statistics, US Bureau ofMines. a Preliminazy figures; b Estimate the US 9% less to 4,626 t, in the UK to 1,242 t (a loss of 12%), andin Japan to 771 (a loss of34%). In West Germany consumption rose only slightly, whereas in France an increase in demand was reported at 25% to 1,200 t. In constant values, the price of cadmium has declined over the past decade -the US producer price was 3 dollars/lb in 1964 when zinc was 10--12 cents/lb. Now it is less than 2 dollars/lb, while zinc is around 30 cents. In the past years, price fluctuations have had the greatest influence on cadmium consumption, but environmental considerations have now assumed greater importance [13].2 Use Electroplating is the largest single use of cadmium amounting to about 34% of all cadmium consumption. Presently it requires about half the cadmium used in the US and about 40% ofthat in the UK, but it is ofless importance in other European countries. became insignificant when legislation strongly reduced 2 Beginning in July 1980, Swedish government will ban the use of cadmium in electroplating, as a stabilizer and as a coloring agent; it is also working to reduce the Ievel of cadmium impurities in phosphoraus fertilizer (ES&T, Dec. 1979, p. 1447) 63 Cadmium the Iimit for cadmium in plating effiuents. Legislation has forced a number of European and US electroplaters to find alternatives to cadmium. However, strong traditional demands from the aerospace and telecommunication industries, requiring the superior resistance of cadmium to alkaline and humid atmospheres is expected to continue, although due to recent developments in zinc plating processes, this metal has replaced cadmium in some instances [13]. Small amounts of cadmium are used for the production of fungicides [16], control rods for nuclear reactors, fluorescent lamps, phosphors for television picture tubes, luminescent dials, compounds used in photography, lithographs, etc. [1 0]. 23% of cadmium consumed is in the form of pigments. In the plastic and ceramic industries cadmium sulfide and cadmium selenide are used; these have bright clear colors in the yellow-orange-red range and are light-fast up to 600 °C. With the increase in plastics production is isreasonable to expect a similar increase in the demand for cadmium-based pigments. Nickel-Cadmium batteries, originally introduced in the 1920's, account for about 15% of total consumption. Growth in use of some 6% per year is expected because of wider application of rechargeable batteries in hand-held electronic devices and in portable domestic electrical appliances. The most rapidly expanding application forthistype ofbattery is in calculators [15, 10]. Use of cadmium as a stabilizer in PVC presently accounts for some 15% of consumption and 8% of cadmium consumption goes for the production of solders, brazing materials, and other alloys. The low melting point cadmium alloys are used in fire prevention devices, and Cu-1% Cd is widely used for overhead conductors for trains, trams and trolleybuses, and occasionally for overhead telephone wires [13]. Current consumption trends have been determined by recent detailed studies in the FRG [14]. Investigations of sources provide results which are quite variable (Table 2). Since 1973, there has been a steady decrease in the production of stabilizers, batteries, galvanized products and cadmium-containing glass. The production of pigments increased slightly in 1974, but fell off and even decreased in 1975. A similar development was _determined_als_o_for _alloys_and_rectifiers, whereby _the _1975 values_for alloys remained 39% above that for 1973. The use of cadmium in galvanizing processes decreased markedly in 1975 [11]. Table 2. Consumption of Cd in the Fed. Republic Germany according to useage [14] Pigments Stabilizers Batteries Galvanizing Glass products Alloys Rectifiers By-products Cadmium salts lmpurities in Zn 1973 t Cd % 1974 t Cd % 1975 t Cd % 719 349 335 386 30 54 15 10 72 59 100 100 100 100 100 100 100 100 100 100 771 248 150 354 22 87 24 28 25 30 107 71 45 92 73 161 160 280 35 51 424 207 150 196 13 75 18 22 59 59 45 51 43 130 107 220 18 36 13 21 64 U. Förstner General Chemistry, Mineralogy, Geochemistry, Aquatic Chemistry Chemistry Cadmium has the atomic number 48, an atomic weight of 112.40 and consists of eight stab1e isotopes of the following relative abundances [17]: 111 Cd= 12.75%, 106Cd= 1.22%, 11 °Cd= 12.39%, 108Cd= 0.88%, 116Cd= 7.58%. 112Cd=24.07%, 114Cd=28.86%, 113Cd= 12.26%, Its specific gravity is 8.65g/cm\ melting point 320.9 oc and boiling point 767 oc. Cadmium, a white metal with a bluish tinge, was discovered in 1817 by Strohmeyer in Germany. It is soft,_ easily worked and is of considerable ductivity [10]. Like zinc and mercury, cadmium is a transition metal in Group Ilb of the periodic table of elements. Cadmium and zinc, however, differ from mercury in that the latter has 14 additional electrons in the fourth orbital, which probably accounts for the high stability of compounds of mercurycarbon bonds, whereas the similar alkyl-cadmium compounds are extremely unstable and react rapidly with water and moist air under normal environmental conditions [7]. As a result they are not expected to be of importance as environmental pollutants [10]. Cadmium and zinc show only valence + 2 in their compounds. These metals are also generally similar in reactivity, zinc being the more reactive, and cadmium showing a slightly greater tendency to form covalent bonds, especially with sulfur. The ionic radius of Cd2 + has been found to be 1.03 A (Goldschmidt, 1926) and 0.97 A (Pauling, 1927); the observed coordination number of Cd in its compounds are usually four and six, in a few cases also five, seven, eight, nine, and twelve. Cd-compounds are often isotypic with the corresponding compounds of Zn2 +, Mg2 +, Fe2 +, Co2 +, Ni2 + andin some cases ofCa2 + [18]. In air, cadmium vapor oxidizes quickly to cadmium oxide, and the metal dissolves in weak dilute acids, a property which has been responsible for acute oral intoxication in man. The sulfide (CdS), the carbonate (CdC0 3), the oxide (CdO), and the hydroxide (Cd(OH)2) are insoluble in water (negative logarithms of solubility products - pH = 7 at 25 oc - are 27.8 for CdS, 11.3 for CdC03, and 14.4 for Cd(OH)2 [19]. Cadmium sulphide is decomposed by acids with the Iiberation of hydrogen sulfide gas. The fluoride, chloride, bromide, iodide, nitrate, and sulfate of cadmium are relatively soluble compounds. Cadmium forms also a wide variety of soluble complexes, notably with cyanides and amines [10]. Mineralogy The common cadmium minerals are greenockite, hexagonal CdS; hawleyite = cubic CdS; otavite = CdC03; monteponite = CdO, and cadmoselite, hexagonal CdSe. Significant amounts of naturally occurring cadmium are found only in association with zinc ores, in which the amount of cadmium varies considerably. U sually 0.1--0.5% (maximum 5%) is present in zinc blende (sphalerite) and calamine (zinc spar, smithonite). Greenockite occurs usually 65 Cadmium as an earthy coating on zinc minerals, especially sphalerite; crystals found in amygdaloidal cavities in basic igneous rocks are rare [20]. Geochemistry Cadmium is a strongly chalcophilic element, i.e., it is concentrated in sulfide deposits together with zinc and mercury, and to a much lesser extent with lead and copper [7]. The abundance of cadmium in the earth's crust is generally estimated tobe 0.11 ppm [21 ].lts concentration is low in all igneous rocks and shows no clear relation to any major element, not even to zinc; the ratio Zn/Cd varies widely in all types of igneous rocks, with recorded extremes of 27-7,000 [7]. An accumulation of cadmium takes place in the sedimentary environment, in addition to the amount contributed by rock weathering. If the shale + clay concentration (Table 3) is taken tobe representative ofthe average abundance of cadmium in sedimentary rocks, a 2.4-fold increase is measured as compared to magmatic rocks of the upper continental crust. This accumulation has been explained as due to a degassing of the earth [21]. A further slight increase of cadmium concentration is observed in pelagic clay, which is partly explained as the high adsorptive capacity of sedimentary iron and manganese compounds [22]; Mn-nodules from the North and South Pacific were found to contain 8.40 and 5.06 ppm Cd, respectively [23]; in Fe/Mn concretions from Quinte Bay, Ontario, Cd concentrations of 0.8-6.4- average 3.0 ppm Cd were measured [24]. Compared to magmatic rocks and shales, concentrations Table 3. Cd in magmatic rocks, sediments, phosphorites, coal and oil [33] Example Granite rocks Basaltic rocks Ultramafic rocks Shales Pelagic clays Sandstones Limestones Limestones Limestones Diatomaceous ooze Globigerina ooze Radiolarian ooze Red clay Greenmud Calcaerous ooze Organic mud Phosphorites Coal Oil Cd (ppm) Ref. avg. 0.075-0.100 [21] 0.130 0.026 0.300 0.405 0.020 0.035 0.048 0.090 0.39 (n = 5) 0.42 (n = 3) 0.45 (n = 4) 0.56 0.27 0.57 0.39 -15 0.2-30 avg.-1 0.01-16 [21] [21] [21] [21] [26] [25] [26] [21] [23] [23] [23] [23] [23] [23] [23] [27] [30] [32] [ 7] 66 U. Förstner of cadmium in sandstones and Iimestones are diluted (Table 3). Due to the crystal chemical properties, the cadmium concentrations in Iimestones are expected tobe controlled by their clay fraction [21]; this explains the relatively large differences in the average values for cadmium in Iimestones shown in Table 3 [25, 26]. Generally, cadmium seems tobe enriched in some organicrich sedimentary rockssuch as dark shales, while depleted in others such as red shales relative to igneous and metamorphic rocks and the crust. Such enrichment occurs primarily through the adsorption and/or complexation of cadmium onto organic matter followed by the accumulation of organic debris in the depositional environment. Since such an environment will also be reducing, formation of cadmium sulfides during or after deposition can be expected [22]. Recent data on unpolluted environments are rather controversial. Different sediment types in pelagic deposits do not seem to affect cadmium concentrations (Table 3). Significant enrichment has only been observed from marine phosphorites [27]. For lake sediments higher concentrations of cadmium are reported from the organic fractions [28, 29]. It is assumed that gelatinous colloidal substances, which are formed from dissolved organic acids, spores, pollen and decayed leaves take up the metal ion from water. Cd concentrations in fossil organic substances are highly variable. Cadmium content of coal ranges between 0.04 and 30 ppm, with an average concentration of approximately 1 ppm [30-32]. Even larger variations are found for oil, but no average value can be given accurately. Aquatic Chemistry Reliable data on the distribution of cadmium in natural waters have only recently become available. Typical depth profiles from mid-Pacific sampling stations [34] show that cadmium together with phosphate and silicate is seawater) relative to the depleted in the surface water (approx. 0.01 ~g/k deeper ocean water ("' 0.07 ~g Cd/kg). Such a distribution indicates uptake by organisms at the surface and regeneration from sinking biologic debris deeper in the water column. The particularly high covariance of Cd with phosphate suggests that cadmium occurs in a shallow cycle like the labile nutrients, rather than deeper in the ocean in silicates [34]. Natural Cd concentrations in were deterfreshwater are about the same as in deeper seawater: 0.07 ~g/1 in water samples from the mined from Amazon River water [34], and 0.1 ~g/1 lower Mississippi River [35]. Differentiation of chemical species of metals has been performed by analytical procedures (see Section on Analytical Methods) and by computation of equilibrium models. Calculations of equilibrium solubilities with Cd(OH)2 or CdC03 indicate minimum solubility at pH 9.0-10.0 [36]. The pH/Eh systems for Cd+ S + C02 + H 20 is given in Fig. 2 after Hem [36]; the and those of activity of dissolved Cd is 1o-7·05M, which is equivalent to 10 ~g/1, dissolved carbon dioxide and sulfur species are 10-3 M. In the system as or 10 defined, cadmium solubility is below the actual standard limit (5 ~g/1 last section) only at high pH (between pH 8.9 and 10.7) or in reduced ~g/1; 67 Cadmium 1.2 1.0 0.8 0.6 Cd 2 + 0.4 0.2 0.0 -0.2 -0.4 -0.6 Eh (V} pH 2 4 6 8 10 12 14 Fig. 2. Fields of stability of solids and predominant dissolved cadmium species in the system Cd+ C0 2 + S + H 20 at 25 oc and 1 atm pressure in relation to Eh and pH. Dissolved cadmium activity 10·7·05 molesjl; dissolved carbon dioxide and sulfur species 10·3 moles/1 (Hem (36]) systems in which oxygen is severely depleted [36]. Most natural waters are unsaturated with respect to hydroxide or carbonate; about 20% of the waters have carbonate contents in reasonable agreement with values calculated assuming CdC0 3 as the equilibrium solid phase [7]. Computermodels have provided the following data for inorganic speciation ofCd in aquatic systems (listed according to increasing pH): Cd2 +, CdC0 3(s), Cd(OH)2(s), accounting for more than 90%, CdS04 and Cd Cl+, accounting for a few percent [37]. Dissolved Cd species in aerated seawater (CdC0 3 as solid species) are calculated as follows [38]: CdCl+ =56%, Cdl~= 15%, CdI~-= 10%, CdClt-=9% other calculations: Cdl~=38%, CdCI+ =29%, CdC13=28% [39]; Cd1~=50%, CdCl+=40%, CdCl3=6% [40]. In reducing marine environments Cd(HS)0 is the predominant dissolved species [37]. In oxidizing fresh waters sulfato- (45%), carbonato- (42%), and chlorospecies (13%) represent the dissolved fraction of Cd, whereas in the respective reducing environment free (aquo- = 88%) and chloro-species (11 %) are calculated as the dominant dissolved species [37]. However, organic ligands can play a major role in natural waters by complexing trace metal ions and keeping them in solution [41]. In addition, organic 1igands are likely to mediate large interactions among metal ions; for example, the presence of NTA couples the free concentrations of copper and cadmium [37]. Labaratory experiments on natural samples suggest that "labile forms" (free cationic species and complexes with low stability constants, both organic 68 U. Förstner and inorganic) of Cd predominate in natural and polluted freshwater systems [42, 43]. Because ofthe tendency ofthese species tobe sorbed on particulates, there is the need to determine the cadmium content of the suspended matter and sediments for assessing the sources, distribution, and fate of Cd contamination in aquatic systems [6, 33]. Analytical Methods Determination ofCd in air, water, food, and organisms has been performed by different methods. Some major methods arecolorimetry (dithizone method), emission spectroscopy, atomic absorption spectrophotometry (flame and flameless), electrochemical methods (anodic stripping voltametry, polarography), X-ray fluorescence methods, neutron activation analysis, isotope dilution, spark source mass spectrometry, and fluorimetry. Some ofthese methods have been summarized by Fleischer et al. [7], and are partially reproduced here in Table 4 (for references, see original paper by Fleischer). Currently, Table 4. Cadmium analyses [7) Material Method Granite and diabase rocks Atomic absorption, mass spectrometry, neutron activation, optical spectrography, polarography, spectrophotometry Water Atomic absorption, spectrophotometry, polarography, stripping voltammetry, x-ray fluorescence Air Low-temperature ashing of glass fiber filters, atomic absorption Neutron activation Anodic stripping voltammetry Food Dry ashing, atomic absorption Shellfish Homogenization, wet digestion, atomic absorption Blood Wet ashing, dithizone extraction, atomic absorption Dry ashing, dithizone extraction, optical spectrography Blood, tissues, and hair Wet digestion, anodic Stripping voltammetry Renal tissue Drying, neutron activation Dry ashing, atomic absorption Urine Dithizone, atomic absorption atomic absorption spectrometry is probably the most widely used method for Cd analyses [44]. Anodic ·stripping voltammetry is ·extremely sensitive ·and ·is especially useful for determinations in natural waters and its use may yield information on the nature of binding of cations in water ([7]; see section on Aquatic Chemistry). A neutron activation method has been developed for Cd determination in vivo for organs, mainly the liver [10, 45]. Pretreatment ofthe 69 Cadmium samples is often required either for concentration (to improve sensitivity or to eliminate interferences, such as for NaCl matrices) or for extraction of distinct solid or aqueous species. For such differentiation in natural waters analytical schemes have recently been proposed [42, 43], which, however, need further development and standardization (e.g., Chelex-100 resin for the differentiation oflabile and more stable species oftransition metals). Digestion of biological samples and sediments is often performed with conc. nitric acid [46] or HC1-HN03 = 1:1 [47]. The aqua regia digestion (HCl:HN03 = 3: 1) is commonly used in the extraction of metals from polluted sediments. The fact that more volatile elements such as cadmium are not boiled as is the case for hydrofluoric acid in combination with nitric, perchloric or sulfuric acids is a major advantage in their analysis. A compilation of laboratory methods which have been proposed to measure Cd in soil, sewage sludge, and sediments, with special emphasis on methods for the assessment of cadmium available to plants, has been given by Symeonides and McRay [48] and is partially reproduced in Table 5. Other extraction processes for determining chemical associations of cadmium in particulate phases will be discussed in the section on chemical reactions. Table 5. Suggested methods for assessing Cd in soil [48, 49] Extraction solution Strength Ref. Acetic acid Hydrochloric acid 2.5% (0.41 N) 1N 0.1 N 1N 2N 1 N,pH7 [50] [51] [52] [53] [54] [51, 53] 1 N,pH4.8 1 N, unbuffered 1N 0.05 M,pH6 0.5 M,pH 6.5 [53] [48] [54] [49] [55] Nitric acid Ammonium acetate Ammonium acetateacetic acid buffer Ammonium nitrate Ammonium chloride Calcium chloride Sources, Pathways, and Reservoirs in the Environment With the attempt to determine the material flow of cadmium in the environment it becomes clear that complete knowledge of the processes involved is lacking. In the following section we refer to the detailed survey compiled in respect to the fluxes in US and adjacent marine waters by A.F. Sarofim in the framework of the Subpanel on Cadmium, US National Institute of Environmental Health [7]. Estimates by Davis et al. [56] on atmospheric emissions and data from Chizikhov [57] on Cd fluxes in smelters and refineries are the basic information source for this study. The most important data is compiled in Table 6. U. Förstner 70 Table 6. Estimated rates of emission of cadmium during production and disposal of cadmium products for 1968 in the USA [7] Losses during use and disposal Primary Electrocadmium plating, propigment duction, and plastic Uyr formulation, Uyr Coal and oil combustion Uyr Air contamination 955 120 Water contamination 3,000 240 Mining and ore concentration, Uyr Cadmiumplated metals, Uyr Pigment, plastics, and miscellaneous, Uyr Alloys and batteries, Uyr 500 90 40 1,420 2,080 380 500 490 220 300 Soll contamination (140: Phosphate fertilizers) Accumulation in service Land disposal (dumps, land füls, slag pits, mine tailings) 300 310 360 Sources Coa/ and Oil Combustion, Cement Production, Incineration. lf it is assumed that for the 455 million tons of coal consumed in 1968, the averagepartiewate collection efficiency was 80% and the average cadmium concentration was 1.0 ppm {Table 3), an estimated 100 t of cadmiumwas emitted forthat year [56]. Cement manufacture will provide approximately 3 tfyr, assuming that 80% of the emissions were captured in partiewate collection devices, and approx. 100 t of cadmium were released during incineration of waste products in the same year [56]. With the closing of open-burning dumps and stricter air pollution regu1ations on incineration, this source of emission showd be significantly reduced [7]. Industrial Emissions. Ofthe major industries employing cadmium, e/ectroplating shops, pigment plants and producers of al/oys and batteries can be expected to be major sources of cadmium pollution. The contribution of Cd to New York City wastewater treatment plants (approximately 25 t year) is shared by the following activities [60]: electroplaters 33%, other industrial sources 6%, stormwater runofT 12%, residential waste 49%. 71 Cadmium From non-metallurgical industries the following metal concentrations in wastewaters (!J.g/1) were recorded in the same study [60]: laundry 134, for dressing and dyeing 115, ice cream production 31, textile dyeing 30, miscellaneous chemieals 27, car washing 18, fish processing 14, meat processing 11 (mass fluxes are not noted). Sewers. Cd input to sewers stems both from industrial processing and domestic sources (see above). Major domestic sources are atmospheric fallout on residential areas- which can produce particularly adverse effects by shock load during stormwater events [61]- and corrosion of zinc-containing roof fittings and household pipes [62]. On the average, approximately 50% of the cadmium discharge in the waste stream is retained within thesewer [63]. Phosphate Fertilizers. Significant amounts ofCd will reach the agricultural soils during application of phosphate fertilizers, which, on the average, contain approx. 5 ppm Cd [64]. For the entire US an input of 140 t yr has been estimated ([7] Table 6). The cadmium contents in superphosphate fertilizer is normally in the range of 2-50 ppm [1 0], but concentrations as high as 5~ 170 ppm have been determined [64]. At moderate Cd levels (5-10 ppm) in phosphate fertilizers no significant correlation between the rates of application and cadmium concentrations in the plow layer soil was found after 20 years [65]. Pathways Majortransport of Cd from the various sources to the reservoirs (see appropriate sections) and eventually to organisms can be via atmosphere, water, suspended sediment, land application (of sewage material and phosphate fertilizer) and waste dumping. Table 7. Selected examples of Cd concentration in the atmosphere (see [10]) Northern Norway Swiss Alps Erlangen (FRG) Munich (FRG) Tokyo (Japan) EI Paso (USA) 0.10 ng/m3 0.28 ng/m3 1.50 ng/m3 6.90 ng/m3 10-53 ng/m3 120 ng/m3 Atmosphere (Table 7). Naturalbackground values lie below 0.1 ng Cd/m3 air. In urban areas an average of approximately 2 ng Cd/m3 has been determined [7]. Under unfavorable conditions these values may increase by an order of magnitude. Even higher concentrations, up to 500 ng Cd/m3 and more have been recorded in the vicinity of Zn smelters [10]. Water. A compilation ofwater data was made by Fleischer [7]; more recent examples from freshwater bodies are given in Table 8. The highest concentration of cadmium in surface waters is usually found in areas of high U. Förstner 72 Table 8. Se1ected references for Cd concentrations in river waters 0.0711g/1 0.10 llg/1 0.80 llg/1 5.5 .IJ.g/1 [34] [35] [69] [70] Meuse River (Be1giurn) 10.4 f.ig/1 Neckar River max.220 llg/1 [71] [72] Amazon River Mississippi River Missouri River Lower Rhine River (FRG) (FRG) Coeur d'A1ene R. (Idaho) max. 450 llg/1 [73] population density [66]. In a study sponsored by CIPS in Belgium, the following distribution of cadmium concentration in 480 surface water samples has been determined [10]: < 1 Jlg/1-35.4%, 1-5 Jlg/1-22.5%, 6-10 Jlg/1-10.8%, 11-20 Jlg/1-13.8%, 21-30 Jlg/1-7.7%, 31-40 Jlg/1-6.2%, 41-50 Jlg/1-3.6%. With an upper limit of 5 Jlg Cd/1 for drinking water and water for irrigation purposes approximately 40% of these cannot be used. This allowable limit is also valid for surface waters reclaimed by conventional methods, such as bank filtration and artificial groundwater recharge. The average concentration of cadmium in water from public water supplies in 7 large cities of the European Communities was 1.1 Jlg/1 with a range of0.2-4.0 J.lg/1 [67]. These values, however, apply mainly to the water quality at the pumping station and the actual concentration of cadmium at the tap can be expected tobehigher [10]. The galvanized pipes which are sometimes used in plumbing are a potential source of cadmium in drinking water [68]. If the water is soft and somewhat acid the cadmium can conceivably remain in solution. Suspended Sediment. While the concentrations of Cd in water seem to be largely unaffected by the water discharge (despite flushing effects at the onset of stormwater runofi), this parameter is of great influence for the transport behavior ofCd (and other trace metals) in suspended particles. Figure 3 gives an example from the middle section of the Rhine River near Koblenz; values for the springfsummer period are marked by open circles, those for the autumnfwinter period by dots. For both categories, a clear depencency on the river flow can be seen. However, even the moderate Cd-concentrations in the winter period exceed the maximum values for the spring/summer period, despite the much higher water discharge. Such a development is due to the suspended sediments rich in cadmium being heldback in periods oflow flow, e.g., in Germany mainly in the summer period in the lock-regulated Neckar and Main Rivers; in autumn/winter the tendency is reversed when the material is carried by high water flow into the Rhine in increased quantities. Data on Cd concentrations on particulates from natural and polluted aquatic systems are listed below ("sediments"). 73 Cadmium E a. a. .. ~ October- December 12 • E ••• 'C GI Ul 'C GI 'C 8 .• •.. c • GI a. Ul :I Ul c 4 May- October 'C () 0 1000 2000 3000 water discharge 4000 (m 3/s) Fig. 3. Water discharge vs. cadmium concentration in suspended sediments of the Rhine River near Koblenz (Schleichert [74]) Sewage Sludge. Despite large local variations, which are mostly due to industrial input, the average (median) values of Cd in sewage sludges from different countries, such as Sweden [75, 76], USA [77, 78], and the FRG [79, 80] have a narrow range of 7-15 ppm. It has been calculated that annual application of a few tons of sewage sludge containing 20 ppm (and more) to unpolluted agricultural soils will raise the concentration ofthe ploughed layer ofthe soil to levels between 1.2 and 6 ppm Cd [81, 10]. Waste Dumping. Disposal of metal-bearing waste material may affect aquatic biota and groundwater quality. Investigations on soils beneath sewage sludge, effiuent disposal and stormwater retention ponds showed movement ofzinc, cadmium, copper, and chromiumin various manners [82, 83]. It was shown that the distribution of metals with depth was closely related to changes in chemical oxygen demands, suggesting that the metals moved as soluble metal-organic complexes (see Section on Chemical Reactions). Waste waters infiltrated into the substratum under a zinc processing plantat Nievenheim in the lower reaches of the Rhine River led to an increase of Cd concentration of up to 600 J.Lg/1 [84]. Similar effects were reported from sanitary landfillleachates [85, 86]. Waste material is discharged to the sea via sewer outfalls and barges. From an annual input of approx. 50 t of cadmium into New Y ork Bight, 82% is delivered by barges, 5% from municipal wastewater effiuents, 2% stems from atmospheric sources, and approx. 10% is introduced by surface runoff [87]. Similarly, approximately 50 t of cadmium are annually discharged into the Southern California coastal zone from the five largest municipal effiuents; about two-thirds ofthat reach the sea via Hyperion's Joint Water Pollution Control Plant [86]. Maximum factors of enrichment of dissolved trace metals 74 U. Förstner in the main outfall plume, as compared with the surface water composition, are 13 for lead and zinc, 8 for copper, and 4 for cadmium [89). To show what quantities of cadmium and other heavy metals from anthropogenic sources enter this relatively confined area, it is pointed out that the Mississippi River carries in its suspended substances approx. 50 t of cadmium - mainly from geochemical sources- into the Gulf of Mexico each year [90]. Reservoirs The various reservoirs for cadmium contamination can be characterized by the mean residence times and the concentration. tCd has been evaluated for natural environments as follows [91]: air, 20-30 days; river water, a few days; lake water (example: Lake Washington), 1-2 yr [92); humans, 20-30 yr [6]; soil (example: upper Thames valley) 280 yr [91); ocean water, 250,000 yr; pelagic sediment, 2-5 x 108 yr [93). Concentrations of cadmium have been measured: air, 0.1-500 ng/m 3 ; water, 0.01-42,000 Jlg/1; aquatic, organisms 0.001-1120 mgfkg; soil, 0.01-500 ppm; sediments, 0.01-50,000 ppm. 1t has been proved effective in making conclusions as to the origin of cadmium enrichment in a certain reservoir to refer to the Cd/Zn ratio. This ratio in contaminated samples is frequently below that in geological reserves, due to the selective vaporization of cadmium during incineration processes. This is particularly valid for smelter emissions. Values for smelter recovery efficiencies of75% for cadmium and 89-97.5% for zinc have been quoted [94], suggesting that, for an ore with a Zn/Cd ratio of 200, the ratio in the unrecovered portions (losses in the atmosphere and the slag) will range from 20 to 88 [7). As the level of cadmium increases in a distinct compartment, generally the Zn/Cd ratio decreases. Among the major reservoirs for cadmium contamination which may be available to organisms, soils, stagnant waters, and sediment are considered now in more detail. Soil. A selection of representative data on the cadmium content of soils is given by Fleischer [7]. Recent analysis of contaminated soil profiles indicate that the normal, average content of cadmium in soils is about 0.4 ppm. There is a clear increase near highways [95). Typical regional differences are shown from investigations performed by Klein [96) in the Grand Rapids, Michigan area: average Cd concentrations in residential areas was 0.41 ppm, in an agricultural area 0.57 ppm, in a industrial area 0.66 ppm and near an airport 0.77 ppm Cd. In these experiments each sample was collected from the top 5 cm ofsoil. 75 Cadmium Table 9. Cadmium content and Zn/Cd ratios in uncu1tivated soll surrounding East He1ena Stack. (From Miesch and Huffmann [97], modified by[7]) Depth of soll, cm 1.8 kmfrom stack Cd Zn/Cd ppm 0-2.5 5-10 15-25 68 30 3 16 33 70 3.6 kmfrom stack Cd Zn/Cd ppm Z2 kmfrom stack Zn/Cd Cd ppm 17 7 2 4 2 1 14 25 42 12 15 33 The largest contamination originates from smelters and metallurgical plants, and it has been shown that a major fraction of the emissions from smelters accumulates in the surrounding soil [97]. Data from an extensive study of the pollution in areas near lead and zinc smelters in Bast Helena, Montanaare summarized in Table 9. A significant decrease of Cd concentrations in all three profiles is recorded. Simultaneously the Zn/Cd ratios increase, but are lower than those in geological reserves. Low Zn/Cd ratios were also observed in settled dust and suspended particulate, supporting the postulation that the relatively low Zn/Cd ratio in the soil reflects the low ratio at the source and not a selective leaching of zinc from the soil [7]. The effects of phosphatic fertilizers are obviously less important than generally thought. Kloke [16] has estimated that the increase in cadmium concentrations of soil resulting from the application of phosphatic fertilizers (50 kg P20 5/ha) would be at the most 0.016 ppm per year ([10]; seealso the chapter on availability). On the other hand, the application ofsewage material and polluted stream sediments can significantly effect the Cd-content of soils. Water. Obviously, the physico-chemical conditions in normal surface waters prevent large-scale dispersion of dissolved cadmium, even in cases where strong contamination is observed from sediment studies (see be1ow). Only in acidic waters, e.g., from mine tailing effluents, has significant enrichment been found at greater distances from the source. The Cd concentrations in most lake waters studied are relatively less enriched, even for examples with significant overall pollution effects. Average concentration of Cd in Lake Ontario has beendeterminedas 0.09 Jlg (range: 0.03-0.15 Jlg/1) by Chau and colleagues [98]; for Lake Michigan 0.30 J..Lg/1 Cd has been recorded [99]. It seems that during periods of higher biologic activities the Cd-concentrations may fall below the general geochemical background values. At the same time, the input from polluted rivers can significantly affect the Cd contents of coastal waters. Investigations performed by Abdullah and co-workers [100] on the distribution of transition metals in Welsh rivers clearly reflect the influence of mineralization zones. The rivers and lakes in regions where no mineral deposits are known show cadmium levels ranging between 0.1 and 0.6 Jlg/1, whereas the annual average cadmium levels in rivers of the mineralized regions are found to range between 1.2 and 4. 7 Jlg/1, with the highest recorded concentration being 20 Jlg/1. These waters characteristi- U. Förstner 76 Cd March 1971 s• Fig. 4. Cd concentrations in surface waters ofthe Severn Estuary, Cardigan Bay, Liverpool Bay, and the adjacent Irish Sea [101] cally influence the composition of the adjacent Irish Sea (Fig. 4). The highest concentrations of cadmium in the Bristol Channel are probably derived from industrial effluents entering the area from the Avon and Severn Estuary. In Cardigan Bay, however, which is relatively free from industrial effluent, little Table 10. Mean values of Cd in coastal waters of Great Britain [102] English Channe1 Atlantic Ocean (Iceland-Faroes Ridge) Irish Sea (offshore) Liverpool Bay North Sea (nearshore) Firth of Clyde Conway Bay (nearshore) Cardigan Bay Bristol Channel 0.06 f.Lil [103] 0.07 f.L/1 0.11 f.L/1 0.27 f.L/1 0.5 f.L/1 0.5 f.L/1 0.76 f.L/l 1.11 f.L/1 1.94 f.L/1 [104] [103] [101] [105] [106] [107] [101] [108] 77 Cadmium domestic waste is produced due to a low population density; runoff from the mineralized zones and sites of former mining activity is the main source of cadmium and of other trace metals [101]. A compilation of data from coastal waters off Great Britain (Table 10) shows particularly strong enrichment of dissolved cadmium in the Firth of Clyde, Conway Bay, Liverpool Bay, Cardigan Bay and Bristol Channel, which amounts to more than a 10-fold increase compared to the normal values from the open ocean or even from the Irish Sea [102-108]. Sediment. Analyses of sediment are useful when selecting critical sites for routine water sampling. Their profiles (cores) often unique1y preserve the historical sequence of pollution intensities, and lateral distributions (quality profiles) are used to deterrnine and evaluate local sources of pollution [2]. A general description of sediment properties is given in Volume 1 of the Handbook; a chapter on "Cadmium in Poiluted Sediments" has been prepared for Table 11. Cd contamination in natural and polluted sediments Investigation site Cd value (ppm) Sources Ref. Rudson River, Foundry Cove (N.Y) Palestine Lake (Indiana) Derwent Estuary (Tasmania) Los Angeles River (Calif.) Sörfjord (Norway) Vesdre River (Belgium) Hitachi area, NE Tokyo (Japan) River Tawe (Wales) Neckar River (ER.Germany) Takahara River (Karnioka Mine, Japan) Meuse River (Belgium) Tennessee River (Tennessee) South Esk River (Tasmania) Main River (ER.Germany) Milwaukee River (Wisconsin) New Bedford Rarbor (Massachusetts) Corpus Christi Rarbor (Texas) River near Himeji City (Osaka, Japan) Gadura River (Israel) Stola River (Poland) River Conway (Wales) Ginsheimer Altrhein (ER.Germany) Coeur d1\lene River (ldaho) Upper Rhöne River (Switzerland) Voglajna River (Yugoslavia) Santa Monica Canyon (Calif.) max. 50,000 3.2-2,640 0.8-862 max. 860 16-850 max. 430 max. 368 max.355 max. 340 4.1-238 max. 230 max. 227 max. 153 max. 151 rnax. 149 max. 130 2-130 max. 129 max. 123 max. 116 3-95 2-95 max. 80 0.1-73 max. 66 max. 65 Cd-Ni battery Electroplating Zinc smelter [111] [112] [113] [114] [115] [71] [116] [117] [118] [119] [71] [120] [121] [122] [123] [124] [125] [126] [127] [128] [129] [130] [131] [132] [133] [134] Unpolluted lakes in South America, Asia, Africa, and Australia (n = 72) avg. 0.35 (0.04-0.84) [109] Very slightly influenced sea sediments (Japan) avg. 0.45 [110] Mississippi River (clayey sediment; n = 4) avg. 0.49 Pb-Zn smelter Braun tube factory Metal processing Pigment industry Mine effiuents (Suspended solids) Mine effiuents (Suspended solids) Metal processing Battery plant Zn smelter Mine effiuents Mine effiuents [90] 78 U. Förstner J. 0. Nriagu's "Cadmium in the Environment" [12, 33]. Table lllists examples ofless polluted sediments (background data) from lacustrine, marine, and fluviatile environments and examples of sediment investigations on 25 of the most heavily polluted aquatic systems. Various sources- effiuents from mine tailings, processing plants, smelters, the pigment industry, battery plants, the electroplating industry, and from a Brauntube factory- are responsible for very strong accumulations of cadmium in aquatic sediments. Core sediment studies from moderately polluted lacustrine and coastal marine environments indicate that next to mercury, concentrations of cadmium generally prove tobe the most important metal enrichment of all heavy metals under investigation [33]. The order of enrichment in anthropogenically influenced sediments Cd > Pb > Zn > Cu found in these studies corresponds to the accumulation ofmetals in fossil fuel residues [135], and the above sequence is also reflected in the metal enrichment of air-bome particulates. The pollution by cadmium in many aquatic systems is apparently still increasing. By analyzing sediments collected since 1922, Salomons and DeGroot [136] were able to trace the development of metal pollution in the Rhine River in the Netherlands over a period of more than 50 years (Fig. 5): the samples taken from the flood plain at the beginning of this century showed already anthropogenic influences of Iead, cadmium, and mercury, as can be shown by comparison with sediment data from polders reclaimed in the 15th and 18th centuries. Between 1920 and 1958 all trace element concentrations studied have increased in the sediment from the Rhine River. Whereas the concentrations of mercury and lead decreased between 30 Cd .. 10 c Cll E -a .... Cll Ul c ,~: .. ......,·~,,, .. ....·· ,,' ia 3 Cll E E Q. Q. ,,",, ',,..., ,,,,"" ,,,''Hg ,,,, ,, ,, _,i'"···· ,, . ' .············································ ......... Pb X 100 ........··· ,,,,' "" " " ±1900 1920 1940 1960 year Fig. 5. History oftrace metals in sediments from the Dutch Rhine River. (After Salomons and De Groot [136]) 79 Cadmium 1958 and 1975, cadmium continued to increase up to a level100 times greater than that in sediments of Rhine polders reclaimed in 1977. Biota. In order to determine the major compartments of an enrichment of cadmium, biological systems cannot be ignored. Because of their generally greater biomass, plants are of special interest; their capacity for storing cadmium is normally an order of magnitude greater than the corresponding Table 12. Estimates of cadmium concentrations in some plants and plant parts [7] Reported or estimated cadmium concentrations or ranges in concentrations; ppm in dry material Plant or plant part Environments presumably having normal cadmium Ievels Environments having greater than normal cadmium Ievels Marine algae Mosses (bryophytes) Lichens (fruticose type) Grasses Alfalfa Grains Com (Zea mays) Rice (polished) Barley, wheat, and oats Vegetables Asparagus Beetroot Cabbage leaves Carrots Chinese cabbage Eggplant fruit Kaie 0.1 -1 0.7 -1.2 0.1 -1.4 0.03-0.3 0.02-0.2 n.d. 8-340 1 0.6-40 0.2-2.4a 0.1 n.d. 0.1 -0.5 2 0.5 O.l-1.5b n.d. 0.05 0.05 <0.35 n.d. n.d. 1 4 0.24 6-12 8 41 8 n.d. Leafy vegetables used as pot herbs or salads 0.3 -0.5 3-50 Leeks Lettuce Potatoes Spinach Tumip, roots leaves Tomatoes Trees, deciduous leaves stems (branches) Epiphytes (Spanish moss) Floating aquatic plants (duckweed) Marine flowering plant (Zostera marina) n.d. 0.3 -0.5 0.05-0.3 0.6 -1.2 n.d. n.d. n.d. 14 4-16 6-20 n.d. 5 15 2 0.1 -2.4 0.1 -1.3 0.1 4-17 0.03-1.5 1 n.d. 17 0.23 n.d. n.d. = no data; a = Original data given in wet weight; converted to concentration in dry material ba assuming 25% water in original sample U. Förstner 80 reservoir capacity of animal organisms (see section on cycling of cadmium in natural systems). The latter will be discussed in more detail in connection with aspects of metabolism and food chain behavior. A detailed review of the cadmium concentrations in plants is given by Shacklette [137] and Shacklette and Nisbet in Fleischer et al. [7]; Table 12 is excerpted from the latter study. Important observations are summarized as follows: (i) Where environmental cadmium Ievels are low, cadmium concentrations in plant tissue frequently vary more with species than with soil type. (ii) Where cadmium content of soils are higher than background amounts, the cadmium content of plant tissue tends to increase with increased concentrations of soil cadmium. (iii) The cadmium contents of many species of plants reflect above-normal amount of cadmium that are introduced into the environment both from natural sources [138] and from cadmium pollution of soils [139], water [4] and air [140]. Cadmium from these sources may be absorbed by the plant through roots or leaves or both, and thus be incorporated into the tissues. Airborne particulate matter containing cadmium may be deposited on the surface of 1eaves ([7]; see section on biological uptake and accumulation). Cycling of Cadmium in Natural Systems Examples for the transfer pathway of cadmium in the environment are reproduced from results ofinvestigations undertaken by Windom et al. [141] in a coastal marine ecosystem. The annual rate of input of cadmium by nine major river systems emptying into the Georgia Embayment was determined based on analyses of river water samples collected bimonthly and by integrating the meta1 concentrations with flow rates [142]. Once these metals are Annual input by rivers -- 52·10 3 kg (21% particulate) l I Loss to sediments (17%) ~ Transfer through f-------- Through higher organisms spartins r ~ Transfer through estuary (83%) Transfer through uca (0.2%) 1 (3%) ~ Transfer through littorina (0.5%) ' Fig. 6. Transfer pathways of cadmiurn through the estuarine zone ofGeorgia Embayment [141]. Estimated percentages of the total input passing through biological compartments are given in parentheses Cadmium 81 delivered to the estuarine zone they follow various pathways within the environment entering major biologic and nonbiologic reservoirs (Fig. 6). Once the sedimentation rates in the salt marshes are known, the rate of loss of the metals to the sediment can be determined. After the production rates of major biological components of the estuarine zone and their metal concentration is found it is possible to determine to what extent the biota transfers metals through this interface. For example, the annual production rate of Spartina a/terniflora, the major primary producer in the estuarine zone, is approximately 700 g/m2 dry weight; approx. 3% of the total cadmium annually delivered by rivers is transferred through marsh grass. A similar approach can be taken using the major primary consumers (Littorina and Uca), since data exist on their annual production [141]; these transfer routes are less important than for the primary producer (0.5% and 0.2%, respectively). The data show that the loss of cadmium to sediments is roughly equivalent to the total amount of Cd supported by rivers in particulate form, whereas the soluble fraction is ultimately transferred through the estuarine zone either in solution or in biological compartments. lt has been demonstrated by Windom et al. [141] that the importance of the biological cycles of cadmium in the estuarine system is much less than for mercury; transfer of Hg through Spartina was found to be 17% of the total mercury input into the system. Chemical Reactions: Sorption and Release of Cd on Particulates The availability of trace metals for metabolic processes is closely related to the chemical species involved, bothin solutionandin particulate matter. The type of chemical association between metals and particulates has therefore become of interest in connection with problems arising from the disposal, i.e. land application, of sewage sludge and contaminated dredged sediments. In this respect cadmium presently seems to pose the greatest problems of all metals [33]. Leaching Methods In pedo1ogy, for the determination of plant-available concentrations of cadmium and other heavy metals (e.g. Cu, Ni, and Zn) numerous leaching tests have been introduced. Table 5lists several reagents that are important for the assessment of plant-available cadmium in soils. Recently, attempts are made to standardize extraction procedures with respect to the use of sludgejsoil mixtures [143]. Chumbley [144] proposed that 0.5 N HOAc be used to assess zinc and nickel availabilities, and 0.05 MEDTA (ethylene-diamine-tetraacetic acid) to assess copper availability for regulating sludge application to soils. 0.05 M EDTA has been used for cadmium extraction, for example, by Webher [145], Davies [146, 147] and Symeonides and McRae [48]. Similarly, 0.005 M DTPA (diethylene-triamine-penta-acetic acid) has been used to determine the solubilities in soils ofboth nutrient metals [148, 149] and non-nutrient metals, such as Cd [143, 150, 151]; the proposed DTPA-solution by Lindsay and 82 U. Förstner Norvell [148] contains 0.005 M DTPA, 0.01 M triethanol amine, and 0.01 M CaC12 adjusted to pH 7.3. The 0.01 M CaC12 also is used separately as neutral salt solution and seems to provide an index of metal solubility in the soil solution, which, at least for zinc, is related to plant uptake [152]. Increased Ievels of CaC12-extractable cadmium from the sludgefsoil mixtures with heavily contaminated sludges indicate a potential movement of this metal into groundwater [143]. In a comparative study on the relationship between Cd in soils and radish plants it was shown [48] that the most sensitive of several possible indices to Cd uptake by plants is the amount extracted by a 1-h shaking with 1 N ammonium nitrate solution at a soil/solution ratio of 1:10 (wtfvol). Generally, however, these correlations do not give information from which chemical compound the metals are preferably adsorbed into the plants. Plant-available metal concentrations in particulates may be investigated by determining the chemical association of cadmium by successive chemical leaching processes. A number ofthese methods are summarized in Table 13. Both extraction procedures were done, for comparison, on a heavily Cd-contaminated sediment sample (14.8 ppm Cd) from the Neckar River Table 13. Extraction of Cd-forms in particulate matter (examples) Chemical fraction Leaching method Soluble fraction H20; elutriate test (disposal site water) [153] Soluble organics/ soluble free cations Supematant solution is passed through cation exchange resin [154] Exchangeahle cations + easily extractable phases 0.2 MBaCI2 triethanolamine pH 8.1 1 M sodium acetate, pH of suspension 1 N ammonium acetate, adjusted to sedimentpH [157, 158] Reducible phases 0.3 N HCI [159] Carbonate fraction C02-treatment Acidic cation exchange resirr [156] [160] Easily reducible fraction (carbonate, Mn-oxide, amorphaus Fe-oxide) 0.1 Mhydroxylamine hydrochloride + 0.01 M HN03 0.15 M oxalic acid + 0.25 M ammonium oxalate [161] Moderately reducible phases (hydrous Fe-oxide) Sodium dithionite/citrate 1 M hydroxylamine hydrochloride + 25% acetic acid ("acid-reducible agent") Oxidizable phases 30% H202 (95 °C), extraction with 1 N ammonium acetate 0.5 N NaOH, 0.1 N N aOH Benzene/dichloromethane/methano1 Sodium hypochlorite, citratedithionite extraction Humic acids Lipids, asphalt Organic residues Ref. [155, 156] [162, 154) [163] [164] [158] [165, 166] [167] [168] 83 Cadmium [169]. The determination of the compound phases by successive chemical leaching according to the methods compiled in Table 13 had the following results: Soluble fraction Cation exchange Carbonate fraction Easily reducible fraction Moderately reducible fraction Residual fraction disposal site water 0.2 N BaCI2 -triethanolamine acidic cation exchangers 0.1 MNH 20H · HCl + 0.01 MHN0 3 0.3 NHCl HF/HC104 0.2% 28% 48% 12% 4% 8% The pedological experiments (simultaneous) showed the following Cd amounts being released: 1 M ammonium nitrate unbuffered pH 4 I M ammonium acetate acetic acid pH 7 Mixed reagent 0.005 M DTPA, 0.1 M TEA, 0.01 M CaC12 2.5% 17% 38% The exact amounts of metal actually being extracted and the preferred chemical phases from which the plant-available Cd originates have not yet been determined. From the available data, however, it can be clearly seen that there is a general decrease ofthe residual bonding forms of cadmium (and of other heavy metals), i.e., the predominantly inertly fixed cadmium content as the anthropogenic metal enrichment increases [170, 171]. There is a characteristic affinity of cadmium for organic substances and sulfides in polluted sediment at lower carbonate concentrations. At higher carbonate contents the association with carbonate minerals - either as discrete carbonate or as coprecipitates with calcite seems to provide the major process of immobilization of elevated Cd concentrations in the effiuents [171, 172]. Remobilization Processes Trace metals temporarily immobilized in the bottom sediments and suspended matter of aquatic systems may be released as a result of physicochemical changes such as: (i) increased salinity, (ii) lowering of pH (iii) increased input of organic che1ators, (iv) microbia1 activity, and (v) change in the redox conditions. Increased salinity in a water body leads to competition between dissolved cations and adsorbed trace metal ions and can result in partial replacement of the latter. Such effects can be expected particularly in the estuarine environment [173, 174]. Experiments performed by Van der Weijden [175] with artificial seawater indicate desorption oftrace metals from particulate matter, presumably by inorganic complex formation, which was highest for cadmium. From experiments on desorption of metals from sludge material diluted in seawater Rohatgi and Chen [176] found that 93% of the original Cd-content of sewage particles were released during 4 weeks treatment. 84 U. Förstner A lowering of pH Ieads to the dissolution of carbonate and hydroxide minerals and - as a result of hydrogen ion competition - to an increasing desorption of metal cations. Long-term changes of the pH conditions have been observed from waters poor in bicarbonate ions, which are effected by atmospheric so2 emissions. Significant increases of Cadmium were reported from water of the Sudbury mining area [177]. Cd-enrichment in acidic mine effluents by factors of I ,000 and more, in respect to normal surface waters, has been observed [178, 119, 121]. On the other hand, Cd-availability is greatly reduced in soils rich in carbonate [179, 199, 300]. Significant impacts on remobilization from polluted sediments may result from the growing use of synthetic complexing agents (e.g. NT A, nitrilotriacetic acid) in detergents replacing polyphosphates. Experiments performed by Banat et al. [180] with polluted river sediments indicate a high percentage of cadmium mobilization. The results of a test made by Chau and Shiomi [181] in various NTA metal complexes in natural waters of Lake Ontario show a very delayed degradation of Cd chelates. Thus, a potential danger seems to arise for drinking water obtained from bank filtration or artificial recharge processes [182]. Oxygen deficiency in sediments Ieads to an initial dissolution ofmanganese oxides followed by that of hydrous iron oxides. Since these metals are readily soluble in their divalent states, any coprecipitates with metallic coatings become partially remobilized. However, the presence of sulfide under anoxic conditions will precipitate toxic metals. These are released by conversion of sulfide to sulfate under oxidizing conditions [38]. Isotope studies performed by Gambrell et al. [154] with Mississippi River sedimentmaterial indicated that exchangeable 109Cd Ievels are strongly pH-redox-potential dependent. A 64 ~56 0 ~48 "0 ~40 "'0 32 1; ~ a:"' 24 Cl 16 8 0 Fig. 7. EfTects of pH and redox potential on exchangeable Suspensions [154] Cd in Mississippi River sediment 109 Cadmium 85 much greater proportion of the incubated cadmium isotope was recovered in the readily bioavailable forms than for any other potentially toxic heavy metal studied. Figure 7 indicates the influence of pH and redox potential on exchangeable 109Cd in the Mississippi River sediment suspensions. It is suggested that considerable cadmium release to relatively mobile forms may occur as cadmium-contaminated sediment is transported from a near-neutral pH, reducing environment to a moderately acid, oxidizing environment. Under these conditions, cadmium levels of subsurface drainage water from upland disposal of dredged materials may be increased, and cadmium availability to plants growing on the material enhanced [154]. The burial of Cd-rich sediments under succeeding layers, whereby the sediments become anaerobic, may be very effective mechanisms to fix Cd onto solid phases [183, 154]. With log Ksp = 27.8 [19], CdS is one of the least soluble metal sulfides (following HgS and CuS). For example in San Francisco Bay dredged sediments [184], about 92% of the totalcadmiumwas found in the organic and sulfide phases. It has therefore been proposed by Jackson [185] that sewage together with heavy metal effiuents should be introduced into settling ponds to achieve an effective method for preventing heavy metal pollution of natural waters; such mechanisms involve the Stimulation of algal blooms with attendant H 2S production. Dredging activities and other physical perturbations of the surface sediment layers in contaminated deposits may have adverse effects on both water quality and aquatic biota due to the release of cadmium [2, 186]. Holmeset al. [125] reported that the cadmium introduced into Corpus Christi Bay Rarbor from industrial effiuents in the summer when the harbor water was stagnant reach the surface sediment with the sulfide ions and precipitate as CdS. In the winter months, however, the increased flow of oxygen-rich water into the bay results in the desorption of some of the precipitated metal [154]. Older data suggesting insignificant effects of dredging and other activities on the release of heavy metals [187-191) should be reexamined with respect to the behavior of cadmium [33]. Biological Uptake and Accumulation of Cadmium in Organisms Cadmium is not an essential trace element for organisms, but rather a typical contaminant [68, 192]. Its concentration and the typicallog-normal distribution in organisms [193] is influenced by the concentration in the environment. Uptake in Plants Table 12 reproduces cadmium concentrations in some plantsandplant parts. In Cd-contaminated environments, very high enrichment rates are found in mosses, cabbage, carrots, radishes, lettuce, potatoes, and turnip roots. Obviously cadmium is readily taken up by roots and distributed throughout the plant [194]. The amounts ofCd in plants grown in contaminated soils lie in the same order of magnitude as the cadmium concentrations in the substrates. 86 U. Förstner The amount of uptake is influenced by soil factors such as cation exchange capacity [195, 196], pH [195], phosphorous Ievels [195], fertilizers [197], other heavy metals [195], soil temperature [195], and organic matter [195]. The accumulation of Cd varies with the plant tissue and the species of plant investigated [194]. Corn roots contained 2-4 times as much Cd as did shoots when grown in solution culture for 12 days [198]. Uptake and translocation into shoots from a loamy sand soil was 48 mg/kg, while uptake from a silty clay loam soil having a high cation exchange capacitywas 8.4 mgfkg [194]. F o-liar and root uptake of Cd are equally effective [195]; in both methods of application the quantity of Cd distribution had the sequence: stems > leaves > pots > beans [194]. Uptake into crops is significantly higher from acid than from calcareous soils [199]. Uptake, Absorption, Storage, and Excretion in Animals The removal of Cd from the water by organisms occurs by external adsorption as well as internal uptake through organs such as the gills. Heavy metals appear to be accumulated by ion-exchange processes involving organic molecules such as proteins, e.g. in phytoplankton, seaweed, etc. (200]. These processes are responsible for the typical distribution of Cd in vertical ocean water profiles [34, 201]. Agents such as moulted exoskeletons and faeces of zooplanktonic animals affect the vertical distribution of metals in the sea mainly in coastal areas where nutrients for high biological productivity are available from upwelling ofthe ocean water or from runofffrom the land [202, 200]. The following are factors for the uptake of metals from solution (Prosi, in [2]): temperature and oxygen content [203], water hardness [204], pH values [204], salinity, and the concentration of organic compounds [205]. With regard to the latter factor, it has become evident that the environmental impact of a particular metal species may be actually more important than the total metal concentration. Organic ligands, such as fulvic acids, NTA and EDT A, can inhibit the uptake of metals and raise the toxic threshold [206, 207]. Free ion activity, i.e. Cd (H 2 0)~+ is considered as an approximate measure for toxic effects ofmetals [208], especially in respect to phytoplankton [209]. Experiments ofRamamoorthy and Kushner [210] indicate that the metal affinity toward the different microbial growth media largely follows the availability of free cations, i.e. Cd2+ ~ Cu2+ ~ Pb2+ > Hg2+ (the reverse of the Irving-Williams series of stability constants of metals to organic ligands). In addition to these factors, age ofthe organism plays a role in the metal concentration, as well as a number of species-specific effects, that are, however, little known as yet. Particularly in large animals, the adsorption ofheavy metalsfromfoodmay be very important. In oysters, for example, metals such as Zn are obtained from ingested particles rather than from solution; differences in the availability of metals in foodstuffs depend on factors such as the Iacility with which Cadmium 87 the material is digested, the chemical form of the metal, and the relative binding capacities ofthe animaland the products of digestion in its gut [200]. Excretion of abnormal concentrations of heavy metals can take place in a nurober of ways [200]: through the gills such as in the crab and in rainbow trout; in a particulate form from the mantle edge via the byssus gland, such as in mussels; into the gut such as in the cyprid larva of barnacles; and removal in the faeces, such as in most of the higher organisms. Liverand kidney usually are the major storage organs. In the bivalve mollusc Pecten maximus high concentrations ofFe, Cd, and Cu are found in the liver, whereas Zn, Mn, and Pb are stored in the kidney. Storageproteins such as metallothionein for Cd, Zn, and Cu have been found in terrestrial mammals as well as in aquatic animals [200]. The behavior of Cd compared to that of Zn is interesting especially in respect to the fact that zinc is more readily removed from sea water, probably because it is better regulated- as an essential element- by organisms than Cd [200]. Bryan and Hummerstone [211] have shown that in the polychaete Nereis diversicolor, Cd is adsorbed from solution more slowly than Zn, butthat with increasing levels of both metals the rate of adsorption of Cd increases more rapidly than that of Zn. Investigations of Peden et al. [212] on the carnivorous gastropod Nucella suggest that once having been adsorbed, Cd is less readily excreted than Zn, so that ultimately a higher concentration factor is present for Cd [200]. Table 14. Geometrie mean concentrations.of cadmium in different groups of organisms [200] mg Cd/kg dry weight Seaweed (all types) 0.5 mg/kg Phytoplankton 2 Filter-feeding groups Zooplankton (copepods) Tunicates (mainly ascidians) Bivalve molluscsa Oysters mg/kg 4 mg/kg mg/kg 2 mg/kg 10 mg/kg 50:50 carnivorous and herbivorous or particulate feeders Gastropod molluscs Echinoderms Basically carnivorous groups Decapod ernstaceans Coelenterates Cephalopod molluscs Fish aExcludes Pectinidae 6 mg/kg 2 mg/kg 1 mg/kg 1 mg/kg 5 mg/kg 0.2 mg/kg U. Förstner 88 Data on cadmium concentrations in animals are summarized, among other authors, by Prosi in [2], in the contribution of Shacldette and Nisbet in the "Subpanel Report on Cadmium" [7] and by Bryan [200]. The geometric mean concentrations of cadmium in different groups of marine organisms from the latter work are given in Table 14. Food Chain Effects In field investigations dealing with heavy metal enrichment in organisms, it is imperative to group the organisms according to their habitat and ecologic behavior, i.e., feeding habits (phytophageous, carnivorous, omnivorous, filter feeding, sediment feeding, detritus feeding, etc. ), life cycle, life history, sessility and wandering [213]. In addition, the physiological response ofvarious organisms towards metal pollution may be different with respect to organ distribution of the metal, synergistic or antagonistic effects of other metals on metal uptake, heavy metal resistance, etc. [214]. When all these factors are considered in respect to heavy metal amplification in the food chain, it becomes clear that, in many cases, elevated heavy metal concentrations in higher trophic Ievels do occur but not necessarily in the classical sense of food chain enrichment [213]. In an urban-influenced river section, Prosi [215] determined a significant increase of Cd in the food web ofbenthic invertebrates compared to fish (Fig. 8): It was generally found that according to feeding habits, sediment-depenCd 62 20 t • 10 • 5 • E'0. 2.0 0. • 1.0 0.5 • • 0.1 s r •T 0~?·1 A L F Fig. 8. Cd distribution in two sections of the Elsenz River (light column: rural; shaded: urbanindustrial influenced) at different trophic Ievels. S=sediment < 2 Jliil, T=tubificid worms, A=isopods (Ase/lus aquaticus), L=leeches, F=fish (roaches, sticklebacks). Mean concentrations (dry weight); arrows indicate minimum and maximum values [215] 89 Cadmium dent organisms (Tubificidae) has greater metal concentrations than other biota. Metal contents ofthe benthic food web, sludgeworms (Tubifex tubifex and Limnodrilus hoffmeisteri), isopods ( Asellus aquaticus), and leeches (Herpobdella octoculata) constantly decrease, so that the lowest concentrations appear in the fish. Investigations performed by Butterworth et al. [216] in the Severn Estuary demoostrate the effects of pollution on the concentrations of Cd in aquatic organisms. Coastal waters bordering the southern shore of the Bristol Channel contain abnormal amounts of cadmium, zinc and lead, which are probably introduced from the Bristol area via the River Avon. In the water samples the effects ofthe pollution have been traced as far away as Rarland Quay, some 150 km to the west from A vonmouth into the Bristol Channel. Table 15 Table 15. Cadmium concentrations in water, seaweeds and shore animals of four collecting stations on the southem side of Sevem Estuary and Bristol Channel [216] Collecting point Distance from Seawater Avonmouth IJ.g Cd/! Fucus mg Cd/kg Patella Thais 4km 5.8 220 550 Brean 25 km 2.0 50 200 425 Minehead 60km 1.0 20 50 270 Lynmouth 80km 0.5 30 50 65 Portishead indicates that the contamination in the water by cadmium is obvious1y transmitted to the living material inhabiting the shore - at relatively low levels in seaweed Fucus (the producer), at higher levels in limpets Patella (a primary consumer), and greatest concentrations in the dog whelk, Thais (secondary consumer). There are significant differences of the Cd contents of different tissues. Mullin and Riley [23] found that in molluscs, levels ofCd were ofthe order of 1.5 mgfkg in muscle, and up to 550 mgfkg in digestive glands and renal organs. Brooks and Rumsby [217] found that in oysters cadmiumwas strongly concentrated in the gills, visceral mass, and the heart. The same authors found 2000 mg Cdfkg dry weight in the liver of the scallop Pecten novae-zelandiae. Analyses made by Bryan [200] from Pecten maximus revealed 32 mg Cdfkg dry weight for the whole animal, 321 mg Cd/kg for the liver, 79 mg Cd/kg for the kidney and 2.2 mg Cdfkg for the muscle and for other tissue. Schroeder and Balassa [197] found that in lobster, levels of cadmium were 14 times higher in the digestive gland than in muscle; analyses from Topping [218] on the lobster Homarus gammarus reveal 0.3 mg Cd/kg for the abdominal muscle, 17 mg Cd/kg for the gills and 12 mg Cdfkg dry weight for the liver. Fish tissues from teleost Scombresox saurus [219] contain 0.05 mg Cd/kg dry weight in the muscle and 0.62 mg Cd/kg in the liver. Jaakkola et al. [220] 90 U. Förstner analysed pike from polluted and other areas in Finland; whereas the Cd content in muscle was similar for both areas (0.026 mgfkg dry weight in polluted areas, 0.041 in other areas), there is a significantly higher concentration of Cd in the kidney of fish from the polluted area (1.52 mgfkg) compared to those from other, less polluted areas (0.95 mg/kg). lndicator Organisms Due to their wide distribution in the marine environment Mytilus sp. (especially Mytilus edulis) and oyster species ( Ostrea edulis and Crassostrea sp.) have proved tobe especially useful indicator organisms3 • In his "mussel watch" Goldberg [222] has even suggested that as a long-term indicator, bivalves can Cd 140 .....= • 120 ~ 0 E 100 • 0.. 0.. .... $Cl) >. .!: -c u • 80 0 • • • •• • • • • • 60 .,.. 40 • • • •I •• •• 20 0 1a. 0 2.0 4.0 Cd in mud 6.0 Fig. 9. Cadmium concentrations in dried mud and in oysters (Crassostrea gigas) in Tamar Estuary, Tasmania. Values in ppm [223] 3 According to Bryan [200], poor regulators, i.e., organisms having very little ability to regulate the total concentration in their body and which tolerate metals in the tissues or their storage in an inactivated form are suitable for use as biological indicators. Phillips [221] has proposed that the best-studied indicator types to date are the bivalve molluscs and the macroalgae. Among the former group, "Mytilus edulis may be the appropriate candidate because of its extensively-studied physiology, its world - wide distribution in temperate waters and the amount of accumulated knowledge concerning its uptake of metals and its meta) content in various waters" 91 Cadmium make certain water and sediment sampling procedures unnecessary. There is a distinct straightline relationship for cadmium in molluscs and sediment, as shown for the example from the Tamar Estuary in Tasmania (Fig. 9 [223]). Values of Cd concentrations in mussels and oysters from both less and more strongly contaminated examples are listed in Table 16. The higher Cd concentrations reached in some mussels and oysters are suspected ofbeing dangerous for humans upon consumption. According to Ratkowsky et al. [234] cases of nausea and vomiting in consumers who had eaten oysters from the Derwent Estuary in Tasmania was probably caused by the contamination of these kg of body bivalves. The admissible daily intake of cadmium of 100 ~g/70 weight (see below) is reached with approximately 50 g of oyster (wet weight) from a moderately polluted area, and only 10 g of oyster from some parts of Derwent Estuary and several other areas listed in Table 16. Table 16. Cadmium concentrations in musse1s and oysters (mg/kg dry weight) Bivalve mussels Mediterranean Sea NW coast France/ltaly SW Spain/Portugal Mediterranean Sea Trondheimsfjorden Norway Irish Sea Tasman Bay, New Zealand Bristo1 Channe1 Derwent Estuary, Tasmania Port Phillip Bay, Australia Oysters San Antonio Bay, Texas Knysna Estuary, R.S.A. SW England Estuaries Sevem Estuary, U.K. Tamar Estuary, Tasmania Tasman Bay, New Zealand Port Phillip Bay, Australia 1.9 (0.4-5.9) 1.7-3.6 [224] 2 (1-5) 5.1 10 18 (4-60) 18.6 (4.3-38) 24.6 ± 21.9 [226] 3.2 3.7 2.2-26.7 [230] [231] [232] [233] [223] [217] 17-40 33.2 35 (10-43) 91.6 ± 73.1 [225] [227] [217] [228] [113] [229] [229] Human Intake, Absorption, and Excretion of Cadmium Food Concentrations Relative to the data on the concentration of cadmium in plants and animals, a short summary is given here on the contents of cadmium in several foodstuffs relevant for human nutrition. Characteristic data are excerpted from the U. Förstner 92 CEC-Study on Cadmium, which is one of the most recent and up-to-date compilations4 in that respect (Table 17) [1 0]. Table 17. Cd contents in major foodstuffs (examples) [10] Cereals and vegetables mg/kg dry weight Country Ref. Wheat flour Wheatflour Potatoes Potatoes Carrots Tomatoes Cabbage Radishes Rhubarb Lettuce Spinach Onion 0.029-0.108 0.05 -0.10 0.02 -0.05 0.039 0.016-0.088 0.015 0.022-0.094 0.011-0.027 0.010-0.057 0.031-0.198 0.055-0.063 0.018-0.040 Sweden Canada New Zealand ER.Germany ER.Germany ER.Germany ER.Germany ER.Germany ER.Germany ER.Germany ER.Germany New Zealand [6] [235] [236] [237] [238] [237] [238] [238] [238] [238] [238] [236] Fruit Apples Prunes 0.005-0.027 0.014-0.067 ER.Germany ER.Germany [238] [238] Dairy products Milk Butter Eggs, whole 0.010-0.076 0.02 0.04 ER.Germany New Zealand New Zealand [238] [236] [236] Meats Beef Pork Chicken Kidney (beet) Kidney (pork) Kidney (beet) Kidney (beet) 0.02 0.03 0.03 0.17 0.07 0.27 4.10 [236] [236] [236] [236] [236] [239] [239] Kidney (elk) Liver (elk) Liver (horse) 8.0 1.5 7.5 New Zealand New Zealand New Zealand New Zealand New Zealand ER.Germany ER.Germany (Stolberg) Finland (Poorvool) Finland (Poorvool) UK (industrial) Seafood Museie of various fish Oyster Oyster Oyster (canned) Crab Crab Molluscs Various seafish Freshwater fish 0.08 0.1 0.2 3.31 5.0 22 2 0.1 0.2 -0.10 -0.10 -0.08 -0.27 -0.18 -1.67 -7.8 -2.1 -33.1 -50 -0.6 -1.2 UK USA (eastem) USA (westem) New Zealand UK Europe Europe Europe Europe [220] [220] [240] [212] [241] [241] [230] [212] [71] [71] [71] [71] 4 A review of the effects of cadmium in mammalian systems has just been published: J.H. Mennear (ed.) Cadmium Toxicity, Marcel Dekker, Inc. New York, 224 p. (1979) 93 Cadmium Most foodstuffs from less contaminated areas contain less than 0.1 mg Cd/kg, whereas liver, kidney and shellfish can show much higher concentrations. Investigations on cattle from southern Germany [242] show average values of < 0.005 mg Cdjkg for meat parts, whereas the contents in the liver increase to 0.08 mg Cd/kg (0.005-0.3 mgjkg) andin the kidneys to 0.9 mg Cd/kg (0.04--1.41 mgjkg). Some vegetables and cereals concentrate cadmium when cultivated in polluted soil [10]. It is suggested that upon conditions of general air contamination or through the accumulative effect of fertilizers there should be a significant tendency for the Cd concentrations in foodstuffs to increase [243]. Intake from Food, Water, and Air Representative studies on the dietary intake of cadmium and other noxious substances were first carried out in the United States and 1ater in many other countries (basically involving the analyses of samples representative of food at the point of ingestion [244]). Such studies usually reflect the composition of the diet ofthe average person. Table 18 shows the data ofthe cadmium intake Table 18. Cadmium in United States and Canada market basket survey [245, 246] Milk and dairy products Meat, fish and poultry Grain and cereal Potatoes Leafy vegetables Legumes 7. Root vegetables 8. Garden fruits 9. Fruits 10. Oils and fats 11. Sugar and adjuncts 12. Beverages 1. 2. 3. 4. 5. 6. USA (1968/69) Canada (1969) Range mg Cd/kg Range mg Cd/kg 0.01-0.09 0.01-0.06 0.02-0.08 0.02-0.13 0.01-0.23 0.01-0.03 0.01-0.08 0.01-0.38 0.01-0.38 0.01-0.13 0.01-0.07 0.01-0.04 l!g daily intake <0.02-0.06 0.05-0.08 < 0.02-0.14 <0.03-0.22 <0.02-0.05 <0.02-0.06 0.03-0.09 < 0.02-0.06 <0.01-0.02 0.03-0.07 <0.02-0.03 <0.01-0.04 5 4 14 7 5 1 1 5 2 2 1 4 50 l!g l!g daily intake 15 19 13 19 2 1 3 3 1 1 3 2 80 l!g in the US Market Basket Study for June 1968 to April 1969 [245] and for Canada in the Hull-Ottawa area during 1969 [246]. Similar data were calculated for West Germany (48 llg/personjday [237]), for Romania (38-64 llg/personjday [248]), for Czechoslovakia (59 llg/day [248]), for Japan (59 llg/day [249]- for Japan there are other figures ranging from 25-120 llg/person/day [250, 251])- for New Zealand (21 llg/personjday [252]), for France 94 U. Förstner (20--30 J.lgfday [253]), for the UK (15-30 J.lgfday - the average individual consuming 1.5 kg offood per day [244]), and for Sweden (10--17 J.lgfperson/ day [254, 255]). In areas where the soil has been found highly contaminated by cadmium, oral intake (mainly through contaminated rice) has been calculated tobe as high as 600 J.lgfday [10]. If these data are excluded, the total daily intake in non-polluted areas ranges between 6 and 94 J.lg/personjday with a median of approx. 43 J.lgfday from food and 3 J.lgfday from water [10]. From air intake the following figures have been calculated [10], differentiating smokers from non-smokers (assuming a daily inhalation of 20m3 at 25% deposition equal to 3 p;g/day from 40- cigarettes): Rural areas Urban areas Industrial areas Non-smokers Smokers 0.0005- 0.215 0.01 - 3.5 0.05 -25 3.0005- 3.215 j.lg/day 3.01 - 6.5 llg/day llg/day 3.05 -28 It can be seen that most of the cadmium intake originates from foodstuffs. Cadmium intake from water on the average is equivalent to that inhaled during smoking. In industrial areas approx. 30% of the total intake of cadmium stems from air pollution. Absorption Inhalation. The effect of cadmium absorption by the lungs depends on the amount retained (Q deposited- Q rapidly eliminated via the upper respiratory tract) and probably also on the chemical form of the retained particles [10]. From experiments with dogs it has been demonstrated that cadmium oxide dusts and cadmium chlorides were more readily absorbed than cadmium sulphide [256, 257]. Male non-smokers (up to 60 yr old) were found to have an average 6.6 mg Cd in their kidney, liver, and lungs, whereas smokers (one pack of cigarettes per day for 40 yr) showed 14 mg in these organs [258]. Assuming that 64% of the cadmium deposited can be absorbed, this means that in the general environment 13-19% of the inhaled cadmium is absorbed; with regard to residents in urban areas the amount of Cd absorbed has been calculated as 0.006--2.24 J.lgfday for non-smokers and 1.9-4.2 J.lg/day for smokers [10]. Absorption by the Gastrointestinal Tract. Experiments on human volunteers (19 to 50 yr old) given labelled Cd orally, indicate that the absorptionrate ranges between 4. 7 and 7% [259]. Age may be important in the rate of gastrointestinal absorption of cadmium in humans [10], since the uptake rate is higher for adolescents than for adults. Assuming 6% absorption, the total amount of Cd absorbed approx. 2.6 J.lg/day from food and 0.2 J.lgfday from water [10]. Cadmium 95 Body Distribution Normal concentration ofCd in blood averages below 1 J.lg/100 ml, with a large variation between 0.06-15.9 J.lg/100 ml [10]. Smokers have higher blood cadmium concentrations than non-smokers [260]. In workers no correlation was found between cadmium in blood and exposure time [261] and all present results would suggest that cadmium in blood is probably not a reflection of the body burden, but is rather influenced mainly by current exposure [10]. Cadmium accumulates with age (at least until age 50) and about 50% of the accumulated cadmium is found in the kidney and liver. In these tissues this toxic metal is mainly bound to metallothionein, a protein of low molecular weight (10,000-12,000), very rich in cysteine residues, and deficient in aromatic amino acids [10, 262].1t is suggested that cadmium acts as a highly specific inducer of metallothionein [263] and that Cd toxicity occurs when available metallothionein is insufficient to bind all the cadmium [6, 107]. When the cadmium-metallothionein complex is synthesized within the cell it may protect temporarily against cadmium toxicity [10]. However, the protective role of metallothionein against acute toxicity of cadmium has been questioned by recent work of Goyer et al. [264]. The Cd contents in the kidney, liver, and lungs of 172 US adults were for non-smokers 4.16 mg, 2.28 mg, and 0.36 mg, respectively, and for smokers 10.28 mg, 3.06 mg, and 0.81 mg [265]. The average body burden for adult non-smokers in the USA has been estimated at 19.2 mg, for smokers 32.4 mg [266]. Friberg et al. [6] have estimated the total body burden of a 50-year-old adulttobe approx. 15-20 mg in the UK and Sweden and 80 mg in non-polluted areas of Japan. The total body burden of cadmium in workers could exceed 1200 mg [6]. lt is possible, however, to find lower cadmium concentrations in kidney of exposed workers and ltai ltai patients than in normal individuals because when renal darnage is present, urinary cadmium excretion is increased and therefore renallevels may decrease [10, 267, 268]. Decreased cadmium contents in the k:idney cortex with age has been exp1ained, among other hypotheses [10], with the fact that 70% of the total world production of cadmium has occurred within the last 20 yr [6]. Excretion In normal adults the amount of Cd excreted daily via the urine is probably below 2 J.lg/day (range 0.2-3.1 J.lg/1); in workers exposed to cadmium urinary Cd excretion can reach several hundred J.lg/day (compilation by Friberg et al. [6]). 1t is still not known conclusively whether Cd concentration in urine reflects body burden or current exposure (see discussion in [10]): "at low exposure levels the amount of cadmium absorbed may be insufficient to saturate all the body binding sites (e.g. induced metallothionein) and urinary excretion does not increase proportionally to the exposure levels; in high exposure conditions (e.g. workers, adult itai-itai patients) andin the absence of renal lesion the urinary concentration would be more a reflection of exposure levels- all the binding sites are now saturated" [10, 261]. 96 U. Förstner Biological Half-Time in Rumans Information on the various aspects of biologic half-time of cadmium in organisms is summarized by Friberg et al. [6]. With regard to the determination of these parameters for humans, calculations have been performed by Sudo and Nomiyama [269] from urine samples of former cadmium workers with proteinuria. A value of about 200 days was determined. From whole blood data among workers exposed Ionger to cadmium a half-time value was roughly calculated to be about 6 months [6]. Tsuchiya [270] estimated Cd half-life in workers to be about 1 yr. Summarizing a large number of whole body burden data involving measurements of uptake and excretion, the biological half-time is suggested to correspond to values of 13-47 yr [6]. Theoretical models of cadmium metabolism imply that the half-timein whole body are 9-18 yr [271 ]. With inclusion ofliver accumulation into these models, the biological half-time for the human kidney is estimated tobe 17.6 yr on an average and for the Iiver 6:2 yr (272]. Toxicological Aspects of Cadmium Pollution Summaries on the toxicology of cadmium have been given, among others, by Flicket al. [273], Fulkerson and Goehler [65], Friberg et al. (6], Nordberg [9], in the CEC Study [10], and during the First International Cadmium Conference in San Francisco ([11; 17 articles on pp. 167-255]). Review articles in the German language were compiled by Rosmanith [192] and Ohnesorge [274], in French by Godfr.ainetal. !275). The human-toxicological.aspects ofcadmium will be shortly summarized here in accordance with the description by the Commission of the European Communities' study "Criteria (Dose/Effect Relationships) for Cadmium" [10]. Toxic Effects on Aquatic Organisms As for most other heavy metals, the toxicity of cadmium towards organisms is generally associated with the inhibition of enzyme systems. Cd interferes severely with metalloproteins, metalloenzymes, metallothioneins and phospholipids [276]. There is a wide range of Cd concentrations in the water phase which can be tolerated or which may show Iethai effects for individual groups of organisms. Examples ofthe Iethai toxicity data for marine organisms were given in Table 19 (from Bryan [200]). 1t should be noted that the toxicity may change according to influences of the metal's form in water, by the presence of other metals or poisons, by physiological factors, such as temperature, pH, dissolved oxygen, light and salinity, by the condition ofthe organism, and by special behavioral responses [200]. Data on sublethat effects of cadmium are still rare with the exception of investigations on morphological changes [277]. Further parameters to be considered include inhibitory effects on growth, settlement, reproduction, and (1 I>' 0. 2. e 8 Table 19. Lethai toxicity ofCd on marine organisms [200]. References seeoriginal paper ofßryan [200] Minimum observed LCso LCso (ppm) Group Species 24 h Fish Fundulus heteroc/itus Agonus cataphractus 140 - Mytilus edulis >200 Bivalves Mya arenaria Cardium edule >200 Crangon crangon Crangon septemspinosa Pagurus longicarpus Carcinus maenas Uca pugilator - 2-4 >200 100 - 96 h LCso Time (h) 264 - 49 33 23 165 25 - 33-100 Molluscs Crustaceans Shrimps Crabs 43 h 50 10-33 2·2 3·3 3·3-10 0·5 3·7 16·6 - 1·0 0·32 0·32 4·1 47 0·82 Echinoderm Asterlas jorbesi 12 1 Annelids Nereis virens Nereis diversicolor Ophryotrocha /abronica - 25 25 11 - 8 - - - - - - - 9·5 240 - - 10 1 816 410 Conditions -- salinity oc Form S%o = 20 - 20 15 Chloride Chloride S%o = 20 20 Chloride S%o = 20 20 15 Chloride Chloride 20 20 20 20 15 20 20 20 20 Chloride Chloride Chloride Chloride Chloride S%o = 20 20 Chloride S%o = 20 50% s.w. 20 Chloride Sulphate Sulphate - S%o S%o S%o S%o - = = = = 13 20 'J:> -J 98 U. Förstner metabolism and on behavioral processes such as feeding, learning, and swimming activities (200]. Toxic Effects on Humans [10] In humans the two main target organs of acute and short-term cadmium exposure are the gastrointestinal tract after ingestion and the lungs after inhalation. Long-term exposure mainly affects these two organs; the critical organ is the kidney: Lung. Repeated or prolonged inhalation (no-effect Ievels are probably close to 2 J.tg/m 3 -24 h/day exposure for 70 yr) by exposed workers may produce an obstructive pulmonary syndrome and emphysema, probably from direct action of cadmium on the lungs. Kidney. K.idney lesions (mainly on the proximal tubules) have been observed in workers exposed to airborne cadmium (usually preceding lung damage) andin personsthat have ingested contaminated food. Epidemological surveys in Japan suggest that a continuous oral daily intake of 200 J.tg cadmium could cause an increased prevalence of kidney darnage in persons over 50 yr of age.lt is estimated that such an intake corresponds to an average urinary excretion of6 J.tg Cd/1 (see discussion in [10]). In humans renal darnage is likely to occur when the cadmium concentration in the renal cortex exceeds 200 mg/kg wet weight. Evidence of renal darnage is shown by mild proteinuria, in which mainly low molecular weight globulins and some albumin, glycosuria, amino-aciduria are lost; reduced ability to concentrate urine or to excret acid also occur. A high phosphate clearance with hypercalcutia and stone formation has been likewise found [278]. Cardiovascular System. Controversial results have been reported regarding the hypertensive action of cadmium. A significant increase in blood pressure could be measured in rats fed with a diet containing 0.56-0.63 mg Cd/kg [279].1t has been suggested that the cadmium to zinc ratio in the kidney p1ays a more important role in the development of hypertension than the concentration of cadmium [68]. Bones. Osteomalacia and osteoporosis with a tendency to fracture and bond deformation accompanied by lumbar pains, leg myalgia and pains on bone pressure as well as gait have been described in itai-itai patients, principally in women after menopause who have born several children [280]. Whether cadmium exerts a direct toxic action on bone tissue (which can even precede kidney darnage) or is due to disturbed calcium and phosphorous metabolism secondary to the kidney lesion is a point of discussion (see [11], pp. 211-250). Hematopoietic System. Slight hypochromic anaemia has been observed in most itai-itai patients as well as among workers exposed to cadmium. Carcinogenic, Mutagenic, and Teratogenic Effects. Insufficient epidemological data are available on the potential effects of these influences on Cadmium 99 humans [10]. In animals, cadmium and its compounds have been shown to induce sarcoma at injection sites [281-285], whereas on oral administration cadmium does not seem to act as carcinogen [279, 286]. Likewise, no data relevant to industrial societies (e.g. from Japan) is available to suggest that non-occupational exposure to cadmium constitutes a carcinogenic hazard. Chromosomic an omalies of peripheralleucocytes of workers exposed to Iead and cadmium suggest a synergistic effect between both metals [287]. Regulations Guidelines and standards for environmentallevels of cadmium summarized here may be subject to changes. Food. Approximately 200 mg Cd/kg wet weight has been proposedas a tentative critical concentration in the human kidney cortex. If the total intake of cadmium does not exceed 1 jlgjkg body weight per day, it is likely that the Ievels of cadmium in the renal cortex won't exceed 50 mg/kg, assuming an absorption rate of 5% and a daily excretion of only 0.005% of the body Ioad (reflecting the long half-life of cadmium in the body). The Joint FAO/WHO Expert Committee ofFood Additives [288] has therefore proposed a provisional tolerable weekly intake of 400-500 jlg Cd per individual. As the amount of food consumed by an adult person is about 10 kg per week the mean content of cadmium in food should not exceed 0.04-0.05 mg/kg [289]. The provisional guidelines set by the Japanese Ministry of Health and Welfare, whereby detailed investigations of cadmium environmental pollution are initiated, are 0.4 mg/kg in rice or concentrations in drinking water up to 10 11g/l [10]. Reference values for Cd concentrations in various food items have been published by the German Ministry of Health in 1974, i.e. meat- 0.008 mg Cdjkg; vegetables - 0.1 mgjkg and fruit - 0.05 mg Cd/kg. Among the most important sources of contamination by food are pork (1. 7 jlg Cd/personjday), milk (3 jlg), potatoes (11.6 jlg), vegetables (5.6 jlg) and beer (3 jlg [237]). Kidneys from animals destined for slaughter and subsequent human consumption as weil as uncultivated mushrooms may be rich in cadmium [290]. In Britain, the Toxicity Sub-Committee [278] found that a persistently high dietary consumption of shellfish or brown crab meat with a high cadmium content may conceivably constitute a risk for selected individual consumers. Food Containers. The national regulations concerning the amount of cadmium released from ceramies (for a compilation see [289]) are mostly based on tests involving treatment with 3% or 4% acetic acid (at room temperature) and subsequent determination of dissolved Cd. Limits range from 0.1 mg Cd/1 in Sweden, 0.2 mg/1 in the UK (hollow ware, kitchen utensils ), 0. 5 mg Cd/1 in I taly and the USA, 0. 7 mg/1 in the UK (ceramic ware) to 1.0 mg Cd/1 in Denmark (90 min at 100 oq and South Africa (all types of food containers; [10]). Proposed EEC directives differentiate according to the surface area and use (tableware, child's plate, cookingware, hollow ware). 100 U. Förstner Water. A tentative upper limit for Cd in drinking water was initially set at 10 J..Lg/1 by the US Public Health Service (1962), USSR Government (1970), the W orld Health Organization (European: 1970; International: 1971 ), USA National Academy ofSciences (1972), Australian Government (1973) and the US Environmental Protection Agency (1975) [291]. The WHO Regional Office ofEurope recommended 5 J..Lg/1 in 1973. The Council ofMinisters ofthe EEC fixed the maximum concentration of Cd in "surface water used after purifivation as drinking water" at 5 J..Lg/1 and the goal is less than 1 J..Lg/1 [292]. The "DVGW W 151" (FRG) limits the use of surface waterat 5 J..Lg/1 (purification by bank: filtration and artificial recharge) and 10 J..Lg/1 (physico-chemical treatment). Cd contents in irrigation water usually are limited at 5 J..Lg/1 [293]. Efjluent standards are 0.1 mg Cd/1 in Japan (assuming minimal in-river dilution offactor 10 [294]; US standardspermit 40 J.lg Cd/1 when the receiving streams' low flow exceeds ten times the waste flow [295]. Air. In Japan the provisional guideline for cadmium in ambient air is 0.88-2.92 J..Lg/m3 • The industrial maximum allowable concentrations are significantly higher: in the USA the time weighted average value (TLV or TWA) for cadmium dust is 0.05 mg/m3, and in the USSR the TLV for cadmium fumes is 0.1 mgfm3 • In Finland the TLV for cadmium dust and fumes is 10 J..Lg/m3 and 20 J..Lg/m3 in Sweden. In Germany the MAK for cadmium oxide is 0.1 mg/m\ and the TLVs proposed by the Health and Safety Executive in the UK are 0.2 mg/m3 for cadmium as metal dust or soluble salt and 0.05 mgfm3 for cadmium oxide fumes. The TLV for cadmium dust in Switzerland is 0.2 mg/m3 [10]. Sewage Sludge. Unlike zinc, copper, nickel, or boron, cadmium accumulations in vegetation can reach Ievels that are toxic to animals before the vegetation itself shows any signs of damage. In that respect, guidelines limiting cadmium application, e.g. sludge in agriculture, tend to be very conservative [296]. U.S. Department of Agriculture data [297] suggest a guideline of0.1 kg Cd/hafy, assuming a 20-year lifespan, for use of municipal sludges (103 metric tonsfhectarjlife) containing up to 2000 mg Zn/kg, 1000 mg Cu/kg, 200 mg Ni/kg, and 20 mg Cd/kg (1% of Zn). A draft copy of the Illinois Environmental Protection Agency [298] guideline specifies that cadmium application aretobe limitedas follows: (i) a maximum application rate of0.33 kgfhafy; (ii) a maximum of 6. 7 kg/hafy if the Zn: Cd ratio of the sludge exceeds 100; (iii) a maximum application of 3.4 kg/ha/life if the Zn:Cd ratio of the sludge is less than 100. The provisional Ontario guidelines [299] limit the application to a maximum of 1.6 kg Cd/ha/life. Regulations presently do not consider soil characteristics, such as cation exchange capacity (C.E.C., [300]), pH [301, 302], and carbonate content [199, 303]. Acknowledgements Thanks are due to Dr. F. Prosi, who contributed substantial material for the chapter on biological uptake and accumulation, and to Mr. D. Godfrey for his aid in preparing the English version of the text. Cadmium 101 References 1. Bryan, G.W.: Proc. Royal Soc. B (London) 177, 389 (1971) 2. Förstner, U., Wittmann, G.T.W.: Meta! Pollution in the Aquatic Environment. BerlinHeidelberg-New York: Springer-Verlag 1979 3. Friberg, L.: In: Proc. 9th Int. Congr. Ind. Med. 9, 641 (1948) 4. Kobayashi, J.: In: Proc. 5th Int. Conf. Adv. Water Pollut. Res. /-25 (1971) 5. Hagino, N., Yoshioka, K.: J. Japan Orthoped. Assoc. 35, 812 (1961) 6. Friberg, L. et al.: Cadmium in the Environment. Cleveland: CRC-Press 1974, 248p. 7. 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The quantity of coal tar which is co-produced in the coal carbonization process amounts presently to about 16M tonnes p.a.; three-quarters ofthis raw material is processed in the existing 127 coal tar refineries throughout the world [1]. The exposure of the environment with these chemieals occuring in the up-grading of coal tar can, however, be neglected in comparison with the exposure with PAH and hetero-aromatic systems from other sources. PAH are always formed when organic material containing carbon and hydrogen is subjected to temperatures exceeding 700 oc, i.e., in pyrolytic processes and with incomplete combustion. If the starting material also contains hetero-atoms, e.g., oxygen, nitrogen, and sulphur, then heteroaromatics are formed in addition to PAH. Since pyrolysis or incomplete combustion are processes that take place everywhere in our ecological system, it is not surprising that PAH's and related hetero-aromatics occur everywhere in the environment. However, no reliable data on the total emission of polycyclic organic matter (POM) are available, but Table 1 gives a summary of the estimated benzo[a]pyrene emission in the United States [2]. Global emission of benzo[a]pyrene has been estimated as 5,000 tons/yr [3]. The question as to whether the omnipresence ofPOM is related solely to civilization or has, in addition, a biogenic origin [4, 5] ist still under discussion. However, M. Zander 110 Table 1. Estimated benzo [a] pyrene emission in the United States [2]. The data characterize the situation in the late sixties Source Tons/year Transportation sources Gasoline-powered Automobiles Trucks Diese1-fue1-powered Trucks and buses 10 12 0.4 22 ~2% Heat and power generation sources Coal Hand-stoked residential furnaces Intermediate units Coal-frred steam power p1ants 420 10 1 Oil Low-pressure air-atomized and others Gas Wood 2 2 40 475 ~38% Refuse burning Enclosed incineration Open burning Forestand agricu1tural Vehicle disposal Coal refuse fires 33 140 50 340 563 Iudustrial p1ants Cracking units Asphalt air-b1owing Coke production ~45% 6 <1 192 198 ~16% 1,260 tons/year arguments pointing to an origin conditioned solely by civilization are on the increase [6]. The POM contents of fossil materials, andin particular of mineral oil [7], are not given consideration in this respect. Although the mechanism ofPOM formation in combustion and pyrolysis processes is very complex and variable, a relatively clear picture ofthe overall reaction has outlined [8-10]. POM formation proceeds by free radical mechanism. Radical species containing one, two or many carbon atoms can combine rapidly at the high temperatures (500-800 oq prevailing in the flame 111 Polycyclic Aromatic and Heteroaromatic Hydrocarbons front or under pyrolytic conditions. Highly reactive transient species formed in the first steps of the reaction are stabilized by ring closure, condensation, dehydrogenation, Diels-Alder reactions, ring expansions and other path ways yielding a manifold of polycyclic systems. Pyrosynthesis is obviously a function of many variables, not the least of which is the presence of a chemically reducing atmosphere, common in the center of flames, where radical chain propagation is enhanced, allowing the build-up of complex PAH molecules. Although methane itself can Iead to PAH [11] the formation of these large molecules is favoured by the presence of higher-molecular-weight radicals. lt has also been shown that specific aromatic systems can serve as precursors for higher-molecular PAH [9, I 0]. Figures 1 and 2 give examples for the formation ofPAH under pyrolytic conditions according to the work by Badgerand Lang et al. [8-10] respectively. c I c - c....C I c..... c - ac,c - - oS9 h h (J.C......cI c.....-c J CO Fig. 1. Mechanism ofbenzo[a]pyrene formation under pyrolytic conditions [8] / Fig. 2. Pyrolysis of2-methy1-naphthalene (700 oq [9] - 112 M. Zander Chemistry Nomenclature Although severals systeros of noroenclature for PAH have been applied in the past [12, 13] the IUPAC systero [14] has becoroe generally accepted now. Common naroes for the basicring systeros, however, arestill in use. For the construction of roore highly condensed P AH the basic systero is lettered (a, b, c, etc.) and the added rings nurobered in the roanner shown in Fig. 3. The IUPAC naroes for roany known PAH are given in the Ring Index [15]. naphtho [2,1-a ]anthrc~ anthrc~ Fig. 3. IUPACmode ofPAH nomenclature Building Principles The nurober of isoroeric PAH increases treroendously with the enlargeroent of roolecular size. In the case of systeros limited to benzene units, the nurober of isoroers can be obtained very easily on a graph theoretical basis [16]. Table 2 shows several relevant examples. As to the 5-ring· systems, all theoretically possible isoroers have been described [17]. Of the theoretically possible 6-ring systeros, only approx. 45% have been described in the Iiterature to date. The ratio ofthe number ofsystems known today to the nurober ofthe theoretically possible systeros decreases rapidly with increasing nurober ofrings. However, in connection with environroental pollution very high-molecular PAH are less relevant due to their low volatility and solubility. The host of PAH can be subdivided according to different building prinTable 2. Mo1ecular size and number ofisomers ofPAH[16]" Number ofbenzene units 3 4 5 6 7 10 12 Number of isomers, PAH 3b 7 22 82 333 30,490 683,101 • Including phenaleny1 radica1 structures b As an example: phenanthrene. anthracene, phenalenyl radical Polycyclic Aromatic and Heteroaromatic Hydrocarbons 113 ciples. - Extremely useful, in particular from the theoretical point of view, is the classification ofPAH according to the altemance principle [18]. In altemant PAH the carbon centres can be subdivided in two sets g (starred) and u (unstarred) in such a way, that each g-carbon is directly linked to u-carbons only and vice versa, while in non-altemant PAH C-C-bonds between carbons ofthe same parity occur. For altemant PAH the pairing theorem is valid while this is not the case with non-altemant PAH [18]. Thus, many differences conceming the physical and chemical properties of altemant and nonaltemant PAH can be correlated with the different HOMO-LUMO situation in these two classes of n-electronic systems as given in Fig. 4. .&. • 0. • • • • LUMO LUMO HOMO HOMO "'alternant"' "non-alternant "' Fig. 4. HOMO/LUMO representation of alternant and non-alternant PAH annellated systems tetracene benzo[a)anthracene chrysene peri- condensed systems benzo(e]pyrene coronene Fig. 5. Examples for annellated and peri-condensed P AH fluoranthene M. Zander 114 Most common is the subdivision of the whole amount of PAH into annellated and peri-condensed systems. In annellated systems the tertiary carbons are centers of two inter-linked rings, whereas in peri-condensed systems some of the tertiary carbons are centers of three inter-linked rings. Examples of these two classes of PAH are given in Fig. 5. Relationships Between Topology, Stability, and Reactivity of P AH The resonance energy per n-electron ("specific resonance energy") ofbenzene, naphthalene, anthracene, and tetracene (acenes) is plotted versus the total number of n-electrons ofthese systems in Fig. 6. The diagram shows a stability 6 10 14 18 22 total number of 'lf- electrons Fig. 6. Correlation between resonance energy per n-electron (eV) and total number of n-electrons ofPAH decrease of these systems with increasing annellation. However, it is even more remarkable that triphenylene 1 has the same specific resonance energy as benzene, and perylene 2 the same as naphthalene. On a purely formal basis, one must conclude that triphenylene consists of three localized benzene units, and perylene, on the other hand, of two localized naphthalene units. This is shown in the formulae, in which the thicker lines denote "quasi-single bonds". The existence of localized n-electron ranges in PAH was first postulated by Clar [19, 20], who developed this concept on the basis ofnumerous experimental results to build up a comprehensive structural model of PAH. Polansky and Derflinger [21] developed the quantumchemical verification of this model by using the MO's of benzene and other partial structures Po1ycyclic Aromatic and Heteroaromatic Hydrocarbons 115 recognizeable in PAH (butadienoids, allyloids, and ethylenoids) as bases in the Hückel approximation instead ofthe usual atom orbitals. In this way, they obtained the "character orders" of these partial structures, the character orders being a measure for the contribution of the bonding orbitalsofapartial structure to the bonding orbitals of the entire molecule. According to the analogy principle, the bonding properties of a partial structure equate the more with those of the reference compound (benzene, butadiene etc.) the greater the character order of the partial structure under cconsideration. By way of example, the formula of dibenzo[b,n]perylene (Fig. 7) includes the Fig. 7. Dibenzo[b,n]pery1ene- Representation of bonding properties [19, 21] _________ .,. all-benzoid -----· quasi all-alkenoid Fig. 8. The "All-benzoid" and quasi "all-a1kenoid" building princip1e of P AH relevant character orders. The figures in the hexagons are the benzoid character orders, the remaining figures the butadienoid character orders of the cisoid C4 partial structures. The formula on the right gives the qualitative bonding properties in dibenzo[b,n]perylene according to Clar [19, 20]. The Polansky character orders correlate well with the experimental data such as NMR coupling constants, magnetic susceptibilities, half-wave potentials, and reaction rates [22]. 116 M. Zander The model indicates that in a series of isomeric PAH stability increases with the number ofbenzoid partial systems. In fact, PAH such as triphenylene 1 (see Fig. 6), which can also be thought of as purely benzoid partial systems linked by "quasi single bonds", are the most stable PAH known [23]. Such "allbenzoid" PAH can be thought as having been formed by innermolecular ring closures ofpolyphenyls (Fig. 8). The acenes are found at the other end of the stability scale ofPAH (Fig. 8). An infinitely long acene is formally created by linking two polyene chains; it could be described as being an "all-alkenoid". The extremely unstable heptacene, with 7linear annellated rings, is the mosthigh molecular representative of this structural principle known so far. Heptacene could be designated as being a "quasi all-alkenoid". Between these two extreme cases - the all-benzoid and the quasi all-alkenoid structural principle- numerous different structural principles of differing stability exist; the simple relationship between the number of benzoid partial systems and stability is tobe noted in all cases. The localization energy concept [24] has proved useful to describe reactivity ofPAH's quantitatively. The localization energy Lu is the energy- mostly given in units of resonance integral ß-required to isolate a n-electron at the centre u from the remaining n-system. The smaller Lu, the greater the relative reaction rate constant of an addition step under consideration. Of the various known reactivity indices, which are a measure for the localization energy [25], Dewar's reactivity number Nu [26] has two striking advantages: the Nu values correlate extremely well with experimental data, and the Nu values can also be calculated extremely easily for systems having a very large number of centres. One disadvantage of Dewar's method lies in the fact that in its simple form it is applicable only to even-alternant n-electron systems. In principle, the relevant Nu value can be calculated in respect of each carbon centre of a PAH. Accordingly, there is a Nu pattern for each PAH. Initially, no relationships aretobe recognized between the Nu pattern and the topology ofthe systems. Recently, however, it has been shown, that Nu values and Polanskys character orders [21] correspond to some extent in a significative way; since Polansky's approach gives a good description of the bonding properties of PAH an understanding of the relationship between the Nu pattern and the topology ofthe systems has also been derived (27, 28]. Synthetic Methods In most cases the synthesis of a definite P AH consists in adding new rings to an easier available starting PAH. If the starting system shall be enlarged by one benzene unit and (6-m) carbons ofthis benzene unit arealready present in the starting system then m carbons have tobe added by the synthesis. The most important cases are those for which m=O, 2 or 4 respectively. The case m = 0 corresponds to an innermolecular ring closure. The case m = 2 can be advantageously verified by Diels-Alder reactions. Several suitable methods exist for the case m = 4 one of which is cyclialkylation. The respective examples for the synthesis of PAH by innermolecular ring closure [29, 30], DielsAlder reaction [31 ], and cyclialkylation [32] are depicted in Fig. 9-11. 117 Polycyclic Aromatic and Heteroaromatic Hydrocarbons h'l ~ Fig. 9. PAH synthesis by (photochemical) intramolecular ring closure [29, 30] 0 II -8 H reduction Fig. 10. PAH synthesis by Diels-Alder reaction [31] - Pd /C Fig. 11. P AH synthesis by cyclialkylation [32] The most important methods for the purification of PAHs obtained by synthesis are liquid chromatography, high vacuum sublimation, complexation using electron acceptor compounds like picric acid following decomposition ofthe adducts and crystallization [33]. The standard reference books on the synthesis ofPAHs arestill Clar's two volumes (17J. 118 M. Zander Analytical Methods Since the environmental analytical ehernist is interested, generally, in the analytical determination of numerous compounds present only in low concentrations in complex mixtures separation techniques with high sensitivity have to be employed to gain the required information. Thus, gas liquid chromatography (GLC) and high pressureliquid chromatography (HPLC) are the most widely used methods in P AH analysis. What is possibly the most complete G LC separation so far of PAHs up to mole masses of approximately 300, was obtained by using 92 m glas capillary columns with polyphenylether sulfones as the stationary phase [34]. Capillary GLC mostly in combination with mass spectrometry has been widely used for the determination ofPAHs and related heterocyclic systems in environmental samples. Thus, Lao et al. [35] detected 150 different PAH in city air using the capillary GLC/mass spectrometry combination and Grimmer et al. [36] succeeded in characterizing 150 components of vehicle exhaust gas as PAH, 73 of these could be positively identified by comparison with test substances. Stationary phases with high selectivity are of continuing interest, in particular for high-temperature GLC. Among thermally stable liquid crystal phases that can be used in combination with mass spectrometry without bleeding, N,N' -Bis-(p-phenylbenzylidene)a,a' -bi-p-toluidine can be employed at working temperatures up to 290 °C. Excellent separations of PAH with 4-7 rings have been achieved with this phase [37]. lnorganic salts such as LiCl or CaCI2 comprise another group of selective stationary phases that have been used in high temperature GLC ofPAHs [38]. Snowdon [39] reported on the application of column packings consisting of eutectic salt mixtures (KN03/LiN0 3fNaN0 3) on Chromosorb for the separation ofPAH and PAH homologues. The application of HPLC to the analysis of PAHs has been reviewed recently [40]. Since, compared with HPLC, GLC is much easier to apply in quantitative analysis and has an excellent separation capacity, GLC should always be used where this is possible as regards the volatility ofthe PAH and their thermal stability. On the other hand, HPLC should be employed in the range ofvery high molecular and/or thermally unstable PAH which cannot, at present, be analysed satisfactorily by GLC. Soluble PAH with mole masses of up to approximately 600 have been separated by HPLC [41, 42). The of thermally advantages of HPLC have also been exploited for the anlysi~ unstable PAH metabolites [43, 44]. HPLC can be performed with high selectivity not only by using suitable stationary phases [45], but also by the application of selective detection methods, for example fluorescence quenching [46]. On-line coupling of HPLC with UV spectroscopy has proved useful for the identification ofPAH in complex mixtures [41]. Although GLC and HPLC are superior to thin layer and paper chromatography regarding the separation capacity these methods are useful for obtaining rapid information on the concentration of distinct PAHs in complex samples and are therefore still widely applied in water analysis [47]. Compared with the Chromatographie analytical methods spectroscopic Polycyclic Aromatic and Heteroaromatic Hydrocarbons 119 methods are less important in environmental PAH analysis, although fluorimetry [48] and phosphorimetry [49] because oftheir excellent sensitivity and selectivity proved useful in special cases. Transport Behaviour in the Environment Air PAH in the atmosphere are predominantly associated with particulate matter, especially soot [50].1t has been speculated that the PAH appear tobe adsorbed primarily on the surface of soot by hydrogen bonding [51]. The benzene soluble portion of this material is approximately 10% by weight but the PAH component is much smaller than that. Partide size is the physical property with the greatest influence on the behaviour ofPAH-containing aerosols. Generally, the particle size spectrum of atmospheric aerosols extends from less than 0.01 Jlm to greater than 10 Jlm, but PAHs appear to be associated largely with particles less than 5 Jlm in diameter. There is considerable variation in the size-concentration distribution of particles with location in space and time, but the large-particle portion of the spectrum (greater than 0.1 Jlm) often tends, on the average, to follow a power law form: dN dDp -- = const. 0 D -4 P where N is the number of particles, DP the particle diameter, 0 the volume fraction and the constant is approximately 0.40. Although there is no reliable theoretical explanation, as yet, for the apparent regularity observed, some speculation has been reported [52, 53]. Particle shape [54] and density are important factors which determine the rate of aerosol deposition. F or urban aerosols a density range from 1.8 to 2.1 gfcm3 has been reported [55]. Particulate PAH released in the atmosphere in one location may be transported to very distant areas [56], whereby the local meteorology as well as the wind fields must be taken into account. In connection with the removal of particulate PAH from the atmosphere deposition of large particles by gravitational settling is important. For some time, it is possible to estimate roughly the deposition rate of particles on an obstacle if the air-flow field near it is known [57, 58] and some limited data have been reported for deposition rates on vegetation [59-61]. The development of rain clouds, on the other hand, influences aerosols containing PAH significantly, whereby the size distribution and the chemical composition as a function of size may be modified. When precipitation begins to fall from clouds, smaller particles will be deposited as in the case of the dry scavenging mechanism. This washout is believed to be significant in removing PAH from the atmosphere (50]. 120 M. Zander Water Although the solubility of pure PAH in water is extremely low (for example benzo[a]pyrene 4·10- 3 mg/1, dibenzo[a,h]anthracene 5·10- 4 mg/1) these compounds can be solubilized by other organic substances in particular detergents [62]. Besides that PAH are capable to form associates with colloids present in water andin this form can be transported through natural occuring water. Thus, PAH have been detected in tissues of organisms from marine habitats far removed from intensive human activity [63]. Chemical and Photochemical Reactions P AH can undergo various types of ground state reactions such as electrophilic and nucleophilic substitution, 1,2- and 1,4-cycloaddition reactions, oxydation, hydrogenation and intra- as well as intermolecular condensation reactions. The reactivity behaviour ofPAH has been comprehensively reviewed by Clar [17]. Most ofthese reactions are known for a long time, but only recently it has been well documented that various PAH and structurally related hetero-aromatic systems can also undergo Lewis acid catalyzed innermolecular skeleton rearrangements [64] this being a type ofPAH transformation that can complicate the synthesis of polycyclics. Since most of the reactions of P AH are addition reactions in the rate determining step or true addition reactions the reactivity behaviour can be quantitatively described by using the localization energy concept (see section on p. 116). In this context Dewar's pertubationa1 mo1ecular orbital method [26] and Fukui's frontierorbital method [65] are extremely useful to provide the organic ehernist with the relevant data. Since the Bell-Evans-Polanyi principle is valid for various reaction types of n-electronic systems free-energy relationships playadominant role in PAH chemistry. For the environmental ehernist reactions of electronically excited PAH (photochemical reactions) are ofparticular interest because the fate ofPAH under environmental conditions is determined to a high degree by its photochemical behaviour. Tricyclic or larger P AH and related heterocyclic systems have strong UV absorption at wavelenghts Ionger than 300 nm (present in solar radiation) and most are readily photooxidized. Photooxidation is probably one of the most important proccesses in the removal of polycyclics from the environment. Photooxidation of P AH in solution involves energy transfer from the triplet state of the aromatic system, producing singlet oxygen, which reacts with the compound, yielding its peroxide. Normally, for endoperoxide formation two anthracene 9,1 0-like positions are required (Fig. 12). Photolysis or pyrolysis of PAH endoperoxides yields a variety of reaction products via dealkylation and ring cleavage [66] as shown by the possible pyrolysis products of9,10-dimethyl-anthracene peroxidein Fig. 12.1t is important to note that quinones can also be produced when no endoperoxide can be formed for Po1ycyclic Aromatic and Heteroaromatic Hydrocarbons 121 0 ~ +~ u 0 !:::, or h'i- + Fig. 12. Pyro1ysis (photo1ysis) of 9, 10-dimethy1-anthracene peroxide [66] 0 0 II II + I 0 + II 0 Fig. 13. Quinone formation from benzo[a]pyrene by irridation (roJ r Fig. 14. Anthracene photodimer formation steric reasons. Thus, benzo[a]pyrene yields a mixture of three quinones by irridation in solution (Fig. 13). It has been speculated that photooxidation of PAH in the adsorbed state does not proceed via endoperoxides [67]. Although only few reports ofphotooxidation of adsorbed PAH have appeared it is quite evident that PAH are photoxidized with higher rates in the adsorbed state than in solution [50]. Since POM in the environment is mostly associated with particulate matter photooxidation studies on PAH in the adsorbed state [68] have gained particular relevance for the environmental ehernist Half-lives for photooxidation of PAR in particular of benzo[a]pyrene under various conditions have been estimated and on average are Iess than one day. 122 M. Zander Non-substituted acenes form easily photodimers by reaction of one PAH molecule excited to its singlet state with another PAH molecule in its ground state (Fig. 14). The primary step of this reaction is the formation of an excimer; since the reaction proceeds via the singlet state excited molecule dimerization does not take place under conditions that enhance intersystem crossing into the triplet manifold. Polansky [69] demonstrated how within the pars orbital concept the character orders of partial structures of electronically excited molecules can be derived. Here, the character orders are so defined that they describe the analogy between the reference compound in the ground state and the relevant partial structure in the electronically excited molecule. The pars orbital concept can be of use in the interpretation and prediction -of the photochemical behaviour ofPAH. However, the method has so far been applied only in the simple HMO fashion as it is restricted to PAH whose first absorption transition is the 1La transition (Clar's para band). For the environmental ehernist the behaviour of PAH towards agents commonly applied for the purification of dtinking water is of interest. While PAH dissolved in water will be oxidized by ozone, chlorinating agents mostly yield chlorine substituted PAH besides oxidation products. Benzo[a]pyrene, for example, during chlorination is transformed into a variety of products, none ofwhich is carcinogenic [70, 71]. Metabolism The metabolism of PAH in terrestrial mammals has been extensively studied, above all to achieve an understanding how PAH act as carcinogens. Although arene epoxides as the primary metabolites of PAH have been postulated already in 1950 [72] this was not definetively proven before 1968 [73, 74]. The mechanism of the carcinogenic activity of benzo[a]pyrene according to our present knowledge is summarized in Fig. 15. With the participation of cytochrome P 450 which is present in the endoplasmic reticulum of the cell benzo[a]pyrene is oxidized yielding the arene oxide J. Thus, in contrast with older theories [75] not the so-called "K-region" (K for Krebs) is attacked but a ring belonging to the "bay-region" of the PAH. During the next step the enzyme epoxide-hydratase [76] transforms 1 into the trans-dihydro diol 2, which then again undergoes an epoxidation with the participation of cytochrome P 450 yielding the "ultimate carcinogen" 3. The trans-diol epoxide 3 exists in two stereo isomers 4 and 5 and each of them can be separated into two pairs ofenantio isomers ( + )-4, (- )-4, ( + )-5, (- )-5.1t was fotind only recently that from these four molecules ( + )-5 exhibits a high carcinogenic activity in experiments with mice while the other three compounds and benzo[a]pyrene itself have minor or no activity at all (77]. The electrophile 3 ([ + ]-5) reacts with nucleophilic bases ofthe DNA, guanine being the favoured base for the attack of the diol epoxide. By using the simple PM 0 method [26] it could be shown that the formation of carbonium ions is energetically favoured in the bay-region of PAHs [78]. 123 Polyeyelie Aromatie and Heteroaromatie Hydrocarbons oB&-~ 7 0 1 OH 2 - 1;&9 J)86? ~lÖJ OH (-)-4 OH (+) -4 HO-lyiOOJ OH 3 ~ Ho·.vvur OH(+)-5 ~JÖl) H)JQTQT ' ÖH (-) -5 Fig. 15. Metabolie aetivation ofbenzo[a]pyrene to the "ultimate carcinogen" OH H OH OH Fig. 16. Metabolieformation of benzo[a]pyrene phenols Consequently, an epoxidering in the bay-region of a PAH should be the critical structure element of the ultimate carcinogen. This has been definetively proven with several modell substances in mutagenecity screening tests and animal experiments [79-81]. Urinary metabolites of PAH are usually phenols that arise by isomerization of arene oxides (NIH-shift) [73, 74] or by elimination of water from trans-dihydro diols (Fig. 16). But none of these mechanisms has been definitely proven. Moreover only one of the two possible isomeric phenols which in principle can be formed from the arene epoxide or trans-dihydro diol has been detected. Frequently transdihydro diols in vivo will be conjugated to sulphuric acid and then excreted as the monosulphates [82]. Also the excretion of phenols and dihydro diols conjugated to glucoronic acidwas observed [83, 84J. M. Zander 124 Quinones are formed insmall amounts as metabolites ofPAH. Benzo[a]anthracene-7,12-quinone and the 1,6-, 3,6-, and 6,12-quinone ofbenzo[a]pyrene have been observed [85], but the mechanism of quinone formation needs additional clarification. The formation of 6-hydroxymethyl-benzo[a]pyrene from benzo[a]pyrene in rat liver rnicrosomes was described by several authors [86-89]. However, neither the origin of the Crfragment nor the enzyme that catalyzes the C 1-transfer are known so far (Fig. 17). Fig. 17. C 1-transfer during benzo[a]pyrene metabolism With methylated PAH oxidation of the methyl groups occurs independently of the ring oxidation. In vivo oxidation of 7, 12-dimethyl-benzo[a]anthracene to 7-hydroxymethyl-12-methyl-benzo[a]anthracene and also of 3-methyl-cholanthrene to 3-hydroxymethyl-cholanthrene was observed [90]. N ormally the oxidation proceeds by formation of carboxyclic acids that occur as urinary metabolites (Fig. 18). COOH Fig. 18. In vivo oxidation ofmethylated PAH The conjugation of arene oxides with glutathione is considered to be a detoxification reaction ofPAH. This reaction is catalyzed by the glutathioneS-epoxide transferase system present in the cytoplasma (Fig. 19) and consisting of several isoenzymes [91-94]. K region epoxides are favourably conjugated with glutathione [95]. Polycyclic Aromatic and Heteroaromatic Hydrocarbons 125 Fig. 19. In vivo conjugation of arene oxides with gluthathione The conjugation of PAH with proteines has been well documented. The rate of conjugation decreases in the order epoxides, phenols, dihydro diols. There is some evidence that K region epoxides are favourably conjugated with proteines [96] and this obviously is a parallel to the reaction with glutathione. lt is well established that the reaction of PAH metabolites with proteines predominates the conjugation to nucleic acids [97] this latter reaction being the ultimate step in the malignant cell transformation. Biodegradation Bacteria can oxidize PAH that rangein size from benzene to benzo[a]pyrene but for more highly condensed PAH this is not clear (98]. Benzo[a]pyrene is oxidized by microorganisms, for example Beijerinckia sp., to form 7,8-dihydroxy-7 ,8-dihydro-benzo[a]pyrene and 9, 10-dihydroxy-9, 10-dihydrobenzo[a]pyrene with further moreextensive degradation [99] (Fig. 20). Microbial degradation is a major mechanism for compound removal in sediments and bacteria capable of degrading hetero-aromatic systems have also been isolated. Generally, benzo[a]pyrene metabolizing cultures are microorganisms with long exposure to benzo[a]pyrene [100]. Growth rates ofbacteria on PAH are directly related to the solubilities ofthe PAH. H OH Fig. 20. Oxidation ofbenzo[a]}Jyrene by microorganisms [99] The Overall Environmental Fate of PAH The overall environmental fate of PAH depends on several factors. - By way of example, Table 3 gives half-lives of dissolved benzo[a]pyrene for individual transformation or removal processes. Half-life for photolysis is two orders of magnitude smaller than the half-lifes of the other transformation pathways. M. Zander 126 Table 3. Half-lives of dissolved benzo [a] pyrene for individual transformation and removal processes [100] Process Photolysisa· Oxidation Volatilization All processes, except dilution River Pond (eutrophic) Lake (eutrophic) Lake (oligotrophic) (h) 3.0 >340 140 7.5 >340 350 7.5 >340 1.5 >340 2.9 7.3 7.4 1.5 700 700 a Summer sunlighf Sorption half-lives have not been measured, but they are probably at least 100 times smaller than the half-lives for photolysis. The degradation ofPAH in the atmosphere by photooxidation has been discussed in an earlier section (see section on chemical and photochemical reactions). The rate of the various transformation and removal processes which Iead to a reduction of P AH and hetero-aromatic systems in the environment is considered to depend on the physical and chemical properties of the individual compounds. Solubility and adsorbtivity are the most important physical properties in this context, while amongst the chemical properties photochemical reactivity is particularly relevant. These properties depend on size and topology of the systems and it is by no means surprising, that some PAH are readily destroyed under environmental conditions while others are remarkably stable. Moreover, it is interesting to note, that special types ofPAH are suspected to occur in interstellar matter [101].- Several mechanisms involving P AH under environmental conditions may cause reactive species to be delivered to genetic and other biological material, but this aspect needs further clarification. Since benzo[a]pyrene and various other PAH are lipophilic these compounds can be bioaccumulated to high Ievels. For example, high accumulation ofbenzo-[a]pyrene in clams was reported (24 h exposure to 0.0305 ppm 14C-benzo[a]pyrene in seawater) the tissue concentration of7 .2 ppm indicating a 236 fold bioaccumulation over exposure water concentration. This result is consistent with the observed slow release of benzo[a]pyrene : 29% remained after 10 days [63]. Lower bioaccumulation factors were observed with other species. Concentrations of benzo[a]pyrene in air, water, soil, sediments, aquatic organisms and food are listed in Table 4 and give a rough idea of the overall contamination of the environment with PAH. Toxicology In general, the acute toxicity of PAH and related heteroaromatic systems is low and is restricted to necrotisation of the glandulae suprarenales. Polycyclic Aromatic and Heteroaromatic Hydrocarbons 127 Table 4. Concentrations of benzo [a] pyrene in the environment and biota Occurrence Air (ng/m3) Sydney (Australia) year: Kopenhagen (Denmark) Salford (Great Britain) Paris (France) Toronto (Canada) Taschkent (Soviet Union) Los Angeles (USA) Berlin (Germany) Concentration 1962/63 1956 1953 1958 1961162 1965 1971/72 1970 Water (ng/1) tap-water ground water rain water Surface water Soil (f.Lg/kg) beach (Baltic sea) forest (Baltic sea) near high-way (Germany) Sediments (f.Lg/kg) Greenland (depth 0.2 m) Italy (highly industrialized area, depth 15-45 m) French mediterranean coast (depth 14m) Bodensee (Gerrnany) Aquatic organisms (f.Lg/kg) Oysters (French Coast) Musseis (France) Codfish and shellfish (Greenland) Plankton (Greenland) Algae (Italy) Food (f.Lglkg) Smoked meat and sausages Smoked fish Kaie winter: 8 17 summer: 0.8 5 210 300-500 5.4 6.4 110 ca. 1,3 18 2.5-9 1.0-10 2.2-7.3 130-350 (Thames river) 8 · 10-4 3.5 · 10·2 3.0 5 1,000-3,000 400 max.1,600 1-70 2-30 16-60 5 2 0.2-0.9 0.1-9.8 4 -16 Various PAH, heterocyclics and derivatives have been examined as for carcinogenicity in short-term screeningtestssuch as the Ames test [102-105], cell transformation test [106] and sebaceous gland supression test [107, 108]. Although a 90--95% correlation between the results from mutagenicity screening tests and carcinogenicity observed in animal experiments was claimed [109] it becomes increasingly evident that these in vitro tests cannot replace in vivo experiments with animals if reliable information on carcinogenic effects is required. M. Zander 128 It is well established that certain PAH and related heterocyclics produce skin cancer in test animals and there are various hints that the same might be true for lung cancer [110]. Some proven polycyclic carcinogens are listed in Fig. 21. Although cocarcinogens such as long chain alkanes [111] play an H;P CH3 7.12-Dimethyl-benz(a]anthracene Di benzo[a.h [ pyrene Dibenz[a,j [acridine Benzo[a] pyrene 3 Methyl-cholanthrene Dibenzo[a,i ]pyrene Benzo [ b] fluoranthene N H Dibenz [ c,g] carbazole Fig. 21. Some proven carcinogenic PAH and related heterocyclic systems important ro1e in PAH carcinogenesis the effects ofthe PAH's itselfprobably exhibit an additive behaviour [2]. A carcinogenic risk to man from chemieals in the environment can in principle evaluated only from epidemio1ogical studies [2, 50] or under certain circumstances from case studies. Since man is exposed to a host of chemieals in the environment unambigous causejeffect relations cannot be deduced from epidemiological studies for specific compounds and even notforadass of compounds. Similarities of metabolism of benzo[a]pyrene in human and mouse cells cultured in vitro have been reported. However, the relevance of this finding for evaluating the risk to man cannot yet be assessed [110]. References I. Collin, G., Mildenberg, R.: Chem. and Ind. 1978, 567 2. Grimmer, G.: in: Luftqualitätskriterien für ausgewählte polycyclische aromatische Kohlenwasserstoffe, ed. Umweltbundesamt (FRG), 1979 3. Suess, M.J.: Sei. Total Environ. 6, 239 (1976) 4. Gräf, W.: Med. Klinik, 15, 561 (1965) 5. Borneff, J., Selenka, F., Kunte, H., Maximos, A.: Environ. Res. 2, 22 (1968) Polycyclic Aromatic and Heteroaromatic Hydrocarbons 129 6. Blumer, M.: Sei. American 1976, März, p. 34 7. Arcos I.C., Argus, M.F.: Chemical Induction of Cancer, vol. IIA. Academic Press, New Y ork - London 1974 8. Badger, G.M.: Proc. Roy. Soc. 1956, 87 9. Lang, K.F., Buffieb, H., Zander, M.: Erdöl/Kohle 16, 944 (1963) 10. Zander, M.: Intern. J. Environ. Anal. Chem. 4, 109 (1975) II. Burrows, I.E., Lindsey, A.J.: Chem. Ind. 19611395 12. Stelzner, R., Kuh, A.: Lit. Reg. Org. Chem. 3, 21 (1921) 13. Patterson, A.M.: J. Amer. Chem. Soc. 47, 543 (1925) 14. IUPAC-Rules 1957 15. The Ring Index. Amer. Chem. Soc., Washington 1960 16. Lunnon, W.F.: in: Graph Theory and Computing. Academic Press, New Y ork- London 1972 p. 87 17. Clar, E.: Polycyclic Hydrocarbons. Academic Press New York-London 1964 18. Coulson, C.A., Rushbrooke, G.S.: Proc. 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Purchase, I.F.H. et al.: Nature 264, 624 (1976) llO. IARC Monögraphs, Vol. 3, Lyon 1973 111. Bingham, E., Falk, H.L.: Arch. Environm. Health 19, 779 (1969) Fluorocarbons J. Russow Hoechst Aktiengesellschaft D-6230 Frankfurt am Main, Federal Republic ofGermany lntroduction "Fluorocarbons" is a terro used to designate partially or coropletely halogenated alkanes containing fluorine, and with one or roore, but priroarily 1 and 2, C atoros. The general forroula is CnH2n+ 2-x-y ClxFy, where X+ y~ 2n + 2. Table 1. Physical data of fluorcarbons Formula Short symbol bp [0 CCI3F CCI2F2 CCIF3 CHCIF2 CCizF · CCIFz CCIF2 · CCIF2 Fll F 12 F13 F 22 F113 F 114 Cj at 1 bar 23.8 -29.8 -81.4 -40.8 47.6 3.6 Vapor pressure (bar) [20 °C] 0.9 5.8 32.4 9.4 0.4 1.9 The roost iroportant roerobers of this group are listed in Table 1 along with their physical characteristics [1]. Because of their technical iroportance and wide-spread use the coropounds are also indicated by the short designation used here, which consists of a series of three digits. 1st digit froro right nurober ofF atoros 2nd digit froro right -1 nurober ofH atoros 3rd digit froro right + 1 nurober of C atoros (oroitted where nurober ofC-atoros = 1). J. Russow 134 Cl atoms contained in a fluorocarbon molecule are not indicated in this system. The compounds which are by far the most important are the compounds F 11 and F 12; they are used principally as aerosol propellants, refrigerants and as blowing agents for plastic foams. Furthermore, F 22 has assumed greater importance as a refrigerant and an intermediate product for the manufacture of the technically extremely important polymeric perfluorinated hydrocarbons polytetrafluoroethylene (PTFE) and its various copolymers. The fluorocarbons first attained environmental significance when they were detected in extremely low concentrations, in the ppt range, in the atmosphere [2, 3]. Owing to their stability they can be used as inert tracers in studies of air flow and dispersion conditions in the lower atmosphere [4]. In recent years fluorocarbons have attained greater topical importance through the hypothesis of Rowland and Molina, according to which the fluorocarbons are responsible for depleting the stratospheric ozone layer, which may have adverse effects on the life inhabiting the earth's surface [5-7]. The scientific discussion of this set of problems, which has led to intensive research on the chemical, photochemical, and dynamic transport processes in the stratosphere, has by no means come to an end, there are still very serious discrepancies between theory and measurement, so that final judgment cannot yet be made [8-11 ]. Production and Use Very recently there havebeen indications that fluorocarbons may be produced in nature, although the amounts detected are negligibly small [12, 13, 69]. Large-sca1e industria1 production began in the 1930s. Tab1es 2 and 3 summarize world production figures for F 11 and F 12. These figures are based on a survey made by the Manufacturing Chemists Association (MCA), WashingTable 2. Fluorocarbon 11 production and release (world total) in million kilograms Year 1940 1950 1960 1970 1971 1972 1973 1974 1975 1976 1977 Production Release Peryear Cumulative Peryear Cumulative 0.2 6.6 49.7 241.1 266.6 310.5 354.3 377.6 322.5 349.9 330.7 0.7 18.6 286.9 1699.1 1965.7 2276.2 2630.5 3008.1 3330.6 3680.5 4011.2 0.1 5.4 39.7 205.1 225.4 253.8 290.6 320.9 312.4 303.6 305.6 0.4 14.4 24.79 1435.2 1660.6 1914.5 2205.0 2526.0 2838.3 3141.9 3447.5 Fluorocarbons 135 Table 3. Fluorocarbon 12 production and release (world total) in million kilograms Year Production 1940 1950 1960 1970 1971 1972 1973 1974 1975 1976 1977 Release Peryear Cumulative Peryear Cumulative 4.5 34.6 99.4 336.9 360.5 401.7 447.5 473.6 419.7 449.8 424.4 18.8 198.3 828.3 2957.5 3318.1 3719.7 4167.2 4640.8 5060.5 5510.4 5934.7 1.7 27.1 83.2 296.2 319.1 348.3 386.2 420.3 412.6 396.3 376.5 13.3 129.5 631.9 2475.2 2494.3 3142.6 3528.8 3949.1 4361.7 4758.0 5134.5 ton, D.C., (USA), among 20 manufacturers within the countfies outside the Bastern bloc, which account for 95% of the total world production [14-16]. The tables also show the emissions into the atmosphere which are accounted for by the use ofthese products. F 11 and F 12 enter the atmosphere unaltered after a Ionger (e.g., refrigerants) or shorter (e.g., aerosols) delay. According to a careful analysis of their use 85% of current production will be emitted into the atmosphere within a year. Of the total cumulative production today more than 95% has been released into the atmosphere [15, 16]. The applications of the fluorocarbons are a consequence of their special chemical and physical properties: they are chemically and thermally very stable and thus practically inert, and they predominantly occur as gases which can be easily liquefied under pressure. The main applications are: a) propellant for aerosols (inert, non-flammable, practically non-toxic) b) refrigerant for air conditioning and refrigeration systems (favourable thermodynamic properties) c) blowing agent for the manufacture ofplastic foams (good heat insulating properties ofthe entrapped gas) Table 4. Fluorocarbon 11 and fluorocarbon 12 sales and uses for year 1977; 20 reporting companies outside of the eastem block representing 95% of the total world production and sales in million kilograms Application field Fll F 12 Refrigeration Blowing agent Aerosol propellant Other uses Totalsales 24.7 106.9 164.6 19.1 315.3 154.3 20.6 174.5 33.5 382.9 J. Russow 136 d) solvent for dry cleaning and for cleaning of high-grade electronic components (non-corrosive, high dissolving power). Table 4 shows the 1977 sales for the Western world broken down by areas of application [14, 15]. Chemistry The starting materials for the manufacture ofthe fluorocarbons are generally the corresponding chloroalkanes, which are produced by the chlorination of alkanes, e.g. 2 C~+3l -CHC13 +3HCI CH4 + 4 Cl2 - - - + CC4 + 4 HCI. (1) (2) In the presence of a catalyst fluorine from hydrogen fluoride is substituted for chlorine in the chloroalkanes, forming fluorocarbons and hydrogen chloride, e.g. CHC13 + 2 HF CC4 + 2 HF CC4 + HF (SbXsl - (AIF3J (CrOFJ CHCIF2+ 2 HCI (3) F22 CCI2F2 + 2 HCl (4) F 12 CC13F + HCI. (5) Fll The catalysts used are a) antimony(V) halides in a homogenous Iiquid-phase reaction and b) solid fluorides like AIF3 or CrOF in a heterogeneaus gas-phase reaction [1]. For the technically important products the conversion rates and yields are practically 100%. The mixture of reactants must be carefully prepared. Apart from the catalyst, which must be renewed occasionally, there are practically no unusable by-products. The Iosses of this type of gas, which is liquefiable under pressure, are also very low; at the present state of the art total Iosses during production, storage, filling and transpoft are expected to be about Y2%. There is little to be said conceming the chemical reactivity of these compounds because the commercially most important compounds are manufactured and used just because of their excellent thermal and chemical stability, which increases with increasing fluorine content. The compounds are nonflammable and practically non-toxic provided they are, like those mentioned in the introduction, fully halogenated. In view of their applications these compounds are usually handled in a very pure form (absolutely acid-free, water content <0.001 %, evaporation residue <0.005 Vol. %), so that their handling creates no problems, such as corrosion due to impurities. They are not known to react chemically in any way. Fluorocarbons 137 Analytical Methods With two exceptions the fluorocarbons are in gaseous form at room temperature and normal pressure, but they are stored and transported in liquefied form under pressure. In practice the only quantitative determination of fluorocarbons that is of any importance is for the purpose of determining the purity and the concentration in gas mixtures, andin both cases the fluorocarbons are present in relatively high concentrations. In recent years quantitative determination in the trace range has become of major interest, as, for example, for the determination of concentrations in the atmosphere. 1971 Lovelock succeeded in quantitatively determining F 11 and F 12 in the range from 50 to 100 ppt (1 ppt corresponds to 1: 1012) by gas chromatography employing an electron capture detector [2, 3]. Several analytical methods have been developed which are similar in principle, but are distinguished from one another by the difference in the enrichment steps taken before GC analysis. An important problern of quantitative flurocarbons determination is the preparation of suitable standard mixtures, and different techniques are proposed for doing this. The present state of the art, which only a few project groups have mastered, permits a high precision in analytical results with standard deviations of a few percent. In contrast the accuracy is much lower, with a margin of error of probably more than ± 10%, which above all makes comparison difficult among the data of the various work groups [17]. A paper by Sze and Wu [18] summarizes the older determinations of concentration in the atmosphere which were available at about mid 1976; Table 5 was taken from this paper. A cautious approach to these data, made necessary by the level of analytical accuracy and comparability and the degree to which samples are free of contamination, allows us to draw three conclusions: in the observation period from 1971 until1974 the concentrations ofF 11 and F 12 in the air near the ground increased by a factor of about 1. 5; the concentration exhibits a distinct decline from the northern to the southern hemisphere, which is related to the higher rate of emission in the northern hemisphere (see below); the concentration is constant within the troposphere and decreases with increasing altitude in the region ofthe tropopause (8-13 km altitude) andin the stratosphere. A more recent compilation by Sandalls [29] contains some additional concentration data. Transport Behaviour in the Environment The fluorocarbons are emitted into the lower atmosphere (troposphere) during or after their use. In accordance with their areas of application and the given economic conditions, the fluorocarbons are released largely in the northern hemisphere. Sales surveys have yielded the emission figures for F 11 and F 12, the two most important types, as shown in Tables 2 and 3 [14--16]. Table 5. Measurements of fluorocarbon concentrations units Author = ppt in vol (from [18]) w 00 F-ll F-12 Time and location Remarks 40-45 34.0 31.0 28.6 24.5 23.0 16.9 13.0 12.0 18.3 <0.2 3 <5 35 48 Sept. 1973 June 1975 By balloon-bome cryogenic sampling system 12.2 75 140 80±3 125 ± 7 Attitude (km) Heidt et al. [19] Hester et al. [20] 6.4 Krey and Lagomarsino [21] 13.7 15.2 16.8 18.3 19.2 12.2 13.7 15.3 16.8 18.3 19.2 9 ll 18 45 95 83 94 60±4 59a 70 65 57 47 69 77 77 66 67 41 - TX (32 "N) 86 133 78 98 ± 18 Sept. 1973 23 May 1974 36.15-39.30 Lat 106.17-106.45 Long 23 May 1974 33.10-34.34 Lat 104.30-105.10 Long May 1974 34.45-33.50 Lat 106.20-105.00 Long l Aprill974 (60 "N-37 "S) Average value from 2 flights Average value from 2 flights Compressed air sample, data analyzed by gas chromatography !-< October 1974 (75 "N-10 "S) ~ e "'0"' :E Author Altitude (km) Lovelock [22] Sutface F-11 F-12 Time and location Remarks äl... 101.7 June/July 1974 W. Ireland October 1973 N. Atlantic June 1974 Central England Sept 1974 Capetown, S. Africa Nov.-Dec. 1971 50 "N-60 "S June 1975 By gas chromatography ...c:r 0 0 t') 79.8 88.6 115.2 101-119 57 49b Lovelock et al. (1973) [2] SehrneUekopf et al. [23] Goidan et al. [24] Williams et al. [25] Rasmussen [26] Wilkniss et al. [27, 28] 26.2 ± 1 22.3 ± 0.7 17.7 ± 0.5 Surface 1-4km 10km 14km 18.5 km 22km 25.5 km 21 18.5 18.5 15 Surface at surface at surface at surface at surface at surface <20 30±3 80± 10 48±5 150 160 ± 15 120 ± 15 1.3 ±83 10 ± 1 7.6±1 23 49 20 110 120-130 43 53 72 61 80 75±5 135 ± 10 210 ± 10 90±10 330 230 ±30 225 ±30 25±3 84±8 100 ± 10 120 160 50-60 140 210-230 = "' Oceanographic cruise of Shackleton Latitude Stratospheric measurements August 1975 Sept. 1975 August 1975 August 1975 August 1975 August 1975 August 1975 26 Sept. 1975 Preliminary results 26 Sept. 1975 12 Aug. 1968 26 Sept 1975 May 1975 Nov. 1971 (10 "S) By gas chromatography Nov.-Dec. 1972 (10 "S) Mar.-Apr. 1974 (10 "S) Dec. 1972 (20 "S-80 "N) Mar.-Apr. 1974 (20 "S-20 "N) Mean concentration averaged over latitudes b Mean aerial concentration averaged over 50° N -60° S. Concentration rang es from 70 ppt at 50° N to 38 ppt at 60° S a ~ 0 .... ~ 1,0 J. Russow 140 These tables reveal the increase in F 11 and F 12 emissions up until1974. Since that time emissions have remained constant or declined slightly; the reason for this is the decline in the use of fluorocarbons in aerosols for economical (substitution by propane/butane) and environmental (fluorocarbons-ozone hypothesis) reasons. The emission of fluorocarbons is also reflected in the concentrations measured in the atmosphere. Whereas the first measurements from the year 1971 yielded concentrations of 40-50 ppt F 11 [3, 27, 28] (cf. Table 5), those made in mid 1978 show a concentration of about 160 ppt F 11 and about 280 ppt F 12 [30]. For the two-year period from November 1975 to November 1977 concentration measurements between 35° and 65° north latitude determined an average growth rate of 12.9 pptfyear or 12%/year for F 11 and 18.5 pptfyr or 10%/yr for F 12 [31]. With regard to sampling it is always difficult to know whether the sample contains the true background concentration or a very much higher concentration because, owing to meteorological conditions, the sample originated in an area with high industrial and population densities, so that it exhibits excessive F 11 and F 12levels. For example, Packet al. [4] have shown that occasionally greatly increased fluorocarbon concentrations at a very remote sampling point are due to these air masses originating in industrial areas. Figure 1, taken from Fraser [32], is .. .. 140 A 0 > 120 ä. a. ~ r -_ 80~-.r April 0 -_ _:.o·~-;0 0 • _.-r;·~ __ ~ 0 I • July Aug. 1976 Sept. Oct. Nov. 6 #~" _., !· i§i<>o,p• o o /..Q.ooo• <> <> <> a> Aircraft type o 0 June ~<l._>·- 0 <> <\> 0 o • • May r_...8 • 001?-A"-·-·-·-·-·<> -·""'0·-·-·(5" ü"' 100 • • • Dec. F 27 o • 8 727 PA 39, CV 990 8 747 A 8 707 o Surface-Cape Grim Jan. Febr. March 1977 Fig. 1. Variation with time of observed CC13F concentration in surface air at Cape Grim andin aircraft air samp1es. Best fit curves: aircraft data. In [CC13F) = 4.65 ( ± 0.02) + 5.0 ( ± 0. 7) 4 t, r2 (coefficient of determination) = 0.51; -·-·- Cape Grim data. In [CC13F] = 4.60 X (±0.03) + 5.1 (±0.7) x 1(}" t, r2 = 0.49; t = days since 31/12/75. Data points are averages (from (32]) w- a typical example for an increase in F 11 concentrations over time. At the same time it also shows the scatter of the empirical data. Within the troposphere the fluorocarbons disperse very quickly, so that the concentration within the troposphere is uniform (cf. Fig. 2 from Fraser [32]). This diagram also clearly reveals the emission, and hence the increased concentration, over heavily populated and industrialized areas. Above the troposphere, however, the concentration decreases with altitude, which is due to stratospheric decomposition mechanisms. Figure 3 is given as a typical example of concentration Fluorocarbons 141 N.W. Tasmania area MelbourneWesternport area 12 0 tropopause ht. CV 990 (in area) 1:!. CV 990 (near area) Apr~. 0 10 0 PA 39 + Ground station (Aspendale) @ a 8 E ~ -~ Ground station (Cape Grim) Cape Grim flasks ~ 0 -·6 ... J: 0 4 0 2 120 130 140 150 200 250 300 CCI3F concentration, pptv Fig. 2. Vertical profiles of CCI 3F observed over NW Tasmania and Melbourne on 11/11/76; air samples from Cv990 by co-operation with NASA-Ames (from [32]) LAT. LONG. 073.78°N 159±3° w <>60-61°N 2-D o 29-34° N 20° l>. 8.16°N 75°----- 25 20 E ""'<i 15 Tropopause 10° N "0 ::::> <( 10 Tropopause 71° N 5 0~-r., 1 5 10 20 30 50 CFCI3 Mixing Ratio, pptv 100 200 Fig. 3. Lower stratosphere mixing ratios of CFCI 3 at various latitudes in the Northern Hemisphere. Tropopause altitudes estimated from National Meteorological Center analysis. Tropospheric mixing ratios from airborne (CV -990) measurements. 2-D curves from Borucki [68] (from [33]) J. Russow 142 profiles for F 11; it was taken from Vedder et al. (33]. Further profiles can be found in [18, 19, 34]. Since fluorocarbons are emitted largely in the northern hemisphere and the period of atmospheric exchange between the two hemispheres is about 1 to 2 years, the concentration measurements also exhibit a north-south gradient. Figure 4 is such a north-south profile, taken from Lovelock [3], and is presented as a characteristic example. Very recent measurements yield flatter gradients [31], because emission has ceased its rise and is now declining (cf. Tables 2 and 3). 80 70 . ·'~ . ...,. \ 60 ...... > ~ iso c .2 ~40 d ·~ •' '!t• •• • 0 c c"' 830 u ......... ..... 0 0 0 0 0 20 oo 0 0 40 20 N 0 Latitude 20 40 000 60 5 Fig. 4. Distribution of CCI3F in and over the North and South Atlantic Ocean. e Aerial concentrations ( x Hl12) by volume. o Seawater concentration ( x to-12); as aerial concentrations in equilibrium with water. Theoretical prediction. --- Best fit third degree polynomial (from [3]) Chemical and Photochemical Reactions The compounds being considered here are characterized by high chemical stability. Forthis reason the fluorocarbons arenot involved, for example, in the chemical processes of smog formation in the lower troposphere. In connection with the ozone depletion hypothesis, which maintains that after the fluorocarbons diffuse into the stratosphere they are photochemically degraded there, and the reaction products formed react with the stratospheric ozone through a chain mechanism, an intensive search has been conducted for degradation mechanisms in the lower atmosphere ( troposphere). 143 Fluorocarbons Any decomposition of fluorocarbons within the troposphere even if this is only at a rate of a few percent per year diminishes the flux of chlorine containing compounds into stratosphere and reduces the predicted ozone depletion attributed to fluorocarbons drastically because of non linear relations between ozone depletion and tropospheric residence time. Thus, for example, hydrolysis is just barely measurable in aqueous solution (,.., I0- 7 mol/l.yr), and it can be catalytically accelerated by traces of heavy metals [35]. Solubility in waterat 20 ac amounts to 1,500 ppm for F 11 and 600 ppm for F 12. Solubility in seawater is somewhat lower [36). Because of the low partial pressure in the atmosphere, the concentration in seawater, at 0.1--0.4 x I0- 9 g/1, is very low [31, 37], so that the oceans do not play any appreciable role as a sink [31, 36, 38]. Various studies have led to the conclusion that F 11 or F 12 is degraded in a heterogeneously catalyzed photolytic process on solid surfaces like those of silica gel, sand or desert dust under exposure to ultraviolet light [39-42). According to more recent studies decomposition also occurs without light (dark reaction) if the substrates are very dry [43, 44]. Decomposition is accelerated by the presence of oxygen. Carbon dioxide has been definitely detected as a decomposition product [40, 41].1t is entirely unknown whether such decomposition mechanisms also play a role under natural conditions, as in desert areas, in determining the residence time in the troposphere [45, 46], as long as no quantitative evaluation of the experiments is available for atmospheric conditions. Comparison ofthe amounts ofF 11 and F 12 emitted into the troposphere, the amounts which diffuse into the stratosphere and the amounts just analytically detectable in the troposphere, which, however, are very unreliable because of the low accuracy, leads to the conclusion that no decomposition occurs in the troposphere [31]. As a result the atmospheric residence time is calculated tobe 40-45 yr for F 11 and 65-70 yr for F 12. On the other hand, evaluation of the north-south gradient in the troposphere, the uncertainty in the profile used for diffusion in the stratosphere and an overall stratospheric chlorine balance support the view that the possibility of shorter atmospheric residence times and thus also of decomposition in the troposphere cannot be excluded at this time [18, 47-49]. In an experiment designed on a grand scale (Atmospheric Lifetime Experiment, ALE) the concentration is monitored constantly at four measuring stations scattered around the globe in order to obtain information on the atmospheric residence time [50]. On the basis ofthe facts that are known now it must be assumed that a major part of the fluorocarbons emitted so far has accumulated in the atmosphere, even if the major part is 95% this still has a very significant effect on the predictions concerning stratospheric ozone depletion. Whereas the fluorocarbons in the lower atmosphere (troposphere) are relatively stable, in the upper atmosphere ( stratosphere) they can be decomposed photochemically by the more energetic (shorter wavelength) UV -radiation [5, 6]. In the first step (6) CF2Cl2 + h v - "CF2Cl +"Cl (..1.<230nm) CFC13 + hv - "CFC12 +"Cl (..1.<230nrn) (7) J. Russow 144 Cl radicals are split off. lt is highly probable that the reaction with oxygen molecules follows this step: "CF2C1 + 02 - CF20 + ·c10 (8) .CFC12 + 0 2 - CFC10 + .C10 (9) At present little is known about the subsequent fate of the compounds CF20 and CFClO [51]. From these reactions the following average photodissociation lifetimes of F 11 and F 12 in the atmosphere have been calculated [6]: Average photodissociation lifetimes Fll F 12 Altitude 3 X 105 yr 6.6 yr 1.3 months 4.7 days 2.1 days 4XJ06yr 66 yr 11 months 1.3 months 16 days lOkm 20km 30km 40km 50 km These figures Iead to average atmospheric residence times of 50 and 100 yr, respectively. Photochemical decomposition influences the shape of the concentration profile in the stratosphere, as Fig. 3 shows. Above the tropopause at an altitude of 8-13 km the concentration declines relatively rapidly with increasing altitude. The "Cl radicals formed by photolytic decomposition may- as stated by the fluorocarbon ozon hypothesis - interfere with the dynamic equilibrium, which is determined essentially by the ozone formation reaction (10, 11), the decomposition reaction (12) and the catalytic decomposition cycles (13 + 14 = 15), in which the cata1ytically effective radicals can be X= "NO, "Cl, and "HO. Ozone formation: Ozone decomposition: 0 2 +hv - 2 0 (A.<240nm) (10) 0+0 2 +M - 03 +M (11) 03+0 -202 (12) -o2+·xo (13) - o2 +·x (14) -202 (15) 03+ X ·xo+o 03+0 There can be no doubt that fluorocarbons diffusing into the stratosphere and their subsequent photolysis increase the total amount of chlorine in the stratosphere, but whether the fluorocarbons cause depletion of the ozone or not remains an open question. As scientific research on the atmosphere has progressed, the margin of error of the calculated hypothetical ozone depletion Fluorocarbons 145 by fluorocarbons has increased: for the equilibrium condition in 50-100 yr at the present emissionrate of 0.33 Mt F 12/yr and 0.27 Mt F 11/yr the ozone depletion is currently calculated to be 15-2(}% with a probability range of 0-40% [10]. Furthermore, there are obviously several irreconcilable discrepancies between theoretical calculations and measurements and they have a bearing on factors which regulate the equilibrium system, which suggests that it is not yet possible to completely describe the chemistry of the stratosphere. On the other hand the ozone measurements and the statistical evaluation of the data, which are subject to large natural fluctuations, have not yet supplied proofthat ozone depletion has occurred, although according to the calculations this should already be the case [52-56]. Metabolism Only a few studies of fluorocarbon metabolism are known. Such studies are probably oflittle interest because, for one, we know that if these compounds have been inhaled in high concentrations and preferentially absorbed by the blood, they will be practically completely eliminated within an hour if fresh air is supplied and do not accumulate [57-61]. Although experiments with Cl 4-labeled F 11 and F 12 have demonstrated a low residual activity in the organism, it is unknown whether metabolites are actually involved in any way [62]. Biodegradation Studies made so far of decomposition on soil samples under natural conditions (microorganisms and flora) have not yielded any clear-cut results. If there is any degradation at all, it must be very little at normal atmospheric concentrations [63]. Accumulation There are no known indications that the fluorocarbons accumulate in living organisms. It is also hardly tobe expected because oftheir high volatility. Biological Effects and Toxicity The compounds exhibit only very low toxicity. The toxicological data are well known because relatively high local concentrations can occur when they are used. Their potential hazard when they are abused, for example, as in "sniffing", has been carefully studied. Gulden [64] and Paulet [58] have summarized information about fluorocarbon toxicity. An acute two-hour inhalation test has yielded a tolerable inhalation concentration of 1.25 vol.% for F 11 and 10 vol.% for F 12 (rat, 146 J. Russow guinea pig). If these concentrations are exceeded, narcotic effects appear, which are tolerated by the experimental animals without any permanent darnage [65]. Only very high concentrations can cause death: LC 50 over 30 min for F 11: 10-25 vol. %, depending on the animal species; LC 50 over 30 min for F 12: 76-80 vol. %, depending on the animal species [58]. In subacute inhalation trials with the compounds F 11, 12, 22, 113 and 114 in rats (1 vol. %) or dogs (0.5 vol. %) over 90 days with 6 h exposure daily no effects could be determined in either the external or histological findings [66]. According to these findings these compounds must be classified as practically non-toxic or relatively harmless under the Hodge and Sterner system [67]. The threshold Iimit value (TL V) has been set at 1,000 ppm for most fluorocarbons. A number of studies have been concerned with intake via the respiratory route. A few minutes after exposure the fluorocarbons can already be detected in the blood in concentrations of a few llgfml [60]. Arrhythmias appear only when the blood Ievel exceeds 40 llg/ml, a value which can be attained only when extremely high concentrations are inhaled [58]. After cessation ofinhalation exposure the fluorocarbons are eliminated within an hour via the expired air. At present it is not yet certain whether to some small extent these compounds are metabolized (cf. the section on Metabolism). The LD 50 value in rats for the two liquid fluorocarbons F 11 and F 113lies above 15 gjkg, which means that these compounds fall into toxicity dass 6 relatively harmless [64]. The corresponding values for the other fluorocarbons, which are gases under normal conditions, cannot be determined. References 1. Ullmanns, Encyklopädied. Techn. Chem. 4th ed. (1976) Vol.11, 632, Verlag Chemie GmbH, Weinheim. Topics Curr. Chem. 14, 129 (1970) 2. Lovelock, J.E.: Nature230, 379 (1971) 3. Lovelock, J.E., Maggs, R.J., Wade, R.J.: Nature, 241, 194 (1973) 4. Pack, D.H., Lovelock, J.E., Cotton, G., Curthoys, Ch.: Atmospheric Environment 11, 329 (1977) 5. Molina, M.J., Rowland, F.S.: Nature 249, 810 (1974) 6. Rowland, F.S., Molina, M.J.: Atomic Energy Commission Rep. No. 1974-1, Irvine, California USA, Sept. 5, 1974 7. Crutzen, P.J.: Geophysical Res. Lett. 1, 205 (1974) 8. US-National Acad. Sei.: Response to the Ozone Protection Sections of the Clean Air Act Amendments of 1977: An Interim Report. Washington 1977 9. Sze, N.D., McElroy, M.B., Wofsy, S.C., Kong, D., Daesen, R.: Atmospheric and Environmental Res., Inc., Cambridge, USA; Report Oct 1978 10. Ehhalt, D.H.: Chlorfluormethane und ihr Einfluß auf die stratosphärische Ozonschicht. Bericht im Auftrag des Umweltbundesamtes Jülich, Germany, Sept. 1978 11. Boville, B.W., Evans, W.F.J.: Current Status ofthe Stratospheric Pollution Problem; UNEP Co-ordinating Committee on the Ozone Layer, Second Session Bonn, Germany Nov 28-Dec 1, 1978 Fluorocarbons 147 12. Prinn, R.G., Barshay, S.S.: Halocarbons and other minor species in volcanic emissions: Theoretical considerations. Rep. CAP Associates to the Manufacturing Chemists Assoc., Bedford, Mass. USA, May 18, 1978 13. Rasmussen, R.A.: Update on EX-GC and GC-MS Analyses of Trace Gases in Hawaiian Voleanie Emissions. Spec. Rep., Beaverton/Oregon USA, Dec 21, 1978 14. Alexander Grant & Company, Washington: 1977 World Production and Sales ofFluorocarbons FC-11 and FC-12, Manufacturing Chemists Assoc., June 26, 1978 15. Manufacturing Chemists Assoc., Fluorocarbons Technical Panel, D.C.: World Production and Release of Chlorfluorocarbons 11 and 12 Through 1977, July 17, 1978 16. McCarthy, R.L., Bower, F.A., Jesson, J.P.: Atmospheric Environment 11,491 (1977) 17. Rasmussen, R.A.: Interlaboratory Comparison of Fluorocarbon Measurements. Preprint, submitted to Atmospheric Environment (1978) 18. Sze, N.D., Wu, M.F.: Atmospheric Environment 10, 1117 (1976) 19. Heidt, L.E., Lueb, R., Pollock, W., Ehhalt, D.H.: Geophys. Res. Lett. 2, 445 (1975) 21. Krey, P.W., Lagomarsino, R.J. (1975): Health and Safety Lab., ERDA, Environmental Quarterly Rep., HASL-294, 97-123 22. Lovelock, J.E.: Nature 252, 292 (1974) 23. Schmeltekopf, A.L., Goldan, P.O., Henderson, W.R., Harrop, W.J., Thomson, T.L., Fehsenfeld, F.S., Schiff, H.I., Crutzen, P.J., Isaksen, I.S.A., Ferguson, E.E.: Geophys. Res. Lett. 2, 393 (1975) 24. Goldan, P.D. (1975): Private communication 25. Williams, W.J., Kosters, J.J., Goldman, A., Murcray, D.G.:· Paper presented at AGU meeting, San Francisco, Dec. 1975 26. Rasmussen, R. (1975): Private communications 27. Wilkniss, P.E., Swinnerton, J.W., Bressan, D.J., Lamontagne, R.A., Larson, R.E.: J. Atmos. Sei. 22, 158 (1975) 28. Wilkniss, R.A., Swinnerton, J.W., Lamontagne, R.A., Bressan, D.J.: Science 187,832 (1975) 29. Sandalls, F.J., Hatton, D.B.: Atmospheric Environment 11,321 (1977) 30. Krasnec, J.: Paper at the AGU meeting Dec. 1978, San Francisco 31. Singh, H.B., Salas, L.J., Shigeishi, H., Scribner, E.: Science 203, 899 (1979) 32. Fraser, P.J.B., Pearman, G.I.: Atmospheric Environment 12, 839 (1978) 33. Vedder, J.F., Tyson, B.J., Brewer, R.B., Boitnott, C.A., Inn, E.C.Y.: Geophys. Res. Lett. 5, 33 (1978) 34. Seiler, W., Müller, F., Oeser, H.: Pure and Applied Geophys. 116, 554 (1978) 35. Grewer, T.: Interna! Rep. Hoechst AG, Frankfurt- Germany 1962, 1966 36. Junge, C.: Naturforsch. 31a, 482 (1976) 37. Rasmussen, R.A., Pierotti, D., Krasnec, J., Halter, B.: Trip Report on the Cruise ofthe Alpha Helix Research Vessel from San Diego, California to San Martin, Peru March 5 to 20. 1976 38. Liss, P.S., Slater, P.G.: Nature 247, 181 (1974) 39. Ausloos, P., Rebbert, R.E., Glasgow, L.: J. Res. Nat. Bur. Standards, Washington, 82, 1 (1977) 40. Gäb, S., Schmitzer, J. Thamm, H.W., Korte, F.: Angew. Chem. 90, 398 (1978) 41. Gäb, S., Schmitzer, J., Thamm, H.W., Parlar, H., Korte, F.: Nature 270, 331 (1977) 42. Rasmussen, R.A.: Halocarbon Photolysis over Sand in Ambient Air. Special report, Nov. 1978 43. Gäb, S.: personal communication 44. Ausloos, P.: personal communication to MCA 45. Alyea, F.N., Cunnold, D.M., Prinn, R.G.: Atmospheric Environment 12, 1009 (1978) 46. Pierotti, D., Rasmussen, L.E., Rasmussen, R.A.: Geophysical Research Lett. 5, 1001 (1978) 47. Jesson, J.P., Meakin, P., Glasgow, L.C.: Atmospheric Environment 11,499 (1977) 48. Meakin, P., Gumerman, P.S., Glasgow, L.C., Jesson, J.P.: Atmospheric Environment 12, 1271 (1978) 49. Glasgow, L.C., Gumerman, P.S., Meakin, P., Jesson, J.P., Atmospheric Environment 12, 2159 (1978) 50. Cunnold, D., Alyea, F., Prinn, R.: J. Geophys. Res. 83, 5493 (1978) 51. National Academy of Sciences: Halocarbons: Effects on Stratospheric Ozone. Washington, D.C., Sept. 1976 148 J. Russow 52. Hili, W.J., Sheldon, P.N.: Geophysical Research Lett. 2, 541 (1975) 53. Hill, W.J., Sheldon, P.N.: Geophysical Research Lett. 4, 21 (1977) 54. Parzen, E., Pagano, M., Newton, H.J.: Report to MCA: Statistical Time Series Analysis of Worldwide Total Ozone for Trends. March 1977 55. Tiede, J.J., Sheldon, P.N., Hili, W.J.: Preprint of a Paper at the WMO-Symposium June 26--30, 1978, Toronto 56. Angell, J.K., Korshover, J.: Paper at the WMO-Symposi\Jm June 26--30, 1978, Toronto 57. Paulet, G., Lanoe, J., Thos, A., Toulouse, P., Dasonville, J.: Toxicol. Appl. Pharmacol. 34, 204 (1975) 58. Paulet, G.: Europ. J. Toxicology. Supplement 9, 385 (1976) 59. Morgan, A., Black, A., Walsh, M., et al.: Int. J. Appl. Radiat. lsot. 23, 285 (1972) 60. Williams, F.W., Draffan, G.H., Dollery, C.T.: Thorax 29, 22 (1974) 61. Cox, P.J., King, L.J., Parke, D.V.: Biochem. J. 130, Proc. Biochem. Soc. 13 p, 87 p. (1972) 62. Blake, D.A., Mergner, G.W.: Toxikol. Appl. Pharmakol. 30, 396 (1974) 63. Seiler, W.: personal communication 64. Gulden, W.: Aerosol Rep.l2, 248 (1973) 65. Scholz, J.: in: Fortschr. biolog. Aerosol-Forschg. 1957--61. Nückel, H. (ed.), Schattauer Verlag, Stuttgart 1961 66. Leuschner, F.: unpublished 67. Hodge, H.C., Sterner, J.H., cit. by Spector, W.S.: Handbook ofToxikology Vol. I, Acute Toxicities ofSolids, Liquidsand Gases to Laboratory Animals, W.B. Saunders Comp. 1958 68. Borucki et al.: Amer. lnst. of Aeronautics and Astronautics J. 14, 1738 (1976) 69. Kranz, R.: Naturwissenschaften 53, 593 (1966) Chlorinated Paraffins V. Zitko Fisheries and Environrnental Sciences, Fisheries and Oceans, Biological Station St. Andrews, N. B. (EOG 2XO), Canada Chlorinated parartins discussed in this chapter are compounds obtained by chlorination of C 10-C30 parartins to a chlorine content between 10 and 70%. The most frequent types of chlorinated parartins are based on C 12, C15 , and C24 feedstocks and are chlorinated to 40-70% chlorine. Depending on the chlorine content, chlorinated parartins range from mobile through highly viscous liquids to solids. Production and Applications Although first uses of chlorinated parartins were reported during World War I, a large-scale commercial production started only around 1930. Production figures for this period are notavailable. During World Warll, the U.S. annual production was about 23 x 106 kg. After a decrease during the late 1940's, the annual production ofthe 1950's was about 18 x 106 kg, and increased to about 21 x 106 kg through the 1960's. Between 1970 and 1975, the production increased to about 45 x 106 kgfyr and is currently an estimated (50-70) x 106 kgfyr. The world production of chlorinated parartins is probably 3-4 times the U.S. production. According to Mills [18], the 1978 annual production of chlorinated parartins was 105 x 106 kg in western Europe, 60 x 106 kg in N orth America, and 65 x 106 kg in other free-world countries. Chlorinated parartins are manufactured by a number of companies and marketed under a variety of tradenames. Some of these are given in Table 1. Additional manufacturers are listed in Table 2. It is understood that chlorinated parartins arealso produced in China, Czechoslovakia, East Germany, Poland, Romania, and the USSR [18]. In addition to tradenames, chlorinated parartin preparations are further characterized by numbers. Theseare often related to the chlorine content of V. Zitko 150 Table 1. A partial Iist of tradenames and manufacturers of chlorinated paraffins Tradename Manufacturer Arubren CP Cereclor Farbenfabriken Bayer (West Germany) ICI Limited (UK, France,Jtaly, Spain, Australia, USA, Canada) Dover (USA) Dover(USA) Diamond Shamrock (USA) Farbwerke Hoechst (West Germany) Chemische Werke Hüls (West Germany) Hercules (USA) Pear8all (USA) Keil (USA) Pearsall (USA) Keil (USA) Dover (USA) Neville (USA) Deutsche Dynamit Nobel (West Germany) Chlorez Chloroflo Chlorowax Chloroparaffme Hoechst Chloroparaffme Hüls Clorafm CPF CW FLX Kloro Paraoll Unichlor Witaclor Table 2. Additional manufacturers of chlorinated paraffins [18] AECI (South Africa) Ajinomoto (Japan) Asahi Denka (Japan) Barm Quimica (Brazil) Caffaro (Italy) Ciclomeros (Mexico) Electrochlor (Argentina) Excel Ind. (India) FPQ (Brazil) Kop (South Africa) Meltur Chem. (India) Plasticlor (Mexico) Rhone Pulenc (France) Rio Rodano (Spain) Sintesis Quimica (Argentina) Toyo Soda (Japan) Ugimica (Spain) Ugine Kuhlmarm (France) Wintershall (West Germany) the preparation, but, as a rule it is necessary to consuJt the manufacturer's technical data sheets to obtain the average formula of the preparation, either directly or by calcuJation from the average molecuJar weight and the chlorine content. Properties of some Cereclor chlorinated paraffins are given in Table 3. Applications of chlorinated paraffins include plasticizers, additives to paints, adhesives, mastics and caulks, additives to lubricants, particularly cutting oils and heavy duty gear oils, and additives to printing inks. According to Howard et al. [11], lubricating oil additives, solvent and plasticizer uses, secondary vinyl plasticizers, and traffic paints represented 45, 27, 24, and 4% respectively, of the U.S. market in 1973. In 1976, the amount of chlorinated paraffins used in plastics in the U.S. (35 x 106 kg) exceeded the amount of non-halogenated phosphoric acid esters (21 x 106 kg). Chlorinated paraffins thus became the second largest group of flame retardant for plastics, exceeded only by aluminum hydroxide [12]. According to Mills [18], the major use of chlorinated paraffins in 1978 was as a plasticizer in flexible PVC (PVC cables, flooring, shoes, extrusions, etc.), 151 Chlorinated Paraffins Table 3. Properties of some typical Cereclor (ICI) chlorinated paraffins [18] Molecular Appearance weight ofthe CP Grade Chlorine content % 42 42 600 48 48 700 54 54 780 70 70 1100 S45 45 390 S52 52 440 S58 58 500 50LV 49 320 56L 56 370 60L 60 400 63L 63 430 65L 65 440 70L 70 500 Colour Density hazen at 25 °C g/mL units Viscosity at 25 °C poises Clear very pale yellow liquid Clear viscous yellow liquid Clear viscous yellow liquid White powder 250 1.16 22 300 1.24 250 450 1.32 5000 100 1.63 Clear water white mobile liquid Clear water white liquid Clear viscous pale yellow liquid Clear almost colourless liquid Clear very pale yellow liquid Clear very pale yellow liquid Clear very pale yellow liquid Clear viscous yellow liquid Clear very viscous liquid 80 1.16 Solid/ Softening point 90 oc 2 100 1.25 16 150 1.35 350 100 1.19 0.8 125 1.30 8 125 1.36 35 125 1.42 150 150 1.44 300 200 1.54 3500 and the relative use of chlorinated paraffinswas PVC (45%), lubricants (25%), paints (13%), flame retardants (10%), and others (7%). Three reviews of the manufacture, properties, applications, analytical chemistry, etc. of chlorinated paraffins are available [11, 15, 27]. Chemistry Chlorinated paraffins are produced by Iiquid-phase chlorination of paraffinic stocks at 50-150 oc, often in the presence of a solvent such as carbon tetrachloride. Both batch and continuous processes are used. In the latter, the chlorination unit may consist of 3-4 cylindrical reactors, connected in series, with chlorinated paraffins flowing counter currently to chlorine. Different stocks, ranging from the kerosine-gas oil fraction of petroleum to very pure straight chain paraffins, may be used. As discussed below, the 152 V. Zitko presence ofbranched paraffins in the feedstockdecreases the thermal stability of the final product and also causes its yellow to dark brown color. Many chlorinated paraffins produced at present are based on very pure straight chain feedstocks, are very stable thermally and practically colorless [7]. Aromatic hydrocarbons in the feedstock may result in chlorinated aromatic hydrocarbons in the product. The content of aromatics in the usual C12 feedstock (range C9-C 14) is typically about 0.5-1% and is reported by the supplier [5]. The content of aromatics in the C24 feedstock (normal range C20-C30) is less than 0.1% [5]. No information is available on the content of aromatics in the C 15 feedstock. It is probably somewhat below that of the C12 feedstock. European feedstocks are treated to reduce aromatics to 50-100 J.lg/g [18]. The chlorination is a radical reaction and the reactivity ofhydrogen atoms decreases with their increasing acidity. At 300 oc in the gas phase, the relative chlorination rates on tertiary, secondary, and primary carbon atoms are approximately 4:2:1 [6]. These reactivities and statistical factors determine the distribution of chlorirre atoms along the chain of the paraffin [9, 20]. As a result, chlorinated paraffins are extremely complex mixtures of chloroparaffins with different chain lengths, degree of chlorination, and distribution of chlorirre atoms along the chain. Consequently, it is difficult to elucidate experimentally their structural details. It is interesting to note that a hexakontane, containing 33% chlorine, still contains 1.8% of monochloro- and 4% of dichlorohexakontane [13]. The stability of chlorirre in chlorinated paraffins is inversely related to the ease of chlorination. Thus the stability decreases in the order primary > secondary > alicyclic > benzylic > allylic > tertiary chlorine. Since the presence of other than primary and secondary chlorirres is due to impurities, chlorinated paraffins of high thermal stability require very pure straight chain paraffins as feedstocks. The loss of chlorirre from chlorinated paraffins occurs primarily by dehydrochlorination which may startat temperatures above 250 oc [8]. Dehydrochlorination followed by a complete volatilization of products was observed on heating of C22 chlorinated paraffins containing up to 25% chlorine. U nder similar conditions, appreciable residues were formed from chlorinated paraffins containing 31 and 38% chlorine. The mechanism of color formation and the degradation products have not been identified [21]. Ions C4H 3, C5H 5, C6H6, C7H7, CsH4Cl, C3H3Clz, C9H7, C 7H 6Cl, and C5H 6Cl2 were observed on direct pyrolysis of chlorinated paraffin (Cereclor 70)- (Bi0)2C0 3or Sb20 3 mixtures in the ion source of a mass spectrometer at 260-350 oc [2]. Binding ofhydrochloric acid, formed by dehydrochlorination, retards the process and stabilizers such as epoxides and organotin compounds are usually added to chlorinated paraffin preparations. Traces of iron also promote dehydrochlorination and complexing agents such as EDT A and NT A may be used as stabilizers. The preparation of chlorinated paraffins containing butoxy, hydroxy, nitrile, quinolinyl or dialkylphosphoryl groups has been described. An acetoxylated chlorinated paraffin preparation had a slightly better plasticizing Chlorinated Paraffins 153 erticiency than chlorinated parartin [3]. The modified chlorinated parartin preparations do not appear to be widely used at the moment. Determination The determination of chlorine appears to be the only technique for the quantitation of trace Ievels of chlorinated parartins. Since many other chlorine-containing compounds, both natural and anthropogenic, may be encountered, several cleanup procedures have been used. The situation is somewhat less complex for technical products, containing chlorinated parartins as additives. The Ievels of chlorinated parartins are much higher than those likely present in environmental samples. Aseparation of chlorinated parartins from polymers, based on gel permeation chromatography, has been described recently [17]. Extraction of Biological Samples. Solventmixtures such as acetone-hexane [1] or cyclohexane-isopropanol [23] have been used. Routine extraction conditions, as used for the extraction of PCB's and common organochlorine pesticides, appear tobe applicable as well [16, 25]. Cleanup of Extracts. Solvent partitioning, column and thin-layer chromatography are the basic steps used in the cleanup. Partitioning between hexane and dimethyl formamide [1], or between hexane and acetonitrile [16] is the usual preliminary step for the separation of chlorinated parartins from the bulk of co-extracted Iipids. The pair hexane-dimethyl sulfoxide may be possibly used toseparate highly chlorinated (70% chlorine) from less chlorinated (40% chlorine) preparations [28]. Alumina, silica, and Florisil are the common adsorbents for the column Chromatographie cleanup and separation of chlorinated parartins not only from Iipids but also from other organochlorine compounds. These procedures may be used in combination with solvent partitioning or independently. On elution with hexane, chlorinated parartins are on1y partially eluted from alumina, but the elution becomes quantitative if the Iipid Ioad on the column exceeds 10 mg/g alumina [24]. Chlorinated parartins are eluted from silica by 10% ether in hexane [24] or by benzene, carbon tetrachloride or carbon disulfide [1], and from Florisil by 6% ether in hexane [16]. The chromatography on silica thus separates chlorinated parartins from organochlorine compounds such as PCB's, DDE, and hexachlorobenzene. Since these experiments have been carried out with C24 chlorinated parartins, there is some uncertainty about the behaviour of the C 12 preparations under these chromatographic conditions. Friedman et al. [10] used UV irradiation of the 6% ether fraction from Florisil to eliminate the interference by PCB's, DDE, and many other organochlorine pesticides in the subsequent quantitation. Svanberg et al. [23] cleaned up extracts by treatment with concentrated sulfuric acid. Quantitation. C24 chlorinated parartins are too non-volatile tobe subjected to gas chromatography without decomposition and have been quantitated !54 V. Zitko either by direct microcoulometry of cleaned-up extracts [24] or by sophisticated thin-layer chromatography, including forward andreversesolvent development of the plates, heat transfer from silica to alumina, and detection by silver nitrate [1]. Gas chromatography with microcoulometric [16] or mass spectrometric [14] detection can be used to quantitate C12 chlorinated paraffins. OV-1 or OV-101 columns, programmed from about 100 to 300 oc were used. Separation of chlorinated paraffins into distinct peaks was not accomplished. Chlorinated paraffins were eluted as a broad envelope with some indication of individual peaks. The data of Lahaniatis et al. [14] indicate that the retention time increases both with increasing carbon chain length and with increasing degree of chlorination. The components identified in a Witaclor 50 preparation ranged from C12H 24Cl2 to C16Hz1Cl7. Svanberg et al. [23] quantitated chlorinated paraffins by determining chlorine by neutron activation analysis. Confirmation. Reductive dechlorination to the parent paraffinic stock can be used to confirm the presence of chlorinated paraffins. Zitko [26] and Panzel and Ballschmitter [19] used sodium bis (2-methoxyethoxy) aluminum hydride; Lahaniatis et al. [14] used sodium in ammonia. Sosa [22] described the application of infrared spectrophotometry to detect potentially toxic chlorinated kerosine and gas-oil fractions in chlorinated paraffins. These fractions may be used to decrease the viscosity of C24 chlorinated paraffins, and may be present in amounts up to 15%. Smaller additions may not be detectable by this technique. The possibility of such blending in the formulation of chlorinated paraffin preparations compounds the problems in the determination of these materials. Solvent partitioning may be used to detect differences between chlorinated paraffin preparations due to either different stocks ofthe samenominal carbon chain length or to blending [27]. Chlorinated Paraffins in the Environment Because of the lack of selective and sensitive analytical techniques for chlorinated paraffins, very littleis known about their occurrence and fate in the environment. Many of the applications of chlorinated paraffins (additives to plastics, oils, and paints) are similar to the past "open-ended" applications of PCB's and are potentially likely to lead to environmental contamination. On the other hand, chlorinated paraffins are much less stable thermally than PCB's and, in all probability, arenot released unchanged on incineration of solid waste, although some chlorinated fragments oflow molecular weight might be released. No data on the solubility of chlorinated paraffins in waterare available. lt can be expected that the solubility decreases with increasing carbon chain length and thatitis generally extremely low (1 Jlg/L or less). Consequently, the 155 Ch1orinated Paraffins tendency of chlorinated paraffins to be adsorbed on suspended solids in the aquatic environment should be very high. Many compounds with these properties tend to accumulate in aquatic fauna. Although the data [25] are somewhat inconclusive, they certainly indicate that C 24 chlorinated paraffins with 40 and 70% chlorine are not accumulated, or accumulated much less than PCB's by fish exposed to contaminated solids or food. On the other hand, C 12 chlorinated paraffins with 50% chlorine were detected in fish following administration of contaminated food [16]. The accumulation coefficient was about 0.1. Chlorinated paraffins (Chloroparaffin Hüls 70C, C 1z, 70% chlorine) were also taken up by fish (bleaks, Alburnus alburnus) from water [23]. The accumulation coefficient, based on nominal concentration in water, was about 460. Accumulation coefficients and excretion half-lives of chlorinated paraffins, estimated from the data of Bengtsson et al. [4], are given in Table 4. Tab1e 4. Accumu1ation coefficients and excretion half-1ives of Witaclor ch1orinated paraffins in b1eaks (Albumus alburnus). Ca1cu1ated from the data ofBengtsson et al. [4] Formu1ation Carbon chain 1ength % ch1orine Accumu1ation coefficienta Ha1f-life (days) Witaclor 149 Witaclor 159 Witaclor 171P Witaclor 350 Witaclor 549 10-13 10-13 10-13 49 59 71 50 49 770 740 140 40 10 34 7 30 7 a 14 14-17 18-26 13 days exposure The data are insufficient to place much reliability on the estimated excretion half-lives. The accumulation coefficients indicate that the accumulation decreases with increasing carbon chain 1ength and chlorine content. The laboratory data are in agreement with the environmental survey for ch1orinated paraffins presented by Ba1dwin and Bennett [1]. Out of 52 samp1es, representing eggs of 4 species of aquatic birds, 6 species of fish, and 2 species of shellfish, C24 ch1orinated paraffins were detected in on1y one samp1e at about 0.06J.!g/g, close to the detection 1imit ofthe method. c12 ch1orinated paraffins were detected in 13 samples at levels around 0.05 Jlg/g. Little is known about the persistence of chlorinated paraffins in the environment. Biodegradation data quoted by Howard et al. [11] indicate a limited biodegradation of chlorinated paraffins by an acclimated sewage seed, but the data are quite inconclusive. The levels ofC24 chlorinated paraffins with 40 and 70% chlorine decreased in marine sediments to about 20% ofthe initial concentration after 28 days under anaerobic conditions, and the decrease was somewhat less pronounced under aerobic conditions [27]. The degradation products have not been identified and products containing polar groups would not have been extracted. Consequently, the degree of degradation might have been less than the data would indicate. 156 V. Zitko From the available information it appears that chlorinated paraffins have a lower potential than PCB's for contamination of the environment. However, the data are limited and a comprehensive assessment is impossible at the moment. References 1. Baldwin, M.K., Bennett, D.: Analysis ofBiological Sampies for Chlorinated Straight-Chain Paraffins. Group Research Report TLGR.0058.74. Tunstall Laboratory 1974 2. Ballistreri, A., Foti, S., Montaudo, G., Pappalardo, S., Scamporrino, E.: Thermal Decomposition Products from Mixtures ofChlorinated Paraffin with Sb20 3 and (Bi0hC03• Chem. Ind. (Milan), 60, 501 (1978) 3. Bellorin, C., Sosa, J. M.: J. Appl. Polym. Sei. 22, 851 (1978) 4. Bengtsson, B.-E., Svanberg, 0., Linden, E., Lunde, G., Bauman Ofstad, E.: Ambio 8, 121 (1979) 5. Borror, J .A.: Personal Communication, Diamond Shamrock Organics Research & Development, Electro Chemieals Division, 1979 6. Bratolyubov, A.S.: Chem. Revs. (Russian) 30,602 (1961) 7. Caesar, H.J.: Chem. Ind. (London) 1978, 615 8. Camino, G.C.O., Costa, L., Guaita, M.: J. Calorim. Anal. Therm. [Prepr.]9A, B8, 59 (1978) 9. Frensdorf, H.K., Ekiner, 0.: J. Polym. Sei. 42, 1157 (1967) 10. Friedman, D., Lombardo, P.: J. Assoc, Offic. Anal. Chem. 58, 703 (1975) 11. Howard, P.H., Santodonato, J., Saxena, J.: Investigation of Selected Potential Environmental Contaminants: Chlorinated Paraffins. EPA-560/2-75-007, Office ofToxic Substances, U. S. Environmental Protection Agency, Washington, D.C. 20460, 1975 12. lkeda, K.: Chem. Economy Engng. Rev. 9, 31 (1977) 13. Koennecke, H.G., Hahn, P.: J. prakt. Chem. 16, 37 (1972) 14. Lahaniatis, E.S., Parlar, H., Klein, W., Korte, F.: Chemosphere 1975, 83 15. Linden, E., Svanberg, 0.: Chlorinated Paraffins in the Environment. A Literature Survey, Statens Naturvärdsverk, SNV PM 1035, NBL Rapp. 62, Nyköping 1978 16. Lombardo, P., Dennison, J.L., Johnson, W.W.: J. Assoc. Offic. Anal. Chem. 58,707 (1975) 17. Migliori, F .: The Application of Gel Permeation Chromatography to the Analysis of Binders in Paints Used for Road Marking. Rapp. Rech. LPC No. 77, Min. L'Equip. L'Amenag. Territoire, Laboratoire Central Des Ponts et Chaussees, Paris 1978 18. Mills, J.F.D.: Personal Communication, ICI Mond Division (1979) 19. Panzel, H., Ballschmitter, K.: Fresenius Z. Anal. Chem. 271, 182 (1974) 20. Saito, T., Matsumura, Y.: Polymer J. 4, 124 (1973) 21. Sosa, J.M.: J. Polym. Sei., Polym. Chem. Edit. 13, 2397 (1975) 22. Sosa, J.M.: Brit. Polym. J. 7, 161 (1975) 23. Svanberg, 0., Bengtsson, B.-E., Linden, E., Lunde, G., Baumann, E.: Ambio 7, 64 (1978) 24. Zitko, V.: J. Chromatogr. 81, 152 (1973) 25. Zitko, V.: Bull. Environ. Contam. Toxicol. 12,406 (1974) 26. Zitko, V.: J. Assoc. Offic. Anal. Chem. 57, 1253 (1974) 27. Zitko, V., Arsenault, E.: Chlorinated Paraffins: Properties, Uses, and Pollution Potential. Environment Canada Fisheries and Marine Service Technical Rep. 491, St. Andrews, N. B. (1974) 28. Zitko, V., Arsenault, E.: Adv. Environm. Sc. Technol., I.H. Suffet (Ed.), Vol. 8.2, Wiley Interscience, New York 1977, p. 409 Chloroaromatic Compounds Containing Oxygen Phenols, Diphenyl Ethers, Dibenzo-p-dioxins and Dibenzofurans C. Rappe Department ofOrganic Chemistry, University ofUmeä S-901 87 Umeä, Sweden Chlorophenols A monograph covering the chemistry, pharmacology and environmental toxicology of pentachlorophenol has recently been published [1]. Production, Use, Contaminants Chlorinated phenols are manufactured in large amounts and the most widely used methods by which they are prepared, are direct chlorination and alkaline hydrolysis of the appropriate chlorobenzene, the particular method used is depending on the isomer desired. 2,4-Di-, 2,4,6-tri-, 2,3,4,6-tetra, and pentachlorophenol (PCP) are prepared by direct chlorination, while 2,4,5-tri- and pentachlorophenol are manufactured via the hydrolysis of chlorobenzenes. hcl ~a Cl 2, 4-Dichlorophenol 2, 4, 6-Trichlorophenol Isomers manufactured by direct chlorination a~CI ClyCl Cl ClyCI Cl 2, 3, 4, 6-Tetrachlorophenol Pentachlorophenol 158 C. Rappe Cl~ Let Cl*OHCl Cl Cl Cl Cl 2, 4, 5-Trichlorophenol Pentachlorephenol Isomers manufactured by hydrolysis of chlorobenzenes Chlorinated phenols have found use in a great diversity of applications. The most important use of2,4,6-tri, 2,3,4,6-tetra-, and pentachlorophenol is as wood perservatives, in addition they are used as bactericides, insecticides and herbicides; pentachlorophenol also being used as slimicide in pulp and paper mills and for curing hides. In the United States PCP is the second heaviest in use of all pesticides [2]. 2,4-Di- and 2,4,5-trichlorophenol are mainly used as starting materials for the phenoxy herbicides 2,4-D and 2,4,5-T. The US production of PCP is about 25,000 tons [2] and the total world production of all chlorophenols is estimated to be in the range of 150,000 tons [3]. The annual consumption of chlorophenols in Canada has been estimated to 1,500 tons [4]. InSweden they arebannedas wood perservatives from January 1, 1978, previously the annual consumption was estimated to 150 tons [5]. Most chlorophenol preparations are contaminated to a greater or lesser extent with a variety of products. In addition to other chlorophenols, they contain up to 8% ofpolyhalogenated phenoxyphenols, "predioxins" [6,8]. Polychlorinated diphenyl ethers, dioxins and dibenzofurans are often present in the range of ten to thousands ofmg/kg, the Ievels and identity ofthese toxic contaminants depending on the route of synthesis [9-11] see also section "Occurrence of PCDDs and PCDFs in industrial chemicals". Lindström and Nordin [12] have recently identified 2,4,6-trichlorophenol, chloroguajacols and chlorocatecols in spent bleach Iiquors from sulphate kraft mills. Analytical Methods A variety of analytical techniques have been used to detect chlorophenols. These have included colorimetric methods, ultraviolet and infrared absorption, and paper, thin layer and gas chromatography. After suitable derivatization, gas chromatography seems to be the most sensitive, rapid and specific methods [13, 14]. Physical and Chemical Properties All chlorophenols are solids at room temperature and they all have a pungent odor. The lower chlorinated chlorophenols areinsoluble in water, ethanol, ether and acetone while the higher chlorinated ones are soluble in ethanol, ether and acetone. The volatility of the compounds generally decreases and Chloroaromatic Compounds Containing Oxygen 159 the melting and boiling point generally increase as the number of chlorine atoms substituted on the benzene ring increases. Transport Behaviour Mass balances and flow diagrams of chlorophenol transport through the environmentarenot available because ofthe generallack ofmonitoring data. Their moderate volatility would suggest that atmospheric transport may be a significant route. However, in general they are considered as water and soil contaminants. Chemical and Photochemical Reactions The chlorophenols are weak acids and they will form ethers, esters and salts due to the phenol function. The lower chlorinated isomers easily undergo substitution reactions such as halogenation, nitration, alkylation and acylation. When a dilute aqueous solution of pentachlorophenol was irradiated with sunlight or UV -light the main reaction was dechlorination to lower chlorinated phenols, but tetrachlorodihydroxyl benzenes and non-aromatic products like dichloromaleic acid could also be identified [15]. Irradiation of an aqueous solution of the sodium salt of PCP yielded octachlorodioxins [16]. Thermal degradation of salts of chlorinated phenols results in the formation of high yields of dioxins [17, 18] seealso section "Formation of PCDDs and PCDFs". Metabolism and Biodegradation The microbial metabolism ofPCP has been studied by Reiner et al. [19]. Three metabolites were isolated and identified as tetrachlorohydroquinone, tetrachlorobenzoquinone and trichlorohydroxybenzoquinone. PCP absorbed by goldfish and rainbow trout was quickly excreted into the surrounding water, mostly in a conjugated form accompanied with small amount offree PCP. The conjugate was identified as the PCP sulphate [20, 21 ]. In rats it has been shown that rapid dechlorination of PCP occurs. The dechlorination is mediated by liver microsomal enzymes and the dechlorination products formed are tetraand trichloro-p-hydroquinone [5]. Autoradiographie studies after administration of 14C PCP in rats and mice showed high levels in liver, kidney, blood and the gastrointestinal tract. These investigations also showed that PCP is fairly rapidly eliminated from the body. In the rat 90% of a single dose is eliminated within 3 days [5, 22]. Accumulation and Persistence Pierce and others studied the fate of PCP in an aquatic ecosystem after an accidental spill [23]. PCP was found to persist in the water andin fish for over six months. Sediment samples retained high concentrations of PCP through- 160 C. Rappe out the two-year period of investigation. Pentachloroanisol was found to be a major conversion product. 2,4,6-Trichlorophenol, tetra- and trichloroguaiacol were shown to bioaccumulate in liver fat offish cought in the vicinity of a pulp mill producing full bleach sulphate pulp, the levels were in the range 0.4-11.5 mg/kg fat [24]. Dougherty and Piotrowska found an average of20 Jlg PCP jkg urine in 60 university students in USA [25]. An human monitaring program, also in USA, has detected a range of 1-193 J.tg/kg ofPCPin the urine of 34 occupationally nonexposed individuals [26]. Similar values was found for PCP in plasmaandin fat [27, 28]. Allthese studies suggest that contamination of human population with PCP at a level of 10-20 Jlg/kg is quite general in the USA. Biological Effects Chlorinated phenols are acute toxic for mollusks, fish and mammals, the toxicity in general increased for the higher chlorinated compounds. Most sensitive are mollusks, a concentration of 15.8 J.tg/1 caused marked reduction in the numbers of individuals [29]. The LC 50 values for goldfish and fathead minnow is about 200 J.tg/1 [30]. The chlorinated phenols have an effect of uncoupling the oxidative phosphorylation. Studies on the chronic toxicity of chlorophenols must take into consideration the degree of contamination by chlorinated dimers. Fahrig et al. [31] have shown that carefully purified 2,4,6and pentachlorophenol have a weak mutagenic effect in a mammalian spot test. Halogenated Diphenyl Ethers The polychlorinated diphenyl ethers (PCDPEs) have physical and chemical properties similar to the PCBs, and they have been suggested as substitutes for PCBs as hydraulic fluids and as additives to pesticides. A 4-monochloro-4'isobutyl-diphenyl ether is now in use as dielectric fluid in capacitors. The PCDPEs also occur as impurities in commercial chlorophenols in levels as high as 1,000 ppm [9, 10]. A number of the polybrominated analogoues (Br 4 -Br 10) are now in use as flame retardants [32, 33]. For environmental purpose the most important chemical reactions for the PCDPEs is the thermal and photochemical conversion ofthese compounds to polychlorinated dibenzofurans and dioxins [34-38], see also section "Formation ofPCDDs and PCDFs". No report seems to be available on the environmental contamination with PCDPEs. Tulp et al. [39] have recently studied the metabolism ofthe PCDPEs. In the rat they found two metabolic pathways, the predominant reaction bring aromatic hydroxylation, primarily in the ortho position. Scission of the ether band was found tobe a minor metabolic pathway. In two studies PCDPEs have shown to have the tendency to bioaccumulate [40, 41]. Chloroaromatic Compounds Containing Oxygen 161 Chlorinated Dibenzo-p-dioxins and Dibenzofurans Polychlorinated dibenzo-p-dioxins (PCDDs) and dibenzofurans (PCDFs) are two series oftricyclic compounds, which exhibit similar physical and chemical properties. Some ofthese compounds have extraordinary toxic properties and were the subject of much concern. They have been involved in accidents like the Yusho accident in Japan in 1968 [42], the intoxication at horse arenas in Missouri, USA, in 1971 [43], and the accident near Seveso, Italy, in 1976 [44]. Because they are both chemically stable and lipophilic in nature, they have a potential for accumulation in food chains and therefore they present a threat for man and the environment. There is no known technical use or production of PCDDs and PCDFs, and they do not occur naturally. Environmental contamination with hazardous PCDDs and PCDFs can result from industrial and agricultural chemieals containing these toxic impurities, from the accidental formation and release of these compounds into the environment, from burning or incineration of industrial chemieals acting as precursors and by generation from suitable precursors under environmental conditions. Chemical and Physical Data The structures the numbering ofPCDDs and PCDFs are given below. :~ 9 I 8~2 7~03 4 Cly Clx 6 C\6 X PCDDs 0 4Cl y PCDFs The number of chlorine atoms in these compounds can vary between one and eight, the number of positional isomers is quite large. In all, there are 75 PCDD and 135 PCDF isomers as shown in Table 1. Table 1. Possible nurober of positional PCDD and PCDF isomers Chlorine substitution monoditritetrapentahexaheptaocta Number ofisomers PCDDs PCDFs 2 10 14 22 14 2 1 4 16 28 38 28 16 4 1 75 135 10 C. Rappe 162 A large number of the individual PCDDs and PCDFs have been synthesized by various methods and characterized mainly by gas chromatography/ mass spectrometry [45, 46]. As a general trend in both series, solubility and volatility decreases with the incraesing number of chlorine atoms. The most toxic and most extensively studied representative of these compounds is 2,3,7,8-tetra-CDD or TCDD. Cl~01()YI Cl~O TCDD Although TCDD is lipophilic, it is only slightly soluble in most organic solvents and very, very slightly soluble in water, see Table 2. Table 2. Solubility of 2,3, 7,8-tetrachlorobenzo-p-dioxin in various solvents at 25 oc [47] Solvent Solubility g/1 o-dichlorobenzene chlorobenzene benzene chloroforrn acetone Herbicide Orange lard oil methanol n-octanol water 1.4 0.72 0.57 0.37 0.11 0.058 0.04 0.01 0.0048 0.0000002 (0.2 ~g/k) The melting point of TCDD is 305-306 oc [48] but no boiling point has been given for this compound. The volatility must be quite low, although it can be analyzed by gas chromatography like the other PCDD and PCDF isomers. The first synthesis of TCDD was reported by Sandermannet al., by the catalytic chlorination of the unchlorinated dioxin [49]. It has also been prepared in good yields by the pyrolytic dimerization of 2,4,5-trichlorophenol salts [50], and this method has been successfully used for the preparation of a large number ofPCDD standards [17]. The main reactions for the synthesis of the PCDF Standards are Pd (II) acetate catalyzed cyclization of chlorinated diphenyl ethers, Sandmeyer reactions [51] and pyrolytic conversion of individual PCB isomers [46]. Chloroaromatic Compounds Containing Oxygen 163 Occurrence of PCDDs and PCDFs in Iudustrial Chemieals 2,4,5-T ( Phenoxy Acids). The dioxinproblern was first observed in connection with teratogenic effects found for the phenoxy acid 2,4,5-T. These effects were shown to be caused by 30 ~g/ of 2,3,7,8-tetra-CDD present in this particular sample [52]. The Ievels of2,3,7,8-tetra-CDD in drums ofHerbicide Orange placed in storage in the USAandin the Pacific before 1970 have been found to vary between 0.1 and 47 ~g/ [53]. Since Herbicide Orange was formulated as a 1:1 mixture ofthe butyl esters of2,4-D and 2,4,5-T, the Ievels of 2,3,7,8-tetra-CDD in individual 2,4,5-T preparations used in the 1960's could be as high as 100 ~g/. As a result of governmental regulations, efforts were made during the 1970's to control and to minimize the formation of 2,3,7,8-tetra-CDD, and now all producers claim that their products contain less than 0.1 ~g/ of2,3,7,8-tetra-CDD. In addition to 2,3,7,8-tetra-CDD it has also been reported that samples of Herbicide Orange as well as 2,4,5-T presently on the market may contain penta-CDDs, tetra- and penta-CDFs at similar Ievels [54]. However, the analytical technique used in this investigation (low-resolution GC/MS) does not allow the identification or quantification of the individual PCDD and PCDF isomers. In analyses using high-resolution GC/MS and MS confirmation, Rappe et al., have reported that in other samples of Herbicide Orange, as well as in European 2,4,5-T formulations from the 1960's, 2,3,7,8-tetra-CDD was the dominating compound ofthis group. Only minor amounts of other PCDDs and PCDFs could be found, primarily lower chlorinated PCDDs in samples of Herbicide Orange [55]. U p to now most of the interest from scientists, regulatory agencies and the public has been concentrated on 2,4,5-T as the only source of PCDDs and PCDFs. However, using the analytical techniques now available, it has been possible to identify other important sources of these hazardous products. Chlorophenols. Chlorophenols have been found to contain a variety of contaminants, including PCDDs and PCDFs [3]. In female rats the hepatic effects oftechnical and pure grade pentachlorophenol have been investigated. The technical product, which was heavily contaminated, produced a number ofhepatic effects that cannot be attributed to the pentachlorophenol itself, but were consistent with effects associated with PCDDs and PCDFs, present at the 100-1000 ~g/ Ievel [56]. It has been reported that several positional isomers ofPCDDs and PCDFs can be found in chlorinated phenols. The presence of these isomers, as well as the relative ratio ofPCDDs to PCDFs, seems tobe dependent on the synthetic route, by which the chlorophenol is prepared. A series of pentachlorophenols from commercial sources in Switzerland, and possibly prepared by the alkaline hydrolysis of hexachlorobenzene, was found to have a ratio for PCDFs/ PCDDs of about 1 and an almost identical pattern of hexa- and hepta-CDD isomers [1 0]. On the contrary, in commercial Scandinavian 2,4,6-tri- and 2,3,4,6-tetrachlorophenols, both prepared by chlorination of phenol, the ratio PCDFs/PCDDs was found tobe greater than 20. The toxic 1,2,3,7,8,9-hexa- 164 C. Rappe CDD, which was the major hexa-CDD in the pentachlorophenol samples, was only a minor isomer in these lower chlorinated samples. A difference was also seen in the case of the hepta-CDDs [18, 57]. Concerning the PCDFs, all products contained the same major PCDF isomers (1,2,4,6,8penta-, 1,2,3,4,6,8-, 1,2,4,6,7,8- and 1,2,4,6,8,9-hexa-, 1,2,3,4,6,7,8- and 1,2,3,4,6,8,9-hepta-CDF), but the ratiowas different between the two groups of chlorophenols. In general the suspected most toxic isomers were present only as minor components. However, in the Scandinavian 2,4,6-trichlorophenol the toxic 2,3,7,8-tetra-CDF was found at a Ievel of 0.5 mg/kg [11]. Obviously, different reaction conditions in the synthesis of these chlorophenols are the cause for the changed ratio between the amounts ofPCDDs and PCDFs, and for the different PCDD and PCDF isomers being formed. These data are of importance in evaluating the risks associated with the production and use of these products. Polychlorinated Biphenyls ( PCBs). In 1970 Vos et al. identified PCDFs as toxic impurities in European PCBs at the Jlg/g Ievel. The toxic effects of these PCBs were found to be parallel to the Ievels of PCDFs [58]. The same impurities have also been reported in American and Japanese PCBs [59]. U sing packed column GC and MS Bowes et al. found that the mostabundant PCDFs had the same retention times as 2,3,7,8-tetra- and 2,3,4,7,8-penta CDF [60]. Using high-resolution GC and MS it has been shown that commercial PCBs contain quite a complex mixture of PCDFs (up to 40 different isomers) [61]. A PCB used for two years in a heat exchangesystemwas found to have a four-fold increase in PCDFs (15-20 Jlg/g). The dominating isomer was identified as 2,3, 7,8-tetra-CDF at a Ievel of 1.25 Jlg/g [62]. Hexachlorophene. The bactericide hexachlorophene is prepared from 2,4,5-trichlorophenol, the key intermediate in the production of2,4,5-T. Due to additional purification, the Ievel of 2,3,7,8-tetra-CDD in this product is usually <0.03 Jlg/kg. However, hexachlorophene also contains about 100 mg/kg of a hexachloroxanthene, the 1,2,4,6,8,9-substituted isomer [63, 64]. PCDDs and PCDFs in Fly Ash. Recently, Olie et al. reported on the occurrence of PCDDs and PCDFs in fly ash and flue gases of municipal incinerators in the Netherlands [65]. No quantitative data were given in this report, but more recently Buser and Bosshardt made a quantification that the total amount of PCDDs and PCDFs in fly ash from a municipal incinerator in Switzerland was 0.2 Jlg/g and 0.1 Jlg/g, respectively, and the fly ash from an industrial heating facility, also in Switzerland, was found to have 0.6 Jlg/g and 0.3 Jlg/g, respectively [66]. In additional studies it has been shown that the number of individual isomers was quite large with up to 30 PCDD and over 60 PCDF isomers [67, 68]. The highly toxic PCDDs (2,3,7,8-tetra-, 1,2,3,7,8-penta-, and 1,2,3,6,7,8- and 1,2,3,7,8,9-hexa-CDD) were only minor constituents whereas the known toxic PCDFs (2,3,7,8-tetra-, 1,2,3,7,8- and 2,3,4, 7,8-penta-CDF) were major constituents. 165 Chloroaromatic Compounds Containing Oxygen Formation of PCDDs and PCDFs The photochemical dimerization of chlorophenols to PCDDs has been studied by Crosby et al. [16]. The only PCDD formed in this study was the octa-CDD. Other PCDDs can be formed by photochemical cyclization of chlorinated o-phenoxyphenols, so called predioxins [69]. These predioxins are very common impurities (1-8%) in commercial chlorophenols [6-9], but the cyclization is only a minor reaction pathway, the main reaction being the dechlorination of the predioxin [69]. Cl~- Cl ~Cl ~oN-lcJ I, 2, 3, 8-tetra-CDD Another photochemical process of potential environmental importance is dechlorination ofthe higher chlorinated PCDDs and PCDFs, octa-CDD and octa-CDF [70]. The products formed in solution photolysis of octa-CDD have now been identified. By comparison with authentic standards it was found that the main tetrachloro isomer was the 1,4,6,9-tetra-CDD; the major pentachloro compound is expected tobe the 1,2,4,6,9-isomer. The main hexa- and heptachloro compounds were the 1,2,4,6,7,9- (or 1,2,4,6,8,9-) and the 1,2,3,4,6, 7,9-isomer, respectively. The reaction scheme deduced from this data shows that the chlorine atoms are removed preferably from the lateral positions on the carbon rings. Consequently the most toxic PCDD isomers such as 2,3,7,8-tetra-CDD are not likely to be formed from the solution photolysis ofthe higher PCDDs [45]. In the case ofthe octa-CDF the photochemicalloss of chlorine seems tobe a non-specific reaction: all four possible hepta-CDFs were formed in about similar amounts [71]. The identification ofPCDDs and PCDFs in fly ash and flue gases indicates that these hazardous compounds can be formed in pyrolytic processes or by burning. Arecent report [72] referred to these compounds as possibly ubiquitous products of all combustion processes. Salts of chlorophenols have been found to undergo a pyrolytic dimerization yielding PCDDs. This reaction has been used for the preparation of a large number ofPCDD standards [45]. The main PCDDs found in the fly ash are the same as those formed in a laboratory pyrolysis of a mixture of 2,4,5-tri-, 2,3,4,6-tetra- and pentachlorophenol, the most commonly used commercial chlorophenols, suggesting these products to be precursors to PCDDs in fly ash [67]. In a recent publications Rappe et al. have studied the burning ofmaterials impregnated with various chlorophenates [18]. The burning of materials impregnated with commercial 2,3,4,6-tetrachlorophenates yielded a total of 150-1,000 jlgPCDDs/g ch1orophenate ranging from the tetra- to hepta-CDDs. The 2,3,7,8-tetra-CDD (peak 3, Fig. la) and 1,2,3,7,8-penta-CDD (peak 13) were both present as minor constituents. These two compounds are not the expected dimerization products of any ch1orophenate present in these 166 C. Rappe / 10...._ 22 16 ~3 2 } J[ _c__ ~-+1·l_ m/a 456 40 21 ' 1---- 1--- 15 / lU ....._____. - ,-~.b 30 m/a 422 18 '!7 -- m/e 388 20 1 12 14 3.._ 13 24 a 1 / 11 19 20 ,9 ~ L}..J l) -- ~ ' ~ t..p.L 'L ~ 4ll u m/a 354 76 m/• 320 1 min Fig. la-c. Mass fragmentograms showing elution of PCDDs from the burning of birch leaves of a commercial2,3,4,6-tetrachlorophenate, b purified pentachlorophenate, c purified 2,4,6-trichlorophenate. (From Rappe et al. [18]) commercial mixtures and it was suggested that they are formed in cyclization reactions ofimpurities present in the commercial formulation [18]. The mass fragmentog rams in Fig. 1b, c are from the burning of purified penta- and 2,4,6-trichlorophenate. In the case of the 2,4,6-trichlorophenate, the only tetra-CDDs were the 1,3,6,8- and 1,3,7,9-isomers (peaks 1 and 2), the expected dimerization products. In the case of the pure pentachlorophenate, the octa-CDD was the main product, but unexpectedly in addition, both hepta-CDDs, eight hexa-CDDs and several penta- and tetra-CDDs were detected and identified. Of special interest is the observation of the highly toxic 2,3,7,8-tetra-(peak 3) and 1,2,3,7,8-penta-CDD (peak 13) in these burning extracts, see Fig. 1. Although they are found tobe minor constituents only, in individual burnings both have been found at Ievels exceeding 10 mgjkg chlorophenate. In this case, the formation of the lower chlorinated PCDDs takes place in a so far unknown nonspecific dechlorination process [18]. In other studies we have found that PCBs can be converted to PCDFs under pyrolytic conditions [68, 73]. The pyrolysis of commercial PCBs (Aroclor 1254 and 1260) yielded about 30 major and more than 30 minor PCDFs (see Fig. 2). One of the main constituents was 2,3,7,8-tetra-CDF, the most toxic ofthe PCDFs (peak 44). Taking into consideration the amount ofPCB recovered after the pyrolysis, the yield of PCDFs was between 3-25% calculated on the amount of PCB decomposed. Consequently, uncontrolled burning of PCB can be an important environmental source of the hazardous PCDFs. Moreover, a comparison showed a striking similarity between the Chloroaromatic Compounds Containing Oxygen 167 17 15 14 44 47 19 40 53+54 8 7 5 1 32 Fig. 2a, b. Mass fragmentograms showing elution of PCDFs in a pyrolyzed Aroclor 1254, b pyrolyzed Aroclor 1260. (From Buser et al. [68]) Reaction I 0~1 0 2, 3, 7, 8-tetra-CDF Cl~ .:lT Reaction 2 Cl Cl Reaction 3 Cl~ Cl~ Cl 0 2, 3, 4, 7, 8-penta-CDF Cl 0 1, 3, 4, 7, 8-penta-CDF Fig. 3. Reaction routes leading to tetra- and penta-CDFs from the pyrolysis of2,4,5,2' ,4' ,5' -hexachlorobiphenyl 168 C. Rappe pattern of the PCDFs in the fly ash and those formed in the pyrolysis of the commercial PCBs [68]. Therefore it is recommended that all destruction of PCB-containing wastes using incinerators must be carefully controlled, including monitoring ofPCDFs in the exhaust. Pyrolysis of individual synthetic PCB isomers showed that the formation of PCDFs can follow several reaction pathways [68]. In Fig. 3 the reaction routes leading to tetra- and penta-CDFs from 2,2' ,4,4' ,5,5' -hexachlorobiphenyl are illustrated, The reactions involve the loss of ortho-Cl2 and ortho-HCl with and without a 2,3 chlorine shift. A fourth reaction route (loss of ortho-H2) was later found to occur with some other PCB isomers [46]. Pyrolysis ofthe commercial flame retardant Fire Master BP 6, mainly 2,2',4,4',5,5' -hexabromobiphenyl, yielded the very toxic 2,3,7,8-tetra-BDF at a yield in the percent range [73]. Consequently the use ofthis compound as a flame retardant should be discontinued. In the temperature range 300-400 oc however, the yield of the conversion of the halogenated biphenyls to the halogenated dibenzofurans seems tobe only in the ppm range [74, 75]. In addition to PCBs, polychlorinated diphenyl ethers (PCDPEs) are found tobe precursors to PCDFs. lt has been shown that the thermal conversion of the ethers into PCDFs is of the same order as that from PCBs [38, 57]. In addition to PCDFs, some ofthe PCDPEs also yielded PCDDs [38], see Fig. 4. It has recently been shown by Buser that the pyrolysis of chlorobenzenes yielded small amounts ofboth PCDFs and PCDDs. In both cases the known toxic isomers (see Biological effects) were present but not as main components [76]. The mechanisms for the synthesis ofPCDDs and PCDFs from Cl-benzenes are not yet fully understood. I, 2, 4, 6, 8, 9-hexa-CDF Cl Cllß!Cl Cllß!Cl .1T -HCI Cl Cl~ I, 2, 4, 7, 8-penta-CDF ::NOOC 2, 3, 7, 8-tetra-CDD Fig. 4. Reaction routes leading to PCDFs and PCDDs from the pyrolysis of2,4,5,2',4',5'-hexaCDPE Chloroaromatic Compounds Containing Oxygen 169 The possible formation of 2,3,7,8-tetra-CDD and other PCDDs as the result ofthermal reactions of the esters of the phenoxy acid 2,4,5-T has been the subject ofmuch controversy. However, recent investigations have shown that this formation is only of very limited importance, if it takes place at all [77-79]. Recently 2,3,7,8-TCDD and higher chlorinated dioxins have been found in ashes from refuse incinerators, fossil-fueled power plants and fireplaces, charcoal grills, cigarettes and the emission of automobil engines both gasoline and diesei fueled [72, 80, 81]. Theseobservations have been interpreted that dioxins can be generated in the combustion of any material. However, no data has been provided to show that TCDD and other dioxins and dibenzofurans are formed in thermal processes unless suitable precursors like chlorinated phenols, predioxins, diphenyl ethers, PCBs or chlorobenzenes are present, see Table 3. Table 3. PCDFs formed in thermal processes Starting material Products Yield (mg/kg) Ref. PCBs Cl-phenols PCDPEs Cl-benzenes Fly ash Dow PCDFs PCDDs PCDDs + PCDFs PCDDs + PCDFs PCDDs + PCDFs PCDDs 10,000- 250,000 1,000- 100,000 10,000 + > 10,000 2,000 + 10,000 0.1-1 + 0.1-1 < 0.001 [73] [17, 18] [38] [76] [66] [72] The photochemical cyclization ofPCDPEs to PCDFs has been studied by Norström et al. [34, 35] and by Choudhry et al. [36, 37]. A 20% yield of 2,8-di-CDF has been reported from the pyrolysis of 2,4,4' -tri-CDPE [34]. - hv Cl~ HCI 2, 4, 4' -tri-CDPE 2, 8-di-CDF Analytical Methods Due to the extreme toxicity ofsome ofthe PCDDs and PCDFs, very sensitive and highly specific analytical techniques are required. Detection Ievels in environmental and biological samples should be orders of magnitude below the usual detection limits obtained in pesticide analysis. Any analysis at such low levels is complicated by the presence of a multitude of other, possibly 170 C. Rappe interfering compounds. The best available separation techniques followed by highly specific detection means have tobe used for an accurate determination of these hazardous compounds. Differentisomers of the PCDD or PCDF may vary significantly in their toxicological properties and therefore their Separation and identification becomes important. In recent years, many analytical methods were developed for the analysis of PCDDs, PCDFs and especially 2,3,7,8-tetra-CDD in environmental and industrial samples, the most specific methods making use of mass spectrometry [82]. Prerequisites for best analyses are efficient extraction and sample purification followed by good separation, ultra-sensitive detection and very desirably- confirmation. A technique for analyzing individual PCDDs and PCDFs has recently been described and discussed in detail [57]. It involves one or two steps of column chromatographic clean-up followed by highresolution gaschromatography using glass capillary columns and detection and quantification using mass spectrometry. Artifacts usually disturb the analytical work at extreme low concentration levels, but the risk can be minimized by a careful inspection of complete mass spectra. F or a correct structure assignment ofthe PCDDs, a study ofthe low mass ions can be useful [45]. Positive and negative ion chemical ionization mass spectrometry (CI and NICI) has shown tobe a useful complement to the normal electron impact (EI) technique for the quantification oftrace amounts ofPCDDs and PCDFs [25, 83, 84]. Transport in the Environment Owing to analytical problems, no data are available on the transport of TCDD and other PCDDs or PCDFs in air and water. However, the accident at Seveso, Italy, clearly shows that such transport is of importance in local pollution. TCDD is practically insoluble in water (see Table 2), but because of its extreme toxicity even such low concentrations as 0.2 Jlg/1 (0.2 ppb) can be quite important. Most ofthe TCDD, ifpresent in waterways, has been found in the sediments or attached to suspended particles. The half-life ofTCDD in a lake sedimentwas found to be about 600 days [85]. The mobility of TCDD and of a dichlorodioxin in soils has been studied [86]. Both were found to be immobile in all soils and therefore would not be leached out by rainfall or irrigation, though lateral transport during surface erosion of the soil could occur. The US Air Force conducted studies in an area in north-west Florida which had been heavily sprayed with the herbicide "Agent Orange" between 1962 and 1964. This herbicide mixturewas contaminated with TCDD (see above). A 19.3 acre testgridreceived a total of 40 tons of2,4,5-T between 1962 and 1964. When 6-inch core soil samples were taken in 1974, they showed TCDD concentrations ranging from 10 to 710 ppt. This study illustrates that significant levels ofTCDD residues remained 10 years after the last herbicide application [87]. Ch1oroaromatic Compounds Containing Oxygen 171 Chemical and Photochemical Reactions PCDDs and PCDFs are considered tobe stable compounds but due to the extreme toxicity of some of the isomers, their chemistry has not been fully evaluated. However, TCDD and other PCDDs undergoes Substitution reactions [63, 88] and SbC15 has been found to react with PCDDs and PCDFs yielding higher chlorinated or perchlorinated isomers [89]. Thermally PCDDs and PCDFs are quite stable, and decomposition of2,3,7,8-tetra-CDD occurs only at temperatures above 750 oc [90]. Under environmental conditions TCDD, like other PCDDs and PCDFs is not likely tobe degraded at a significant rate by hydrolytic reactions. Earlier work indicated that pure TCDD was rather stable to photochemical degradation. Crosby et al. found the photochemical dechlorination ofTCDD tobe extremely slow on the soil surface [9], and Y oung et al. found the half-time of TCDD tobe about half a year for soil [87]. Pure crystalline TCDD was stable to sunlight wavelengths when applied as thin films to glass or leaves or suspended in water [70, 91, 92]. However, several reports on rapid photochemical degradation ofPCDDs and PCDFs under laboratory conditions make the situation more complicated. Rapiddegradation of2,7-di, 2,3,7,8-tetra-, and octa-CDD in solutions ofmethanol was shown with higher decomposition rates for the lower chlorinated species. Other experiments using 2,4,5-Tester formulations with known amounts of TCDD and exposed to natural sunlight on leaves, soil and glass plates showed that most ofthe TCDD was lost during a single day [91, 93]. In these two experiments a hydrogen donor, such as methanol or 2,4,5-T ester, highly enhanced the photochemical dechlorination [95]. It can be mentioned here that at Seveso the TCDD was released together with salts of2,4,5-trichlorophenol, ethylene glycol and inorganic constituents [96], like water most of these are no potent hydrogen donors. According to Bertoni et al., the addition of a solution of ethyl oleate in xylene enhances the breakdown ofTCDD in soil by UV -light, more than 90% was degraded during 7-days exposure [97]. Similarly a cationic surfactant, 1-hexadecylpyridinium chloride was also reported to enhance photodecomposition [98]. Another experiments have shown that TCDD adsorbed on silica gel undergoes rapid photochemical degradation [99,100]. These experiments might be a good model for TCDD bound to dust particles during air transport. Hutzinger et al. showed rapid dechlorination for di- and octa-CDF in solution and the formation of a series of lower chlorinated PCDFs was observed [101]. Similar results were obtained by Buser [71]. Metabolism and Biodegradation Contrary to the chemical and physical effects, there is a pronounced difference in the biological effects between the different PCDD and PCDF isomers. The 172 C. Rappe metabolic behaviour and biodegradation seem tobe quite different for TCDD and for some ofthe other PCDDs and PCDFs. Metabolism ofTCDD. No metabolites ofTCDD have been identified so far. Matsumura and Benezed [102] reported that most microorganism do not degrade TCDD. Of a total of 100 microbial strains with the abilitytodegrade persistent pesticides, only 5 strains showed .some ability to degrade this compound. Ward and Matsumura [85] have studied the fate ofTCDD using aquatic sediment and lakewaterunder laboratory conditions. Evaporation is suggested to be the major mode of disappearance with metabolism playing only a minor role. The metabolic activities were enhanced under conditions which simulated microbial growth in the presence of sediment, and the unidentified metabolites were found tobe released from the sediment to the ambient water. It has recently been reported by Guenthner et al. [103] that TCDD can be metabolized by the mouse liver cytochrome P-450 system to reactive intermediates, which easi1y bind covalently to cellular proteins. It is suggested that this extreme reactivity inhibits the formation of normal metabolites like phenols, dihydrodiols, or conjugated products. Following a singleoral dose of14C-TCDD in rats, Rose et al. [104] were able to detect 14C activity only in feces and not in urin. The half-life of 14C activity in the body was about 31 days and the major part ofthe TCDD was stored in liver and fat. After repeated oral doses the major route of excretion was again found to be feces, but the urin contained 3-18% ofthe total14C activity. The half-life of 14C activity in theseratswas about 24 days, and most ofTCDD was found in liver and fat. The experiments indicated that materials other than TCDD constituted a significant fraction of the 14C activity excreted in the feces, but no metabolite was identified [104]. Van Miller et al. [1 05] reported on the tissue distribution and excretion of 3H TCDD in monkeys and rats. A marked difference was found in the tissue distribution in the two species. In monkeys, a large percentage of the dose was located in tissues that had a high lipid content, i.e. in skin, muscle, and fat; whereas in rats these tissues had much lower levels ofTCDD. Metabolism of Other PCDDs and PCDFs. Tulp and Hutzinger have studied the rat metabolism of a series ofPCDDs. 1- and 2-Mono-, 2,3- and 2, 7-di, 1,2,4-tri-, and 1,2,3,4-tetra-CDD are metabolized to mono- and dihydroxy derivatives, whilst in the case of the two monochloro isomers, also sulphur containing metabolites are excreted. It has also been shown that the primary hydroxylation exclusively takes place in the lateral positions (2-, 3-, 7- andjor 8-positions) in the molecule. In none ofthe experiments metabolites resulting from a fission fo the C-0-C bonds were detected. No metabolites were found from octa-CDD [105]. The results are rationalized in terms that the metabolism of the PCDDs occurs mainly via 2,3-epoxides. In the octa-CDD as in 2,3,7,8-tetra-CDD these positions are blocked, consequently the reaction is less likely to take place or takes place at a highly reduced rate [105]. Chloroaromatic Compounds Containing Oxygen 173 Isomers retained: Cl~ C~l Cl Cl~ Cl~ Cl 2, 3, 7, 8- 2,3,6,8Cl Cl Cl Cl 2, 3, 4, 7, 8- Cl I, 2, 4, 7, 8- Ci~-(rYI Cl Cl Cl 1,2,3,7,8- CIMOA0Cl Cl Cl I, 2, 3, 4, 7, 8- 1, 2, 3, 6, 7, 8- Isomers excreted: Cl~OY Cl~ Cl 2,3,6,7- Cl~OYI ~0 Cl~ Cl Cl 2, 3, 4, 6, 7- ~OCl c Cl Cl Cl I, 2, 6, 7, 8- I, 2, 3, 4, 8- Cl~A0 Cl~ Cl Cl 1,2,3,4,6,7Fig. 5. PCDF isomers retained and excreted from the liver ofYusho patient. (From Rappe et al. [1061) 174 C. Rappe A similar relationship between PCDF isomers retained and apparently excreted has been observed for patients with the Yusho disease, intoxication by a rice oil contaminated with PCBs and PCDFs. The contaminated rice oil and liver samples from two ofthe patients were analyzed by Rappe et al. [106] and all the major PCDFs were identified, see Fig. 5. A comparison revealed that none of the isomers retained had two vicinal hydrogenated C-atoms in any of the two C-rings of the benzofuran system. Most of these isomers had alllateral positions chlorinated. Contrary, all the PCDF isomers apparently excreted had two vicina/hydrogenated C-atoms in at least one ofthe two rings, and these unblocked positions are involved in the metabolism by forming epoxides, see Fig. 5. Kuroki and Masuda [107] have estimated that 0.37% of 2,3,6,8-tetra, 0.006%--0.03% of2,3,7,8-tetra and 0.9% ofthe 2,3,4,7,8-penta-CDF ingested were retained in the liver of one ofthe Yusho patients when he died 44 months after the use ofthe rice oil had been discontinued. Zitko et al. have shown that in fish, 2,8-di-CDF was metabolized to a hydroxylated derivative [108]. Accumulation and Persistence When 14C 2,7-di- and 2,3,7,8-tetra-CDD was added to soil, Isensee and Jones [92] found that both oats and soya beans accumulated small quantities of the dioxins. A maximum of0.15% ofthe dioxins present in the soil was translocated to the aerial portion, but neither the grain nor the soya beans harvested showed any dioxins. Analyses quoted by Firestone [53] showed that TCDD analyses ofvegetation from Seveso, Italy, gave values up to 50 mg/kg possibly due to direct contamination. The bioaccumulation of 14C-TCDD in aquatic organisms was investigated by lsensee and Jones [109], and the accumulation ratios were 2,000-7,000 times, which is about the same as those reported for many ch1orinated hydrocarbon insecticides. Total amounts accumulated were directly related to water concentrations, and equilibrium concentrations were reached in tissues in 7-15 days. Fish and shellfish taken from areas in South Vietnam that were heavi1y exposed to Herbicide Orange during military defoliation Operations have been reported by Baughman and Meselson [110] to contain 18-810 ng TCDD/kg. Young et al. [87] reported that in two creeks, associated with a military test area in Florida, USA, which also had been heavily sprayed with Herbicide Orange, 10 yr later the silt contained up to 35 ng TCDD/kg where eroded soil entered the water. Fish from the streams showed the presence ofTCDD, the highest value reported was in the gut, 85 ng/kg. Y oung et al. [87] also reported on the analyses ofterrestal animals (rodents reptiles) from the same area also collected 10 yr after the spraying, which were found to contain 130-1,300 ng TCDD/kg. The analytica1 method(s) used in these investigation is not specified. After the accident near Seveso, Italy, over 1,100 animals were killed by direct exposure, and up to 225 J.lg TCDD/kg oftissue was found in the liver of dead rabbits from the most contaminated zone [53]. 175 Chloroaromatic Compounds Containing Oxygen Using a direct probe and NICI technique, Dougherty et al. [111] have identified 2,3,7,8-tetra-CDD and found higher chlorinated PCDDs (Cl 5-Cl8) in fish from dams in Tittabawassee River, Michigan, USA. These dams were located at Dow Chemica1 Inc. Contrary, in fish from Ohio River and Connecticut River, a series of PCDFs (C1cC1 7) was found, but no PCDDs. No identification of the discret isomers or quantification was possible using this analytical technique. The US EPA initiated a TCDD monitoring programme ofbeeffat samples taken from cattle that had grazed on rangeland known to have been treated with 2,4,5-T. Three laboratories were involved in this programme. Of 52 samples, 19 were reported by one or more laboratory to have TCDD, in the range 5-66 ngfkg, and the overallaveragewas 7 ng/kg [53, 82]. Tissues and milk from cattle possibly intoxicated by licking pentachlorophenol-treated timher has been analyzed by Hass et al. [84] for higher chlorinated PCDDs.- Hexa-, hepta-, and octa-CDD were found in the jlg-ng/kg range, the highest values found for the octa-CDD. a~o Cl~O Cl~Oß:JI Cl~O 2, 3, 7, 8-tetra-CDD I, 2, 3, 7, 8-penta-CDD :Ne»: Cl ~Cl cMoß::Jlci Cl I, 2, 3, 6 7~:8-hexaCD Cl~ CI I, 2, 3, 7, 8, 9-hexa-CDD Cl Cl~ CI CIMOMCI CIMOMCI 2, 3, 7, 8-tetra-CDF I, 2, 3, 7, 8-penta-CDF Cl~O Cl~ Cl 2, 3, 4, 7, 8-penta-CDF Fig. 6. The most toxic PCDD and PCDF isomers 176 C. Rappe Biological Effects Several books and reviews covering the toxicology and biological effects of PCDDs (mainly TCDD) and PCDFs have recently been published [42, 112 to 116]. The toxicity and biological properties ofindividual congeners is strikingly depending on number and position of chlorine substituents. The isomers with the highest acute toxicity appear to be the 2,3,7,8-tetra, 1,2,3,7,8-penta-, 1,2,3,6,7,8-, and 1,2,3,7,8,9-hexa-CDD and the 2,3,7,8-tetra, 1,2,3,7,8-, and 2,3,4,7,8-penta-CDF, see Fig. 6. Alltheseisomers have LD 50-values in the range 1-100 J.lg/kg for the most sensitive animal species [112, 113, 117, 118]. The bromine analogaus like 2,3,7,8-tetra-BDD and 2,3,7,8-tetra-BDF seem to be equally toxic as the corresponding chlorine compounds [118, 119]. Recent work has shown that the positional isomers of PCDDs and PCDFs vary highly in their acute toxicity and biological activity. A factor of 1,000-10,000 can be found for closely related isomers such as 2,3,7,8- and 1,2,3,8-tetra-CDD [120, 121]. TCDD poisoning is characterized by loss ofbody weight with delayed lethality. A large variety of sublethal effects have been identified after acute or chronic exposure to PCDDs and PCDFs, mainly TCDD [115, 117, 119]. TCDD has caused serious toxic effects in workers due to industrial exposure in 2,4,5-trichlorophenol plants resulting in irreversible liver damage, severe chlorance, hepatitis, porphyria and darnage to the nervaus system [116]. TCDD isapotent inducer of enzyme systems, particularily in the liver, it has mutagenic, teratogenic, carcinogenic andfor co-carcinogenic effects [115, 117, 119]. An increased incidence of liver cancer in people in Vietnam has been attributed to the TCDD present in Herbicide Orange sprayed in large quantities in this country [122, 123]. Exposure to PCDDs and PCDFs has been discussed in relation to malignant mesenchymal tumors and Iymphomas among workers exposed to phenoxy acids and chlorophenols in Sweden [124-126]. References 1. Ranga Rao, K. (Ed.): Pentachloropheno1; Chemistry, Pharmaco1ogy and Environrnenta1 Toxico1ogy. Plenum, New York 1978 2. Cirelli, D.P.: ibid., p. 13 3. Nilsson, C.-A. et al.: ibid., p. 313 4. Hoos, R.A.W.: ibid., p. 3 5. Ahlborg, U.G.: Metabolism of Chlorophenols; Sturlies on Dechlorination in Mammals. Swed. Environm. Prot. Board PM 895. Stockholm 1977 6. Rappe, C., Nilsson, C.-A.: J. Chromatogr. 67,247 (1972) 7. Jensen, S., Renberg, L.: Ambio 1, 62 (1972) 8. Deinzer, M. et al.: Biomed. Mass. Spectrom. 5, 566 (1978) 9. Firestone, D. et al.: J. Assoc. Offic. Anal. Chem. 55, 85 (1972) 10. Buser, H.R., Bosshardt, H.-P.: ibid. 59, 562 (1976) Chloroaromatic Compounds Containing Oxygen 177 II. Rappe, C., Garä, A., Buser, H.R.: Chemosphere 7, 981 (1978) Lindström, K., Nordin, J.: J. Chromatogr. 128, 13 (1976) Lamparski, L.L., Nestrick, T.J.: ibid.156, 143 (1978) Edgerton, T.R., Moseman, R.F.: J. Agric. Food Chem. 27, 197 (1979) Wong, A.S., Crosby, D.G.: in Ranga Rao K., (Ed.).: Pentachlorophenol; Chemistry, Pharmacology and Environmental Toxicology. Plenum, New York 1978 p. 19 16. Crosby, D.G., Wong, A.S.: Chemosphere, 5, 327 (1976) 17. Buser, H.R.: J. Chromatogr.JJ4, 95 (1975) 18. Rappe, et al.: Chemosphere 7, 269 (1978) 19. Reiner, E.A., Chu, J.P., Kirsch, E.J.: in Ranga Rao, K. (Ed.): Pentachlorophenol; Chemistry, Pharmacology and Environmental Toxicology. Plenum, New Y ork 1978, p. 67 20. Kobayashi, K.: ibid. p. 89. 21. Glickman, A.H. et al.: Toxicol. Appl. Pharm. 41,469 (1977) 22. Larsen, R.V. et al., J. Pharm. Sei. 61,2004 (1972) 23. Pierce, R.H. Jr. et al.: Bull. Environ. Contam. Toxicol. 18, 251 (1977) 24. Landner, L. et al.: ibid. 18, 663 (1977) 25. Dougherty, R.C., Piotrowska, K.: Proc. Natl. Acad. Sei. USA 73, 1977 (1976) 26. Kutz, F.W., Murphy, R.S., Strassman, S.C.: in Ranga Rao, K. (Ed.): Pentachlorophenol; Chemistry, Pharmacology and Endvironmental Toxicology. Plenum, New York 1978, p. 12. 13. 14. 15. 363 27. Pearson, J.E. et al.: Bull. Environ. Contain. Toxicol. 16, 556 (1976) 28. Shafik, T.M.: ibid., 10, 57 (1973) 29. Tagatz, M.E., Ivey, J. M., Tobia, M.: in Ranga Rao, K. (Ed.): Pentachlorophenol; Chemistry, Pharmacology and Environmental Toxicology. Plenum, New York 1978, p. 157 30. Adelman, I. R., Smith, L.L., Jr. Siesennop, G.D.: J. Fish Res. Board Can. 33, 203 (1976) 31. Fahring, R., Nilsson, C.-A., Rappe, C.: in Ranga Rao, K. (Ed.): Pentachlorophenol; Chemistry, Pharmacology and Environmental Toxicology. Plenum, New York 1978, p. 325 32. Sundström, G., Hutzinger, 0.: Chemosphere 5, 187 (1976) 33. Norström, A., Andersson, K., Rappe, C.: ibid. 5, 255 (1976) 34. Norström, A., Andersson, K., Rappe, C.: ibid. 5, 21 (1976) 35. Norström, A., Andersson, K., Rappe, C.: ibid. 6, 241 (1977) 36. Choudhry, G.G. et al.: J. Agric. Food Chem. 25, 1371 (1977) 37. Choudhry, G.G. et al.: Chemosphere 6, 327 (1977) 38. Rappe, C., Lindahl, R., Buser, H.R.: Chemosphere, in press. 39. Tulp, M. Th. M. et al.: Xenobiotica 9, 65 (1979) 40. Neely, W.B., Branson, D.R., Blau, G.E.: Environm. Sei. Techno!. 8, 1113 (1974) 41. Zitko, V., Carlson, W.G.: Chemosphere 6, 293 (1977) 42. Higuchi, K. (Ed.): PCB Poisoning and Pollution. Kodansha, Tokyo 1976 43. Carter, C.D. et al.: Science 188, 738 (1975) 44. Pocciari, F.: Ecol. Bull. (Stockholm) 27, 67 (1978) 45. Buser, H.R., Rappe, C.: Chemosphere 7, 199 (1978) 46. Buser, H.R., Rappe, C.: ibid. 8, 157 (1979) 47. Crummett, W.B., Steh!, R.H.: Environm. Health Perspect. 5, 15 (1973) 48. Pohland, A.E., Yang, G.C.: J. Agric. Food Chem. 20, 1093 (1972) 49. Sandermann, W., Stockmann, H., Carsten, R.: Chem. Ber. 90, 690 (1957) 50. Langer, H.G., Brady, T.P., Briggs, P.R.: Environ. Health Perspect. 5, 3 (1973) 51. Gan1, A. et al.: Chemosphere, in press. 52. Courtney; K.D., Moore, J.A.: Toxicol. Appl. Pharm. 20, 396 (1971) 53. Firestone, D.: Ecol. Bull. (Stockholm) 27, 39 (1978) 54. Huckins, J.N., Stalling, D.L., Smith, W.A.: J. Assoc. Offic. Anal. Chem. 61, 32 (1978) 55. Rappe, C., Buser, H.R., Bosshardt, H.P.: Chemosphere 7, 431 (1978) 56. Goldstein, J. 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Ser. 120,44 (1973) Buser, H.R.: J. Chromatogr. 129, 303 (1976) Dow Chemical Company: The Trace Chemistries ofFire, Report, Nov. 1978 Buser, H.R., Bosshardt, H.-P., Rappe, C.: Chemosphere 7, 109 (1978) Morita, M., Nakagawa, J., Rappe, C.: Bull. Environ. Contam. Toxicol. 19, 665 (1978) O'Keefe, P.W.: Environ. Health Perspect. 23, 347 (1978) Buser, H.R.: Chemosphere 8, 415 (1979) Rappe, C.: Ecol. Bull. (Stockholm) 27, 19 (1978) Steh!, R.H., Lamparski, L.L.: Science 197, 1008 (1977) Ahling, B. et al.: Chemosphere 6, 461 (1977) Rawls, R.L.: Chem. Engng. News Feb. 12, 1979, p. 23 Smith, R.J.: Science 202, 1166 (1978) McKinney, J.D.: Ecol. BuH. (Stockholm) 27, 53 (1978) Hunt, D.F., Harvey, T.M., Russell, J.W.: J.C.S. Chem. Comm. 151 (1975) Hass, J.R. et al.: Ana1yt. Chem. 50, 1474 (1978) Ward, C.T., Matsumura, F.: Arch. Environ. Contam. Toxicol. 7, 349 (1978) Helling, C.S. et al.: J. Environ. Quality 2, 171 (1973) Y oung, A.L. et al.: Fate of 2,3, 7,8,-Tetrachlorodibenzo-p-dioxin (TCDD) in the Environment: Summary and Decontamination Recommendations. USA FA-TR-76-18 Boulder, Colo, USA 1976 Grey, A.P. et al.: J. Org. Chem. 41, 2435 (1976) Buser, H.R., Rappe, C.: unpub1ished results 1978 Steh!, R.H. et al.: Adv. Chem. Ser. 120, 119 (1973) Crosby, D.G. et al.: Science 173, 748 (1971) Isensee, A.R., Jones, G.E.: J. Agric. Food Chem. 19, 1210 (1971) Crosby, D.G., Wong, A.S.: Science, 195, 1337 (1971) Crosby, D.G.: Amer. Chem. Soc. Symposium Ser. 73, I (1978) Liberti, A. et al.: 10, 97 (1978) Rappe, C.: in Cattabeni, F., Cavallaro, A., Galli, G., (Eds.): Dioxin, Toxicological and Chemical Aspects. SP Medical and Scientific Books, Jamaica 1978, Chap. 17 Bertoni, G. et al.: Analyt. Chem. 50, 732 (1978) Botre, C., Memoli, A., Alhairque, F.: Environ. Sei. Techn. 12, 335 (1978) Plimmer, J.R.: Bull. Environ. Contam. Toxicol. 20, 87 (1978) Gebefügi, I., Baumann, R., Korte, F.: Naturwissenschaften 64, 486 (1977) Hutzinger, 0. et al.: Environ. Hea1th Perspect. 5, 267 (1973) Matsumura, F., Benezet, H.J.: ibid. 5, 253 (1973) Guenthner, T.M., Fysh, J.M., Nebert, D.W.: Pharmaco1ogy, in press Rose, J.Q. et al.: Toxicol. Appl. Pharmacol. 36, 209 (1976) Tulp. M.Th.M., Hutzinger, 0.: Chemosphere 7, 761 (1978) Rappe, C. et al.: ibid. 4. 259 (1979) Kuroki, H., Masuda, Y.: ibid. 7, 771 (1978) Zitko, V. et al.: Environ. Health Perspect. 5, 187 (1973) lsensee, A.R., Jones, G.E.: Environm. Sei. Techno!. 9, 668 (1975) Baughman, R., Meselson, M.: Environm. Health Perspect. 5, 27 (1973) Dougherty, R.C. et al.: ibid. in press. IARC Monogr. Evaluation Carcinogenic Risk to Man 15, 41 (1977) IARC Interna! Techn. Rep. 78(001, IARC, Lyon 1978 Ramel, C. (Ed.): Chlorinated Phenoxy Acidsand Their Dioxins. Ecol. Bull. (Stockholm) 27, (1978) 88. 89. 90. 91. 92. 93. 94. 95. 96. 97. 98. 99. 100. 101. 102. 103. 104. 105. 106. 107. 108. 109. 110. II!. 112. 113. 114. Chloroaromatic Compounds Containing Oxygen I79 II5. Cattabeni, F., Cavallaro, A., Galli, G., (Eds.): Dioxin, Toxicological and Chemicai Aspects. SP Medicai and Scientific Books, Jamaica I978 II6. Ann. New Y ork Acad. Sei, 320, I (1979) II7. McConnell, E.E. et al.: Toxicol. Appl. Pharmacol. 44, 335 (1978) II8. Moore, J.A. et al.: Ann. New York Acad. Sei, 320, I (1979) II9. Poland, A., Glover, E., Kende, A.S.: J. Bio!. Chem. 251, 4936 (1976) I20. Poiand, A., Glover. E.: Mol. Pharmacol. 9, 736 (1973) I2I. Bradlaw, J.: FDA Washington, personal communication I22. Tung, T.T.: Rev. Medecine 14, 653 (1977) I23. Wade, N.: Science 204, 8I7 (I978) I24. Hardell, L., Sandström, A.: Läkartidningen 75, 335 (1978) I25. Hardell, L., Sandström, A.: Brit. J. Cancer 39, 7II (1979) I26. Hardell, L.: Lancet 55 (1979) Organic Dyes and Pigments E. A. Clarke, R. Anliker Ecological and Toxicological Association ofthe Dyestuffs Manufacturing Industry (ETAD) CH-4005 BaselS, Switzerland Introduction Color, which contributes so much to the beauty ofNature, is essential to the attractiveness and acceptability ofmost products used by modern society [1]. As long ago as the 25th century BC man colored his surroundings and clothes using a limited range of natural colorants ofboth animaland vegetable origin. Alizarin (18) 1 extracted as the glycoside rubierythric acid from madder, was used by the ancient Egyptians and Persians, the use of indigo (16) obtained from Indigofera datesback to 3000 BC, and Tyrian Purpie (6,6'-dibromoindigo), prepared from the sea snail Murex brandaris, has been used since the Roman era. However, the preparation in 1856 of the first synthetic dyestuff, mauveine (12), by Perkin gave birth to the development of many other important sectors of the modern chemical industry. Compared with natural dyestuffs, synthetic colorants are better able to meet the increasingly rigorous technical demands ofthe present day in terms ofstability, fastness, etc. Color can add not only aesthetic appeal, but frequently provides an almost irreplaceable safety feature (traffic lights and signs, drug identification, control systems) [2]. Dyes and pigments are substances which when applied to a substrate Iead to selective reflection or transmission ofincident daylight2 • Substauces which create the sensation of blackness or whiteness are also regarded as dyes or 1 The chemical structures of a selection of representative colorants are given in Fig. 6, preceding the references at the end of this chapter 2 Throughout this article the terms "dyestuffs" (or "dyes") and "pigments" have been used according to this definition. The term "colorants" is applied collectively when no distinction is required. Descriptive terms such as "pigment preparation", "commercial dyestuff', "technical grade colorants" are used to describe various qualities of commercial product 182 E. A. Clarke, R. Antiker pigments. Characteristic ofpigments is their extremely low solubility in water, and in the application substrate. They generally also exhibit low solubility in organic solvents. For this reason they remain essentially in the solid state during processing and when applied to the substrate. Colorants, not covered by this definition of pigments, are dyestuffs. Synthetic organic dyestuffs and pigments exhibit an extremely wide variety of physical, chemical and biological properties, making any comprehensive review ofthe ecotoxicological properties ofthe several thousand commercially available products difficult. In the following sections an attempt has been made to provide a perspective of the environmental problems posed by synthetic organic colorants, and to outline the current efforts by the manufacturing industry to ensure that its products present no unreasonable risk to the ecosystem, including man. Most commercial dyestuffs and pigments are in fact mixtures in which the content ofthe specific colored component normally lies in the range 10-98%. The other components are necessary to confer the desired physical properties and may of course influence the product's ecotoxicological behaviour. Fluorescent whitening agents, which are frequently classified as dyestuffs, are not treated in detail in this chapter. An assessment of the ecological behaviour and the pollution aspects of these products as well as their chemistry, application and their toxicological properties have been comprehensively reviewed [3]. Chemistry and Uses A detailed description of the chemical processes involved in manufacture of the wide variety of synthetic organic colorants is beyond the scope of this handbook and the authoritative texts edited by Yenkatamaran [4] should be consulted. An essential preliminary to any discussion of the chemistry of synthetic organic colorants is the classification of the numerous products which are commercially available. The Colour Index [5] lists an estimated 38,000 commercial colorants involving 7,000-8,000 different chemical structures. This unique system identifies each product by a generic name (e.g. Acid Yellow 23) which describes the application type (i.e. Acid) and color (Yellow 23) together with dyeing characteristics. In addition where the chemical structure (1) has been made known to the Colour Index, a five-digit constitution nurober is allocated which uniquely identifies the structures (in this case C.I. 19140). The specific constitution numbers are drawn from ranges which have been allocated to the various chemical types and a cross reference is given for each generic name to the known commercial brands (and vice-versa). Many ofthe allocated C.I. constitution numbers refer to products which are no Ionger used commercially. A factor of particular relevance for the ecological behaviour of a dyestuff is its ionic character and Table 1 summarises the major applications of the major dyestuff and structural classes. _ _ _ _ _ _~ Table 1. Summary of ionic character, chemical type and ap~_lictons Dyestuffs and Pigments ~ Basic Acid Mordant Direct Reactive Solubili- Sulphur sed Vat Azo Stilbene Di-, and Triphenylmethane Xanthene X X X X Acridine Quinoline Methine Azine X X X X X X Oxazine Thiazine Sulphur Anthraquinone X Pigment Solvent Mordant Vat X X X X X X X X X X X X X X X X X a "' X X X X X X X X X X X X X X X X X X X X X c c Other substrates Paper Paper Leather Paper Leather Leather Other usages Basic: Acid: Disperse: Solvent: Pigment: p PAm X X X X X p PAm c p c c PE PP, PVC CE CE PAm - p c c p c Textile Printing Leather Inks, manifold record systems Food, drugs, cosmetics, vamishes, inks, plastics, resins Plastics Wood stains, varnishes, lacquers, inks, polishes, plastics, colored smokes Paints, inks, lacquers, plastics, cosmetics, waxes, writing utensils, carbon paper ~- C CE p PAc ~ ~ X X X X X p PAc p PE PAm PAm PAc p Key: § X X X ~ p. X lndigoid Phihalocyanine Textile dyeing substrates Disperse 0 =Cellulose (cotton, linen) = Cellulose ester (acetate, triacetate) = Protein (wool, silk) = Polyacrylics PAm PE PP PVC =Polyamide =Polyester = Polypropylene = Polyvinylchloride 00 w 184 E. A. Clarke, R. Antiker Production Data Production or sales statistics are only available for a few countries, notably USA [6] and Japan, but total world dye production is now estimated tobe 600,000--700,000 tonnes (based on active substance). The breakdowns of this total amount by use, production area, and application type are estimated in Tables 2-4 respectively. Table 2. Estimated world production by use category (1978) Use Production (t) % Textile dyestuffs Paper/leather dyestuffs Organic pigments Fluorescent whitening agents and others 360,000 90,000 150,000 40,000 56 15 23 6 640,000 Table 3. Analysis of estimated world production by industrial region (1978) Production area Production (t) % W. Europe 250,000 147,000 109,000 57,000 51,000 13,000 13,000 COMECON USA China Japan India S. America and Mexico 39 23 17 9 8 2 2 Table 4. Analysis ofworld production by application type (1978) Application type Production (t) % Acid Basic Chrome Direct Disperse Fluorescent whitening agents Organic pigments Reactive Vat Others 60,000 45,000 20,000 100,000 75,000 30,000 150,000 35,000 65,000 60,000 9 7 3 16 12 5 23 6 10 9 Organic Dyes and Pigments 185 Analytical Methods A comprehensive treatment of the analytical chemistry of dyestuffs has been edited by Yenkatamaran [7] and this section seeks only to outline the importance of modern analytical techniques for the monitoring of the environment for dyestuffs or associated impurities which may result in adverse toxicological or ecological effects. Good analytical methodology is essential to most ecological and toxicological investigations and it is ironical that the remarkable advances in sensitivity of analytical techniques should have sharpened so many controversies because these advances far exceed the ability to interpret the results in terms of risk. Although no rational person believes that a "no risk" society can be achieved, such concepts as "zero tolerance" [8-10], "one-hit theory of carcinogenesis" and the "Delaney Amendment" [11, 12] still have a pervasive influence on regulatory decision-making. The major areas of application within the scope of this article are as follows: - monitoring of product quality - monitoring trade effluent quality, (both solid and liquid waste). See also p. 10-11 - monitoring work-place environment - monitoring ecological and toxicological experiments - monitoring environmentallevels of dyestuffs. Apart from the need to control effluents within specific Iimits for color, the required analytical techniques are essentially for measurement oftrace harmful impurities, particularly: (i) Metals. 3 Discharge consents for trace metals are increasingly stringent in most developed countries. Extensive analytical surveys have been conducted on the trace metal content of commercial dyestuffs [13, 14] and pigments [15], and these confirm that with the exception of the metal complex dyestuffs the metal content is low (generally ~ 100 ppm). Amongst the more important modern techniques availab1e for trace metal analysis are atomic absorption spectrophotometry [16-18], neutron activation analysis [19], polarography [20], emission spectrophotometry [20] and spark source mass spectrophotometry [21]. 3 In 1975 ETAD recommended to its member companies that their dyestuffs should meet the following trace meta! Iimits (ppm): As 50 Hg 25 Ba 100 Mn 1,000 Cd 50 Ni 250 Co 500 Pb 250 Cr 500 Sb 50 Cu 250 Se 50 Fe 2,500 Sn 250 Zn 5,000 These Iimits, which do not apply to meta! complex dyestuffs, were derived from currently known legal requirements for metals in effluents, on the basis of a 2% dyeing and a total dilution of the effluents in a ratio of 2500: I in relation to the dyestuff used E. A. Clarke, R. Antiker 186 (ii) Lipophilic aromatic amines. Some aromatic amines are recognized human carcinogens and several have shown carcinogenic activity in animal feeding studies. Some of these products are subject to speciallegislative controls and sensitive analytical methods have been developed to determine the presence of trace amounts in: a) Working environments. Typically the method involves absorption from sampled air followed by gas-liquid chromatography [22]. b) Dilute aqueous solution. As applied to aqueous trade effiuent and urine. Typically the method involves an extractionjenrichment stage from alkaline solution, followed by analysis using gas-liquid chromatography, thin-layer chromatography or high performance liquid chromatography [23-28]. (iii) PCBs. The wide environmental distribution ofpolychlorobiphenyl compounds (PCBs) has generated considerable interest in these persistent products [29]. Although their major usage is as thermally stable dielectric fluids in transformers they are also adventitious impurities in certain commercially important diarylide and phthalocyanine pigments. Although these trace amounts present no significant environmental burden, the development of several new analytical methods was necessary in response to regulatory pressure in the USA to meet 50 ppm PCB limit [30, 31]. a Ecological Aspects Environmental Assessment of Colorants Aceurate data on the quantity of colorants discharged to the environment from the manufacturing and processing operations arenot available. Percentage losses vary from product to product and are dependent on the processes and equipment employed. Table 5 gives a breakdown of estimated losses. Table 5. Estimated Iosses of synthetic organic colorants from manufacturing and processing operations (1978) %Losses in Textile dyestuffs Paper/leather Organic pigments Others Production (t) Production Processing Totalloss (t) 360,000 90,000 150,000 40,000 2 2 1 2 10 6 1-2 10 43,000 7,000 4,000 5,000 In most industrialized countfies only about 20% or less of process losses will reach open waters due to effective adsorption in the primary and the biological treatment stages (see Effiuent Treatment Processes). Possible envi- 187 Organic Dyes and Pigments ronmental darnage does not depend solely on the quantity released, but also on the ecotoxicological properties of the individual products involved, and their environmental transport characteristics. An adequate reassurance that the environmental release of colorants, either individually or collectively, does not lead to any significant disturbance of the ecosystem, particularly man, is required. Because of the multitude of colorants available, most satisfying specific technical requirements, it is simply not practical to undertake exhaustive evaluation of each individual product. A more effective approach is to concentrate available resources on (i) reduction of release to the environment (ii) environmental monitaring to detect any localized high concentrations and possible associated adverse effects (iii) characterisation of certain ecological properties of selected members of important chemical classes, from which the likely environmental behaviour may be predicted (iv) carrying out limited screening tests on individual products as a basis for identifying possible problern products for moreintensive evaluation. ORGANIC COLORANT CHEMICAL DEGRADATION PHOTODEGRADATION ADSORPTION Biomass, Adsorbents, Sediments, Soil, Living Matter, Finished Products. BIODEGRADATION Primary or Functional, Acceptable, Ultimate. ~ DISPOSAL Incineration , Orderly Deposit , Landfill, Soil Conditioner, etc. MINERALIZATION ~ .co2,HP NO], so4-. er: etc. Fig. 1. Processes of elimination and degradation of waste organic colorants 188 E. A. Clarke, R. Antiker Elimination and Degradation Cycle Figure 1 disp1ays the various processes, and their combinations, of elimination and degradation of organic colorants in the environment [32]. The term primary or functional biodegradation means biodegradation of a substance to an extent sufficient to remove some characteristic property ofthe molecule; in biological terms, this would be referred to as a biological transformation [33]. In the case of surfactant biodegradation this is usually taken as loss of surface activity, andin the case of a dyestuff it would typically be loss of color. Under environmentally acceptable biodegradation one understands the biodegradation of a substance to such an extent that environmentally undesirable properties are lost. Although this clearly requires a judgement it may be possible to apply a simple bioassay such as acute toxicity to fish to demonstrate that biodegradation has resulted in degradation products which are oflower acute toxicity. Ultimate biodegradation [33] is the breakdown of an organic compound to carbon dioxide, water, the oxides or mineral salts of other elements present or to small organic compounds which will be utilised for the synthesis of new cell material. This may involve complete mineralization, but does not necessarily do so. Bioelimination [33] is a term applied to the sum ofvarious processes (e.g. sludge adsorption, chemical oxidation, volatilization) contributihg to the removal of a compound from the aqueous phase during sewage treatment and may, or may not, include biodegradation. Effiuent Treatment Processes General Aspects Untreated effiuent from dyestuff production and dyeing plants are usually highly colored and thus particularly objectionable if discharged to open waters. Many dyestuffs are easily visible in waterat concentrations well below 1 ppm. As a result considerable attention has been focussed on polluting discharges with high color content. Extensive testing indicates that dyestuffs are generally adsorbed to the extent of 40--80% by the biomass and are thus partially eliminated in sewage treatment plants. They are not, however, biodegraded in this stage to any significant extent [32, 34, 35]. Furthermore, experience shows that either chemical or biological treatment alone is not sufficiently effective for decolorization and a combination of physical, chemical and biological processes is usually necessary to achieve adequate color removal of a mixture of dyestuffs. ADMI [36] and more recently McKay [37] have reviewed the various treatment processes applied so far for color removal from textile effiuents. In an EPA study [14, 38] the treatment ofwastewaters from selected typical dyebaths by a variety of processes has been examined. Twenty systems were selected to provide a wide cross-section of dye classes, 189 Organic Dyes and Pigments fibres and application techniques. Table 6 summarises the removal capability of the most common treatment processes used to remove color from such wastewaters. Table 6. Summary of effectiveness of effiuent treatment processes for various dyestuff classes Color removal by treatment processes Dyestuffs in dyeing wastewaters Azoic Reactive Acid Basic Disperse Vat Sulphur Direct Coagulation alum Activated carbon Biological Combination Ozone Sludge adphysicosorptions of dyestuffs chemical and biological 0 0 0 0 + + + + + + +(s) 0 0 0 0 0 0 + 0 0 0 + + + + + + + + + +(s) + + 0 + + 0(+) 0 + 0 + Color removal: 0 unsatisfactory, + good, s specially suitable,- not investigated Although in the following text only the elimination of the dyestuff entities themselves from the generally complex composition of effluent from dyestuffs manufacturing or processing plants is considered, the broader aspects of the problern are considered elsewhere [39-52]. Dyestuffs generally exhibit low acute toxicity to warm-blooded animals, fish, and sewage works bacteria and have so far not caused any serious environmental problems. Even in the case of Basic dyestuffs, which show a somewhat higher toxicity compared with other dyestuff classes in acute toxicity studies in mammals, fish, algae and activated sludge bacteria [34, 35, 53], their high degree of exhaustion in the dyebath and their strong adsorption characteristics facilitate the achievement of low concentrations in the effluent. The composition of dyestuff-containing effluents from manufacturing plants and dyeworks, normally characterised by the parameters BOD 5, COD and TOC 4 vary widely depending on the level of production, product mix, dilution procedures etc. In the case of dyestuffs manufacturing plant the effluent composition and pH is dependent on the nature of the particular manufacturing process, whereas in processing plants the effluent typically contains large quantities of surfactants, resins, textile auxiliaries etc. Within the framework of an EPA study [38] 20 different textile dye process effluents were systematically investigated. The average BOD 5 was 280 mg/1 (range 4 BOD 5: biological oxygen demand over 5 days COD: chemical oxygen demand TOC : total organic carbon E. A. Clarke, R. Anliker 190 12-1470 mg/1) and the average TOC was 276 mg/1 (range 55-1120 mg/1). In a raweffiuentfroma tannery[40] theaverage BOD 5 was 1900mgjl, andaverage COD 5200 mg/1 whilst the dyestuff concentration varied between 22-56 ppm (COD ::::;; 112 mg/1); i.e. the COD contribution of the dyestuff was less than 2%. Effiuent from a dyestuff and chemical manufacturing plant [41] had a BOD 5 of 900-1400 mg/1 and a TOC of 600-1000 mg/1. The corresponding values for domestic effiuent are 90-200 and 40-160 mg/1 respectively. In Table 7 the theoretical TOC and COD values for some selected dyestuffs are displayed. Table 7. Theoretical TOC and COD values for selected dyestutrs• Structureb Colour index narne Constitution number Theoretical TOC %C Theoretical COD mg 0/mg dyestuff 2 4 9 11 Disperse Yellow 3 Direct Yellow 12 Disperse Yellow 54 Basic Yellow 11 Basicßlue 3 Sulphur Black 1 Acid Green 25 Direct Blue 86 C.l. 11855 C.I. 24895 C.I. 47020 C.I. 48055 C.I. 51004 C.I. 53185 C.I. 61570 C.l. 74180 67 57 75 68 67 39 58 52 2.56 1.94 1.67 2.44 2.62 1.22 1.94 2.02 13 15 17 20 • Calculation based on 100% active dyestuff bSee Fig. 6 Because dyestuffs arenot readily biodegradable they make little contribution to BOD. The contribution to the measured COD and TOC of textile dyeing plant effiuents may amount to a few percent, but is unlikely to exceed 10%; typically the dyestuff concentration (as active ingredient) lies in the range 10-50 mg/1. Decolorized effiuents contain less than 1 mg/1 dyestuff and the TOC contribution of dyestufffollowing the primary and biological treatment stages is normally considerably less than 0,5 mg/1. Colored wastewaters usually contain a !arge number of individual dyestuffs and very often the identities ofthe colorants responsible for the color of the waterare not known exactly. This makes the determination of the color concentration using individual dyestuff spectra almost impossible. To overcome these difficulties a standard method was introduced by APHA (American Public Health Association). As a refinement of the existing platinumcohalt APHA standards [54], Allen et al. [55] introduced the ADMI color value which provides a measure of the color of aqueous solutions which is independent of hue and can be related to APHA values. The ADMI method requires relatively inexpensive instrumentation allowing the use of a wide Organic Dyes and Pigments 191 variety of color measuring instruments5 . Investigation of 45 commercial dyes showed that solutions measured as 50 ADMI units corresponded to commercial dyestuff concentrations varying from 0,1 mg/1 to 16,5 mg/1, depending upon the inherent tinctorial strength and the active dyestuff content of the commercial dyestuffs involved. Although these values serve well to define color of wastewaters and to monitor decolorization, the ecotoxicological assessment of dyestuffs in wastewaters requires quantification of the actual amounts of individual dyestuffs and more specific analytical methods. Tineher [56] has analyzed the distribution of individual dyestuffs in the Coosa River basin on which over 50% of the US carpet dyeing industry is centred. Physical and Chemical Treatment Processes ( Abiotic Processes) In the commonly used abiotic processes the decolorization is achieved either by removal of the intact dyestuff or by its destruction. Precipitation and flocculation procedures using lime [57], alum [38, 58], ferric chloride and ferric sulphate [59, 60], and organic agents [61] have given good results. Activated charcoals of various origins and qualities seem particularly suitable for Acid, Basic and Reactive dyestuffs [38, 62-69]. Different adsorbents such as activated carbon and organic agents may be added at the biological stage [38, 61, 70]. Silica gel, Fuller's earth, and bauxite have shown good results as adsorbents for Basic dyes [62]. Other adsorbent materials including peat [37, 71] and wood [72] have also been investigated for possible application in decolorizing wastewaters. Ion-exchange resins have been used for the elimination of anionic and cationic dyestuffs [73, 74], but, in the case of dyestuff mixtures recovery is normally not economic andin the disposal of the desorbed concentrate additional high costs may be incurred. Other processes including foam-fractionation, dynamic membrane hyperfiltration (reverse osmosis), arestill in the experimental stage [37, 75], but again the disposal of the concentrate may pose major problems. Of the chemical processes for color removal, ozonisation has achieved the greatest practical importance. However, for an extensive degradation of dyestuff it is necessary to use a large amount of ozone [76]. In order to minimize costs the dyestuffs are only partially oxidized using ozone and then further oxidized catalytically [77, 78]. The oxidation products, mostly polycarboxylic acids, are then either removed by flocculation or subjected to a biological treatment stage. Cheaper than ozonisation, but less satisfactory, is bleaching using chlorine, chlorine dioxide, or chloramine. By heavy chlorination a complete decolorization can be achieved in many cases; however, the 5 The procedure includes the following steps: 1. Measurement of sample on a suitable spectrophotometer or colorimeter 2. Calculation of C.I.E. Tristimulus Values X, Y, Z (may be inherent in instrument) 3. Conversion ofX, Y, Z to Vx, Vy, V, (from published tables) 4. Calculation of Adams-Nickerson Color Difference (DE) 5. Conversion ofDE to ADMI value 192 E. A. Clarke, R. Anliker possible formation of chlorinated compounds, which may be less acceptable in terms of toxicity or biodegradability, should not be overlooked. The y-radiation induced oxidation of dyestuffs in wastewaters is of potential interest in some cases [66, 79, 80]. For example, colored wastewaters of an anthraquinone dyestuffs manufacturing plant are totally decolorized by this means [81 ]. F or some classes of dyestuffs, particularly azo-dyestuffs, decolorization with reductive agents such as hydrosulphite is a workable proposition [82]. In this case the reduction can be reversible and generally involves no real degradation of the dyestuff molecule. A recently developed wet pressure oxidation process [83] has successfully dealt with non-biodegradable by-products from the production of dyestuff intermediates, but the applicability of this technique to dyestuffs has not yet been investigated. Electrochemical oxidation and the electrolytic reductive precipitation have not yet obtained any practical importance [37, 84]. Biological Treatment Processes Of the four most common biological treatment processes: stabilization ponqs, aerated lagoons, trickling filters and activated sludge [37], the last named is the most widely used today. With the possible exception of Basic dyestuffs, these processes have proved in most cases tobe insufficiently effective in removing dyestuffs from wastewaters. As already indicated, dyestuffs are practically not biodegraded in this stage, but may be adsorbed by the sludge to the extent of about 40-80%, or even more, depending on the individual dyestuff and treatment conditions. Complete removal or decolorization can only be achieved by combination with other treatment processes [85-90]. In many cases it is more economical to remove the residual color by a "polishing" treatmentsuch as adsorption or coagulation after the biological stage. Sludge Adsorption and Digestion In the biological treatment plant, dyestuffs are eliminated essentially only by an adsorption process on the sludge. Investigations indicate that the extent of adsorption is determined by the dyestuff structure, the pH and the composition of the wastewater. Lower pH conditions favour adsorption. In the practical concentration range of I 0-50 mg dyestuff per litre, there is an almost linear relationship between the concentration in solution and the amount adsorbed. The adsorptive capacity of activated sludge for the dyes investigated [91] was, in neutral media, in the range of 0.01-4% of dyestuff on dry weight sludge. Table 6, last column, gives some indication ofthe adsorption behaviour of the various dyestuff classes [92]. In general it is found that adsorption is favoured by hydroxyl-, nitro- and azo-groups, as weil as increase in length of the dyestuff molecule. Sulpho-groups reduce adsorption in the case of Acid dyestuffs whereas the number of sulpho-groups does not appear to influence adsorption in the case of Reactive and Direct dyes. Organic Dyes and Pigments 193 Following the biological treatment, the sludge containing adsorbed dyestuffs may be digested under anaerobic conditions. In an investigation of 42 dyestuffs by ADMI [35] only 4 dyestuffs had an inhibitory effect on the sludge bacteria when fed daily at a concentration of 150 mg/1 to anaerobic digestors. Of the 29 soluble dyestuffs studied only 4 (Acid Black 1, Basic Blue 3 (13), Acid Green 25 (17), and Acid Blue 45) showed no signs of decolorization. The rest were either completely decolorized (16 out of25) or underwent significant spectral changes. The mechanism of the reported decolorization has still to be determined. Little knowledge has been accumulated on the chemical steps of anaerobic degradation of dyestuffs. In the case of azo dyestuffs, this loss of color is almost certainly due to reductive cleavage of the azo groups (see section on biodegradation). Environmental Elimination Processes Photochemical Degradation The photochemical properties of some naturally occurring dyestuffs are fundamental to photosynthesis and vision. Similarly the properties of synthetic organic colorants have been exploited by modern technology in photography, organic videcon tubes for television, photochromic plastics, dye-sensitized photo-tendering of textiles, etc. A high degree of stability to photochemical degradation is normally required of commercial dyestuffs when applied to a textile substrate, and it is known that this stability is dependent on such factors as humidity, temperature, presence of oxygen, substrate and spectral distribution [93]. Of relevance to the environmental fate of dyestuffs, however, is the photochemical behaviour of the dyestuff in the low concentrations in aqueous solution which may occur in lakes and streams, and this has not been widely investigated. Arecentreview [94] concluded that in aquatic systems photoreduction of azo dyestuffs to hydrazines and amines is possible, but is likely to be very slow except in oxygen poor water. Porter [95] studied the stability of 36 commercial dyes to visible and ultra-violet light and reported only slow degradation under the experimental conditions chosen. N ormally such studies determine the rate of disappearance of color, but do not identify the reaction products. In the case of BasicGreen 4, some degradation products were identified and the degradation mechanism given in Fig. 2 was proposed. Heitz and Wilson [96] showed that the photodegradation of several xanthene dyes in dilute solution proceeds as a first order reaction, that the rate increases with increasing halogenation of the dyestuff, and that the degradation results in the loss of the phototoxic properties of the dyestuff to both insects and bacteria. Since even low-levels of pollution of waters by dyestuffs can be readily detected visually, due to their high tinctorial value, the increasingly stringent legal restrictions on such pollution are based, not improperly, more on 194 E. A. Clarke, R. Anliker \excitedl Lstate J I + ~ Me 2N©OH H 20/0 2 + Fig. 2. Photodegradation of Basic Green 4 aesthetic than toxicological considerations. Practical considerations Iead to the conclusion that photodegradation does not play a dominant röle in the environmental fate of dyestuffs, although its contribution to the total mineralization ofwidely dispersed trace amounts may be underestimated. Biodegradation ~ Metabolism A knowledge of the biodegradation characteristics of a chemical is of primary importance in selecting what further testing is appropriate to evaluate its possible environmental effects. This is recognized for example in the Japanese Law "Control of Chemical Substances" under which it is only necessary to proceed sequentially to bioaccumulation and to fish toxicity testing if the product is of low biodegradability [97). Dyestuffs, for satisfactory technical performance, must be resistant to change under aerobic conditions in contact with body fluids (perspiration, urine) and are, therefore, highly resistant to aerobic biodegradation. There are some advantages from this situation as the undegraded dyestuff molecules are more effectively removed in the biological treatment plant by adsorption processes, than would smaller, more watersoluble degradation products. Under anaerobic conditions, such as in digesting sewage sludge, the indications are that dyestuffs degradation takes place at least slowly, as shown by decolorization. lt is probable that subsequent total mineralization in the environment proceeds through a series of both aerobic and anaerobic degradation steps. However, studies of degradation pathways are so demanding of 195 Organic Dyes and Pigments resources, that in the case of dyestuffs they have only been undertaken for a few model compounds, e.g. a series of substituted phenylazonaphthalene compounds [98]. The wide variety of biodegradability tests available and their usefulness was recently summarised by Gilbert [99]: Table 8. Summary ofbiodegradation tests (After Gilbert, P.A. [99]) Test classilication I. Diodegradability potential tests a) Ready biodegradability b) Inherent biodegradability II. Simulation tests a) Biological treatment (aerobic) b) Biological treatment (anaerobic) c) River Primary biodegradation Ultimate biodegradation OECD screening test Modilied OECD screening test(DOC) Closed bottle test (0 2) Closed bottle test (0 2 saturation) (00 Sturm C0 2 evolution test (C00 AFNOR T90-302 (DOC) MITI test (0 2) Zahn-Wellens test (DOC)a SCAS with sewage feed (DOC) 8 Modilied activated sludge test Bunch-Chambers test SDA semi-continuous Coupled units test• activated sludge test• OECD confirmatory test• Porous pot test• Anaerobic digestion test• River die-away test General river elimination test d) Estuary e) Sea f) Soil • For these tests it may on occasion be difficult to distinguish between biodegradation and bio-elimination All such tests need to be carefully interpreted as they grossly simplify the complex natural situation. It must of course be borne in mind, that commercial dyestuffs are frequently mixtures of different colorants and other noncolored constituents necessary for the satisfactory use of the dye. These non-colored compounds may be readily biodegradable, in which case the commercial dyestuff may show a substantial degree of biodegradability in a non-specific biodegradability test, even though the colorant itself may be unchanged. 196 E. A. C1arke,.R. Anliker Metabolie sturlies of dyestuffs have essentially been confined to those dyestuffs whose use involves a significant or deliberate human exposure, i.e. dyestuffs used as food additives, or in cosmetics or drugs. Because of the intensely colored nature of dyestuffs, only low concentrations are tolerable in waters, thus imposing a low ceiling on the concentration of subsequently formed biodegradation products, which in any case would generally be expected to undergo more rapid further degradation. Azo Dyestuffs By far the most intensively investigated dass of dyestuffs are the azo dyes, and the Iiterature up to 1969 has been comprehensively reviewed by Walker [100]. The discovery [101] that the ability of the azo dyestuff Prontosil to hea1 Streptococcal infections was due to the in vivo cleavage of the azo linkage to form sulphanilamide, which was the active antibacterial, is of historical importance in chemotherapy. This reductive cleavage is characteristic of the metabolism of azo dyestuffs and has been demonstrated using gut microflora [102], cell-free extracts of gut microflora [103], liver enzymes [104], and environmental bacteria [105], and indeed azo reductase activity has been reported in other tissues [106]. In the case of food dyestuffs, which are typically water-soluble, sulphonated, acid dyestuffs, little absorption from the gastrointestinal tract takes place and it appears that the gut microflora are more important than the liver reductase system in any metabolism which does occur. However, the primary metabolites, even the sulphonated primary amines, can be absorbed to a much greater extent from the gut and tend to appear in the urine or bile rather than the faeces [107]. The apparent generality of the azo reductive cleavage has prompted concern about the potential hazards associated with exposure to azo dyestuffs which could metabolize to recognized carcinogens. For example Rinde and Troll [25] demonstrated that some azo dyestuffs derived from benzidine are metabolized in Rhesus monkeys to benzidine, the human carcinogenicity of which is beyond doubt [1 08-110]. Benzidine ancl dianisidine have been detected [111] in the urine ofworkers exposed to dyestuffs manufactured from these intermediates. Such findings must be interpreted carefully as the detection of a carcinogenic metabolite in the urine does not provide conclusive proof of risk from the compound exposed [112]. Nevertheless it does pose the question of possible environmental risk from reductive cleavage of azo dyestuffs to toxic intermediates6 either in dyeworks reductive stripping operations, or through microbial biodegradation of dyestuffs released to the environment in process effluent streams. 6 Most dyestuff manufacturing firms abandoned the manufacture of dyestuffs derived from benzidine in 1972 Organic Dyes and Pigments Photodegradation 197 nR~/N=- Chem. degradation ~'-/ I Biodegradation k 1 Reduction Hydroxylation, oxidation, hydrolysis, etc. k2 Further biodegradation, mineralization j Fig. 3. Biodegradation of azo dyes. Each R is any ofvarious substituents (typically S03H, COOH, OH, N0 2, NH 2 , NH-, N =, N = N-, alkyl, halogen). General consideration of the kinetics of this simplified scheme indicates that accumulation of the intermediate amines would only arise if the rate of dyestuff degradation (k 1) exceeds the subsequent amine degradation rate (k2). Although there seems little doubt that k2 ~ k 1 in many instances (e.g. dyestuffs which form aniline on degradation), the generality ofthis situation has not yet been substantiated. However, that amines can be expected to be degraded fairly readily by natural ecosystems is supported by the recent studies [113] of the aerobic degradation of benzidine. A1though reduction is undoubted1y the major metabolic step occurring in the gastrointestinal tract other modifications have been reported including hydro1ysis of conjugates [114], acety1ation [115], heterocyclic ring cleavage, e.g. for tartrazine [116]. In the liver, conjugation [117], acety1ation [107], demethy1ation [118], and hydroxy1ation [119] are the main metabolic processes reported in addition to azoreductase activity. Triphenylmethane Dyestuffs Compared with the azo dyestuffs, the triphenylmethane dyestuffs have received little attention in terms of metabolic studies, although this group still contains several products which are used as food additives. These food additives are highly water soluble products containing sulphonic acid groups and they are poorly absorbed from the gastrointestinal tract [120]. The absorption, excretion and distribution ofthe dyestufffollowing oral ingestion and intravenous injection has been reported for Benzyl Violet 4B (C.I. 42640) [121] and Guinea Green B (C.I. 42085) [122]. E. A. Clarke, R. Anliker 198 !n\-cH2J*={)~Cf5> SO~a C2Hs ~ ~ V C2Hs ~ S03 N(CH 3h Fig. 4. Benzyl Violet 4B, C.l. Acid Violet 49, C.l. Food Violet 2, C.I. 42640 Fig. 5. Guinea Green B (FD & C. Green No. 1). C.l. Acid Green 3, C.I. Food Green 1, C.I. 42085 Xanthene Dyestuffs The most important xanthene dyestuffs are fluorescein and its mono- and poly-halogenated derivatives, and it has been reported [123] that the monohalogenated fluoresceins are degraded to fluorescein in rats whereas the poly-halogenated derivatives, e.g. eosin, are metabolically inert. Webband Hansen [124] demonstrated the stepwise de-ethylation ofRhodamine B to 3',6'-diaminofluoran and subsequent studies [125] indicate that this process occurs in the liver cell microsomes. Accumulation and Persistence Dyestuffs in general must be classified as substances which biodegrade only slowly in the environment. This raises the question as to whether they persist to an extent which could present toxic hazards for the environment, andin particular whether there is any propensity to bioaccumulate with the possibility of affecting man by transport through the food cycle. The "Yusho" incident [29], followed by the discovery ofthe wide environmental distribution of polychlorobiphenyls (PCBs) in Japan, prompted the enactment of the Chemical Substauces Control Law which came into force in 1974 [97]. This law requires that all new substances tobe marketed in Japan, which are not demonstrated to be biodegradable, must be subjected to a bioaccumulation test in fish. Although this requirement for an expensive (ca. $ 25,000.-) fish accumulation test is unique to Japan, the possible environmental build-up of a substance was also a concern of the US Environmental Protection Agency in formulating their criteria for identifying hazardous Organic Dyes and Pigments 199 waste [126] and some indication of the tendency of a new product to accumulate will undoubtedly also be a feature of most new product controllegislation. In regard to the assessment of accumulation potential there are indications that in many instances a non-biological screening test may provide a reliable means of identifying products of low bioaccumulation potential. This test involves the determination of the n-octanolfwater partition coefficient (P) [127, 128]. Neely [127] has demonstrated a linear correlation between log (bioconcentration) of several chemieals in trout muscle and log (partition coefficient). The experimental determination of P can be difficult for products which are highly polar, insoluble, or highly lipophilic but it is also possible to derive P mathematically [129, 130] with sufficient reliability to assess whether there is any likelihood ofthe compound bioaccumulating to an unacceptable extent (a factor of 100 or even higher is generally considered acceptable). In compliance with the Japanese Chemical Substances Control Law the dyestuffs manufacturing industry has conducted fish accumulation studies on a large number of new products. An investigation of these results confirmed that no products with Pcaic. value less than 1,000 showed an accumulation factor of over 100. Although it would be premature to use the partition coefficient as a decisive criterion of bioaccumulation potential, it does seem reasonable to conclude that such fish accumulation tests are superfluous in the case of highly polar water-soluble dyestuffs. Products which are almost completely insoluble in water present particular experimental difficulties both in the fish accumulation test and for the measurement of partition coefficient. These difficulties have necessitated, for the purpose of compliance with the Japanese legislation, the development of methods for solubilizing these substances for bioaccumulation testing under conditions that have no relevance to the real situation in nature: because of the application technology and their extremely low solubility these substances do not reach open waters to any significant extent. Toxicological Aspects Toxicity to Aquatic Organisms Fish The fish is a particularly important test animal not only because its well being is a useful indicator ofthe general condition ofwaters, but also because it is an important source of food for human populations. In 1973 Little and Lamb [131] reported extensive studies ofthe toxicity of 46 dyes to the fathead minnow (Pimephales promelas). A survey of available fish toxicity data on over 3,000 commercial products by ETAD member firms indicated that about 98% have LC50 values greater than 1 mg/1, a concentration at which colored pollution of a river would normally be observable. 200 E. A. Clarke, R. Anliker The remaining 2% consisted of 27 different chemical structures including 16 Basic dyestuffs of which 10 were of the triphenylmethane type. In only one case was the LC 50 as low as 0,01 mg/1, which is comparable to DDT (0,006 mg/1) and synthetic pyrethrin (0,025 mg/1) [132]. Although internationally little harmonisation has been achieved in terms of standardization in fish species, it is generally accepted that different species are unlikely to exhibit major differences in sensitivity and indeed a comparative study of three dyestuffs on minnows, trout, and golden orfe showed similar sensitivity [133]. Algae As algae are important components of aquatic ecosystems, and algal photosynthesis is a critical source of oxygen supply, a knowledge of the effect of dyestuffs on algal activity is essential to the evaluation of the possible impact of discharges to oxidation ponds or receiving waters. In 1974 Little and Chillingworth [134] reported the effect of 56 selected dyestuffs on the growth of green alga (Selenastrum capricornutum). The results werein genera1 agreement with the relative toxicities shown in the fish studies [131, 135], with the exception that many of the acid dyestuffs exhibit high toxicity to fish but do not significantly inhibit algal growth. The toxicity of 12 aminoanthraquinone dyes to Selenastrum capricornutum and Pimepha/es promelas has also been reported [136]. Mammalian Toxicity General Aspects The factors to be considered in the assessment of the toxicological risk of a colorant include: the total exposure potential, the seriousness of the toxic effect, and the fact or the probability of its occurrence. The complexity of assessing evidence of toxicity and specially carcinogenicity dictates that the evaluation of the potential human hazard of a given compound must be individualized in terms of the chemical and metabolic aspects ofthat specific agent, its intended uses, the data available at the time that a decision must be made, and other factors pertinent to the case under consideration. One can distinguish the following principal groups of colorants with regard to test requirements: colorants for drugs and food, for cosmetics, and for technical use for such as textiles, plastics, paints, leather, and paper. Toxicological studies of colorants for technical use are designed primarily to evaluate possible effects on exposed populations in the manufacturing and processing plants, and should be adequate to indicate any unacceptable risk of adverse effects at the much lower Ievels experienced by the consumer7, or the general public through environmental pollution. 7 Special consideration is required in the case of higher exposure outlets, e.g. finger-paints, do-it-yourself products, children's and artists' paints 201 Organic Dyes and Pigments Acute Taxicity The product safety data sheets now made available to customers by the major dyestuffs manufacturers [137] include the following information: 1. Acute oral LD 50 in rats 2. Skin irritiation on rabbits 3. Eye irritation on rabbits obtained using standardized methods. These data provide a basis for recommendation of appropriate handling precautions, and under present day minimum acceptable standards of handling and working environment the acute toxicity of dyestuffs is not a problem. Table 9. Determination of single administration toxicity of dyestuffs (commercial brands) by ETAD members• No. of dyes tested 4461 (lOO)b No. of dyes with LD 50 values (mg/kg) <250 250-2,000 2,000-5,000 >5,000 44 (1) 314 (7) 434 (9.7) 3,669 (82.3) • Acute oral toxicity LD 50 in rats. Only data published in the ETAD safety data sheets up to August 1977 b The values in parenthesis are perccntages A survey [34] ofthe data available in mid 1977 indicated that dyestuffs are of generally low acute toxicity (Table 9) as only 1% of the commercial products showed LD 50 values under 250 mgjkg (none was less than 100 mgjkg). Closer investigation revealed that this more toxic 1% involved only 15 different chemical structures (Table 10). Sensitizatian Contact dermatitis or skin sensitization effects have been experienced with several specific dyestuffs bothin terms ofmanufacturing or processing experience [138~40] and as a result ofconsumer exposure to dyed fabrics [138, 141, 142]. There are no known reports of such effects arising as a result of environmental pollution by dyestuffs. Chranic Taxicity / Carcinagenicity The question ofmost concern to managements, workers and public is whether a chemical product can produce a carcinogenic effect in humans as a result of chronic exposure at low Ievels. Epidemiological evidence in general supports the conclusion that, provided sensible working procedures are used, there is no significantly higher cancer incidence among exposed workers [143, 144]. E. A. Clarke, R. Anliker 202 Table 10. Analysis ofLD 50 values lower than 2,000 mg/kg by chemical type• Chemical type Monoazo Monoazo/ quatemary Disazo Trisazo Phthalocyanine Diphenylmethane Triphenylmethane Xanthene Oxazine Methine Anthraquinone lndigoid Stilbene azo Miscellaneous Total Total No. of products No. ofproducts with LDso (mg/kg) 14 16 20 1 2 1 13 6 4 20 3 1 2 11 114 :::250 251-2,000 1 (125) 0 4(199,200,240,200) 0 0 0 1 (100) 2 (220, 250) 1 (210) 4 (133, 213, 222, 224) 0 0 1 (150) 1 (221) 15 13 16 16 1 2 1 12 4 3 16 3 1 1 10 99 • Criteria for analysis: 1. Mixture products were omitted 2. Azoic diazo components (28) were not included 3. Where products with the same basic structure, as identified by Colour Index No., have been tested by different member firms, only the lowest LD 50 value was used 4. Where LD 50 values were quoted as a range, the lower value has been used for classification. By this procedure the total number (358) of tested commercial products was reduced to 114 products with different structures Evidence of adverse effects in human populations exposed to dyestuffs has however been reported [145-149]; for example, among kimono painters who had the habit of pointing their paint-brushes between their lips [148], and workers whose working conditions were apparently inadequate [149]. The primary prophylactic measure is to reduce exposure to a minimum by adopting good working procedures and personal hygiene, whilst at the same time seeking to detect any product which is a strong carcinogen. Evidence of carcinogenicity can be obtained from three sources [150]: (1) Epidemiological evidence from exposed human populations. Although providing persuasive evidence of actual risk to man at the exposure levels experienced, this approach has the disadvantage of being retrospective. Furthermore, negative data cannot normally adequately establish noncarcinogenicity of a product although they can help define the upper limits ofrisk. (2) Experimental animal studies. These provide the best available experimental evidence of carcinogenic potential, but the high cost (ca. $ 200,000.-) of a properly conducted bioassay, limited testing resources, and the problems ofinterpreting positive results obtained at high dosage levels in terms of risk at low practical exposure levels has limited their application in the case oftextile dyestuffs. In this context the proposal by the Ameri- Organic Dyes and Pigments 203 can Iudustrial Health Council to categorize carcinogens by potency is a sensible approach [151]. (3) Evidence from studies of chemical structure, reactivity, and mutagenic effects as detected by various short-term tests. Such additional indication of carcinogenic potential must still be regarded as simply suggestive [152, 153]. However, the rapid developments in mutagenicity testing are of great interest to the dyestuff industry, which like other sectors of the chemical industry, would benefit from the availability of a rapid low-cost method for reliably assessing carcinogenic potential. Several extensive reviews ofthe carcinogenicity of dyestuffs [154] particularly of azo dyestuffs [155-160] and food dyestuffs [120, 161] have been written. Evaluation ofthe available Iiterature is made difficult by the fact that the identity and quality of the test substance are frequently not defined unequivocally, and often the route of exposure is not relevant (e.g. sub-cutaneous injection, pellet implantation). Most studies fall short of the criteria proposed by the lnteragency Regulatory Liaison Groups [150] and indeed of the 58 organic colorants listed by NIOSH [162] as having been reported as carcinogenic or neoplastic only about 20--25% can be regarded as having been adequately tested. Bioassay studies have concentrated largely on food dyestuffs and based on the evaluation of the results several products are no Ionger permitted as food additives. The finding that three benzidine-based dyestuffs produced preneoplastic hepatic lesions in rats in a 13-week subchronic study [149, 163] has led to concern about a possible hazard from azo dyestuffs based on benzidine or its derivatives [164], but no complete bioassay on a benzidine-based dyestuff has yet been reported. Available evidence supports the conclusion that organic pigments are of low toxicity presumably because their generally very low solubility means that they are scarcely available for biological action. Few 2-year feeding studies have been completed although two pigments based on 3,3' -dichlorobenzidine (DCB) - Pigment Yellow 12 (3) and Pigment Yellow 83 - were found to be non-carcinogenic [165, 166]. These results are of particular interest because reductive cleavage of the azo linkages would have led to the regeneration of DCB, an animal carcinogen [167] and therefore a suspected human carcinogen, although epidemiological evidence in plants where DCB was handled for many years did not substantiate that there was a toxic risk under the conditions used [168-170]. No evidence ofmetabolism ofthese pigmentswas found in this work, or in a recent study of Pigment Yellow 13 [24], contrary to the earlier report by Akiyama [171]. No carcinogenic effects were found in feeding studies with Pigment Y ellow 16 (based on o-tolidine) [165], or the monoazo pigmentsPigmentRed 49 [172], and Pigment Red 53:1 [173]. Mutagenicity Although organic colorants have been tested in a variety of short-term mutagenicity procedures [174], notably the Salmonella microsome assay deve1oped 204 E. A. Clarke, R. Anliker by Ames, none of these methods can be regarded as having been established as reliable indicators of carcinogenic potential. Although most carcinogenic azo dyestuffs of the p-dimethylaminoazo-benzene series are mutagenic [17 5], negative mutagenic results have been obtained on several other monoazo dyestuffs which have been reported tobe animal carcinogens [176]. Mutagenic activity has been reported [177-179] in many anthraquinone derivatives and dyestuffs but a series of22 acid dyes (6 anthraquinone, 14 azo and 2 nitro) were non-mutagenic [180] in E. Coli TM4 and S. typhimurium TAlOO. Permitted foodcolors [181, 182] andcosmeticcolors [183] have been found to be non-mutagenic with the exception of Pigment Orange 5 which is used in some lipsticks. A study of colorants important in the graphic arts and painting industry [184] included 16 organic pigments of which 2 (Pigment Orange 5 and Pigment Red 1) were reported to be weakly mutagenic. In both cases the test substance was dissolved in DMSO and the interpretation of results in terms of risk under practical conditions, from products which are extremely insoluble in water and fat, needs clarification. Legislation In recent years there has been a dramatic increase in regulatory activities aimed at achieving safer manufacture, use and disposal of chemicals, including colorants [185]. The complexity of internationallaws and regulations is now suchthat their review is extremely difficult [186]. Regulations in themselves are not opposed by responsible industry provided they effectively serve certain specific needs and achieve their objectives in a cost-effective manner. The regulation of gaseous and liquid effluents from manufacturing and processing plants, as well as the disposal of industrial waste, has evolved gradually in the developed countries and generally operates on a regional basis (State, local authority, or regional waste board) within a framework of nationallegislation. This allows some flexibility in approach and enables the requirements to be geared to the local situation. For the dyestuffs industry, where pollution problems are essentially localised, this approach operates satisfactorily, although an international approach is undoubtedly required to deal with major pollution of international waterways (e.g. Rhine river, W. Europe). Dyestuffs are subject to the laws which apply to chemieals in general. The first stepwise efforts to exercise environmental controls were through national laws which applied to specific limited areas, e.g. the US national water pollution controllaws. Table 11 shows the year of introduction of the major environmentallaws in the OECD countries [187]. The complexity of the interaction of various Acts contributing to water pollution control is best illustrated by the US situation [188], involving the Federal Water Pollution Control Act, Marine Protection, Research and Sanctuaries Act (1972), Safe Drinking Water Act (1974), Resource Conserva- 205 Organic Dyes and Pigments Table 11. Date of introduction of major national water pollution control laws in the OECD countries OECD country Australia Austria Belgium Canada Denmark Finland France Germany Greece leeland Ireland Italy Year 1959 1971 1970 1961, 1965 1964 1957, 1976 1978 1977 1976 OECD country Year Japan Luxembourg N etherlands NewZealand Norway Portugal Spain Sweden Switzerland Turkey United Kingdom United States 1958, 1970 1970, 1975 1967, 1974 1970 1969 1971 1971 1961, 1974 1972, 1977 tion and Recovery Act (1976), Hazardous Materials Transportation Act (1974), Portsand Waterways Safety Act (1972), Federal Insecticide, Fungieide and Rodenticide Act (1972), Taxie Substances Control Act (1976), Atomic Energy Act (1954). The legislative process on an internationallevel is even more complex. For example within the EEC the process of developing a directive for control ofwater quality was protracted over several years [189]. Examples of legislation concerned with the working environment are the Health and Safety at Work Act (1974) in the UK and the Occupational Safety and Health Act (1970) in the USA. Some recent legislation has sought tobe much wider in scope and to tackle the totality of the effects of chemieals on man and the environment. Probably the most far-reaching environmentallegislation introduced so far is the US Taxie Substances Control Act (1976). In order to meet the requirements of this Act a register of existing commercial chemieals has been compiled and the introduction of a new chemical to the market must be notified to the authorities and an extensive dossier submitted which includes appropriate ecological and toxicological data. A similar notification system is foreseen in the EEC Directive "6th Amendment ofthe Council Directive of June 27, 1967 on the approximation of laws, regulations and administrative provisions relating to the classification packaging and labeHing of dangeraus substances" [190]. In Japan the Chemical Substances Control Law (1973) requires an expensive fish accumulation test for new substances which are not readily biodegradable. Provided newly introduced dyestuffs and pigments show no significant accumulation tendency in this accumulation test they are permitted under the terms of the law [97]. The increased incidence of bladder cancer among dyestuff manufacturing workers was attributable to the carcinogenic properties of some amines used as intermediates, particularly benzidine, 2-naphthylamine and 4-aminobiphenyl. These chemieals have been either banned or subjected to strict regulatory controls of handling and use in most industrial countries, but unfortunately N 0 0'\ Table 12. Prohibition or control ofvarious dyes and chemieals as carcinogens Product Australia Belgium Finland W.Ger- Italy many 4-Aminobiphenyl 4-Aminobiphenyl salts Auramine + Benzidine Benzidine salts Dianisidine + + + + + + + + + + + + + + + + +a + + + + + Dianisidine salts 3,3' -Dichlorobenzidine 3,3' -Dichlorobenzidine salts 4-Dimethylaminoazobenzene Magenta 1-Naphthylamine 1-Naphthylamine salts 2-Naphthylamine 2-Naphthylamine salts + + + + + + o-Tolidine + o-Tolidine salts a Addition proposed in draft "Carcinogenic Substarrces Regulations" + + Japan Sweden Switzer- UK land + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + + USSR USA + + + + + + + + + + + ~ ~ Q !» ..., PI" s• !" > 2.. ~ ..., (\) 207 Organic Dyes and Pigments there has been much diversity in these controls from country to country [191] (Tab1e 12). This particular example underlines the urgent necessity for much closer international harmonisation of legislation and testing methodology: if the current tendency to insist on national preferences for test methodology persists, the limited resources and test capacitywill be unnecessarily squandered with minimal benefits in terms ofproduct safety. Acknowledgement The authors gratefully acknowledge the invaluable cooperation ofthe various members of the ETAD committees. Thanks are due in particular to Dr. D. Brown for his critical comments on the manuscript and to Dr. Dorothee Braun-Steinle for assistance in preparation of this contribution. AZO (a) Monoazo (2) Acid Yellow 23 C.I. 19140 tartrazine Disperse Yellow 3 C.I. 11855 (b) Disazo CH 3 0 \._ C~OH j NHCO~= I II -0--0Cl Cl - \._ j \._ j N=~COH CH 3 C~OH I II 0 \._ j (3) Pigment Yellow 12 C.I. 21090 Stilbene S0 3 Na S0 3 Na c,u,oON=N--0-c•-cn-b--N=N-ooc,n, Direct Yellow 12 C.I. 24895 (4) E. A. Clarke, R. Anliker 208 Triphenylmethane Diphenylmethane (6) Basic Yellow 2 C.I. 41000 auramine Acridine Xanthene (7) Basic Orange 14 C.I. 46005 Basic Violet I 0 C.I. 45170 Quinoline disulphonated 0 0 Disperse Yellow 54 C.I. 47020 (9) (10) Acid Yellow 5 Direct Y ellow 5 C.I. 47035 209 Organic Dyes and Pigments Azine Methine ~CH CH3 ~NH=C-9/; CH 3 I Cl- CH OCH3 ~ 3 PhNH ~NH _ 6 N OCH 3 Basic Yellow II C.I. 48055 3 ~CH 3 ~) ~ 0 1) 2 [S04]I/2 ~ I cH 3 (1 2 ) C.I. 50245 mauveine Oxazine Thiazine (C2Hs)N~Oh N~ CX (CH 3)2N ~N BasicBlue9 Solvent Blue 8 C.I. 52015 methylene blue (13) Basic Blue 3 C.I. 51004 Sulphur 3)2 ~N(CH (14) Indigoid heated with sodium polysulphide 0 Vat Blue I Pigment Blue 66 C.I. 73000 indigo (15) Sulphur Black I C.I. 53185 (16) Anthraquinone @ 0 0 NH9CH, o)OOH so~' NH--pcH, S0 3 Na Acid Green 25 C.I. 61570 0 (17) Mordant Red 11 Pigment Red 83 C.I. 58000 alizarin (18) 210 $ E. A. Clarke, R. Anliker 0 NH HN$(19) 0 0 I~ ~I ~ ."-:;::. 0 Pigment B1ue 60 Vat Blue 4 C.I. 69800 in danthrone Phthalocyanine Direct B1ue 86 C.I. 74180 Fig. 6. Se1ection of some representative chemica1 structures References 1. Anliker, R.: Ecotox. Environ. Safety 1, 211 (1977) 2. Wiedhaup, K.: Acta Pharm. Techno!. Suppl. 8, 67 (1979) 3. Anliker, R., Müller, G.: F1uorescent Whitening Agents. Environ. Qual. Safety Suppl. 4. G. Thieme, Stuttgart, Academic Press, New Y ork 1975 4. Venkatamaran, K. (Ed.): The Chemistry of Synthetic Dyes. Academic Press, New York 1952 (Vol. I) to 1978 (Vol. VIII) 5. Co1our Index, 3rd Edition. Soc. Dyers and Co1ourists. Bradford, England 1971 (Vol. 1-4), 1975 (Suppl. Vol. 5,6) 6. US Tariff Commission: Synthetic Organic Chemicals, U. S. Production and Sales. US Government Printing Office, Washington, issued annually 7. Venkatamaran, K. (Ed.): The Analytical Chemistry of Synthetic Dyes. J. Wiley & Sons, New Y ork 1977 8. Weisburger, J.H., Weisburger, E.K.: Food Cosmet. 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Fischer, K.: Melliand Textilber. 59,487, 659 (1978) 52. Games, L.M., Hites, R.A.: Analyt. Chem. 49, 1433 (1977) 53. ADMI: Dyes and the Environment, Vol. 2., Amer. Dye Manufacturers Inst. Inc., New York 1974 54. APHA: StandardMethodsforthe Examination ofWaterand Waste Water. 13th Ed. Amer. Public Health Assoc., New Y ork 1971 55. Allen, W. et al.: Determination of color of water and waste water by means of ADMI color values. In Ref. 5, Chap. III 56. Tincher, W.C.: Survey ofthe Coosa Basin for Organic Contaminants from Carpet Processing. Environmental Protection Division, Dept. ofNatural Resources, Contract No. E-27630; Georgia, USA, Oct. 1978 57. Oswalt, Y.H., Land, Y.G.: Color removal from kraft pulp mill effiuents by massive time treatment, US Environmental Protection Agency, EPA-R2-73-086; Washington, D .C. 1973 58. Rinker, T.L.: Treatment oftextile wastewater by activated sludge and alum coagulation; US Environmental Protection Agency, EPA-600/2/75-055; Washington, D.C. 1975 59. 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Color. 5, 255 (1973) 88. Masselli, J.W., Masselli, N.W., Burford, M.G.: Text. lnd. 135 (10), 84, 108 (1971) 89. Porter, J.J., Snider, E.H.: J. Water Pollut. Control Fed. 48,2198 (1976) 90. Weeter, D.W., Hodgson, A.G.: Am. Dyest. Rep. 66 (8), 32 (1977) 91. Dohänyos, M., Madera, V., Sedläcek, M.: Prog. Wat. Tech.10, 559 (1978) 92. Hitz, H.R., Huber, W., Reed, R.H.: J. Soc. Dyers Colour. 94, 71 (1978) 93. Meier, H.: Photochemistry ofDyes, in: The Chemistry ofSynthetic Dyes, Vol. IV (Venkatamaran, K. (Ed.)), Academic Press, NewYork 1971 94. Ra:dding, S.R. et al.: Review of the Environmental Fate of Selected Chemicals. Office of Toxic Substances, EPA 560/5-77-003; Washington, D.C. 1977 95. Porter, J.J.: A Study of the Photodegradation of Commercial Dyes. US Environmental Protection Agency, EPA-R2-73-058; Washington, D.C. 1973 96. Heitz, J.R., Wilson, W.W.: Photodegradation ofha:logenated xanthene dyes. ACS Sympo- Organic Dyes and Pigments 97. 98. 99. 100. 101. 102. 103. 104. 105. 106. 107. 108. 109. 110. 111. 112. 113. 114. 115. 116. 117. 118. 119. 120. 121. 122. 123. 124. 125. 126. 127. 128. 213 sium Ser. 73, Disposaland Decontamination ofPesticides (Kennedy M.V. (Ed.)), Washington, D.C. 1978 Kubota, Y.: Ecotox. Environ. Safety 3, 256 (1979) Meyer, U., Ovemey, G., von Wattenwy1, A.: Textilveredlung 14, 15 (1979) Gilbert, P.A.: Ecotox. Environ. Safety 3, 111 (1979) Walker, R.: Food Cosmet. Toxicol. 8, 659 (1970) Trefouel, J. et al.: Compt. rend. soc. biol. 120, 756 (1935) Chung, K.T., Fulk, G.E., Egan, M.: Appl. Environ. Microbiol. 35, 558 (1978) Hartmann, C.P., Fulk, G.E., Andrews, A.W.: Mutat. Res. 58, 125 (1978) Manchon, Ph., Lowy, R.: Food Cosmet. Toxicol. 3, 783 (1965) Idaka, E. et al.: J. Soc. Dyers Colour. 94,91 (1978) Juchau, M.R., Krasner, J., Yaffe, S.J.: Biochem. Pharmacol.J7, 1969 (1968) Daniel, J.W.: Toxicol. Appl. Pharmacol. 4, 572 (1962) Case, R.A.M. et al.: Brit. J. Ind. Med. 11, 75 (1954) Barsotti, M., Vigliani, E.C.: A.M.A. Arch. Ind. Hyg. Occup. Med. 5, 234 (1952) Scott, T.S.: Brit. J. Ind. Med. 9, 127 (1952) Genin, V.A.: Vopr. Onkol. 23 (9), 50 (1977) Batten, P.L., Hathway, D.E.: Brit. J. Cancer 35, 342 (1977) Tabak, H.H., Barth, E.F.: J. Water Pollut. Control Fed. 50, 552 (1978) Williams, R.T., Milburn, P., Smith, R.L.: Ann. N. Y. Acad. Sei. 123, 110 (1965) Scheline, R.R., Longberg, B.: Acta Pharmacol. Toxicol. 23, 1 (1965) Roxon, J.J. et al.: Food Cosmet. Toxicol. 5, 447 (1967) Williams, R.T.: Detoxication Mechanisms. John Wiley Ltd, London 1959, p. 4 Müller, G.C., Miller, J.A., Glassner, M.: J. Biol. Chem. 202, 579 (1953) Larsen, J.C., Tarding, F.: Acta Pharmacol. Toxicol. 39, 525 (1976) Radomski, J.L.: Am. Rev. Pharmacol. 14, 127 (1974) Minegishi, K.l., Yamaha, T.: Toxicology 7, 367 (1977) Minegishi, K.I., Yamaha, T.: Chem. Pharm. Bull. 22, 2042 (1974) Webb, J.M., Fonda, M., Brouwer, E.A.: J. Pharmacol. Exp. Ther. 137, 141 (1962) Webb, J.M., Hansen, W.H.: Toxicol. Appl. Pharmacol. 3, 86 (1961) Webb, J.M. et al.: ibid. 3, 696 (1961) EPA Proposed Criteria for identifying hazardous waste; Standards for generators, and standards for management facilities, Sect. 250.14. US Fed. Register 43, 58946 (Dec. 18, 1978) Neely, W.B., Branson, D.R., Blau, G.E.: Environ. Sei. Techno!. 8, 1113 (1974) Chiou, C.T. et al.: ibid.JJ, 475 (1977) 129. Fujita, T., Iwasa, J., Hansch, C.: J. Am. Chem. Soc. 86, 5175 (1964) Hansch, C. et al.: J. Med. Chem. 16, 1207 (1973) 130. 131. 132. 133. 134. 135. 136. 137. 138. 139. 140. 141. 142. 143. 144. 145. 146. 147. Little, L.W., Lamb, J.C.: in ref. 35, Chap. V Tooby, T.E., Hursey, P.A., Alabaster, J.S.: Chem. Ind. (12), 523 (1975) Hamburger, B., Häberling, H., Hitz, H.R.: Arch. Fisch Wiss. 28, 45 (1977) Little, L.W., Chillingworth, M.A.: in ref. 53, Chap. II Little, L.W., Chillingworth, M.A.: in ref. 53, Chap. IV Chillingworth, M.A.: in ref. 53, Chap. V Anliker, R.: Swiss Chem. 1 (9), 34 (1979) Cywie, P.L. et al.: Les eczemas allergiques professionnels dans l'industrie textile; Inst. Nat. Rech. de Securite, Rep. No. 244/Rl. Paris, March 1977 Gardiner, J.S. et al.: Brit; J. Dermatol. 85,264 (1971) Alanko, K. et al.: Clinical Allergy 8, 25 (1978) Cronin, E.: Trans. St. John's Hosp. Dermatol. Soc. 54, 156 (1968) Sim-Davies, D.: ibid. 58, 251 (1972) Ferber, K.H., Hili, W.J., Cobb, D.A.: J. Am. Ind. Hyg. Assoc. 37, 61 (1976) · Proportional mortality study on members of the National Union of Dyers, Bleachers, and Textile Workers (NUDBTW), Bradford, England, unpublished Anthony, H.M.: J. Soc. Occup. Med. 24, 110 (1974) Cole, P., Hoover, R., Friedell, G.H.: Cancer 29, 1250 (1972) Yoshida, 0. et al.: lgaku no Ayurni 79,421 (1971) 214 E. A. C1arke, R. Antiker 148. Yoshida, 0., Miyakawa, M.: Etio1ogy of b1adder cancer: Metabolie aspects, in: Ana1ytic and Experimental Epidemio1ogy of Cancer (Nakahara, W. et al. (Eds.)), University Park Press, Ba1timore 1973 149. NIOSH/NCI Joint Current Intelligence Bulletin 24. Direct B1ack 38, Direct B1ue 6, and Direct Brown 95, benzidine-derived dyes. US Dept. of Health, Education, and Welfare, Aprill7, 1978 150. IRLG Work Group Rep. on the Sei. Basis for identification ofpotential carcinogens and estimation ofrisks. US Fed. Register 44, 39858 (July 6, 1979) 151. AIHC Recommended Alternatives to OSHA's Generic Carcinogen Proposal. Amer. Ind. Health Council. Jan. 9, 1978. Published in part in Chem. Eng. News 56 (5), 30 (1978) 152. Nat. Cancer Adv. Board. General criteria for assessing the evidence for carcinogenicity of chemical substances. Nat. Cancer lnst. 58,461 (1977) 153. Environmental Health Criteria 6. Principles and methods for evaluating the toxicity of chemicals. Part. I. WHO, Geneva 1978 154. IARC Monographs on the Evaluation of the Carcinogenic Risk of Chemieals to Man. Vol. 16. IARC, Lyon 1978 155. lbid. Vol. 8. IARC, Lyon 1975 156. A Iiterature survey oriented towards adverse environmental effects resultant from the use of azo compounds, brominated hydrocarbons, EDTA, formaldehyde resins and o-nitrochlorobenzene. US Environmental Protection Agency, EPA 560/2-76-005, Washington, D.C. 1976 157. Fishbein, L.: Potential Industrial Carcinogensand Mutagens. Elsevier, New York 1979 158. Burg, A.W., Charest, M.C.: An evaluation of the Iiterature concerning the potential for carcinogenic properties of bisazobipheny1 compounds used as dyes. C-82353. A.D. Little Inc., Jan. 25, 1979 159. Auerbach Associates Inc.: Benzidine derived dyes andfor pigments, AAI-2434-100-TR-2, May 2, 1978. (Submitted to US Consumer Product Safety Commission under contract CPSC-C-77-0088) 160. Calculon: Monoazo dyes and pigments, CALC-2434-300-TR-2, Febr. 9, 1979. (Submitted to US Consumer Product Safety Commission under contract CPSC-C-77-0088) 161. Drake, J.J.P.: Toxicology 5, 3 (1975) 162. NIOSH: Registry ofToxic Effects ofChemical Substances, Vol. I & II; US Dept. ofHealth, Education, and Welfare 1977 163. NCI: 13-week subchronic toxicity studies of Direct Blue 6, Direct Black 38 and Direct Brown 95 dyes, NCI-CG-TR-108, NCI Technical Rep. Ser. No. 108, 1978 164. Petition to OSHA by US Unions for emergency temporary standard for benzidine-derived dyes, May 16, 1978 165. Leuschner, F.: Toxicology Lett. 2, 253 (1978) 166. NCI/NIH Report: Bioassay of Diary1anilide Yellow for possib1e carcinogenicity, DHEW publication no. (NIH) 77-830, 1977 167. IARC Monographs on the Evaluation ofCarcinogenic Risk ofChemicals to Man. Vol. 4. IARC, Lyon 1974, p. 49 168. Gerarde, H.W., Gerarde, D.F.: J. Occup. Med.J6, 322 (1974) 169. Gadian, T.: Chem. Ind. (19), 821 (1975) 170. Maclntyre, 1.: J. Occup. Med. 17, 23 (1975) 171. Akiyama, T.: Iikei Med. J. 17, I (1970) 172. Davis, K.J., Fitzhugh, O.G.: Toxicol. Appl. Pharmacol. 5, 728 (1963) 173. Davis, K.J., Fitzhugh, O.G.: ibid. 4, 200 (1962) 174. Burg, A.W., Charest, M.C.: Mutagenicity results with implications for carcinogenicity, A.D. Little Inc., C-82875, March 9, 1979 175. Yahagi, T. et al.: Cancer Lett.J, 91 (1975) 176. Garner, R.C., Nutman, C.A.: Mutat. Res. 44,9 (1977) 177. Tamaro, M., Monti-Bragadin, C., Banfi, E.: Boll. Ist. Sieroter. Milanese 54, 105 (1975) 178. Brown, J.P., Brown, R.J.: Mutat. Res. 40, 203 (1976) 179. Brown, J.P., Dietrich, P.S.: ibid. 66,9 (1979) 180. Tamaro, M., Banfi, E.: Boll. Ist. Sieroter. Milanese 55, 191 (1976) 181. Brown, J.P., Roehm, G.W., Brown, R.J.: Mutat. Res. 56, 249 (1978) Organic Dyes and Pigments 182. 183. 184. 185. 186. 187. 188. 189. 190. 191. Viola, M., Nosotti, A.: Boll. Chim. Farm. 117,402 (1978) Muzzall, J.M., Cook, W.L.: Mutat. Res. 67, I (1979) Milvy, P., Kay, K.: J. Toxicol. Environ. Health 4, 31 (1978) An1iker, R.: Aquatic Ecological Chemistry (Japan) 1, 211 (1979) Alston, P.: Ecology Law Quart. 1978, 397 Anonymous: Environ. Sei. Techno!. 13, 786 (1979) Barrett, B.R.: Environ. Sei. Techno!. 12, 154 (1978) Tetlow, J.A.: Chem. Ind. (6), 183 (1979) Smeets, J.: Ecotox. Environ. Safety 3, 116 (1979) Anliker, R.: J. Soc. Dyers Colour. 95,317 (1979) 215 Inorganic Pigments W. Funke II. Institut für Technische Chemie, Universität Stuttgart D-7000 Stuttgart 80, Federal Republic of Germany Introduction In contrast to dyes inorganic colorants are generally used as pigments rather than in a molecular-dispersed state. Besides imparting color for decorative, indicatory and informational purposes, such pigments may serve various other or additional purposes like corrosion protection, filling or reinforcement. Table 1 presents a survey of the more important fields of application of these colorants, but does not claim completeness. From the large number of inorganic colorants some have been shown to cause hazardous or toxic effects when taken up by the human or animal organism. This may in principle, occur via the ingestive and respiratory systems, but sometimes also via skin contact. Important factors in estimating Table 1. lnorganic colorants and powders in various applications Decorative Proteelive Indicatory Compositional Plastics Rubber Coatings Building materials Putties and sealants Printing inks Artist paints Vitreous materials Ceramies Enamels Cosmetics Drugs + + + + + + + + + + + + + + + + + + + + + + + Reinforcing Filler + + + + + + + N 00 Table 2. Inorganic pigments, based on heavy metals Name Colloquial name Formula Density [g/cm3] Solubility [g/100 ml]* cold dil. HCI H20 9,1 5,7 6,9 11,3 1 Lead Red Iead oxide Calcium plumbate Lead molybdate Metalliclead pigments Red Iead, Minium Lead powder Pb304 (+ PbO) Ca2Pb04 PbMo04 Pb Chromate Lead chromates pigments Chrome yellow PbCr04 6,1 6,6 Lead silico chromate Zinc chromate Zinc yellow Zinc potassium chromate Zinc tetraoxychromate Strontium chromate Barium chromate Lemon yellow PbCr04·PbO PbCr04 · PbS04 PbCr04 · Pb(OHh PbCr04 + KFe[Fe(CN)6] PbCr04 + Si02 ZnCr04 ZnCr04 · K2Cr04 ZnCr04 · 4Zn(0Hh SrCr04 BaCr04 4,0 3,9 4,5 0,06-0,15 0,004 0,04-0,1 0,0005 s pigments Cadmium Cadmium sulfide Cadmium zinc sulfide Cadmium sulfo selenide CdS (Cd,Zn)S Cd(S,Se) 4,4-4,8 0,00013 s 5,8 I Chrome red Chrome lemon Chrome orange Chrome green i i i 0,()()()()()58 (gCr03/100cm3) i s s s s dil. CH3COOH EtOH s s s s s s s 4,1 3,5 I ~ "'j = ::I ~ Cl> ~ ~ ...... ~ ~g Name Colloquial name Silicogenous Asbestos pigments Formula Density [g/cm3] Solubility [g/100 ml] cold H20 2,8-3,4 Tale Silica Hydro silicates of var. comp. Mg3 [S4010] (OHh Si02 2,8 2,0-2,3 0,008 Pigments and Fillers for special purposes Antimony trioxide Naples yellow Lead antimonate Emerald green, Copper aceto Parisgreen arsenate Sb203 Pb3(Sb04h Cu(CH3C00)2 3Cu(As02)2 5,2 6,6 3,3 vsls "' dil. HCl dil. CH3COOH EtOH 0,0140,052 vsls s 0,03 * i = insoluble s =soluble vsls = very slightly soluble IV ...... 'Ci 220 W. Funke hazardous effects is the total surface area of exposure where interactions take place. With the exception of some fillers or extenders the average particle diameter of inorganic colorants ranges between 0, 15 and 2 J.lm. Though inorganic colorants are used for many purposes, considering the probability and intensity of possible exposures, it is reasonable to concentrate discussion on the pollution and hazards oftheir use in organic coatings and thin films. A competent and informative survey on legislation, Standards and codes of practice and toxicology for the paint industry also includes inorganic colorants [1]. In the following discussion hazards encountered in the industrial production of inorganic colorants are mentioned only incidentally since these operations are usually subjected to special regulations and safety measures. In Table 2 inorganic colorants are listed which have already been mentioned in the Iiterature as being potentially hazardous, together with some data relevant to pollution aspects. Several surveys on ecological problems with inorganic colorants and metal oxides are known from the Iiterature [2, 3, 4]. Considering the actual danger in handling objects containing hazardous colorants, its relevancy very much depends on their special use. Thus toxic inorganic pigments that have been used in old paintings hardly present any danger because it is highly improbable that non-specialists come into contact with them. On the other hand little attention is paid to possible hazards on paint removal and in the deposition and combustion of coated or painted subjects, which are quite safe when in normal use. It should be also realized that coating materials containing potentially hazardous pigments, which have been used long since by expert workers without significant accidents, are now considered dangerous after they get into the hands of amateurs in Do-itY ourself projects. The two important routes by which hazardous pigments may be absorbed by the body are inhalation or ingestion. The most dangerous is inhalation, because it allows toxic substances to be directly absorbed by the blood stream or deposited on the outer surface ofthe 1ung. In the latter case irritations may be caused with temporary or even permanent reduction in the area available for the gas exchange. As inorganic pigments are only slightly soluble in water and even less in organic solvents, they are scarcely or not at all absorbed by intact skin or mucous membranes. Though acute poisoning is highly improbable (LD 50 in most cases are > 5 gjkg) the danger of chronic effects must be considered. The solubility of pure inorganic colorants may differ from that of corresponding technical pigments. The latter may contain varying amounts of soluble impurities from the production. In using such data it should be ascertained therefore, that soluble impurities are not more toxic than the soluble fraction of the pure substance. The water soluble fraction of a pigment may be determined by extraction with water at room temperature or at elevated temperatures [5]: The sample is dispersed in carbon dioxide-free water, thoroughly stirred, water is added to a fixed volume and the dispersion filtered after vigorous shaking. A part of the filtered solution is evaporated to dryness and afterfurther drying at elevated temperature and cooling to room lnorganic Pigments 221 temperature in a desiccator the residue is weighed. This procedure is repeated until two successive weighings do not differ significantly. In considering hazards by paint films one should allow for the fact that due to the encapsulation in the binder matrix, the amount of extractible or leachable hazardous colorants is normally lower than what is expected from solubility data of the free pigment. To obtain practically significant data, the amount of leachable pigment should be determined by some extraction method [6] applied to the paint film after its formation is completed. Sources of Hazards in Using Inorganic Colorants As far as organic coatings are concerned, exposure to hazardous colorants may occur under various circumstances. Production Process The production of inorganic colorants, paint and coating materials is subject to safety regulations. They are mainly concerned with preventing hazards from inhalation of pigment dusts by the workers during mixing, dispersing, storing and handling of these materials. Application Various methods of paint application differ considerably in ease and probability of exposure to hazardous pigments. Special protection measures are required in spraying to avoid inhalation of the paint spray. In other application methods like brushing, dipping or electro-coating, exposure is possible on cleaning the equipment after the application or on changing the coating material. This is also true for electrostatic powder coating. Usually the overspray is caught by water curtains in the spraying booth or, in powder coating, by the air stream. Before disposal to the waste water system the solid waste material (paint sludge) must be removed from the water by coagulation and separation, bothin spraying of conventional solvent-based paints and in the electro-deposition of water borne paints. In the case of electrostatic powder coatings the overspray powder is recovered for further use. Performance Even more than in paint application, exposure to paint films and coatings containing hazardous pigments differs widely, depending on the special function of the coated subject. Accordingly attention has to be paid to all possible exposures in which children may be involved, e.g. indoor wall paintings, and paints for furniture, toys and pencils. Colorants for food and drugs are subjected to speciallaws 222 W. Funke and regulations, however in these materials inorganic pigments are of minor importance. Removal Removal of old paint layers before repainting or repair painting presents a specialproblern ifhazardous pigments are involved. Mechanical methods like wire brushing or sand blasting may produce dust particles that may be inhaled by the worker or contaminate the soil. Detrimental effects may be expected if these waste materials are incorporated by plants or animals. Another hazard may arise in the chemical or physical removal of paint by paint removers or solvents. Care has to be taken on disposing the waste material produced in these processes. Welding A special kind of danger in using hazardous pigments is the welding and flame cutting of painted iron or steel constructions. Due to the high temperature some pigments may vaporize or decompose to volatile toxic products. Under unfavorable circumstances toxicological hazards may be encountered even with inherently non-toxic inorganic pigments like zinc oxide. Welding of zinc-rich primers may produce zinc-oxide fumes above the TLV-Iimit of 5 mg/m 3 [1, 7, 8, 9]. Waste Disposal Discarded manufacturing batches, waste from cleaning tools and mixing, dispersing or application machines, used paint Containers, coagulated paint from spraying booths as well as various kinds of painted waste materials and painted subjects may provide another hazard when deposited, combusted or disposed of in the waste water system. Some more important possible exposures to potentially toxic colorants by which harmful effects may be expected after absorption or ingestion, are summarized as follows: 1. inhalation of pigment dust, paint spray or mechanical abrasion products on paint removal, and absorption through lung tissue; 2. Cantamination of fingers on processing or manipulating paints; paint may become embedded under the finger nails and may be transmitted to the mouth or absorbed by the skin; 3. putting water color brushes in the mouth, as during artist's work; 4. smoking when handling hazardous pigments or paint containing them; 5. handling food, eating or drinking with contaminated hands resp. dishes; 6. allowing pigments or paints to contact injured or scratched skin where direct absorption into the blood stream is possible; 7. children chewing on toys, furniture or other internal equipment in dwelling houses, that have been painted with hazardous materials. To prevent possible dangers by intoxication in most countries a series of Inorganic Pigments 223 directions and regulationsexist which describe safety measurements on handling hazardous colorants and material containing them. They will be referred to in the following discussion ofthese colorants. The analytical detection and quantitative determination of inorganic colorants is largely identical with that common in respective heavy metals. lt is therefore referred to the respective chapters on these metals. lnorganic Colorants Based on Heavy Metals Lead Pigments The toxic nature of Iead and its compounds has been known for a long time and has been thoroughly studied. Divalent Iead may replace calcium and thus be retained in the body over long periods by accumulation in the hone. However, the mobile fraction in the blood is significant. The totallead content increases as long as the contact with soluble Iead compounds continues [2]. Despite these facts, Iead containing pigmentsarestill widely used, especially as anticorrosive constituents in corrosion protective paints. The total production of Iead containing pigments and oxides in West Germany during 1978 amounted to about 56,000 tons (calculated on PbO), half ofwhich is red Iead (Pb 30 4) [10]. The toxic effect of Iead containing pigments strongly depends on their solubility in water and dilute acids [11]. Basic Iead carbonate 2PbC03 . Pb(OH) 2 (white Iead, flake white or Chremnitz white) is soluble in dilute acids. Despite giving durable paint films e.g. for window frames and walls, it is the complete acid solubility which has made this pigment dangerous to use in any paints. In the past numerous cases have been reported of Iead poisoning mainly in children who have chewed and swallowed flakes of dried Iead paints detached from treated woodwork [1, 12]. One of the most important anticorrosive pigments is red Iead, Pb30 4 (+ PbO). As for all anticorrosive pigments some solubility is essential for its corrosion protective action in primers. Despite many efforts to substitute this pigment, it is still considered to be indispensible for a number of corrosion prevention measures which require protective coatings to be applied at the location of final use. It is less weil known that inhalation oflead or Iead containing pigments is even more dangerous than ingestion. Below 5 J.lm diameter about 80% of the particles containing Iead or its compounds are incorporated on inhalation, whereas only 10% are resorbed on ingestion and most of it is excreted via the bile [13]. The ready absorption oflead from inhaled dustjustifies the stringent regulations covering almost every aspect of industrial handling of compounds containing Iead. Whereas hazards of dust inhalation are mainly limited to pigment and paint manufacturing and to mechanical paint film removal, care must also to be taken to prevent inhalation of Iead containing particles on spraying of paints. However the general regulations for handling and spraying paints are 224 W. Funke also considered to be sufficient for the application of lead containing paints [14]. Besides redleadalso dibasic lead phosphite (2PbO·PbHP0 3 • Y2H 20), lead phosphate (PblP0 4) 2 .3H20), calcium plumbate (Ca 2Pb0 4) and lead powder are used to some extent as corrosion protective pigments. Lead cyanamide, which was manufactured for corrosion protective purposes at the beginning of 1950, is not used any more, but should be considered in connection with removal of paint films from old steel constructions. F or these pigments the regulations and safety measures of working with red lead are also applicable. Some lead pigments used in artists painting, like lead antimonate (Naples Yellow), PblSb04), are specially hazardous due to their high solubility. Another source of danger is the volatilization oflead and lead oxides from paint films during welding. It has been proposed that silica should be added to anticorrosive paints containing red lead in order to transform these substances to the less volatile lead silicates [15]. For general information on numerous national and internationallegislation regulations and standards concerning the use of lead containing pigments the following Iiterature is recommended [1, 4, 16, 17]. According to the Occupational Safety and Health Administration (OSHA) in the USA the permissible exposure level for lead in the air is 50 J.lg/m3 . The lead-in-paints regulations ofthe Consumers Product Safety Commission (CPSC) denotes paints and coatings containing more than 0,06% of lead as "lead containing paints". Such paints are prohibited in toys and other objects used by children. As a consequence of the restrictions in using lead containing pigments, there is a tendency to use lead-free paints and coatings [18). As has been recently stated [1] however, there arestill some questionstobe answered: the relative contributions of the different sources of lead as a poison to man; whether the hazards result from the total amount of lead or only the lead soluble in acid organic fluids; the necessity for the use of small amounts oflead compounds in paints as drying agents in air drying paints. It has been found [11] that under the sameexperimental conditions lead naphthenate, which is a common drying agent, migrates more easily from the film than lead chromate. Attempts have been made to substitute these drying agents by lead-free compounds [19]. In view ofthere being still no equivalent alternative to the use of red lead in anticorrosive paints that can be offered for various protective applications, its future use is justifiable, provided the safety regulations and restrictions are properly observed. Chromate Pigments The worldwide production is estimated tobe about 150,000 tons annually [20] (During 1977 and 1978 almost 12,000 t of chromate pigments were produced in West Germany [10]). The hazardous nature of chromate pigments have been known for many years [21, 22]. Soluble hexavalent chromium is toxic, potentially carcinogenic and a common contact dermatitic. As all chromate pigments used in paints are lnorganic Pigments 225 soluble to some extent in acidic fluids of the body, they are hazardous to the same extent. For chromates used as corrosion protective pigments in primers, like zinc chromate, zinc tetraoxy chromate or strontium chromate, some solubility is necessary for the protective action, as with all anticorrosive pigments. Correspondingly the vehicles used in anticorrosive primers must be slightly swellable in water and more diffusible for the dissolved fraction ofthe anticorrosive pigment than vehicles of other organic coatings. In the case of chromate pigments, such as the Iead chromates, which are used as colorants only, some encapsulation effect may significantly decrease the extraction of soluble pigment on exposure to water or similar liquid media [23]. As chromate pigments are used in anticorrosive primers for car bodies, contact dermatitis has been observed with workers engaged in wet sand papering ofthese primers on car bodies [24]. There is strong evidence that lung cancer may be caused on inhalation of chromate pigment dust [25, 26]. Similar to other chromate pigments lead chromate has also been suspected ofbeing potentially carcinogenic. Above average incidence oflung cancer has been observed in factories producing both zinc chromate and lead chromate. However, it is assumed that only zinc chromate has to be blamed for this hazard [27]. More recent epiderniological studies [28, 29] showed that no increased risk for lung cancer exists in the production and processing of lead chromate pigments if modern regulations for working hygiene are observed. Despite the very low lead release from plastics pigmented with lead chromate, this pigment should not be used in paints for toys, in toys made from plastics and in packing or wrapping material coming in direct contact with food. For these purposes lead-free alternatives of colorants are used. In all other fields where colored plastics or paints are used, no economical alternative to lead chromate exists, and there is also no ecological reason for replacing it [30]. According to the OSHA-regulations in the USA 1 11g/m3 air of chrome containing carcinogenic substances are tolerated. The current TLV -resp. MAK-values for airborne levels of chromate (calculated as Cr0 3) are 0,1 mgjm3 [31, 32]. Cadmium Pigments Both cadmium and selenium in dissolved or soluble form are very toxic when directly taken up by the blood circulation system. Cadmium, especially as cadmium oxide, is a respiratory poison and industrial poisoning has been caused on exposure to fumes or dust. In paints cadmium is used as cadmium sulfide or cadmium sulfoselenide, which present a range ofhigh quality yellow to red pigments [33, 34]. There are also some composite cadmium pigments containing zinc sulphide or mercury sulfide. World production of cadmium pigments is estimated to 8,500 t annually [20], about 80% ofwhich are used in plastics and 10% each in paints and ceramic products [35]. Cadmiumpigments have superior technical properties and are widely applied in plastics, ceramies and some industrial coatmgs. 226 W. Funke In cantrast to some very toxic cadmium compounds like cadmium oxide, a series of more recent investigations have shown that the cadmium sulfide and sulfoselenide pigments are much less hazardous, especially when used in paints or plastics. Of course the usual hygienic regulations for handling industrial dusts must be observed. The solubility of cadmium pigments in 0,1 N hydrochloric acid is less than 0,1% [1, 33, 36]. Purity requirements for cadmium pigments used in plastics and coatings coming in contact with foodstuffs may be checked by a solubility test according to DIN 53 770 [27]. Animal tests with cadmium chloride showed a low Ievel oftoxicity [37, 38] andin toxicological investigations it was found [39, 40] that as much as 30 ppm of Cd given as CdC12 in the food could be tolerated over a three months period without harm. A two-years feeding test with rats resulted in a no-effect-level of as low as 10 ppm. No carcinogenic effects on oral application of water soluble cadmium sulfate to animals were observed [41]. Also no toxic effects have been reported on using cadmium pigments in paints [42]. Obviously there are significant differences in toxicity depending on whether soluble cadmium compounds are taken up orally or via other pathways than the digestive system, e.g. by inhalation. Colorants containing cadmium should not be used in food and cosmetics [43]. The use of cadmium pigments for packaging material- especially plastic packaging material - coming into contact with foodstuffs is, in general, permitted by the various European legislations. The legal regulations vary from country to country. The majority of purity requirements or maximum allowable migration values are, as a rule, met by cadmium-pigmented plastics. This is also true for plastic toys, whereas in a few countries restrictions have been iniposed on the use of cadmium-pigmented paints intended for the coating of toys [33]. Emission oftoxic cadmium compounds on combustion ofwaste or extraction in waste deposits has been estimated tobe unimportant in comparison to other sources. [34]. The regulations on the use of cadmium pigments vary in different countries as do the MAK-TLV Iimits given below. As no corresponding values for pigments are known these values should serve for orientation only: Fed. Rep. Germany England Sweden USA Japan MAK-value CdO (smoke) Inhalable fraction of total dust Total wt. ofCd-compounds soluble in 0,1 N HCl Total Cd lnhalable fraction of Cd TLV CdO (smoke) MAKCd 0,1 mg/m3 0,05 mgfm3 0,2 mg/m3 0,05 mgfm3 0,02 mgfm3 0,05 mg/m3 0,05 mg/m 3 Many countries insist on the "non migration principle" according to which no colorant should migrate visibly to the food or to the test solution. As cadmium pigments do not migrate at all, they can be used as colorants for plastics except in swellable ones like polyamide in contact with acids [34]. Inorganic Pigments 227 Silica, Silicates and Asbestos These powders are used as fillers and extenders rather than as colorants. Although not being toxic in a chemical sense, inhalation of such dust particles on handling, spraying of paints or on mechanical paint film removal should be avoided because of the risk of silicosis or asbestosis and associated lung cancer. The hazards depend on the particle size. Silica is most dangeraus with particle sizes between 0,5 and 5 J.lm [2, 44]. Smaller particles remain suspended in the streaming air and may leave the lung. Larger particles are usually filtered by some other protective mechanism and pass to the intestines where they are harmless. Various kinds of asbestos in fibrous form are specifically harmful to the lung tissues [45]. The most hazardous powder of this group is blue asbestos (crocidolite). Only asbestos fibers above 5 J.Lm in length are dangerous. Shorter ones remain suspended in the respired air [2]. As talc is similar to asbestos it has also become suspicious as being carcinogenic, however evidence is still controversial [46]. Although there is no indication that industrially produced highly disperse amorphaus silica causes silicosis even under extreme working conditions [47], the usual safety measures on handling industrial dusts should be obeyed when paint films containing them have tobe removed mechanically. In West Germany the maximum concentration ofinert fine dusts (MAK) is 8 mg/m3, which can also be considered as a limiting value in using the fillers mentioned above. In England asbestos regulations [48] have to be applied when the average atmospheric concentration of asbestos dust (other then crocidolite) is 2 fibresjcm3 or 0,2 fibres of crocidolitejcm3 • The American Conference of Governmental Industrial Hygienists (ACGIH) has related the TLV to the percentage of quartz to determine the TLV (VSHS Standard) for silica bearing dust by the formula (30 mg/m3) / (% quartz in dust + 2) [2]. The asbestos content in powders may be determined by X-ray diffraction or by a microscopic dispersion staining technique [49]. Miscellaneous Inorganic Colorants Some inorganic pigments for special purposes, containing antimony, arsenic or barium have been also discussed as hazardous compounds. Antimony As soluble antimony compounds are known to be toxic, antimony oxide has been suspected ofbeing hazardous. However, this oxide, which is used in high performance flame retardant paints, is only slightly soluble in water and in hydrochloric acid. No harmful effects could be detected with workers in a plant manufacturing Sb20 3 even on prolonged severe exposure [50, 51]. Possibly the only hazard that may arise is by inhalation of fumes from the burning or welding of painted surfaces [2]. Lead antimonate, also known as Naples 228 W. Funke Yellow, is soluble in acids.lt has been mainly used in artists paints.lts toxicity, which may be equally well be ascribed to the presence of the lead, is well known [3]. Arsenic Copper aceto-arsenate Cu(CH3C00)2 .3Cu(As02) 2, (Emerald or Paris green) is one of the earliest examples of dangerous pigments in paints [3]. However its use in artist paints is negligible today. Barium Soluble barium salts are highly toxic. The lowest toxic dose reported for humans is 80 mg BaC12/kg body weight [2]. Apart from barium chromate, which has been mentioned with the chromate pigments, barium metaborate is the only pigmenttobe considered in this connection [1]. This pigment is used on account ofits fungicidal and anticorrosive properties. As far as the limit of acid soluble barium content is concerned its use would be excluded in most paint specifications. References 1. O'Neill, L.A.: Hea1th and safety environmenta1 pollution and the paint industry- a survey covering 1egis1ation, standards, codes of practice and toxicology. England: Paint Research Association, Jan. 1977 2. Morrison, R.: Hazardous Paint Pigments. Australian OCCA, Proc. and News, Oct. 1975, 5 3. Mansell, H.: ICCM Bulletin, Pigment Toxicity. 3, No. 2, 11, June 1977 4. Dunn, M. J.: Paint and Vamish Production, Aug. 1973, pg. 49 5. Deutsche Norm, DIN 53 197, Nov. 1971 6. Zorll, U.: Dtsch. Farbenztschr. 11,495 (1976) 7. Chmielewski, J. et al.: Bull.-Inst. Mar. Med. Gdansk 25,43 (1974) 8. Inchingo1o, P. et al.: lndustr. Vem. 30, (8), 3 (1976) 9. Douglas, C.P., Plummer, R.M.: Protection 13, (4), 3 (1976) 10. Farbe+ Lack85, 597 (1979/7) 11. Brezinski, D.R.: Coatings Techno!. 48/4, 48 (1976) 12. Gage, J.C., Litchfield, M.H.: J. Oil Co!. Chem. Ass. 52, 236 (1969) 13. Konietzko, H., Elster, I., Reill, G.: Zbl. Arbeitsmed. 1978/6, 163 14. Niemann, E.: I-Lack, 46, 390 (1978/11) 15. Schatz, H.: Korrosionsschutz 1978/8, 13, (Ed. Verein Dtsch. Bleioxid Hersteller, Köln) 16. Niemann, E., ibid, March 1979 17. Umwelt-Bundes-Amt: Ber. 76/3, 116, West Germany 18. Schneider, W.F.: Am. Paint a. Coatings J., Convention Daily, 30. Oct. 1976, pg. 34 19. Mann, A.: Mod. Paint a. Coatings, Febr. 67/2,21 (1977) 20. Farbe + Lack 85, 598 (1979/7) 21. Gross, E., Kölsch, F.: Arch. Gewerbepath 12, 164 (1943) 22. Langärd, S., Norseth, T.: Brit. J. lndustr. Med. 32, 62 (1975) 23. Am. Paint a. Coatings J., 25. Oct. pg 9 (1976) 24. Engel, H.O., Calnan, C.D.: Brit. J. Industr. Med. 20, 192 (1963) 25. Nat. Paint & Coatings Ass., Safety & Health Bulletin, 1975, No. 27 26. Brit. Colour Makers Ass., Polymer Paint Co!. J., 166, 933 (1976) 27. Davies, J.M.: J. Oil Co!. Chem. Assos. 62, 157 (1979) Inorganic Pigments 28. 29. 30. 31. 32. 33. 34. 35. 36. 37. 38. 39. 40. 41. 42. 43. 44. 45. 46. 47. 48. 49. 50. 51. 229 Davies, J.M.: The Lancet, Febr. 18 (1978) Sperfeld, R.: Farbe + Lack 84, 137 (1978) Endriß, H.: Kunststoffe 69,403 (1979) Health and Safety Executive: Threshold Limit Values for 1975, USA. Techn. Data Note 2/75 Deutsche Forschungsgemeinschaft, Maximale Arbeitsplatzkonzentrationen, Mitt. XIII u. Mitt. XIV, 1977/78, West Germany Technical notes on cadmium, Cadmium Pigments. Cadmium Association, London, and Cadmium Council, New Y ork 1978 Endriß, H.: Kunststoffe 69, 39 (1979) Polymer Paint a. Colour J. June 13, 1979, 595 Chem. Ind., XXXI, (6), 369 (1979) Gabby, J.L.: Ind. Hyg. & Occup. Med. 1, 677 (1950) Krynskaya, I.L. et al.: Plast. Massy 1, 65 (1975) Streatfield, G.R.: Pigment Res. Techno!. 6, 18 (1976) Loeser, E., Lorke, D.: Toxicology 7, 215 and 225 (1977) Occup. Hyg. 17, 205 (1975) Pigmente- Toxikologie, Ullmanns Encyklop. techn. Chem. 13, 822 (1962) Umwelt-Bundes-Amt, Luftqualitätskriterien für Cadmium, Berichte 77/4, 149 Über das physiologische Verhalten von hochdispersen Oxiden des Siliciums, Aluminiums und Titans. Schriftenreihe DEGUSSA "Pigmente", 1977, No. 64, (9) Buckup, H.: Zbl. Arbeitsmedizin 16, 203 (1966) Pelfn!ne, A., Shubik, P.: Nouvelle Presse Med. 4, 301 (1975) Hofmann, W.: Gummi, Asbest, Kunststoff, 1974, 624 The Asbestos Regulations 1969. Statutory Instrument No. 690, H.M.S.O., Great Britain Julian, Y., McCrone, W.C.: Microscope 18, I (1970) Oliver, T.: Brit. Med. J., 1, 1094 (1933) Fairhall, L.T., Hyslop, F.: US Treasure Dpmt. Pub!. Health, Rep. 1947, Suppl. No. 195 Radioactive Substances G.C. Butler, C. Hyslop Division of Biological Seiences National Research Council ofCanada Ottawa, Canada KIA OR6 Glossary Radiation Protection Concepts ALl Annual Limit on Intake: the activity of a radionuclide which, taken alone, would irradiate a person, represented by Reference Man, to the Iimit set by the ICRP [7] Class D, W, Y (days), (weeks), (years): a classification scheme for inhaled material according to its rate of clearance from the pulmonary region of the Jung [7] DAC Derived air concentration: equals the ALl for inhalation (of a radionuclide) divided by the volume of air inhaled by Reference Man in a working year (i.e. 2.4 x I (}3 m3) (Bq m-3) [7] IL Investigation Level: a value of dose equivalent or intake above which the results are considered sufficiently important to justify further investigation ([3], par. 151, p. 29) Reference Man A person with the anatomical and physiological characteristics defined in the report ofthe ICRP Task Group on Reference Man [30] Nuclear Reactors AGR BWR CANDU GCR HTGR HWR LMFBR LWR NPD NFS PWR Advanced gas cooled reactor Boiling water reactor Canadian deuterium uranium reactor Gas cooled reactor High temperature gas cooled reactor Heavy water reactor Liquid meta! fast breeder reactor Light water reactor Nuclear Power Demonstration reactor Nuclear Fuel Services reactor Pressurized water reactor Organizations and Groups BEIR Advisory Committee on the Biological Effects of Ionizing Radiations, National Academy ofSciences-National Research Council (USA) BNWL Batteile Pacific Northwest Labaratory (USA) 232 EML HASL IAEA ICRP ICRU LASL MRC NCRP ORNL SCOPE USAEC UNSCEAR USNAS WASH-1400 G. C. Butler, C. Hys1op Environmenta1 Measurements Laboratory (USA) (formerly HASL) Health and Safety Laboratory (USAEC) International Atomic Energy Agency International Commission on Radiological Protection International Commission on Radiation Units and Measurements Los Alamos Scientific Laboratory (USA) Medical Research Council (UK) National Council on Radiation Protection and Measurements (USA) Oak Ridge National Laboratory (USA) Scientific Committee on Problems of the Environment of the International Council of Scientific Unions United States Atomic Energy Commission United Nations Scientific Committee on the Effects of Atomic Radiation United States National Academy ofSciences USAEC Reactor Safety Study (draft) 1974 Introduction The purpose of this chapter is to show how to assess the detriment resulting from the release of radioactive materials to the environment. Because of the wide range of the subject and the Iimitation of space the chapter consists of little more than a listing of principles and concepts. A more adequate examination of these will require consulting the Iiterature cited. The minimum information required for the assessments is given for seven radionuclides of interest from the point of view of environmental contamination. Basic Concepts Radiation Doses and Units Recently new units in the International System of Units (SI) have been introduced to quantify ionizing radiation [1, 2]. They are given below along with the older units they replace. Exposure: Activity: Absorbed dose: 1 roentgen (R) = 2.58 x 104 coulombs per kilo- gram of air (C kg-1) I becquerel (Bq) (new unit) = 1 radioactive transformation per second (tr s-1) = 2.7 x lü-11 curies (old unit) = 3.7 x 1010 radioactive transI curie (Ci) formations per second (tr s-1) 1 gray (Gy) (new unit) = 1 joule perkilogram (J kg-1) = 100 rads (old unit) = lQ-2 joules perkilogram (J kg-1) 1 rad Radioactive Substarrces Dose equivalent: 233 sievert (Sv) (new unit) = 1 gray x quality factor (Q) (J kg-1) = 1 rem 100 rems (old unit) = 1 rad x quality factor (Q). Effects of Radiation and Dose-Effect Functions Radiation doses as low as those usually encountered in the environment result in "stochastic" detrimental effects ([3], Sect. 7, p. 2). These comprise malignant and hereditary diseases for which the probability of occurrence, rather than the severity, is proportional to the dose (e.g. cancers, lethal mutations). For these effects it is assumed that there is a "linear non-threshold" doseeffect relationship ([3], Sect. 27, p. 6; [4, 5]; [6], Sect. 36, p. 366; Sect. 143, p. 592). This means that all doses greater than zero received during a lifetime contribute, according to their magnitude, to causing biological effects. Dose Equivalent (H) All radiations do not have the same effectiveness, gray for gray, in producing stochastic effects, thus the concept of dose equivalent has been introduced and defined as follows [1]: H= DQN where H = the dose equivalent at a point in tissue, expressed in sieverts D = the absorbed dose, expressed in grays Q = a quality factor dependent on density of ionization in tissue, produced by the radiation N = the product of any other modifying factors such as rate of irradiation. Mean values of Q adopted for the purposes of radiation protection are ([3], Sect. 20, p. 4): Type of radiation Q X-rays, y-rays, electrons 1 thermal neutrons 2.3 fission neutrons and protons 10 a-particles and other multiply-charged particles 20 Committed Dose Equivalent (Hso) The total dose equivalent accumulated by a given organ or tissue during an individual's working lifetime of 50 yr, from a single bodily intake of radioactive material, is called the committed dose equivalent. It is defined as follows ([3], Sect. 26, p. 6): Hso = f t(l 'o + 50y H(t) dt, G. C. Butler, C. Hyslop 234 H 50 = the 50-year committed dose equivalent H(t) = the dose equivalent rate at time t t0 the time of intake. The calculations of committed dose equivalent from a single intake are described in ICRP Publication 30 [7] and from multiple intakes in ICRP Publication lOA [8]. where Dose-Equivalent Commitment (llc) The dose-equivalent commitment to an individual (He), resulting from a given decision or practice, istheinfinite time integral of the per caput dose-equivalent rate (H(t)) in a given organ or tissue for a specified population ([3], Sect. 25, p. 6; [6], Sect. 16, p. 28). It may be expressed as: He= J =H(t) dt, 0 Table 1. Global dose equivalent commitments from various radiation sources. (From Table 3 of UNSCEAR ([6], p. 16)) Source of exposure Annual absorbed dose (man-Gy) Natural Irradiation a) One-year exposure to natural sources 2Xl0 6 Natural Irradiation Enhanced by Technology b) One year of commercial air travel 3Xl03 c) U se of one year's production of phosphate fertilizers at the present production rate Annual dose equivalent (mSv) Globaldose equivalent commitment (days)" 1 365 0.4 102 0.04 d) One-year global production of electric energy by coal-fired power plants at the present global installed capacity [10 6 MW(e)] 50 0.02 e) Mining Ca-irradiation oflungs) 50 0.02 Man-Made Sources of Radiation f) One-year exposure to radiation-ernitting consumer products 3 g) One-year production of nuclear power at the present global installed capacity [8X104 MW(e)] h) One year of nuclear explosions averaged over the period 1951-1976 i) One year's use of radiation in medical practice 0.6 0.07 30 70 • The global dose comrnitment is expressed as the duration of exposure of the world population to natural radiation which would cause the same dose commitment. The occupational contribution is included b In the most technologically developed countfies [9] 235 Radioactive Substances Table 1 ([6], p. 16; [9]) showsglobal dose-equivalent commitments from vanous sources. Risk Estimates Estimates of the risk of biological effects of ionizing radiation have been published by UNSCEAR [6, 10, 11], ICRP [12, 13] andin the "BEIR Report" [14]. The most recent estimates of genetic risks are given by UNSCEAR ([6], pp. 425-564) and compared with USNAS va1ues [14] in Table 2 ([6], p. 539). The risks ofmalignancies published by UNSCEAR [6] and ICRP [3] are shown in Table 3. The rates of incidence of these malignancies in Canada in 1975 are given in Table 4 [15]. Effective Dose Equivalent (IIE) To estimate the total harm from an intake of radionuclides it is necessary to know the annual dose equivalents to specific high risk tissues and to multiply these by weighting factors proportional to the risks of stochastic effects ([3], Sect. 104, p. 21; [16]). The sum of all these weighted dose equivalents is called the effective dose equivalent (HE) and is described algebraically as: HE= :EHTwT T the annual dose equivalent for tissue T the weighting factor representing the ratio of the stochastic risk arising from tissue T to the total risk when the whole body is irradiated uniformly. The values ofwT for the tissues at greatest risk, assigned by ICRP ([3], Sect. 105, p. 21) on the basis ofthe risks listed in Table 3, are listed in Table 5. The effective committed dose equivalent {H 50E) is defined by the equation [7]: where HT WT HsOE = :ET HsOT X wT where HsoT = the committed dose equivalent for tissue T. Collective Dose Equivalent The detriment to a population resulting from ionizing radiation may be proportional to the collective dose equiva1ent (S) defined by the equation ([3], Sect. 22, p. 5) S =:EHxP. . I where I I Hi = the per caput dose equivalent to the whole body or an organ or tissue in sub-group i of the exposed popu1ation Pi = the number of people in the sub-group i of the exposed population. 236 G. C. Butler, C. Hyslop Table 2. Estimated effect ono- 2 Gy (I rad) per generation oflow-dose, low dose-rate, low-LET irradiation on a population of one rnillion Iive-born individuals. Assumed doubling dose, 1Gy(100 rad) (Table 50 ofUNSCEAR [16], p. 539) Effect ofl0- 2 Gy (1 rad) per generation Disease classification• Current incidenceb First generationc Equilibrium Autosomal dominant and X-1inked diseases Recessive diseases lO,OOOd 1,100 20 Relatively slight 38[ 100 Very slow increase 40 4,ooo• Chromosomal diseases Congenital anomalies Anomalies expressed 1ater Constitutional and degenerative diseases 90,0008 Total Percentage of current incidence 105,200 5h 45h 63 0.06 185 0.17 20 Relatively slight 100 Very slow increase Recalculated BEIR assessments Autosomal dominant and X-linked diseases Recessive and chromosomal diseases Congenital anomalies Anomalies expressed later Constitutional and degenerative diseases Total Percentage of current incidence I 10,000 10,000 40,000 60,000 100 2-20i 25-40j 0.04-0.07 20-200i 125-300i 0.21-0.50 • Follows that given in the BEIR Report [14] Basedon the results ofthe British Columbia Survey with certain modifications; see Table 9 in [6], p. 519 c The first generation incidence is assumed to be ab out one fifth of the equilibrium incidence for autosomal dominant and X-linked diseases; for those included under the heading "congenital anomalies etc." it is one tenth ofthe equilibrium incidence. For rationale see [14) ct See Table 9 in [6), p. 519 • Based on the pooled values cited in Nielsen and Sillesen (363 in [6), p. 553) includes mosaics but excludes balanced translocations r The first generation incidence is assumed to include all the numerical anomalies and three fifths of the unbalanced trans1ocations (the remairring two fifths being derived from a balanced trans1ocation in one parent) 8 lncludes an unknown proportion of numerical (other than Down' s syndrome) and structural chromosomal anomalies h Based on the assumption of a 5% mutational component ; The range reflects the assumption of 5 and 50% mutational components; see [6] for explanation i Rounded-offfigures b The collective dose equivalent (Sk) resulting from a practice or source (k) is defined by the expression ([3], par. 23, p. 5) Sk = foooH X P(H)dH Radioactive Substauces 237 Table 3. Estimated effects of one unit of low-dose, low dose-rate irradiation on a population of one million persons Risk Tissue Effect UNSCEAR (per 10-2 Gy) ICRP (per Sv) 1. Gorrads 2.Body 3. Breast 4. Red hone marrow 5.Lung 6. Thyroid 7. Bone surfaces 8. Remainder (2-3, 4, 5, 6, 7) Mutations Allcancers Fatal cancer Leukemia Fatal cancer Fatal cancer Fatal cancer Fatal cancer 63 (~) 200 50 (population) 20-50 25-50 10 2-5 35-93 10,000 (~+f2) 2,500 (workers) 2,000 2,000 500 500 5,000 For the qualifications concerning these numerical estimates UNSCEAR [6] and ICRP [3] should be consulted Table 4. Rate of reporting of malignant neoplasms in Canada, 1975 [15] Incidence per million Tissue or effect Cases Deaths Total body Breast Leukemia Lung Thyroid Bone 1,900 350 60 300 20 8 1,500 140 60 300 0 8 Table 5. Values ofwT recomrnended by ICRP [3] Tissue l1'r Gorrads Breast Red hone marrow Lung Thyroid Bone surfaces Remainder 0.25 0.15 0.12 0.12 0.03 0.03 0.30 Total where H P(H) = 1.00 the dose equivalent received the number of individuals receiving a dose equivalent in the range from H to H + dH. 238 G. C. Butler, C. Hyslop Collective Dose Commitment (SO To assess the dose equivalents received by a population and the resulting total detriment, from single exposures to long-lived radionuclides or repeated exposures to short- or long-lived ones, UNSCEAR ([6], Sect. 15, p. 29) has developed the concept of collective dose commitment. The collective dose commitment (SO due to a given event, decision, or finite practice k is defined as: where sk = the collective dose rate from source k. In the case where releases of relatively short-lived radionuclides continue long enough for concentrations in environmental compartments to become constant, the collective dose equivalent resulting from one year of a practice is equal to the collective dose commitment ofthe amount released in one year ([17], pp. 102, 107; [18]). Detriment and Dose Limits One of the bases of the ICRP system of radiological protection is that any human activity should produce more benefit than detriment ([3], Sect. 69, p. 14). Detriment in a population may be defined as ([3], Sect. 16, p. 3): " ... the mathematical 'expectation' of the harm incurred from an exposure to radiation, taking into account not only the probability of each type of deleterious effect, but also the severity of the effect." The most recent recommendations of ICRP on dose Iimitation ([3], Sect. 104, p. 21) are based on the princip1e that, for stochastic effects, the risks resulting from the Iimit of dose should be equa1 for uniform and non-uniform irradiation of the body and its tissues. The annua1limit recommended by ICRP for HE and HsoE for workers is 50 mSv (5 rem). Two other occupationa1 dose Iimits, for non-stochastic effects ([3], Sect. 103, p. 21), are 0.3 Sv (30 rem) for the 1ens ofthe eye and 0.5 Sv (50 rem) for any other tissues; these are reported for completeness only since they would not likely be relevant to an environmental situation. The ICRP ([3], Sect. 119, p. 23) recommends a dose-equivalent Iimit of 5 mSv (0.5 rem) per year for critical groups1 or individual members ofthe public. Transfer to Man The pathways by which man is irradiated as a result of the presence of radioactive materials in the environment are complex and differ depending on I A critical group has been described by ICRP ([3], Sect. 85, p. 17) as a group within the population small enough to be relatively homogeneous, yet representative of those individuals in the population expected to receive the highest dose equivalents Radioactive Substauces 239 whether the radioactivity is airborne or waterborne. The pathways have been described diagrammatically by ICRP Committee 4 [19]; their diagrams are reproduced as Figs. 1 and 2. UNSCEAR ([6], pp. 27-34) has described in a Direct irradiation Deposition Ingestion Deposition Direct radiation Inhalation Inhalation Fig. 1. Simplified pathways between radioactive materials released to atmosphere and man [19] Direct irradiation Ingestion Ingestion Indirect irradiation Fig. 2. Simplified pathways between radioactive materials released to ground or surface waters (including oceans) and man [19] 240 G. C. Butler, C. Hyslop general way the concept of transport through these environmental compartments, and the resulting tissue and organ doses. The transfer factor from compartment i to compartmentj has been defined by UNSCEAR as ([6], Sect. 29, p. 31): 1: cj(t) dt l:ci(t)dt f·cj( t) dt =::: 1=ci(t)dt where C and Cj are the quantities (e.g. activity concentrations) in the respective compartments at timet. Under conditions of constant release and under constant environmental conditions concentrations in the compartments may become constant, when c p .. =...!::l IJ Ci where C and Cj are the constant concentrations in compartments i and j ([18], p. 15). Exposures ofNon-Human Biota The ICRP ([3], Sect. 14, p. 3) " ... believes that ifman is adequately protected then other living things are also likely to be sufficiently protected." IAEA points out that, for humans, great importance is placed on the long-term effects on individual members of a population whereas for other organisms the long-term structure and fate ofthe populations are the main concern [20]. a) Doses received. For large groups the average dose and doserateswill nearly always be less than those due to natural sources, viz., 1 mSv (100 mrem)fy. At such dose Ievels only stochastic effects or late cumulative effects of low dose rates will be involved. Some organisms may receive larger-than-average doses of direct radiation because of their location. Examples of non-human exposure that will be mentioned later in the chapter are: i) terrestrial exposure - contamination ofplants by fallout, e.g.lichens (137Cs) ([11], Vol. 1, pp. 52-53) - plants and animals living in certain regions oflndia and Brazil ([6], pp. 48-49) - plants and animals living around reactors [21] ii) aquatic exposure - organisms inhabiting bottom sediments accumulate relatively high levels of plutonium [22, 23] - organisms near the outfall of nuclear effiuents [24, 25] For indirect radiation, larger doses may be received because of some metabolic factor or a special niche in a food chain: - caribou eating Iichens contaminated with 137Cs ([11], Vol. 1, pp. 52-53) - domestic animals eating grass contaminated with radioiodine. At Wind- Radioactive Substauces 241 scale 131 I contents of cows' and sheep's thyroids were measured following the accidental release in 1958 and the highest total radiation dose to the thyroid gland was around 10 Gy (1,000 rads) ([26], p. 136). - ifthe radionuclide in question is readily absorbed and has a long half-life of retention it will accumulate in higher limnologicallevels such as piscivorous fish [27]. - animals feeding directly off bottom sediments of lakes and rivers, such as molluscs, usually contain high levels of radionuclides such as plutonium [22]. b) Radiosensitivity. Although lethal doses are not encountered in the environment the radiosensitivity of species may be compared in terms of LD 50; 30 (lethal dose for 50% of organisms in 30 days). Mammals are generally more radiosensitive than other vertebrates, including birds, reptiles, amphibians and fish. LD 50; 30 for dogs is about 3.4 Sv (335 rem) (X-rays) while for goldfish it is 6.7 Sv (670 rem) (X-rays) ([28], pp. 299-310). The dose ofX- or y-rays needed to kill an insect is at least 100 x greater than that needed to kill a mammal. Adult Drosophila are not killed by 64,000 R from 6°Co y-rays, but are sterilized. Unicellular organisms may be less sensitive yet. LD 50; 30 for Amoeba is 1,000 Sv (100,000 rem) (X-rays) ([28], pp. 299-310). In aquatic systems, teleost fish (especially developing eggs) are the organisms most sensitive to radiation [20, 29]. The genetic character of a species or strain is a major determinant of the carcinogenic response to radiation exposure ([6], Sect. 334, p. 622). Differences in susceptibility are especially manifest at low doses and tend to disappear with increasing doses and dose rates. In experimental animals such as mice, a dose of at least 50 rad is generally required to detect an increase over the natural tumor incidence ([6], Sect. 328, p. 621). Mutationratesper locus per Gy for low-LET irradiation are in the range of I0-5 to I0-7 for organisms as diverse as mice, Drosophila and barley ([6], Table 44, p. 535). The dose of radiation needed to double the natural mutation rate when given in a single dose (doubling dose) is about 0.3 Gy (30 rads) in mice, 0.5--4 Gy (50--400 rads) in Drosophila and 0.3-0.6 Gy (30-60 rads) in plants ([28], p. 258). Studies of the effects of irradiation in fetal rodents consistently show a reduction in sensitivity with advancement of fetal age ([6], Sect. 342, p. 709; [28], pp. 299-31 0). Insect larvae also become less radiosensitive with age. F or Drosophila eggs 3 hold the LD 50 is 200 R, for 4-h eggs it is 500 Rand for pupae, 2800 R ([28], pp. 299-310). Selected Radionuclides Introduction In this section seven radionuclides have received detailed discussion. The choice was made because of their practical importance, public interest or G. C. Butler, C. Hyslop 242 suitability for illustration. As far as possible, data are given which permit the calculation of the risks to human health resulting from a unit of practice. The reviews depend heavily on the most recent publication of UNSCEAR [6] where, on p. 116, the elements of the assessment are illustrated as Inhalation Input (0) -+ Atrnosphere (ll) -+ Earth's surface (I) . -+ Diet (3) -+ Tissue (4) -+ Dose (5) Extemal irradiation UNSCEAR has its own methods for calculating, from the rate of intake and the equilibrium body content, the resulting tissue concentrations and dose rates (in grays) to the tissues. In the present reviews, the UNSCEAR data are used to calculate intakes and thereafter when the resulting dose equivalents to tissues (in sieverts) are calculated, the data of ICRP Committee 2 are used. According to ICRP Publication 30 [7] the Annual Limits on Intake (by either inhalation or ingestion) give to all the tissues of the body an effective dose equivalent of 50 mSv or to a single tissue a dose equivalent of 500 mSv, whichever is the lesser intake. From these limiting intakes can be calculated the dose equivalent or the effective dose equivalent resulting from unit intake. Tritium Oxide Exposure Due to Natural Sources a) Production and Release. Tritium occurs naturally, principally in the atmosphere where it is produced by cosmic ray protons and neutrons reacting with nitrogen, oxygen and argon. The reaction producing most of the tritium is 14N + n ... 12C + 3Hl where the energy ofthe neutrons is >4.4 MeV ([6], Sect. 82, p. 54). The most recent estimates of production rate of 3ß and corresponding world inventory are 0.20 atoms per cm2 of earth's surface per second and 1 x 10t8 Bq (30 MCi), respectively ([6], Table 11, p. 55). More than 99% of the tritium produced either by natural processes or human technology, when released to the environment, appears as tritiated water (HTO) and hereafter the tritium discussed will be assumed tobe in that form, unless specified otherwise. b) Pathways to Man. As mentioned above, most ofthe HTO produced in nature is found in surface waters of the earth. Concentrations of HTO in continental surface waters before nuclear explosions began were 0.2-0.9 Bq (6-24 pCi)/L ([6], Sect. 84, p. 55) and, assuming that the hydrogen ofthe body of Reference Man had the same proportion of tritium as had the surface waters, this would give a whole body dose of 1 x I0-8 Gy (1 J.Lrad)/yr. (Accord- Radioactive Substances 243 ing to UNSCEAR ([6], Sect. 19, p. 118), 3.7 x 104 Bq (I J.lCi)/L gives 9.5 x 10-4 Gy (95 mrad)/yr.) Exposure Due to Man-Made Sources Tritium arises from temary fission in nuclear explosives or nuclear fuel and also by neutron activation reactions with isotopes of light elements such as Iithium and boron. Whenever water (which contains deuterium) is irradiated with a high flux of neutrons, tritium is produced according to the reaction Thus the chief sources of tritium production by man will be nuclear bomb explosions and nuclear reactors. 1. Exposure Due to Nuclear Bombs a) Production and Release. UNSCEAR ([6], p. 117) has summarized the estimates of total tritium production in nuclear bomb explosions and the resulting world inventory. Since the total production is released the quantities given will serve for estimates of the release. The best estimates for the total release up to 1970 lie between 1.3 x 1020 and 1.7 x 1020 Bq (3,500 and 4,500 MCi), with 20% ofthis in the southem hemisphere and the remaining 80% in the northem hemisphere. b) Deposition. The HTO released from above-ground nuclear explosions is injected into the stratosphere where the average residence time is about one year. It then passes to the troposphere and atmosphere and enters the earth's hydrological cycle. According to UNSCEAR ([6], Sect. 17 and Fig. I, p. 117), between 1963 and 1969, 6.6 x 1019 Bq (1,780 MCi) of HTO were deposited in the northem hemisphere and 1.5 x 1019 Bq (400 MCi) in the southem hemisphere. The latitudinal distribution ofHTO in the top 500 m ofthe Pacific Ocean between 1965 and 1972 was about 4.5 x 1019 atoms per km2 at 30° south latitude and 32 x 1019 atoms per km2 at 30°-40° north. From the data acquired by the IAEA world network for monitoring HTO in precipitation ([6], p. 118) it was calculated that, at marine stations, the concentration ofHTO in raindoubledas the latitude increased by 13° and that the concentration was 3.6 times as high over landasover water. The mean concentrations in surface waters ofthe USA varied from a low of about 0.6 Bq (15 pCi)/L in 1951-1953 to a high ofabout 185 Bq (5,000 pCi)/L in the 1960's. Concentrations in the Ottawa River (approx. 46° north latitude) were about twice as great. c) Pathways to Man. The dose commitments from nuclear explosions have been calculated by UNSCEAR ([6], Sect. 18-27, pp. 118-119) tobe 2 x 10-5 Gy (2 mrad) for the northem hemisphere and 2 x 10--{j Gy (0.2 mrad) for the G. C. Butler, C. Hyslop 244 southem hernisphere. For the population of the USA the dose commitment was calculated to be 1.5 x 10-5 Gy (1.5 mrad) and, for inhabitants of the Ottawa Valley, 2.8 x 10-5 Gy (2.8 mrad). The collective dose commitment for explosions in the northem hemisphere is estimated to be 8 x 104 man-Gy (8 x 106 man-rad), corresponding to 8.1 x 1o-16 man-Gy per Bq (3 x 1o-3 man-rad per Ci) of HTO released. 2. Exposure Due to Nuclear Reactor Operations a) Production. The rates of production given by UNSCEAR for various types ofreactors are ([6], Table 10, p. 178): BWR, PWR, GCR HWR 7.4 x 1011 Bq (20 Ci) per MW(e)y 2.2 x 1013 Bq (600 Ci) per MW(e)y. The reference LWR fuel described in [6], (Table 25, p. 202), irradiated to 33,000 MWd per tonne, and cooled 150 days, contained 7.9 x 1015 Bq (213 kCi) oftritium per tonne which corresponds to a production rate of2.6 x 1014 Bq (7.1 kCi) per MW(e)y. In HWR ofthe CANDU type, operated by Ontario Hydro, the production rate is 8.9 x 1013 Bq (2,400 Ci) per MW(e)y [31]. b) Release and Deposition. The releases of HTO from various types of reactors reported by UNSCEAR ([6], Table 10, p. 178; Table 11, p. 179) are summarized in Table 6. It can be seen that by far the greatest releases come from HWR and this merits some comment. The amounts released depend on the amounts leaking or escaping from the system during normal operations and rninor accidents and on the amount of neutron irradiation received by the heavy water moderator and coolant. These factors will vary from reactor to reactor due to differences in design and operating experience. Ontario Hydro has the greatest body of experience in operating heavy water power reactors andin 1978 [31] they reported the following release rates: Installation Release rate (Bq per MW( e)y] NPD Douglas Point Pickering 2.7 X 1013 (730 Ci) 7 X 1012 (190 Ci) 8.9 X 1011 (24 Ci) They have estimated that 20% of current releases are to liquid effiuents and 80% to airbome effiuents. The Pickering generating station of Ontario Hydro is located on Lake Ontario; for this site it is assumed [31] that all the releases to liquid effiuents, and that half the airbome releases, enter the lake. 245 Radioactive Substarrces Table 6. Releases of tritium from reactor operations Normalized release [Bq per MW(e)y] Type of reactor In liquid effiuent In airborne effiuent (1974) BWR PWR GCR HWR (Pickering) 7.4 X 109 (0.2 Ci) 3.7 X 10 10 (1 Ci) 1.1 X 10 10 (0.3 Ci) 7.4 X 10ll (20 Ci) 1.9 X 109 (0.05 Ci) 7.4 X 10 9 (0.2 Ci) 1.5 X 109 (0.04 Ci) 5.6 X 10 11 (15 Ci) c) Pathways to Man. UNSCEAR ([6], Sect. 103-104, pp. 193-194) estimates that for airborne HTO released from an operating reactor the collective dose commitment is 5.4 x 10-17 man-Gy per Bq (2 x 10-4 man-rad per Ci). Thus, for the releases quoted above, the different types ofreactors would give the following collective dose commitments: Reactor type Collective dose commitment [man-Gy/MW(e)y] PWR GCR BWR HWR 4 X 10-7 (4 X 10-5 man-rad) 4 X 10-7 (4 X 10-5 man-rad) 1 X 10-7 (1 X 10-5 man-rad) 4 X 10-5 (4 X 10-3 man-rad) For HTO released in liquid effluents to a body ofwater providing drinking water, the calculated collective dose commitment is 1.9 x 10-15 man-Gy per Bq (0.007 man-rad per Ci) released, resulting in the following ([6], Sect. 105, p. 194): Reactor type PWR BWR HWR Collective dose commitment [man-Gy/MW(e)y] 7 X 10-5 (0.007 man-rad) 7 X 10-6 (0.0007 man-rad) 3 X 10-4 (0.03 man-rad) On the basis of operating experience, including environmental monitoring, Ontario Hydro has estimated the individual doses resulting from both airborne and waterborne releases at a typical station with CANDU-type reactors. The results of their calculations [31] are reproduced in Table 7, below. Since the major, and unavoidable, part of the population dose resulting from a release of HTO arises through inhalation and skin absorption of airborne activity, this could form the basis of calculations. The most important information required is the dilution factor, Ka = air concentration at the target (Bq/m3) releaserate (Bq/s) G. C. Butler, C. Hyslop 246 for the site in question. Ontario Hydro [31] has measured values of Ka at 1 km from four of their reactor stations and found it to average 2.8 ± 1.3 x 1Q-7 s/m3 • lt can be assumed that, beyond 1 km, the air concentration diminishes as the distance (in km) raised to the power -1.5, or, Cd=Gd-1.5, where Cd is the air concentration at d km, C1 is the air concentration at 1 km, d is the distance downwind from the source in km. Table 7. Estimated annual dose equivalents from HTO released to air and water from a CANDU station Annual individual dose equivalent (IJ.Sv) as a function of distance from the station Pathway 1km 3km 5km 10km 1. Inhalation and skin absorption 2. Ingestion of milk" 3.1ngestion ofhome grown fruits and vegetables 4. Drinking water 5 1 1 1 1 0.2 0.6 1 0.1 0.2 1 0.09 0.6 0.6 0.6 0.6 8 3 2 2 Total a Milk is from farms at a distance ofl0-15 km from Station With a knowledge of the air concentration, the population density and the ICRP dose coefficient [7], viz. 2,000 h exposure to a concentration of 8 x 105 Bqfm3 (DAC) gives an effective dose equivalent of 50 mSv, the collective dose commitment from an accidental release may be calculated. 3. Exposure Due to Fuel Reprocessing a) Production and Release. As reported by UNSCEAR ([6], Table 24, p. 201) operating experience in fuel reprocessing at two plants has given the following results: Tritiumrelease rate [Bq/MW(e)y] Installation Airborne Liquid 4.8 X 1011 Wmdscale (UK.) (13 Ci) NFS(USA) 3.7 X 1010 (1 Ci) 2.2 X 1011 (6 Ci) Grathwohl from Karlsruhe, quoted in [6] (Sect. 146, p. 203), estimated that, for a PWR, the production rate oftritium would be 7 x 1011 Bq (19 Ci)/ MW(e)y and of this about 3.7 x 1010 Bq (1 Ci) would be released during 247 Radioactive Substauces reactor operation and 5.9 x 1011 (16 Ci)/MW(e)y released during fuel reprocessing. Similar amounts were estimated for HWR fuel, one-half of this amount for AGR fuel and one-tenth for HTGR fuel. b) Pathways to Man. Using the same coefficients for collective dose commitment as for reactor operation (man-Gy per Bq released), the following collective dose commitments were calculated by UNSCEAR ([6], Sect. 159, p. 204): Installation Type of release Collective dose commitment [man-Gy/MW(e)y] NFS (USA) Wmdscale (UK) NFS(USA) airborne to salt water to fresh water 2 X 10-6 (2 X 10-4 man-rad) 2 X 10-8 (2 X 10-6 man-rad) 4 X 10-4 (4 X 10-2 man-rad) 4. Exposure Due to Occupation Special mention needs to be made of the occcupational exposure to two groups ofworkers: (a) staff ofHWR, and (b) tritium luminizers. a) Staff of HWR. The most informative statistics on this subject are provided by Ontario Hydro ([6], Table 11, p. 238) from more than 10 yr experience with operative CANDU type reactors. The mean collective occupational dosewas 9 x 10-3 man-Gy (0.9 man-rad)/MW(e)y, similar tothat for other types of power reactors in the USA, where most of the exposure is to external radiation. The exceptional feature of the CANDU statistics is that 26% of the collective dosewas due to internal contamination with HTO. b) Luminizers. Radium has been largely replaced by promethium and tritium for luminizing the dials of watches, although in the USAradium is still much in use for clocks, according to UNSCEAR ([6], Sect. 277 and Table 46, p. 96). These new luminous paints emit only soft ß-particles and thus give smaller doses to the wearer than radium which emits y-rays and also generates radon which leaks out. Some luminous paints contain tritiated organic compounds which may leak out slowly giving internal doses to the wearer of a watch. UNSCEAR ([6], Sect. 286, p. 97) reported a study of the HTO content of the urine of eight persons wearing tritium-luminized watches; the content averaged 1.2 x 102 Bq (3.2 nCi)/L above background which corresponds to a whole-body dose of 3 x 10~ Gy (0.3 mrad)/yr. Workers with tritium luminous paint may be monitored by measuring the HTO content of urine. The results from four countries in 1975 ([6], Sect. 122-123, p. 255; Tables 81-84, pp. 289-290) are as follows: G. C. Butler, C. Hyslop 248 Country No. ofworkers Average annual dose equivalent UK Switzerland France Germany 136 235 80 56 7 x 10-3 Sv (0. 7 rem) 1 x 10-2 Sv (1 rem) 3.5 X 10-3 Sv (0.35 rem) 1.35 x 10-2 Sv (1.35 rem) 5. Annual Limits on Intake ICRP Committee 2 [7] has calculated that the following intakes will give an effective dose equivalent, HE, of 50 mSv: Ingestion Inhalation (Class D) 3 x 109 Bq (50 mCi) 3 x J09 Bq (50 mCi) Krypton-85 Exposure Due to Man-Made Sources 1. Exposure Due to Nuclear Bombs a) Production and Release. The 85 Krj9°Sr ratio of fission yield=0.07; thus about 1.1 x 1017 Bq (3 MCi) have been produced in nuclear explosions ([6], Sect. 41, p. 121). Another estimate [32] has given 2 x 1017 Bq (5 MCi). b) Pathways to Man. Radioactive krypton, being a noble gas, is not deposited and does not enter into metabolic processes in the food chain nor in man. Thus, population doses are calculated by mu1tiplying the air concentration by a coefficient (one for each tissue of interest) to give the corresponding dose rate. To calculate population doses, one needs to know only the concentration in the air surrounding the population. UNSCEAR ([6], Sect. 158, p. 204; Sect. 191, p. 209) and NCRP [32] give the following doserate coefficients for an air concentration of 3.7 x 1010 Bq (1 Ci)/m3 : Organ or tissue Dose rate (Gy/yr) Testes Ovaries Red hone marrow Skin Lung Total body 60 (0.6 X 104 rad) 160 (1.6 x 104 rad) 180 (1.8 x 104 rad) 18,000 (1.8 x 106 rad) 310 (3.1 x 104 rad) 150 (1.5 x 104 rad) W ASH -1400 ([33], Table Vl-17, p. 60) gives a value of 1.1 x 104 Sv (1.1 x 106 rem)/yr (3.6 x I0-4 Sv (0.036 rem)/s) from an air concentration of 3. 7 x 1010 Bq/m3 but the tissue receiving the dose is not specified. 249 Radioactive Substauces Assuming that the 1.1 x 1017 Bq (3 MCi) released are uniformly mixed in the earth's troposphere (5 x 1021 g of air) the resulting concentration at NTP=3 x I0-2 Bq (0.8 pCi)/m3 • Assuming that the 85Kr concentration declines with the radioactive half-life of 10.7 years, the average life is 15 yr and the exposure is 0.4 Bq-yr (12 pCi-yr)/m3 • When this is multiplied by the dose rate coefficients given above the following individual dose commitments and collective dose commitments result ([6], Sect. 42, p. 121): Dose commitments Individual (nGy) Collective• (man-Gy) Gonads 1.4 (0.14 iJiad) 7.7 (770 man-rad) Red hone marrow 2.2 (0.22 iJiad) 12.6 (1260 man-rad) Skin 220 (22 iJiad) 1.26 X 103 (1.26 x 105 man-rad) Lungs 3.7 (0.37 iJiad) 21.7 (2170 man-rad) Organ or tissue • Based on a present world population of 4 X 109 that increases by 2% per year 2. Exposure Due to Nuclear Reactor Operations a) Production. 85Kr is only one ofseveral radioactive noble gases produced in reactor operation; a number of isotopes of krypton and xenon are produced in fission and 41 Ar is a neutron activation product ofthe argon in air ([6], Sect. 50, p. 172). About 25 cm3 of Kr and Xe are produced in reactor fuel per MWd thermal. This creates pressure inside the fuel canister and any cladding failure results in an escape ofthe gas ([6], Sect. 51, p. 172). The thermal fission yield of 85 Kr is 0.29% for 235 U and 0.14% for 239 Pu, corresponding to 1.9 x 1013 and 9.3 x 1012 Bq (500 and 250 Ci)/MW(e)y for the two fuels respectively. More detailed calculations give estimates falling between these two rates ([6], Sect. 144, p. 203; [32]). The amount of 85Kr in the LWR "reference" fuel of UNSCEAR is 1.4 x 10 13 Bq (375 Ci)/MW(e)y ([6], Table 25, p. 202). b) Release. The amount ofradioactive noble gases escaping will depend on the number of fuel cladding failures, the design of the cooling and ventilating systems and operating procedures. Thus there are tremendous individual variations contributing to the overall normalized releases given by NCRP and UNSCEAR ([6], pp. 172-178; [32]) as follows: G. C. Butler, C. Hyslop 250 Type of reactor %oftotal Normalized release [Bq 85Kr per MW(e)y] noblegases PWR BWR 6.3 1.1 GCR AGRandHTGR Insignificant Negligible X X 109 (0.17 Ci) 1012 (30 Ci) 1 2 c) Pathways to Man. Zuker et al. [34] used trajectory analysis based on historic wind data to calculate the 100-day-average ground concentrations of 85 Kr released at a constant rate from the Ontario Hydro Pickering Station. These ground-level concentrations were calculated for every point on a 70 km grid from Toronto to the east coast of North America. Population density figures, obtained from electoral districts and county censuses, were applied to the same grid. Multiplying the concentration at a square on the grid by the population and one of the dose conversion coefficients given above gave the annual collective dose for each square on the grid. When expressed as a function of distance from the source it was found that, for a constant release of3.7 x 10 10 Bq/s, the total annual collective dosewas received within a radius of 1000 km and that more than 95% ofthiswas within 600 km. UNSCEAR ([6], pp. 191-193) reported the collective doses due to radioactive noble gases released from various reactor sites. The collective dose depends on the population density around the reactor site so there will be great variation from site to site. The following table summarizes the dose commitments from the normalized releases reported by UNSCEAR: Reactor type BWR PWR Collective dose [man-Gy (gonad) per Bq released] Collective dose [total man-Gy (whole body) perMW(e)y] 1.2 x 10-16 (4.5 x 10-4 man-rad/Ci) 4.1 x 10- 17 (1.5 X 10-4 man-rad/Ci) 9 x 10-4 (0.09 man-rad) 0.4 2.5 x 10-5 -5 x 10-5 (0.0025-0.005 man-rad) 0.1 %due to 85Kr From these figures it is apparent that 85 Kr is not a significant contributor to local or regional collective doses but, because it has the Iongest half-life of the radioactive noble gases, it may make the greatest contribution to the global dose commitment. 3. Exposure Due to Fuel Reprocessing a) Production and Release. More than 90% ofthe 85Kr generated by fission in fuel with intact cladding is released at the fuel reprocessing plant. The rates of release were 1.5 x 1013 Bq (400 Ci)/MW(e)y for Windscale (UK) and 1.3 x 10 13 Bq (340 Ci)/MW(e)y for NFS (USA) ([6], Table 24, p. 201). Radioactive Substances 251 The NCRP predicted [32] that by the year 2000, when the world nuclear electric power generation would be 4,500 GW, the annual production of 85 Kr would be about 3.7 x 1019 Bq (1,000 MCi) and the amount accumulated in the world about 2.2 x 1020 Bq (6,000 MCi). b) Pathways to Man. UNSCEAR ([6], par. 158, p. 204) estimated that reprocessing spent fuel after cooling 150 d would yield the following collective tissue doses: Organ or tissue Collective dose fman-Gy/MW(e)y] Gonads Red hone marrow Lungs Skin 7 X 10-6 (7 X 10-4 man-rad) 1 X 10-5 (1 X 10-3 man-rad) 2 X 10-5 {2 X 10- 3 man-rad) 1 X 10- 3 (1 X 10- 1 man-rad) The following assumptions were made: i) All the 85 Kr in the fuel was released. ii) Dispersion factor at 1 km= 5 X 1o-7 s m-3 ([32] gives 1 X 10-7 s m-3). iii) Cd = Ct km d-LS, d in km. iv) Population density = 100 km-2 • v) Dose conversion coefficients as above. 4. Maximum Permissihle Concentration (MPC) Since the dose to skin is about two orders of magnitude higher than that to any other tissue the permitted concentration for continuous exposure would probably be that giving an annual dose of0.5 Sv to the skin; this is 9.3 x 105 Bq (25 J.lCi)/m3 • Ifthe dose Iimit is based on an annual dose of0.3 Sv to the lens ofthe eye, the MPC would be 5.5 x 105 Bq/m3 (15 J.1Ci/m3). Strontium-90 Exposure Due to Man-Made Sources 1. Exposure Due to Nuclear Bombs a) Production. 90Sr is a fission product, the yield varying with the fissile material and with the method offission, from about 1-9% [35]. The productioninnuclearbombsisestimated tobe 3.7 x 1015 Bq (0.1 MCi) per megaton of explosive energy [36] but this may vary greatly in individual tests. b) Release. The explosion of nuclear bombs in the atmosphere results in some local fallout of fission products which has not been documented in the open literature. The remaining fission products are carried aloft to the troposphere and stratosphere, circulate around the globe and s1owly deposit on the 252 G. C. Butler, C. Hyslop earth. The half-life of 90Sr in the stratosphere is about one year ([6], Sect. 13, p. 117). The variation ofstratospheric inventory of9°Sr from 1962 to 1975 for the world as well as northern and southern hemispheres has been published by UNSCEAR ([6], Fig. V, p. 121). The total in the stratosphere has declined from 2.3 x 1017 Bq (6.3 MCi) in 1962 to 3.7 x 1015 Bq (0.1 MCi) in 1974. In the northern hemisphere it has declined from 2 x 1017 Bq (5.4 MCi) in 1962 to 1 x 1015 Bq (0.03 MCi) in 1975. c) Deposition. The deposition ofbomb-produced stratospheric 90 Sr varies with latitude ([6], Table 3, p. 122), the maximum occurring at 40°-50° north. The annual worldwide deposition of 90Sr has been tabulated by UNSCEAR ([6], Table 2, p. 122). Part ofthe data are reproduced here as Table 8. The deposition velocity of an airborne material may be calculated by dividing the rate of deposition (Bq/cm2/s) by the air concentration above the surface (Bqjcm3) or by dividing the integrated deposit (Bq/cm2) by the timeintegrated air concentration (Bq-sjcm3). This gives the deposition velocity in Table 8. Annual deposition of strontium-90. (From [6], Table 2, p. 122; [38]) Annual deposition in Bq X 1016 (MCi) Northem hemisphere Southem hemisphere Global Pre-1958 1958 1959 1960 1961 1962 1963 1964 1965 1966 1967 1968 1969 1970· 1971 1972 1973 1974 1975 1976 Integrated deposition 6.7 (1.80) 2.3 (0.63) 3.9 (1.05) 1.0 (0.26) 1.3 (0.35) 5.3 (1.44) 9.7 (2.62) 6.1 (1.66) 2.8 (0.77) 1.2 (0.33) 0.6 (0.17) 0.7 (0.20) 0.6 (0.15) 0.8 (0.21) 0.7 (0.19) 0.3 (0.09) 0.1 (0.03) 0.4 (0.12) 0.2 (0.06) 0.1 (0.03) (0.65) (0.25) (0.18) (0.17) (0.17) (0.26) (0.31) (0.42) 1.3 (0.36) 0.8 (0.21) 0.4 (0.11) 0.4 (0.10) 0.5 (0.14) 0.5 (0.13) 0.6 (0.15) 0.4 (0.10) 0.1 (0.03) 0.1 (0.04) 0.1 (0.03) 0.007 (0.02) 9.1 (2.45) 3.3 (0.88) 4.6 (1.23) 1.6 (0.43) 1.9 (0.52) 6.3 (1.70) 10.8 (2.93) 7.7 (2.08) 4.2 (1.13) 2.0 (0.54) 1.0 (0.28) 1.1 (0.30) 1.1 (0.29) 1.3 (0.34) 1.3 (0.34) 0.7 (0.19) 0.2 (0.06) 0.6 (0.16) 0.3 (0.09) 0.2 (0.05) 45 (12.16) 14.2 (3.83) 59.2 (15.99) Stratospheric inventory 0.9 (0.23) 0.04 (0.01) 0.9 (0.24) 45.8 (12.39) 14.2 (3.84) 60.1 (16.23) Total injection to January 1977 2.4 0.9 0.7 0.6 0.6 1.0 1.1 1.6 Radioactive Substarrces 253 the usual dimensions of cmjs. When data like those contained in [37] for New Y ork Cityare used for such calculations a value of 4 cm/s results. Similar data for other sites may yield somewhat lower values (1-3 cm/s). d) Pathways to Man. The radioactive material released to the stratosphere is transported in the air, from which it may enter the body directly by inhalation or indirectly by deposition on the earth's surface and entry into the body in drinking water and food. This latter indirect route gives most of the dose to tissues. UNSCEAR ([6], Sect. 68-69, p. 131) has published estimates of the transfer offallout 90 Sr to diet in various parts of the world and for a number of foods. The numbers for P23 for total diet range from 0.1 to 0.4 Bq (3-10 pCi)y/g Ca per 3.7 x 107 Bq/km2 (mean value ofP23 is about 5). Assuming that the daily dietary intake of Ca = 1 g for people of all ages ([30], p. 365), the daily ingestion of 90 Sr can be calculated from the rate of deposition of 90 Sr. For population intake by inhalation assume 100 people/km2. 2. Exposure Due to Nuclear Reactor Operations a) Production. Some examples of the estimated amounts of 90 Sr present in irradiated reactor fuel are: - fuel irradiated to 20,000 MWd/tonne contains 1.9 X 10 15 Bq ( 50,000 Ci) 90 Sr per tonne ([39], Table 3, p. 15); - in the "reference" irradiated fuel ofWASH-1400, 0.13% ofthe total activity is due to 90 Sr ([33], Table VI -1, p. 6); - in a LWR, fuel irradiated to 33,000 MWd/tonne contained 2.8 x 1015 Bq (77,000 Ci) 90 Sr per tonne, corresponding to 9.6 x 10 13 Bq (2,600 Ci) per MW(e)y [40]. b) Release. Releases of 90 Sr to the environment from nuclear power production are of current interest; they fall into two categories. i) Normal operations. The releases from normal Operations vary with the type of reactor and its containment but they have been documented by UNSCEAR for BWR ([6], pp. 188-190) as: - to water, 3.7 x 106 Bq (100 JlCi)/MW(e)y; - to air, 1.9 x 10 5 Bq (5 JlCi)/MW(e)y. ii) Accidents. For accidental fuel melt-down the MRC ([39], Table 3, p. 15) has estimated that 1% of the 90 Sr in the fuel would escape to the atmosphere, although this could be reduced to a fraction of a percent by fuel-cladding ([39, p. 16). There could, in an event, be large variations in this number depending on many relevant influences. In WASH-1400 ([33], Table VI-2, p. 9) a wide variety of fractional releases (from a few percent to negligible) of strontium isotopes in irradiated fuel, along with their probabilities, have been estimated for different PWR and BWR. c) Pathways to man. For releases to air the intake may be calculated as for bomb fallout. For releases to water, calculations [38] show that releasestosalt G. C. Butler, C. Hyslop 254 water result in negligible population dose commitments and that for fresh water, most ofthe population dose comes from drinking the water rather than eating the fish that live in it. From the knowledge that Reference Mandrinks 2-3 L ofwater per day ([30], p. 360), the daily intake may be calculated. The principal routes of entry of9°Sr into the body for the population most affected by an aceidentat release of fission products will be air transport resulting in inhalation and ingestion of contaminated milk through the air ---t forage ---t cow ---t milk food chain. Concentrations in air must be measured or calculated according to [34] or [41] or from a knowledge ofKa (see 3 H, Sect. 2c). From the air concentration the amount inhaled can be calculated from the fact that Reference Man inhales about 2.2 x 104 L of air per day ([30], p. 346). The deposition rate may be calculated as described above. The MRC ([39], Table 12, p. 40) has calculated that, following the deposition of 3. 7 x 104 Bq (lj.1Ci)/m2 on pasture the integrated concentration of 90 Sr in milk is 1.7 x 109 tr (0.53j.1Ci-d)/L in the first year and a total of 4.8 x 109 tr (l.5j.1Ci-d)/L. From the daily intake ofmilk (0.7 L for an infant and 0.5 L for an adult) the ingestion of 90 Sr resulting from deposition on pasture can be calculated. 3. Exposure Due to Fuel Reprocessing a) Production and Release. For fuel reprocessing UNSCEAR ([6], p. 201) has published the following typical release rates: UK USA To air 2 X 106 Bq 1.5 X 106 Bq (4 X 10-5 Ci)/MW(e)y (6 X 10-5 Ci)/MW(e)y Towater 2 X !Oll Bq (6 Ci)/MW(e)y 3.7 X 108 Bq (1 X 10-2 Ci)/MW(e)y b) Pathways to Man. See Sect. 2c above. 4. Annual Limits on Intake As described in [7], Committee 2 of ICRP has calculated the effective dose equivalent (HE) resulting from the ingestion or inhalation of 1 Bq of 90 Sr and from this the intakes (ALl) that result in an effective dose equivalent of 50 mSv to all tissues of the body or that give a dose of 0.5 Sv to the most irradiated tissue. For 90 Sr these are: Ingestion Inhalation (Class D) (Class Y) 1 x 106 Bq (27 J.tCi) 8 x 1Q5 Bq (21.6 J.tCi) 1 x 105 Bq (2.7 J.tCi) From these can be calculated the effective dose equivalents resulting from the intakes estimated above. Radioactive Substances 255 Iodine-131 Exposure Due to Man-Made Sources 1. Exposure Due to Nuclear Bombs a) Production. Several radioactive isotopes of iodine are produced in appreciable yields in nuclear fission or as daughters arising from the radioactive transformation of other fission products (e.g. Te). The iodine isotope produced in fission which is of greatest concern in environmental contamination is 131 1. The fission yield of 131 I is 3% ([42], p. 255), halfthat of 137 Cs. b) Release and Deposition. UNSCEAR ([6], Table 16, p. 139) has published integrated milk concentrations of 131 I for severallocations from 1966 to 1976. Reference [43] (Table 1.1, p. 59) reports values of deposition velocity from 0.1-5 cm/s and adopts a nominal value of 1 cm/s. c) Pathways to Man. Ifitis assumed that dairy cows obtain all their fodder by grazing grass contaminated with radioiodine, the food chain air ~ grass ~ cow ~milk child outweighs the inhalation dose by a factor of3 for an adult and 60 for an infant. Thus the dose commitments to human populations are often assessed by monitaring Ievels of 131 I in commercial fresh milk and calculating doses to the thyroid gland of Reference Man. The dose to a "reference child" would beten times higher because the infant thyroid has a mass of about 2 g whereas the adult gland has a mass of about 20 g ([6], Table 18, p. 195). UNSCEAR has published ([6], Table 16, p. 139) calculated dose commitments to infant thyroids for several places in the northern and southern hemispheres for the decade following 1966. 2. Exposure Due to Nuclear Reactor Operations a) Production. Because of the relatively short half-life of 131 I the content in reactor fuel does not continue to increase with time of irradiation but soon reaches a constant equi1ibrium Ievel which has been reported by UNSCEAR as 1 x 1015 Bq (30 kCi)/MW ([6], Sect. 82, p. 181). In the "reference fuel" of WASH -1400 ([33], Table VI -1, p. 6) 2.2% of the fission product activity was due to 131 1. The postulated fuel ofMRC, 1975 ([39], Table 3, p. 15), irradiated to 20,000 MWd/tonne contained 2 x 10 16 Bq (6 x 105 Ci)/tonne of 131 1. As reported in ORNL-4451 the LWR fuel irradiated to 33,000 MWd/tonne, and cooled 150 days, contained 8 x 10 10 Bq (2.2 Ci)/t [40]. b) Release. UNSCEAR ([6], Sect. 84, p. 184; Table 13, p. 185) reported average releases of 7-20 x 107 Bq (2-5 x 10-3 Ci)/MW(e)y for BWR and 2-20 x 106 Bq (5-50 x 10-5 Ci)/MW(e)y from PWR. There were, however, wide variations in release rates between various individual installations. 256 G. C. Butler, C. Hyslop lmportant experience of the environmental effects of radiodiodine released accidentally from a reactor resulted from the "Windscale Accident" in the UK [44]. Irradiated fuel elements became overheated, the cladding ruptured and volatile fission products were released through the stack. It has been estimated that 7 x 1014 Bq (20,000 Ci) of 131 1 were discharged to the environment ([26], p. 129). WASH-1400 ([33], Table Vl-2, p. 9) gives estimated releases of negligible to 60%, with the correspondingprobabilities, for a number of different PWR andBWR. MRC in 1975 ([39], Table 4, p. 16) estimated percentage releases of 0.2-100% of the postulated iodine content depending on the type of cladding and other reactor variables. c) Deposition. The local collective dose commitment for the release of 131 1 from reactors is given by UNSCEAR ([6], Sect. 118, pp. 195-196) as 6 x 10-12 man-Gy per Bq (22 man-rad per Ci), which, on the basis of operating experience gives, for BWR PWR, GCR, HWR I x 10-3 man-Gy (0.1 man-rad)/MW(e)y I x 10-s man-Gy (1 x 10-3 man-rad)/MW(e)y ICRP Committee 4 has calculated the food chain contamination resulting at 1000 m downwind from an assumed continuous atmospheric release of 1311 to terrestrial environment. The following values are given in [43]: Assumed ground-level releaserate Effects at I km: - deposition on grass and vegetables - concentration in cows' milk - rate of intake by infant drinking 0. 7 Lmilkfday 3.7 x 1010 Bq (1 Ci)/yr 11 Bq (300 pCi)/m2 4 Bq (120 pCi)/L 3 Bq (84 pCi)fd. From this can be calculated the annual intake and the annual dose to the thyroid (from the ALl). d) Pathways to Man. In the "Windscale Accident" of 1957 which released 7 x 1014 Bq (20,000 Ci) of 131 1, the maximum concentration in milk was 5 x 104 Bq (1.4 J.1Ci)/L ([26], pp. 129, 132). Forasingle ground-level release of 1311 which was assumed to deposit on pasture 3.7 x 104 Bq (1 J.1Ci)/m2 the MRC ([39], p. 22) calculated an integrated concentration in fresh cows' milk of 5 x 104 Bq-days (1.4 J.1Ci-days)/L and thus an intake by an infant of3.7 x 104 Bq (1 J.1Ci) of 131 1. The resulting dose to the thyroid was estimated tobe 0.2 Sv (16 rems) for a child. ICRP Committee 4 ([43], Table 1.7, p. 65) assumed an acute ground-level release of 3.7 x 1010 Bq (1 Ci) of 131 1 to the atmosphere; the calculated results at a distance of 1,000 m downwind were: Assumed ground-level release Effects at 1 km: - deposition on grass and vegetables - beef -milk 3.7 X 1010 Bq (1 Ci) 1.1 X 1010 tr (3.5 J.!Ci·d)fm2 1.1 x 109 tr (0.34 J.!Ci·d)fkg 4.1 x 109 tr (1.3 J,lCi·d)/L This would give a dose to the infant thyroid of0.15 Sv (15 rems). Radioactive Substauces 257 3. Exposure Due to Fuel Reprocessing a) Production and Release. As mentioned in Sect. 2a the amount of 131 I in the irradiated fuel depends very much on the cooling time. Experience in the UK ([6], Table 24, p. 201; Sect. 161, p. 205) in the 1970's has shown a normalized release of 131 I of 3 x 108 Bq (9 x 10-3 Ci)/MW(e)y andin the USA of 3 x 105 Bq (8 x 10-6 Ci)/MW(e)y. b) Deposition and Pathways to Man. Because irradiated reactor fuel is allowed to "cool" for some months before reprocessing, most of the 131 I will have disappeared by radioactive decay. The environmental effects are therefore due to the longer-lived 129I (half-life 1.6 x 107 yr). Local collective dose commitments from the 129I released in fuel reprocessing have been published by UNSCEAR ([6], Sect. 163, p. 205). 4. Annual Limits on Intake ICRP Committee 2 has calculated that the following intakes of 131 I give an effective dose equivalent of 50 mSv [7]: Ingestion Inhalation (Class D) 4 X 106 Bq (I X 102 jlCi) 6 x 106 Bq (2 X 1Q2 j.!Ci) The ALl calculated by ICRP Committee 2 to give a dose of 0.5 Sv to the thyroid gland of Reference Man are: Ingestion Inhalation (Class D) 1 x I 06 Bq (30 jlCi) 2 x I 06 Bq (55 j.!Ci) For infants these ALI's should be reduced by a factor of 10. Caesium-137 Exposure Due to Man-Made Sources 1. Exposure Due to Nuclear Bombs a) Production. About six atoms of 137 Cs are produced perhundred fissions. Assuming that 1.45 x 1023 fissions yield 1 kt of energy [36] it may be calculated that an explosion of 1 megaton produces 6.3 x 1015 Bq (0.17 MCi) of 137 Cs. b) Release and Deposition. The calculation above shows that the activity yield of 137 Cs from a nuclear explosion is 1. 7 timesthat of 90 Sr (see p. 251 ). In addition, UNSCEAR has reported ([6], Sect. 97, p. 141; [11], Vol. 1, Sect. 222, p. 51) that the ratio of 137 Cs/90 Sr is fairly constant at about 1.6 in fallout deposited at many different times and sites. Thus the moreextensive data on 90 Sr Ievels can frequently be used to compute the 137 Cs Ievels. Since production and deposition ratios are nearly the same, one may conclude that 137 Cs and 90 Sr have equal deposition velocities, when they are averaged over long periods and many different atmospheric conditions. G. C. Butler, C. Hyslop 258 c) Pathways to Man. 137 Cs deposited on the earth from the air finds its way into human diets mainly through grain, meat and milk ([6], Table 17, p. 143). The transfer factor from deposition to total diet is taken by UNSCEAR ([6], Sect. 105, p. 143) tobe 0.1 Bq (4 pCi) per g ofpotassium in food per 3.7 x 107 Bq (1 mCi) deposited per km2 (P 23 =4 x 1019 Bq (gK-1) per Bq km-2 or 4 pCi (gK-1) per mCi km-2). The concentration in milk is representative ofthat in total diet. For residents of Chicago, Gustafson et al. [45] reported that the dietary contributions to 137 Cs intake were approximately as follows: milk, 30%; grain, 25%; meat, 20%; fruits, 10%; vegetables, 10%; other, 5%. The following facts about the potassium metabolism of Reference Man ([30], pp. 327, 403) permit calculation of daily intakes and equilibrium body contents: - Body content of K at all ages ~ 2 gjkg infant - Daily intake in food, 10-year-old adult 0.5 g Kjday 3 gKjday 3.3 g K/day Similar quantitative conclusions were arrived at by the NCRP [46] who reported that at a continued depositionrate of 3.7 x 107 Bq (1 mCi)/km2 per year the dietary level would reach 0.1 Bq (3 pCi) 137 Cs/g K which would lead to a constant body content of 0.3 Bq (9 pCi)/g K which is a total of 48 Bq (1,300 pCi) for a 70 kg Reference Man. 137 Cs that finds its way into fresh water may find its way to man's food through the fish that live in those waters. These fish may have 137Cs concentrations several thousand times higher than the water [47]. In [43] it was calculated that a constant release of 137Cs into surface water sufficient to maintain a constant concentration of 0.04 Bq (1 pCi)/L would Iead to a concentration in fish of 111 Bq (3,000 pCi)/kg and this would be the dominant route of intake by two orders of magnitude. A Special pathway for the transfer offallout 137 Cs to the diet of sub-polar peoples is by way oflichens and reindeer or caribou meat. It is well known that lichens and mosses trap airborne pollutants [48] and that the deposition of bomb-produced fission products is higher in northern latitudes. Since reindeer and caribou graze lichens in winter their intake of 137 Cs may be high with consequent elevation of the concentration in meat. This meat is an important item of diet for native peoples in the Arctic and sub-Arctic who had body contents of 137Cs, and the resulting dose commitments, in the 1960's ([11], Vol. 1, Sect. 230, p. 52; Sect. 233 and Fig. XXI, p. 53), 50-100 times higher than other inhabitants of the northern hemisphere. 2. Exposure Due to Nuclear Reactor Operations a) Production. Some estimates ofthe amount of 137 Cs in irradiated reactor fuel are: - Fuel irradiated to 20,000 MWd per tonne contains 2.5 x 1015 Bq (6.67 x 104 Ci)jtonne of 137Cs ([39], Table 3, p. 15). 259 Radioactive Substances - In the "reference" irradiated fuel ofWASH -1400, O.t5% of the total activity was due to 137 Cs ([33], Table VI-1, p. 6). - The fuel of a light water reactor irradiated to 33,000 MWd per tonne contained 3.7 x 101s Bq (O.t MCi)/t of 137Cs [40]. b) Release. Releases of 137 Cs to the environment from nuclear power production may arise from three different sources, (i) reactor operations, (ii) fuel reprocessing, or (iii) accidents. i) For both PWR and BWR, UNSCEAR has given ([6], pp. t88-t89) the rate of airborne release of 137Cs as 7.4 x tos Bq (20 J.I.Ci)/MW(e)y. The same reference gives waterborne releases of about 7.4 x t 04 Bq (2 J.I.Ci)/MW(e)y for PWR and about 9.3 x tos Bq (25 J.I.Ci)/MW(e)y for BWR. ii) In fuel reprocessing UNSCEAR ([6], Sect. 150-t5t, p. 203) reports airborne releases of 10-7 to t0-10 of the 137Cs in the fuel, and waterborne releases to the sea of 10-2 to t0-3 ofthe activity in the fuel. iii) For aceidentat fuel melt-down the MRC ([39], Table 3, p. 15) has postulated arelease of 100% ofthe 137Cs in irradiated fuel (2.5 x t0 11 (6.7 Ci)/t in fuel irradiated to 20,000 MWd/t). Cladding ofthe fuel may reduce this to a few percent ([39], Table 4, p. t6). In WASH -t400 fractional releases of Cs varying from a few tens of a percent to negligible (along with their probabilities) were postulated for a number of PWR and BWR ([33], Table VI -t, p. 6). c) Deposition. In a short-term releasesuch as that occurring in an accident the deposition velocity could vary from 0.1 to 30 cm/s according to ICRP [43]. d) Pathways to Man. i) Reactor Operations. UNSCEAR ([6], Table 2t, p. 196) estimates a local collective dose equivalent commitment of 1.4 x to-s man-Gy (1.4 x lQ-3 man-rad) per MW(e)y from airborne effluents of PWR and 1.1 x I0-5 man-Gy (1.1 x I0-3 man-rad) per MW(e)y from BWR. For waterborne releases ([6],Table 23, p. 200) the dose commitments are: PWR BWR GCR 1.1 x to--7 man-Gy (1.1 x to-s man-rad)/MW(e)y 1.4 x 1(]6 man-Gy (1.4 x 10--4 man-rad)/MW(e)y 3.8 x 1(]6 man-Gy (3.8 x 10-4 man-rad)/MW(e)y ii) Fuel reprocessing. For effluents from fuel reprocessing plants UNSCEAR ([6], Tables 26-27, p. 206) estimates the following whole-body dose commitments from 137Cs: Airborne Salt waterborne Fresh waterborne 6.6 X w-7 man-Gy (6.6 X w-s man-rad)/MW(e)y 8 x 10-4 man-Gy (0.08 man-rad)/MW(e)y 3 x 10-4 man-Gy (0.03 man-rad)/MW(e)y WASH-1400 ([33], Table VI-t7, p. 60) estimates that a ground deposition of 137Cs of 3.7 x 107 Bq (1 mCi)/km2 gives a whole body dose equivalent rate of7 x tQ-7 Sv (7 x lQ-5 rem)/yr. iii) Accidents. The MRC concluded ([39], pp. 26-30) that when 137 Cs was released to the air the major source of internal contamination is by the pasture ~ cow ~ milk food chain. It was estimated that deposition of 3. 7 x 104 Bq (1 260 G. C. Butler, C. Hyslop l!Ci)/m2 on pasture would give a total integrated concentration in milk of 3.8 x 1010 tr (12l!Ci-days) per litre, 3.2 x 10 10 [10] ofthese being received in the first year. Ifthese integrated concentrations are multiplied by the daily intake of milk (0. 7 L/d for a child and 0.5 L/d for an adult) the number ofBq ingested will be obtained. ICRP Committee 4 ([43], p. 32; Figs. 1-12-1-14, pp. 90-92) calculated the consequences of an acute ground-level release of 3. 7 x 1010 Bq (1 Ci) of 137 Cs. At a distance of 1000 m downwind, computations ofthe results for the 1,000 days after the release gave deposition rates for soil and foliage (3.7 x 104 Bq (l!Ci)/m2), concentration in milk (3. 7 x 104 Bq (l!Ci)/L) and in beef (3. 7 x 104 Bq (l!Ci)/kg). From the results it was concluded that eating leafy vegetables would be the most critical food path for both infants and adults; otherwise drinkingmilk was the most critical path for infants and eating beef for adults. The whole-body dose for adults was calculated tobe 0.09 Sv (9 rems) in the first year, but because ofthelarge uncertainty in deposition velocity (0.1-30 ern/sec) the uncertainty in dose gave a similar range (5 x 10-3-1.8 Sv (0.5-176 rem)/yr). 3. Annual Limits on Intake ICRP has published the following ALl for 137Cs [7]: Ingestion Inhalation (Class D) 4 X 106 Bq (I X 102 1lCi) 6 X 106 Bq (2 X 102 1lCi) Radium-226 Exposure Due to Natural Sources a) Production. Radium-226 is an intermediate member of the radioactive decay chain of uranium-238 found in nature. Most samples of "radium" will contain several short-lived daughters; the principal radioactive emissionswill include the a-particles from 226 Ra, 222 Rn, 218 Po, and 214 Po; the ß-particles from 214 Pb, 214 Bi, and 210 Tl along with a mixture of y-rays mainly from 226 Ra and 214 Bi ([42], p. 246). Since the uranium-238 radioactive family occurs in nature, 226 Ra and its parents and daughters arenormal constituents ofthe earth's crust. They occur in higher concentrations in uranium ores. 226 Ra is important as an environmental contaminant not only because of its ubiquity, leading to daily intakes by inhalation and ingestion, but also because of its first radioactive daughter, radon-222. 222 Rn, being a noble gas, is transported in airtobe breathed by man or to contaminate the environment more widely by deposition ofits radioactive daughters. Environmental contamination by 222 Rn is in itself a large and important subject which will not be dealt with in this section which is restricted to 226 Ra alone. b) Pathways to Man. Small amounts of radium are found in the air due to resuspension of soil particles. In most regions of the earth this is responsible for the daily inhalation of about 3. 7 x 10-5 Bq (10-15 Ci) ([6], Sect. 112, p. 59). 261 Radioactive Substances Uncontaminated surface waters usually contain such small amounts of radium that drinking water is a minor source of intake. Some wells and hot springs, however, may contain 0.04-0.4 Bq (1-10 pCi)/L ofradium ([6], Sect. 114, p. 59). For the population in general food is the main source of radium intake which, for an average diet, may be about 0.04 Bq (1 pCi)/day ([6], Sect. 113, p. 59). Larger dietary intakes may be due to eating some items offood containing higher concentrations of radium (Brazil nuts and Pacific salmon) or to living in areas (found in India and Brazil) with high concentrations of natural uranium and thorium in the soil ([6], Sect. 113, 115, p. 59). For purposes of dosimetry, UNSCEAR ([6], pp. 60-61) gives data on the radium content of various human tissues, especially hone, as a function of dietary content and daily intake. The average activity concentrations in four important tissues are shown in the table below. The resulting a-particle doses per year were calculated assuming that two-thirds of the 222 Rn daughter escaped from the tissue. Organ or tissue 226Ra concentration in Bq/kg (pCi/kg) Yearly a dose inGy(mrad) Lung Gonads Bone Red hone marrow Bone lining cells 4.8 X 10-3 (0.13) 4.8 X 10-3 (0.13) 0.3 (8) 4.8 X 10-3 (0.13) 4.8 X 10-3 (0.13) 1 X 10-7 (0.01) 1 X 10-7 (0.01) The effective dose equivalent (H0 is about 4 ~v 3 X 10-7 (0.03) 2.7 X 10-6 (0.27) (4 X 10-4 rem)/yr. Exposure Due to Man-Made-Sources 1. Exposure Due to Coal-Fired Power Plants ([6], pp. 86--88) a) Production and Release. Coal contains all the radioactive elements found naturally in the earth and when the coal is burned these radionuclides are emitted through the stack in the fly ash. lf the activity concentration of 226 Ra in fly ash is 0.04 Bq (1 pCi)/g and the flow of fly ash through the stack is 0.7 to 30 t (representative mean 10 tonne) per megawatt-year of electrical-energy, the activity of 226 Ra discharged would be 3.7 x 105 Bq (1Q-5 Ci)/MW(e)y. b) Pathways to Man. From assumptions about diffusion ofthe emitted fly ash and population density, the following collective a dose commitments to various tissues have been calculated. G. C. Butler, C. Hyslop 262 Collective a dose commitment in man-Gy per MW(e)y [man-rad per MW(e)y] Lung Radionuclide w-6 w1 x w- 2~ 2x (2 x Total (238U, 226Ra, 210pb, 228Ra, 22sn, 232Th) Red bone marrow Bone lining cells ww-6) 2 x w- 5 (2 x w- 3) ww- 5) 2 x w(2 x w- 2) 4x (4 x 4) 4 (10-2) 8 3 x (3 x 7 4 The effective dose equivalent (HE) is about 5 x 10--{i man-Sv (5 x 10-4 manrem)/MW(e)y for 226 Ra and 6 x 10-4 man-Sv (6 x I0-2 man-rem)/MW(e)y for all nuclides emitted. The collective effective dose commitments from these releases are about 5 man-Sv (5 x 102 man-rem) and 600 man-Sv (6 x 104 man-rem), respectively. 2. Exposure Due to Inhaled Phosphates ([6], pp. 89-91) a) Production and Release. Phosphate-containing rock is mined in appreciable quantities (130 million tonnes in USA in 1973) to provide industrial phosphates, one-halffor fertilizer, the other halffor chemieals and gypsum building materials. The most important natural radionuclide in this rock is 226 Ra which occurs in concentrations from 0.04-5 Bq (1-130 pCi)/g. b) Pathways to Man. The largest collective dose commitments resulting from the various uses are those from the use of phosphogypsum as a building material, viz., 0.01 man-Gy (1 man-rad) ofwhole body dose from y-rays per tonne of rock marketed (38 x 106 tjyr in USA), giving a collective dose commitment of 4 x 105 man-Svjyr. Lesser doses, shown in the following table, result from the use ofthe phosphate rock for fertilizers (2 x 107 tjyr in USA). Collective a dose commitment in man-Gy per tonne (man-rad per tonne) Radionuclide Lung 2~ 4x (4 x Total ( 238U, 2~ 210pb) w-8 w-6) 6 x w-7 (6 X 10-5) Red bone marrow Gonads w- 4 x 8 (4X10-6) 1x (1 x w-6 w- 4) 1 x w-7 o x w1x (1 x 5) w-6 w- 4) Bone lining cells 1x w- o x w1x (7 x 6 4) w-6 w- 4) The resulting collective dose commitments are 30 man-Sv/yr for 226 Ra and 270 man-Sv/yr for all nuclides emitted. 3. Exposure Due to Luminous Timepieces ([6], pp. 96, 97) a) Production and Release. Although largely replaced as a luminizer for watch dials, 226 Ra is still widely used for clocks. 263 Radioactive Substances b) Pathways to Man. UNSCEAR has concluded that the population doses resulting from this application arise mainly from external exposure to y-rays. They have also calculated that the annual dose to the gonads could be 2 x 108 Bq (6 mrad) from a wristwatch and 3.7 x 106 Bq (0.1 mrad) from an alarm clock. 4. Exposure Due to Uranium Milling ([6], pp. 167-170) a) Production and Release. 226 Ra is the most important radionuclide in uranium mill wastes; the concentration in liquid effiuent may vary from 9.3-18.5 Bq (250-500 pCi)/L. In dry tailings the concentration may be 20.7 Bq (560 pCi)jg. A mill processing 6 x 105 t of uranium ore per year in the USA released, airborne, about 3.7 x 108 Bq (10 mCi) of 226 Ra per year, which is equivalent to 3.7 x 104 Bq (1 J.1Ci) per MW(e)y. b) Pathways to Man. UNSCEAR calculated the collective dose commitments from this release as: Collective dose commitments in man-Gy/MW(e)y [man-rad/MW(e)y] Route of exposure Body External 6 x w-7 (6 X 10-5) Lung Bone marrow 5 x w-8 Ingestion x w- 6) 5 x w- 8 (5 1 x w-7 Inhalation (1 X 10-5) (5 X 10-6) Bone lining cells 6 x w- 8 x w- 6) 4 x w- 8 (6 (4 X 10-6) 5. Exposure Due to Uranium Fuel Fabrication ([6], p. 171) In fuel fabrication residual amounts of 226 Ra are removed from the uranium compounds produced in milling. In the USA it was estimated that fuel fabrication operations released about 3. 7 x 109 Bq (0.1 Ci)jyr of 226 Ra in liquid Ci)/MW(e)y. effiuents, equivalent to 1.3 x 105 Bq (3.4 x 10~ 6. Annual Limits on Intake [7] ICRP Committee 2 has calculated that the following intakes will give an effective dose equivalent of 50 mSv: Ingestion Inhalation (Class W) 5 X 104 Bq (1 J.1Ci) 8 x J03 Bq (0.2 JlCi) 264 G. C. Butler, C. Hyslop This ALl for ingestion could be used to calculate a maximum permissible concentration of 226 Ra in soil, using Canada as an example, from the following facts: - ICRP maximum permissible annual intake of 226 Ra for individuals in the population (critical group) = 0.1 x ALl = 4 x 103 Bq (0.1 J.!Ci) ([3], Sect. 119, p. 23); - Annual consumption ofvegetables, other than potatoes = 86 kg [49]; - Vegetable fresh weight/dry weight ~ 11 [50]; 226Ra per g dry weight vegetable 226Ra per g dry soil = 0 25 [Sl]· · ' - Permissihle concentration of 226 Ra in soil to give a daily ingestion of 3.7 x 103 Bq (0.1 J.tCi) = 185 Bq (5 nCi)/kg. Plutonium-239 Exposure Due to Man-Made Sources 1. Exposure Due to Nuclear Bombs a) Production. 239 Pu is produced from neutron capture in 228 U according to the following scheme: mu 92 + n -+23~ 92 L rapid 23~P 93 L rapid 239pu 94 Capture of neutrons by 239 Pu 1eads to isotopes of plutonium with higher atomic weights and other transuranic daughters. A diagram illustrating these relations has been pub1ished by UNSCEAR ([6], Fig. 1, p. 204). The plutonium that occurs in the environment is usually a mixture of 239 Pu and 240 Pu which are difficult to distinguish; therefore, hereafter, "Pu" will be used to indicate the mixture of 239&2 40 Pu. b) Release and Deposition. It has been estimated that 1.5 x 10 16 Bq (400 kCi) of Pu have been released in weapons testing and that 1.2 x 10 16 Bq (320 kCi) have been dispersed around the world [52], 9.3 x 10 15 Bq (250 kCi) in the northern hemisphere and 2.6 x 1015 Bq (70 kCi) in the southern hemisphere ([6], Sect. 127, p. 148). Isotopic analyses have indicated that the ratio of activities 239 Puj2 40 Pu = 60/40 [52]. Mostofthis came from tests conducted before 1963. Many measurements over several years have shown that in the stratosphere andin surface air the activity ratio Puj9°Sr has remained fairly constant at 0.017 ([6], Sect. 127, p. 148). By assuming the same deposition velocity as for 90 Sr (1-4 cm/s) the deposition can be calculated when monitaring data arenot available. Bennett has published the results offallout monitaring and computation for New York for the twenty years 1954--1974 ([52], Table I, p. 368). Radioactive Substances 265 Plutonium deposited on soil moves slowly downward and displays the same depth profile as 137Cs ([52], pp. 375-376). Ninety-five percent of plutoniumentering the sea and freshwater lakes is quickly deposited in sediments where its behaviour is similar tothat of 137Cs [27, 53]. In January 1968 aB-52 aeroplaneloaded with a nuclear bomb crashed at Thule, Greenland, dispersing about 9.3 x 10 11 Bq (25 Ci) of plutonium into the sea. Environmental monitaring carried out between 1968 and 1974 discovered the presence of some plutonium in bottarn Sediments, molluscs and worms but nonein higher vertebrates such as fish, seabirds and marine mammals [54]. c) Pathways to Man. UNSCEAR has concluded ([6], Sects. 131-136, p. 148) that the mostimportantraute to man is by inhalation of contaminated air. Bennett [52] has calculated that residents of New Y ork inhaled a total of about 1.5 Bq (40 pCi) ofPu in the two decades from 1954--1974. Thus, ofthe 9.3 x 1015 Bq (250 kCi) deposited inthe northern hemisphere during the same period, about 10-16 was inhaled by an average individual. This fraction should be kept in mind for assessing some ofthe absurd estimates ofthe consequences ofhaving plutonium fuel in reactors [55]. The calculated body content at the end of the 20-year inhalation intake by New Y ork residents was 0.09 Bq (2.5 pCi) ([6], Sects. 131-136, p. 148). The contents of various argans and tissues were calculated; the results agreed reasonably well for all tissues, except kidneys, with those found by analysis of members ofthe population in the USA [52, 56]. UNSCEAR estimated that the population-weighted dose, up to 2000 A.D., from bomb plutoniumwas 1 x 10-s Gy (1 mrad) in the northern hemisphere and 3 x 10--U Gy (0.3 mrad) in the southern ([6], par. 131-136, p. 148). Ingestion of environmental plutonium by man may result from its deposition on land or its entry into surface waters. Bennett ([52], Table IV, p. 374) has reported measured values ofthe ratio pCi per g fresh weight of vegetables pCi per g of soil for a number of plants including vegetables in the human diet; most of the values ranged from 1 x 10-3 to 1 x 10-4. Miettinen [22] measured Pu in two food chains in Finland. In the terrestrial one the Pu content in Iichens and reindeer liver, respectively, were 8.1 Bq (220 pCi)/kg and 0.7 Bq (20 pCi)/kg in 1963 and 0.7 Bq (20 pCi)/kg and 0.07 Bq (2 pCi)/kg in 1973. In a typical marine food chain from the Gulf afFinland the following concentrations of Pu were found: Sediment Brown algae (fresh wt) Blue mussei (fresh wt, whole animal) Fish (fresh wt) 7.4 Bq (200 pCi)/kg 0.2 Bq (5 pCi)/kg 0.02 Bq (0.6 pCi)/kg 1.5 X 10-3 to 5.2 X 10-3 Bq (0.04--0.14 pCi)/kg Measurements were made of 239 Pu in several marine invertebrates, including mussels, clams, oysters and scallops, from Cape Cod [57]. Mean body concentrations ranged from 4 x 10-3 to 1.8 x 1o-z Bq (0.11 to 0.49 pCi)/kg fresh wt (body), 100-500 times greater than concentrations in the environment. G. C. Butler, C. Hyslop 266 In a limnological study of Lake Michigan and other Great Lakes, Edgington et al. [27] measured the concentration ofPu in sediments, mixed plankton, zooplankton, planktivorous fish, piscivorous fish and water. The concentration declined quite regularly through each stage from 3. 7 Bq (100 pCi)/kg to about 3. 7 x 10-5 Bq (1 0-3 pCi)/kg by about one order of magnitude per stage. From the data available it seems clear that the concentration of plutonium declines as one proceeds along food chains from soil to man. Bennett [52] reported the results of a dietary analysis in New Y ork in 1972 indicating that the annual ingestion ofPu was 0.06 Bq (1.6 pCi). From this the transfer coefficient P 23 was calculated by UNSCEAR ([6], Sect. 138, p. 150) as _ _:.0-=6B~q"y1 6.3 X 105 Bq km- 2 y- 1 or 6 L.P. =C::.. . y._-...." i 1 ,..----:,...-- -1::.:.·.::... 0.017 mCi km- 2 y- 1 = 9.5 X = w-s Bq/Bq km-2 94 pCilmCi km- 2. If this coefficient is multiplied by the estimated deposition for each year from 1954-1974 a total ingestion of 9.25 Bq (250 pCi) is derived. From this UNSCEAR estimates the mean population dose commitment tobe 1.2 x 10-7 Gy (1.2 x 10-2 mrad) and the collective dose commitment from all test explosionstobe 3 x 10-12 man-Gy/Bq (10 man-rad/Ci) ofPu released, to the bone lining cells and to the lungs ([6], Sect. 140, 142, p. 150). A somewhat different estimate of individual dose commitment is obtained using the dosimetry calculations of ICRP Committee 2 [7] as follows: - The ingestion of 7.4 x 105 Bq (0.02 mCi) gives an effective dose equivalent of0.5 Sv to bone surfaces. - The ingestion of 7.4 Bq (200 pCi) would give 5 J.lSV to bone surfaces. - Since ICRP uses Q = 20 for a-particles, 5 J.tSv = 2.5 x 10-7 Gy (2.5 x 10-2 mrad) - This is tobe compared with the 1.2 x 10-7 Gy (1.2 x 10-2 mrad) estimated by UNSCEAR (above). 2. Exposure Due to Nuc/ear Reactor Operations a) Production and Release. At present the greatest production of 239 Pu is for nuclear weapons, but understandably very little information is available concerning the amounts produced or released. The environmental statement for the LMFBR ofthe USA, quoted in [58], postulatesarelease to the atmosphere of3.7 x 106 Bq (0.1 mCi) ofPu per 1,000 MW(e)y. It is also given as 10-9 ofthe Pu made and burned in the fuel cycle [59]. Because of its low volatility and because it gives rather low dose commitments when released to the environment, plutonium is not often considered in assessing the consequences of reactor accidents. For example, in the "reference" fuel ofWASH-1400 ([33], Table VI-I, p. 6; Table VI-2, p. 9) 2.5 x 10-4% 267 Radioactive Substances of the radioactivity is due to Pu and of this, less than 10-3 is released in the postulated accident. b) Pathways to Man. Estimates of the dose commitments to a population in the USA as a result of operating the LMFBR fuel cycle have been referred to in [58] and [59]. In [59] it is assumed that 10-5 ofthe Pu released [1.5 x 103 Bq (4 x I0-5 mCi)/MW(e)y [58]] is inhaled by man. Thus the intake by inhalation = 1.5 x 10-2 man-Bq (0.4 man-pCi)/MW(e)y. According to ICRP Committee 2 the inhalation of 200 Bq (5 nCi) gives an effective dose equivalent to the skeleton of 0.5 Sv [7]. Therefore the inhalation of 1.5 x 10-2 man-Bq (0.4 man-pCi) gives an effective population dose equivalent of 4 x 10-5 man-Sv/ MW(e)y. 3. Exposure Due to Fuel Reprocessing a) Production and Release. Pu is one of the nuclides of greatest concern for internal contamination ofworkers in fuel processing and reprocessing [60]; it is not considered one of the major nuclides for environmental contamination. UNSCEAR ([6], Table 25, p. 202) reports a release to liquid effiuents of 2.6 x 105 Bq (7 x I0-6 Ci)/MW(e)y from the NFS plant in the USA and a normalized release rate, for all plants, into liquid effiuents of 2.2 x 1010 Bq (0.6 Ci)/MW(e)y tosalt water and 2.6 x 105 Bq (7 x 10-6 Ci)/MW(e)y to fresh water ([6], Table 27, p. 206). b) Pathways to Man. UNSCEAR ([6], Table 25, p. 202) has calculated a collective dose from the 2.2 x 10 10 Bq (0.6 Ci)/MW(e)y ofPu, released in fuel processing, tobe 1 x I0-5 man-Gy (1 x 10-3 man-rad)/MW(e)y. 4. Annual Limits an Intake ICRP Committee 2 [7] has calculated that the following intakes ofPu give an effective dose equivalent of 50 mSv: Ingestion Inhalation (Class W) (Class Y) 1 x 106 Bq (30 11Ci) 4 x 102 Bq (10 nCi) 6 x 102 Bq (20 nCi) and that the following intakes give a dose equivalent of 500 mSv to hone lining cells: Ingestion Inhalation (Class W) (Class Y) 8 X 10 5 Bq (20 jlCi) 2 x 102 Bq (5 nCi) 5 x 1Q2 Bq (15 nCi). Larsen and Oldham [61] found that 1 ppm of chlorine in drinking water that contained approximately 0.2 Bq (5 pCi)/mL of 239 Pu (IV) resulted in a 75% oxidation to 239 Pu (VI). The significance of this for the maximum permissible concentration of plutonium in drinking water was nullified when Sullivan et al. [62] found that, in contradiction to Weeks et al. [63], there was no appreciable difference in the absorption of intragastrically injected 238 Pu (IV) and 238Pu (VI) by rats or guinea pigs allowed food ad libitum. 268 G. C. Butler, C. Hyslop References l. Internat. Comm. Rad. Units Measurements: Radiation Quantities and Units. ICRU Rep. 19; ICRU Publications: Washington, D.C. 1971 2. Internat. Comm. Rad. Units Measurements: Phys. Med. Bio!. 20, 1029 (1975) 3. Recommendations ofthe Internat. Comm. on Radiological Protection. ICRP Publication 26, Ann. ICRP 1, l (1977) 4. Mayneord, W.V., Clarke, R.H.: Brit. J. Radio!., Suppl. No. 12, 15 (1975) 5. Brown, J.M.: Health Phys. 31,231 (1976) 6. UN Sei. Comm. Effects of Atomic Radiation: Sources and Effects of Ionizing Radiation. Offic. Rec. ofthe General Assembly, 32nd Session, Suppl. No. 40 (A/32/40); United Nations: NewYork, N.Y.l977 7. Recommendations Internat. Comm. Radiological Protection, Rep. Committee 2: Limits for Intakes of Radionuclides by Workers. ICRP Publication 30, Part 1, Ann. ICRP 2 (3/4), I (1979) 8. Recommendations Internat. Comm. Radiological Protection, Rep. Committee 4: The Assessment of Intern. Contarnination Resulting from Recurrent or Prolonged Uptakes. ICRP Publication 10A; Pergarnon Press: Oxford 1971 9. Pochin, E.E.: Estimated Population Exposure from Nuclear Power Production and Other Radiation Sources; Organization for Economic Cooperation and Development: Paris 1976 10. Rep. UN Sei. Comm. Effects of Atomic Radiation, Offic. Rec. General Assembly, 13th Session, Suppl. No. 17 (A/3838); United Nations: NewYork, N.Y. l958.1bid., 17th Session, Suppl. 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Advisory Committee on the Biological Effects of lonizing Radiations (BEIR Report): The Effects on Populations ofExposure to Low Levels oflonizing Radiation; US Nat. Acad. Sei., Nat. Res. Council: Washington, D.C. 1972 15. New Primary Sites of Malignant Neoplasms in Canada, 1975. Statistics Canada: Ottawa 1978 16. Jacobi, W. in: Radiation Protection Measurement- Philosophy and lmplementation. EUR5397; Comm. Europ. Communities: Luxembourg 1975; p. 63 17. Butler, G.C. (ed.): Principles of Ecotoxicology. SCOPE Rep. 12; John Wiley: ChichesterNewYork-Brisbane-Toronto 1978 18. Internat. Atol'nic Energy Agency: Principles for Establishing Limits for the Release of Radioactive Materialsinto the Environment. Safety Series No. 45, STI/PUB/477; IAEA: Vienna 1978 19. Recommendations Internat. Comm. Radiological Protection, Rep. Committee 4: Principles of Environmental Monitoring Related to the Handling of Radioactive Materials. ICRP Publication 7; Pergarnon Press: Oxford 1965 20. Internat. Atomic Energy Agency: Effects of lonizing Radiation on Aquatic Organisms and Ecosystems. Techn. Rep. Ser. No. 172, STI/DOC/10/172; IAEA: Vienna 1976 21. Rogers, L.E., Rickard, W.H. (eds.): Ecology of the 200 Area Plateau Waste Management Environs: A Status Report. PNL-2253; Batteile PaeificNorthwest Laboratories: Richland, Washington 1977 22. Miettinen, J.K. in: Miller, M.W., Stannard, J.N. (eds.): Environmental Toxicity of Aquatic Radionuclides: Modelsand Mechanisms; Ann Arbor Seience Publishers Inc.: Ann Arbor, Michigan 1976; pp. 29-43 Radioactive Substances 269 23. Dahlman, R.C., Bondietti, E.A., Eyman, L.D. in: Friedman, A.M. (ed.): Actinides in the Environment; Amer. Chem. Soc., Washington, D.C. 1976; pp. 47-80 24. Singh, H., Marshall, J.S.: Health Phys. 32, 195 (1977) 25. Nat. Res. Council, Rep. by the Panel on Radioactivity in the Marine Environment of the Committee on Oceanography: Radioactivity in the Marine Environment; US Nat. Acad . . Sei., Washington, D.C. 1971 26. 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EML-334; Dept. of Energy, Environmental Measurements Lab.: New York, N.Y. 1978 38. Bennett, B.G.: Private communication 1978 39. Medical Res. Council: Criteria for Controlling Radiation Doses to the Public after AceidentalEscape ofRadioactive Material. Her Majesty's Stationery Office: London 1975 40. Oak Ridge Nat. Laboratory: Siting of Fuel Reprocessing Plantsand Waste Management Facilities. ORNL-4451; Oak Ridge Nat. Laboratory: Oak Ridge, Tennessee 1970 41. Bryant, P.M., Jones, J.A. in: Environmental Behavior of Radionuclides Released in the Nuclear Industry. STI/PUB/345; Internat. Atomic Energy Agency: Vienna 1973 42. Wilson, B.J. (ed.): The Radiochemical Manual, 2nd ed.; The Radiochemical Centre: Amersham 1966 43. Recommendations Internat. Comm. Radiological Protection, Rep. Committee 4: Radionuclide Release into the Environment: Assessment ofDoses to Man. ICRP Publication 29. Ann. ICRP 2 (2), 1 (1979) 44. 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Atomic Energy Agency: Vienna 1976; pp. 367-383 53. Hetherington, J.A. in: Miller, M.W., Stannard, J.N. (eds.): Environmental Toxicity of Aquatic Radionuclides: Models and Mechanisms; Ann Arbor Science Publishers Inc.: Ann Arbor, Michigan 1976; pp. 81-106 54. Aarkrog, A.: Health Phys. 32, 271 (1977) 55. Gofman, J.W.: J. Am. Med. Assoc. 236,284 (1976) 56. Campbell, E.E. et al. in: Annual Report of the Biomedical and Environmental Research Program ofthe LASL Health Division, Jan.-Dec. 1973. LA-5633-PR; Los A1amos Scientific Laboratory: Los A1amos, New Mexico 1974; pp. 27-33 57. Noshkin, V.E.: Health Phys. 22, 537 (1972) 58. Cuddihy, R.G. et al. in: Transuranium Nuclides in the Environment. STI/PUB/410; Internat. Atomic Energy Agency: Vienna 1976; pp. 657-669 59. Barr, N.F. in: Transuranium Nuclides in the Environment. STI/PUB/410; Internat. Atomic Energy Agency: Vienna 1976; pp. 649-656 60. Cohen, B.L.: Health Phys. 32, 359 (1977) 61. 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Res. 4, 339 (1956) Subject Index absorption of cadmium 94 acceptable daily intake (ADI), methyl mercury 42 accumulation of cadmium 85 ~ of chlorinated paraffins 155 ~ of mercury 29 Acid Blue 1 208 ~ Green 209 ~ Yellow 5 208 23 207 ~ ~ Agent Orange 170 air, transport of PAH 119 algae 200 alkoxyalkyl mercury compounds 27 alkylmercurials 26 alkyloxyalkyl mercury compounds 24 Ames test for PAH 127 amines, carcinogenic activity 205 analysis, cadmium 68 153 ~ of chlorinated paraffins 137 ~,fluorcabns mercury 12 ff. ~, ofPAH 118 ~ animals, uptake and excretion of cadmium 86 annellated systems 113 anthracene 112 antimony, colorants 227 aryl mercury compounds 26, 27 asbestos 217, 227 atmosphere, fluorocarbon concentration 138 atomic absorption 14 auramine 108 azo dyestuffs 196 barium salts 228 Basic Blue 209 Orange 14 208 ~ ~ Violet 10 208 benzidine 196 benzo[a]pyrene, chlorination 127 ~ concentrations 109, 110 ~, emJsswn 123 ~, metabolism 122 phenols 123 solubility in water 120 bioconcentration, PAH 126 biodegradation of chlorinated paraffins 125 ~ of PAH biotransformation, mercury 26 biota, cadmium concentration 79 ~ ~, cadmium 59fT., 217 absorption 94 ~, analytical methods 68 ~, aquatic chemistry 66 ~, biological half-life 96 ~, body distribution and excretion 95 ~, chemistry 64 ~, concentration in the environment 71 ~, concentrations in organisms 87 ~, consumption 61 70 ~, emission ~, food chain effects 88 ~, ~ concentration 91 65 ~, geochemistry ~, indicator organisms 90 ~, leaching 82 64 ~, minerals ~ in mussels and oysters 91 80 ~, natural cycle ~ pigments 225 production 60 ~, regulations 99 ~, 83 ~, remobilization ~, residence time 74 ~, storage and excretion 86 ~, toxic effects on humanes 98 ~, toxicological aspects 96 ~, transport in the environment 69 ~, ~ in estuarine zone 80 ~, uptake and accumulation 85 use 62 ~, caesium-137 257 carcinogenicity of dyestuffs 201 128 ~ of PAH carcinogens, dyestuffs 206 ~, 153 272 cell transformation test for PAH 127 chloroarornatic compounds, containing oxygen 157fT. chlorinated parallins 149fT. - -, accumulation 155 - -, chemistry 151 - -, persistence 155 - -, transport in the invironment 154 chlorine radicals in fluorocarbon photolysis 144 -, stability in chlorinated parallins 152 chlorophenols 157 -, analytical chemistry 165 -, analytical chernistry 165 -, biodegradation 159 -, metabolism 159 -, persistence 159 chromate 217 - pigments 224 cigarette smoke, cadrnium concentration 94 cinnabar 1 coal tar 109 coal-fired power plants, radioactive elements 261 color 181 colors 217 Colour Index 182 coastal water, cadrnium concentration 76 copper aceto-arsenate 228 decomposition on soil 145 decontamination of mercury 28 dianisidine 196 dibenzo[a,h]anthracene, solubility in water 120 dibenzofurans 157fT., 161 dibenzo[b,n]perylene, representation of bonding properties 115 dibenzo-p-dioxin 157fT., 161 Diels-Aider reaction 117 7,12-dimethyl-benzo[a]anthracene, in vitro oxidation 124 dimethyl mercury 24, 27 dioxinproblern 163 dioxins 160 -, analysis 169 diphenyl ethers 157fT. Direct Blue 210 - Yellow 12 207 Disperse Yellow 3 207 - Yellow 54 208 dithizone 14 dose Iimits, radiation 238 drinking water, upper Iimit for Cd 100 dye production 184 dyes, inorganic 217 - organic 181 dyestuffs 181, 183 Subject Index -, analytical methods 185 -, biodegradability 194 -, biological treatmentplant 192 -, degradation cycle 188 -, ecological aspects 186 -, effiuent treatment 188, 189 -, rnammalian toxicity 200 -, toxicity (fish) 199 -, wastewater treatment 191 effiuent standarts for cadmium 100 - treatment, dyestuffs 189 epoxide-hydratase 122 F 11, average photodissociation lifetimes 144 F 12, photodissociation lifetimes 144 fate of cadmium 71 - of chlorinated parallins 154 - ofPAH 125 fluorimetry 119 fluorocarbon ozon hypothesis 144 flurocarbons 133 ff. -, analytical methods 137 -, atmospheric residence time 144 -, biological effects and toxicity 145 -, chemical and physical properties 135 -, chemistry 136 -, concentrations in atmosphere 138 -, LD 50 146 -, metabolism 145 -, physical data 133 -, production and use 134 -, transport in the environment 137 food chain see aquatic food chain see terrestrial food chain - - effects, cadrnium 88 foodstuffs, mercury Ievels 36 fuel reprocessing 246 - -, 131 I 257 --,Kr 250 - -, 239 Pu 267 - -, 90 Sr 254 Greenland ice 8 Herbicide Orange 163 hexachlorophene 164 indanthrone 210 indicator organisms for cadmium 90 indigo 181,209 inorganic colorants, hazards 221 iodine-131 255 irradiation, natural 234 isotopes 231 ltai-Itai 95 - disease 59 Subject Index krypton-85 248 leaching tests, cadmium 81 Iead 217 - pigments 223 legislations, pigments 226 luminous timepieces 262 mauveine 209 maximum allowable concentrations, mercury 41 mercuric sulfide 29 mercury 1ff. -, anthropogenic discharge 3 - in aquatic media 31 - - organisms 35 - in the atmosphere 29 -, biological methylation 24 -, chemistry 8 - compounds, MAC values 43 - compounds 10, 16 -, contaminated rooms 9 - cycle 18 - distribution, soil 35 - emission 7 -, environmental release by burning and smelting 6 -, fate 22 -, food chain 21 - in hydrosphere 30 -, interconversion in the aquatic environment 20 -, intoxication 40 -, Ievels in air 30 - in marine biota 31 -, naturally released 8 -, persistence 38 -, photochemical reactions 23 - in plankton 33 -, plant uptake 22 -, production and consumption 2 - in sediments 30, 33 - in soil 21, 34 - in terrestrial animals and man 37 - - plants and fruits 36 -, thereshold Iimit values 32 -, toxicity 39 - transport 19 -,- in the environment 17 -, uptake 25 -, use 4fT. metabolic activation, benzo[a]pyrene 123 metabolism of fluorocarbons 145 - of mercurials 25 methyl mercury, ADI 42 - -, blood brain barrier 41 - - chloride 26 273 - - dicyandiamide 26 methylene blue 209 2-methyl-naphthalene, protolysis 111 microorganisms, oxidation of benzo[a]pyrene 125 Minamata disease 39 Mordant Red 209 mutagenicity of dyestuffs 203 naphtho [2,1-a]anthracene 112 neoplasms from radiation 237 nuclear bombs 243, 248 - -, 137 Cs 257 - -, 131 1 255 - -, 239 Pu 264 - -, 90 Sr 251 - reactor operations 244 - - -, 137 Cs 258 - - -, 131 1 255 - - -, Kr and Xe 249 - - -, 239 Pu 266 - - -, 90 Sr 253 organic mercury compounds 27 organomercury compounds 11, 24 ozone, decomposition by fluorocarbons 144 PAH, analyticar methods 118 -, carcinogenicity 128 -, chemical reactions 120 -, metabolism 122 -, synthetic methods 116 -, topology, stability, and reactivity 114 -, toxicology 126 -, transport 119 paraffin, chlorinated 149fT. PCB 164 PCDD- polychlorodibenzo-p-dioxin(s) 161 -, biodegradation 171 - in the environment 170 -, photochemical reactions 171 -, toxicity 175 PCDF- polychlorodibenzofuran(s) 161 -, biodegradation 171 - in the environment 170 - isomers 173 -, photochemical reactions 171 -, toxicity 175 pentachlorophenol 158 peri-condensed systems 113 persistence, dyestuffs 198 - of mercury 38 perylene 114 phenols 157 ff. phenyl mercury acetate 6 phosphorimetry 119 photochemical degradation of dyestuffs 193 274 Subject Index photolysis of tluorocarbons 143 Pigment Blue 209 - Yellow 12 207 pigments 183, 217 -, heavy metals 217 plant uptake, mercury 22 plants, uptake of cadmium 85 plutonium-239 264 polychlorinated biphenyls 164 - diphenyl ethers 160 polycyclic aromatic hydrocarbons (PAH) - heteroaromatic hydrocarbons 109 polytetrafluoroethylene (PTFE) 134 predioxin 158, 165 promethium 247 quinone, formation soil, cadmium 74 Solubilisation of mercury compounds 23 sorption of cadmium 81 spectrophotometry 14 stratosphere, photochemical fluorocarbon decomposition 143 strontium-90 251 -, annual deposition 252 Sulphur Black 209 109 ff. 121 radioactive materials, atmosphere 239 - -, water 239 - substances 231 ff. radiation diseases 236 - dose 232 - estimates 235 - sources 234 radionuclides 241 radiosensitivity 241 radium 247 radium-226 260 residence time, fluorocarbons in atmosphere 144 Rhine River, cadmium 78 - -, history oftrace metals in sediments 78 - -, Iead 78 - -, mercury 78 sediments, cadmium concentrations 77 -,- contents 72 selenium, protection against mercury toxicity 40 sewage sludge, cadmium 73 sillcates 227 TCDD- tetrachlorodibenzo-p-dioxin(s) 162 - in the environment 170 -, metabolism 172 -, persistence 174 TL V for cadmium 100 TOC, dyestuffs 190 tolerable weekly intake for cadmium 99 toxicity of cadmium 96 - of dyestuffs 202 -,fluorocarbons 145 -, PAH 126 transport see environmental transport 17 - of chlorinated paralTins 154 -,fluorocarbons 137 triphenylene 114 triphenylmethane dyestuffs 197 tritium 245, 246 - luminous paints 247 - oxide 242 Tyrian Purpie 181 uranium fuel fabrication 263 - milling 263 Wastewater, dyestuffs 188 water, cadmium concentration 75 - pollution controllaws 205 -, transport of PAH 120 Windscale Accident 256 xanthene dyestuffs Zeeman effect 14 198 Archives of TOXICOLOGV Archiv für Toxikologie Edited on behalf ofthe Deutsche Pharmakologische Gesellschaft and the Deutsche Gesellschaft ftir Rechtsmedizin Official Organ ofthe European Society ofToxicology Editor in Chief: H. 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