INTERNATIONAL PROGRAMME ON CHEMICAL SAFETY ENVIRONMENTAL HEALTH CRITERIA 135 CADMIUM - ENVIRONMENTAL ASPECTS This report contains the collective views of an international group of experts and does not necessarily represent the decisions or the stated policy of the United Nations Environment Programme, the International Labour Organisation, or the World Health Organization. Published under the joint sponsorship of the United Nations Environment Programme, the International Labour Organisation, and the World Health Organization First draft prepared by Dr S. Dobson, Institute of Terrestrial Ecology, United Kingdom World Health Orgnization Geneva, 1992 The International Programme on Chemical Safety (IPCS) is a joint venture of the United Nations Environment Programme, the International Labour Organisation, and the World Health Organization. The main objective of the IPCS is to carry out and disseminate evaluations of the effects of chemicals on human health and the quality of the environment. Supporting activities include the development of epidemiological, experimental laboratory, and risk-assessment methods that could produce internationally comparable results, and the development of manpower in the field of toxicology. Other activities carried out by the IPCS include the development of know-how for coping with chemical accidents, coordination of laboratory testing and epidemiological studies, and promotion of research on the mechanisms of the biological action of chemicals. WHO Library Cataloguing in Publication Data Cadmium : environmental aspects. (Environmental health criteria ; 135) 1.Cadmium - toxicity 2.Environmental exposure I.Series ISBN 92 4 157135 7 (NLM Classification: QV 290) ISSN 0250-863X The World Health Organization welcomes requests for permission to reproduce or translate its publications, in part or in full. Applications and enquiries should be addressed to the Office of Publications, World Health Organization, Geneva, Switzerland, which will be glad to provide the latest information on any changes made to the text, plans for new editions, and reprints and translations already available. (c) World Health Organization 1992 Publications of the World Health Organization enjoy copyright protection in accordance with the provisions of Protocol 2 of the Universal Copyright Convention. All rights reserved. The designations employed and the presentation of the material in this publication do not imply the expression of any opinion whatsoever on the part of the Secretariat of the World Health Organization concerning the legal status of any country, territory, city or area or of its authorities, or concerning the delimitation of its frontiers or boundaries. The mention of specific companies or of certain manufacturers' products does not imply that they are endorsed or recommended by the World Health Organization in preference to others of a similar nature that are not mentioned. Errors and omissions excepted, the names of proprietary products are distinguished by initial capital letters. CONTENTS ENVIRONMENTAL HEALTH CRITERIA FOR CADMIUM - ENVIRONMENTAL ASPECTS 1. SUMMARY 2. IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, AND ANALYTICAL METHODS 2.1. Physical and chemical properties 2.2. Analytical procedures 2.2.1. Sampling and preparation 2.2.2. Quantitative instrumental methods 3. NATURAL OCCURRENCE AND SOURCES OF ENVIRONMENTAL CONTAMINATION 3.1. Natural occurrence 3.2. Industrial uses 3.3. Sources of environmental cadmium 3.3.1. Sources of atmospheric cadmium 3.3.2. Sources of aquatic cadmium 3.3.3. Sources of terrestrial cadmium 3.4. Environmental transport and distribution 3.4.1. Atmospheric deposition 3.4.2. Transport from water to soil 3.5. Concentrations in various biota 3.5.1. Concentrations in fish 3.5.2. Concentrations in sea-birds 3.5.3. Concentrations in sea mammals 3.6. Concentrations adjacent to highways 3.7. Concentrations from industrial sources 4. KINETICS AND METABOLISM 4.1. Uptake 4.1.1. Uptake from water by aquatic organisms 126.96.36.199 Microorganisms 188.8.131.52 Aquatic molluscs 184.108.40.206 Other aquatic invertebrates 220.127.116.11 Fish 18.104.22.168 Model aquatic ecosystems 22.214.171.124 Uptake from aquatic sediment 126.96.36.199 Uptake from food relative to uptake from water 4.1.2. Uptake by terrestrial organisms 188.8.131.52 Uptake into plants 184.108.40.206 Terrestrial invertebrates 220.127.116.11 Birds 4.2. Distribution 4.2.1. Aquatic organisms 4.2.2. Terrestrial organisms 18.104.22.168 Terrestrial plants 22.214.171.124 Terrestrial invertebrates 4.3. Elimination 4.4. Bioaccumulation and biomagnification 5. TOXICITY TO MICROORGANISMS 5.1. Aquatic microorganisms 5.1.1. Freshwater microorganisms 5.1.2. Estuarine and marine microorganisms 5.2. Soil and litter microorganisms 6. TOXICITY TO AQUATIC ORGANISMS 6.1. Toxicity to aquatic plants 6.2. Toxicity to aquatic invertebrates 6.2.1. Acute and short-term toxicity 126.96.36.199 Effects of temperature and salinity on acute toxicity 188.8.131.52 Effect of water hardness 184.108.40.206 Effect of organic materials and sediment 220.127.116.11 Lifestage sensitivity 18.104.22.168 Other factors affecting acute and short-term toxicity 6.2.2. Long-term toxicity 6.2.3. Reproductive effects 6.2.4. Physiological and biochemical effects 6.2.5. Behavioural effects 6.2.6. Interactions with other chemicals 6.2.7. Tolerance 6.2.8. Model ecosystems 6.3. Toxicity to fish 6.3.1. Acute and short-term toxicity 6.3.2. Reproductive effects and effects on early life stages 6.3.3. Metabolic, biochemical and physiological effects 6.3.4. Structural effects and malformations 6.3.5. Behavioural effects 6.3.6. Interactions with other chemicals 6.4. Toxicity to amphibia 7. TOXICITY TO TERRESTRIAL ORGANISMS 7.1. Toxicity to terrestrial plants 7.1.1. Toxicity to plants grown hydroponically 7.1.2. Toxicity to plants grown in soil 7.1.3. In vitro physiological studies 7.2. Toxicity to terrestrial invertebrates 7.3. Toxicity to birds 7.3.1. Acute and short-term toxicity 7.3.2. Reproductive effects 7.3.3. Physiological effects 7.3.4. Behavioural effects 7.4. Toxicity to wild small mammals 8. EFFECTS IN THE FIELD 8.1. Tolerance 8.2. Effects close to industrial sources and highways 8.3. Effects on fish 8.4. Effects on sea-birds 9. EVALUATION 9.1. General considerations 9.2. The aquatic environment 9.3. The terrestrial environment 10. RECOMMENDATIONS FOR PROTECTING THE ENVIRONMENT 11. FURTHER RESEARCH REFERENCES APPENDIX 1 APPENDIX 2 APPENDIX 3 APPENDIX 4 APPENDIX 5 RESUME RESUMEN WHO TASK GROUP ON ENVIRONMENTAL HEALTH CRITERIA FOR CADMIUM - ENVIRONMENTAL ASPECTS Members Dr L.A. Albert, Consultores Ambientales Asociados, S.C., Xalapa, Veracruz, Mexico Dr J.K. Atherton, Toxic Substances Division, Directorate for Air, Climate and Toxic Substances, Department of the Environment, London, United Kingdom Dr R.W. Elias, Trace Metal Biogeochemistry, Environmental Criteria and Assessment Office, US Environmental Protection Agency, Research Triangle Park, North Carolina, USA Dr A.H. El-Sebae, Faculty of Agriculture, Alexandria University, Alexandria, Egypt Dr R. Koch, Bayer AG, Leverkusen, Germany Professor Y. Kodama, Department of Environmental Health, University of Occupational and Environmental Health, Japan School of Medicine, Yahata Nishi-ku, Kitakyushu City, Japan Dr P. Pärt, Department of Zoophysiology, Uppsala University, Uppsala, Sweden Dr J.H.M. Temmink, Department of Toxicology, Agricultural University, Wageningen, The Netherlands ( Chairman) Secretariat Dr S. Dobson, Institute of Terrestrial Ecology, Monks Wood Experimental Station, Abbots Ripton, Huntingdon, Cambridgeshire, United Kingdom ( Rapporteur) Dr M. Gilbert, International Programme on Chemical Safety, World Health Organization, Geneva, Switzerland ( Secretary) Mr P.D. Howe, Institute of Terrestrial Ecology, Monks Wood Experimental Station, Abbots Ripton, Huntingdon, Cambridgeshire, United Kingdom NOTE TO READERS OF THE CRITERIA DOCUMENTS Every effort has been made to present information in the criteria documents as accurately as possible without unduly delaying their publication. In the interest of all users of the Environmental Health Criteria documents, readers are kindly requested to communicate any errors that may have occurred to the Director of the International Programme on Chemical Safety, World Health Organization, Geneva, Switzerland, in order that they may be included in corrigenda. * * * A detailed data profile and a legal file can be obtained from the International Register of Potentially Toxic Chemicals, Palais des Nations, 1211 Geneva 10, Switzerland (Telephone No. 7988400 or 7985850). ENVIRONMENTAL HEALTH CRITERIA FOR CADMIUM - ENVIRONMENTAL ASPECTS A WHO Task Group on Environmental Health Criteria for Cadmium - Environmental Aspects met at the Institute of Terrestrial Ecology (ITE), Monks Wood, United Kingdom, from 13 to 17 May 1991. Dr M. Roberts, Director, ITE, welcomed the participants on behalf of the host institution and Dr M. Gilbert opened the meeting on behalf of the three cooperating organizations of the IPCS (UNEP/ILO/WHO). The Task Group reviewed and revised the draft criteria document and made an evaluation of the risks for the environment from exposure to cadmium. The first draft of this document was prepared by Dr S. Dobson (ITE). Dr M. Gilbert and Dr P.G. Jenkins, both members of the IPCS Central Unit, were responsible for the technical development and editing, respectively. The efforts of all who helped in the preparation and finalization of the document are gratefully acknowledged. ABBREVIATIONS ALAD delta-aminolevulinic acid dehydratase DPTA diaminopropanoltetraacetic acid EDTA ethylenediaminetetraacetic acid EEC European Economic Community EIFAC European Inland Fisheries Advisory Commission of FAO FAO Food and Agriculture Organization of the United Nations GESAMP Group of Experts on the Scientific Aspects of Marine Pollution MATC maximum acceptable toxicant concentration NOEL no-observed-effect level NTA nitrilotriacetic acid NTEL no-toxic-effect level 1. SUMMARY Cadmium (atomic number 48; relative atomic mass 112.40) is a metallic element belonging, together with zinc and mercury, to group IIb of the periodic table. Some cadmium salts, such as the sulfide, carbonate, and oxide, are practically insoluble in water; these can be converted to water-soluble salts in nature. The sulfate, nitrate, and halides are soluble in water. The speciation of cadmium in the environment is of importance in evaluating the potential hazard. The average cadmium content of sea water is about 0.1 µg/litre or less. River water contains dissolved cadmium at concentrations of between < 1 and 13.5 ng/litre. In remote, uninhabited areas, cadmium concentrations in air are usually less than 1 ng/m3. In areas not known to be polluted, the median cadmium concentration in soil has been reported to be in the range of 0.2 to 0.4 mg/kg. However, much higher values, up to 160 mg/kg soil, are occasionally found. Environmental factors affect the uptake and, therefore, the toxic impact of cadmium on aquatic organisms. Increasing temperature increases the uptake and toxic impact, whereas increasing salinity or water hardness decreases them. Freshwater organisms are affected by cadmium at lower concentrations than marine organisms. The organic content of the water generally decreases the uptake and toxic effect by binding cadmium and reducing its availability to organisms. However, there is evidence that some organic matter may have the opposite effect. Cadmium is readily accumulated by many organisms, particularly by microorganisms and molluscs where the bioconcentration factors are in the order of thousands. Soil invertebrates also concentrate cadmium markedly. Most organisms show low to moderate concentration factors of less than 100. Cadmium is bound to proteins in many tissues. Specific heavy-metal-binding proteins (metallothioneins) have been isolated from cadmium-exposed organisms. The concentration of cadmium is greatest in the kidney, gills, and liver (or their equivalents). Elimination of the metal from organisms probably occurs principally via the kidney, although significant amounts can be eliminated via the shed exoskeleton in crustaceans. In plants, cadmium is concentrated primarily in the roots and to a lesser extent in the leaves. Cadmium is toxic to a wide range of microorganisms. However, the presence of sediment, high concentrations of dissolved salts or organic matter all reduces the toxic impact. The main effect is on growth and replication. The most affected of soil microorganisms are fungi, some species being eliminated after exposure to cadmium in soil. There is selection for resistant strains after low exposure to the metal in soil. The acute toxicity of cadmium to aquatic organisms is variable, even between closely related species, and is related to the free ionic concentration of the metal. Cadmium interacts with the calcium metabolism of animals. In fish it causes hypocalcaemia, probably by inhibiting calcium uptake from the water. However, high calcium concentrations in the water protect fish from cadmium uptake by competing at uptake sites. Zinc increases the toxicity of cadmium to aquatic invertebrates. Sublethal effects have been reported on the growth and reproduction of aquatic invertebrates; there are structural effects on invertebrate gills. There is evidence of the selection of resistant strains of aquatic invertebrates after exposure to cadmium in the field. The toxicity is variable in fish, salmonids being particularly susceptible to cadmium. Sub-lethal effects in fish, notably malformation of the spine, have been reported. The most susceptible life-stages are the embryo and early larva, while eggs are the least susceptible. There is no consistent interaction between cadmium and zinc in fish. Cadmium is toxic to some amphibian larvae, although some protection is afforded by sediment in the test vessel. Cadmium affects the growth of plants in experimental studies, although no field effects have been reported. The metal is taken up into plants more readily from nutrient solutions than from soil; effects have been mainly shown in studies involving culture in nutrient solutions. Stomatal opening, transpiration, and photosynthesis have been reported to be affected by cadmium in nutrient solutions. Terrestrial invertebrates are relatively insensitive to the toxic effects of cadmium, probably due to effective sequestration mechanisms in specific organs. Terrestrial snails are affected sublethally by cadmium; the main effect is on food consumption and dormancy, but only at very high dose levels. Birds are not lethally affected by the metal even at high dosage, although kidney damage occurs. Cadmium has been reported in field studies to be responsible for changes in species composition in populations of microorganisms and some aquatic invertebrates. Leaf litter decomposition is greatly reduced by heavy metal pollution, and cadmium has been identified as the most potent causative agent for this effect. 2. IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, AND ANALYTICAL METHODS 2.1 Physical and chemical properties Cadmium (atomic number 48; relative atomic mass 112.40) is a metallic element belonging, together with zinc and mercury, to group IIb in the periodic table. It is rarely found in a pure state. It is present in various types of rocks and soils and in water, as well as in coal and petroleum. Among these natural sources, zinc, lead, and copper ore are the main sources of cadmium. Cadmium can form a number of salts. Its mobility in the environment and effects on the ecosystem depend to a great extent on the nature of these salts. Since there is no evidence that organocadmium compounds, where the metal is covalently bound to carbon, occur in nature, only inorganic cadmium salts will be discussed. Cadmium may occur bound to proteins and other organic molecules and form salts with organic acids, but in these forms, it is regarded as inorganic. Cadmium has a relatively high vapour pressure. The vapour is oxidized quickly to produce cadmium oxide in the air. When reactive gases or vapour, such as carbon dioxide, water vapour, sulfur dioxide, sulfur trioxide or hydrogen chloride, are present, the vapour reacts to produce cadmium carbonate, hydroxide, sulfite, sulfate or chloride, respectively. These salts may be formed in stacks and emitted to the environment. Some of the cadmium salts, such as the sulfide, carbonate or oxide, are practically insoluble in water. However, these can be converted to water-soluble salts in nature under the influence of oxygen and acids; the sulfate, nitrate, and halogenates are soluble in water. The physical and chemical properties of cadmium and its salts are summarized in Table 1. Equilibrium data for complexes of group IIB cations, comparing cadmium with zinc and mercury, can be found in Table 2. A diagrammatic representation of the capacity of soil types for metals is given in Fig. 1. The speciation of cadmium in soil water (Fig. 2) and surface water (Fig. 3) is important for the evaluation of its potential hazard. Most of the cadmium found in mammals, birds, and fish is probably bound to protein molecules. Table 1. Physical and chemical properties of cadmium and its salts Cadmium Cadmium Cadmium Cadmium Cadmium Cadmium Cadmium Cadmium chloride acetate oxide hydroxide sulfide sulfate sulfite CAS number 7440-43-9 10108-64-2 543-90-8 1306-19-0 1306-23-6 10124-36-4 Empirical formula Cd CdCl2 C4H6CdO4 CdO Cd(OH)2 CdS CdSO4 CdSO3 Relative atomic or molecular mass 112.41 183.32 230.50 128.40 146.41 144.46 208.46 192.46 Relative density 8.642 4.047 2.341 6.95 4.79 4.82 4.691 Melting point (°C) 320.9 568 256 < 1426 300 1750 1000 decomposes (decomposes) Boiling point (°C) 765 960 decomposes 900-1000 (decomposes) Water solubility insoluble 1400 very soluble insoluble 0.0026 0.0013 755 slightly soluble (g/litre) (20 °C) (26 °C) (18 °C) (0 °C) Table 2. Equilibrium data for complexes of group IIB cations a System Metal log K1 DELTA H1 DELTA S1 (kJ mol-1) (J K-1 mol-1) zinc 5.0 b 0 b 105 M2+-OH- cadmium 3.9 b 0 79 mercury 10.6 b - - zinc 0.8 7.5 42 M2+-F- cadmium 0.6 4.2 25 mercury 1.0 c 4.2 c 33 c zinc - 0.2 5.4 16 M2+-Cl- cadmium 1.5 - 0.4 29 mercury 7.1 - 24.3 54 zinc - 0.6 1.7 - 4 M2+-Br- cadmium 1.7 - 4.2 21 mercury 9.4 - 40.1 46 zinc - 1.5 - - M2+-I- cadmium 2.1 - 9.2 8 mercury 12.9 c - 75.3 c - 8 c zinc 5.3 - - M2+-CN- cadmium 5.6 - 30.5 b 13 b mercury 18.0 c - 96 b 0 b zinc 0.7 d - 5.9 d - 4 d M2+-SCN- cadmium 1.3 d - 9.6 d - 8d mercury 9.1 d - 49.7 d 8 zinc 1.9 - - M2+-S2O32- e cadmium 4.7 - 6.3 d 67 d mercury 29.9 d - - zinc 2.4 f - 10.9 f 8 f M2+-NH3 cadmium 2.7 f - 14.6 f 4 f mercury 8.8 f - - zinc 4.8 c - 11.3 g 59 g 2+ - cadmium 4.1 d - 8.8 b 50 g (glycinate)- mercury 10.3 c - - zinc 16.4 - 20.5 247 M2+-(EDTA)4- cadmium 16.4 - 38.1 184 mercury 21.5 - 79.0 146 a From: Aylett (1979). Data, which refer to first stepwise stability constant, [ML]/[M][L], unless otherwise stated, are from Sillen (1964) and Smith & Martell (1974, 1975, 1976); see also Christensen et al. (1975). All values refer to measurements in water at 25 °C; the ionic strength is 3 mol/litre unless otherwise stated. b ionic strength 0 c ionic strength 0.5 mol/litre d ionic strength 1.0 mol/litre e Data refer to overall stability constant, ß2 = [ML2]/[M][L]2 f ionic strength 2.0 mol/litre g ionic strength 0.1 mol/litre 2.2 Analytical procedures The following is an outline of the analytical procedure for cadmium; further information is given in Environmental Health Criteria 134: Cadmium (WHO, 1992). 2.2.1 Sampling and preparation Only a few nanograms, or even less, of cadmium may be present in collected samples of air or water, whereas hundreds of micrograms may be present in small samples of kidney, sewage sludge, and plastics. Different techniques are, therefore, required for the collection, preparation, and analysis of the samples. In general, the techniques available for measuring cadmium in the environment and biological materials cannot differentiate between cadmium species. With special separation techniques, cadmium-containing proteins can be isolated and identified. In most studies, the concentration or amount of cadmium in water, air, soil, plants, and other environmental or biological material is determined as the element. Standard trace element methods can generally be used for the collection of samples (LaFleur, 1976; Behne, 1980). During the handling and storage of samples, particularly liquid samples, special care must be taken to avoid contamination; coloured materials in containers, especially plastics and rubber, should be avoided. Glass and transparent, cadmium-free polyethylene, polypropylene or teflon containers are usually considered suitable for storing samples. All containers and glassware should be precleaned in dilute nitric acid and deionised water. In order to avoid possible adsorption of cadmium onto the container wall, water samples or standards with low cadmium concentrations should not be stored for long periods of time. To prepare samples for analysis, inorganic solid samples (such as soil or dust samples) are usually dissolved in an acid, e.g., nitric acid. Organic samples need to be subjected to wet ashing (digested) or dry ashing. When the cadmium concentration is low, special treatment is sometimes needed. The procedures for separating cadmium from interfering compounds and concentrating the samples are very important steps in obtaining accurate results. 2.2.2 Quantitative instrumental methods The most commonly used methods, at present, are atomic absorption spectrometry, electrochemical methods, neutron activation analysis, atomic emission spectrometry, atomic fluorescence spectrometry and proton-induced X-ray emissions (PIXE) analysis. Analytical methods for cadmium have been reviewed by Friberg et al. (1986). Detection limits of some of the methods are given in Table 3. Table 3. Analytical procedures a Method Detection limit Matrix Atomic absorption 1 to 5 mg/litre water spectrometry 0.1 mg/kg biological samples electrothermal a few pg atomization Electrochemical method (potentiometric stripping analysis) 0.1 mg/litre urine Neutron activation 0.1 to 1 mg/litre biological analysis samples/fluids X-ray atomic 17 mg/kg biological samples fluorescence a From: Friberg et al. (1986) 3. NATURAL OCCURRENCE AND SOURCES OF ENVIRONMENTAL CONTAMINATION 3.1 Natural occurrence A comparison of natural and anthropogenic sources of trace metals is given in the Appendix 1. Cadmium is widely distributed in the earth's crust at an average concentration of about 0.1 mg/kg and is commonly found in association with zinc. However, higher levels are present in sedimentary rocks: marine phosphates often contain about 15 mg/kg (GESAMP, 1984). Weathering and erosion result in the transport by rivers of large quantities of cadmium to the world's oceans and this represents a major flux of the global cadmium cycle; an annual gross input of 15 000 tonnes has been estimated (GESAMP, 1987). In background areas away from ore bodies, surface soil concentrations of cadmium typically range between 0.1 and 0.4 mg/kg (Page et al., 1981). The median cadmium concentration in non-volcanic soil ranges from 0.01 to 1 mg/kg, but in volcanic soil levels of up to 4.5 mg/kg have been found (Korte, 1983). Volcanic activity is a major natural source of atmospheric cadmium release. The global annual flux from this source has been estimated to be 100-500 tonnes (Nriagu, 1979). Deep sea volcanism is also a source of environmental cadmium release, but the role of this process in the global cadmium cycle remains to be quantified. The average cadmium content of sea water is about 0.1 µg/litre or less (Korte, 1983), while river water (Mississippi, Yangtze, Amazon, and Orinoco sampled between 1976 and 1982) contains dissolved cadmium at concentrations of < 1.1-13.5 ng/litre (Shiller & Boyle, 1987). Cadmium levels of up to 5 mg/kg have been reported in river and lake sediments and from 0.03 to 1 mg/kg in marine sediments (Korte,1983). Current measurements of dissolved cadmium in surface waters of the open oceans give values of < 5 ng/litre. The vertical distribution of dissolved cadmium in ocean waters is characterized by a surface depletion and deep water enrichment, which corresponds to the pattern of nutrient concentrations in these areas (Boyle et al., 1976). This distribution is considered to result from the absorption of cadmium by phytoplankton in surface waters and its transport to the depths, incorporation to biological debris, and subsequent release. In contrast, cadmium is enriched in the surface waters of areas of upwelling and this also leads to elevated levels in plankton unconnected with human activity (Martin & Broenkow, 1975; Boyle et al., 1976). Oceanic sediments underlying these areas of high productivity can contain markedly elevated cadmium levels as a result of inputs associated with biological debris (Simpson, 1981). In remote, uninhabited areas, cadmium concentrations in air are usually less than 1 ng/m3 (Korte,1983). 3.2 Industrial uses The principal applications of cadmium fall into five categories: protective plating on steel; stabilizers for PVC; pigments in plastics and glass; electrode material in nickel-cadmium batteries; and as a component of various alloys (Wilson, 1988). The relative importance of the major applications has changed considerably over the last 25 years. The use of cadmium for electroplating represented in 1960 over half the cadmium consumed worldwide, but in 1985 its share was less than 25% (Wilson, 1988). This decline is usually linked to the introduction of stringent effluent limits from plating works and, more recently, to the introduction of general restrictions on cadmium consumption in certain countries. In contrast, the use of cadmium in batteries has shown considerable growth in recent years from only 8% of the total market in 1970 to 37% by 1985. The use of cadmium in batteries is particularly important in Japan and represented over 75% of the total consumption in 1985 (Wilson, 1988). Pigments and stabilizers accounted for 22% and 12% of the total world consumption in 1985. The share of the market by cadmium pigments remained relatively stable between 1970 and 1985 but the use of the metal in stabilizers during this period showed a considerable decline, largely as a result of economic factors. The use of cadmium as a constituent of alloys is relatively small and has also declined in importance in recent years, accounting for about 4% of total cadmium use in 1985 (Wilson, 1988). 3.3 Sources of environmental cadmium 3.3.1 Sources of atmospheric cadmium Estimates of cadmium emissions to the atmosphere from human and natural sources have been carried out at the worldwide, regional, and national level; examples of such inventories are shown in Table 4. The median global total emission of the metal from human sources in 1983 was 7570 tonnes (Nriagu & Pacyna, 1988) and represented about half the total quantity of cadmium produced in the same year. In both the European Economic Community (EEC) and on a worldwide scale (Nriagu, 1989), about 10-15% of total airborne cadmium emissions arise from natural processes, the major source being volcanic action. Municipal refuse contains cadmium derived from discarded nickel-cadmium batteries and plastics containing cadmium pigments and stabilizers. The incineration of refuse is a major source of atmospheric cadmium release at country, regional, and worldwide level (Table 4). Steel production can also be considered as a waste-related source, as large quantities of cadmium-plated steel scrap are recycled by this industry. As a result, steel production is responsible for considerable emissions of atmospheric cadmium. 3.3.2 Sources of aquatic cadmium Non-ferrous metal mines represent a major source of cadmium release to the aquatic environment. Contamination can arise from mine drainage water, waste water from the processing of ores, overflow from the tailings pond, and rainwater run-off from the general mine area. The release of these effluents to local watercourses can lead to extensive contamination downstream of the mining operation. Mines disused for many years can still be responsible for the continuing contamination of adjacent watercourses (Johnson & Eaton, 1980). At the global level, the smelting of non-ferrous metal ores has been estimated to be the largest human source of cadmium release to the aquatic environment (Nriagu & Pacyna, 1988). Discharges to fresh and coastal waters arise from liquid effluents produced by air pollution control (gas scrubbing) together with the site drainage waters. Table 4. Estimates of atmospheric cadmium emissions (tonnes/year) on a national, regional and worldwide basis Source United EEC b Worldwide c Kingdom a Natural sources ND 20 150-2600 d Non-ferrous metal production mining ND ND 0.6-3 zinc and cadmium 20 920-4600 copper 3.7 6 1700-3400 lead 7 39-195 Secondary production ND 2.3-3.6 Production of cadmium-containing substances ND 3 ND Iron and steel production 2.3 34 28-284 Fossil fuel combustion coal 1.9 6 176-882 oil 0.5 41-246 Refuse incineration 5 31 56-1400 Sewage sludge incineration 0.2 2 3-36 Table 4 (contd). Source United EEC b Worldwide c Kingdom a Phosphate fertilizer manufacture ND ND 68-274 Cement manufacture 1 ND 8.9-534 Wood combustion ND ND 60-180 TOTAL EMISSIONS 14 130 3350-14 640 a From: Hutton & Symon (1986); data apply to 1982-1983 b From: Hutton (1983); data apply to 1979-1980 (the EEC consisted, at that time, of Belgium, Denmark, Federal Republic of Germany, Italy, Luxembourg, The Netherlands, Republic of Ireland, and the United Kingdom) c From: Nriagu & Pacyna (1988); data apply to 1983 d From: Nriagu (1979) ND Not determined The manufacture of phosphate fertilizer results in a redistribution of the cadmium present in the rock phosphates between the phosphoric acid product and gypsum waste. In many cases, the gypsum is disposed of by dumping in coastal waters, which leads to considerable cadmium inputs. Some countries, however, recover the gypsum for use as a construction material and thus have negligible cadmium discharges (Hutton, 1982). The atmospheric fallout of cadmium to fresh and marine waters represents a major input of cadmium at the global level (Nriagu & Pacyna, 1988). A GESAMP study of the Mediterranean Sea indicated that this source is comparable in magnitude to the total river inputs of cadmium to the region (GESAMP, 1985). Similarly, large cadmium inputs to the North Sea (110-430 tonnes/year) have also been estimated, based on the extrapolation from measurements of cadmium deposition along the coast (van Alst et al., 1983a,b). However, another approach based on model simulation yielded a modest annual cadmium input of 14 tonnes (Krell & Roeckner, 1988). Acidification of soils and lakes may result in enhanced mobilization of cadmium from soils and sediments and lead to increased levels in surface and ground waters (WHO Working Group, 1986). 3.3.3 Sources of terrestrial cadmium Solid wastes are disposed of in landfill sites, resulting in large cadmium inputs at the national and regional levels when expressed as total tonnage (Hutton, 1982; Hutton & Symon, 1986). Sources include the ashes from fossil fuel combustion, waste from cement manufacture, and the disposal of municipal refuse and sewage sludge. Of greater potential environmental significance are the solid wastes from both non-ferrous metal production and the manufacture of cadmium-containing articles, as well as the ash residues from refuse incineration. These three waste materials are characterized by elevated cadmium levels and as such require disposal to controlled sites to prevent the contamination of the ground water. The agricultural application of phosphate fertilizers represents a direct input of cadmium to arable soils. The cadmium content of phosphate fertilizers varies widely and depends on the origin of the rock phosphate. It has been estimated that fertilizers of West African origin contain 160-255 g cadmium/tonne of phosphorus pentoxide, while those derived from the southeastern USA contain 35 g/tonne (Hutton, 1982). The annual rate of cadmium input to arable land from phosphate fertilizers has been estimated at 5 g/ha for the countries of the EEC (Hutton 1982). This only represents about 1% of the surface soil cadmium burden. Despite the relatively small size of this input, long-term continuous application of phosphate fertilizers has been shown to cause increased soil cadmium concentrations (Williams & David, 1973, 1976; Andersson & Hahlin, 1981). The application of municipal sewage sludge to agricultural soils as a fertilizer can also be a significant source of cadmium; a value of 80 g/ha has been estimated for the United Kingdom (Hutton & Symon, 1986). On a national or regional basis, however, these inputs are much smaller than those from either phosphate fertilizers or atmospheric deposition (see section 3.4). Polluted soils can contain cadmium levels of up to 57 mg/kg (dry weight) resulting from sludge applied to soil and up to 160 mg/kg in the vicinity of metal-processing industry (Fleischer et al., 1974). The highest cadmium levels reported appear to be from ancient mining areas with levels of up to 468 mg/kg. 3.4 Environmental transport and distribution 3.4.1 Atmospheric deposition Cadmium is removed from the atmosphere by dry deposition and by precipitation. In rural areas of Scandinavia, annual deposition rates of 0.4-0.9 g/ha have been measured (Laamanen, 1972; Andersson, 1977). Similarly, in a rural region of Tennessee, USA, a deposition rate of 0.9 g/ha was observed (Lindberg et al., 1982). Hutton (1982) suggested that 3 g/ha per year was a representative value for the atmospheric deposition of cadmium to agricultural soils in rural areas of the EEC. The corresponding input for these areas from the application of phosphate fertilizers is 5 g/ha per year (see section 3.3). Many industrial sources of cadmium possess tall stacks which bring about the wide dispersion and dilution of particulate emissions. Nevertheless, cadmium deposition rates around smelter facilities are often markedly elevated nearest the source and generally decrease rapidly with distance (Hirata, 1981). Soil cadmium concentrations in excess of 100 mg/kg are commonly encountered close to long established smelters (Buchauer, 1972). Crop plants growing near to atmospheric sources of cadmium may contain elevated cadmium levels (Carvalho et al., 1986). However, it is not always possible to distinguish whether the cadmium is derived directly from surface deposition or originates from root uptake, since soil levels in such areas are generally higher than normal. 3.4.2 Transport from water to soil Rivers contaminated with cadmium can contaminate surrounding land, either through irrigation for agricultural purposes, by the dumping of dredged sediments, or through flooding (Forstner, 1980; Sangster et al., 1984). For example, agricultural land adjacent to the Neckar River, Germany, received dredged sediments to improve the soil, a practice that produced soil cadmium concentrations in excess of 70 mg/kg (Forstner, 1980). Much of the cadmium entering fresh waters from industrial sources is rapidly adsorbed by particulate matter, where it may settle out or remain suspended, depending on local conditions. This can result in low concentrations of dissolved cadmium even in rivers that receive and transport large quantities of the metal (Yamagata & Shigematsu, 1970) 3.5 Concentrations in various biota Table 5 indicates the levels of cadmium found in various biota (Eisler, 1985). Eisler (1985) concluded that there are at least six trends evident from the abundant residue data available for cadmium. * Marine organisms generally contain higher cadmium residues than their freshwater and terrestrial counterparts. * Cadmium tends to concentrate in the viscera of vertebrates, especially the liver and kidneys. * Cadmium concentrations are generally higher in older organisms. * Higher cadmium residues are generally associated with industrial and urban sources, although this does not apply to sea birds and sea mammals. * Cadmium residues in plants are normally less than 1 mg/kg. However, plants growing in soil amended with cadmium (e.g., from sewage sludge) may contain significantly higher levels. * The species analysed, season of collection, ambient cadmium levels, and the sex of the organism probably all affect the residue level. Table 5. Concentrations of cadmium in biota Organisms Parts of the Cadmium concentration organisms (mg/kg dry weight) Marine organisms Algae < 1 to 16 Molluscs soft parts up to 425 kidney up to 547 liver up to 782 digestive gland up to 1163 Crustaceans whole body < 0.4-6.2 Annelids whole body 0.1-3.6 Fish whole body up to 5.2 Birds kidney up to 231 Mammals kidney up to 300 Freshwater organisms Plants whole plant 0.5-1.8 roots up to 6.7 Molluscs soft parts; fresh weight 0.2-1.4 Annelids whole body; fresh weight 0.5-3.2 Fish whole body; fresh weight 0.01-1.04 Table 5 (contd). Organisms Parts of the Cadmium concentration organisms (mg/kg dry weight) Terrestrial organisms Plants whole plant up to 27.1 grain up to 257 Annelids whole body 3-12.6 Birds whole body; fresh weight < 0.05-0.24 kidney; fresh weight up to 7.4 Mammals kidney up to 8.1 3.5.1 Concentrations in fish May & McKinney (1981) monitored freshwater fish from the USA in 1976 and 1977 and found cadmium concentrations ranging from 0.01 to 1.04 mg/kg (wet weight), the mean being 0.085 mg/kg. This represented a significant decline from the mean 1972 concentration of 0.112 mg/kg. The authors pointed out that this decline parallels a decline in cadmium metal production and consumption over the same period. Hardisty et al. (1974a) sampled flounder ( Platichthyes flesus) from the Severn estuary, United Kingdom, and found mean cadmium concentrations of 3.4-7.3 mg/kg (dry weight). No overall correlation between cadmium concentration and length or age was observed, although the largest (27-29 cm) and the oldest („ 5 years) fish gave the highest mean concentrations. Hardisty et al. (1974b) found a positive correlation between the cadmium content of a variety of fish species and the crustacea content of their diet. Lovett et al. (1972) sampled fish from New York State, USA, and reported mean cadmium concentrations of < 10-142.7 µg/kg (fresh weight). There was no relationship between total residues and size, sex or age of lake trout ( Salvelinus namaycush). 3.5.2 Concentrations in sea-birds Cadmium has been found in a wide variety of birds, and particularly high levels have been reported in pelagic sea-birds. Much of the cadmium occurs in the kidney and liver, and relatively little is transferred to the eggs. A review of the uptake of cadmium and of the factors that affect it can be found in Scheuhammer (1987). Interestingly, the concentrations of cadmium in sea-birds are often higher in areas with little or no contamination from industrial sources (Bull et al., 1977; Hutton, 1981; Osborn & Nicholson, 1984). 3.5.3 Concentrations in sea mammals High levels of cadmium have been reported in sea mammals from areas around the world, which they are assumed to take up from their diet of fish. Roberts et al. (1976) showed that kidney levels of cadmium in the common seal off the United Kingdom coast were age related. Drescher et al. (1977) showed a similar relationship in seals off the German coast and Hamanaka et al. (1982) in stellar sea lions off the coast of Japan. Similar trends in dolphins and porpoises have been reported (Falconer et al., 1983; Honda & Tatsukawa, 1983; Honda et al., 1986). Muir et al. (1988) sampled white-beaked dolphins ( Lagenorhynchus albirostris) and pilot whales ( Globicephala melaena) from the coast of Newfoundland, Canada, and reported mean cadmium levels in kidney (dry weight) of 13.6 mg/kg and 108 mg/kg, respectively. Cadmium concentrations were age related in pilot whales. The lower levels found in dolphins were probably related both to species differences and to the fact that they were all young animals. 3.6 Concentrations adjacent to highways Muskett & Jones (1980) monitored levels of cadmium adjacent to a heavily used road. The concentrations in air were highest at a distance from the road of 0-10 m, and a similar pattern was found in soil. Cadmium levels in earthworms sampled at known distances from a highway revealed levels of 12.6 mg/kg (dry weight) within 3 m falling to 7.1 mg/kg approximately 50 m from the highway. The level in earthworms from control sites was 3 mg/kg (Gish & Christensen, 1973). The land snail Cepaea hortensis accumulates cadmium from roadside verges (Williamson, 1980). The highest concentration of cadmium was found in the digestive gland (40.3 mg/kg dry weight) and kidney (12.8 mg/kg dry weight). There was little metal in the head and foot, which make up most of the body tissue. The author showed that age accounted for 80% of the total variance of soft tissue body burdens. The cadmium body burdens were found to be effectively immobile, accumulating progressively with age. 3.7 Concentrations from industrial sources Burkitt et al. (1972) analysed the cadmium content of ryegrass at various distances from a zinc smelter and found 50, 10.8, and 1.8 mg/kg dry weight at distances of 0.3, 1.9, and 11.3 km, respectively, from the smelter. Teraoka (1989) found that cadmium levels in rice roots were significantly higher in industrial urban and roadside areas of Japan compared to sparsely populated areas. The mean level in industrial areas was 10 mg/kg (dry weight). Beyer et al. (1985) monitored biota from the vicinity of two zinc smelters in eastern Pennsylvania, USA. Cadmium concentrations were highest in carrion insects (25 mg/kg dry weight), followed by fungi (9.8 mg/kg), leaves (8.1 mg/kg), shrews (7.3 mg/kg), moths (4.9 mg/kg), mice (2.6 mg/kg), songbirds (2.5 mg/kg), and berries (1.2 mg/kg). Van Hook (1974) sampled soil and earthworms from soil that had not been disturbed for 30 years and reported mean cadmium levels in the soils and earthworms of 0.35 and 5.7 mg/kg dry weight, respectively. Ma et al. (1983) analysed soil and earthworms ( Lumbricus rubellus) at varying distances from a zinc-smelting plant. Cadmium concentrations ranged from 0.1 to 5.7 mg/kg for the soil and 20 to 202 mg/kg for the worms, and there was a correlation between decreasing distance from the smelter and increasing cadmium levels. Pietz et al. (1984) sampled soil and earthworms ( Aporrectodea tuberculata) and ( Lumbricus terrestris) from mine soil and non-mine soil, either amended or not with sewage sludge. Soil and worms from mine soil gave residues of 0.6 and 3.8 mg/kg dry weight, respectively, in non-amended soil and 2 and 22 mg/kg in sludge-amended soil. Residues in soil and worms from non-mined soil were 1 and 12 mg/kg for non-amended and 3.5 and 36 mg/kg for sludge-amended soil, respectively. The much lower capacity of worms from areas already contaminated with cadmium to take up the metal suggests some selection for varieties that control metal uptake. Morgan & Morgan (1988) sampled earthworms ( Lumbricus rubellus and Dendrodrilus rubidus) from one uncontaminated site and fifteen metal-contaminated sites (in the vicinity of disused non-ferrous metalliferous mines) in the United Kingdom. Cadmium concentrations in the worms ranged from 8 mg/kg (dry weight) to 1786 mg/kg; they were generally higher than soil levels, and the total soil cadmium explained 82% to 86% of the variability in earthworm cadmium concentrations. The authors found some evidence that cadmium accumulation was suppressed in extremely organic soils. Martin et al. (1980) reported cadmium levels in a variety of invertebrates sampled from sites contaminated by airborne cadmium. The woodlouse was shown to accumulate cadmium principally in the hepatopancreas. Van Straalen & van Wensem (1986) analysed 13 species of arthropods from an area polluted by zinc factory emissions. They found no effect of body size or trophic level on the cadmium content of the arthropods. Roberts & Johnson (1978) sampled invertebrates and their diet from the area of an abandoned lead-zinc mine in the United Kingdom. They found cadmium levels higher in herbivorous invertebrates than in the vegetation on which they fed (but not markedly so). There were much higher levels of cadmium in carnivorous invertebrates, suggesting that cadmium might have a capacity for accumulation in food chains. In contrast to mercury levels, total cadmium body burdens were higher in sparrows ( Passer domesticus) caught in industrialised areas of Poland than in those caught in agricultural regions (Pinowska et al., 1981). Pigeon brain, liver, and kidney sampled in rural, suburban, and urban areas gave a good indication of the level of environmental pollution with cadmium (Hutton & Goodman, 1980). Hunter & Johnson (1982) monitored small mammals near to an industrial works complex and found that cadmium accumulated particularly in the liver and kidney. Cadmium levels in the liver ranged rom 1.5 to 280 mg/kg (dry weight) and in the kidney from 7.4 to 193 mg/kg. Small mammals from unpolluted sites contained liver levels ranging from 0.5 to 25 mg/kg and kidney levels of 1.5-26 mg/kg. The insectivorous common shrew ( Sorex areneus) was found to be a more prominent accumulator of cadmium than omnivorous and herbivorous small mammals, based on body burden to dietary metal concentration ratios. Similar results were obtained by Andrews et al. (1984) who monitored cadmium levels in the herbivorous short-tailed field vole ( Microtus agrestis) and the insectivorous common shrew ( S. araneus) from a revegetated metalliferous mine site. Mean cadmium concentrations were 1.84 mg/kg (dry weight) and 52.7 mg/kg for voles and shrews, respectively, values that were significantly higher than those found in control sites. 4. KINETICS AND METABOLISM Appraisal In aquatic systems, cadmium is most commonly taken up by organisms directly from water, but may also be ingested with substantially contaminated food. The free metal ion, Cd2+, is the form most available to aquatic species. Uptake from water may be reduced by the concentration of calcium and magnesium salts (water hardness). Cadmium uptake from sea water may be greatly reduced by the formation of less available complexes with chloride. Organic complexes with cadmium can be classified in three groups: those that are unavailable (e.g., EDTA, NTA, DPTA), those that are available but less so than the free Cd2+ (e.g., fulvic acids of low relative molecular mass), and those that form readily available hydrophobic complexes with cadmium (xanthates and dithiocarbamates). Organisms in the freshwater environment are contaminated according to their ability to absorb or adsorb cadmium from the water, rather than to their position in the food chain. Consequently, differences in cadmium concentration between species at the same trophic level are common and there is no evidence for biomagnification. Conversely, marine organisms take up cadmium principally from food. The primary source of cadmium in terrestrial systems is the soil, and uptake follows the typical food chain pathway, although deposition of cadmium on plant and animal surfaces can account for some additional contamination at each trophic level. Variations in uptake and retention occur, and there is some evidence for biomagnification in carnivores. Organisms that feed on sediment or detritus may accumulate more cadmium than those in the grazing food chain. High levels of cadmium have been reported in sea mammals, pelagic sea-birds, and terrestrial invertebrates. Within a variety of organisms, cadmium is distributed throughout most tissues, but tends to accumulate in the roots, gills, livers, kidneys, hepatopancreas, and exoskeleton. Cadmium in the cell is often bound to cytoplasmic proteins, a possible detoxifying mechanism. Elimination probably occurs primarily via the kidney but also via moulting of the exoskeleton. There is some evidence of an interaction between cadmium and other metals, especially calcium and zinc. Cadmium may replace calcium on the calcium-specific protein calmodulin and is affected by other physiological processes that regulate the uptake of calcium. In certain circumstances, zinc increases cadmium retention in the liver and kidneys of aquatic vertebrates. In terrestrial systems, high soil zinc levels can reduce cadmium uptake appreciably. Selection can lead to cadmium-tolerant populations in both the aquatic and terrestrial environments. 4.1 Uptake 4.1.1 Uptake from water by aquatic organisms Several studies have shown that the free metal ion, Cd2+, is the form of cadmium most available to aquatic organisms (Sunda et al., 1978; Borgmann, 1983; Part et al., 1985; Sprague, 1985). Inorganic cadmium complexes appear not to be taken up, at least by fish (Part et al., 1985). This is particularly important in marine water where cadmium is mainly present in soluble chloride complexes (Zirino & Yamamoto 1972). It is most probable that chloride complexation is responsible for the reduced cadmium accumulation and toxicity in a variety of organisms observed with increasing salinities (Coombs, 1979). In the case of organic cadmium complexes, the chemical properties are of importance with respect to bioavailability. Three categories can be distinguished. The first comprises cadmium complexes with EDTA, NTA, and DPTA, which are unavailable to aquatic organisms (Sunda et al., 1978; Part & Wikmark, 1984). The second consists of complexes that to some extent contribute to the total metal uptake, i.e. uptake is higher than predicted from the actual Cd2+ activity, but the complex is still less available than the free Cd2+ ion. This group includes fulvic acids of low relative molecular mass (Giesy et al., 1977; John et al., 1987), the amino acid histidine (Pecon & Powell, 1981), and carboxylic acids like citric acid (Guy & Ross Kean, 1980; Part & Wikmark, 1984). The third category includes compounds such as xanthates and dithiocarbamates that form hydrophobic complexes with heavy metals. These hydrophobic complexes act as metal carriers across biological membranes and they lead to a greater uptake of cadmium in aquatic organisms than when the metal is present as the free ion (Poldoski, 1979; Block & Part, 1986; Gottofrey et al., 1988; Block, 1991). This latter observation is of particular environmental concern because xanthates are used in the mining industry in the enrichment of metals from sulfide ores by flotation. Xanthate concentrations of between 4 and 400 µg/litre have been measured in waters receiving effluent from metal refineries (enrichment plants) (Waltersson, 1984). Another water quality parameter affecting cadmium uptake is the Ca2+ and Mg2+ concentration (hardness) of the water. Increasing Ca2+ concentration reduces cadmium uptake through fish gills (Part et al., 1985; Wicklund, 1990), cadmium accumulation (Carroll et al., 1979), and cadmium toxicity for fish (Calamari et al., 1980). Two mechanisms can be distinguished for the Ca2+-mediated reduction in cadmium uptake. The first is an inhibitory effect on uptake into gill tissue, while the second is related to the adaptive response of the fish to increased Ca2+ concentrations (Calamari et al., 1980, Wicklund 1990). Mg2+ also reduces cadmium uptake through fish gills but at 5 times higher concentrations than Ca2+ (Part et al., 1985). Cadmium uptake in fish is not strongly pH dependent; uptake in rainbow trout gills was not affected over the pH range 5-7 (Part et al., 1985). Recent data from fish gills indicate that, to some extent, Cd2+ shares uptake mechanisms with Ca2+; these two ions are about the same size and also form complexes with the same kind of ligands. Thus Cd2+ can replace Ca2+ in the calcium-specific protein calmodulin (Flik et al., 1987). In the gills, Cd2+ is assumed to enter the epithelial cells down its concentration and electrical gradient by facilitated diffusion through a calcium channel in the apical membrane (Verbost et al., 1989). Several lines of evidence support this assumption. Firstly, increasing water Ca2+ concentrations reduce cadmium uptake. Secondly, cadmium in the water inhibits Ca2+ uptake in the gills (Verbost et al., 1987; Reid & McDonald, 1988). Thirdly, La3+, a calcium channel blocker in cell membranes, inhibits both Ca2+ and Cd2+ uptake in the gills. Fourthly, the hypocalcaemic hormone stanniocalcin reduces both Ca2+ and Cd2+ uptake in the gills (Verbost et al., 1989). Stanniocalcin has been shown to close the apical calcium channel in the gill epithelial cells thereby reducing Ca2+ uptake from the water (Lafeber et al., 1988). The hormone is secreted when the fish has a surplus of Ca2+, i.e. hypercalcaemic. The two-fold effect of Ca2+ on cadmium uptake in fish discussed previously can be well explained by this model. A direct competition between Ca2+ and Cd2+ at the apical calcium channel reduces the uptake of cadmium into the cells, while the adaptive response in Ca2+-rich water probably involves an increased stanniocalcin level, which closes the apical calcium/cadmium channel. The transport mechanism from the epithelial cells to the blood is unclear. Cadmium is not transported by the high affinity Ca-ATPase in the basolateral epithelial membrane which transports Ca2+ (Verbost et al., 1988). The possible involvement of the Na+/Ca2+ exchange mechanism, where Cd2+ replaces Ca2+, has recently been suggested as a translocation mechanism to the blood (personal communication to the IPCS by G. Flik). Zinc also has been shown to reduce cadmium uptake through the gills (Wicklund, 1990). Like cadmium, zinc is assumed to enter the epithelial cell by facilitated diffusion (Spry & Wood, 1989) and, furthermore, Ca2+ acts antagonistically on zinc uptake. Taken together, these data suggest that the apical epithelial membrane of fish gills contains an ion channel shared by cadmium and calcium, and probably also zinc. The movement of metals through this channel is controlled both by external factors such as the Ca2+ content of the water and internal factors such as hormones. Increasing temperature increases the uptake of cadmium from water (Vernberg et al., 1974; Zaroogian & Cheer, 1976; Denton & Burdon-Jones, 1981). 22.214.171.124 Microorganisms In the alga Chlorella pyrenoidosa, uptake of cadmium was completely blocked by 0.2 mg manganese/litre and inhibited by 2 to 5 mg iron/litre, but calcium, magnesium, molybdenum, copper, zinc, and cobalt had no effect on uptake (Hart & Scaife, 1977). Cultures of Chlorella accumulate twice as much cadmium at pH 7.0 as at pH 8.0 when exposed to 0.5 mg cadmium/litre (Hart & Scaife, 1977). 126.96.36.199 Aquatic molluscs Hardy et al. (1984) found greater uptake of cadmium from sea water into oysters given an uncontaminated phytoplankton food source than into those without food. The authors explain their findings on the basis that the presence of phytoplankton increases the flow of water through the oysters. Studies on oysters without a food source may thus underestimate cadmium uptake. Oysters fed phytoplankton containing cadmium retained only 0.59% of this cadmium; the majority of the cadmium in molluscs is taken up directly from the water. The oyster accumulates about twice as much cadmium in summer as in the winter. This is presumed to reflect the increased flow of water through the animal at higher temperatures (Zaroogian & Cheer, 1976). Hardy et al. (1981) showed that clams ( Protothaca staminea) took up much less cadmium from water in the presence of sediment at 3.6 g/litre. The uptake was only 17% of that measured in sediment-free water. Langston & Zhou (1987a,b) found no evidence of cadmium uptake into the bivalve Macoma balthica involving metallothionein or metallothionein-like proteins. Accumulation in soft tissues was linear throughout a 29-day exposure period, whereas uptake onto the shell was characterized as saturation kinetics. In contrast, the gastropod Littorina littorea did show induction of specific cadmium-binding proteins, which contributed to uptake and storage of cadmium. Watling & Watling (1983) demonstrated uptake of cadmium in a dose-dependant manner into sandy beach gastropod molluscs in laboratory experiments. Much of the cadmium (as chloride) accumulated in the gill. The rate of cadmium uptake was 0.01 mg/kg per day for Donax serra and 0.16 mg/kg per day for the smaller Bullia rhodostoma after exposure to cadmium at 20 µg/litre. The freshwater snail Physa integra took up more cadmium as exposure increased, concentrations ranging between 1 and 40 µg/litre. The highest concentration factors were found with the lowest exposure concentration (Spehar et al., 1978a). Wier & Walter (1976) exposed the freshwater snail Physa gyrina to 1.3 mg cadmium/litre (as the chloride) and found an average cadmium uptake rate of 0.55 mg/kg per hour over 24 h. Heavier snails took up less cadmium, after the same exposure, than lighter individuals. 188.8.131.52 Other aquatic invertebrates Rainbow & White (1989) investigated uptake of cadmium and zinc in three marine crustaceans, Palaemon elegans (Decapoda), Echinogammarus pirloti (Malacostraca), and Elminius modestus (Cirripedia) at water concentrations of cadmium between 0.5 and 1000 µg/litre and zinc between 2.5 and 4000 µg/litre. All three crustaceans accumulated the non-essential cadmium at all dissolved cadmium concentrations without regulation. Differences between species were interpreted by the authors in terms of differences in cuticle permeability and way of life. All three species took up zinc more rapidly than cadmium; the ratios between molar uptake rates of zinc to cadmium were 11.4:1, 2.7:1, and 3.7:1 for the three species, respectively, following an exposure to a molar ratio of 1.7:1. 184.108.40.206 Fish Cadmium uptake in fish continues for some considerable time in fish exposed to the metal. The peak of tissue residues may not be reached for several weeks, particularly after exposure to low concentrations of the metal (Cearley & Coleman, 1974; Benoit et al., 1976; Sullivan et al., 1978a). Douben (1989a) exposed the stone loach Noemacheilus barbatulus to cadmium in water (as the sulfate) at a concentration of 1 mg/litre and monitored uptake and loss at different temperatures with fed and starved fish. The size of the fish affected both uptake and loss of cadmium, bioconcentration factors decreasing with size. Uptake of cadmium increased with temperature up to about 16 °C and decreased as the concentration of cadmium in the water increased. Feeding the fish increased the rate of uptake of cadmium from the water. The author concluded that metabolic rate was an important factor in the uptake of cadmium into the fish and in its subsequent loss. 220.127.116.11 Model aquatic ecosystems Ferard et al. (1983) investigated the transfer of cadmium through a model food-chain consisting of an alga, a daphnid, and a fish. Concentration factors relative to food were low, indicating that cadmium is mainly taken up directly from water. Daphnids fed algae containing cadmium at between 4.5 and 570 mg/kg dry weight showed a maximum concentration factor of 1. Fish fed contaminated daphnids or algae showed concentration factors of 0.0038 and 0.0018, respectively. Nimmo et al. (1977) reported low concentration factors, ranging from 0.018 to 0.027, for grass shrimp fed on brine shrimp containing cadmium at between 27 and 182 mg/kg. Rehwoldt & Karimian-Teherani (1976) fed zebrafish on food containing cadmium acetate at 10 mg/kg over a period of 6 months. Maximum residues, in males and females respectively, were 5.92 and 13.64 mg/kg, the median residue levels after 6 months of exposure being 5.19 and 12.95 mg/kg (on a dry weight basis). 18.104.22.168 Uptake from aquatic sediment Ray et al. (1980b) exposed the ragworm Nereis virens to sediment to which cadmium chloride had been added. Smaller worms took up more cadmium relative to body weight than larger worms. The cadmium was taken up in a dose-related manner and no equilibrium was reached during the 24-day experiment. The rate of uptake directly from sea water also increased with exposure concentration over the range of 0.03 to 9.2 mg/litre. For the range of sediment cadmium concentrations used (1 to 4 mg/kg), the corresponding concentrations in the overlying sea water were 0.03 to 0.1 mg/litre. Comparing uptake into the ragworms from water with these concentrations to the uptake from the spiked sediment produced identical concentrations of cadmium in the worms. Rate of uptake from sediment was between 16 and 39 times less than the uptake from the corresponding exposure to cadmium in water. The authors concluded that all of the uptake of cadmium from sediment derived from desorbed metal ions in the interstitial water. 22.214.171.124 Uptake from food relative to uptake from water Fish can take up cadmium from the surrounding water and from ingested food. The main uptake route in fresh water is from the water via the gills (Williams & Giesy, 1978). However, the relative importance of food and water to the body burden depends very much on the cadmium content of the food organism. In contaminated areas with an increased cadmium content in food organisms, the relative importance of food as a cadmium source may increase. In the marine environment, where cadmium is mainly present in chloride complexes not available to fish, the relative importance of food as a cadmium source increases. Consequently food has been shown to be the main cadmium source in marine fish (Pentreath, 1977; Dallinger et al., 1987). 4.1.2 Uptake by terrestrial organisms 126.96.36.199 Uptake into plants The uptake of cadmium into plants generally depends upon the availability of the metal in soil solution. The soil pH and composition, particularly the nature of soil clays, the organic matter content, and, obviously, the soil cadmium level, affect this availability. The relationship between soil cadmium level and plant uptake is not a simple one because of the wide variety of soil characteristics that affect the extent of cadmium uptake. Cataldo & Wildung (1978), Peterson & Alloway (1979), and Page et al. (1981) have reviewed this subject. Plants grown in a greenhouse or a container take up more cadmium than the same plants grown in soil with the same cadmium levels in the field. This is due to greater root development in a confined volume in containers and to the fact that all the roots are in contact with cadmium-contaminated soil. In the field, roots may grow down below the cadmium-contaminated level (Page & Chang, 1978; De Vries & Tiller, 1978). Mahler et al. (1978) cultured lettuce and chard on acid or calcareous soils to which cadmium sulfate had been added at levels up to 320 mg/kg. For both types of soil there was a dose-related uptake of cadmium from soil into leaves. The uptake of the metal was much greater in acid than in calcareous soils, particularly at higher rates of cadmium application (over 40 mg/kg). At the highest soil concentration of 320 mg/kg, lettuce leaves contained cadmium at a concentration of 800 mg/kg and chard leaves 1600 mg/kg when grown in acid soil. Leaves of lettuce cultured on calcareous soils with cadmium at 320 mg/kg contained a lower cadmium concentration of 200-300 mg/kg and chard, similarly cultured, contained 300 mg/kg or less. Bingham et al. (1980) showed an effect of soil pH on cadmium (as sulfate) uptake in rice; more metal was incorporated as acidity increased. Chaney et al. (1975) reported that liming of soil in which soybeans were growing decreased the concentrations of cadmium in leaves from 33 to 5 mg/kg dry weight as pH increased from 5.3 to 7.0. Eriksson (1988) investigated the effect of pH on the uptake of cadmium into perennial ryegrass ( Lolium perenne) and winter rape ( Brassica napus). The more soluble fractions of cadmium in soil increased as the pH was lowered; increasing the pH from 5 to 7 with calcium oxide invariably reduced the cadmium content of ryegrass plants, but this decrease was less consistent when the pH was increased from 5 to 6. The cadmium content of rape plants was markedly higher at pH 4 than pH 5. Adding more cadmium to the soil increased the amount of cadmium in the plants in direct proportion to the increased concentration of the metal in soil over the range 0 to 5 mg/kg. Eriksson (1988) found that soil organic matter decreased the availability of cadmium to perennial ryegrass and winter rape grown in pots. Addition of organic material to sand and clay soils reduced cadmium uptake to a greater extent in the sand. When Mitchell & Fretz (1977) cultured seedlings of three species of tree (red maple, white pine, and Norway spruce) hydroponically or in soil with added cadmium, the concentration in roots was greater than that in leaves. Cadmium added to soil was less readily taken up than cadmium added to nutrient solutions. Similarly, Root et al. (1975) reported greater cadmium concentration in roots than in shoots of maize grown hydroponically in a medium containing cadmium chloride. Harkov et al. (1979) found the highest uptake of cadmium into hydroponically grown tomatoes in the roots, while stems had lower cadmium concentrations than leaves. Lepp et al. (1987) measured high concentrations of cadmium in the sporophores (fruiting bodies) of the fungus Amanita muscaria growing in birch woodland. The fungus sporophores contained 29.9 mg/kg dry weight, compared to a cadmium level of 0.4 mg/kg in the soil on which they grew. The cadmium was released from the rotting sporophore, after it had shed its spores, in a form which was readily available to other plants growing on the woodland soil; this was shown experimentally with lettuce plants grown in pots. The authors calculated that an abundant population of sporophores could recycle 1.4% of the total cadmium load in leaf litter to higher plants over a period of 14 days (the mean lifespan of the sporophores). 188.8.131.52 Terrestrial invertebrates Beyer et al. (1982) demonstrated that earthworms concentrated cadmium from soils amended with sewage sludge containing cadmium oxide. Cadmium concentrations were as high as 100 mg/kg in worms exposed to soils containing cadmium at 2 mg/kg, a concentration factor of 50. Adding calcium carbonate to soils decreased the cadmium uptake of worms slightly, while high soil zinc levels decreased the cadmium uptake appreciably. Results were variable with different sludge treatments. Hartenstein et al. (1980) amended sludge with 10, 50, and 100 mg/kg cadmium (as cadmium sulfate) and added earthworms ( Eisenia foetida). The worms accumulated 3.9, 2.04, and 1.44 times the respective sludge levels of cadmium over a period of 5 weeks. In field trials on non-amended soils containing 12 to 27 mg cadmium/kg, worms sampled during a 28-week period gave levels of cadmium ranging from 8 to 46 mg/kg. Terrestrial pulmonate snails retained up to 59% of cadmium administered in their diet as the chloride (Russell et al., 1981). The highest retention was after dosing at 25 mg cadmium/kg diet. The higher the dose (up to 1000 mg/kg diet) the lower the percentage retention of the metal. Ireland (1981) noted that in the terrestrial slug Arion ater most of the cadmium was located in the digestive gland without association with any particular sub-cellular organelles, and isolated a specific cadmium-binding protein from the animals. 184.108.40.206 Birds In a study by White & Finley (1978), adult mallard ducks were fed a diet containing cadmium chloride at levels of 2, 20 or 200 mg/kg and killed at 30-day intervals. The cadmium content increased with dose level and time (except in the case of the highest dose where body burden peaked after 60 days), and the highest concentrations occurred in the liver and kidney. The highest levels overall occurred after dosing for 60 days at 200 mg/kg; cadmium concentrations were 109 mg/kg in the liver and 134 mg/kg in the kidney. Nicholson & Osborn (1983) dosed starlings ( Sturnus vulgaris) with cadmium chloride at a concentration of 2 mg/kg body weight, three times weekly for 6 weeks, and reported a wide range of kidney concentrations (from < 10 to > 200 mg/kg dry weight). 4.2 Distribution 4.2.1 Aquatic organisms In higher organisms, cadmium can be bound in several different tissues, whereas in plants cadmium is bound to the cell wall in roots. Brooks & Rumsby (1967) measured the cadmium taken up by the oyster ( Ostrea sinuata) from water containing 115Cd (50 mg/litre). The soft parts of the oyster contained 100 mg cadmium/kg after 100 h. Concentrations in tissues were, in decreasing order, 360 mg/kg for gills, 285 mg/kg for heart, 141 mg/kg for the visceral mass, 83 mg/kg for the mantle, 53 mg/kg for white muscle, and 25 mg/kg for striated muscle. Nimmo et al. (1977) reported that in the pink shrimp the hepatopancreas took up more cadmium than other tissues. Lower concentrations were found in the exoskeleton, muscle, and serum. Short-term exposure of the crab Uca pugilator to cadmium chloride led to the hepatopancreas and gill concentrations of the metal being similar after a 24-h exposure to 1 mg cadmium/litre (Vernberg et al., 1974). Sangalang & Freeman (1979) determined the cadmium in tissues of brook trout exposed to the metal (added as the chloride) via the water or by injection. After water exposure to cadmium chloride at 1 µg/litre, the trout showed greatest uptake of the metal in the gills, kidney, and liver. The gills and the posterior kidney revealed a higher metal content than any other tissues. Levels of cadmium in whole blood and plasma, heart, spleen, testis, stomach, and skin were higher than control levels after 77 and 93 days of exposure. Smith et al. (1976) found the greatest accumulation of the metal in the kidney of catfish exposed to cadmium (as sulfate) in the water. In an autoradiographic study of cadmium distribution in rainbow trout exposed to cadmium in water, Tjalve et al. (1986) confirmed the general picture of cadmium distribution, the metal being found in the gills, liver, and kidney. However, they also observed heavy labelling of the olfactory rosette and the olfactory nerve, an observation not reported earlier. In a detailed study they later showed that cadmium was transported axonally from the olfactory rosette to the bulbus olfactorius but not further into the brain (Gottofrey, 1990). The significance of this observation with respect to the olfactory responses of fish in cadmium-contaminated environments remains to be investigated. The few studies that have been conducted on the subcellular distribution of cadmium indicate that, while much is located in the cytosol, a significant proportion can be found in the nucleus and the mitochondria. Cadmium is bound in the cytosol to proteins of low relative molecular mass, metallothioneins, and other cadmium-binding proteins. These proteins are rich in the sulfur-containing amino acid cysteine but poor in aromatic amino acids. Metallothioneins have been isolated and characterized in a number of aquatic and terrestrial organisms. Fish metallothioneins have received considerable interest in recent years as tools in monitoring metal pollution in the environment (Hamilton & Mehrle, 1986; Hogstrand & Haux, 1990a). Simple methods to analyse fish metallothionein have been developed, including differential pulse polarography (Olson & Haux, 1986) and radioimmunoassay based on specific antibodies to fish metallothionein (Hogstrand & Haux, 1990b). Olson & Haux (1986) found a strong correlation between hepatic metallothionein and cadmium accumulation in perch collected from cadmium-contaminated water. 4.2.2 Terrestrial organisms 220.127.116.11 Terrestrial plants Jones & Johnston (1989) analysed cereal grain and herbage from long-term experimental plots at Rothamsted, United Kingdom, and found that uptake of cadmium into herbage was greatest where phosphate fertilizer had been applied. It was also greater from unlimed soils than from limed soils. However, the authors concluded that there was little evidence of a long-term (1840-1986) increase in crop cadmium concentrations. Byrne et al. (1976) analysed higher fungi from Slovenia, Yugoslavia, and found levels of cadmium ranging from 0.53 to 39.9 mg/kg dry weight (average 5.0 mg/kg). This is an order of magnitude higher than in most other plants. Although the fungi were collected from industrial, urban, and uncontaminated sites, the levels found in the fungi were not very different between sites. The authors suggested geological rather than industrial sources for the cadmium in these soils. The high uptake by mushrooms and related species is probably due to a cadmium-binding phosphoglycoprotein, cadmium-myco-phosphatin, which has been isolated from the mushroom Agaricus macrosporus (Meisch & Schmitt, 1986). 18.104.22.168 Terrestrial invertebrates Hopkin & Martin (1985) investigated the storage of cadmium in the woodlouse Oniscus asellus from heavily contaminated woodland 3 km downwind from a smelter. The hepatopancreas was found to contain up to 5 g cadmium/kg dry weight without apparent ill effects upon the organism. Cadmium was reported to be stored intracellularly in the copper- and sulfur-containing granules of epithelial S cells. In a later study (Hopkin, 1990) it was found that considerable interspecies differences exist with regard to storage in the hepatopancreas. Oniscus asellus stored five times more cadmium than Porcello scaber under the same conditions. The carnivorous centipede Lithobius variegatus, when fed on cadmium-contaminated hepatopancreas from woodlice, accumulated cadmium which was likewise stored in the midgut (Hopkin & Martin, 1984). Berger & Dallinger (1989) studied the distribution of cadmium between several organs of the terrestrial snail Arianta arbustorum during a 20-day feeding experiment on cadmium-enriched agar. Of the cadmium in the medium, 54% was taken up, of which 66% was distributed to the hepatopancreas, leading to a concentration of more than 500 mg/kg dry weight. In other organs (intestine, foot/mantle, gonads), the cadmium concentration was considerably lower. In the earthworm Lumbricus rubellus taken from heavy-metal- polluted soil, more than 70% of the cadmium burden was found in the posterior alimentary canal (Morgan & Morgan, 1990). This distribution prevented dissemination of large concentrations of cadmium into other tissues and, according to the authors, may represent a detoxification strategy. 4.3 Elimination Information on loss of cadmium from organisms is relatively scarce. The information that does exist suggests that this is very variable, and has been reviewed by Coombs (1979) and Taylor (1983). Organisms that accumulate cadmium also tend to retain the metal for long periods. The main excretory route appears to be via the kidney, except in the case of organisms that moult, where loss from the shed exoskeleton can be significant. Robinson & Wells (1975) administered a single oral dose of cadmium acetate to softshell turtles ( Trionyx spinifer) and killed and dissected the animals either 48 h or 96 h later. After 48 h, 9.43% of the total dose was recovered from tissues, while turtles killed after 96 h had retained 4.02% of the dose. The greatest retention of cadmium, after both time periods, was in the liver. Cadmium was also retained in the small intestine for the first 48 h, but the amount had decreased by 96 h. Harrison & Klaverkamp (1989) exposed rainbow trout ( Salmo gairdneri) and lake whitefish (Coregonus clupeaformis) to cadmium in water, via a continuous-flow system, or the diet, via pelleted food, for 72 days. The fish were then kept in clean water on a cadmium-free diet for a further 56 days. In the case of water-exposed fish, the majority of the cadmium was present in the gill and kidney, but food-exposed fish retained cadmium principally in the kidney, gut, and liver. Bioconcentration factors for exposure via the water were 55 for the trout and 42 for the whitefish, whereas concentration factors from the food were less than 1 for both species. However, both species accumulated a greater proportion of the cadmium that was in the food than that in water (1% as against 0.1%). Equilibrium bioconcentration factors were estimated to be 161 for trout and 51 for whitefish. In the same model, the half-times for depuration of accumulated cadmium ranged from 24 to 63 days. Douben (1989b) investigated the kinetics of cadmium in freshwater fish (the stone loach Noemacheilus barbatulus) exposed to cadmium via the diet (tubifex worms previously contaminated with cadmium by uptake from water). The body burden of cadmium declined after the period of feeding with contaminated diet more rapidly in starved than in fed fish. Rate constants for loss of cadmium appeared to be greater during the exposure period than after exposure. Both uptake and loss of cadmium were influenced by the body weight of the fish. Janssen et al. (1991) investigated uptake and loss of cadmium from contaminated soil by four species of soil arthropod and developed kinetic models that gave good predictions of the degree of accumulation in a variety of species. They also reviewed data on other soil arthropods (Tables 6 and 7). The kinetics of cadmium in different arthropods is related to taxonomy and reflects the different physiological characteristics of the different organisms. Some, notably isopods and molluscs, take up and retain cadmium in their tissues with little or no excretion. These species are capable of holding large quantities of the metal in the hepatopancreas without apparent ill effect. There is no direct correlation between assimilation capacity and the capacity to excrete or eliminate cadmium. Figure 4 illustrates the uptake of cadmium (measured as total body burden) and its subsequent loss in four species of arthropods. Elimination half-lives of 53, 8, and 2 days, respectively, have been reported for Platynothrus peltifer, Orchesella cincta, and Notiophilus biguttatus; no elimination took place over 130 days in Neobisium muscorum. Sawicka-Kapusta et al. (1987) investigated the effect of keeping the vole Clethrionomys glareolus at different temperatures on the rate of loss of cadmium from body tissues. Although the different temperatures (10 °C and 20 °C) affected the metabolic rate of the voles, there was no difference in the rate of loss of cadmium. 4.4 Bioaccumulation and biomagnification Bioaccumulation occurs when the concentration in the organism exceeds the concentration in the nutrient medium and is expressed quantitatively as a bioconcentration factor. Progressive bioaccumulation at each trophic level is termed biomagnification. Table 6. Cadmium assimilation efficiencies in different soil invertebrates Species Food Cadmium concentration Assimilation efficiency Reference in food (µmol/g) (%) Snail Arianta arbusloruma agar 1.48 55-92 Berger & Dallinger (1989) b Centipede Lithobius variegatus isopod 1.21-10.2 0-7.2 Hopkin & Martin (1984) hepatopancreas Millipede Clomeris marginata maples leaves 8.2-40.6 Hopkin et al. (1985) Pseudoscorpion Neobisium muscorum collembolans 0.20 58.9 Janssen et al. (1991) Mite Platynothrus peltifer green algae 0.15 17.2 Janssen et al. (1991) Insects Orchesella cincta green algae 0.09 8.3 Van Straalen et al. (1987) Orchesella cincta green algae 0.15 9.4 Janssen et al. (1991) Notiophilus biguttatus collembolans 0.23 35.5 Janssen et al. (1991) a assimilation value for midgut gland b recalculated from the data Table 7. Excretion constants (k) for cadmium in different soil invertebrates Species Taxonomic k Reference group (day-1) Helix pomatia snail 0 Dallinger & Wieser (1984) b Cepaea nemoralis snail 0.007 Williamson (1980) b Oniscus asellusa isopod 0.002 Hopkin (1989) b Neobisium muscorum pseudoscorpion 0 Janssen et al. (1991) Lycosa spp spider 0.007 Van Hook & Yates (1975) Platynothrus peltifer oribatid mite 0.013 Janssen et al. (1991) Orchesella cincta collembolan 0.061 Van Straalen et al. (1987) Orchesella cincta collembolan 0.087 Janssen et al. (1991) Acheta domesticus cricket 0.090- Van Hook & Yates (1975) 0.110 Notiophilus biguttatus carabid beetle 0.375 Janssen et al. (1991) a k value for midgut gland or hepatopancreas b recalculated from the data Bioconcentration factors (the ratio between the cadmium concentration in the organism and the concentration in the medium) for several groups of organisms studied under laboratory conditions are shown in Table 8. They range from 16 to 130 000 and do not seem to show any consistent pattern. Table 8. Bioconcentration of cadmium in laboratory studies Organism Size Stat/ Organ a Temperature Duration Exposure Bioconcentration Reference flow (°C) (days) (µg/litre) factor b Freshwater alga 10 10 3000 dw c Ferard et al. (1983) (Chlorella vulgaris) Freshwater alga stat 20-22 14 250 4940 daw Cain et al. (1980) (Scenedesmus obliquus) Freshwater diatom flow WB 23 10 40 000 Conway (1978) (Asterionella formosa) Submerged plant WP 25 30 25 1730 dw Nakada et al. (1979) (Elodea nuttallii) Water hyacinth leaves 28 500 16 dw c Kay & Haller (1986) (Eichhornia crassipes) American oyster 4.9-5.1 g flow WB 16-20 21 10 116 ww Eisler et al. (1972) (Crassostrea virginica) 4280 aw 8.1 g flow ST 2.8-22.6 280 5 2376 ww Zaroogian & Cheer (1976) 18 472 dw Mussel 32-34 mm flow ST 13 166 10 50 802 dw Riisgard et al. (1987) (Mytilus edulis) Scallop 6.8-7.7 g flow WB 16-20 21 10 131 ww Eisler et al. (1972) (Aquipecten irradians) 3970 aw Bay scallop 0.51-0.73 g flow ST 9.5-16 42 60 20 400 Pesch & Stewart (1980) (Argopecten irradians) Crab 2-4 g WB 10 14 37 152 dw Ray et al. (1980a) (Pandalas montagui) Grass shrimp 20-33 mm flow WB 9.5-16 42 60 223 Pesch & Stewart (1980) (Palaemonetes pugio) Table 8 (contd). Organism Size Stat/ Organ a Temperature Duration Exposure Bioconcentration Reference flow (°C) (days) (µg/litre) factor b Lobster 160-169 g flow WB 16-20 21 10 21 ww Eisler et al. (1972) (Homarus americanus) 10 aw Mummichog 2.3-2.4 g flow WB 16-20 21 10 15 ww Eisler et al. (1972) (Fundulus heteroclitus) 200 aw Fathead minnow flow WB 13.9-15.3 21 49 190 Sullivan et al. (1978a) (Pimephales promelas) Red maple leaves 15-27 45 0.5 14 400 dw d Mitchell & Fretz (1977) (Acer rubrum) roots 15-27 45 0.5 131 800 dw d Mitchell & Fretz (1977) leaves 15-27 101 2.6 mg/kg 0.76 dw e Mitchell & Fretz (1977) roots 15-27 101 2.6 mg/kg 12.5 dw e Mitchell & Fretz (1977) White pine leaves 15-27 66 0.5 3400 dw d Mitchell & Fretz (1977) (Pinus strobus) roots 15-27 66 0.5 118 400 dw d Mitchell & Fretz (1977) leaves 15-27 36 52.6 mg/kg 1.2 dw e Mitchell & Fretz (1977) roots 15-27 36 52.6 mg/kg 10.4 dw e Mitchell & Fretz (1977) a WB = whole body; WP = whole plant; ST = soft tissues b Chloride salt used unless stated otherwise; bioconcentration factor = concentration in the organism divided by concentration in the medium; dw = dry weight; ww = wet weight; aw = ash weight; daw = dry ash weight c Nitrate salt used d The medium was a cadmium-enriched nutrient solution e The medium was a cadmium-amended soil mix Microorganisms generally exhibit a high capacity to take up cadmium from water and retain the metal in their cells. The highest bioconcentration factors reported have been for micro-organisms, the greatest value being 40 000 in a freshwater diatom (Conway, 1978). In this diatom, 58% of the cadmium was located in the cellular content with 25% in the organic coating of the frustule and 17% in the silicaceous frustule. The bioconcentration factor of 3000 for the alga Chlorella (Ferard et al., 1983) is typical of the value for microorganisms. Flatau et al. (1988) demonstrated the uptake (it was not specified whether this referred to absorption or adsorption) of cadmium from sea water by marine bacteria; the uptake of the metal increased with its concentration in the water, and the accumulation rate was a logarithmic function of the dose. Sorption was only observed with exposure concentrations above 10 µg Cd/litre, suggesting that a threshold had to be exceeded for cadmium uptake to occur. Dongmann & Nurnberg (1982) showed that the bioconcentration factor for a marine diatom, Thalassiosira rotula, decreased with increasing metal concentration, suggesting a saturation effect. Their reported concentration factors, which vary between 1000 and 2000, reflect the reduced sorption of cadmium by marine microorganisms compared with their freshwater relatives. Hart & Scaife (1977) reported a direct relationship between the level of cadmium in the medium and sorption to the alga Chlorella exposed to cadmium concentrations ranging from 0.25 to 1.00 mg/litre. After water hyacinths had been exposed for 4 weeks to water containing 0.5 or 1.0 mg cadmium/litre, added as cadmium nitrate, the leaves had accumulated 8.00 and 17.20 mg/kg, respectively (Kay & Haller, 1986). Molluscs concentrate cadmium to a high degree over a period of time, but uptake is often slow. Oysters showed a concentration factor of only 149 over a 10-day period (Eisler et al., 1972) but a factor of 2714 after 40 weeks (Zaroogian & Cheer, 1976). Elliott et al. (1985) examined the accumulation of cadmium, copper, lead, and zinc in the tissues of the mussel, Mytilus edulis. Under simultaneous exposure to all four metals, both lead and cadmium were accumulated in direct proportion to the exposure time, whereas copper and zinc were not. Accumulation of cadmium was influenced by the presence of other metals. Compared with oysters, the related bay scallop shows greater accumulation of cadmium when exposed to low concentration of the metal as the chloride over 6 weeks (Pesch & Stewart, 1980). Short-term exposure of the same scallop to higher concentrations of cadmium resulted in very much lower concentration factors. Exposure for 96 h to cadmium (as the chloride) at up to 2.0 mg/litre led to a bioconcentration factor of around 50 (Nelson et al., 1976). Bioconcentration factors (from water and food) and biomagnification factors (from food alone) were calculated for the freshwater isopod Assellus aquaticus by van Hattum et al. (1989). Much of the cadmium (added as the chloride) was taken up from the water (bioconcentration factor 18 000), but there was little uptake from food (bioconcentration factor 0.08). Direct uptake from water accounted for between 50 and 98% of the body burden after 30 days of exposure (based on dry weight measures). Cadmium was readily taken up by the isopod even at exposure concentrations of 1 µg/litre. Experiments conducted at two different pHs (5.9 and 7.6) revealed no significant effect of pH on uptake of cadmium by the isopod. Wright & Frain (1981b) demonstrated that adult intermoult amphipods ( Gammarus pulex) accumulated only half as much cadmium from a solution of 5 mg/litre in the presence of 200 mg calcium/litre as with 20 mg calcium/litre. Ramamoorthy & Blumhagen (1984) investigated the uptake of cadmium, mercury, and zinc by rainbow trout (Salmo gairdneri) in a model system which simulated the presence of other competing compartments that would be found in nature. The system consisted of either a simple sediment/water model or a more complex series of compartments in dialysis bags of suspended sediment, cation and anion exchange resins (to represent naturally occurring polyelectrolyte materials of plant origin), and fish. River water was used as the fluid transfer medium, and the system was continuously stirred. Equilibrium with one heavy metal ion did not inhibit the uptake of other metal ions; cadmium and zinc were taken up after equilibrium with mercury. The authors calculated approximate partition coefficients (fish/substrate) to be 2.8 for sediment, 550 for water, and 2 and 3.6 for the cation and anion exchange resins, respectively. The problem of expressing changes in concentration between trophic levels is that the units are not compatible. There is no significance to a bioconcentration term that expresses a ratio of cadmium in soil moisture to cadmium in plant tissue, or cadmium in plant tissue to cadmium in herbivore tissue. Therefore, it is difficult to assess the impact of cadmium on the environment in terms of bioconcentration factors. An alternative method is to measure both cadmium and calcium at each trophic level and express these measurements as a molar ratio of these two elements. (The molar ratio should be used to account for the movement by atoms, not grams.) Differences between trophic levels are calculated as the ratio of the higher trophic level to the lower. This approach, called biopurification, recognizes that the flow of the non-nutrient cadmium through successive trophic levels follows a pathway similar to that of nutrients such as calcium, and that calcium must pass natural chemical and physiological barriers, such as membranes and selective enzymes, that progressively purify the pool of the nutrient calcium relative to the non-nutrient cadmium. In the case where two similar ecosystems are compared, and where one is believed to be more contaminated than the other, the relative degree of contamination can be calculated as the difference between molar ratios at the same or similar trophic levels. It is unfortunate that the absence of concentration data on nutrients such as calcium or, alternatively, zinc, prohibits the calculation of biopurification factors for any of the studies discussed in this monograph. 5. TOXICITY TO MICROORGANISMS Appraisal Cadmium is toxic to a wide range of microorganisms in culture (effects of cadmium on microorganisms in the field are discussed in chapter 8). However, the presence of sediment, organic matter or high concentrations of dissolved salts reduces the availability of cadmium to microorganisms and, therefore, reduces the toxic impact. Freshwater microorganisms in culture are thus affected by cadmium at lower concentrations than marine species (for example, 50 µg/litre affects growth in many freshwater species of algae while at least 100 µg/litre, and often 1000 µg/litre, is required to reduce growth in marine species). Soil microorganisms are partially protected from the toxic effects of cadmium by the presence of clay. 5.1 Aquatic microorganisms 5.1.1 Freshwater microorganisms Canton & Slooff (1982) exposed the bacterium Salmonella typhimurium and the alga Chlorella vulgaris to cadmium in the form of the chloride, and calculated an 8-h EC50 (growth inhibition) of 10.4 mg/litre for the bacterium and a 96-h EC50 of 3.7 mg/litre for the alga. No-toxic-effect levels of 0.65 and 1.5 mg/litre were estimated for the bacterium and alga, respectively. Jana & Bhattacharya (1988) found significant inhibition of population growth in the faecal coliform bacterium Escherichia coli during exposure to cadmium concentrations of 1, 2 or 5 mg cadmium/litre for 7 or 28 days. Cadmium was the most toxic of the metals tested. Norberg & Molin (1983) exposed the bacterium Zoogloea ramigera (abundant in sewage treatment plants) to cadmium chloride concentrations of 1, 3, 5, and 10 mg cadmium/litre for 30 h. A prolonged lag phase and decrease in growth resulted, the length of the lag phase being proportional to the concentration of cadmium in the medium. Babich & Stotzky (1977a) showed that the presence of clay particles protected bacteria from the toxic effect of cadmium added to culture medium. The degree of protection was related to the cation exchange capacity of various clays tested. Chapman & Dunlop (1981) estimated the 8-h LC50 for the freshwater protozoan Tetrahymena pyriformis to be less than 1 mg/litre. However, this value increased with increasing water calcium concentration; at a value of 500 mg calcium/litre, the LC50 was 19 mg/litre. Magnesium also exerted a protective effect against cadmium when mixed with calcium. Cadmium was consistently more toxic to Tetrahymena in the presence of magnesium alone. Berk et al. (1985) calculated a 15-min EC50 (inhibition of ciliate chemotactic response) for Tetrahymena sp. of 0.35-0.7 mg. When Skowronski et al. (1988) exposed the green microalga Stichococcus bacillaris to cadmium chloride concentrations of 45 and 90 µmol/litre for 4 days, growth rate was inhibited by 28% and 45% at the two respective concentrations. At both exposure levels, dry weight and chlorophyll a content were reduced in a dose-related manner. Addition of manganese at concentrations of between 45 and 1800 µmol/litre had a dose-related antagonistic effect on cadmium toxicity. Bennett (1990) found that the addition of cadmium (1.8 µmol/litre) to a turbidostat culture of Chlorella pyrenoidosa caused a decrease in the maximum specific growth rate (toxicity was expressed after a lag of 5 generations). A gradual decrease in the maximum specific growth rate was also noted during a 40-day exposure to stepwise increases in the cadmium concentration (0.96 to 1.68 µmol/litre). The author found that the addition of manganese (10.4 µmol/litre) had an antagonistic effect, causing the maximum specific growth rate to increase. Cadmium is toxic to the growth of the freshwater alga Chlorella pyrenoidosa (Hart & Scaife, 1977). In cultures maintained at pH 7.0, doubling times were 11, 21, 22, and 35 h for cadmium concentrations of 0, 0.25, 0.5, and 1.0 mg/litre medium, respectively. At a pH of 8.0, the effect was somewhat lessened; doubling times were 11, 16, 17, and 25 h for the same range of cadmium doses. There was also a pronounced effect on carbon dioxide fixation, which was reduced from 0.738 to 0.720, 0.558, and 0.283 µmol HCO3- fixed per hour with cadmium exposures in the culture medium of 0, 0.246, 0.554, and 1.090 mg/litre, respectively. There was less of an effect on oxygen evolution over the same dose range. Zinc offered no protection against cadmium effects. Wong et al. (1979) exposed four different species of freshwater algae to cadmium and measured the uptake of 14C-carbonate. Scenedesmus quadricaudata was the most sensitive species, carbonate uptake being inhibited by 80% at a cadmium concentration of 20 µg/litre. Chlorella pyrenoidosa showed 70% inhibition of carbonate uptake at about 100 µg/litre, while Chlorella vulgaris showed only 50% inhibition at about 500 µg/litre. The least sensitive of the four species tested was Ankistrodesmus falcatus variety acicularis where an effect on carbonate uptake started only at concentrations higher than 500 µg/litre. There was no observed effect on the growth of A. falcatus at cadmium concentrations lower than 5 mg/litre. Rebhun & Ben-Amotz (1984) demonstrated that the chlorophyll content of cells of Chlorella stigmatophora was reduced in a dose-dependant manner across a range of cadmium concentrations of between 1 and 10 mg/litre medium. Laegreid et al. (1983) studied the effects of cadmium on the alga Selenastrum capricornutum cultured in the laboratory in water taken from two lakes at various times throughout the year. The two lake waters contained different amounts of organic material. The first lake, a dystrophic bog lake, had a high organic content, while the second, an eutrophic lake, had a low organic content. In the dystrophic lake, which had a low pH (4.4), the toxicity of cadmium was related to the free ionic concentration of the metal, as suggested by many laboratory experiments. In the eutrophic lake, where there was less influence from organic material, there was a pronounced seasonal effect. In the summer, when growth and productivity of the algae were highest, there was a much greater effect of the metal than predicted. The toxicity of cadmium, at this time, was far greater than would be expected even if all of the metal was in the free ionic form and none was bound to organic compounds. On the basis of their field evidence, the authors questioned the generally held assumption that organic binding is the major factor in determining cadmium toxicity to microorganisms. They considered that the presence of certain organic compounds of low relative molecular mass could increase cadmium toxicity. This conclusion is supported by the work of Giesy et al. (1977), who found that uptake of cadmium into zooplankton could be increased in the presence of organic compounds of low relative molecular mass. Chin & Sina (1978) investigated the cellular basis of cadmium toxicity in microorganisms using cultures of Physarum polycephalum. The organism was cultured, in plasmodial form, on the surface of liquid medium, and replicate discs, cut from the protoplasmic sheet of the organism, were used for the tests. The discs maintain mitotic synchrony with each other and, therefore, cadmium could be introduced at specific points in the cell cycle. The cultures were exposed to cadmium sulfate (5 x 10-4 mol/litre), which was floated onto the surface of the culture medium. Exposure to cadmium immediately prior to early prophase of mitosis extended the normal DNA replication period from 3 h to 4 h; this was monitored using measurements of uptake of 3H-thymidine. Two stages of the cell cycle were particularly susceptible to cadmium. Exposure either at the beginning of the cycle or 80% of the way through the cycle caused delays in the completion of mitosis. A 30-min exposure to cadmium at the onset of early prophase inhibited incorporation of 3H-uridine into RNA for the following 3 h by 51% and stimulated the incorporation of 3H-thymidine into DNA, for the same period, by 85%. Later in the cycle, DNA synthesis was inhibited and DNA content was depressed by 12.5%. There was an ultrastructural effect on the nucleoli (less dense material centrally giving nucleoli in section a "ring" structure), which was the only structural effect of the metal even after 4 h of exposure. Accommodation occurred after pre-treatment with sub-toxic doses of cadmium. Treatment with cadmium at the less sensitive periods of the cell cycle led to reduced effect at the more sensitive phases; the organism, in some way, compensated. There was no accommodation by Physarum after pre-treatment with zinc ions. This result contrasts with that of a similar study on Escherichia coli where pre-exposure to zinc reduced the effects of cadmium (Mitra et al., 1975). 5.1.2 Estuarine and marine microorganisms Chan & Dean (1988) exposed the marine bacterium Pseudo-monas marina to cadmium sulfate at concentrations of 1 to 25 mg cadmium/litre. Effects were exposure related and included a lengthened lag time, reduced growth rate, reduced biomass and oxygen uptake, and a decrease in the activity of dehydrogenase and alkaline phosphatase. The IC50 values for inhibition of biomass and of growth rate were 11 and 11.5 mg/litre, respectively. Flatau et al. (1987) progressively adapted the marine bacterium Pseudomonas sp. to cadmium concentrations from 1 to 80 mg/litre. Although the length of the lag phase was not linearly or logarithmically correlated with the cadmium concentration, it was significantly longer at cadmium concentrations of 70 mg/litre or more. The growth rate was reduced at 10 mg/litre but remained constant at cadmium concentrations of 10 to 50 mg/litre; no growth was observed at 75 or 80 mg/litre. Oxygen consumption was not different from that of controls at 1 or 5 mg/litre but at concentrations of 10 mg/litre or more respiratory activity decreased. Bressan & Brunetti (1988) exposed the marine microalgae Dunaliella tertiolecta and Isochrysis galbana to cadmium concentrations of 13.8 and 0.2 mg/litre, respectively, for up to 8 days. Cadmium significantly reduced the population growth, expressed as the mean number of cells/ml, for both species. The addition of nitrilotriacetic acid (NTA, a sequestering agent) at ratios of 1:1, 1:2 or 1:3 did not modify the toxic effect of cadmium. Berk et al. (1985) exposed the marine ciliates Paranophrys sp. and Miamiensis avidus to cadmium and monitored the inhibition of the chemotactic response. The 15-min EC50 values were 2-3.1 mg cadmium/litre and 5.1-7 mg cadmium/litre for the two species, respectively. Dongmann & Nurnberg (1982) investigated the effects of cadmium on the marine diatom Thalassiosira rotula (a chain-forming diatom native to shallow temperate waters) in petri dish and in batch liquid cultures. They calculated generation times from the dish cultures and reported a toxicity threshold for generation time of 30 µmol cadmium/litre (3.4 mg/litre). At 50 µmol/litre, cadmium increased the generation time from 24 h to 28 h. Using chain length as a parameter, the toxicity threshold was 15 µmol/litre (1.7 mg/litre). Cell density was found to be affected in batch culture. Cell chlorophyll content and chain length were too variable to show significant effects of the metal in batch culture. 5.2 Soil and litter microorganisms Babich & Stotzky (1977b) showed that several species of fungi tolerated cadmium to a greater degree when grown on cultures amended with clays than in pure culture medium. Sato et al. (1986) exposed the soil bacterium Nitrosomonas europaea both to cadmium concentrations of 0.05, 0.1, and 0.4 mg/litre and to a range of ammonia concentrations (1 to 100 mg/litre as N). Growth was markedly inhibited at the highest cadmium concentration, especially when this was coupled with ammonia concentrations greater than 10 mg/litre. Cadmium toxicity caused a characteristic growth response. Ammonia oxidation proceeded at a reduced rate for approximately 3 days and then fell sharply. The subsequent oxidation proceeded at a diminished but constant rate. At cadmium concentrations of 0.1 mg/litre or less, the toxic effect of cadmium could be partially offset by increasing the ammonia concentration. Prahalad & Seenayya (1988) found that cadmium concentrations of 3.5 mg/litre inhibited the growth of the bacterium Alcaligenes faecalis. However, if the nutrient broth was diluted by one half then growth was inhibited at 2 mg/litre, and with quarter-strength nutrient broth inhibition occurred at 0.5 mg/litre. Hartman (1974) reported less species diversity of soil fungi in areas of high cadmium contamination than in control areas. Samples of Fusarium oxysporium isolated from contaminated and control soils showed different tolerances to cadmium. This suggested the development of some resistance, presumably by selection of more resistant strains. Bond et al. (1976) incubated coniferous forest soil and litter (mixed with a pre-prepared homogeneous soil) in microcosm units and monitored oxygen consumption, carbon dioxide evolution, and bacterial and fungal populations. With cadmium added to a concentration of 0.01 mg/kg soil and, in the initial stages of incubation, with cadmium at 10 mg/kg soil, there was a stimulation of oxygen consumption, suggesting an effect of the metal on uncoupling of respiratory phosphorylation. In the later stages of incubation with cadmium at 10 mg/kg there was a reduction in both oxygen consumption and carbon dioxide evolution. No effect was seen on numbers of organisms in the microcosms. Bewley & Stotzky (1983) investigated the effect of cadmium (100 and 1000 mg/kg soil) on carbon mineralization and on the mycoflora in glucose-supplemented soils amended with clays (kaolinite or montmorillonite) at 9%. Cadmium had no significant effect on the length of the lag period, carbon dioxide evolution or on the amount of carbon mineralized. When subsequent cadmium additions of 2400 and 4000 mg/kg were made to soils previously treated with 100 or 1000 mg/kg, the rate of glucose degradation decreased more in the clay-amended soils than in control soil not amended with clay. Clay protected fungi from the toxic effect of cadmium at 5000 mg/kg. Fungi from clay-treated soils were more sensitive to cadmium in the culture media after they had been isolated from soil pre-treated with cadmium. The authors interpreted these results as showing a reduction in the availability of cadmium to organisms in clay-amended soils. The overall effect would be a prevention of selection of more tolerant strains. Thus, when the organisms were challenged with high doses of cadmium, they would have been more susceptible than organisms from non-amended soil. Naidu & Reddy (1988) incubated black cotton soil (0.8% organic carbon; 55% clay) for up to 8 weeks in the presence of cadmium chloride at concentrations of between 50 and 500 mg cadmium/kg. The ammoniacal nitrogen (NH4-N) concentration increased for the first week at all treatment levels and then decreased at concentrations of 50 mg/kg or less. The initial rise in NH4-N levels led to an increase in nitrate nitrogen (NO3-N) levels, the accumulation of NO3-N being inversely proportional to the cadmium exposure. The authors pointed out that at cadmium concentrations of 100 and 500 mg/kg, hydrolysis of urea was significantly poorer than at other treatment levels, as shown by the lower concentration of NH4-N observed after 1 week. At all cadmium concentrations there was significant accumulation of nitrite nitrogen (NO2-N) at every sample time, compared to control soil, suggesting that cadmium might be toxic to soil nitrification. At all exposure levels cadmium significantly depressed both bacterial and fungal populations. Cadmium concentrations of 10 or 50 mg/kg had no effect on soil actinomycetes, but both 100 and 500 mg/kg significantly reduced the population. Tyler et al. (1974) incubated a mull soil for 7 weeks. Cadmium chloride concentrations of 9 to 18 µmol/g and cadmium acetate concentrations of 9 to 22 µmol/g caused decreases in soil ammonium content and significantly increased nitrate accumulation. 6. TOXICITY TO AQUATIC ORGANISMS Appraisal Cadmium uptake from water by aquatic organisms is extremely variable and depends on the species and various environmental conditions such as water hardness (notably the calcium ion concentration), salinity, temperature, pH, and organic matter content (see chapter 4). The majority of chelating agents decrease cadmium uptake but some, such as dithiocarbamates and xanthates, increase uptake. As a consequence of the variability in cadmium uptake, the toxic impact to aquatic organisms also varies across a wide range of concentrations and is dependent on the species of organism and on the presence of other metal ions, notably calcium and zinc. The lowest recorded 96-h LC50 in a flow system is 16 µg cadmium per litre for the adult shrimp Mysidopsis bahia. A nominal no-observed-effect level (NOEL) of 0.6 µg cadmium/litre was found for Daphnia magna, reproductive rate being the most sensitive parameter. A nomi-nal NOEL has been noted at a similar level (1.7 to 3.4 µg cadmium per litre) with respect to the reproductive effects on brook trout. The available results indicate that the embryonic and larval stages of aquatic organisms are more sensitive than the adult stage. Spinal malformations are induced in cadmium-exposed fish. In addition to causing reproductive effects, cadmium influences the behaviour of aquatic organisms. At low concentrations, cadmium inhibits ion transport systems (10 µg cadmium/litre) and induces metallothionein synthesis (< 1 µg cadmium/litre) in freshwater fish. 6.1 Toxicity to aquatic plants Hutchinson & Czyrska (1975) exposed two floating aquatic weeds, the common duckweed Lemna minor and a floating fern Salvinia natans, to cadmium concentrations of between 0.01 and 1.0 mg/litre for up to 3 weeks. Growth was reduced at all concentrations but especially at 0.05 mg/litre or more. The effect of cadmium on growth became more marked with time. Loss of green coloration (chlorosis) was a common symptom of cadmium toxicity, and at concentrations of 0.5 and 1.0 mg/litre Lemna plants died. When the two species were grown in competition, growth at lower cadmium concentrations (0.01 and 0.05 mg/litre) was greater in Salvinia but less in Lemna than when the plants were grown alone. In a study by Nir et al. (1990), water hyacinth plants were exposed to cadmium concentrations of 0, 0.05, 0.1, 0.4 or 1.0 mg/litre for 7 days. Concentrations of 0.1 mg/litre or less had no significant effect on wet or dry biomass gain or on chlorophyll a content. Concentrations of 0.4 or 1.0 mg/litre significantly reduced both wet biomass gain and chlorophyll a content but had no significant effect on dry biomass gain. The chlorophyll a content of leaves decreased with time in plants exposed to 0.4 mg/litre. After 3 weeks of exposure, the chlorophyll a levels were 75% lower than in control plants. 6.2 Toxicity to aquatic invertebrates Cadmium is moderately to highly toxic to aquatic invertebrates (see Tables 9 and 10). Its toxic effect is dependent on a variety of environmental variables. Factors that reduce the free ionic concentration, e.g., water hardness, salinity, chelating agents, and high organic content of water, tend to reduce the toxic effect of cadmium. The presence of zinc increases the toxic effect of cadmium on invertebrates. 6.2.1 Acute and short-term toxicity The acute toxicity of cadmium to aquatic invertebrates, as assessed in laboratory tests, is summarized in Tables 9 and 10. The most notable features are the variability in cadmium toxicity between different organisms and the effects of temperature, salinity, and water hardness. There is considerable variation even amongst closely related species. In a study by Canton & Slooff (1982), the water flea Daphnia magna was exposed to cadmium over a 48-h period. At a water hardness of 1 mmol/litre there was no mortality at 16 µg cadmium per litre, while at a water hardness of 2 mmol/litre there was no mortality or abnormal behaviour at 17 µg/litre. Clubb et al. (1975b) investigated the toxicity of cadmium to nine species of aquatic insects, but seven of the species tested were too insensitive to the effects of the metal for the LC50 to be determined. The insensitive species were Atherix variegata, Hexatoma sp., Holorusia sp., Acroneuria pacifica, Arcynopteryx signata, Pteronarcys californica, and Brachycentrus americanus; these species represented Dipterans (true flies), Plecoptera (stoneflies), and Tricoptera (caddis flies). Table 9. Toxicity of cadmium to marine or estuarine invertebrates Organism Size/ Stat/ Temperature Salinity pH Duration LC50 Reference age flow a (°C) (%) (h) (mg/litre)b Purple sea urchin embryo stat 8.2-8.4 30 7.8-8.1 120 0.5 (0.4-0.6) m Dinnel et al. (1989) (Strongylocentrotus purpuratus) Green sea urchin embryo stat 8.2-8.4 30 7.8-8.1 120 1.8 (1.5-2.2) m Dinnel et al. (1989) (Strongylocentrotus droebachiensis) Sand dollar embryo stat 12.5-13.0 30 8.0-8.1 72 7.4 (5.2-10.8) m Dinnel et al. (1989) (Dendraster excentricus) Starfish 24.5 g stat 20 20 8.0 24 12 n Eisler (1971) (Asterias forbesi) 24.5 g stat 20 20 8.0 48 1.0 n Eisler (1971) 24.5 g stat 20 20 8.0 96 0.82 n Eisler (1971) 11.2 g stat 20 20 7.8 24 71 n Eisler & Hennekey (1977) 11.2 g stat 20 20 7.8 48 7.1 n Eisler & Hennekey (1977) 11.2 g stat 20 20 7.8 96 0.7 n Eisler & Hennekey (1977) American oyster embryo stat 26 25 7.0-8.5 48 3.8 (2.85-4.48) n Calabrese et al. (1973) (Crassostrea virginica) Mussel stat 18.5 32.9 7.9 96 1.62 (1.19-2.22) m Ahsanullah (1976) (Mytilus edulis planulatis) Blue mussel 4 g stat 20 8.0 24 > 200 n Eisler (1971) (Mytilus edulis) 4 g stat 20 8.0 48 165 n Eisler (1971) 4 g stat 20 8.0 96 25 n Eisler (1971) Table 9 (contd). Organism Size/ Stat/ Temperature Salinity pH Duration LC50 Reference age flow a (°C) (%) (h) (mg/litre)b Soft-shell clam 5.2 g stat 20 20 8.0 24 > 20 n Eisler (1971) (Mya arenaria) 5.2 g stat 20 20 8.0 48 50 n Eisler (1971) 5.2 g stat 20 20 8.0 96 2.2 n Eisler (1971) 4.6 g stat 20 20 7.8 24 32 n Eisler & Hennekey (1977) 4.6 g stat 20 20 7.8 48 2.5 n Eisler & Hennekey (1977) 4.6 g stat 20 20 7.8 96 0.7 n Eisler & Hennekey (1977) Bay scallop 20-30 mm stat 20 25 8.0 24 8.2 n Nelson et al. (1976) (Argopecten 20-30 mm stat 20 25 8.0 48 3.21 n Nelson et al. (1976) irradians) 20-30 mm stat 20 25 8.0 72 2.18 n Nelson et al. (1976) 20-30 mm stat 20 25 8.0 96 1.48 (0.95-2.31) n Nelson et al. (1976) Atlantic oyster 0.6 g stat 20 20 8.0 24 158 n Eisler (1971) drill 0.6 g stat 20 20 8.0 48 28 n Eisler (1971) (Urosalpinx 0.6 g stat 20 20 8.0 96 6.6 n Eisler (1971) cinerea) Eastern mud snail 0.56 g stat 20 20 8.0 24 > 200 n Eisler (1971) (Nassarius 0.56 g stat 20 20 8.0 48 125 n Eisler (1971) absoletus) 0.56 g stat 20 20 8.0 96 10.5 n Eisler (1971) Ragworm 8 g stat 20 20 8.0 24 25 n Eisler (1971) (Nereis virens) 8 g stat 20 20 8.0 48 25 n Eisler (1971) 8 g stat 20 20 8.0 96 11 n Eisler (1971) 7.6 g stat 20 20 7.8 24 56 n Eisler & Hennekey (1977) 7.6 g stat 20 20 7.8 48 9.3 n Eisler & Hennekey (1977) 7.6 g stat 20 20 7.8 96 0.7 n Eisler & Hennekey (1977) Copepod nauplius stat 22 10 96 0.06 (0.001-0.2) m Roberts et al. (1982) (Eurytemora affinis) Copepod adult stat 22 10 96 0.38 (0.006-1.52) m Roberts et al. (1976) (Acartia tonsa) Table 9 (contd). Organism Size/ Stat/ Temperature Salinity pH Duration LC50 Reference age flow a (°C) (%) (h) (mg/litre)b Harpacticoid copepod adult stat 20-22 3 96 0.43 (0.31-0.55) Bengtsson & Bergstrom (1987) (Nitocra spinipes) adult stat 20-22 7 96 0.66 (0.53-0.82) Bengtsson & Bergstrom (1987) adult stat 20-22 15 96 0.78 (0.41-120) Bengtsson & Bergstrom (1987) Marine amphipod young stat 10 96 3.5 m Wright & Frain (1981a) (Marinogammarus adult stat 10 96 13.3 m Wright & Frain (1981a) obtusatus) Mysid shrimp adult flow 22 20 7.3 96 0.036 (0.022-0.081) m Roberts et al. (1982) (Neomysis adult flow 22 20 7.8 96 0.02 (0.015-0.027) m Roberts et al. (1982) americanus) Mysid shrimp adult stat 22 20 7.3 96 0.017 m Roberts et al. (1982) (Mysidopsis bahia) adult stat 22 20 7.7 96 0.029 (0.013-0.043) m Roberts et al. (1982) flow c20-28 15-23 96 0.016 (0.013-0.02) m Nimmo et al. (1978) Shrimp stat 18.7 32.1 8.0 120 2.3 (1.05-5.06) m Ahsanullah (1976) (Palaemonetes sp.) stat 18.7 32.1 8.0 168 1.85 (1.32-2.59) m Ahsanullah (1976) 0.38 g flow 16.8 96 6.8 (5.2-9.76) m Ahsanullah (1976) 0.16 g flow 17.8 8.1 96 6.4 (5.73-7.19) m Ahsanullah (1976) Sandworm 0.37 g stat 18.5 32.7 8.1 168 6.4 (5.82-7.1) m Ahsanullah (1976) (Neanthes vaali) Sand shrimp 0.25 g stat 20 20 8.0 24 2.4 n Eisler (1971) (Crangon 0.25 g stat 20 20 8.0 48 0.5 n Eisler (1971) septemspinosa) 0.25 g stat 20 20 8.0 96 0.32 n Eisler (1971) Sand shrimp adult flow 10.2 28.6 7.9 96 2.3 (1.7-5.1) m Dinnel et al. (1989) (Crangon spp.) Table 9 (contd). Organism Size/ Stat/ Temperature Salinity pH Duration LC50 Reference age flow a (°C) (%) (h) (mg/litre)b Grass shrimp 0.33 g stat 20 20 8.0 24 43 n Eisler (1971) (Palaemonetes 0.33 g stat 20 20 8.0 48 3.7 n Eisler (1971) vulgaris) 0.33 g stat 20 20 8.0 96 0.32 n Eisler (1971) Shrimp stat 35 96 2.07 (± 0.22) m McClurg (1984) (Penaeus indicus) Pink shrimp flow 25 20 96 4.6 m Bahner & Nimmo (1975) (Penaeus duorarum) Grapsid crab 1.47 g stat 17.8 32.6 8.1 168 14 (11.2-17.5) m Ahsanullah (1976) (Paragrapsus 1.08 g stat 17.1 168 16.7 (15.11-18.45) Ahsanullah (1976) quadridentatus) Hermit crab 0.47 g stat 20 20 8.0 24 > 200 n Eisler (1971) (Pagurus 0.47 g stat 20 20 8.0 48 3.7 n Eisler (1971) longicarpus) 0.47 g stat 20 20 8.0 96 0.32 n Eisler (1971) 0.5 g stat 20 20 7.8 24 15 n Eisler & Hennekey (1977) 0.5 g stat 20 20 8.0 96 1.3 n Eisler & Hennekey (1977) Shore crab 5.9 g stat 20 20 8.0 24 100 n Eisler (1971) (Carcinus maenus) 5.9 g stat 20 20 8.0 48 16.6 n Eisler (1971) 5.9 g stat 20 20 8.0 96 4.1 n Eisler (1971) Dungeness crab zoea stat 8.5 30 8.1 96 0.2 (0.1-0.4) m Dinnel et al. (1989) (Cancer magister) Squid larva stat 8.6 30 8.1 96 > 10.2 m Dinnel et al. (1989) (Loligo opalescens) a stat = static conditions (water unchanged for duration of test); flow = flow-through conditions (cadmium concentration in water continuously maintained) unless stated otherwise b organisms exposed to cadmium added as cadmium chloride; m = measured; n = nominal c intermittent flow-through conditions Table 10. Toxicity of cadmium to freshwater invertebrates Organism Size/ Stat/ Temperature Hardness c pH Duration LC50d Reference age flow a (°C) (mg/litre) (h) (mg/litre) Snail adult stat 20-22 6.7 24 7.6 n Wier & Walter (1976) (Physa adult stat 20-22 6.7 48 4.25 n Wier & Walter (1976) gyrina) adult stat 20-22 6.7 96 1.37 n Wier & Walter (1976) adult stat 20-22 6.7 228 0.83 n Wier & Walter (1976) immature stat 20-22 6.7 48 0.69 n Wier & Walter (1976) immature stat 20-22 6.7 96 0.41 n Wier & Walter (1976) Snail flow b 15 7.1-7.7 168 0.114 m Spehar et al. (1978a) (Physa integra) Snail 10-12 mm flow 12 128-176 7.7 24 4.4 m Williams et al. (1985) (Physa 10-12 mm flow 12 128-176 7.7 48 2.1 m Williams et al. (1985) fontinalis) 10-12 mm flow 12 128-176 7.7 96 0.8 m Williams et al. (1985) Isopod 8-10 mm flow 12 128-176 7.7 24 13 m Williams et al. (1985) (Asellus 8-10 mm flow 12 128-176 7.7 48 3.6 m Williams et al. (1985) aquaticus) 8-10 mm flow 12 128-176 7.7 96 0.6 m Williams et al. (1985) Scud 10 96 0.12 m Wright & Frain (1981b) (Gammarus 8-12 mm flow 12 128-176 7.7 24 1.6 m Williams et al. (1985) pulex) 8-12 mm flow 12 128-176 7.7 48 0.4 m Williams et al. (1985) 8-12 mm flow 12 128-176 7.7 96 0.02 m Williams et al. (1985) Water flea adult stat 10 0.85 meq/ 7.2 48 0.055 e (0.032-0.095) n Baudouin & Scoppa (1974) (Daphnia hyalina) litre Water flea 1 day stat 20 7.6-7.7 24 3 n Kuhn et al. (1989) (Daphnia < 1 day stat 20-22 110-130 7.8 48 0.04 (0.02-0.07) n Hall et al. (1986) magna) < 1 day stat 20-22 190-210 7.7 48 0.08 (0.06-0.1) n Hall et al. (1986) < 1 day stat 18-20 1 mmol/ 48 0.03 m Canton & Slooff (1982) litre Table 10 (contd). Organism Size/ Stat/ Temperature Hardness c pH Duration LC50d Reference age flow a (°C) (mg/litre) (h) (mg/litre) Water flea < 1 day stat 20-22 110-130 7.8 48 0.07 n Hall et al. (1986) (Daphnia < 1 day stat 20-22 110-130 7.7 48 0.1 (0.07-0.12) n Hall et al. (1986) ulex) < 1 day stat 19-22 7.7 96 0.047 (0.042-0.052) n Bertram & Hart (1979) Copepod adult stat 10 0.85 meq/ 7.2 48 3.8 e (2.3-6.3) n Baudouin & Scoppa (1974) (Cyclops abyssorum) litre Copepod adult stat 10 0.85 meq/ 7.2 48 0.55 e (0.39-0.77) n Baudouin & Scoppa (1974) (Eudiaptomus padanus) litre Crayfish flow 19-21 24-28 6.7-7.0 96 6.1 (4.7-7.9) m Mirenda (1986) (Orconectes virilis) Mayfly flow 10 7.8 96 28 m Clubb et al. (1975b) (Ephemerella grandis grandis) Midge 10-12 mm flow 12 128-176 7.7 96 300 m Williams et al. (1985) (Chironomus riparius) Stonefly flow 10 7.8 96 18 m Clubb et al. (1975b) (Pteronarcella badia) Stonefly 10-15 mm flow 12 128-176 7.7 96 520 m Williams et al. (1985) (Hydropsyche angustipennis) a stat = static conditions (water unchanged for duration of test); flow = flow-through conditions (cadmium concentration in water continuously maintained) unless stated otherwise b intermittent flow-through conditions c hardness expressed as mg CaCO3/litre unless stated otherwise d organisms exposed to cadmium added as cadmium chloride unless otherwise stated; m = measured concentration; n = nominal concentration e organism exposed to cadmium as cadmium sulfate Spehar et al. (1978a) reported a decreased survival of the water snail Physa integra at a cadmium concentration of 85.5 µg/litre after 7 days of exposure. After 21 days of exposure, survival was significantly decreased at 27.5 µg/litre, the next lowest concentration tested. Some snails exposed to these concentrations developed a condition in which the animal was extended from the shell but unable to attach the foot or crawl. A white mucus layer covered the exposed foot region of some snails and these subsequently died. Concentrations tested were not high enough to obtain a 4-day LC50 value but the 7-day LC50 of 114 µg/litre was approximately 11 times higher than the 28-day LC50 value of 10.4 µg/litre. Mirenda (1986) reported a 14-day LC50 of 0.7 mg cadmium per litre for the crayfish Orconectes virilis under flow-through conditions. Pesch & Stewart (1980) estimated the 10-day LC50 for bay scallops Argopecten irradians to be 0.53 mg/litre in flowing sea water. The EC50 (for growth) for the same species over 42 days was 0.078 mg/litre. Byssal thread detachment, which precedes death, showed an EC50 of 0.54 mg/litre of cadmium 8 days into the test and before there was any appreciable mortality. Robinson et al. (1988) compared 10-day LC50 values for freshly collected and cultured infaunal amphipods ( Rhepoxynius abronius). Cultured amphipods appeared normal and survived well (93%) under control toxicity test conditions, but were more sensitive to cadmium in sediment (10-day LC50 = 4.4 mg/kg) than were freshly caught amphipods (10-day LC50 = 8.7 mg/kg). When Winner (1988) exposed Daphnia magna and Ceriodaphnia dubia to cadmium for 7 days, the most sensitive indicators were mean body length of primiparous females in D. magna, which was significantly reduced at 2 µg cadmium/litre, and the total young per female in C. dubia, significantly reduced at 1 µg/litre. 22.214.171.124 Effects of temperature and salinity on acute toxicity An increase in toxicity as temperature increases and as salinity decreases is valid for all organisms that have been tested with these variables (Tables 9 and 10). O'Hara (1973) investigated the effects of temperature and salinity on the toxicity of cadmium to adult male and female fiddler crabs ( Uca pugilator). Mortality was greatest at high temperatures and low salinities in tests lasting 240 h. LC50 values varied from 2.9 mg/litre for the lowest salinity (10%) and highest temperature (30 °C) to 47.0 mg/litre for the highest salinity (30%) and lowest temperature (10 °C). Frank & Robertson (1979) exposed the blue crab ( Callinectes sapidus) to cadmium chloride at salinities of 1, 15, and 35%. Like O'Hara, they found a decrease in cadmium toxicity with increase in salinity. For example, 96-h LC50 values were 0.32, 4.7, and 11.6 mg cadmium/litre for the three salinities, respectively. Rosenberg & Costlow (1976) reported increased cadmium toxicity during larval development of two estuarine crab species as salinity decreased and increased toxicity as temperature increased. Voyer & Modica (1990) found the same pattern with the mysid shrimp Mysidopsis bahia. For salinities of 10 and 30%, the 96-h LC50 values ranged from 15.5 to 28 µg cadmium/litre at a temperature of 25 °C and from 47 to 84 µg/litre at a temperature of 20 °C. At 30 °C the 96-h LC50 was < 11 µg/litre at both salinities. However, when Robert & His (1985) exposed embryos and larvae of the Japanese oyster Crassostrea gigas to cadmium concentrations of up to 50 µg/litre at various salinities (20 to 35%), decreasing the salinity severely affected the development of the oysters but cadmium had no effect. At temperatures higher than 11 °C, the combined effect of temperature and cadmium caused a heavy stress to the copepod Tisbe holothuriae so that the effects of salinity were masked (Verriopoulos & Moraitou-Apostolopoulou, 1981). 126.96.36.199 Effect of water hardness Using either artificial hard water (hardness: 180 mg CaCO3 per litre) or dechlorinated tap water (hardness: 60 mg/litre), Niederlehner et al. (1984) conducted short-term tests on the effects of cadmium on the freshwater oligochaete Aeolosoma headlyi. Mortality and population growth/maintenance were assessed over 10 to 14 days. The authors established NOEL values for population growth of 32.0 and 53.6 µg/litre for hard water (two replicate tests), whereas the NOEL for the softer water was 17.2 µg/litre. The 48-h LC50 values were 4.98 and 1.2 mg/litre, respectively, for hard and softer water. 188.8.131.52 Effect of organic materials and sediment When Schuytema et al. (1984) exposed Daphnia magna to cadmium for a period of 48 h, the mean LC50 value was 39 µg/litre in water and 91 µg/litre in a water-sediment slurry. Giesy et al. (1977) found that cadmium was more toxic to water fleas ( Simocephalus serrulatus) exposed in well water with a low organic content than to those exposed in pond water with a high organic content. The authors isolated a series of organic fractions from the pond water by ultrafiltration. Protection against cadmium toxicity was afforded by fractions of intermediate relative molecular mass (ranging from approximately 500 to 300 000 daltons). The fraction with a relative molecular mass in excess of 300 000 daltons marginally increased the toxicity of cadmium. Kemp & Swartz (1988) compared the acute toxicity of interstitial and particle-bound cadmium to the marine infaunal amphipod Rhepoxynius abronius. The cadmium concentration in interstitial water was kept constant whereas the sediment cadmium level was varied by using perfusion through the sediment with peristaltic pumps. The principal cause of toxicity was found to be cadmium dissolved in interstitial water, between 70 and 88% of the toxicity being predictable from interstitial water concentrations. 184.108.40.206 Lifestage sensitivity When Calabrese et al. (1973) investigated the toxicity of cadmium to embryos of the American oyster Crassostrea virginica, there was no mortality at 1 mg/litre and the 48-h LC50 and LC100 values were 3.8 and 6 mg/litre, respectively. Johnson & Gentile (1979) found the larva of the American lobster Hommarus americanus to be sensitive to cadmium; the 96-h LC50 in static bioassays was 78 µg/litre. The mortalities after 96 h at concentrations of 10 and 30 µg/litre were 3% and 10%, respectively. There is a very steep increase in toxicity of cadmium to lobster larvae between 24 and 96 h. The 24-h LC50 is approximately 1 mg/litre; at this concentration the mortality reaches 100% within 48 h. Verriopoulos & Moraitou-Apostolopoulou (1982) found that the different life-stages of the copepod Tisbe holothuriae showed differences in sensitivity to cadmium. One-day-old nauplius larvae of the copepod were the most sensitive with an LC50 of 0.538 mg/litre, expressed as ions of cadmium, while 5-day-old nauplii showed an LC50 of 0.645 mg/litre. The value for 10-day-old copepodids (0.906 mg/litre) was not significantly different from that for adult females (0.916 mg/litre). Females with ovigerous sacs were slightly more sensitive, with an LC50 of 0.873 mg/litre. When Robinson et al. (1988) exposed the infaunal phoxocephalid amphipod Rhepoxynius abronius to sediment contaminated with cadmium, the 10-day LC50 values were 8.2 mg/kg for juveniles and 11.5 mg/kg for adults. Nebeker et al. (1986) exposed Daphnia magna of different ages to cadmium for a period of 48 h. Mean EC50 (immobilization) values ranged from 23 µg/litre for 6 day-old water fleas to 164 µg/litre for 2-day-old Daphnia. Tests on Daphnia of different ages, conducted in water of different hardness (32 or 76 mg CaCO3 per litre), with or without feeding and in two different sizes of container, resulted in a wide range of EC50 values (4 to 307 µg/litre). There was no consistent effect of any of these variables other than the age of the test animals. Very young animals were relatively tolerant, with a mean EC50 value of 109 µg/litre. McCahon et al. (1989) exposed both cased and uncased 1st, 3rd, and 4th larval instars of the caddis fly Agapetus fuscipes to cadmium chloride. The LC50 values ranged from 295 to > 1000 mg cadmium/litre at 24 h and from 50 to 320 mg/litre at 96 h. First instar larvae were significantly more sensitive than 3rd or 4th instar larvae and at all ages cased animals were more resistant than uncased. 220.127.116.11 Other factors affecting acute and short-term toxicity Chandini (1988; 1989) found that increasing the food source of the cladocerans Daphnia carinata and Echinisca triserialis greatly reduced the toxicity of cadmium, expressed as the 48-h or 96-h LC50, and the effect of cadmium on various other life history parameters, such as fecundity, growth, and age at first reproduction. Verriopoulos & Moraitou-Apostolopoulou (1981) exposed adult females of the copepod Tisbe holothuriae to cadmium and found that the oxygen concentration in the water was negatively related to mortality and that population density was positively related after cadmium exposure. Clubb et al. (1975a) showed that the toxicity of cadmium to aquatic insects decreased with decreasing dissolved oxygen levels in the test water. McCahon et al. (1988) exposed the amphipod Gammarus pulex to cadmium chloride under static conditions and reported a 96-h LC50 of 0.05 mg cadmium/litre. The acute toxicity of cadmium to G. pulex parasitized by the acanthocephalan Pomphorhynchus laevis at several levels of infection was investigated. Toxicity, expressed as LC50, did not differ significantly between uninfected and infected amphipods. Kay & Haller (1986) fed water hyacinth weevils Neochetina eichhorniae on water hyacinths containing cadmium from previous exposure to the metal. There was no mortality in weevils fed on leaves containing up to 17.2 mg cadmium/kg. At this exposure level, the weevils accumulated a cadmium body burden of 36.67 mg/kg over 20 days. 6.2.2 Long-term toxicity In a study by Zaroogian & Morrison (1981), adult and larval oysters of the species Crassostrea virginica were exposed to cadmium at concentrations of 5 or 15 µg/litre. Some adults were exposed (for 35 or 37 weeks) to cadmium at these two concentrations prior to spawning, and larvae from these pre-treated adults and from control treated adults were reared in either control sea water or sea water containing 5 or 15 µg/litre cadmium. In all there were 11 treatment combinations. The highest larval mortality occurred when larvae from parents treated with 15 µg/litre were reared in sea water containing 15 µg/litre for 3 weeks. However, larvae that survived this treatment grew to lengths not significantly different from controls. The effects observed with other treatment combinations were only temporary. Growth and development were slowed but those larvae that survived ultimately developed normally and to the same size as controls. When Holcombe et al. (1984) exposed snails, from embryos through to adult sexual maturity, to cadmium chloride, there was a delayed hatch, a reduction in percentage hatch and survival, and reduced growth when compared to controls. Based on these effects, the authors suggested a maximum acceptable cadmium concentration in water of between 4 and 8 µg/litre from one test and between 2 and 5 µg/litre from a replicate test. Lussier et al. (1985) conducted life-cycle tests, over 35 days, on the mysid shrimp Mysidopsis bahia. Cadmium affected survival primarily and no reproductive effects were noted at sublethal concentrations. Following a long-term investigation in the laboratory and in the field, Marshall (1979) suggested a chronic LC10 for the water flea Daphnia galeata mendotae of 0.15 µg cadmium/litre. Winner (1986) exposed Daphnia pulex chronically (21 or 42 days) to cadmium, added as cadmium sulfate, under different water conditions. Increasing the water hardness from 58 to 116 mg/litre reduced the toxic effect of the cadmium but a further increase to 230 mg/litre had no further effect. The most sensitive aspect of the Daphnia life history to cadmium was the abortion rate of young. Humic acid had no effect on this parameter in soft or medium-hard water but increased the toxic effect of cadmium in hard water. Mortality was increased by humic acid (0.75 or 1.50 mg/litre) at all water hardness levels. van Leeuwen et al. (1985) calculated 14-day and 21-day LC50S for Daphnia magna of 24 and 14 µg cadmium/litre, respectively. No effect on mortality was seen at 3.2 µg/litre. The lowest concentrations producing significant mortality of young amphipods during 6-week exposures to cadmium were 1 µg/litre for Hyalella azteca and 3.2 µg/litre for Gammarus fasciatus (Borgmann et al., 1989b). 6.2.3 Reproductive effects Mysing-Gubala & Poirrier (1981) conducted laboratory experiments on the effect of cadmium on the freshwater sponge Ephydatia fluviatilis. Sponge cuttings were exposed to cadmium concentrations ranging from 0.001 to 1.0 mg/litre for 1 month. At the end of the experimental period, the cuttings were classified in four groups depending on whether the sponge survived, whether it produced asexual reproductive gemmules, and whether the silicaceous spicules of the gemmules were normal or malformed. There was some effect of cadmium even at concentrations as low as 0.001 mg/litre, 17% of the sponge cuttings showing no gemmule production and 33% showing malformed spicules. At concentrations of 0.5 and 1.0 mg/litre, all of the sponge cuttings died. Lee & Xu (1984) investigated the effects of cadmium at 0.5 and 0.1 mg/litre on the fertilization and development of sea urchin eggs and the development of Amphioxus. At both cadmium concentrations, sea urchin development was normal to the gastrulation stage but all the plutei were abnormal. The effects on Amphioxus development were different; cleavage of eggs was not affected by cadmium at 0.5 or 0.1 mg/litre but neurulation was. Dinnel et al. (1989) exposed sperm from various sea urchin species to cadmium chloride for 60 min and assessed fertilization success; EC50 values ranged from 12 to 26 mg/litre. An EC50 of 8 mg/litre was calculated for the sand dollar Dendraster excentricus. Den Besten et al. (1989) exposed the sea star Asterias rubens to cadmium chloride at a concentration of 25 µg cadmium/litre for 5 months. No effect on spermatozoa was found, but maturation of oocytes was delayed and early development of embryos was adversely affected. Conrad (1988) studied the effect of cadmium on newly fertilized eggs of the mud snail Ilyanassa obsoleta. No apparent effect was observed at a cadmium concentration of 10-6 mol/litre. The minimum concentrations producing abnormal veliger development and abnormal late cleavage and stopping early cleavage were 10-5-10-4 mol/litre, 10-3 mol/litre, and 10-3 mol/litre, respectively. Biesinger & Christensen (1972) estimated a 3-week 16% reproductive impairment concentration of 0.17 µg cadmium/litre. In a 21-day reproduction test on Daphnia magna, Kuhn et al. (1989) determined a nominal NOEL of 0.6 µg Cd2+/litre, reproduction rate being the most sensitive parameter. A nominal NOEL of 1 µg/litre was found when daphnids were exposed to the cadmium chloride salt. Reproduction of Daphnia magna was completely inhibited at concentrations exceeding 3.2 µg/litre and time-dependant survival and reproduction were significantly reduced at 1.8 µg/litre. No effects on reproduction were observed at 1 µg/litre (van Leeuwen et al., 1985). When Bertram & Hart (1979) exposed the cladoceran Daphnia pulex to cadmium concentrations of 1 to 30 µg/litre, there was no significant effect on the number of days required for onset of reproduction or on its frequency. At 1 µg/litre no significant effect on longevity of individuals was observed, but at concentrations of 5 µg/litre or more there was a significant reduction. All cadmium concentrations caused a reduction in the average brood size, the average number of broods per adult, and the total number of progeny. The authors also exposed daphnids to cadmium-contaminated food in the form of the phytoplankton Chlorella, the cadmium concentration being 0.3-0.6 µg cadmium/litre of medium. This resulted in a significant reduction in the average number of broods per adult and the average brood size. However, there was no significant effect on average longevity of individuals, the percentage of adults producing broods, the average number of days to the first brood or the average number of days between broods. Bengtsson & Bergstrom (1987) exposed newly fertilized female harpacticoids ( Nitocra spinipes) to cadmium chloride at two salinities for 13 days. At 3% the fecundity EC50 was 37-46 µg/litre at a salinity of 3% and 6-15 µg/litre at 15%. Williams et al. (1987) provided water containing nominal cadmium chloride concentrations of 0, 0.3, 30, 100 or 300 mg cadmium/litre cadmium to newly-emerged adults of the midge Chironomus riparius in which they could lay their eggs. The females preferred to lay their eggs in water with no cadmium or in the lower concentrations; significant preferences were recorded. Eggs of Chironomids are laid within a protective gelatinous matrix. Eggs exposed to cadmium after complete formation of the matrix in control water were unaffected (the hatch was 80-100%), whereas those exposed after removal of the matrix had a reduced hatching rate (60%) at all test concentrations. Eggs laid directly in water containing cadmium were unaffected by a concentration of 0.3 mg/litre, but hatching rates were reduced to 45% at 30 mg/litre, 8% at 100 mg/litre, and 0% at 300 mg/litre. Clearly, cadmium only affects the unprotected egg, either when it is newly laid and before its gelatinous protection has developed or when this gelatinous matrix is removed. In a flow-through study, Hatakeyama (1987) exposed the chironomid Polypedilum nubifer, from the egg stage, to cadmium chloride. There was no effect on total number of adults, emergence success, sex ratio, total number of egg clusters, oviposition success or hatchability at 10 µg cadmium/litre. There was a significant decrease in the total number of adults and emergence success at 20 µg/litre or more and in the total number of egg clusters at 40 µg/litre or more. At 80 µg/litre survival and reproduction were very significantly depressed. In a separate experiment midges were fed cadmium-contaminated food (dried yeast) at concentrations of 22, 220, and 1800 mg cadmium/kg. Emergence success was decreased at 220 and 1800 mg/kg but no other effects were observed. 6.2.4 Physiological and biochemical effects Berglind (1986) investigated the effect of cadmium alone and in combination with other metals on the delta-aminolevulinic acid dehydratase (ALAD) activity of Daphnia magna. ALAD activity was enhanced by cadmium alone (at 2 µg/litre but not at 0.2 µg/litre) but this enhancement was abolished in the presence of zinc at 200 µg/litre. Vernberg et al. (1977) investigated sub-lethal concentrations of cadmium and their effects on the adult shrimp Palaemonetes pugio under static and flow-through conditions. Adult shrimps were exposed to cadmium at 50 µg/litre. This shrimp is highly tolerant to cadmium and after exposure to 23 mg/litre the mortality rate was only 10%. The authors found increased uptake of cadmium into the body of the shrimps with decreasing salinity. Similarly, at higher salinity levels, toxicity was lower. In shrimps kept at the lowest salinity level (5%), where cadmium body burden reached 40 mg/kg, there was inhibition of moulting. At more moderate cadmium body burdens of 23 mg/kg and 10 mg/kg, observed at salinities of 10% and 20%, respectively, moulting was stimulated by the presence of cadmium. An investigation of the effects of cadmium on respiratory rate was inconclusive because of considerable variation. The authors considered that the flow-through system more nearly approximated field conditions than the static system. The relationship between salinity and cadmium uptake was eliminated in a static system at 15 °C, but was quite clearly present when the system was flow-through. After an extensive study on the effects of cadmium on three species of shrimps, Penaeus duorarum, Palaemonetes pugio, and Palaemonetes vulgaris, Nimmo et al. (1977) reported sublethal histological effects and blackening and damage to gill filaments. Placing shrimps with blackened gills in cadmium-free water resulted in the sloughing-off of blackened portions of the branchia and the shrimps appeared normal within 14 days. The effect on the gills occurred with exposure to cadmium concentrations approaching the LC50 which, with an exposure duration of 96 h, was 0.76 mg/litre for Palaemonetes vulgaris and 3.5 mg/litre for Penaeus duorarum. Thurberg et al. (1973) exposed two species of crabs ( arcinus maenus and Cancer irroratus) to various concentrations of cadmium chloride for 48 h and at five different salinities. At the end of each exposure period, tests of blood serum osmolarity and gill tissue oxygen consumption were performed. Cadmium increased the osmolarity of Carcinus serum above its normal hyperosmotic state and reduced oxygen consumption by the gill tissue of both species. Effects on oxygen consumption were dose related over a range of cadmium concentrations from 0 to 4 mg/litre. Cadmium has been shown to cause an increase in oxygen consumption rates in the mud snail Nassarius obsoletus at concentrations of between 0.5 and 4.0 mg/litre over a 72-h exposure period (MacInnes & Thurberg, 1973). A similar increase in oxygen consumption rate was observed in the marine snail Murex trunculus during chronic exposure to 0.05 mg cadmium/litre (Dalla Via et al., 1989) and in crabs ( Callinectes similis) exposed to cadmium concentrations of between 2.48 and 10.05 mg/litre for up to 96 h (Ramirez et al., 1989). 6.2.5 Behavioural effects Olla et al. (1988) monitored the burrowing behaviour of three polychaete species, Nereis virens, Glycera dibranchiata, and Nephtys caeca during a 28-day exposure to a sediment concentration of 40 mg cadmium/kg. Most comparisons of burrowing times and rates between exposed and unexposed worms were not statistically significant. Four out of 15 comparisons gave significant results but these were randomly spread amongst species and exposure periods. The authors concluded that these results would probably have little ecological significance. The feeding behaviour of G. dibranchiata was also monitored but no significant effect of the cadmium treatment was found. 6.2.6 Interactions with other chemicals Sunda et al. (1978) carried out experiments in diluted sea water with various concentrations of the chelating agent nitrilotriacetic acid (NTA) to determine the relationship between the chemical speciation of cadmium and the toxicity of the metal to the grass shrimp Palaemonetes pugio. After 4 days of exposure to a given concentration of cadmium chloride, shrimp mortality decreased with increasing salinity and increasing concentration of the chelating agent. The protective effect of high salinity or NTA was attributable to complexation of cadmium; mortality was related to the measured free cadmium ion concentration, which was determined by measuring total concentration of cadmium and deducting the calculated level of complexation by either the chloride ion or NTA. The mortality at a free cadmium ion concentration of approximately 4 x 10-7 mol/litre was 50%. In a study of the combined effects of zinc and cadmium on the shrimp Callianassa australiensis by Negilski et al. (1981), there was interaction between the two metals; in combination they gave greater mortality than would be expected if there were no interaction. The authors demonstrated that each metal increased the accumulation of the other. 6.2.7 Tolerance Khan et al. (1988) exposed two different populations of grass shrimp Palaemonetes pugio to cadmium under static conditions and calculated 96-h LC50 values of 3.28 mg/litre for shrimps from an industrialized area and 1.83 mg/litre for those from a non- industrialized area. Pre-exposure of shrimps to 0.05 mg cadmium per litre caused an increase in the LC50 values to 6.81 and 3.89 mg/litre for the two respective populations. Moraitou-Apostalopoulou et al. (1982) collected the shrimp Palaemon elegans from two different areas with different natural concentrations of cadmium and investigated the sublethal effects of the metal. They found that cadmium decreased respiration rates in the shrimp at sublethal concentrations. Shrimps sampled from an area with a natural cadmium concentration of 0.6 µg/litre were more tolerant to the metal than were those from an area with 0.1 µg/litre. There was also a difference in the acute toxicity of cadmium to the two populations, indicating that tolerance had developed. In an earlier study (Moraitou-Apostolopoulou et al., 1979), similar development of tolerance was reported for a copepod Acartia clausi. In a study of the toxicity of cadmium to the freshwater cyclopoid copepod Tropocyclops prasinus mexicanus, Lalande & Pinel-Alloul (1986) sampled animals from three Quebec lakes, one polluted and two unpolluted with cadmium. The cultures from the two unpolluted lakes showed lower LC50 values in 48-h tests than the culture from the polluted lake. However, the polluted lake had a significantly higher hardness (120 mg calcium carbonate per litre) than the other two lakes (10 mg/litre). 6.2.8 Model ecosystems Borgmann et al. (1989a) established a Daphnia-phytoplankton model ecosystem and exposed it to cadmium sulfate at concentrations of 1, 5 and 15 µg cadmium/litre. At 5 µg/litre the Daphnia population collapsed after 9 weeks of treatment and chlorophyll levels increased. At 15 µg/litre the Daphnia population collapsed after 5 weeks, but chlorophyll levels remained low. There appeared to be no effect on this model system at 1 µg/litre. 6.3 Toxicity to Fish The toxicity of cadmium has been studied in a variety of fish species in both fresh and sea water at various temperatures and dissolved oxygen concentrations. Generally, increasing dissolved salt concentration decreases the toxicity, whereas increasing temperature increases it. An increase in the dissolved oxygen content of the water decreases the toxicity of cadmium to freshwater fish. Salmonids appear to be particularly susceptible to the metal. Sublethal effects have been reported, notably malformation of the spine. 6.3.1 Acute and short-term toxicity The acute toxicity of cadmium to fish is summarized in Tables 11 and 12. Pickering & Henderson (1966) calculated the LC50 values for five species of warm-water fish, some of which were tested in both soft and hard water. There was surprisingly little difference in fathead minnows ( Pimephales promelas) between the 24-h, 48-h, and 96-h LC50 values, which in soft water, were 1.09, 1.09, and 1.05 mg/litre, respectively. This species was very much more resistant to cadmium in hard water with 24-h, 48-h, and 96-h LC50 values of 78.1, 72.6, and 72.6 mg/litre, respectively. The hardness for the two types of water was 20 and 360 mg CaCO3 per litre and the alkalinity 80 and 300 mg/litre. The dissolved oxygen concentration was similar in the two types of water. However, the pH was 7.5 for soft water and 8.2 for hard water, this being an area of the pH range where speciation of cadmium undergoes major change. Bluegill sunfish, goldfish, and guppies showed a decrease in LC50 with increase in exposure duration from 24 to 96 h (Table 12), but these were tested only in soft water. The green sunfish Lepomis cyanellus showed LC50 values in soft water of 7.84, 3.68, and 2.84 mg/litre with exposure durations of 24 h, 48 h, and 96 h, respectively. This species was also tested in hard water, where, like the fathead minnow, it showed considerably less cadmium toxicity. The 24-h LC50 rose from 7.84 in soft water to 88.6 mg/litre in hard water. Pickering & Gast (1972) determined a maximum acceptable toxicant concentration (MATC) for the fathead minnow Pimephales promelas of between 37 and 57 µg cadmium/litre. The experimental concentration of 57 µg/litre decreased survival of the developing embryos, this being the most sensitive life-stage, but at lower concentrations (between 4.5 and 37 µg/litre) no adverse effect was found on survival, growth or reproduction. Carroll et al. (1979) investigated the protective effects of various constituents of hard water on the toxicity of cadmium to the brook trout ( Salvelinus fontinalis) and concluded that calcium, added as either the sulfate or carbonate, was the most significant source of protection. This protective effect was observed in the absence of significant cadmium precipitation. Magnesium, sulfate, and sodium ions and the carbonate system provided little or no protection. Calamari et al. (1980) found an influence of water hardness on the toxicity of cadmium to Salmo gairdneri; 48-h LC50 values increased from 91 µg/litre, with a hardness of 20 mg/litre in the test water, to 3700 µg/litre at a water hardness of 320 mg/litre. The 48-h LC50 of fish acclimatized to a hardness of 320 mg/litre and then tested at a hardness of 20 mg/litre was about 7 times higher than that of fish acclimatized and tested in the same soft water. There are two types of biological effects of hardness on the availability of cadmium to fish; one of them persists after acclimatization in hard water. Table 11. Toxicity of cadmium to marine or estuarine fish Organism Size/ Stat/ Temperature Salinity pH Duration LC50 Reference age flow a (°C) (%) (h) (mg/litre) b Striped killifish 0.95 g stat 20 20 8.0 24 125 n Eisler (1971) (Fundulus majalis) 0.95 g stat 20 20 8.0 48 59 n Eisler (1971) 0.95 g stat 20 20 8.0 96 21 n Eisler (1971) Mummichog 0.89 g stat 20 20 8.0 24 > 100 n Eisler (1971) (Fundulus 0.89 g stat 20 20 8.0 48 > 100 n Eisler (1971) heteroclitus) 0.89 g stat 20 20 8.0 96 55 n Eisler (1971) 1.3 g stat 20 20 7.8 24 220 n Eisler & Hennekey (1977) 1.3 g stat 20 20 7.8 96 22 n Eisler & Hennekey (1977) 1.3 g stat 20 20 7.8 168 22 n Eisler & Hennekey (1977) 1 day stat 20 20 48 16.2 (12.7-21.2) n Middaugh & Dean (1977) 7 days stat 20 20 48 9 (6.4-12.5) n Middaugh & Dean (1977) 14 days stat 20 20 48 32 (24.6-41.6) n Middaugh & Dean (1977) adult stat 20 20 48 60 (40-90) n Middaugh & Dean (1977) 1 day stat 20 30 48 23 (19.2-27.6) n Middaugh & Dean (1977) 7 days stat 20 30 48 12 (9.2-15.6) n Middaugh & Dean (1977) 14 days stat 20 30 48 7.8 (5.6-10.3) n Middaugh & Dean (1977) adult stat 20 30 48 43 (33-56) n Middaugh & Dean (1977) Sheepshead minnow 1.1 g stat 20 20 8.0 24 100 n Eisler (1971) (Cyprinodon 1.1 g stat 20 20 8.0 48 50 n Eisler (1971) variegatus) 1.1 g stat 20 20 8.0 96 50 n Eisler (1971) Coho salmon smolt flow 11.2 28.3 7.9 96 1.5 (1.2-2.4) m Dinnel et al. (1989) (Oncorhynchus kisutch) Yellow-eye mullet 1.49 g flow 18.5 34.5 7.8 58 15.5 (12.2-19.8) m Negilski (1976) (Aldrichetta 1.15 g flow 18.6 34.8 7.8 120 14.3 (8.4-24.3) m Negilski (1976) forsteri) Small-mouthed 1.36 g flow 17.9 34.5 7.9 168 12.7 (8.3-19.4) m Negilski (1976) hardyhead 1.12 g flow 18.0 34.5 7.8 168 16.6 (13.3-20.7) m Negilski (1976) (Atherinasoma microstoma) Table 11 (contd). Organism Size/ Stat/ Temperature Salinity pH Duration LC50 Reference age flow a (°C) (%) (h) (mg/litre) b Atlantic silverside adult stat 20 20 48 13 (9-20) n Middaugh & Dean (1977) (Menidia menidia) adult stat 20 30 48 12 (8-16) n Middaugh & Dean (1977) Tidewater silverside larva stat 26 22 96 0.31 (0.25-0.38) n Mayer (1987) (Menidia peninsulae) Shiner perch adult flow 13 30.1 7.8 96 11 (5-20) m Dinnel et al. (1989) (Cymatogaster aggregata) a stat = static conditions (water unchanged for duration of test); flow = flow-through conditions (cadmium concentration in water continuously maintained) b organisms exposed to cadmium added as cadmium chloride; m = measured concentration; n = nominal concentration Table 12. Toxicity of cadmium to freshwater fish Organism Size/ Stat/ Temperature Hardness d pH Duration LC50e Reference age flow a (°C) (mg/litre) (h) (mg/litre) Chinook salmon juvenile flow 11-13 20-22 7.0-7.3 96 0.001 (± 0.0007) f m Finlayson & Verrue (1982) (Onchorhynchus tshawytscha) Rainbow trout juvenile flow 6.4-8.3 96 0.0066 g m Hale (1977) (Salmo 5-15 g stat 8.5-10.7 61-65 7.4 48 2.9 m Pascoe et al. (1986) gairdneri) 5-15 g stat 8.5-10.7 283-317 7.4 48 5.7 m Pascoe et al. (1986) 5-15 g stat 8.5-10.7 61-65 7.4 96 1.3 m Pascoe et al. (1986) 5-15 g stat 8.5-10.7 61-65 7.4 96 2.6 m Pascoe et al. (1986) Fathead minnow adult stat 25 20 7.5 24 1.09 (0.79-2.91) n Pickering & Henderson (1966) (Pimephales adult stat 25 360 8.2 24 78.1 (57.2-117) n Pickering & Henderson (1966) promelas) adult stat 25 20 7.5 48 1.09 (0.79-2.91) n Pickering & Henderson (1966) adult stat 25 360 8.2 48 72.6 (52.7-105) n Pickering & Henderson (1966) adult stat 25 20 7.5 96 1.05 (0.7-4.43) n Pickering & Henderson (1966) adult stat 25 360 8.2 96 72.6 (52.7-105) n Pickering & Henderson (1966) adult stat 18-22 190-210 7.7 48 0.1 (0.07-0.17) n Hall et al. (1986) adult stat 18-22 190-210 7.7 96 0.09 (0.07-0.14) n Hall et al. (1986) Bluegill sunfish adult stat 25 20 7.5 24 4.56 (3.64-6.08) n Pickering & Henderson (1966) (Lepomis adult stat 25 20 7.5 48 2.76 (2.02-3.46) n Pickering & Henderson (1966) macrochirus) adult stat 25 20 7.5 96 1.94 (1.33-2.35) n Pickering & Henderson (1966) Goldfish adult stat 25 20 7.5 24 3.46 (2.85-4.82) n Pickering & Henderson (1966) (Carassius adult stat 25 20 7.5 48 2.62 (2.04-3.68) n Pickering & Henderson (1966) auratus) adult stat 25 20 7.5 96 2.34 (1.81-3.16) n Pickering & Henderson (1966) Table 12 (contd). Organism Size/ Stat/ Temperature Hardness d pH Duration LC50e Reference age flow a (°C) (mg/litre) (h) (mg/litre) Guppy adult stat 25 20 7.5 24 3.37 (2.73-4.81) n Pickering & Henderson (1966) (Poecilia adult stat 25 20 7.5 48 2.31 (1.78-3.11) n Pickering & Henderson (1966) reticulata) adult stat 25 20 7.5 96 1.27 (0.97-1.71) n Pickering & Henderson (1966) 3-4 weeks flow c 23-25 1 mM 24 10.4 m Canton & Slooff (1982) 3-4 weeks flow c 23-25 1 mM 48 5.7 m Canton & Slooff (1982) 3-4 weeks flow c 23-25 1 mM 72 4.3 m Canton & Slooff (1982) 3-4 weeks flow c 23-25 1 mM 96 3.8 m Canton & Slooff (1982) 3-4 weeks flow c 23-25 2 mM 24 33 m Canton & Slooff (1982) 3-4 weeks flow c 23-25 2 mM 48 20.5 m Canton & Slooff (1982) 3-4 weeks flow c 23-25 2 mM 72 14.4 m Canton & Slooff (1982) 3-4 weeks flow c 23-25 2 mM 96 11.1 m Canton & Slooff (1982) Green sunfish adult stat 25 20 7.5 24 7.84 (6.13-14.2) n Pickering & Henderson (1966) (Lepomis adult stat 25 360 8.2 24 88.6 (74-106) n Pickering & Henderson (1966) cyanellus) adult stat 25 20 7.5 48 3.68 (2.89-4.69) n Pickering & Henderson (1966) adult stat 25 360 8.2 48 71.3 (56.3-92.2) n Pickering & Henderson (1966) adult stat 25 20 7.5 96 2.84 (2.1-3.56) n Pickering & Henderson (1966) adult stat 25 360 8.2 96 66 (51.7-84.4) n Pickering & Henderson (1966) Golden shiner flow 72.2 7.5 96 2.8 (1.9-4.3) m Hartwell et al. (1989) (Notemigonus crysoleucas) Puntius 2.4 g stat b 23-27 60-70 7.5 24 59.99 (58.5-61.5) n Shivaraj & Patil (1988) arulius 2.4 g stat b 23-27 60-70 7.5 48 45.7 (43.9-47.5) n Shivaraj & Patil (1988) 2.4 g stat b 23-27 60-70 7.5 72 41.7 (39.7-43.8) n Shivaraj & Patil (1988) 2.4 g stat b 23-27 60-70 7.5 96 39 (36.5-41.7) n Shivaraj & Patil (1988) Table 12 (contd). Organism Size/ Stat/ Temperature Hardness d pH Duration LC50e Reference age flow a (°C) (mg/litre) (h) (mg/litre) Killifish 4-5 weeks stat b 23-25 1 mM 48 > 2.8 m Canton & Slooff (1982) (Oryzias 4-5 weeks stat b 23-25 1 mM 72 0.35 m Canton & Slooff (1982) latipes) 4-5 weeks stat b 23-25 1 mM 96 0.35 m Canton & Slooff (1982) 4-5 weeks stat b 23-25 2 mM 24 > 2.6 m Canton & Slooff (1982) 4-5 weeks stat b 23-25 2 mM 48 1.8 m Canton & Slooff (1982) 4-5 weeks stat b 23-25 2 mM 72 0.17 m Canton & Slooff (1982) 4-5 weeks stat b 23-25 2 mM 96 0.13 m Canton & Slooff (1982) Zebra fish 6 months flow 19-21 1.7 mM 24 7 m Canton & Slooff (1982) (Brachydanio 6 months flow 19-21 1.7 mM 48 4.2 m Canton & Slooff (1982) rerio) a stat = static conditions (water unchanged for duration of test unless stated otherwise); flow = flow-through conditions (cadmium concentration in water continuously maintained unless stated otherwise) b static conditions but water renewed every 24 h c intermittent flow-through conditions d Hardness was expressed as mg CaCO3/litre unless stated otherwise e fish were exposed to cadmium added as the chloride unless stated otherwise; m = measured concentration; n = nominal concentration f cadmium was added as the sulfate g cadmium was added as the nitrate In a large scale study of the toxicity of cadmium to the mummichog Fundulus heteroclitus, Voyer (1975) examined effects of salinity, pre-exposure to high salinity and different concentrations of dissolved oxygen on the tolerance of fish to cadmium over 96 h. He found no significant influence of dissolved oxygen levels between 4.0 mg/litre and saturation regardless of salinity during acclimatization or during the test. By contrast, Voyer et al. (1975) showed a distinct effect of dissolved oxygen concentration on toxicity of cadmium to the same species of fish in fresh water. Median tolerance concentrations at 96 h ranged upwards from 1.3 to 3.0 mg cadmium/litre with 2.3 and 8.5 mg/litre of dissolved oxygen, respectively. They demonstrated statistically an independent effect of dissolved oxygen and time against cadmium toxicity. It should be noted that cadmium is 10 times more toxic to this species in fresh water than in sea water. Toxicity of cadmium to both marine (Eisler, 1971) and freshwater (Roch & Maly, 1979) fish has been shown to be greater at higher temperatures. Canton & Slooff (1982) exposed several fish species to cadmium in short-term toxicity tests. At a water hardness of 1.7 mmol/litre they found the no-observed-adverse-effect level (NOAEL) for mortality in the zebra fish ( Brachydanio rerio) to be 2 mg/litre over a 48-h exposure period. For the killifish ( Oryzias latipes), the 96-h NOAEL was 0.06 mg/litre for mortality and 0.03 mg/litre for mortality and abnormal behaviour at a water hardness of 2 mmol/litre. The corresponding values at a water hardness of 1 mmol/litre were 0.055 and 0.006 mg/litre. For the guppy ( Poecilia reticulata), the 96-h NOAEL for mortality and abnormal behaviour was 5.2 mg/litre at a water hardness of 2 mmol/litre and 0.6 mg/litre at 1 mmol/litre. The authors calculated a 24-h NOAEL for the rainbow trout ( Salmo gairdneri) of 0.01 mg/litre for inhibition of opercular movements. Abel & Papoutsoglou (1986) studied the toxicity of cadmium to Cyprinus carpio and Tilapia aurea and reviewed data on other species of freshwater fish. The found for all species examined that the median survival time changed little over a wide range of cadmium concentrations and that a toxic threshold was clear in most studies. For Tilapia this threshold lay between 0.1 and 0.5 mg cadmium/litre; negligible mortality was recorded in fish exposed to 0.1 mg/litre for 3 months. 6.3.2 Reproductive effects and effects on early life-stages Meteyer et al. (1988) exposed sheepshead minnow ( Cyprinodon variegatus) eggs to cadmium concentrations of between 0.39 and 1020 µg/litre from approximately 4 h after fertilization. Hatching was delayed by up to 3 days at the highest cadmium concentration. All treated larvae were shorter than controls, but there was no dose-related effect of cadmium on growth. Middaugh & Dean (1977) examined the toxicity of cadmium to various life-stages of the mummichog Fundulus heteroclitus. In 48-h tests, eggs were highly resistant to cadmium. The greatest effect (54% non-emergence) occurred at a cadmium concentration of 32 mg/litre; the control non-emergence was 17%. Newly-emerged larvae were less sensitive to cadmium than were 7-day-old larvae. There was an effect of salinity on the sensitivity of 14-day-old larvae to the metal (Table 11). Adults were less sensitive to cadmium than were larvae. Similar results were obtained with the Atlantic silverside ( Menidia menidia), larvae being the most sensitive life-stage. Weis & Weis (1977) found no effect of cadmium, at concentrations up to 10 mg/litre, on embryos of Fundulus heteroclitus. Rombough & Garside (1982) found the most sensitive indicator of cadmium toxicity to early life-stages of the Atlantic salmon ( Salmo salar) to be inhibition of growth of alevins, where significant reductions occurred with cadmium concentrations of 0.47 µg/litre. The LC50 for the interval between fertilization and viable hatch lay between 300 and 800 µg per litre. Newly hatched alevins showed a 24-h LC50 of between 1.5 and 2.4 mg/litre. Sensitivity increased sharply in late alevins and significant mortality was recorded at a concentration of 8.2 µg/litre. Eaton et al. (1978) exposed embryos and larvae of seven freshwater fish species to nominal cadmium concentrations of 0, 0.4, 1.2, 3.7, 11.6, 33.3, and 100 mg/litre for periods ranging from 3 to 126 days. Actual concentrations were monitored and used in the assessment of the results. Results were expressed in terms of "standing crop", which the authors estimated as the product of the proportion of fish surviving and their total biomass. The lowest concentration of cadmium at which the standing crop was significantly different from controls was approximately 12 µg/litre for the white sucker, northern pike, smallmouth bass, west coast coho salmon, lake trout, and brown trout exposed as embryos or larvae for up to 64 days. The value was lower for brook trout (0.48 µg per litre) after exposure of larvae/juveniles for 65 days and for Lake Superior coho salmon (3.4 µg/litre) after exposure of larvae/juveniles for 27 days. The highest cadmium concentration at which standing crop was not significantly different from controls varied for the seven species between 1.1 and 4.2 µg/litre. Woodall et al. (1988) exposed rainbow trout ( Salmo gairdneri) fry to cadmium concentrations of 0, 0.1, 1.0 or 5.0 mg/litre for up to 90 h. In preliminary experiments, they calculated that the 90-h LC50 lay between 0.1 and 1.0 mg/litre. Pre-treatment of trout fry with cadmium initially had little effect (up to 30 h). However, with exposure periods of between 45 and 90 h, some protection was induced by pre-treatment. Benoit et al. (1976) exposed three generations of brook trout ( Salvelinus fontinalis) to concentrations of total cadmium varying between 0.06 and 6.4 µg/litre. Significant numbers of first and second generation adult males died during spawning when exposed to 3.4 µg/litre. This concentration also significantly retarded the growth of juvenile second and third generation offspring, but at a concentration of 1.7 µg/litre these effects were not seen. In a study by Borgmann & Ralph (1986), white sucker larvae Catostomus commersoni and young common shiners Notropis cornutus were exposed to cadmium chloride at concentrations ranging from 6.24 to 200 µg cadmium/litre. The relative growth rate of fish was significantly reduced at concentrations of 36 µg/litre or more in the case of suckers and 63 µg/litre or more in the case of shiners. Cadmium had no effect on the relative feeding rates. Hatakeyama & Yasuno (1987) studied the chronic effects of cadmium on the reproduction of the guppy Poecilia reticulata. The fish were exposed to cadmium via cadmium-accumulated midge larvae used as their food source. The cumulative numbers of fry produced by guppies fed midge larvae containing 500, 800 or 1300 mg cadmium/kg for 6 months decreased to 79%, 65%, and 55% of the controls, respectively. At the highest dose, the mortality of females was significantly elevated at 6 months, but no such effect was observed with the males. Michibata (1981) reported a protective effect of water hard-ness against the effects of cadmium on the eggs of Oryzias latipes. 6.3.3 Metabolic, biochemical and physiological effects Protective metal-binding proteins (metallothioneins) are induced by cadmium in fish (chapter 4). A manifest symptom of cadmium toxicity in freshwater fish is ionic imbalance with reduced plasma Ca2+, Na+, and CL-. The probable explanation is that cadmium is a potent inhibitor of ion-transporting enzymes. Verbost et al. (1988) showed that cadmium inhibited Ca-ATPase in the cell membranes of fish gut. It probably does the same in the gills because cadmium exposure has been shown to inhibit calcium uptake in the gills of adults (Verbost et al., 1987; Reid & McDonald, 1988) as well as in larvae (Wright et al., 1985). Similarly, cadmium has been shown to inhibit Na/K-ATPase in fish gills (Watson & Benson 1987), which, taken with the fact that cadmium probably also affects the production of ATP in the gills (Dickson et al., 1982), could explain the reduction of plasma Na+. Huiguang Fu (1989) studied the role of the hormones prolactin and cortisol in correcting cadmium-induced impairment of the calcium balance in tilapia ( Oreochromis mossambicus). Fish were able to recover from initial hypocalcaemia during a 35-day exposure to a concentration of 10 µg cadmium/litre. This recovery involves prolactin-induced stimulation of active Ca2+ uptake and reduction of passive Ca2+ efflux, together with cortisol-induced changes in chloride cells and stimulation of metallothionein synthesis in the liver, kidney, and gills. The author stated that the capacity to survive prolonged exposure to 10 µg cadmium/litre through physiological adaptation does not indicate that this cadmium concentration is acceptable to tilapia. Adaptation may be achieved at the expense of other essential processes like growth and reproduction. Arillo et al. (1984) investigated the effect of cadmium at levels of 1-10 µg/litre (concentrations above the water quality criteria value of 0.75 µg/litre proposed for salmonids by EIFAC/FAO) on a wide variety of biochemical parameters in the rainbow trout (Salmo gairdneri). The exposure period was 4 months. Only at the highest test concentration of 10 µg/litre were there any effects on the fish, and then only on liver aminolevulinic acid dehydratase activity. The authors concluded that the water quality criterion was realistic. Dawson et al. (1977) exposed juvenile striped bass ( Morone saxatilis) to cadmium chloride concentrations of 0.5, 2.5 or 5 µg/litre for 30 to 90 days, and the fish were then allowed to recover for 30 days in clean, running sea water. There was an inhibition of gill tissue respiration at 30 and 90 days, which recovered during the 30-day period with clean water. The activities of the various enzymes measured were not affected. MacInnes et al. (1977) reported reduced gill tissue oxygen consumption in cunner ( Tautogolabrus adspersus) exposed to cadmium (0.05 or 0.1 mg/litre) for 30 or 60 days. They also reported a reduced activity of aspartate aminotransferase and an increased activity of glucose-6-phosphate dehydrogenase in the liver of the fish after 30 days of exposure to cadmium. Gill & Pant (1983) measured levels of various blood and tissue constituents after acute (24 h) or chronic (90 day) exposure to cadmium. Acute exposure to 12.65 mg/litre led to significant hyperglycaemia and an increase in liver, kidney, and ovarian cholesterol levels. Chronic exposure to 0.63 or 0.84 mg/litre, by contrast, led to an enduring hypoglycaemia and diminished levels of cholesterol in tissues. Both acute and chronic exposure to cadmium caused marked hypocholesterolaemia, glycogenolysis in the liver and brain and a rise in myocardial glycogen. Testis cholesterol was depleted after 60 days in both acute and chronic exposures. Sastry & Subhadra (1983) exposed the catfish Heteropneustes fossilis to cadmium in the water at the sublethal concentration of 2.3 µmol/litre for 15 or 30 days. The cadmium caused reduced absorption of glucose and fructose from the gut, this effect being more pronounced after 30 days of exposure than after 15 days. Filling intestinal sacs, in vivo, with cadmium solutions (1 µmol per litre) reduced absorption of the sugars significantly over 1 h. Merlini (1978) pre-treated immature sunfish ( Lepomis gibbosus) with cadmium (0.004 mg/litre) and then fed treated and control fish with a single ration containing 58Co-labelled vitamin B12. The fish were subsequently fed non-radioactive food for 31 days before sacrifice. The author reported that cadmium-treated fish stored significantly less vitamin B12 in the liver than did controls. Carrier & Beitinger (1988a) studied the effect of cadmium on the critical thermal maximum (the temperature at which loss of equilibrium is coupled with loss of righting response) in the red shiner ( Notropis lutrensis) and fathead minnow ( Pimephales promelas). The red shiner was exposed to sublethal cadmium concentrations of 4.88, 5.07, and 5.46 mg/litre and the fathead minnow to 0.09, 0.48, and 1.26 mg/litre. Critical thermal maxima were significantly reduced for both species over a 10-day exposure period. The effect was found to be dose and time dependant. In the green sunfish ( Lepomis cyanellus) there was no effect of cadmium concentrations of 2.76, 4.22 or 5.17 mg/litre on the critical thermal maximum (Carrier & Beitinger, 1988b). 6.3.4 Structural effects and malformations When Bengtsson et al. (1975) exposed 180 minnows ( Phoxinus phoxinus) to various concentrations of cadmium ranging from 7.5 µg/litre to 4.8 mg/litre for 70 days, 31 of the 101 fish that survived developed lesions in the spinal column. Fractured vertebrae occurred in the caudal end of the abdominal section of the spine or in the caudal section; 64% of all fractures occurred in the first 7 caudal vertebrae and 21% occurred in abdominal vertebrae numbers 8 to 14. Hiraoka & Okuda (1984) cultured medaka ( Oryzias latipes) eggs in a cadmium solution of 0.01 mg/litre for 1 week and then investigated abnormalities in the vertebrae of the hatched fry, which were reared in clean water. There was no damage to the centra of the vertebrae in newly hatched fry. However, centrum-damaged fish were found in the first week and the numbers increased rapidly up to the 4th week after hatch. The cumulative frequency of vertebral-damaged fish was 13% and 14% in the 5th and 6th weeks, respectively, and seemed to remain constant after this. Muramoto (1981b) reported that fish showing malformations of the spine after cadmium treatment had significantly less calcium in the vertebral column than did control fish. Voyer et al. (1975) found no short-term histopathological effects of cadmium, but long-term exposure to 28 mg/litre caused necrosis and sloughing of the mucosa of gills. Tissue damage was also evident in the nasal passages and buccal cavity. This histopathological effect was seen between exposure times of 512 and 612 h, but not before. 6.3.5 Behavioural effects Hartwell et al. (1989) studied the avoidance response of the golden shiner ( Notemigonus crysoleucas) to cadmium concentrations of up to 68 µg/litre, but found no significant avoidance. Sullivan et al. (1978b) subjected fathead minnows ( Pimephales promelas) to acute (24 h) or subacute (21 days) exposure at sublethal cadmium concentrations and then placed them in experimental chambers with largemouth bass ( Micropterus salmoides), a fish which preys upon them. The minnows displayed altered behaviour patterns, including abnormal schooling behaviour, and were more vulnerable to predation than controls. The lowest acute cadmium exposure level that increased vulnerability was 0.375 mg/litre and the lowest subacute level 0.025 mg/litre. The authors pointed out that the subacute value of 0.025 mg/litre was well below reported no-effect levels of cadmium for fathead minnows established with respect to survival and reproductive effects. The avoidance threshold for cadmium in water by the rainbow trout ( Salmo gairdneri) is 50 µg/litre (Black & Birge, 1980), which is 50 times higher than the 96-h LC50 value reported for this species. 6.3.6 Interactions with other chemicals Muramoto (1981a) showed that the chelating agents EDTA and DPTA afforded some protection against the effects of cadmium on carp ( Cyprinus carpio) exposed to 0.05 or 0.1 mg cadmium/litre for 3 months. There were effects on the vertebrae of exposed fish. Chelating agents that form hydrophobic complexes with heavy metals increase the bioavailability of the metal to aquatic organisms. Examples of these are xanthates and dithiocarbamates. Xanthates are used in the mining industry in the flotation process to refine metal from sulfide ores. As discussed in section 4.1, these compounds increase the uptake rate of cadmium through the gills in fish (Block & Part, 1986; Gottofrey et al., 1988; Block, 1991). Furthermore, they change the tissue distribution of the metal in such a way that more cadmium is found in lipid-rich tissues such as nervous tissue (brain) and adipose tissue than when fish are exposed to the metal alone. Finlayson & Verrue (1982) reported 96-h LC50 values for juvenile chinook salmon ( Onchorynchus tshawytscha) ranging from 0.6 to 1.6 µg/litre. They found no synergistic or antagonistic toxic effects after combining cadmium and zinc in their test system. The results of tests were additive, the overall LC50 being a simple combination of individual metal effects at a zinc:cadmium ratio of 1:0.008. Spehar et al. (1978b) exposed the flagfish Jordanella floridae to cadmium and zinc, both individually and as a mixture, at concentrations ranging from 4.3 to 8.5 µg cadmium/litre and 73.4 to 139 µg zinc/litre, through one complete life-cycle of the fish. There was no additive effect of cadmium and zinc at sublethal concentrations in mixed exposure. Effects on survival showed that the toxicity of cadmium and zinc mixtures was only slightly, if at all, greater than the toxicity of zinc alone. Anadu et al. (1989) reported that pre-exposure of rainbow trout ( Salmo gairdneri) to zinc (100 µg/litre) for 17 days increased the subsequent 120-h LC50 for cadmium from 1.1 µg/litre to 4.1 µg/litre. 6.4 Toxicity to Amphibia Francis et al. (1984) exposed eggs of the leopard frog ( Rana pipiens) to cadmium-enriched sediments during their development and for 4 days after the larvae had hatched. Measured concentrations of cadmium ranged from 1.04 to 1074 mg/kg in sediment and from 1.0 to 76.5 mg/litre in water above the sediment. Cadmium concentrations in the tissues of the tadpoles at the end of the experiment ranged from 0.08 to 12.55 mg/kg. There was no mortality as a result of cadmium exposure. The LC50 for this species has been reported to be 50 µg/litre for water without sediment by Westerman (1977). Slooff & Baerselman (1980) determined 48-h LC50 values for the neotenous larval mexican axolotl ( Ambystoma mexicanum) and larval South African clawed toad ( Xenopus laevis) of 1.3 mg/litre and 32 mg/litre, respectively, after exposure to cadmium nitrate. Canton & Slooff (1982) exposed Xenopus laevis to cadmium, added as cadmium chloride, and obtained 24-h and 48-h LC50 values of 4 and 3.2 mg/litre, respectively. A NOEL of 2.2 mg/litre was reported for both exposure periods. In a longer-term exposure (100 days), by the same authors, an LC50 value of 1500 µg/litre was found and the EC50 value for inhibition of larval development was 650 µg/litre. For mortality and larval development, the NOELs were 30 and 9 µg/litre, respectively. De Zwart & Sloof (1987) determined a 48-h LC50 of 20.2 mg/litre for tadpoles of the clawed toad. Khangarot & Ray (1987) established a 96-h LC50 value of 8.18 (6.96-9.53) mg/litre for the tadpoles of the toad Bufo melanosticus. The test water was obtained from a well and had a hardness of 185 mg per litre, pH 7.4, and temperature of 31 °C. Solids were present at a level of 920 mg/litre. Muino et al. (1990) calculated the 48-h and 96-h LC50 for Bufo arenarum tadpoles to be 2.52 and 2.08 mg cadmium/litre for the two exposure periods, respectively, under semi-static conditions. In tests to study the effect of a sublethal cadmium concentration (1.0 mg cadmium/litre) on the water balance of the animals, the authors found that all animals died within a few hours in ion-free media. Tadpoles exposed in ionic solutions showed mortality of less than 10% (equivalent to control groups). Woodall et al. (1988) exposed Xenopus laevis tadpoles to cadmium concentrations of 0, 50, 80, and 100 mg cadmium/litre for up to 90 h. In preliminary experiments, they calculated the 90-h LC50 to lie between 80 and 100 mg/litre. The authors found that pre-treatment of tadpoles with cadmium induced protection, which decreased with an increase in the subsequent exposure concentration. Cadmium pre-treatment induced maximum protection to cadmium at a concentration of 50 mg/litre at both 45 and 90 h. Perez-Coll et al. (1986) exposed developing Bufo arenarum embryos to cadmium chloride concentrations of 6 x 10-7 to 1.5 x 10-5 mol Cd2+/litre during gastrulation at 20 °C and 30 °C. Initial failures at gastrulation resulted mainly in axial incurvations, microcephaly, hydropsy, and abnormal tail development. At the higher temperature, high concentrations of cadmium caused a significant increase in early malformations and at low concentrations the high temperature prevented alterations. 7. TOXICITY TO TERRESTRIAL ORGANISMS Appraisal Both terrestrial plants and animals accumulate cadmium, but the rate of accumulation is much higher under experimental conditions, where cadmium is available in solution, than it is with plants grown in soil, when part of the cadmium is bound and less available. Cadmium has adverse effects on hydroponically grown plants at concentrations in the mg/litre range, whereas plants grown in soil only show reduced growth in contaminated soils with hundreds of mg cadmium/kg. Terrestrial invertebrates are relatively insensitive to cadmium-induced toxic effects, probably due to effective sequestration mechanisms in specific organs. When toxic effects do occur, they consist of reduced growth and reproduction. 7.1 Toxicity to terrestrial plants Cadmium has been shown to have an adverse effect on plant growth and yield in laboratory experiments. However, plants grown in soil are generally insensitive to the effects of cadmium except at high doses. Effects are only seen when cadmium is given in nutrient solutions rather than in soil, where the cadmium is bound and is therefore less available to the plants. Cadmium is only available to plants in solution in soil. There is considerable evidence from field studies that plants are able to develop tolerance to various heavy metals in their growth medium. Research into cadmium tolerance has been more limited than for other metals, but there is some evidence of tolerance developing. 7.1.1 Toxicity to plants grown hydroponically Mitchell & Fretz (1977) cultured seedlings of three species of tree, the white pine ( Pinus strobus), red maple ( Acer rubrum), and Norway spruce ( Picea abies), in sand. Plants were irrigated with a nutrient solution containing cadmium at concentrations of 0, 0.5, 1, 2, 4, 8, and 16 µg/litre; white pine seedlings were also treated with 32 and 64 µg/litre. Both roots and foliage were affected by the cadmium. In the red maple symptoms of cadmium toxicity began with interveinal chlorosis and stunting of leaves in most cases. As exposure increased cadmium caused wilting and then death. The first observed effect of the metal on the pine was inhibition of needle expansion and, in the case of the spruce, chlorotic tips to new growth. Tissue accumulation of cadmium correlated well with exposure. Red maple, white pine, and spruce exhibited foliage effects at leaf cadmium concentrations of 22.8, 61.3, and 7.5 mg/kg, respectively, corresponding to nutrient solutions of 8.0, 32.0, and 4.0 µg/litre. There was an increasing effect on root development with increasing exposure to cadmium. At the higher doses, there was severe reduction in the number of roots initiated and stunting of those that grew. Accumulation of cadmium was greater in roots than in leaves. Root et al. (1975) grew maize ( Zea mays) in hydroponic solutions containing cadmium chloride at concentrations ranging from 1 to 40 mg/litre. Uptake of cadmium into the plants increased with time, and cadmium was present at higher concentrations in roots than in shoots. Leaf chlorophyll concentration and yield (as dry weight) of both roots and shoots decreased with increasing cadmium concentration. As the cadmium concentration in the leaves increased, the concentration of zinc decreased and the concentration of iron increased. This gave a linear correlation between cadmium in the leaf and iron/zinc ratio. Chlorosis resulting from cadmium in the leaves (seen at a cadmium concentration of 1 mg/litre nutrient solution) appeared comparable to iron-deficiency chlorosis. However, in this case, the chlorosis was not due to iron deficiency, as previously suggested by other workers, but was associated in some way with an increasing iron/zinc ratio. Harkov et al. (1979) found no effect on the yield of tomatoes grown in vermiculite and cultured with a nutrient solution containing cadmium at concentrations of 0.25 or 0.75 mg/litre. The plants were more susceptible to damage by ozone, under conditions where ozone damage would have been slight, after exposure to cadmium. Where ozone damage was heavy or when conditions were not conducive to ozone damage, there was no effect of cadmium. When Wong et al. (1988) exposed pea ( Pisum sativum) seeds to cadmium concentrations of 1, 5, 10, and 20 mg/litre in culture solutions, germination was significantly reduced at 20 mg/litre and radicle elongation to 1 cm was significantly reduced at 5 mg/litre and to 2 cm at 1 mg/litre. Early development (shoot elongation, leaf development, root and shoot growth) was inhibited in a dose-dependant manner. Slight inhibition was observed at 1 mg/litre, increased inhibition at 5 mg/litre, and significant inhibition at 10 and 20 mg/litre. 7.1.2 Toxicity to plants grown in soil Mitchell & Fretz (1977) showed that the effects on the growth of red maple, white pine and Norway spruce plants in soil, amended with cadmium, were similar but less severe, owing to reduced uptake of the metal, than in the case of the same plants grown hydroponically. Cadmium only affected current growth of the plants, except where it was present in excess. Mahler et al. (1978) treated eight soils, the pH values of which ranged from 4.8 to 7.8, with 1% (by weight) sewage sludge containing added cadmium sulfate, leading to cadmium concentrations in the soil ranging from 0.1 to 320 mg/kg. Two plants, lettuce ( Lactuca sativa variety longifolia) and Swiss chard ( Beta vulgaris variety cicla), were grown in the soils in pots. The EC50 (yield) for lettuce was 214 and 139 mg/kg soil for acid and calcareous soils, respectively, whereas the values for chard were 175 and 250 mg/kg for acid and calareous soils. The corresponding tissue concentrations of cadmium associated with these effects were 470 and 160 mg/kg for lettuce and 714 and 203 mg/kg for chard. Thus, a markedly lower tissue concentration of cadmium produced 50% yield reduction on calcareous soils than on acid soils. Alloway et al. (1990) reported stunted growth and toxic signs on leaves of lettuce, cabbage, carrot, and radish plants, but only at the highest concentrations of cadmium tested (which resulted in a cadmium content of around 20 mg/kg in the upper parts of the plants). 7.1.3 In vitro physiological studies Bazzaz et al. (1974) demonstrated an effect of cadmium on transpiration and photosynthesis in excised sunflower heads. The heads (15 cm diameter) were placed in flasks containing distilled water or cadmium salt solutions (2, 20, 100 or 200 mg/litre) and transpiration and photosynthesis were measured at daily intervals over 4 to 5 days. Cadmium reduced transpiration and photosynthesis at concentrations of 100 or 200 mg/litre. Excised epidermal peels floating on solutions of cadmium salts showed a log-linear relationship between metal concentration and stomatal opening. The stomata opened less with increasing cadmium concentration; this accounted for the effect on transpiration and, hence, on photosynthesis. 7.2 Toxicity to terrestrial invertebrates Haight et al. (1982) calculated 24-h, 48-h, and 72-h LC50 values of 36, 15.1, and 5.85 mg cadmium/litre, respectively, for juvenile free-living nematodes ( Panagrellus silusiae) and 111, 26.3, and 13.2 mg/litre for adults. Williams & Dusenbery (1990) exposed the free-living nematode Caenorhabditis elegans to cadmium and found values of 904, 22, and 1.5 mg/litre, respectively, for the same three exposure periods. They also calculated a 96-h LC50 of 0.06 mg/litre. Popham & Webster (1979) found that a 6-h exposure to 3.26 x 10-7 moles of cadmium significantly decreased the fecundity of C. elegans. A 3.5-day exposure to 10-8 moles caused the same effect. Nematodes exposed to 4 x 10-6 moles of cadmium never grew to the same length as controls and resembled worms from starved, overcrowded cultures. Van Kessel et al. (1989) exposed juvenile C. elegans to various concentrations of cadmium chloride and found that growth and subsequent reproduction were significantly reduced at 1 µmol/litre. At levels of 160 and 320 µmol/litre the nematodes did not reach the adult stage and, therefore, did not reproduce. Doelman et al. (1984) exposed the soil nematodes Mesorhabditus monhystera and Aphelenchus avenae to cadmium via food for up to 22 days and monitored the size of the population. M. monhystera was exposed, via bacteria and fungi, to concentrations of 0.23, 4.4, and 12.7 mg cadmium/kg, and a significant reduction in the size of the population was found at all doses. A. avenae was exposed, via fungi alone, to concentrations of 1, 10, and 25 mg/kg. At 1 mg/kg there was no effect on the population size, whereas at 10 mg/kg, a reduction was observed until the final day. A concentration of 25 mg/kg significantly reduced the size of the population. The authors noted that exposure via fungi alone gave far more variable results. In studies by Van Straalen et al. (1989), the collembolan Orchesella cincta and the oribatid mite Platynothrus peltifer were exposed to cadmium in the food. The 9-week LC50 values were 1.6 and 7.27 µmol/g and the no-observed-effect levels (NOEL) were 0.042 and 0.026 µmol/g for O. cincta and P. peltifer, respectively. The most sensitive parameters were female growth for O. cincta and reproduction in P. peltifer. Russell et al. (1981) fed subadult garden snails ( Helix aspersa) on diets containing six different levels of cadmium ranging from 10 to 1000 mg/kg diet over a period of 30 days. There was little mortality (two animals out of 350 died, one at an exposure level of 50 mg/kg and the other at 1000 mg/kg) but food consumption declined with each increase in cadmium dose. Food consumption was strongly depressed at cadmium doses of 100 mg/kg or more. Relative weight loss was the most pronounced effect of cadmium treatment; this was dose related and directly attributable to reduced feeding rates. At doses of 25 mg/kg or more, shell growth and reproductive activity were depressed while the incidence of sealing response (the sealing of the operculum with a disc of mucus and a dormancy reaction) increased markedly. These three effects were all related to the dietary cadmium concentration. 7.3 Toxicity to birds 7.3.1 Acute and short-term toxicity The acute and short-term toxicity of cadmium salts to birds in laboratory studies is summarized in Table 13. Dosing for 5 days, followed by 3 days of clean diet, resulted in LC50 values generally in excess of 2000 mg/kg diet. Only the pheasant showed greater sensitivity to cadmium but, even in this species, the LC50 was close to 1000 mg/kg diet. All the birds used were between 10 and 14 days old. Table 13. Toxicity of cadmium to birds a Species Age Salt LC50 Reference (days) (mg/kg diet) Japanese quail 14 cadmium chloride 2440 (1807-3294) Hill & (Coturnix coturnix 14 cadmium succinate 2052 (1621-2598) Camardese japonica) (1986) Pheasant 10 cadmium chloride 767 (651-898) Hill et al. (Phasianus 14 cadmium succinate 1411 (1202-1657) (1975) colchicus) Bobwhite quail 14 cadmium succinate 1728 (1381-2132) Hill et al. (Colinus (1975) virginianus) Table 13 (contd). Species Age Salt LC50 Reference (days) (mg/kg diet) Mallard duck 10 cadmium chloride > 5000 Hill et al. (Anas 10 cadmium succinate > 5000 (1975) platyrhynchos) a Birds were fed with a dosed diet for 5 days and then a "clean" diet for 3 days. In a study by Pritzl et al. (1974), 2-week-old leghorn chicken chicks were dosed with dietary cadmium chloride for 20 days. In the first experiment, chicks dosed with 700 mg cadmium/kg diet showed an increase in the weight of the gastrointestinal tract, kidney, and gizzard expressed as a ratio to the body weight. In a second experiment, chicks were fed diets containing 400, 600, 800 or 1000 mg cadmium/kg. Weight gain and food consumption were decreased, relative to controls, at all dose levels, and at levels higher than 400 mg/kg the birds lost weight. All the birds fed diets containing 800 or 1000 mg/kg died within 20 days. The LC50 was calculated to be 565 mg/kg diet. When Cain et al. (1983) fed 1-day-old mallard ducklings ( Anas platyrhynchos) a diet containing cadmium chloride at concentrations of 5, 10 or 20 mg cadmium/kg for 12 weeks, significant effects were only noted at the highest dose. These included a significant reduction in packed cell volume and haemoglobin concentrations and a significant increase in serum glutamic-pyruvic transaminase. Mild to severe kidney lesions were evident in ducklings fed 20 mg/kg for 12 weeks. Body weight, liver weight, and femur weight-to-length ratio were unaffected by the cadmium treatment. No other haematological or histological effects were found. 7.3.2 Reproductive effects Lofts & Murton (1967) injected 0.2 ml of a solution of cadmium chloride (0.04 mol/litre) intramuscularly into wood pigeons ( Columba palumbus). When the cadmium was given to birds with regressed testes, which were then stimulated into reproductive condition by long photoperiods (16 h of light per day), there was a reduction in the numbers of birds showing full testicular development in the treated group. Only one out of six birds given cadmium had developed spermatozoa in the testis by the end of the experiment. The remainder had not even produced spermatids; two birds had only spermatogonia in the seminiferous epithelium while the other three had secondary spermatocytes. Of the six control birds, four had spermatozoa, one had spermatids, and one primary spermatocytes. Injection of cadmium into birds late in the season had no effect on the autumnal regression of the testes. There was no sign of testicular necrosis in the treated birds. An intratesticular injection of cadmium caused local necrosis in the testis of feral pigeons. A dietary concentration of 200 mg cadmium/kg, in the form of cadmium chloride, reduced spermatogenesis in male mallards and egg production in females, but a lower dose of 20 mg/kg produced no effects (White & Finley, 1978; White et al., 1978). 7.3.3 Physiological effects Mayack et al. (1981) found that the survival and growth of the wood duck ( Aix sponsa) were unaffected by a cadmium chloride dietary level of 100 mg/kg, although some kidney damage was reported. Nicholson et al. (1983) compared the ultrastructure of the kidneys of sea-birds contaminated with cadmium in the wild, sea-birds from uncontaminated colonies, starlings dosed with cadmium in the laboratory, and control starlings. They found damage to kidney cells to be comparable between wild sea-birds and dosed starlings having kidney cadmium levels of 60-480 and 95-240 mg per kg, respectively. Damage was greatest in the proximal tubule of the kidney, and included cell necrosis, nuclear pyknosis, mitochondrial swelling, and some tubulorrhexis. The tubulorrhexis would be irreversible. There was some indication of regeneration, judging from the number of undifferentiated cells present in the tubule, in both dosed and naturally contaminated birds. Debris was found in the distal nephron lumen and there was some damage in the distal tubule and the renal corpuscles. Necrosis of kidney cells was very rare in control birds or uncontaminated sea-birds. 7.3.4 Behavioural effects Heinz et al. (1983) assessed the avoidance response to a visual fright stimulus of ducklings fed a diet containing cadmium chloride at a concentration of 4 or 40 mg/kg. The parents of the ducklings had also been fed this cadmium-containing diet. Ducklings fed 4 mg/kg were hypersensitive to the fright stimulus, whereas those fed the higher dose reacted as did controls. The authors could offer no explanation of why the higher dose had no effect but pointed to similar results from other materials. They were of the opinion that hypersensitivity to behavioural signals could be as deleterious to the organism in the wild as a failure to respond. 7.4 Toxicity to wild small mammals Shore et al. (1991) fed herbivorous bank voles ( Clethrionomys glareolus) and granivorous wood mice ( Apodemus sylvaticus) a pelleted diet contaminated with cadmium chloride and collected urine and faeces in metabolism cages. Bank voles fed diets containing 10.3 mg/kg for 40 days and then 4.5 mg/kg for 35 days suffered significant net daily loss of calcium and sodium, and reduced net gain of potassium and magnesium compared to controls. Assimilation of the macroelements was not significantly altered in wood mice fed 10.3 mg/kg for 75 days. 8. EFFECTS IN THE FIELD Appraisal Tolerance to cadmium has been demonstrated in soil fungi, plants, aquatic invertebrates, and fish from cadmium-contaminated sites. Some field evidence suggests that cadmium is responsible for reduced leaf litter degradation and a failure to recycle nutrients due to adverse effects on populations, particularly of microorganisms. However, no studies have identified cadmium as the sole cause of the effect, since it is always associated with other metals. Although soil invertebrates in contaminated sites accumulate cadmium and other metals, there is evidence that most populations are not affected. A field study has shown that fish from a cadmium-contaminated river have physiological abnormalities. Kidney damage has been found in pelagic sea-birds from areas away from industrial or other anthropogenic sources of cadmium, but there was no effect on survival or reproduction of populations. In industrially contaminated areas, kidney damage has been observed in several species of birds found to contain cadmium plus other metals. 8.1 Tolerance Tolerance to cadmium has been demonstrated in soil fungi (section 5.2), aquatic invertebrates (section 6.2.7), and plants collected from sites with high cadmium levels, such as those in the vicinity of metalliferous mines and smelters. Coughtrey & Martin (1977) experimentally demonstrated tolerance of the grass Holcus lanatus to cadmium to be greater in plants collected from an area subject to high fall-out of cadmium than in plants collected from a control site. Growth of tolerant plants was reduced in uncontaminated culture solutions, relative to controls, but was similar to that of control plants in solutions where cadmium salts had been added at levels similar to the field exposure. Simon (1977) reported cadmium tolerance in the grasses Festuca ovina and Agrostis tennuis. The tolerant grasses were collected from areas contaminated by mining and aerial fall-out of cadmium. 8.2 Effects close to industrial sources and highways There have been several reports of effects of heavy metal deposition on the accumulation rate of leaf litter in deciduous woodlands. These effects are restricted to areas close to, or down wind from, smelter sites. The separation of the effects of cadmium from those of other heavy metals present in the litter is difficult. A study by Coughtrey et al. (1979) attempted to do this by detailed analysis of the litter itself and by the use of statistics to separate effects of different components of the pollution fall-out. Seven areas of woodland in the vicinity of, or up to 28 km away from, a lead-zinc-cadmium smelter in Avonmouth, United Kingdom, were studied. The leaf litter contained lead, zinc, copper, and cadmium (in that order of concentrations), and metal levels were high in samples taken from within 3 km of the smelter. Litter from a wood 6.8 km from the smelter had similar levels to control litter collected 30 km away, but the prevailing wind would not have carried much of the fall-out in the direction of this wood. For the four woods within 3 km of the smelter, cadmium levels ranged from 23 to 98 mg/kg litter (lead levels were between 721 and 2179 mg/kg, zinc levels between 764 and 2814 mg/kg, and copper levels between 47 and 135 mg/kg). The weight of litter accumulated per unit area was markedly greater in the contami-nated sites than in the uncontaminated ones; litter standing crop ranged from 7.91 to 13.16 kg/m2 in contaminated and from 0.913 to 3.104 kg/m2 in uncontaminated sites. The litter accumulation correlated well with levels of all metals but not with the pH of the litter, which varied between 3.88 and 6.3 over the sites. Partial correlation analysis showed that cadmium and zinc interrelated; with both cadmium and zinc, partial correlation coefficients were highly significant when lead, copper or pH effects were accounted for. However, correlations were low for lead and copper when cadmium or zinc were accounted for. Of further interest was an analysis of cadmium and zinc in leaf litter from various sites. There was an increase in the smaller particle sizes in contaminated sites relative to uncontaminated ones. These smaller particles contained a disproportionate amount of the metals present, particularly in the case of cadmium. Litter standing crop and cadmium concentrations were highly correlated; the correlation between cadmium content and different particle sizes of litter was better for small particles sizes, and the slope of the regression line between cadmium concentration and litter weight decreased with increasing particle size. The authors argued that litter degradation was not affected at the early stages but only when breakdown had progressed to much smaller particle sizes. This perhaps supports the view that microorganisms are inhibited by metals to a greater extent than invertebrates, which would produce the initial reduction in size of litter fragments. Taking the figure of 900 g/m2 as the normal leaf litter level for woodland, the extra accumulation in contaminated woods represented 25 to 30 years of litter accumulation (the smelter in question had been operating for 48 years) and, possibly a large proportion of the total capital of nutrients normally recycled to the plants. Other authors have also reported accumulation of leaf litter in areas contaminated by metals (Tyler, 1972; Strojan, 1978), although they disagree about the probable cause. Strojan (1978) proposed that the effect relates to the absence of some groups of invertebrates, while Jordan & Lechavalier (1975) suggested that the effect is on microorganisms. There is no direct evidence that invertebrates in leaf litter are adversely affected by metals, although they do accumulate all the metals found in litter (Martin et al., 1976; Coughtrey & Martin, 1976). Both Tyler (1972) and Strojan (1978) argued that the productivity of woodland may be adversely affected by the failure to recycle nutrients in areas contaminated by heavy metals. Coughtrey et al. (1979) considered that the litter is an important sink for heavy metals, and that the result of litter organisms developing tolerance to the metals and, therefore, in the long-term increasing the rate of degradation is unpredictable since metals would be released at the same time as nutrients. In a study by van Straalen et al. (1987), the metal excretion efficiency of the collembolan Orchesella cincta collected from various contaminated forest soils was monitored. The authors found that moderate to high soil cadmium contamination of industrial origin did not evoke increased cadmium excretion. In fact contamination initiated in this century from a zinc factory caused a significant decrease in excretion efficiency. Soils that had been contaminated with cadmium for many years (lead/zinc factory) or to an extreme degree (lead smelter) were inhabited by Collembola able to increase metal excretion. Muskett & Jones (1980) found levels of cadmium to be higher than normal within 10 m of a road with a heavy traffic, but no effect on the numbers of invertebrates caught or their species diversity was observed. 8.3 Effects on fish Field studies in Sweden showed that perch ( Perca fluviatilis) from a cadmium-contaminated river (0.1 to 0.2 µg cadmium/litre) had physiological abnormalities similar to those shown in laboratory experiments (Sjobeck et al., 1984). 8.4 Effects on sea-birds The reported effects on the kidney of sea-birds are not always a result of exposure to cadmium as an industrial pollutant, since the individuals most affected come from areas where there is no industrial effluent. This is often, therefore, a response to naturally occurring cadmium presumed to derive from the oceans. The birds appear to cope with this damage to the kidney and suffer no effects on survival or breeding success. No damage resulting from exposure to strictly anthropogenically derived cadmium appears to have been reported on the same scale as that from exposure to naturally occurring cadmium. Nicholson et al. (1983) compared the ultrastructure of the kidneys of sea-birds contaminated with cadmium in the wild, sea-birds from uncontaminated colonies, starlings dosed with cadmium in the laboratory, and control starlings. They found damage to kidney cells to be comparable between wild sea-birds and dosed starlings having kidney cadmium levels of 60-480 and 95-240 mg/kg, respectively (see section 7.3.3). Nicholson & Osborn (1983) reported kidney lesions (described in section 7.3.3) in several different species of sea-bird caught in contaminated areas, although other pollutant metals such as mercury were also present in the tissues. 9. EVALUATION 9.1 General considerations In evaluating the environmental hazard of cadmium, it is necessary to extrapolate from laboratories studies to ecosystems. This must be done with extreme caution for a number of reasons. a) The availability of cadmium to organisms in the environment is limited by its strong adsorption to environmental components such as soil, sediment, and organic matter. Organisms in contaminated areas accumulate high body burdens of cadmium. b) Environmental variables such as temperature, pH, and the chemical composition of water or soil have been shown to affect both the uptake and the toxic impact of cadmium. c) Available, rather than nominal or total, cadmium is the determinant in assessing uptake by, and effects on, organisms. d) There are limited data from controlled experimental studies on the effects of mixtures of metals. Organisms in the environment are exposed to mixtures of pollutants. Acid deposition can release metals, including cadmium, into the environment. e) Little experimental work has been carried out on species or communities that are either representative or key components of natural communities and ecosystems. Studies have not considered all of the interactions between populations and all of the environmental factors affecting these populations. As a result, the impact of cadmium on ecosystems may have been underestimated. f) Results from laboratory studies based on very sensitive parameters may be indicative of physiological impacts on individuals rather than impacts on ecosystems. 9.2 The aquatic environment Cadmium input to the aquatic environment is through dis-charge of industrial waste, surface run-off, and deposition. It is strongly adsorbed onto sediments and soils. The average cadmium content of sea water is about 0.1 µg/litre or less, while fresh waters contain < 0.01 to 0.06 µg/litre in unpolluted areas. Cadmium levels of up to 5 mg/kg and 0.03 to 1 mg/kg have been reported for freshwater sediments and marine sediments, respectively. The rate of uptake and the toxic impact of cadmium on aquatic organisms is greatly affected by physicochemical factors such as temperature, ionic concentration, and organic matter content. Cadmium is translocated by aquatic plants and concentrated in roots and leaves. It is also taken up and accumulated by various aquatic animals. The toxicity of cadmium to freshwater organisms varies considerably depending on the exposure duration, species, and life-stage. The early life-stages and the reproductive system are the most vulnerable. Cadmium is, by comparison, one of the most toxic heavy metals in the freshwater environment. Manifest responses of certain organisms to cadmium are observed at environmental concentrations lower than 1 µg/litre. Cadmium-induced kidney damage has been reported in sea-birds sampled from the field. However, this damage is present in both cadmium-polluted areas and areas remote from industrial contamination. The effect is probably, therefore, due to natural cadmium in certain species and areas. 9.3 The terrestrial environment Cadmium is introduced into the terrestrial environment from mining, non-ferrous metal production, landfill sites and from the application of sewage sludge, phosphate fertilizers, and manure. Background concentrations of cadmium are in the range of 0.1 to 0.4 mg/kg soil and can reach 4.5 mg/kg in volcanic soils. Levels up to 160 mg/kg soil have been found close to metal processing sources. Reduced breakdown of leaf litter and recycling of nutrients has been attributed to metal pollution in the field. Cadmium appears to be the most potent metal at inhibiting litter degradation. The effect is thought to be due largely to reduced populations of microorganisms, which are responsible for the final stages of litter decomposition. Plants take up cadmium and can translocate and accumulate it. However, uptake from soil is limited. Where there is high-level exposure to cadmium (in the range of hundreds of mg/kg), growth reduction is the major effect. Plants exposed to cadmium in the field for long periods can develop tolerance to the metal. There is no evidence of adverse effects of cadmium on plant populations in the field. Terrestrial invertebrates vary considerably in their sensitivity to cadmium. Some species can take up and store cadmium to levels of up to 5000 mg/kg body weight without apparent ill effects, while others show population effects at levels of a few mg/kg soil. Populations of some terrestrial invertebrates could be adversely affected at levels of cadmium contamination seen in the field. Isopods and earthworms are useful biomonitors for cadmium contamination. Invertebrates with high body burdens may pose a threat to predators. Kidney damage was found in experimental birds fed 20 mg cadmium/kg diet for 12 weeks, but not at lower doses. Reproductive effects have been observed at 200 mg/kg diet. A dose of 4 mg/kg affected the behaviour of ducklings. No effects of cadmium have been seen in terrestrial birds sampled from the field, although the cadmium level in the brain, kidney, and liver of pigeons has proved to be a good indicator of urban cadmium contamination. Small mammals accumulate cadmium in the vicinity of mining spoil. The ionic balance was affected in voles exposed experimentally to a concentration of 10 mg/kg diet. Populations of terrestrial organisms may also develop tolerance to cadmium after long-term exposure. 10. RECOMMENDATIONS FOR PROTECTING THE ENVIRONMENT To eliminate environmental effects, emissions of cadmium from the following sources should be reduced as far as is practicable: * smelters * incinerators * sewage sludge applied to the land * phosphate fertilisers * cadmium-containing manure 11. FURTHER RESEARCH a) More study is needed to clarify the effects of cadmium on the decomposition process of plant debris. Effects on the degree of nutrient cycling and long-term plant growth and the exact nature of the inhibition of decomposition require further attention. b) The adsorption of cadmium to soil and sediment requires further study and quantification of coefficients. 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Global emissions of trace metals from natural sources (x 1000 tonnes/year) a Arsenic Cadmium Copper Mercury Lead Selenium Zinc Wind-borne soil particles range 0.3-5.0 0.01-0.4 0.9-15 0-0.01 0.3-7.5 0.01-0.35 3.0-35 median 2.6 0.21 8.0 0.05 3.9 0.18 19 Sea salt spray range 0.19-3.1 0-0.11 0.23-6.9 0-0.04 0.02-2.8 0-1.1 0.02-0.86 median 1.7 0.06 3.6 0.02 1.4 0.55 0.44 Volcanoes range 0.15-7.5 0.14-1.5 0.9-18 0.03-2.0 0.54-6.0 0.10-1.8 0.31-19 median 3.8 0.82 9.4 1.0 3.3 0.95 9.6 Forest fires range 0-0.38 0-0.22 0.1-7.5 0-0.05 0.06-3.8 0-0.52 0.3-15 median 0.19 0.11 3.8 0.02 1.9 0.26 7.6 Biogenic continental particulates range 0.2-0.5 0-0.83 0.1-5.0 0-0.04 0.02-2.5 0-0.25 0.3-5.0 median 0.26 0.15 2.6 0.02 1.3 1.12 2.6 Biogenic continental volatiles range 0.03-2.5 0-0.8 0.01-0.62 0.02-1.2 0.01-0.38 0.15-5.0 0.02-5.0 median 1.3 0.04 0.32 0.61 0.20 2.6 2.5 Appendix 1 (contd). Arsenic Cadmium Copper Mercury Lead Selenium Zinc Biogenic marine range 0.16-4.5 0-0.1 0.02-0.75 0.04-1.5 0.02-0.45 0.4-9.0 0.04-6.0 median 2.3 0.05 0.39 0.77 0.24 4.7 3.0 Total emission range 0.86-23 0.15-2.6 2.3-54 0.1-4.9 0.97-23 0.66-18 4.0-86 median 12 1.3 28 2.5 12 9.3 45 a From: Nriagu (1989) Appendix 2. Natural and anthropogenic emissions of trace metals to the atmosphere in 1983 (x 1000 tonnes/year) a Trace metal Anthropogenic source Natural source Total emission Natural/total emissions Arsenic 19 (12-26) 12 (0.86-23) 31 (13-49) 0.39 Cadmium 7.6 (3.1-12) 1.3 (0.15-2.6) 8.9 (3.2-15) 0.15 Copper 35 (20-51) 28 (2.3-54) 63 (22-105) 0.44 Mercury 3.6 (0.91-6.2) 2.5 (0.10-4.9) 6.1 (1.0-11) 0.41 Lead 332 (289-376) 12 (0.97-23) 344 (290-399) 0.04 Selenium 6.3 (3.0-9.7) 9.3 (0.66-18) 16 (2.5-24) 0.58 Zinc 132 (70-194) 45 (4.0-86) 177 (74-280) 0.34 a From: Nriagu (1989) Appendix 3. Sources of global emissions of trace elements to the atmosphere in 1983 (tonnes/year) a Arsenic Cadmium Copper Mercury Lead Selenium Zinc Coal combustion - electric utilities 232-1550 77-387 930-3100 155-542 775-4650 108-775 1085-7750 Coal combustion - industry and domestic 198-1980 99-495 1390-4950 495-2970 990-9900 792-1980 1485-11 880 Oil combustion - electric utilities 5.8-29 23-174 348-2230 - 232-1740 35-290 174-1280 Oil combustion - industry and domestic 7.2-72 18-72 179-1070 - 716-2150 107-537 358-2506 Pyrometallurgical non-ferrous metal production - mining 40.0-80 0.6-3 160-800 - 1700-3400 18-176 310-620 Pyrometallurgical non-ferrous metal production - Pb production 780-1560 39-195 234-312 7.8-16 11 700-31 200 195-390 195-468 Pyrometallurgical non-ferrous metal production - Cu-Ni production 8500-12 750 1700-3400 14 450-33 600 37-207 11 050-22 100 427-1280 4250-8500 Pyrometallurgical non-ferrous metal production - Zn-Cd production 230-690 920-4600 230-690 - 5520-11 500 92-23 46 000-82 800 Secondary non-ferrous metal production - 2.3-3.6 55-165 - 90-1440 3.8-19 270-1440 Steel and iron manufacturing 355-2480 28-284 142-2840 - 1065-14 200 0.8-2.2 7100-31 950 Refuse incineration - municipal 154-392 56-1400 980-1960 140-2100 1400-2800 28-70 2800-8400 Refuse incineration - sewage sludge 15-60 3-36 30-180 15-60 240-300 3-30 150-450 Appendix 3 (contd). Arsenic Cadmium Copper Mercury Lead Selenium Zinc Phosphate fertilizers - 68-274 137-685 - 55-274 0.4-1.2 1370-6850 Cement production 178-890 8.9-534 - - 18-14 240 - 1780-17 800 Wood combustion 60-300 60-180 600-1200 60-300 1200-3000 - 1200-6000 Mobile sources (gasoline) - - - - 248 030 - - Miscellaneous 1250-2800 - - - 3900-5100 - 1724-4783 Total emissions - range 12 000-25 630 3100-12 040 19 860-50 870 910-6200 288 700-376 000 1810-5780 70 250-193 500 Total emissions - median 18 820 7570 35 370 3560 332 350 3790b 131 880 a From: Nriagu & Pacyna (1988) b This value applies to particulate selenium only. Since volatile selenium accounts for about 40% of the selenium released, the total selenium emission is estimated to be 6300 tonnes per year. Appendix 4. Anthropogenic inputs of trace elements to aquatic ecosystems (x 1000 tonnes/year) a Arsenic Cadmium Copper Mercury Lead Selenium Zinc Atmospheric fallout 3.6-7.7 0.9-3.6 6.0-15 0.22-1.8 87-113 0.54-1.1 21-58 Other industrial sources 8.4-62.3 1.2-13.4 29-75 0.08-7.0 10-67 9.5-70.9 56-317 Total - range 12-70 2.1-17 35-90 0.3-8.8 97-180 10-72 77-375 Total - median 41 9.4 62 4.6 138 41 226 Atmospheric fallout/total anthropogenic inputs (%) 14 24 17 22 72 2 17 a From: Nriagu & Pacyna (1988) Appendix 5. Anthropogenic inputs of trace elements to soils (x 1000 tonnes/year) a Arsenic Cadmium Copper Mercury Lead Selenium Zinc Atmospheric fallout 8.4-18 2.2-8.4 14-36 0.63-4.3 202-263 1.3-2.6 49-135 Other industrial sources 43.6-94 3.4-29.6 527-1331 0.97-10.7 277-850 4.7-73.4 640-1919 Total input - range b 52-112 5.6-38 541-1367 1.6-15 479-1113 6.0-76 689-2054 Total input - median b 82 22 954 8.3 796 41 1372 Atmospheric fallout/total anthropogenic inputs (%) b 16 24 3 30 29 5 7 a From: Nriagu & Pacyna (1988) b These data do not include inputs from mine tailings, smelter slags and wastes to land. RESUME Le cadmium (numéro atomique 48; masse atomique relative 112,40) est un élément métallique qui appartient, avec le zinc et le mercure, au groupe IIb du tableau périodique. Certains sels de cadmium, tels que le sulfure, le carbonate et l'oxyde sont presque insolubles dans l'eau; ils peuvent cependant ętre transformés en sels solubles dans l'environnement naturel. La formation de différents dérivés du cadmium dans l'environnement est importante pour l'évaluation du risque. La teneur moyenne en cadmium de l'eau de mer est d'environ 0,1 µg/litre ou moins. L'eau des cours d'eau contient du cadmium dissous ŕ des concentrations allant de < 1 ŕ 13,5 ng/litre. Dans les régions écartées et inhabitées, on observe en général des concentrations de cadmium dans l'air inférieures ŕ 1 ng/m3. Dans les régions oů l'on ne connaît pas de pollution, la concentration médiane en cadmium dans le sol se situerait entre 0,2 et 0,4 mg/kg. Toutefois, on rencontre occasionnellement des valeurs beaucoup plus élevées pouvant aller jusqu'ŕ 160 mg/kg de terre. Certains facteurs environnementaux influent sur la fixation, et par voie de conséquence, sur les effets toxiques du cadmium sur les organismes aquatiques. Une élévation de température augmente la fixation et l'effet toxique du cadmium qui sont en revanche réduits lorsque la salinité de l'eau ou sa dureté augmentent. Les organismes d'eau douce sont sensibles ŕ des concentrations plus faibles de cadmium que les organismes marins. Plus la teneur de l'eau en matičres organiques est élevée, plus la fixation et les effets toxiques du cadmium sont réduits car ces matičres se lient au cadmium et en réduisent la biodisponibilité. Toutefois, on est fondé ŕ croire que certains dérivés organiques pourraient avoir un effet inverse. Le cadmium s'accumule facilement dans de nombreux organismes, en particulier les microorganismes et les mollusques, pour lesquels le facteur de bioconcentration peut atteindre plusieurs milliers. Les invertébrés terricoles concentrent égale-ment assez fortement le cadmium. Pour la plupart des organismes, le facteur de concentration est faible ŕ modéré, généralement inférieur ŕ 100. Dans de nombreux tissus, le cadmium est lié aux protéines. On a isolé d'organismes exposés au cadmium, des pro-téines qui fixent spécifiquement les métaux lourds (métallothio-néines). Le cadmium se concentre préférentiellement dans les reins, les branchies et le foie (ou leurs équivalents). L'élimination du métal s'effectue probablement par le rein, encore que chez les crustacés, il puisse ętre éliminé en quantités notables en passant dans l'exosquelette. Chez les plantes, le cadmium se concentre principalement dans les racines et ŕ un moindre degré dans les feuilles. Le cadmium est toxique pour de nombreux microorganismes. Toutefois, la présence de sédiments et de fortes concentrations de sels ou de matičres organiques en solution réduit ces effets toxiques. Ces effets s'exercent principalement sur la croissance et la réplication. Parmi les microorganismes terricoles, les plus affectés sont les champignons, dont certaines espčces peuvent ętre éliminées lorsque le sol est contaminé par du cadmium. Une faible exposition au cadmium présent dans le sol peut entraîner une sélection des souches résistantes. La toxicité du cadmium pour les organismes aquatiques est variable, męme lorsqu'il s'agit d'espčces trčs proches et elle est liée ŕ la concentration du métal sous forme ionique. Le cadmium perturbe le métabolisme du calcium chez l'animal. Chez les poissons, il provoque une hypocalcémie, probablement en inhibant la fixation du calcium ŕ partir de l'eau. Toutefois, la présence de fortes concentrations de calcium dans l'eau protčge les poissons par inhibition compétitive de la fixation de cadmium. Le zinc accroît la toxicité du cadmium pour les invertébrés aquatiques. On a fait état d'effets sublétaux sur la croissance et la reproduction d'invertébrés aquatiques; des effets ont été également observés sur la structure des branchies d'invertébrés. Le cadmium est plus ou moins toxique pour les poissons, les salmonidés étant particuličrement sensibles. On a signalé des effets sublétaux chez les poissons, en particuliers des malformations de l'épine dorsale. Les stades les plus sensibles sont les embryons et les jeunes larves, les moins sensibles étant les oeufs. Chez les poissons, on n'observe pas d'interaction systématique entre le cadmium et le zinc. Le cadmium est toxique pour certaines larves d'amphibiens, mais on a constaté que la présence de sédiments dans les aquariums expérimentaux apportait une certaine protection. Le cadmium perturbe la croissance des végétaux au laboratoire, mais aucun effet n'a été observé dans la nature. Les plantes captent plus facilement le métal lorsqu'il est présent dans les solutions nutritives que lorsqu'il est dans le sol; les effets observés l'ont été essentiellement lors d'études portant sur des cultures hydroponiques. Il semblerait que le cadmium présent dans les solutions nutritives affecte l'ouverture des stomates, la transpiration et la photosynthčse. Les invertébrés terrestres sont relativement insensibles aux effets toxiques du cadmium, probablement ŕ cause de l'intervention de mécanismes de séquestration efficaces au niveau des divers organes. Les gastéropodes terrestres peuvent subir des effets sublétaux, principalement en ce qui concerne la consommation de nourriture et la dormance, mais uniquement ŕ des doses trčs élevées. Męme ŕ forte dose, le cadmium n'entraîne pas la mortalité des oiseaux mais peut provoquer des lésions rénales. Des études effectuées sur le terrain ont montré que le cadmium pouvait entraîner la modification de la proportion relative des diverses espčces dans les populations de microorganismes et de certains invertébrés aquatiques. La décomposition des feuilles mortes est fortement entravée par une forte pollution due aux métaux lourds, et le cadmium en serait le principal responsable. RESUMEN El cadmio (número atómico 48; masa atómica relativa 112,40) es un elemento metálico que pertenece, junto con el zinc y el mercurio, al grupo IIb de la tabla periódica. Algunas sales de cadmio, como el sulfuro, el carbonato y el óxido, son prácticamente insolubles en agua; pueden convertirse en sales hidrosolubles en el medio natural. El sulfato, el nitrato y los haluros son hidrosolubles. La especiación del cadmio en el medio ambiente tiene importancia para evaluar su potencial de riesgo. El contenido medio de cadmio en el agua de mar es de alrededor de 0,1 µg/litro o menos. El agua de los ríos contiene cadmio disuelto en concentraciones que varían entre < 1 y 13,5 ng/litro. En zonas aisladas y deshabitadas, las concentraciones de cadmio en el aire suelen ser inferiores a 1 ng/m3. En zonas que se suponen no contaminadas, se ha comunicado que la concentración mediana de cadmio en el suelo se encuentra entre 0,2 y 0,4 mg/kg. No obstante, a veces se encuentran valores mucho más altos, que pueden llegar hasta 160 mg/kg de suelo. Los factores ambientales influyen en la captación y, por ende, en los efectos tóxicos del cadmio en los organismos acuáticos. Al aumentar la temperatura aumentan la captación y los efectos tóxicos, mientras que el aumento de la salinidad o de la dureza del agua los hace disminuir. Los organismos de agua dulce sufren los efectos del cadmio en concentraciones inferiores a las que afectan a los organismos marinos. La materia orgánica contenida en el agua suele reducir la captación y los efectos tóxicos fijando el cadmio y reduciendo su disponibilidad para los seres vivos. Sin embargo, hay pruebas de que cierto tipo de materia orgánica puede ejercer el efecto contrario. El cadmio se acumula fácilmente en numerosos seres vivos, particularmente microorganismos y moluscos, en los que los factores de bioconcentración son del orden de varios millares. Los invertebrados del suelo también concentran este metal en grado considerable. La mayoría de los organismos presentan factores de concentración bajos o moderados (inferiores a 100). El cadmio está ligado a proteínas en numerosos tejidos. En organismos expuestos a ese metal se han aislado proteínas fijadoras de metales pesados (metalotioneínas). La concentración de cadmio es más elevada en el rińón, las branquias y el hígado (o sus equivalentes). La eliminación se produce probablemente por vía renal, si bien los crustáceos pueden eliminar cantidades importantes con la muda del exoesqueleto. En las plantas, el cadmio se concentra principalmente en las raíces y, en menor medida, en las hojas. El cadmio tiene efectos tóxicos para muy diversos microorganismos. No obstante, la presencia de sedimentos y las concentraciones elevadas de sales disueltas o materia orgánica reducen esos efectos. Los procesos más afectados son el crecimiento y la replicación. Los organismos del suelo más vulnerables son los hongos: algunas especies desaparecen tras la exposición al cadmio en el suelo. Tras exposiciones reducidas al metal en el suelo, se observa una selección a favor de las cepas resistentes. La toxicidad aguda del cadmio para los organismos acuáticos es variable, incluso entre especies estrechamente emparentadas, y guarda relación con la concentración de iones libres del metal. El cadmio interacciona con el metabolismo del calcio en los animales. En los peces provoca hipocalcemia, probablemente al inhibir la captación de calcio a partir del agua. No obstante, las concentraciones elevadas de calcio en el agua los protegen de la ingestión de cadmio por competencia en los lugares de captación. El zinc aumenta la toxicidad del cadmio para los invertebrados acuáticos. Se han notificado efectos subletales en el crecimiento y la reproducción de invertebrados acuáticos, así como modificaciones estructurales en las branquias de invertebrados. Hay pruebas de la selección de estirpes resistentes de invertebrados acuáticos tras la exposición al cadmio sobre el terreno. La toxicidad es variable en los peces; los salmónidos son especialmente susceptibles. Se han notificado efectos subletales en los peces, en particular malformaciones de la espina dorsal. Las fases biológicas más susceptibles son el embrión y la larva joven; los huevos son los menos vulnerables. No se ha observado una interacción homogénea entre el cadmio y el zinc en los peces. El cadmio resulta tóxico para algunas larvas de anfibios, si bien los sedimentos en el recipiente de ensayo confieren cierta protección. El cadmio afecta al crecimiento de las plantas en estudios experimentales, pero sobre el terreno no se ha observado efecto alguno. El metal es absorbido por las plantas con más rapidez a partir de soluciones de nutrientes que a partir del suelo; los efectos se han observado sobre todo en estudios de cultivo en soluciones de nutrientes. En éstas se ha notificado que el cadmio influye en la apertura de los estomas, la transpiración y la fotosíntesis. Los invertebrados terrestres son relativamente insensibles a los efectos tóxicos del cadmio, probablemente debido a la existencia de mecanismos eficaces de captura y fijación en ciertos órganos. Los gasterópodos terrestres sufren efectos subletales; los principales procesos afectados son el consumo de alimentos y el letargo, pero sólo con dosis muy elevadas. El metal no produce efectos letales en las aves, ni siquiera con dosis elevadas, si bien se observan lesiones renales. En estudios de campo se ha comunicado que el cadmio induce cambios en la composición de especies en las poblaciones de microorganismos y ciertos invertebrados acuáticos. La descomposición del mantillo de hojas se ve notablemente reducida por la contaminación con metales pesados; se ha identificado al cadmio como el principal causante de este efecto.
See Also: Cadmium (EHC 134, 1992) Cadmium (ICSC) Cadmium (PIM 089)