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 Organization or the World Health Organization.

First draft prepared by Dr R. Liteplo and Ms R. Gomes, Health Canada, Ottawa, Canada and Mr P. Howe and Mr H. Malcolm, Centre for Ecology and Hydrology, Cambridgeshire, United Kingdom

Published under the joint sponsorship of the United Nations Environment Programme, the International Labour Organization and the World Health Organization, and produced within the framework of the Inter-Organization Programme for the Sound Management of Chemicals.

World Health Organization           Geneva, 2002

The International Programme on Chemical Safety (IPCS), established in 1980, is a joint venture of the United Nations Environment Programme (UNEP), the International Labour Organization (ILO) and the World Health Organization (WHO). The overall objectives of the IPCS are to establish the scientific basis for assessment of the risk to human health and the environment from exposure to chemicals, through international peer review processes, as a prerequisite for the promotion of chemical safety, and to provide technical assistance in strengthening national capacities for the sound management of chemicals.

The Inter-Organization Programme for the Sound Management of Chemicals (IOMC) was established in 1995 by UNEP, ILO, the Food and Agriculture Organization of the United Nations, WHO, the United Nations Industrial Development Organization, the United Nations Institute for Training and Research and the Organisation for Economic Co-operation and Development (Participating Organizations), following recommendations made by the 1992 UN Conference on Environment and Development to strengthen cooperation and increase coordination in the field of chemical safety. The purpose of the IOMC is to promote coordination of the policies and activities pursued by the Participating Organizations, jointly or separately, to achieve the sound management of chemicals in relation to human health and the environment.

WHO Library Cataloguing-in-Publication Data

Fluorides.

(Environmental health criteria ; 227)

1.Fluorides - adverse effects 2.Environmental exposure 3.Occupational exposure

4.Risk assessment I.International Programme for Chemical Safety II.Series

ISBN 92 4 157227 2           (NLM classification: QV 282)

ISSN 0250-863X

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©World Health Organization 2002

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This publication was made possible by grant number 5 U01 ES02617-15 from the National Institute of Environmental Health Sciences, National Institutes of Health, USA, and by financial support from the Federal Ministry for the Environment, Nature conservation and Nuclear Safety, Germany.

In 1973, the WHO Environmental Health Criteria Programme was initiated with the following objectives:

1. to assess information on the relationship between exposure to environmental pollutants and human health, and to provide guidelines for setting exposure limits;
2. to identify new or potential pollutants;
3. to identify gaps in knowledge concerning the health effects of pollutants;
4. to promote the harmonization of toxicological and epidemiological methods in order to have internationally comparable results.

The first Environmental Health Criteria (EHC) monograph, on mercury, was published in 1976, and since that time an ever-increasing number of assessments of chemicals and of physical effects have been produced. In addition, many EHC monographs have been devoted to evaluating toxicological methodology, e.g., for genetic, neurotoxic, teratogenic and nephrotoxic effects. Other publications have been concerned with epidemiological guidelines, evaluation of short-term tests for carcinogens, biomarkers, effects on the elderly and so forth.

Since its inauguration, the EHC Programme has widened its scope, and the importance of environmental effects, in addition to health effects, has been increasingly emphasized in the total evaluation of chemicals.

The original impetus for the Programme came from World Health Assembly resolutions and the recommendations of the 1972 UN Conference on the Human Environment. Subsequently, the work became an integral part of the International Programme on Chemical Safety (IPCS), a cooperative programme of UNEP, ILO and WHO. In this manner, with the strong support of the new partners, the importance of occupational health and environmental effects was fully recognized. The EHC monographs have become widely established, used and recognized throughout the world.

The recommendations of the 1992 UN Conference on Environment and Development and the subsequent establishment of the Intergovernmental Forum on Chemical Safety with the priorities for action in the six programme areas of Chapter 19, Agenda 21, all lend further weight to the need for EHC assessments of the risks of chemicals.

The criteria monographs are intended to provide critical reviews on the effects on human health and the environment of chemicals and of combinations of chemicals and physical and biological agents. As such, they include and review studies that are of direct relevance for the evaluation. However, they do not describe every study carried out. Worldwide data are used and are quoted from original studies, not from abstracts or reviews. Both published and unpublished reports are considered, and it is incumbent on the authors to assess all the articles cited in the references. Preference is always given to published data. Unpublished data are used only when relevant published data are absent or when they are pivotal to the risk assessment. A detailed policy statement is available that describes the procedures used for unpublished proprietary data so that this information can be used in the evaluation without compromising its confidential nature (WHO (1999) Revised Guidelines for the Preparation of Environmental Health Criteria Monographs. PCS/99.9, Geneva, World Health Organization).

In the evaluation of human health risks, sound human data, whenever available, are preferred to animal data. Animal and in vitro studies provide support and are used mainly to supply evidence missing from human studies. It is mandatory that research on human subjects is conducted in full accord with ethical principles, including the provisions of the Helsinki Declaration.

The EHC monographs are intended to assist national and international authorities in making risk assessments and subsequent risk management decisions. They represent a thorough evaluation of risks and are not, in any sense, recommendations for regulation or standard setting. These latter are the exclusive purview of national and regional governments.

The layout of EHC monographs for chemicals is outlined below.

• Summary — a review of the salient facts and the risk evaluation of the chemical
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• Effects on humans
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• Evaluation of human health risks and effects on the environment
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• Further research
• Previous evaluations by international bodies, e.g., IARC, JECFA, JMPR

Since the inception of the EHC Programme, the IPCS has organized meetings of scientists to establish lists of priority chemicals for subsequent evaluation. Such meetings have been held in: Ispra, Italy, 1980; Oxford, United Kingdom, 1984; Berlin, Germany, 1987; and North Carolina, USA, 1995. The selection of chemicals has been based on the following criteria: the existence of scientific evidence that the substance presents a hazard to human health and/or the environment; the possible use, persistence, accumulation or degradation of the substance shows that there may be significant human or environmental exposure; the size and nature of populations at risk (both human and other species) and risks for the environment; international concern, i.e., the substance is of major interest to several countries; adequate data on the hazards are available.

If an EHC monograph is proposed for a chemical not on the priority list, the IPCS Secretariat consults with the cooperating organizations and all the Participating Institutions before embarking on the preparation of the monograph.

The order of procedures that result in the publication of an EHC monograph is shown in the flow chart on the next page. A designated staff member of IPCS, responsible for the scientific quality of the document, serves as Responsible Officer (RO). The IPCS Editor is responsible for layout and language. The first draft, prepared by consultants or, more usually, staff from an IPCS Participating Institution, is based initially on data provided from the International Register of Potentially Toxic Chemicals and from reference databases such as Medline and Toxline.

The draft document, when received by the RO, may require an initial review by a small panel of experts to determine its scientific quality and objectivity. Once the RO finds the document acceptable as a first draft, it is distributed, in its unedited form, to well over 150 EHC contact points throughout the world who are asked to comment on its completeness and accuracy and, where necessary, provide additional material. The contact points, usually designated by governments, may be Participating Institutions, IPCS Focal Points or individual scientists known for their particular expertise. Generally, some four months are allowed before the comments are considered by the RO and author(s). A second draft incorporating comments received and approved by the Director, IPCS, is then distributed to Task Group members, who carry out the peer review, at least six weeks before their meeting.

The Task Group members serve as individual scientists, not as representatives of any organization, government or industry. Their function is to evaluate the accuracy, significance and relevance of the information in the document and to assess the health and environmental risks from exposure to the chemical. A summary and recommendations for further research and improved safety aspects are also required. The composition of the Task Group is dictated by the range of expertise required for the subject of the meeting and by the need for a balanced geographical distribution.

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When the Task Group has completed its review and the RO is satisfied as to the scientific correctness and completeness of the document, the document then goes for language editing, reference checking and preparation of camera-ready copy. After approval by the Director, IPCS, the monograph is submitted to the WHO Office of Publications for printing. At this time, a copy of the final draft is sent to the Chairperson and Rapporteur of the Task Group to check for any errors.

It is accepted that the following criteria should initiate the updating of an EHC monograph: new data are available that would substantially change the evaluation; there is public concern for health or environmental effects of the agent because of greater exposure; an appreciable time period has elapsed since the last evaluation.

All Participating Institutions are informed, through the EHC progress report, of the authors and institutions proposed for the drafting of the documents. A comprehensive file of all comments received on drafts of each EHC monograph is maintained and is available on request. The Chairpersons of Task Groups are briefed before each meeting on their role and responsibility in ensuring that these rules are followed.

Professor Peter Aggett, Lancashire Postgraduate School of Medicine and Health, University of Central Lancashire, Preston, Lancashire, United Kingdom

Dr Roberto Belmar, Environmental Health Division, Ministry of Health, Santiago, Chile (Chairman)

Dr John Bucher, Environmental Toxicology Program, National Institute of Environmental Health Sciences, National Institutes of Health, Research Triangle Park, NC, USA

Dr Julio Camargo, Ecology Department, Faculty of Science, University of Alcalá, Madrid, Spain

Dr Jane Cauley, Department of Epidemiology, University of Pittsburgh, Pittsburgh, PA, USA

Professor Jan Ekstrand, Department of Basic Oral Sciences, Karolinska Institute, Stockholm, Sweden

Mr Paul Howe, Centre for Ecology and Hydrology, Monks Wood, Abbots Ripton, Huntingdon, Cambridgeshire, United Kingdom (Co-Rapporteur)

Dr Gopalakrishnan Karthikeyan, Department of Chemistry, Gandhigram Rural Institute, Gandhigram, Tamil Nadu, India

Dr Uwe Kierdorf, Institute of General and Systematic Zoology, Justus-Liebig-University of Giessen, Giessen, Germany

Dr Päivi Kurttio, Radiation and Nuclear Safety Authority, Helsinki, Finland

Dr Robert Liteplo, Existing Substances Division, Bureau of Environmental Contaminants, Health Canada, Ottawa, Ontario, Canada (Co-Rapporteur)

Dr Akiyoshi Nishikawa, National Institute of Health Sciences, Tokyo, Japan

Mr Daryl Stevens, Land and Water, Commonwealth Scientific and Industrial Research Organisation, Adelaide, Australia

Professor Paolo Vineis, Department of Biomedical Science and Human Oncology, University of Torino, Torino, Italy

Dr Jin Yinlong, Institute of Environmental Health and Engineering, Chinese Academy of Preventive Medicine, Beijing, People’s Republic of China

Dr Antero Aitio, International Programme on Chemical Safety, World Health Organization, Geneva, Switzerland (Secretary)

Dr Bing Heng Chen, Department of Environmental Health, School of Public Health, Shanghai Medical University, Shanghai, People’s Republic of China

Mr John Fawell, Director, Environmental Division, Warren Associates, Devizes, Wiltshire, United Kingdom

Ms Rose Gomes, Existing Substances Division, Bureau of Environmental Contaminants, Health Canada, Ottawa, Ontario, Canada

Dr Philip Jenkins, International Programme on Chemical Safety, World Health Organization, Geneva, Switzerland

A WHO Task Group on Environmental Health Criteria for Fluorides met at the Institute of Environmental Health and Engineering of the Chinese Academy of Preventive Medicine in Beijing, People’s Republic of China, on 28 May – 1 June 2001. The group reviewed the draft document and the peer review comments and revised and further updated the draft, including the evaluation of the risks for human health and the environment from exposure to fluorides.

The first and second drafts of this monograph were prepared by Dr R. Liteplo, Health Canada, Canada, and Mr P. Howe, Centre for Ecology and Hydrology, United Kingdom. The document was sent for peer review to the IPCS contact points and additional experts on fluoride. The authors, in collaboration with the IPCS Secretariat, revised the document based on the comments received. Following an updating at the end of 2000, the document was sent for review to the Task Group members and further revised based on these comments.

Peer review comments were received from the following:

Dr J. Ahlers, Umwelt Bundes Amt, Germany

Dr R. Benson, Region VIII, US Environmental Protection Agency, USA

Professor G.B. Bliss, N.N. Petrov’s Research Institute of Oncology, Russian Federation

Dr M. Bolger, US Food and Drug Administration, USA

Dr J. Bucher, National Institute of Environmental Health Sciences, USA

Dr J. Camargo, University of Alcalá, Spain

Dr S. Cao, Chinese Academy of Preventive Medicine, People’s Republic of China

Dr F.M. Carpanini, European Centre for Ecotoxicology and Toxicology of Chemicals, Belgium

Dr J. Cauley, University of Pittsburgh, USA

Dr L.K. Cohen, National Institute of Dental Research, USA

Dr A. Conacher, Health Canada, Canada

Dr P. Dargan, National Poison Information Service, United Kingdom

Professor I. Dési, Albert Szent-Györgyi University, Hungary

Dr J. Donohue, US Environmental Protection Agency, USA

Dr J. Ekstrand, Karolinska Institute, Sweden

Dr L. Friberg, Karolinska Institute, Sweden

Dr R. Hertel, Federal Institute for Health Protection of Consumers and Veterinary Medicine, Germany

Dr C. Hiremath, National Center for Environmental Assessment, US Environmental Protection Agency, USA

Dr B.L. Johnson, Agency for Toxic Substances and Disease Registry, USA

Dr G. Karthikeyan, Gandhigram Rural Institute, India

Dr U. Kierdorf, Justus-Liebig-University of Giessen, Germany

Dr J. Kriz, National Institute of Public Health, Czech Republic

Dr P. Kurttio, Radiation and Nuclear Safety Authority, Finland

Dr P. Lundberg, National Institute for Working Life, Sweden

Dr E. Ohanian, Office of Water, US Environmental Protection Agency, USA

Dr Y.A. Rakhmanine, Sysin Research Institute of Human Ecology and Environmental Health, Russian Federation

Dr D. Renshaw, Department of Health, United Kingdom

Dr J.M. Rice, International Agency for Research on Cancer, France

Dr T.G. Rossman, New York University School of Medicine, USA

Dr U. Schlottman, Federal Ministry for the Environment, Nature Conservation and Nuclear Safety, Germany

Dr P.A. Schulte, National Institute for Occupational Safety and Health, USA

Dr D. Stevens, Commonwealth Scientific and Industrial Research Organisation, Australia

Dr G. Ungváry, National Institute of Occupational Health, Hungary

Dr P. Vineis, University of Torino, Italy

Ms J. Walter, Swedish Poisons Information Centre, Sweden

Dr G. Whitford, Medical College of Georgia, USA

Dr A. Aitio of the IPCS central unit was responsible for the scientific aspects of the monograph, and Ms M. Sheffer, Ottawa, Canada, for the technical editing.

The efforts of all who helped in the preparation and finalization of the monograph are gratefully acknowledged.

This document focuses on environmental exposure to fluoride derived mostly from inorganic sources and its effects on humans, animals and other biota. Data on hydrogen fluoride, calcium fluoride, sodium fluoride, sulfur hexafluoride and silicofluorides are covered, as these compounds are considered to be the most relevant of the inorganic fluorides on the basis of quantities released to the environment, environmental concentrations and toxicological effects on living organisms.

Hydrogen fluoride (HF) is a colourless, pungent liquid or gas that is highly soluble in organic solvents and in water, in which it forms hydrofluoric acid. Calcium fluoride (CaF₂) is a colourless solid that is relatively insoluble in water and dilute acids and bases. Sodium fluoride (NaF) is a colourless to white solid that is moderately soluble in water. Sulfur hexafluoride (SF₆) is a colourless, odourless, inert gas that is slightly soluble in water and readily soluble in ethanol and bases.

The most common procedure used to quantify free fluoride anion is the fluoride ion-selective electrode. Microdiffusion techniques are considered to be the most accurate methods of sample preparation (i.e., liberation of free ionic fluoride from organic and inorganic complexes).

Fluorides are released into the environment naturally through the weathering and dissolution of minerals, in emissions from volcanoes and in marine aerosols. Fluorides are also released into the environment via coal combustion and process waters and waste from various industrial processes, including steel manufacture, primary aluminium, copper and nickel production, phosphate ore processing, phosphate fertilizer production and use, glass, brick and ceramic manufacturing, and glue and adhesive production. The use of fluoride-containing pesticides as well as the controlled fluoridation of drinking-water supplies also contribute to the release of fluoride from anthropogenic sources. Based on available data, phosphate ore production and use as well as aluminium manufacture are the major industrial sources of fluoride release into the environment.

Hydrogen fluoride is an important industrial compound that is used mainly in the production of synthetic cryolite (Na₃AlF₆), aluminium fluoride (AlF₃), motor gasoline alkylates and chlorofluorocarbons, with an annual world consumption in excess of 1 million tonnes. It is also used in etching semiconductor devices, cleaning and etching glass, cleaning brick and aluminium and tanning leather, as well as in commercial rust removers. Calcium fluoride is used as a flux in steel, glass and enamel production, as the raw material for the production of hydrofluoric acid and anhydrous hydrogen fluoride, and as an electrolyte in aluminium production. Sodium fluoride is used in the controlled fluoridation of drinking-water, as a preservative in glues, in glass and enamel production, as a flux in steel and aluminium production, as an insecticide and as a wood preservative. Sulfur hexafluoride is used extensively in various electronic components and in the production of magnesium and aluminium. Fluorosilicic acid (H₂SiF₆) and sodium hexafluorosilicate (Na₂SiF₆) are used for the fluoridation of drinking-water supplies.

Fluorides in the atmosphere may be in gaseous or particulate form. Atmospheric fluorides can be transported over large distances as a result of wind or atmospheric turbulence or can be removed from the atmosphere via wet and dry deposition or hydrolysis. Fluoride compounds, with the exception of sulfur hexafluoride, are not expected to remain in the troposphere for long periods or to migrate to the stratosphere. Sulfur hexafluoride has an atmospheric residence time ranging from 500 to several thousand years.

The transport and transformation of fluoride in water are influenced by pH, water hardness and the presence of ion-exchange materials such as clays. Fluoride is usually transported through the water cycle complexed with aluminium.

The transport and transformation of fluoride in soil are influenced by pH and the formation of predominantly aluminium and calcium complexes. Adsorption to the soil solid phase is stronger at slightly acidic pH values (5.5–6.5). Fluoride is not readily leached from soils.

The uptake of fluoride by biota is determined by the route of exposure, the bioavailability of the fluoride and the uptake/excretion kinetics in the organism. Soluble fluorides are bioaccumulated by some aquatic and terrestrial biota. However, no information was identified concerning the biomagnification of fluoride in aquatic or terrestrial food-chains.

Terrestrial plants may accumulate fluorides following airborne deposition and uptake from soil.

Fluoride levels in surface waters vary according to location and proximity to emission sources. Surface water concentrations generally range from 0.01 to 0.3 mg/litre. Seawater contains more fluoride than fresh water, with concentrations ranging from 1.2 to 1.5 mg/litre. Higher levels of fluoride have been measured in areas where the natural rock is rich in fluoride, and elevated inorganic fluoride levels are often seen in regions where there is geothermal or volcanic activity (e.g., 25–50 mg fluoride/litre in hot springs and geysers and as much as 2800 mg/litre in certain East African Rift Valley lakes). Anthropogenic discharges can also lead to increased levels of fluoride in the environment.

Airborne fluoride exists in gaseous and particulate forms, which are emitted from both natural and anthropogenic sources. Fluoride released as gaseous and particulate matter is deposited in the general vicinity of an emission source, although some particulates may react with other atmospheric constituents. The distribution and deposition of airborne fluoride are dependent upon emission strength, meteorological conditions, particulate size and chemical reactivity. In areas not in the direct vicinity of emission sources, the mean concentrations of fluoride in ambient air are generally less than 0.1 µg/m³. Levels may be slightly higher in urban than in rural locations; however, even in the vicinity of emission sources, the levels of airborne fluoride usually do not exceed 2–3 µg/m³. In areas of China where fluoride-rich coal is used as a source of fuel, reported concentrations of fluoride in ambient air have reached 6 µg/m³.

Fluoride is a component of most types of soil, with total fluoride concentrations ranging from 20 to 1000 µg/g in areas without natural phosphate or fluoride deposits and up to several thousand micrograms per gram in mineral soils with deposits of fluoride. Airborne gaseous and particulate fluorides tend to accumulate within the surface layer of soils but may be displaced throughout the root zone, even in calcareous soils. The clay and organic carbon content as well as the pH of soil are primarily responsible for the retention of fluoride in soils. Fluoride in soil is primarily associated with the soil colloid or clay fraction. For all soils, it is the soluble fluoride content that is biologically important to plants and animals.

Fluorides can be taken up by aquatic organisms directly from the water or to a lesser extent via food. Fluorides tend to accumulate in the exoskeleton or bone tissue of aquatic animals. Mean fluoride concentrations of >2000 mg/kg have been measured in the exoskeleton of krill; mean bone fluoride concentrations in aquatic mammals, such as seals and whales, ranged from 135 to 18 600 mg/kg dry weight.

Fluoride levels in terrestrial biota are higher in areas with high fluoride levels from natural and anthropogenic sources. Lichens have been used extensively as biomonitors for fluorides. Mean fluoride concentrations of 150–250 mg/kg were measured in lichens growing within 2–3 km of fluoride emission sources, compared with a background level of <1 mg fluoride/kg.

Most of the fluoride in the soil is insoluble and, therefore, less available to plants. However, high soil fluoride concentrations or low pH, clay and/or organic matter can increase fluoride levels in soil solution, increasing uptake via the plant root. If fluoride is taken up through the root, its concentrations are often higher in the root than in the shoot, due to the low mobility of fluoride in the plant. Most fluorides enter plant tissues as gases through the stomata and accumulate in leaves. Small amounts of airborne particulate fluoride can enter the plant through the epidermis and cuticle. Vegetation has been widely monitored in the vicinity of anthropogenic fluoride emission sources. Correlations between fluoride concentrations in vegetation and annual growth increments, wind pattern, distance from fluoride source and hydrogen fluoride concentrations in aerial emissions have been observed.

Fluoride accumulates in the bone tissue of terrestrial vertebrates, depending on factors such as diet and the proximity of fluoride emission sources. For example, mean fluoride concentrations of 7000–8000 mg/kg have been measured in the bones of small mammals in the vicinity of an aluminium smelter.

Fluoride is ubiquitous in the environment; therefore, sources of drinking-water are likely to contain at least some small amount of fluoride. The amount of fluoride present naturally in non-fluoridated drinking-water (i.e., drinking-water to which fluoride has not been intentionally added for the prevention of dental caries) is highly variable, being dependent upon the individual geological environment from which the water is obtained. Levels may range up to approximately 2.0 mg/litre; however, in areas of the world in which endemic fluorosis of the skeleton and/or teeth has been well documented, levels of fluoride in drinking-water supplies range from 3 to more than 20 mg/litre. In areas in which drinking-water is fluoridated (i.e., fluoride is intentionally added for the prevention of dental caries), the concentration of fluoride in drinking-water generally ranges from 0.7 to 1.2 mg/litre.

Virtually all foodstuffs contain at least trace amounts of fluoride. Elevated levels are present in fish. Tea leaves are particularly rich in fluoride; the amount of fluoride in brewed tea is dependent upon the concentration of soluble fluoride in the tea leaves, the level of fluoride in the water used in its preparation and the length of the brewing period. The concentration of fluoride in food products is not significantly increased by the addition of superphosphate fertilizers, which contain significant concentrations of fluoride (1–3%) as impurities, to agricultural soil, due to the generally low transfer coefficient from soil to plant material. However, a recent study suggests that, given the right soil conditions and application of sufficient fluoride as an impurity in phosphate fertilizers to soils, plant uptake of fluoride can be increased. The use of water containing relatively low (<3.1 mg/litre) levels of fluoride for crop irrigation generally does not increase fluoride concentrations in foodstuffs. However, this is dependent on plant species and fluoride concentrations in soil and water. The level of fluoride in foods is significantly affected by the fluoride content of the water used in preparation or processing, most notably in beverages and dry foodstuffs — for example, powdered baby formula — to which water is added prior to consumption. The concentrations of fluoride in unwashed or unprocessed foods grown in the vicinity of industrial sources (emissions) of fluoride may be greater than the levels in the same foods grown in other non-industrially exposed areas. In commercially available infant formulas sold in the USA, soy-based ready-to-use and liquid concentrate formulas contained higher levels of fluoride than the equivalent milk-based products; however, no significant difference was observed between soy- and milk-based powdered infant formulas. Fluoride has been detected in breast milk; reported levels range from <2 to about 100 µg/litre, with most values being between 5 and 10 µg/litre.

Available data on the concentrations of fluoride in indoor air are limited. In the Netherlands, concentrations of gaseous fluoride ranged from <2 to 49 µg/m³ in the indoor air of five homes constructed with wood treated with a preservative containing 56% fluoride. In China, concentrations as high as 155 µg/m³ have been reported for samples of indoor air collected from homes where coal containing high amounts of fluoride was burned indoors.

Dentifrice products for adults that are commercially available in many countries generally contain fluoride at concentrations ranging from 1000 to 1500 µg/g; some products designed for use by children contain lower levels, ranging from 250 to 500 µg/g. Dental products such as toothpaste, mouthwash and fluoride supplements have been identified as significant sources of fluoride. Mouth rinses marketed for daily home use usually contain between 230 and 500 mg fluoride/litre, whereas mouthwash products intended for weekly or biweekly use may contain 900–1000 mg fluoride/litre.

Although individual exposure to fluoride is likely to be highly variable, the inhalation of airborne fluoride generally makes a minor contribution to the total intake of this substance. For adults, the consumption of foodstuffs and drinking-water is the principal route for the intake of fluoride. In areas of the world in which coal rich in fluoride is used for heating and food preparation, the inhalation of indoor air and consumption of foodstuffs containing increased levels of fluoride also contribute to elevated intakes. Infants fed formula receive 50–100 times more fluoride than exclusively breast-fed infants. The ingestion of dentifrice by young children makes a significant contribution to their total intake of fluoride. In general, estimated intakes of fluoride in children and adolescents do not exceed approximately 2 mg/day. Although adults may have a higher absolute daily intake of fluoride in milligrams, the daily intake of fluoride by children, expressed on a milligram per kilogram body weight basis, may exceed that of adults. In certain areas worldwide in which the concentration of fluoride in the surrounding environment may be exceedingly high and/or where diets are composed of foodstuffs rich in fluoride, estimated intakes of fluoride in adults as high as 27 mg/day have been reported, the principal source being drinking-water obtained from groundwater sources located in geological areas rich in fluoride.

Occupational exposure to fluoride via inhalation or dermal contact likely occurs in individuals involved in the operation of welding equipment or in the processing of aluminium, iron ore or phosphate ore. In relatively recent studies, reported concentrations of airborne fluoride in the potrooms of aluminium smelters have been in the order of 1 mg/m³.

In humans and laboratory animals, the absorption of ingested fluoride into the general circulation occurs primarily in the stomach and intestine and is dependent upon the relative aqueous solubility of the form consumed. Soluble fluorides are almost completely absorbed from the gastrointestinal tract; however, the extent of absorption may be reduced by complex formation with aluminium, phosphorus, magnesium or calcium. There is partial to complete absorption of gaseous and particulate fluorides from the respiratory tract, with the extent of absorption dependent upon solubility and particle size.

Fluoride is rapidly distributed by the systemic circulation to the intracellular and extracellular water of tissues; however, in humans and laboratory animals, approximately 99% of the total body burden of fluoride is retained in bones and teeth. In teeth and skeletal tissue, fluoride becomes incorporated into the crystal lattice.

Fluoride crosses the placenta and is transferred from mother to fetus. Fluoride is eliminated from the body primarily in the urine. In infants, about 80–90% of a fluoride dose is retained; in adults, the corresponding figure is approximately 60%. These values can be altered by alterations in urinary flow and urinary pH.

Fluoride is present in body organs, tissues and fluids. Concentrations of fluoride in whole blood of individuals residing in a community in the USA receiving fluoridated drinking-water ranged from 20 to 60 µg/litre. The mean plasma level in 127 subjects with 5.03 mg fluoride/litre in their drinking-water was 106 ± 76 (SD) µg/litre. Serum and plasma contain virtually the same amount of fluoride. Levels of fluoride in calcified tissues are generally highest in bone, dentine and enamel. The concentration of fluoride in bone varies with age, sex and the type and specific part of bone and is believed to reflect an individual’s long-term exposure to fluoride. The concentration of fluoride in dental enamel decreases exponentially with the distance from the surface and varies with site, surface attrition, systemic exposure and exposure to topically applied fluoride. The concentration of fluoride in soft tissues is reflected by that in blood. Levels of fluoride in the urine of healthy individuals are related to the intake of fluoride. Increased levels of urinary fluoride have been measured in individuals following occupational exposure to airborne fluoride and among those residing in areas associated with endemic fluorosis.

Effects on the skeleton, such as inhibition of bone mineralization and formation, delayed fracture healing and reductions in bone volume and collagen synthesis, have been observed in a variety of studies in which rats received fluoride orally for periods of 3–5 weeks. In medium-term exposure studies, altered bone remodelling, hepatic megalocytosis, nephrosis, mineralization of the myocardium, necrosis and/or degeneration of the seminiferous tubules in the testis were observed in mice administered fluoride in drinking-water (>4.5 mg/kg body weight per day) over a period of 6 months.

In a comprehensive carcinogenicity bioassay in which groups of male and female F344/N rats and B6C3F₁ mice were administered drinking-water containing up to 79 mg fluoride/litre as sodium fluoride for a period of 2 years, there was no statistically significant increase in the incidence of any tumour in any single exposed group. There was a statistically significant trend of an increased incidence of osteosarcomas in male rats with increasing exposure to fluoride. However, the incidence was within the range of historical controls.

Another 2-year carcinogenicity bioassay involving Sprague-Dawley rats exposed to up to 11.3 mg/kg body weight per day in the diet also found no statistically significant increase in the incidence of osteosarcoma or other tumours. Another study, which reported an increased incidence of osteomas in mice receiving up to 11.3 mg/kg body weight per day, is difficult to interpret, because the animals were infected with Type C retrovirus.

In general, fluoride is not mutagenic in prokaryotic cells. Although fluoride has been shown to increase the frequency of mutations at specific loci in cultured mouse lymphoma and human lymphoblastoid cells, these mutations are likely due to chromosomal damage rather than point mutations. Fluoride has been shown to be clastogenic in a variety of cell types. The mechanism of clastogenicity has been attributed to the effect of fluoride upon the synthesis of proteins involved in DNA synthesis and/or repair, rather than direct interaction between fluoride and DNA. In most studies in which fluoride was administered orally to rodents, there was no effect upon sperm morphology or the frequency of chromosomal aberrations, micronuclei, sister chromatid exchange or DNA strand breaks. However, cytogenetic damage in bone marrow or alterations in sperm cell morphology were reported when the substance was administered to rodents by intraperitoneal injection.

Reproductive or developmental effects were not observed in recent studies in which laboratory animals were administered fluoride in drinking-water. However, histopathological changes in reproductive organs have been reported in male rabbits administered (orally) 4.5 mg fluoride/kg body weight per day for 18–29 months, in male mice administered (orally) >4.5 mg fluoride/kg body weight per day for 30 days and in female rabbits injected subcutaneously with >10 mg fluoride/kg body weight per day for 100 days. Adverse effects on reproductive function have been reported in female mice administered (orally) >5.2 mg fluoride/kg body weight per day on days 6–15 after mating and in male rabbits administered (orally) >9.1 mg fluoride/kg body weight per day for 30 days.

Epidemiological investigations on the effects of fluoride on human health have examined occupationally exposed workers employed primarily in the aluminium smelting industry and populations consuming fluoridated drinking-water. In a number of analytical epidemiological studies of workers occupationally exposed to fluoride, an increased incidence of lung and bladder cancer and increased mortality due to cancer of these and other sites have been observed. In general, however, there has been no consistent pattern; in some of these epidemiological studies, the increased morbidity or mortality due to cancer can be attributed to the workers’ exposure to substances other than fluoride.

The relationship between the consumption of fluoridated drinking-water and morbidity or mortality due to cancer has been examined in a large number of epidemiological studies, performed in many countries. There is no consistent evidence of an association between the consumption of controlled fluoridated drinking-water and increased morbidity or mortality due to cancer.

Fluoride has both beneficial and detrimental effects on tooth enamel. The prevalence of dental caries is inversely related to the concentration of fluoride in drinking-water. The prevalence of dental fluorosis is highly associated with the concentration of fluoride, with a positive dose–response relationship.

Cases of skeletal fluorosis associated with the consumption of drinking-water containing elevated levels of fluoride continue to be reported. A number of factors, such as nutritional status and diet, climate (related to fluid intake), concomitant exposure to other substances and the intake of fluoride from sources other than drinking-water, are believed to play a significant role in the development of this disease. Skeletal fluorosis may develop in workers occupationally exposed to elevated levels of airborne fluoride; however, only limited new information was identified.

Evidence from several ecological studies has suggested that there may be an association between the consumption of fluoridated water and hip fractures. Other studies, however, including analytical epidemiological investigations, have not supported this finding. In some cases, a protective effect of fluoride on fracture has been reported.

Two studies permit an evaluation of fracture risk across a range of fluoride intakes. In one study, the relative risks of all fractures and of hip fracture were elevated in groups drinking water with >1.45 mg fluoride/litre (total intake >6.5 mg/day); this difference reached statistical significance for the group drinking water containing >4.32 mg fluoride/litre (total intake 14 mg/day). In the other study, an increased incidence of fractures was observed in one age group of women exposed to fluoride in drinking-water in a non-dose-dependent manner.

Epidemiological studies show no evidence of an association between the consumption of fluoridated drinking-water by mothers and increased risk of spontaneous abortion or congenital malformation. Other epidemiological investigations of occupationally exposed workers have provided no reasonable evidence of genotoxic effects or systemic effects upon the respiratory, haematopoietic, hepatic or renal systems that may be directly attributable to fluoride exposure per se.

Fluoride did not affect growth or chemical oxygen demand degrading capacity of activated sludge at concentrations of 100 mg/litre. The EC₅₀ for inhibition of bacterial nitrification was 1218 mg fluoride/litre. Ninety-six-hour EC₅₀s, based on growth, for freshwater and marine algae were 123 and 81 mg fluoride/litre, respectively.

Forty-eight-hour LC₅₀s for aquatic invertebrates range from 53 to 304 mg/litre. The most sensitive freshwater invertebrates were the fingernail clam (Musculium transversum), with statistically significant mortality (50%) observed at a concentration of 2.8 mg fluoride/litre in an 8-week flow-through experiment, and several net-spinning caddisfly species (freshwater; family: Hydropsychidae), with "safe concentrations" (8760-h EC_0.01s) ranging from 0.2 to 1.2 mg fluoride/litre. The brine shrimp (Artemia salina) was the most sensitive marine species tested. In a 12-day static renewal test, statistically significant growth impairment occurred at 5.0 mg fluoride/litre.

Ninety-six-hour LC₅₀s for freshwater fish range from 51 mg/litre (rainbow trout, Oncorhynchus mykiss) to 460 mg/litre (threespine stickleback, Gasterosteus aculeatus). All of the acute toxicity tests (96 h) on marine fish gave results greater than 100 mg/litre. Inorganic fluoride toxicity to freshwater fish appears to be negatively correlated with water hardness (calcium carbonate) and positively correlated with temperature. The symptoms of acute fluoride intoxication include lethargy, violent and erratic movement and death. Twenty-day LC₅₀s for rainbow trout ranged from 2.7 to 4.7 mg fluoride/litre in static renewal tests. "Safe concentrations" (infinite hours LC_0.01s) have been estimated for rainbow trout and brown trout (Salmo trutta) at 5.1 and 7.5 mg fluoride/litre, respectively. At concentrations of >3.2 (effluent) or >3.6 (sodium fluoride) mg fluoride/litre, the hatching of catla (Catla catla) fish eggs was delayed by 1–2 h.

Behavioural experiments on adult Pacific salmon (Oncorhynchus sp.) in soft-water rivers indicate that changes in water chemistry resulting from an increase in the fluoride concentration to 0.5 mg/litre can adversely affect migration; migrating salmon are extremely sensitive to changes in the water chemistry of their river of origin. In laboratory studies, fluoride seems to be toxic for microbial processes at concentrations found in moderately fluoride polluted soils; similarly, in the field, accumulation of organic matter in the vicinity of smelters has been attributed to severe inhibition of microbial activity by fluoride.

Signs of inorganic fluoride phytotoxicity (fluorosis), such as chlorosis, necrosis and decreased growth rates, are most likely to occur in the young, expanding tissues of broadleaf plants and elongating needles of conifers. The induction of fluorosis has been clearly demonstrated in laboratory, greenhouse and controlled field plot experiments. A large number of the papers published on fluoride toxicity to plants concern glasshouse fumigation with hydrogen fluoride. Foliar necrosis was first observed on grapevines (Vitis vinifera) exposed to 0.17 and 0.27 µg/m³ after 99 and 83 days, respectively. The lowest-observed-effect level for leaf necrosis (65% of leaves) in the snow princess gladiolus (Gladiolus grandiflorus) was 0.35 µg fluoride/m³. Airborne fluoride can also affect plant disease development, although the type and magnitude of the effects are dependent on the specific plant–pathogen combination.

Several short-term solution culture studies have identified a toxic threshold for fluoride ion activity ranging from approximately 50 to 2000 µmol fluoride/litre. Toxicity is specific not only to plant species, but also to ionic species of fluoride; some aluminium fluoride complexes present in solution culture may be toxic at activities of 22–357 µmol fluoride/litre, whereas hydrogen fluoride is toxic at activities of 71–137 µmol fluoride/litre. A few studies have been carried out in which the fluoride exposures have been via the soil. The type of soil can greatly affect the uptake and potential toxicity of fluorides.

In birds, the 24-h LD₅₀ was 50 mg/kg body weight for 1-day-old European starling (Sturnus vulgaris) chicks and 17 mg/kg body weight for 16-day-old nestlings. Growth rates were significantly reduced at 13 and 17 mg fluoride/kg body weight (the highest doses at which growth was monitored). Most of the early work on mammals was carried out on domesticated ungulates. Fluorosis has been observed in cattle and sheep. The lowest dietary level observed to cause an effect on wild ungulates was in a controlled captive study with white-tailed deer (Odocoileus virginianus) in which a general mottling of the incisors characteristic of dental fluorosis was noted in the animals at the 35 mg/kg diet dose.

Aluminium smelters, brickworks, phosphorus plants and fertilizer and fibreglass plants have all been shown to be sources of fluoride that are correlated with damage to local plant communities. Vegetation in the vicinity of a phosphorus plant revealed that the degree of damage and fluoride levels in soil humus were inversely related to the distance from the plant. Average levels of fluoride in vegetation ranged from 281 mg/kg in severely damaged areas to 44 mg/kg in lightly damaged areas; at a control site, the fluoride concentration was 7 mg/kg. Plant communities near an aluminium smelter showed differences in community composition and structure due partly to variations in fluoride tolerance. However, it must be noted that, in the field, one of the main problems with the identification of fluoride effects is the presence of confounding variables such as other atmospheric pollutants. Therefore, care must be taken when interpreting the many field studies on fluoride pollution.

The original findings of fluoride effects on mammals were from studies in the field on domestic animals such as sheep and cattle. Fluoride can be taken up from vegetation, soil and drinking-water. Tolerance levels have been identified for domesticated animals, with the lowest values for dairy cattle at 30 mg/kg feed or 2.5 mg/litre drinking-water. Incidents involving domesticated animals have originated both from natural fluoride sources, such as volcanic eruptions and the underlying geology, and from anthropogenic sources, such as mineral supplements, fluoride-emitting industries and power stations. Symptoms of fluoride toxicity include emaciation, stiffness of joints and abnormal teeth and bones. Other effects include lowered milk production and detrimental effects on the reproductive capacity of animals. The lowest dietary concentration of fluoride to cause fluorosis in wild deer was 35 mg/kg. Investigations of the effects of fluoride on wildlife have focused on impacts on the structural integrity of teeth and bone. In the vicinity of smelters, fluoride-induced effects, such as lameness, dental disfigurement and tooth damage, have been found.

Fluoride has both positive and negative effects on human health, but there is a narrow range between intakes that are associated with these effects. Exposure to all sources of fluoride, including drinking-water and foodstuffs, is important.

There is little information to characterize the dose–response relationships for the different adverse effects. In particular, there are few data on total exposure, particularly with respect to intake and fluoride absorption.

The most serious effect is the skeletal accumulation of fluoride from long-term excessive exposure to fluoride and its effect on non-neoplastic bone disease — specifically, skeletal fluorosis and bone fractures. There is clear evidence from India and China that skeletal fluorosis and an increased risk of bone fractures occur at total intakes of 14 mg fluoride/day and evidence suggestive of an increased risk of bone effects at total intakes above about 6 mg fluoride/day.

In the freshwater environment, natural fluoride concentrations are usually lower than those expected to cause toxicity in aquatic organisms. However, aquatic organisms might be adversely affected in the vicinity of anthropogenic discharges. Fluoride toxicity is dependent on water hardness.

Sensitive plant species growing near anthropogenic sources of fluoride are at risk. The release of fluoride from anthropogenic sources is associated with damage to local terrestrial plant communities, but it is often difficult to attribute these effects to fluoride alone, due to the presence of other atmospheric pollutants. Fluoride is generally strongly adsorbed by soils. Consequently, plant uptake via this pathway is relatively low, and leaching of fluoride through soil is minimal.

Concentrations of fluoride in vegetation in the vicinity of fluoride emission sources, such as aluminium smelters, can be higher than the lowest dietary effect concentration reported for mammals in laboratory experiments. Fluorosis in domesticated animals has been reported. There are still some areas reporting fluorosis incidents in livestock due to uptake of fluoride-rich mineral supplements and drinking-water. Furthermore, there is a potential risk from fluoride-contaminated pasture and soil ingestion due to the long-term use of phosphate fertilizers containing fluoride as an impurity. Fluoride-induced effects, such as lameness and tooth damage, have also been reported in wild mammals close to anthropogenic sources.

All organisms are exposed to fluoride from natural and/or anthropogenic sources. Very high intakes have been observed in areas worldwide in which the environment is rich in fluoride and where groundwater high in fluoride is consumed by humans. Increased exposure might occur in the vicinity of point sources. Fluoride in dental products is an additional source for many people.

Fluoride has both beneficial and detrimental effects on human health, with a narrow range between the intakes at which these occur.

Effects on the teeth and skeleton may be observed at exposures below those associated with the development of other organ- or tissue-specific adverse health effects.

Effects on the bone (e.g., skeletal fluorosis and fracture) are considered the most relevant outcomes in assessing the adverse effects of long-term exposure of humans to fluoride.

Skeletal fluorosis is a crippling disability that has a major public health and socioeconomic impact, affecting millions of people in various regions of Africa, China and India.

Intake of fluoride in water and foodstuffs is the primary causative factor for endemic skeletal fluorosis. In some regions, the indoor burning of fluoride-rich coal also serves as an important source of fluoride.

There are few data from which to estimate total exposure to and the bioavailability of fluoride, and there are inconsistencies in reports on the characterization of its adverse effects.

There is clear evidence from India and China that skeletal fluorosis and an increased risk of bone fractures occur at a total intake of 14 mg fluoride/day and evidence suggestive of an increased risk of bone effects at total intakes above about 6 mg fluoride/day.

Excess exposure to bioavailable fluoride constitutes a risk to aquatic and terrestrial biota.

Fluoride-sensitive species can be used as sentinels for the identification of fluoride hazards to the environment.

There is a need to improve knowledge on the accumulation of fluoride in organisms and on how to monitor and control this.

The biological effects associated with fluoride exposure should be better characterized.

This document focuses on environmental exposure to fluoride derived mostly from inorganic sources and its effects on humans, animals and other biota. Data on hydrogen fluoride, calcium fluoride, sodium fluoride, sulfur hexafluoride and silicofluorides are emphasized, as these compounds are considered the most relevant of the inorganic fluorides on the basis of quantities released to the environment, environmental concentrations and toxicological effects on living organisms.

There is one stable isotope of fluorine (F), with an atomic mass of 18.9984. There are also several radioactive isotopes (¹⁷F, ¹⁸F, ²⁰F, ²¹F and ²²F), with ¹⁸F having the longest half-life (109.7 min) (Weast, 1986).

At room temperature, hydrogen fluoride (HF) (relative molecular mass 20.01; density 0.991 g/litre; CAS No. 7664-39-3) is a colourless, pungent, acrid liquid or gas with a melting point of -83 °C and a boiling point of 19.5 °C. Hydrogen fluoride is highly soluble in many organic solvents and in water, in which it forms hydrofluoric acid (Neumüller, 1981; Weast, 1986).

Calcium fluoride (CaF₂) (relative molecular mass 78.08; CAS No. 7789-75-5) is a colourless solid with a melting point of 1403 °C and a boiling point of 2513 °C. It is relatively insoluble in water — approximately 3000 times less soluble in water than sodium fluoride (McIvor, 1990) — as well as in dilute acids and bases (Neumüller, 1981). Calcium fluoride is also known as fluorite. Fluorspar is a mineral containing 60–97% calcium fluoride, depending on the grade.

Sodium fluoride (NaF) (relative molecular mass 41.99; CAS No. 7681-49-4) is a colourless to white solid with high melting (988–1012 °C) and boiling (1695 °C) points. It is moderately soluble in water (Neumüller, 1981).

Fluorosilicic acid (H₂SiF₆) (relative molecular mass 144.08; CAS No. 16961-83-4), which is also known as hexafluorosilicic acid, hydrofluorosilicic acid, fluosilicic acid or fluorosilicic acid, is a colourless solid that is highly soluble in water.

Sodium hexafluorosilicate (Na₂SiF₆) (relative molecular mass 188.05; CAS No. 16893-85-9), also known as disodium hexafluorosilicate or sodium silicofluoride, is a colourless solid that is moderately soluble in water.

Sulfur hexafluoride (SF₆) (relative molecular mass 146.05; density 6.16 g/litre; CAS No. 2551-62-4) is a colourless, odourless, tasteless, chemically inert and non-flammable gas. It is slightly soluble in water but readily soluble in ethanol and bases (Weast, 1986).

Methods employed for the quantification of fluoride in biological samples and environmental media generally rely on the detection of fluoride ion (F^–). Perhaps the most widely used method of fluoride quantification has involved potentiometry employing the fluoride ion-selective electrode (Neumüller, 1981; Harzdorf et al., 1986; ATSDR, 1993). This method has been used for the quantification of fluoride in biological tissues and fluids (e.g., urine, serum and plasma, organs, bone, teeth), foodstuffs and environmental media (e.g., air, water, soil). Owing to variations in the efficacy of sample preparation procedures, detection limits using the fluoride ion-selective electrode may range from 0.1 to 300 ng/m³ in air, from 1 to 1000 µg/litre in water and from 0.05 to 20 mg/kg in tissues (Harzdorf et al., 1986; ATSDR, 1993). Other approaches for the quantification of fluoride have included spectrophotometry, gas chromatography, ion chromatography, capillary electrophoresis, atomic absorption and photon activation (Neumüller, 1981; ATSDR, 1993; Wen et al., 1996).

Appropriate sample preparation is a critical step in the accurate quantification of fluoride, especially where only the free fluoride ion is measured. For analyses involving biological materials, the most accurate method is the microdiffusion technique, such as the acid-hexamethyldisiloxane (HMDS) diffusion method by Taves (1968), since methods involving acid or alkali digestion may not convert all complex inorganic and organic fluorides into an ionic form that can be conveniently measured (Venkateswarlu, 1983). Open ashing methods may result in the loss of volatile fluoride compounds or of fluoride itself at temperatures in excess of 550 °C, or they may result in contamination with extraneous fluoride (Venkateswarlu, 1975; Campbell, 1987).

Fluorides are released into the environment naturally through the weathering of minerals, in emissions from volcanoes and in marine aerosols (Symonds et al., 1988; ATSDR, 1993). Estimates of the annual global release of hydrogen fluoride from volcanic sources through passive degassing and eruptions range from 60 to 6000 kilotonnes, of which approximately 10% may be introduced directly into the stratosphere (Symonds et al., 1988). Annually, approximately 20 kilotonnes of fluoride may be released in marine aerosols (Symonds et al., 1988).

The main natural source of inorganic fluorides in soil is the parent rock (WHO, 1984). During weathering, some fluoride minerals (e.g., cryolite, or Na₃AlF₆) are rapidly broken down, especially under acidic conditions (Fuge & Andrews, 1988). Other minerals, such as fluorapatite (Ca₅(PO₄)₃F) and calcium fluoride, are dissolved more slowly (Kabata-Pendias & Pendias, 1984). The mineral fluorophlogopite (mica; KMg₃(AlSi₃O₁₀)F₂) is stable in alkaline and calcareous soils (Elrashidi & Lindsay, 1986). However, its solubility is affected by pH and the activities of silicic acid (H₄SiO₄) and aluminium (Al³⁺), potassium (K⁺) and magnesium (Mg²⁺) ions.

3.2.1 Production and use

3.2.1.1 Hydrogen fluoride

Hydrogen fluoride (hydrofluoric acid) is an important industrial compound, with an estimated annual world consumption in excess of 1 million tonnes (Greenwood & Earnshaw, 1984). Hydrogen fluoride is manufactured from calcium fluoride and is used mainly in the production of synthetic cryolite, aluminium fluoride (AlF₃), motor gasoline alkylates and chlorofluorocarbons; however, the demand for chlorofluorocarbons is decreasing as a result of efforts to restrict their use. Hydrogen fluoride is also used in the synthesis of uranium tetrafluoride (UF₄) and uranium hexafluoride (UF₆), both of which are used in the nuclear industry (Neumüller, 1981). It is also used in etching semiconductor devices, cleaning and etching glass, cleaning brick and aluminium and tanning leather, as well as in petrochemical manufacturing processes. Hydrogen fluoride may also be found in commercial rust removers (Upfal & Doyle, 1990).

3.2.1.2 Calcium fluoride

Industrially, calcium fluoride is the principal fluoride-containing mineral used (WHO, 1984). Identified production data were confined to the USA, where the average annual production of calcium fluoride was estimated to range from 118 000 to 225 000 tonnes during 1972–1978 (ATSDR, 1993). The consumption of calcium fluoride (as fluorspar) in Canada in 1989 was estimated at 180 000 tonnes (Government of Canada, 1993); in 1977, the estimated consumption of calcium fluoride in the USA was 1 063 000 tonnes (ATSDR, 1993). Calcium fluoride is used as a flux in steel, glass and enamel production and as the raw material for the production of hydrofluoric acid and anhydrous hydrogen fluoride (Neumüller, 1981). Calcium fluoride is also used as a molten electrolyte for the separation of oxygen and alumina in aluminium production.

3.2.1.3 Sodium fluoride

Data concerning the total annual consumption or production of sodium fluoride worldwide were not identified. Sodium fluoride is usually prepared from hydrofluoric acid and sodium carbonate or sodium hydroxide (Neumüller, 1981); it is used in the controlled fluoridation of drinking-water, as a preservative in certain glues, in glass and enamel production, as a flux in steel and aluminium production, as an insecticide and as a wood preservative (Neumüller, 1981).

3.2.1.4 Fluorosilicic acid

Fluorosilicic acid is an aqueous solution that is most commonly manufactured as a co-product from the manufacture of phosphate fertilizers. It is used widely for the fluoridation of drinking-water, in which it hydrolyses to release fluoride ions. When used for the fluoridation of drinking-water, fluorosilicic acid should meet appropriate standards, such as those published by the American Water Works Association and the European Committee for Standardization or other approved schemes for drinking-water chemicals.

3.2.1.5 Sodium hexafluorosilicate

Sodium hexafluorosilicate, like fluorosilicic acid, is used in the fluoridation of drinking-water. It is normally completely dissolved in water prior to dosing, when it hydrolyses to give fluoride ions. When used for drinking-water fluoridation, it too should meet appropriate standards of purity for drinking-water chemicals.

3.2.1.6 Sulfur hexafluoride

More than 110 tonnes of sulfur hexafluoride are imported into Canada annually (Government of Canada, 1993). This substance is used extensively as an insulation and current interruption medium in electrical switchgear, such as power circuit breakers, in various components in electrical substations (Government of Canada, 1993) and as a protective inert gas over molten metals, such as magnesium and aluminium (Neumüller, 1987). Over 90% of the total amount of sulfur hexafluoride imported into Canada is used in the production of magnesium; the remainder is used in electrical switchgear (Government of Canada, 1993).

3.2.1.7 Fluorapatite

Fluorapatite, an important calcium- and fluoride-containing mineral, is used as a source of phosphates in the fertilizer industry (Neumüller, 1981).

3.2.1.8 Phosphate fertilizers

Phosphate fertilizers are the major source of fluoride contamination of agricultural soils. They are manufactured from rock phosphates, which generally contain around 3.5% fluorine (Hart et al., 1934). However, during the manufacture of phosphate fertilizers, part of the fluoride is lost into the atmosphere during the acidulation process, and the concentration of fluoride in the final fertilizer is lowered further through dilution with sulfur (superphosphates) or ammonium ion (ammoniated phosphates); the final product commonly contains between 1.3 and 3.0% fluorine (McLaughlin et al., 1996). In Australia, an average annual addition of fluoride to soil through fertilization has been estimated to be 1.1 kg/ha.

3.2.2 Emissions

Available quantitative information concerning the release of fluoride into the environment (air, water and soil) from industrial sources is limited. Fluoride is released into the environment via exhaust fumes, process waters and waste from various industrial processes, including steel manufacture, primary aluminium, copper and nickel production, phosphate fertilizer production and use, glass, brick and ceramic manufacturing, and glue and adhesive production. The use of fluoride-containing pesticides as well as the fluoridation of drinking-water supplies also contribute to the release of fluoride from anthropogenic sources.

The total annual amount of fluoride released to the environment from industrial sources was estimated to be in excess of 23 500 tonnes in Canada and 46 600 tonnes in the Netherlands (Sloof et al., 1989; Government of Canada, 1993). The relative contribution of various anthropogenic sources to total emissions of fluoride to air, water and soil in Canada are estimated at 48% for phosphate fertilizer production, 20% for chemical production, 19% for aluminium production, 8% for steel and oil production and 5% for coal burning (Government of Canada, 1993). In the Netherlands, 93% of total fluoride emissions to air, water and soil are derived from phosphate ore production and use, with smaller amounts emitted via mineral processing (2%), the metal industry (4%) and "other industry" (1%) (Sloof et al., 1989). The total amounts of hydrogen fluoride released to air, surface water, underground injection and land in the USA during 1999 were 33 000, 7.7, 1800 and 64 tonnes, respectively. Total amounts of fluorine released to air, surface water and land were 39, 24 and 500 tonnes, respectively (US EPA, 1999).

The cycling of fluoride through the biogeosphere is summarized in Figure 1.

4.1.1 Atmosphere

The fate of inorganic fluorides in the atmosphere is primarily influenced by vaporization, aerosol formation, wet and dry deposition and hydrolysis (Environment Canada, 1994). Non-volatile inorganic fluoride particulates are removed from the atmosphere via condensation or nucleation processes.

Atmospheric fluorides emitted from both natural and anthropogenic sources may be in gaseous or particulate form (Kirk & Lester, 1986). Gaseous forms include hydrogen fluoride, silicon tetrafluoride (SiF₄), fluorosilicic acid and sulfur hexafluoride. Particulate forms include sodium aluminium fluoride (cryolite), aluminium fluoride, calcium fluoride, sodium hexafluorosilicate, lead fluoride (PbF₂) and calcium phosphate fluoride (fluorapatite). Globally, hydrogen fluoride and inorganic fluoride particulates (sodium and calcium fluoride) account for approximately 75% and 25%, respectively, of inorganic fluorides present in the atmosphere (Health Council of the Netherlands, 1990). Fluorine and the silicon fluorides are hydrolysed in the atmosphere to form hydrogen fluoride. Hydrogen fluoride may combine with water vapour to produce an aerosol or fog of aqueous hydrofluoric acid.

Fluorides adsorbed on particulate matter in the atmosphere are generally stable and are not readily hydrolysed, although they may be degraded by radiation if they persist in the atmosphere (US NAS, 1971).

Hydrofluoric acid is approximately 5 orders of magnitude less soluble than hydrochloric acid and will therefore be degassed from marine aerosols more readily than hydrochloric acid. Hydrofluoric acid is expected to be depleted in aged marine aerosols, and this may be a significant source of hydrogen fluoride in the troposphere (Brimblecombe & Clegg, 1988).

Schotte (1987) used a dispersion model to predict the formation and behaviour of the fog formed from the release of hydrogen fluoride to the atmosphere. Initially, the hydrogen fluoride will cool significantly due to depolymerization. The fog will therefore stay near ground level, since it is more dense than ambient air. As the fog mixes with more air, it will begin to warm up and it may rise, depending on the ambient air temperature and the relative humidity.

Davison et al. (1973) reported that between 60 and 74% of atmospheric fluoride in urban coal-burning areas in the United Kingdom was in gaseous form. Similarly, approximately 60% of the fluorides in the atmosphere in the Netherlands are in the gaseous state (Sloof et al., 1989).

Based upon available data, inorganic fluoride compounds, with the exception of sulfur hexafluoride, are not expected to remain in the troposphere for long periods or to migrate to the stratosphere. Estimates of the residence time of sulfur hexafluoride in the atmosphere range from 500 to several thousand years (Ramanathan et al., 1985; Chu, 1991).

Fluoride in aerosols can be transported over large distances by wind or as a result of atmospheric turbulence. The distance travelled is determined by the deposition velocity of both the gaseous hydrogen fluoride and the fluorides in particulate form. The transportation of particles with a diameter greater than 10 µm is determined by the particle falling speed, and the dispersion of such particles is generally limited to the immediate vicinity of the source. Smaller particles are less restricted by the falling speed and can be transported over larger distances (Sloof et al., 1989).

Atmospheric fluorides may be transported to soils and surface waters through both wet and dry deposition processes (US NAS, 1971). Seasonal climatic conditions are expected to influence the rate at which and mode by which atmospheric fluorides are deposited; for example, in the Tamar Valley, Tasmania, wet deposition dominates during winter (high precipitation; June to August), and dry deposition dominates during summer (low precipitation; December to February) (Low & Bloom, 1988).

Wet deposition of fluoride may occur as washout from plumes below cloud or rainout of particulates taken up by clouds. The washout process is of particular importance for the removal of soluble fractions such as hydrogen fluoride aerosols at short distances from the source. It is assumed that all irreversibly soluble gases such as hydrogen fluoride are washed out during showers. The rainout process is more important for the removal of fluorides distant from the source when the plume is situated at least partially in the clouds. The scavenging ratio, the ratio between measured concentrations in rainwater and the atmosphere, was calculated to be 0.15 × 10⁶ (Sloof et al., 1989). For large-scale dispersion of fluorides, the annual average wet deposition rate was 1.4% per hour for fluoride aerosol and 5.9% per hour for gaseous fluorides. These values give an atmospheric residence time of 12 h for gaseous fluoride and 50 h for particulates.

The dry deposition rate for fluoride in the Agra region of India was highest between December and June, when atmospheric fluoride concentrations were highest (Saxena et al., 1994). Seasonally averaged dry deposition rates at the four sites ranged from 0.14 to 0.15 mg/m² per day for summer (March to June), from 0.08 to 0.21 mg/m² per day for winter (October to February) and from 0.008 to 0.03 mg/m² per day for the monsoon season (July to September). Similar patterns of dry deposition were recorded by Chandrawanshi & Patel (1999) for central India during 1995; however, higher deposition rates were reported, with values of up to 1.1 mg/m² per day being recorded during the winter months. The overall mean fluoride flux deposited with dust and rainwater during 1995 in central India was 474 kg/km².

Several studies have been conducted to determine whether fluoride in rainwater was derived from anthropogenic emissions or natural sources such as sea salt cycling. Barnard & Nordstrom (1982) stated that fluoride should not be regarded as a cyclical sea salt, because the fluoride concentrations in rain from areas with no local anthropogenic emissions were not correlated with sea salt availability (as determined by the sodium concentration). Mass balance considerations suggested that the majority of fluoride samples in the rainwater were of anthropogenic origin. Similarly, Saether et al. (1995) calculated that more than 90% of fluoride in precipitation samples collected in southern Italy were of non-marine origin.

The ratio between total fluorine and chloride in rainwater from Wales was greater than the ratio in seawater (Neal et al., 1990). This implied enrichment of total fluorine relative to chloride, reflecting complex fractionation processes in the transport of fluorine from the sea to the atmosphere and back to land as precipitation. The total fluorine/chloride ratio in streamwater was higher than it was in rainwater, suggesting a net release of total fluorine from the catchment to the stream. The source of the release was uncertain, since the total fluorine concentration in baseflow waters was not significantly higher than stormflow values.

Mahadevan et al. (1986) reported a strong correlation between fluoride and sodium concentrations in precipitation samples collected from marine, coastal and inland sites in India. The authors suggested that fluoride in precipitation was derived from the cycling of sea salt. The correlation was not as strong in samples from urban areas, where the majority of fluoride was derived from anthropogenic sources.

The deposition of fluoride emitted from a phosphorus plant was reported to decrease with increasing distance from the source (Sidhu, 1982). The rate of deposition at 1.4 km from the source was calculated to be 2.61 g fluoride/ha per millimetre of rain and 3.10 g fluoride/ha per millimetre of snow water. These data corresponded to an annual deposition rate of 3.43 kg/ha. Fluoride deposition on soil from leaf litter also decreased with increasing distance from the source. Fluoride input to soil ranged from 10 to 720 g/ha per year. Input from precipitation was 5–10 times greater than it was for leaf litter.

Davison & Blakemore (1980) determined the deposition of fluoride at field sites near areas of industrial and urban sources of fluorides. The mean total fluoride deposited from wet and dry deposition and sedimentation was 38.0 µg/dm² per week. Deposition of gaseous fluoride was 23.4 µg/dm² per week.

The average large-scale deposition velocity for total soluble fluoride in the Netherlands was calculated to be 1.4 cm/s (Sloof et al., 1989). This figure was based upon 70% of the soluble fluoride being in a gaseous state and an atmospheric residence time of 14 h for gaseous fluorides and 12 days for aerosol fluorides. The average deposition velocity calculated does not apply to the area surrounding a point source. Under stable atmospheric conditions, a low deposition velocity will be accompanied by high atmospheric concentrations. The deposition velocity of fluoride depends heavily on atmospheric conditions. The deposition velocity for hydrogen fluoride can vary by more than 7 orders of magnitude; for particulate fluoride, it varies by less than 10%. The annual average effective deposition velocity varies with height of the emission source and was calculated to be 1.2 and 2.5 cm/s for low and high source heights, respectively, in the Netherlands.

4.1.2 Water and sediment

In water, the transport and transformation of inorganic fluorides are influenced by pH, water hardness and the presence of ion-exchange materials such as clays (Environment Canada, 1994). Fluoride is usually transported through the water cycle complexed with aluminium (Ares, 1990).

In areas of extreme acidity and alkalinity, inorganic fluorides may leach from fluoride-containing minerals into surface water or groundwater (Cuker & Shilts, 1979). Solubilization of inorganic fluorides from minerals may also be enhanced by the presence of ion-exchange materials (e.g., bentonite clays and humic acid) (Pickering et al., 1988). Once dissolved, inorganic fluorides remain in solution under conditions of low pH and hardness and in the presence of ion-exchange material (Cuker & Shilts, 1979; Sahu & Karim, 1989). Soluble inorganic fluorides may also form aerosols at the air–water interface or vaporize into the atmosphere (Brimblecombe & Clegg, 1988), whereas undissolved species generally undergo sedimentation (Drury et al., 1980).

Kudo et al. (1987) calculated the fluoride mass balance for the once severely polluted Maurienne Valley in the French Alps. Fluoride emission into the valley from aluminium production plants was 500 tonnes per year in 1980. Fluoride output by the river was calculated to be 680 tonnes per year, of which 665 tonnes were due to water flow and 15 tonnes to sediment movements.

Chamblee et al. (1984) analysed approximately 100 estuarine water samples and reported that 10.5% of fluoride originated from fluoride complexes with trivalent cations such as Fe³⁺ and Al³⁺. The proportion of fluoride in the form of magnesium fluoride (MgF₂) ranged from 0.4 to 33.7%. Fluoride concentrations were reported to increase with salinity (within a salinity range of 0.1–16%).

Fluoride in seawater is divided between the following fractions (Stumm & Morgan, 1981):

In seawater, fluorides are removed by the formation of complexes with calcium compounds, principally carbonate and phosphate (Carpenter, 1969). Undissolved fluoride is generally removed from the aquatic phase by sedimentation (US EPA, 1980). Carpenter (1969) calculated a residence time for fluoride in ocean sediment to be 2–3 million years.

Fluorosilicic acid and hydrofluoric acid in high aquatic concentrations such as may be found in industrial waste ponds may volatilize, releasing silicon tetrafluoride and hydrogen fluoride into the atmosphere (US NAS, 1971).

4.1.3 Soil

Factors that influence the mobility of inorganic fluorides in soil are pH and the formation of aluminium and calcium complexes (Pickering, 1985; Environment Canada, 1994).

In more acidic soils, concentrations of inorganic fluoride were considerably higher in the deeper horizons. The low affinity of fluorides for organic material results in leaching from the more acidic surface horizon and increased retention by clay minerals and silts in the more alkaline, deeper horizons (Davison, 1983; Kabata-Pendias & Pendias, 1984). This distribution profile is not observed in either alkaline or saline soils (Gilpin & Johnson, 1980; Davison, 1983). The fate of inorganic fluorides released to soil also depends on the chemical form, rate of deposition, soil chemistry and climate (Davison, 1983).

Fluoride in soil is mainly bound in complexes. The maximum adsorption of fluoride to soil was reported to occur at pH 5.5 (Barrow & Ellis, 1986). In acidic soils with pH below 6, most of the fluoride is in complexes with either aluminium or iron (e.g., AlF²⁺, AlF₂⁺, AlF₃⁰, AlF₄^–, FeF²⁺, FeF₂⁺, FeF₃⁰) (Perrott et al., 1976; Murray, 1984b; Elrashidi & Lindsay, 1986). Fluoride in alkaline soils at pH 6.5 and above is almost completely fixed in soils as calcium fluoride, if sufficient calcium carbonate is available (Brewer, 1966).

Fluoride binds to clay by displacing hydroxide from the surface of the clay (Huang & Jackson, 1965; Bower & Hatcher, 1967; Meeussen et al., 1996). The adsorption follows Langmuir adsorption equations and is strongly dependent upon pH and fluoride concentration. It is most significant at pH 3–4, and it decreases above pH 6.5.

Pickering et al. (1988) determined changes in free fluoride ions and total fluoride levels following equilibration of either poorly soluble fluoride species, such as calcium fluoride and aluminium fluoride, or wastes from aluminium smelters. The experiments were carried out on materials that had different cation-exchange capacities, such as synthetic resins, clay minerals, manganese oxide and a humic acid. Increased amounts of fluoride were released from fluoride salts and fluoride-rich wastes when solids capable of exchanging cations were present. The effect was greatest when there were more exchange sites available and when the fluoride compound cation had greater affinity for the exchange material. In a few cases, soluble complex ions were formed when the released fluoride attacked the substrate, such as illite or alumina wastes.

Fluoride is extremely immobile in soil, as determined by lysimeter experiments. MacIntire et al. (1955) reported that 75.8–99.6% of added fluoride was retained by loam soil for 4 years. Fluoride retention was correlated with the soil aluminium content. The leaching of fluoride occurred simultaneously with the leaching of aluminium, iron and organic material from soil (Polomski et al., 1982). Soil phosphate may contribute to the mobility of inorganic fluoride (Kabata-Pendias & Pendias, 1984). Oelschläger (1971) reported that approximately 0.5–6.0% of the annual addition of fluoride (atmospheric pollution and artificial fertilizers) to a forest and agricultural areas was leached from the surface to lower soil horizons. Arnesen & Krogstad (1998) found that fluoride (added as sodium fluoride) accumulation was high in the upper 0–10 cm of soil columns, where 50–90% of the accumulated fluoride was found. The B-horizons sorbed considerably more fluoride than the Ah-horizons, due to higher content of aluminium oxides/hydroxides. A study by McLaughlin et al. (2001) involving long-term application of phosphate fertilizers has shown a large portion of fluoride applied as impurities in the fertilizer to remain in the 0- to 10-cm depth of the soil profile.

In sandy acidic soils, fluoride tends to be present in water-soluble forms (Shacklette et al., 1974). Street & Elwali (1983) determined the activity of the fluoride ion in acid sandy soils that had been limed. Fluorite was shown to be the solid phase controlling fluoride ion activity in soils between pH 5.5 and 7.0. At pH values below 5.0, the fluoride ion activity indicated supersaturation with respect to fluorite. These data indicate that liming of acid soils may precipitate fluorite, with a subsequent reduction in the concentration of fluoride ion in solution.

Murray (1984b) reported that low amounts of fluoride were leached from a highly disturbed sandy podzol soil of no distinct structure. Even at high fluoride application rates (3.2–80 g per soil column of diameter 0.1 m with a depth of 2 m), only 2.6–4.6% of the fluoride applied was leached in the water-soluble form. The pH of the eluate increased with increasing fluoride application, and this was probably due to adsorption of fluoride, releasing hydroxide ions from the soil metal hydroxides. Over time, the concentration of water-soluble fluoride decreased due to increased adsorption on soil particles.

Mean soil concentrations in Pennsylvania, USA, were 377, 0.38 and 21.7 mg/kg for total fluoride, water-soluble fluoride and resin-exchangeable fluoride, respectively. The authors suggested that fluoride is relatively immobile in soil, since most of the fluoride was not readily soluble or exchangeable (Gilpin & Johnson, 1980).

The water-soluble fluoride in sodic surface soil treated with gypsum increased with increasing exchangeable sodium per cent (Chhabra et al., 1979). The increase in exchangeable sodium per cent also caused an increase in soil pH, which in turn caused an increase in water-soluble fluoride. Incubation studies revealed that a major portion of the added fluoride was adsorbed to soil within the first 8 days. Adsorption to soil followed Langmuir isotherms up to an equilibrium soluble fluoride concentration (11.4 mg/litre), with precipitation at higher concentrations.

Calcium fluoride was formed in soils irrigated with fluoride solutions. Calcium fluoride is formed when the fluoride adsorption capacity is exceeded and the fluoride and calcium ion activities exceed the ion activity product of calcium fluoride (Tracy et al., 1984). Less than 2% of applied fluoride was measured in the leachate, and between 15 and 20% of added fluoride was precipitated as calcium fluoride. Fluoride was precipitated in the upper profile, although the authors expected that once the adsorption mechanisms were exceeded, soluble fluoride would leach deeper into the soil with continued irrigation.

A large fraction of the fluoride in topsoil sampled at a distance of 0.5–1.0 km from an aluminium smelter was reported to be in water-soluble form (Polomski et al., 1982). The authors concluded that the fluoride was present as calcium fluoride.

Breimer et al. (1989) determined the vertical distribution of fluoride in the soil profiles sampled near an industrial region. In calcareous soils, fluoride (as extractable with hydrochloric acid) was restricted to the top 40–50 cm, probably due to the precipitation of calcium fluoride in the presence of lime. A slight leaching of fluoride into the Bt and C horizons was reported in non-calcareous soils. Water-extractable fluoride showed an increase with depth in the A horizons and subsequently decreased to base levels in the lower subsoil.

The adsorption of fluoride from the water phase may be an important transport characteristic in calcareous soils at low flow rates, but this exchange may be rate-limited at high flow rates (Flühler et al., 1982). Dissolved fluoride concentrations may be high around the root zone in soils with a high fluoride input such as from atmospheric deposition. The high concentrations exist only for a limited time until the fluoride is withdrawn from the solution. The adsorption isotherm was reported to be non-linear between initial concentrations of 10 and 50 mg fluoride/litre. Retention of fluoride in uncontaminated calcareous soil was higher than retention in calcareous soil from areas with fluoride contamination. The adsorption and desorption of fluoride in acidic soil were not related to previous fluoride contamination.

Fluoride-containing solutions increased the mobilization and leaching of aluminium from soils. Leaching of aluminium was reported to be greater from soil contaminated from an aluminium smelter than from uncontaminated soil (Haidouti, 1995). In the uncontaminated soil, losses of aluminium from the acid soil were higher than those from the calcareous soil. Arnesen (1998) also found that fluoride can solubilize aluminium, iron and organic material and can increase soil pH through exchange with hydroxide ions.

Unlike other soluble salts, fluoride was not leached from naturally salinized salt-affected soil. It was redistributed within the soil profile (Lavado et al., 1983). The adsorption of fluoride to soils increased with decreasing pH within the pH range 8.5–6. Retention of fluoride in the soil was positively correlated with ammonium acetate extractable iron.

4.2.1 Atmosphere

Molecular fluorine hydrolyses to form hydrogen fluoride, the most stable form of fluorine in the atmosphere (ATSDR, 1993). Silicon tetrafluoride reacts with moisture in the atmosphere, forming hydrated silica and fluorosilicic acid. Uranium hexafluoride, which is used in nuclear power applications, is hydrolysed to hydrogen fluoride and uranyl fluoride (UO₂F₂), which are removed from the atmosphere by condensation or nucleation processes.

4.2.2 Water

In dilute solutions at neutral pH, dissolved fluorides are predominantly present as the fluoride ion (Bell et al., 1970). As the pH decreases below pH 5.5, the proportion of fluoride ions decreases, and the proportion of non-dissociated hydrogen fluoride increases. However, if sufficient aluminium is present in solution, aluminium–fluoride complexes (AlF²⁺, AlF₂⁺ and AlF₃) generally dominate below pH 5.5 until pH 1, where hydrogen fluoride begins to dominate as pH decreases further (Parker et al., 1995). Aluminium solubility is low between pH 5.2 and pH 8.8, whereas at higher pH values, aluminium solubilizes as the aluminate ion (Al(OH)₄^–).

Soluble fluorides are bioaccumulated by some aquatic and terrestrial biota. However, no information was identified concerning the biomagnification of fluoride in aquatic or terrestrial food-chains (Hemens & Warwick, 1972; Barbaro et al., 1981; ATSDR, 1993). Inorganic fluorides tend to accumulate preferentially in the skeletal and dental hard tissues of vertebrates, exoskeletons of invertebrates and cell walls of plants (Hemens & Warwick, 1972; Davison, 1983; Michel et al., 1984; Kierdorf et al., 1997; Sands et al., 1998).

Bioconcentration factors greater than 10 (expressed on a wet weight basis) were reported in both aquatic plants and animals following exposure to solutions of up to 50 mg fluoride/litre (Sloof et al., 1989). Twenty-four hours after sodium fluoride (22 mg fluoride/litre) was released into an experimental pond, the concentration of fluoride in aquatic vascular plants was increased 35-fold, and uptake was also increased in algae (14-fold), molluscs (12-fold) and fish (7-fold) (Kudo & Garrec, 1983).

The water hyacinth (Eichhornia crissipes) has been reported to take up fluoride from water. Plants were exposed to fluoride solutions of 6–26 mg/litre for 4 weeks, and fluoride uptake was related to exposure, ranging from 0.8 to 5 mg/kg (Rao et al., 1973).

The fluoride concentration in Sydney rock oyster (Saccostrea commercialis) sprat increased with increasing fluoride exposure. The sprat were exposed to 0.7–30.7 mg fluoride/litre for 3 weeks and had whole-body concentrations ranging from 45 to 204 mg/kg dry weight (Nell & Livanos, 1988).

Following exposure to inorganic fluoride levels of 52 mg/litre for 72 days, prawn (Penaeus indicus), crab (Tylodiplax blephariskios), shrimp (Palaemon pacificus) and mullet (Mugil cephalus) had whole-animal (ash) concentrations of 3248, 1414, 3116 and 7743 mg fluoride/kg, respectively (Hemens & Warwick, 1972). The fluoride was accumulated from the water and not via ingested plant food. No differences in total-body fluoride levels were reported in the same species when exposed to 5.7 or 5.9 mg/litre for either 68 or 113 days (Hemens et al., 1975). Fluoride was accumulated mainly in the calcified skeletal structures of fish and crustaceans, with higher accumulation rates reported during the early stages of growth in fish and during periods of deposition of new skeletal material (ecdysis) in crustaceans.

Wright (1977) exposed brown trout (Salmo trutta) fry to inorganic fluoride (5, 10 or 20 mg/litre) for 200 h (12 °C; pH 6.8; hardness 73 mg/litre). Fry accumulated 10, 18 and 30 mg fluoride/kg at the three fluoride concentrations, respectively.

Terrestrial plants may accumulate inorganic fluorides following airborne deposition and uptake from the soil (Davison, 1983). Sloof et al. (1989) reported that the main route of uptake of fluoride by plants is from aerial deposition on the plant surface. Plant uptake from soil is generally low (except for accumulators) unless the fluoride has been added suddenly, such as following amendment with sludge or phosphate fertilizer. The availability to plants tends to decrease with time following application of fluoride. The degree of accumulation depends on several factors, including soil type and, most prominently, pH. At acidic pH (below pH 5.5), fluoride becomes more phytoavailable through complexation with soluble aluminium fluoride species, which are themselves taken up by plants or increase the potential for the fluoride ion to be taken up by the plant (Stevens et al., 1997). Plant uptake of fluoride from solution culture is dependent on plant species and positively related to the ionic strength of the growth solution. Once a threshold fluoride ion activity in nutrient solutions is reached, fluoride concentrations in plants increase rapidly (Stevens et al., 1998a).

The uptake of fluoride by ryegrass (Lolium multiflorum) grown in soil amended with a fluoride-rich sludge increased with increasing fluoride application (84–672 kg/ha) (Davis, 1980). Fluoride concentrations in the plant tissue exceeded the 30 mg/kg dry matter threshold of toxicity to cattle in the first cuts of grass exposed to 300 kg sludge/ha or greater. Concentrations in the second cut of grass exceeded the threshold value at 672 kg sludge/ha, and the third cut of grass was below the threshold at all concentrations tested.

The amount of fluoride taken up from soil by sunflower seedlings (Helianthus annuus) was proportional to the fluoride concentration of the solution added to the soil (10–100 µg/ml). Highest concentrations were found in the roots, with reduced concentrations in successive leaf ranks (Cooke et al., 1978). Similarly, concentrations of fluoride were high in roots and low in shoots in tomatoes (Lycopersicon esculentum) and oats (Avena sativa) grown on fluoride-enriched media (Stevens et al., 2000).

Reynolds & Laurence (1990) found that exposure of kidney bean (Phaseolus vulgaris) plants to intermittent high levels of atmospheric hydrogen fluoride (3 µg fluoride/m³ for 5 consecutive days or 5 µg fluoride/m³ for 3 consecutive days) caused higher accumulation in leaves than did exposure to a continuous dose at a lower level (1 µg fluoride/m³ for 15 consecutive days), even though the plants received the same total dose. They also noted that the rate of uptake from the air over a 15-day period, particularly for the highest dose (5 µg fluoride/m³), was most significant at first exposure and declined during subsequent exposures.

Walton (1987a) demonstrated that untreated earthworms (Lumbricus terrestris) from which gut soil had been removed (depurated) contained only 6% of the amount of fluoride reported in whole worms. Worms kept in soil were exposed to a sodium fluoride concentration of 1000 mg/litre for 30 days. The fluoride content of the worms was established either at the end of exposure or following an 8-day starvation period. The fluoride concentration in treated worms with their gut contents washed out was more than 6 times greater than the fluoride concentration of untreated worms. Even after an 8-day starvation period, the concentration in exposed worms was 4 times greater than the concentration in the unexposed worms. The fluoride concentration in the worms was always lower than the concentration in the soil.

The deposition of fluoride in the bones of white leghorn (Gallus gallus) chicks exposed to 252 mg fluoride/kg in the diet from hatching for 3–10 weeks was 3–7 times greater than for birds fed a basal diet. Fluoride was deposited in the bones of exposed females following sexual maturity and was shown to be associated with physiological factors associated with egg production, since birds with restricted ovulation had fluoride levels similar to those of males (Michel et al., 1984).

The bioavailability of fluoride to sheep (Ovis aries) from plants such as sea aster (Aster tripolium), sand couch (Elymus pycnanthus), fescue (Festuca rubra) and common salt-marsh grass (Puccinellia maritima) grown in an industrial area was calculated to range from 65 to 90% (Baars et al., 1987).

Dairy calves (Bos sp.) were fed on continuous (1.5 mg/kg body weight), periodic (1.5 mg/kg body weight for 6 months; control diet for 6 months) and alternate (3 mg/kg body weight for 4 months; 0.75 mg/kg body weight for 8 months) diets for 6 years. Vertebral fluoride levels were 550 mg/kg ash weight for controls and 15 000, 7700 and 14 000 mg/kg for the three treatment groups, respectively (Suttie, 1983).

Suttie et al. (1985) maintained white-tailed deer (Odocoileus virginianus) on diets containing fluoride at 25 and 50 mg/kg for 1 year. Fluoride levels in antlers ranged from 3000 to 5800 mg/kg ash weight and in vertebrae from 4200 to 7400 mg/kg.

5.1.1 Surface water

Fluoride levels in surface waters vary according to geographical location and proximity to emission sources. Surface water concentrations generally range from 0.01 to 0.3 mg/litre (ATSDR, 1993). Ambient surface water fluoride levels are summarized in Table 1. Seawater contains more fluoride than fresh water, with concentrations ranging from 1.2 to 1.5 mg/litre.