Fluoride Action Network

The Associate Committee on Scientific Criteria for Environmental Quality was established by the National Research Council of Canada in response to a mandate provided by the Federal Government to develop scientific guidelines for defining the quality of the environment. The concern of the NRC Associate Committee is strictly with scientific criteria. Pollution standards and objectives are the responsibility of the regulatory authorities and are set for the purpose of pollution control. These may be based on scientific criteria starting point but they also take into account the optimal socioeconomic impact of proposed measures as well as the state of existing technology.

The Associate Committee’s program includes the evaluation of available information on the probability of effects of contaminants on receptors together with the related fundamental principles and scientific knowledge. In this work particular attention is directed to receptors and contaminants (and their interactions) important to Canada. This Canadian approach is necessary because evaluations made in other countries or regions will not always be applicable to the particular circumstances prevailing in Canada.

Members of the Associate Committee, its Subcommittees and Expert Panels, serve voluntarily and are selected for their individual competence and relevant experience with due consideration for a balance among all sectors in Canada. Responsibility for the quality of study documents rests with the Associate Committee. Each report is carefully reviewed according to a multi-stage procedure established and monitored by the National Research Council of Canada in order to preserve objectivity in presentation of the scientific knowledge. Publication and distribution of the report are undertaken only after completion of this review process.

Comments on Associate Committee documents are welcome and will be carefully reviewed by the Expert Panels. It is foreseen that these scientific criteria may be revised from time to time, as new knowledge becomes available.

All documents published by the Associate Committee are published in both French and English.


This report was requested by the Management Subcommittee of NRC’s Associate Committee on Scientific Criteria for Environmental Quality. Dr. Dyson Rose (retired), formerly of the National Research Council’s Division of Biological Sciences, undertook the task of preparing this report, with assistance from J.R. Marier of NRC’s Environmental Secretariat

The report emphasizes Cause/Effect interrelations of environmental fluoride, and also attempts to identify deficiencies in the current scientific knowledge. The compilation covers the scientific literature that came to the authors’ attention prior to June 30, 1977.

The report has been reviewed by the members of the Management Subcommittee of NRC’s ACSCEQ, and by the following individuals:

Dr. J. Franke, Orthopedics Clinic, Martin Luther University, Halle, Wittenburg, DDR;

Drs. C.C. Gordon and P.C. Tourangeau, Environmental Studies Laboratory, University of Montana, Missoula, U.S.A.;

Dr. E. Groth, Environmental Studies Board, National Research Council, Washington, D.C., U.S.A.;

Dr. R.J. Hall, Analytical Chemistry Department, U.K. Ministry of Agriculture, Fisheries, and Foods, Newcastle-upon-Tyne, England;

Dr. S.S. Sidhu, Newfoundland Forest Research Centre, Canadian Forestry Service, Environment Canada, St. John’s, Newfoundland, Canada.

The authors wish to express their thanks to the members of the Management Subcommittee, and to the other reviewers, for the valuable comments received. However, we must emphasize that the viewpoints expressed in this report represent our own assessment of the environmental fluoride situation.

The authors also wish to express their gratitude to Miss Lynda Boucher and Miss Pat Moss, for their sustained cooperation in typing this report.

Dyson Rose and J.R. Marier
October 4, 1977






1.1.1 Atmospheric Emissions
1.1.2 Aqueous Discharges
1.1.3 Solid Wastes


1.2.1 Airborne Fluoride
1.2.2 Water-borne Fluoride



2.1.1 Aquatic Vegetation
2.1.2 Terrestrial Vegetation Ecological Effects Fluoride-Induced Effects on Agricultural and Forest Crops
2.1.3 Criteria for Crop Injury


2.2.1 Aquatic Species
2.2.2 Insects
2.2.3 Wildlife
2.2.4 Livestock



3.1.1 Fluoride Content of Blood
3.1.2 Effect of Fluoride on Blood Components


3.2.1 Fluoride Content of Urine
3.2.2 Effect of Fluoride on Urine Components


3.4 BONE

3.4.1 Fluoride Content of Bone
3.4.2 Fluoride Induced Changes in Bone






5.1.1 Intake From Foods and Beverages
5.1.2 Intake From Air


5.9.1 Criteria Based on Bone Fluoride and Plasma F-
5.9.2 Assessment of Fluoride Intake From Air




1 – Total Fluoride Emissions to the Atmosphere by Canadian Industrial Sources in 1972
2 – Estimated Soluble Fluoride Emission Rates and Total Emissions for U.S. Industries, 1968 or 1970 Data
3 – Fluoride Emission Rates Collected from Various Reports on the Aluminum Industry
4 – Comparison of Fluoride Emission Rates in the Primary Aluminum Industry in Canada and the U.S.
5 – Fluoride Emissions from Phosphate Fertilizer Plants
6 – Volumes and Fluoride Contents of Some Industrial Waste Waters
7 – Fluoride Content of Water from the East Gallatin River, Montana
8 – Influence of Domestic and Industrial Sewage on the Fluoride Content of Rhine and Ham River Water
9 – Regression Equations Relating Airborne Fluoride Concentration to Plant Response
10 – Fluoride Content of Insects from Polluted and NonPolluted Areas of Montana
11 – Fluoride Content of Bones of Animals Collected in Non-polluted Areas of Montana
12 – Fluoride Levels in the Bones of Wild Birds from Non-polluted Areas
13 – Regression Equations Relating Plasma Ionic Fluoride Levels to Age in Adult Humans
14 – Mean Plasma Ionic Fluoride Levels for Humans Residing in Non-fluoridated and Fluoridated Communities
15 – Effect of Fluoride on the Levels of Various Blood Components in Experimental Animals
16 – Effect of Fluoride on the Levels of Various Blood Components in Humans
17 – Effect of Fluoride on Levels of Metabolites in, and Physiological Activities of, Animal Soft-Tissue Organs
18 – Effect of Fluoride on Some Physical Parameters of Animal Bones
19 – Recent Data Illustrating the Effects of Environmental Factors on the Range of Fluoride in Some Foods
20 – Recent Data on the Daily Intake of Fluoride by Children
21 – Recent Data on the Daily Intake of Fluoride by Adults
22 – The Percentage Contribution of Water and of Various Foods to the Fluoride Ingested by Humans
23 – Health Problems Among Residents Near Fluoride-Emitting Sources
24 – Symptoms Common to Both Fluoride Intoxication and Magnesium Deficiency
25 – Fluorosis in Persons Who Have the Diabetes Insipidus Syndrome


1 – An illustration of the atmospheric distribution of gaseous fluoride, in relation to elevation and distance from an industrial point source. (see figure)
2 – Influence of airborne gaseous fluoride on the yield of beans, strawberries, and oranges, as plotted from the data of several authors. (see figure)
3 – Influence of dietary fluoride on the weight-gain of young swine. (see figure)
4 – Interrelation between rib bone fluoride content, blood plasma F-, and fluoride intake from three daily meals. Data are for 55-year-old lifetime residents. (see figure)


“Environmental Fluoride” (Marier and Rose 1971) was largely completed before the National Research Council, Canada, Associate Committee on Scientific Criteria for Environmental Quality had become operational. The document thus differs somewhat in format from later Associate Committees’ documents. The relevant Subcommittee therefore requested another document on this topic.

Comprehensive reviews on fluoride were published by the World Health Organization (WHO 1970), by the U.S. National Academy of Sciences (NAS 1971), and as “a non-experimental dissertation on a topic dealing with political aspects of public policy-making on scientific issues” (Groth 1973). These three documents differ from one another in intent, but all agree on the need for further research on the effect of environmental fluoride. Thus the WHO (1970) report states:

“Little is known about the in vivo effects of fluoride at the low levels occurring naturally in body-fluids and soft tissues on enzymes and the various facets of general metabolism in the living organism…”

“However, the indices of early intoxication are poorly defined and this has resulted in an element of speculation and confusion about the toxic potentialities of the fluoride ion”.

Similar statements emphasizing the lack of precise knowledge are found elsewhere in the document.

Similarily, the National Academy of Sciences document (NAS 1971) states:

“The available information is insufficient in depth and scope to allow unequivocal statements to be made about the effects on plants of fluoride at low atmospheric concentrations. One reason for the lack of information is the paucity of experiments designed to relate air quality to effects on plants. A second is the lack of sufficient ambient-air monitoring in connection with field studies and surveys, due in part to the lack of accurate and precise methods for the separation and collection of particulate and gaseous fluorine compounds. A third reason is the inadequacy of present experimental techniques for long-term studies in which field conditions can be simulated”.

“Unfortunately, many studies for a better evaluation of the effects of airborne fluoride on human health remain to be done. Not many authors have investigated the incidence and magnitude of effects on the thyroid gland, the hematopoietic system, the cardiovascular system, and the central nervous system. However, these systems respond readily to a number of stresses, not only to fluoride, and a causal relation to airborne fluoride has been established only poorly or not at all. More careful studies are required, with better attention being paid to the nature of the responses, the presence or absence of other medical or physical conditions that might contribute to the occurrence of the responses, and the proper control groups”.

“The airborne fluorides to which subjects are exposed must be better evaluated with respect to amounts of fluoride-containing material, proportions of gaseous and particulate fractions, chemical and physical properties (including particle size) of the particulate fraction, and meteorologic conditions in the surrounding community when resident populations are being studied”.

The third document (Groth 1973) presents the need for further research even more emphatically. Thus

“…there have been very few studies of potential non-lethal effects of chronic accumulation of fluoride on populations exposed to lifetime ingestion”.

“Amounts of fluoride ingested by average adults are sufficient to produce chemical and structural changes in the mineral of the bones, and the long-term health significance of these changes is not known”.

“In short, there are a great many unanswered questions in regard to long-term potential adverse effects of fluoridation, and a number of indications of potential harm which have not been shown yet to be unfounded. In view of the seriousness of some of the possible consequences if fluoridated water is in fact harmful to a fraction of the population, extensive, continuing research would seem imperative. However, there are no ongoing large-scale efforts being made to carry out such research”.

During the seven year period (1970 to 1977) covered in the present document, there has been a voluminous output of literature related to fluoride pollution and fluoride toxicity to plants, animals and man. This has increased our general knowledge of the multiple effects of chronic exposure to fluoride, and has confirmed and possibly augmented the difficulties attending attempts to relate quantitatively exposure and time factors to effect. Nevertheless, a prime purpose of the present review is to identify criteria (dose-response relations) that may assist in establishing limits of exposure. A second purpose is to identify areas where additional research is urgent.

In environmental studies, it is often necessary or convenient to investigate individual sources of fluoride and to focus on the level of fluoride acting through a particular pathway. For example, the pathway involving airborne fluoride, forages, and domestic cattle has been studied extensively. However, it is essential to remember that living organisms respond to the total fluoride impact from all sources: plants are affected by fluorides in soil, water, and air; animals by fluorides in their forages, feed supplements and water; and man by fluorides in his foods, beverages, drugs and prophylactic agents, cigarettes, and air. Therefore a comprehensive assessment of the cumulative impact of fluorides on man’s environment requires consideration of the total fluoride contributed by multiple sources.

A serious effort has been made to consider all papers published since 1970 that are relevant to environmental fluoride. Because of the voluminous literature on the dental aspects of fluoride and on the freon-ozone argument, these two areas have been intentionally left for others to summarize and develop criteria. Papers on fluoride therapy in humans have been included only because data on high-dose, short-term effects appear relevant to chronic exposure (low-dose, long-term) situations. Reports on pollution control technology are considered to be outside the scope of this review. Sampling and analytical methodology are discussed only in relation to the interpretation of environmental effects.

Undoubtedly we have overlooked valuable research papers particularly among those published in languages other than English and French; for this we apologize to the authors concerned. For conciseness and brevity, we have omitted specific reference to about half the papers we examined.



The sources of fluoride in man’s environment have been discussed by numerous authors (e.g. Marier and Rose 1971; NAS 1971; Bittel and Vaubert 1971; Prival and Fisher 1974; Bojic et al.1975). Sources of fluoride include natural sources such as volcanic gases, and soluble fluorides in the earth’s crust. However, the pre- ponderance of pollution problems have been caused by modern-day man-made sources which singly, or in combination, occasionally lead to the presence of harmful levels of fluoride compounds in air, water, food or forage. In this section, we present data on the amounts of fluoride discharged from major man-made sources, and attempt to indicate the extent of the geographical areas affected by the fluoride discharges.

Fluoride emission data from industrial sources are often circumscribed by industrial secrecy and by industries’ ability to have environmentally-relevant data classified as proprietary to the industry. Also, governments have sometimes been loath to release data gathered at public expense as well as those submitted by industry. The rationale often given for this secrecy is that it allows decisions to be made in the absence of public clamor and emotionalism. Less rationally, it also denies the public’s right to take part in decisions involving a balance of economic and environmental objectives. The secrecy situation in Great Britain has been discussed by Tinker (1972).

Industrial and governmental secrecy has been detrimental to Canadian efforts to develop criteria relating the concentration of pollutants to their effects. Thus the studies of LeBlanc and his students (1971, 1972) on the effects of air-borne fluorides on epiphytes and bryophytes could not be related to existing but secret data on fluoride concentrations in the air. Similarly, the author of a report on pollution in the Shawinigan and other areas of Quebec (Pellissier 1973) repeatedly comments on the non-availability of results from related air-monitoring programs. Sidhu and Roberts (1976) encountered a parallel situation in Newfoundland.

1.1.1 Atmospheric Emissions

In spite of the secrecy discussed above, some information on atmospheric fluoride emissions by industry has become available during recent years. Environment Canada (1976) published data on fluoride emissions to the atmosphere in Canada during 1972. A portion of the data is reproduced in Table 1 and shows that, with the exception of aluminum production, the fluoride emissions are preponderantly in gaseous form. The U.S. Environmental Protection Agency (EPA 1972) reported the corresponding U.S. data in considerable detail, and we present summarized data in Table 2.

Unfortunately, the data in Table 1 and 2 are not directly comparable. Fluoride emissions into the atmosphere occur in gaseous and particulate forms, and the particulates vary in solubility. The solubility of the particulate matter has a marked influence on its toxicity to plants and animals (NAS 1971). Thus, the “Total soluble fluoride emissions” as recorded in Table 2 are more directly relevant to environmental-impact criteria than are either the “total” or “percent gaseous” data of Table 1.

The primary aluminum reduction industry, which is the largest single-industry source of atmospheric fluoride pollution in Canada (Table 1), and the third largest in U.S., has been the subject of several studies. Data on the rates of fluoride emission (i.e. the amount of fluoride released to the atmosphere per unit of aluminum produced) are presented in Table 3. The low emission rates for recently constructed smelters are indicative of the progress being made in controlling atmospheric emissions by this industry.

An interesting comparison can be made between the emissions from U.S. primary smelters in 1970 and those of Canadian smelters in 1972 (Table 4). Effluent fluorides, (i.e. total fluoride at source, before passage through emission control units) per unit of aluminum produced, are similar for Canadian and U.S. reduction lines. However, the average amount of fluoride emitted to the atmosphere, per ton of aluminum produced, is markedly higher for Canadian than for U.S. smelters.

The steel industry, which is the major source of atmospheric fluorides in U.S. and third largest in Canada (Tables 1 and 2), does not appear to have been studied as intensively, regarding fluoride emissions, as the aluminum industry. In part, this is probably related to the presence of other pollutants besides fluoride in emissions from steel mills, and to the fact that attention has been primarily focussed on pollution by sulfur dioxide and particulate matter.

In relation to the phosphate industry, which is also a major source of fluoride emissions, Osag et al. (1976) have presented a comparison of “industry wide” and “best controlled” atmospheric emissions (Table 5). It is difficult to relate these data to the rate of emission (3.1 to 4.1 lb/ton of P205 equivalent) in Table 2, but it would appear that the data of Osag et al. refer only to specific steps in the process and not to overall emissions. They probably do not include emissions from the surface of gypsum ponds (King and Ferrell 1975).

Table 1. Total fluoride emissions to the atmosphere by Canadian industrial sources in 1972 (Environment Canada 1976)

Sector Total fluorides released (US tons) % of Canadian total % of gaseous fluoride in effluent
Primary aluminum production 8,852 56.6% 55
Phosphate fertilizer and elemental phosphorous plants 2,668 17.1% >96
Primary iron and steel production 2,418 15.5% 80-85
Miscellaneous sources  534 3.4%  70-75
Power generation  1,006  6.4%  >90
 Industrial and commercial  162  1.0%  >90
 TOTAL EMISSIONS  15,644  100.0




Table 2. Estimated soluble fluoride emission rates and totals for United States industries, 1968 or 1970 data (EPA 1972).

Total US tons
Reference page or table
0.99 lb/ton ore
p. 3-64, Table 3-23
Coal combustion for power
0.16 lb/ton coal
p. 3-131, p. 3-132
Phosphate rock processing

3.1 to 4.1

lb/ton P205 equiv

Table 3-46
Primary aluminum
8.1 lb/ton prod.
p. 3-21, Table 3-6
Heavy clay products
0.81 lb/ton prod
Table 3-87, p. 3-249
Hydrofluoric acid prod
4.1 lb/ton HF
Table 3-104
HF alkylation process
0.15 lb/bbl alkylate
Table 3-101
Expanded clay aggreg.
1.14 lb/ton aggreg.
Table 3-93
Glass manufacture
up to 17 lb/ton glass
p. 3-220, calc. from
Tables 3-73 and 3-75
Frit smelting
180 lb/ton CaF2
Table 3-81, p. 3-235
Cement manufacture
0.008 lb/ton cement
Table 3-97
Non-ferrous metals,CopperZincLead




p. 3-307p. 3-314p. 3-311
55 + 18
p. 3-321
Aluminum anoding
up to 668
p. 3-322


Table 3. Fluoride emission rates, in kg/metric ton, collected from various reports on the aluminum industry

Reported emissions rate,kg/metric tons
Sweden, newest installations
1.0 total F
Linberg 1971
U.S. new control technology
0.25 gaseous F
Rosano and Pilet 1971
0.64 solid F
OECD countries, actual emission
6.1 total F
OECD 1972
OECD, obtainable emissions
2.3 total F
4.1 soluble F
EPA 1972
U.S. best primary system
1.2-4.7 total F
Rush et al. 1973
Best primary & secondary system
0.8-2.0 total F
U.S., weighted average
5.1 total F
U.S., weighted average
2.1 gaseous F
Singmaster and Breyer 1973,
Table 7-1d
U.S. new construction
1.0 total F
EPA 1976



Table 4. Comparison of fluoride emission rates in the primary aluminum industry in Canada and the U.S.

Canada 1972

United States 1970

Aluminum production, metric tons
904,491 (1)
3,614,545 (2)
Effluent fluoride, pre-abatement, kg/metric


14.1 (3)
13.1 (4)





Fluoride atmospheric emissions, kg/metric


4.9 (5)
2.7 (6)




(1) Personal communication, Statistic
Canada.(2) Singmaster and Breyer 1973, Table 7.3.(3) Environment Canada 1976, p. 4.(4) Singmaster and Breyer 1973, Table 7. 1d, weighted average.(5) Calculated from data of Table 1 (this document), and total production.(6) Calculated from Singmaster and Breyer 1973, p. 7-12 totals.


Table 5. Fluoride emissions from phosphate fertilizer plants (Osag et al. 1976).

lb.F/U.S. ton of P2O5 Input

Fluoride Source
Best Controlled
Wet Process Phosphoric Acid
Superphosphoric Acid
Diammonium Phosphate
Triple Superphosphate
Granular Triple Superphosphate

In discussing the lesser sources of fluoride emissions shown
in Table 1, the Environment Canada (1976) report notes that “It is not
possible to rationalize” differences in the fluoride effluent data reported
by the Canadian “clay products” industry and by the U.S. Environmental
Protection Agency. Environment Canada’s estimates of possible fluoride emissions
by this source therefore vary from 274 to 2463 tons (249 to 2239 metric tons)
in 1972 (Environment Canada 1976, their Table 7). The lower figure was used
to calculate the “misc. sources” total shown in Table I .

Glass manufacturing firms in Canada are reported to have “almost totally
phased out by 1972” the use of fluorspar as a flux. Fluoride emissions
by this industry are therefore thought to be low, i.e. 5 tons (Environment Canada
1976, p. 14, 15).

The Environment Canada (1976) report on fluoride emissions by the petroleum
industry (hydrofluoric acid alkylation process) indicates an “HF consumption” of 0.3
to 0.8 lb HF/barrel of alkylate. Available information does not enable us to
relate “consumption” to emission. However, if we assume that emissions
occur at the same rate as in U.S. plants (Table 2), the estimated total 1972
emissions in Canada of “less than one ton” (Environment Canada 1976,
D. 19) indicates a Canadian production of alkylate of less than 37 barrels per
day. Data published by Energy, Mines and Resources of Canada (EMR 1973) indicate
that Canadian HF-alkylation capacity was 13,470 barrels per day in 1972, and
had increased to 24,620 barrels per day in 1975 (EMR 1976).

Data on fluoride added to the atmosphere by domestic burning
of coal
in Canada are not available, but the amounts are probably small
because of the extensive use of other fuels for domestic heating in Canada.
The potential impact of domestic fuel burning on fluoride pollution should be
considered if changes in fuel consumption patterns occur. Baum et al. (1972)
report that 34 to 72% of the fluoride in coal, which varied from 0.0025 to 0.039%
in the coals tested, was contained in the flue gases of an industrial type furnace.
We have been unable to locate similar data for domestic-type furnaces.

1.1.2 Aqueous Discharges

Data on the volumes and concentrations of fluoride wastes being discharged to
rivers, lakes and oceans are not plentiful. All wet-scrubbing systems for control
of atmospheric emissions probably contribute some fluoride to the aqueous discharge,
but economic factors often favor recovery of fluoride from the scrubbers (e.g.
as precipitated calcium fluoride) and re-use of the water. Effluents and overflows
from limed settling-ponds contribute fluoride to the aqueous environment. General
discussions of problems related to pollution of waterways have been published
by McCaull (1972) and Cheremisinoff and Habib (1973).

Recent data on the volumes and fluoride contents of industrial waste waters
(Table 6) make it evident that large quantities of fluoride are being discharged
to waterways. For example, it can be calculated that if all North American plants
discharge fluoride at the rate (14 kg/metric ton) reported by Teworte (1972),
the total discharge by the aluminum industry would exceed 63,000 metric tons,
or about 4-fold the amount discharged into the atmosphere.

The production of uranium tetra- and hexa-fluorides involves the discharge of
significant quantities (625 to 1134 tons per year in U.S.) of hydrofluoric acid
by way of aqueous sewage (EPA 1972).

Rak (1969) presented data on the discharge of fluoride in waste waters during
production of some inorganic fluoride compounds. The reported discharges ranged
from 5.7 kg per metric ton of product for aluminum fluoride, to 55 kg/metric
ton of product for cryolite.

Pettyjohn (1975) has reported environmental damage caused by an unsuitable aqueous
disposal method applied to steel industry “pickling wastes”.

Table 6. Volumes and fluoride contents of some industrial waste waters.

Waste water

Industry and location
F-content ppm
Aluminum, Germany
200,000 litres per metric
ton A1
Teworte 1972
Phosphate fertilizer, U.S.
400 gpm (=90,800 1/hr)
Cheremisinoff and Habib 1973
Phosphate fertilizer, India
13,240 1/hr
Arora and Chattopadhya 1974
Stainless steel, U.K.
Jenkins 1972
Steek, U.S.
0.17 kg/metric ton of product
McCaull 1972


1.1.3 Solid Wastes

Information on the disposal of solid wastes containing fluoride has not been found
in any of the papers reviewed in the preparation of this document. Presumably,
large quantities are used as landfill or buried (Williams 1975) and, since this
practice is considered to be nonpolluting, the quantities involved are rarely
reported. However, Stepanek et al. (1972) have reported contamination of surface
and groundwaters by fluoride from solid wastes.

Williams (1975) has given a brief report on solid wastes from the aluminum industry;
individual smelters are reported to produce from 15 to 30 kg of calcium fluoride
sludge per metric ton of aluminum produced (30 to 60 lb/ton).

The disposal of high-fluoride solid wastes from the reprocessing of nuclear fuels
has been studied by Emma et al. (1968) and by Fitzgerald et al. (1969). Combined
chemical treatments to reduce fluoride volatility, along with sintering or canning,
appear to be prerequisites to safe longterm disposal of these wastes.

Polluted soil can also be considered as a form of solid waste. For land-locked
factories, all of the air-borne emissions discussed above can be considered as
eventual soil pollutants, except for the portion that is carried to rivers and
lakes by run-off. This amounts to about 18,000 tons per year in North America
(Tables 1 and 2).

Soils can also become contaminated with fluoride when fertilizers containing fluoride
are used. The fluoride content of fertilizers varies widely (Ammerman 1974) depending
on the method of processing and on the fluoride content of the phosphate raw material
used (Forster 1969). Ammerman (1974) reported the following fluoride concentrations:

Dicalcium phosphate 0.14%

Triple superphosphate 1.87%

Diammonium phosphate 2.00%

Gordon (1970b) lists fluoride contents ranging from 0.58
to 2.34% for fertilizers sold in Montana.


Reviews on fluoride and fluoride effects (WHO 1970; NAS 1971) usually stress
that “fluoride is well-nigh ubiquitous: detectable traces occur in almost
all substances” (Hodge and Smith 1977). This can be said about a great
number of pollutants; nevertheless, this fact is relevant to a discussion of
fluoride for two reasons: (1) it emphasizes the need to consider total fluoride
from all sources when investigating fluoride injury to plants, animals and man;
and (2) it often makes estimation of the role played by industrial fluoride
pollution more difficult. Manmade fluoride pollution nearly always arises from
a small geographic area or point-source and is detectable above the natural
or background fluoride over a definable area. Assessment of the distribution
and extent of these man-made fluoride anomalies is considered in this section.
Attention will, of course, be focussed on soluble fluoride as this is the most
environmentally-relevant form (cf. p. 10).


1.2.1 Airborne Fluoride

The presence of fluoride in rainwater collected in areas remote from human settlements
(Carpenter 1969) suggests that air which has not been contaminated by human
activity does contain some fluoride. However, ambient-air fluoride is usually
below the level of detection, which can be roughly defined as less than 0.05
ug F/m3 air (Thompson et al. 1971). Natural phenomena such as dust storms and
forest fires can contribute small amounts of soluble fluoride to the atmosphere.
Volcanic activity can contribute larger amounts. However, except for unusual
circumstances (e.g. volcanic activity), all soluble fluoride found in the atmosphere
in excess of 0.05 ug/m3 can be assumed to have originated from man-made sources.

From the above discussion, it miqht be concluded that the distribution and extent
of abnormal fluoride concentrations arising from point-sources would be relatively
easy to monitor. Unfortunately, however, man’s activities are so widespread
that background levels exceeding 0.05 ug/m3 are not rare, even in rural areas
of industrialized countries. The spread of pollution from a major source often
must be determined against a somewhat variable background level arising from
multiple minor sources (e.g. domestic coal burning) and from distant major sources
(Fischer and Brantner 1972). Thompson et al. (1971) reported data on 9,175 air
samples collected in various non-industrial urban sites and 2,164 samples from
non-urban sites. The distributions, as percentages found within the limits (ug/m3)
shown, were: urban = 88% < 0.05; 12% between 0.05 and 1.0; 0.2% > 1.0;
non-urban 98.5% < 0.05; 1.5% between 0.05 and 1.0; 0.14% > 1.0. Davison
et al. (1973) reported that only a small percentage of urban air samples from
Northumberland contained < 0.05 ug F/m3, and that the average fluoride level
was 0.28 pg/m3. On the other hand, “most” air samples from rural sites
contained < 0.05 ug F/m3 even though 19% of the samples exceeded the 0.1
ug/m3 level.

Data which have become available since 1970 confirm the presence of abnormally
high airborne fluoride concentrations in association with many of the industries
for which fluoride emissions are shown in Table 2. Peak fluoride concentrations
within these high-fluoride zones are rarely available, because they occur over
company-owned land. In a study of the effectiveness of potroom ventilization,
Sutter (1973) recorded mean daily atmospheric concentrations (aluminum industry)
of 540 to 3700 ug F/m3. In a study of fluoride emissions from an openhearth
(steel smelter) furnace with an electrostatic precipitator, Brown et al. (1971)
presented the data shown in Fig. 1. These data are no longer representative of this particular smelter, because the operating
procedure has been changed (Schuldt 1977). They do, however, indicate the high
concentrations and the atmospheric stratification that can occur within a few
hundred feet of a point-source of fluoride emissions. The stratification was
still apparent 12,000 feet (3.6 km) from the source (Fig.
). These data for airborne fluoride concentrations in areas surrounding fluoride-emitting

factories have been presented in numerous reports. These include data gathered
by static and dynamic air sampling devices (IJC 1971; Linzon 1971; Bourbon et
al. 1971) and by analysis of vegetation (Linzon 1971; Gilbert 1971; Carlson
1972; Gordon 1970a, 1976; Keller 1975; Jacobson and Weinstein 1977; Sidhu 1977a).

The studies of C.C. Gordon and his co-workers at the University of Montana (Gordon
and Tourangeau 1977; Tourangeau et al. 1977) are particularly important because
of their contribution to our knowledge of “shielding” effects. These
studies clearly demonstrate that vegetation tends to impede or intercept fluoride
in air that is moving through the foliage, thus creating an adjacent down-wind
area of lower airborne fluoride concentration. (Little or no effect of this
sort was observed with sulfur dioxide). The effect is so marked with airborne
fluoride that samples of needles taken from the upper, windward side of a pine
tree exposed to atmospheric fluoride will consistently contain 2- to 4-fold
more fluoride than found in equal-age needles from the lower, lee side of the
same tree. The effect becomes even more marked when windward and leeward sides
of a small grove are compared; also, groundcover vegetation under a stand of
pines may contain little fluoride, even in areas that are obviously polluted.
Terrain elevations that allow unimpeded impact by airborne fluoride result in
an increased amount of fluoride in exposed vegetation (Note also the stratification
effect illustrated in Fig. 1).

Sidhu (1977b) has similarly observed the effects of terrain elevation and shielding
in a fluoride-polluted Canadian coniferous forest. However, to ensure consistency
of sampling, he recommends collection of foliage samples from the windward side
of the mid-crown, “because defoliation occurred in the upper crown”.
In a study of the fluoride content of lichens, Gilbert (1971) observed that
even a boulder provided some shielding from fluoride carried by prevailing winds.

These observations make the siting of air-sampling devices and the collecting-points
for vegetation extremely critical. Gordon and Tourangeau (1977) recommend that
the sites for air-sampling devices for Maryland
be “in the middle of open fields, … one to two feet (0.3
to 0.6 m) above the height of corn crops and away from stands of hard woods
which impede or intercept the fluoride-polluted winds”. Samples of agricultural
crops should be taken from parts of the field that are 50 ft (15 m) or more
from hedgerows or other vegetation that is taller than the crops. In non-agricultural
areas, sampling should be from near the top of windward slopes, at a height
sufficient to be clear of any screening by vegetation.

The fluoride content of vegetation varies with the plant species and variety,
and with the stage of development (Chang 1975; Weinstein 1977). It is also influenced
by the plant tissue sampled (leaf, fruit, etc.), the age of individual leaves
or needles (Chang 1975; Gordon 1976), the location of sampled foliage on the
plant (Gordon 1976), and the season (Harris 1974). All these factors must be
considered when sampling vegetation as a means of monitoring fluoride in air.
[See Guderian and Schoenbeck (1971), Teulon (1971), and Sidhu (1977a) for a
discussion of other aspects of the methodology.] Uptake from the soil must also
be considered (Weinstein 1977).

When the above factors are taken into consideration in the planning of a study,
reliable data on the extent, concentration and distribution of a man-made atmospheric
fluoride anomaly can be determined with reasonable accuracy, even against an
urban fluoride background. These factors are influenced by pollution loading,
wind velocity and constancy, other meteorological conditions, and geographic
factors. Some examples of airborne fluoride discharges are given herein. Preference
has been given to Canadian data.

Gilbert (1971) studied fluoride levels around a small (20,000 tons per year)
aluminum smelter in Scotland. On the basis of an average rate of emission for
smelters in O.E.C.D. countries of 6.1 kg F per metric ton of aluminum (OECD
1972), the total discharge would have been only 123 metric tons (135 short tons)
per year of total fluoride. The smelter was surrounded by a “bryophyte
desert” about 0.5 mile (0.8 km) wide and extending about 1 mile (1.6 km)
downwind and 0.7 mile (1.1 km) upwind. This, in turn, was surrounded by a further
area of damage, and elevated fluoride levels in vegetation were observed 4.3
miles (6.9 km) downwind.

LeBlanc and co-workers (1971, 1972, 1975) studied epiphytes in the proximity
of a Canadian aluminum smelter with an unspecified (proprietary company data)
amount of atmospheric fluoride discharge. The area of vegetative disturbance,
as indicated by an “Index of Atmospheric Purity” based on species
frequencies, extended 10 km (6.2 miles) downwind.

Carlson and Dewey (1971), Carlson (1972), and Harris (1974) have reported extensively
on the distribution of atmospheric fluoride discharged by an Anaconda aluminum
smelter in Flathead County, Montana. In spite of assurances by the company that
vegetation damage would not occur (Burk 1972), this smelter had a 10-year history
of causing foliage injury in the surrounding territory. Nevertheless, the smelter
capacity was greatly expanded between 1965 and 1970. By 1970, foliar material
from various species contained fluoride levels in excess of background values
(i.e. greater than 10 ppm, dry weight basis) over a 213,760 acre (86,570 hectare)
area. Extensive injury, and foliar fluoride concentrations above 30 ppm, were
observed over a 69,120 acre (27,994 ha) area. During 1970, this Anaconda plant
installed fluoride emission control equipment that reportedly reduced emissions
from 7,500 to 2,500 lb (3,410 to 1,136 kg) per day. A subsequent survey showed
above-normal (> 10 ppm) fluoride in 1971 foliage over an area of 179,200
acres (72,575 ha) along with serious injury and > 30 ppm fluoride in foliage
over 15,200 acres (6,156 ha).

Sidhu and Roberts (1976) reported damage and high foliar fluoride concentrations
in the vicinity of a Canadian phosphorus plant. The total area affected was
11,434 ha (28,242 acres), but fluoride emission data were “confidential
to the industry”. However, in a subsequent paper, Sidhu (1977a) reported
airborne fluoride concentrations ranging from 0.8 to 20.8 ug/m3 at a 3 0.7 km
distance from this factory, and concentrations of 0.06 to 0.34 ug/m at 18.7

Preliminary data have also become available concerning fluoride distribution
around an aluminum reduction site at Kitimat B.C. (Gordon 1976). At a production
rate of 250,000 tons aluminum per year, and a reported emission rate of 5 to
7 lb F/ton Al, total fluoride emissions are estimated at 625 to 875 tons (568
to 795 metric tons) per year or 3,425 to 4,795 lb/day (1,556 to 2,180 kg/day).
The data available are insufficient to define the totality of the area affected
by these emissions, but “a twenty-plus square mile ‘death band’ of dead
timber trees” (5,180 ha) is reported. Foliage collected from coniferous
trees 5, 10, 11, and 20 miles north of the smelter contained higher fluoride
levels than Gordon had observed at these distances around other aluminum plants.

Fischer and Brantner (1972) studied the fluoride content of beech (Fagus
) leaves in Austrian urban areas of heavy and moderate air pollution,
and in open country. Fluoride levels of less than 10 ppm were common in leaves
from unpolluted areas. Fluoride levels in leaves from urban areas were up to
47 ppm. Even in wooded areas outside the city limits, fluoride levels well above
10 ppm were encountered at “fronts of collision which were caused by the
particular meterological conditions” in the area.

1.2.2 Water-borne Fluoride

The “average dissolved fluoride content of the major rivers of the world
is fairly well defined at 0.01 to 0.02 ppm” (Carpenter 1969). Atmospheric
dusts are thought to be the major sources of this “background” fluoride,
although the source of a large portion of the fluoride-containing atmospheric
dust is a subject of some dispute (Carpenter 1969, Bressan et al. 1974). Leaching
of fluoride from rocks increases the fluoride content of ground waters, but
under the conditions observed by Jacks (1973), this source contributed little
fluoride to surface waters.

The contribution of domestic sewage from cities to the fluoride content of rivers
was studied by Masudo (1964). The amount of fluoride found in effluent sewage,
in excess of the amounts present in the cities’ water supplies, were as follows:

Raw sewage ( 4 cities) 1.30 mg/l

After primary treatment (23 cities) 1.28 mg/l

After secondary treatment (29 cities) 0.39 mg/l

Fluoride is considered to be a “difficult to treat”
industrial waste (Environment Canada 1975).

Soltero (1969) and Bahls (1973) reported fluoride concentrations in the East
Gallatin river (Table 7). These data show that elevated fluoride caused by sewage
discharge from the city of Bozeman was detectable for a distance of 4 km below
the sewage outlet.

Table 7. Fluoride content of water from East Gallatin River, Montana.
Fluoride content, mg/l
Sampling Location
Soltero 1969
Bahls 1973
Bahls 1973
Above sewer outlet
3.8 *
0.3 km below outlet
1.8 km below outlet
2.2 km below outlet
4.2 km below outlet
5.3 km below outlet
8.2 km below outlet
* Soltero (1969) reported the
data in meq/l; it appears probable that his data are too high by a factor
of 10.

A study of fluoride input into Narragansett Bay, Rhode Island
(reviewed by Groth 1975b) indicated that “36%
of the fluoride entering the Bay was due to fluoridation of water supplies in
five communities on rivers feeding into the estuary”.

Data on the fluoride content of the Rhine (Teworte 1972) and Ham (Lee and Whang
1972) rivers (Table 8) also indicate that both domestic and industrial sewage
contribute significantly to the total fluoride content. Seepage and leaching
from solid and liquid waste disposal sites can also cause serious pollution
of run-off and ground waters (Stepanek et al. 1972; Pettyjohn 1975).

Table 8. Influence of domestic and industrial sewage on the fluoride content of Rhine and Ham River water.

F-content, mg/l
Sampling sites
Lee and Whang 1972
Main Stream
City water reservoirs
Tributary water, residential areas
Tributary water, industrial areas
RHINE RIVER Teworte 1970
at Rheinfeld
below Al smelter
at Dutch border
0.30 to 0.35

The distribution of fluoride released into flowing bodies
of water such as rivers is usually detectable on the basis of differentials
between the fluoride content of samples taken above and below the known or suspected
source of pollution. However, lakes, bays, and inlets can present a more difficult
problem, although comparative analyses (i.e. in relation to input-sources) can
provide meaningful information on the degree and extent of a contaminated zone.
Ocean water has a nearly constant fluoride content of 1.35 to 1.4 mg total fluoride/litre,
(Carpenter 1969; Bewers 1971), and a fluoride-to-chloride ratio of 6.71 0.07
x 10 5:1 (Warner and Jones 1975). Theoretically, inflow of fluoride-contaminated
river water should be detectable as a change in the F:Cl ratio. However, if
an ion-specific electrode is used to determine fluoride in brackish or ocean
water, it is necessary to correct the observed fluoride ion activities for the
complexing effect of magnesium (Thompson 1967; Brewer et al. 1970).

Use of the F:Cl ratio has provided considerable information showing fluoride
pollution of estuaries and ocean-bays. For example, Kitano and Furukawa (1972)
determined the fluoride-to-chloride ratio, to estimate fluoride pollution in
Tokyo Bay. Fluoride concentrations in contaminated inflowing waters ranged from
0.15 to 1.07 mg/l, with F:Cl ratios of 1.4 x 10-4 to 3.6 X 10-2:1. Surface samples
from the bay contained from 0.63 to 1.28 mg F/kg water, and the F:Cl ratio varied
from normal (i.e. 6.71 X 10-5) up to 9.05 x 10-5:1. Values above 7.1 x 10-5
were encountered at 11 sampling points (mostly surface) in the western half
of the Bay, but not at sampling points in the eastern half which is influenced
by incoming seawater.

The distribution of waterborne fluoride discharged from the aluminum reduction
plant at Kitimat, B.C., has led to abnormally-high F:Cl ratios throughout the
surface waters of Kitimat Harbour (Harbo e,t al. 1974). Observed F.-Cl ratios
ranged from 13 to 1,500 x 10-5 (av. 158 X 10-5) and fluoride concentrations
ranged from 0.10 to 11.0 mg/l (av. 1.17). Occasional high F:Cl ratios were also
encountered in subsurface waters at depths from 10 to 100 m (av. 7.61 x 10-5;
range 6.64 to 15.0 X 10-5). Comparable samples taken from Howe Sound “where
input of non-natural fluoride is not known to occur” had F:Cl ratios ranging
from 7.8 to 66 X 10-5, av. 14.5 x 10-5 for surface samples; and from 6.55 to
7.42 x 10-5, av. 6.83 x 10-5 for subsurface samples. No investigation of the
factors causing high F:Cl ratios in surface waters of Howe Sound was reported.

An interesting but incomplete study of water-borne fluoride has been reported
for Tampa Bay, Fla. (Taft and Martin 1974).
In July 1973, a phosphate plant was discharging an estimated 24,000 lbs (10,900
kg) of fluorine daily, along with quantities of phosphate and nitrate, into
Tampa Bay. This resulted in deposition of solid calcium fluoride at the point
of discharge and for about 1,000 ft. (300 m) into the bay. The precipitate accounted
for only a small portion of the total fluoride in the discharge. Fluoride concentrations
in samples of surface water above the fluorite deposit varied between 16.3 and
36.5 ppm. No data were presented for a more extended area of the Bay. Nevertheless,
a severe thermal effect was generated at the fluorite:water interface, and this
caused a significant increase in the temperature of the surface water. The authors
also reported the absence of all living organisms in the afflicted area.

The effect of fluoride on aquatic life is discussed in Sections 2.1.1 and 2.2.1.



The sources of man-made fluoride pollution discussed in Section I result in
above-normal concentrations which impinge on terrestrial and aquatic flora and
fauna, and on man. The exposure of living organisms to above-normal concentrations
of fluoride, which induces fluoride accumulation by the organism, may result
in an alteration of the organism’s biochemistry and morphology. Directly or
indirectly, such changes may restrict the organism’s ability to maintain its
ecological position. In the plant kingdom, an example of this has been provided
by McLaughlin and Barnes (1975) who observed that fluoride accumulation in the
foliage of some pines and hardwoods reduced photosynthesis and stimulated dark
respiration, thus undoubtedly reducing the amount of carbohydrate available
for growth and seed production. In the animal kingdom, Gerdes et al. (1971b)
report that exposure of fruit flies to low levels of atmospheric fluoride significantly
reduced the fecundity and egg hatchability of the descendants who were not themselves
exposed to fluoride.

Some published data suggest that exposure to low levels of airborne fluoride
can stimulate the growth of some plants (cf. Weinstein 1977). Bennett et al.
(1974) suggest that a low level of fluoride and of ozone was the norm under
which plants evolved, and that in tests on the effects of exposure to fluoride
the “control” plants should not be grown in fluoride-free air. However,
growth that occurs as a result of fluoride stimulation is often abnormal (Weinstein
1977). Even the growth stimulation that resulted in an increased fresh weight
in bean plants (Pack 1971a) did not result in an increased yield of beans, and
the ripened beans produced by exposed plants developed less vigorous seedlings
than did beans from control plants (Pack 1971b). It is thus doubtful that the
apparent growth stimulation occasionally observed on exposure of plants to low
levels of atmospheric fluoride is of any evolutionary, ecological, or economic



2.1.1 Aquatic Vegetation

Available data on the responses of aquatic vegetation to fluoride pollution
have been briefly reviewed by Groth (1975a, b).
The data are insufficient to allow firm conclusions to be drawn, but do indicate
that levels as low as 2 ppm in water can decrease the growth of one species
of Chlorella. The data also show that many aquatic plants accumulate fluoride
to concentrations that may be many-fold higher than the external concentration.

Ishio and Makagawa (1971) report that Potphyxix tenaa, an alga, was killed
by a 4-hour laboratory fumigation with fluoride (1.8 ppm in head space of growth
chamber) and that the critical concentration appeared to be 0.9 ppm.

Kilham and Hecky (1973) have discussed possible ecological effects of relatively
high natural fluoride levels in African lakes.

The accumulation of fluoride by aquatic plants and plankton is of interest because
of its potential impact on animals that consume these organisms. In an unpolluted
area of New Zealand, Stewart et al. (1974) observed fluoride levels from 31
to 209 ppm in the shells of species feeding on plankton, and from 1,425 to 1,882
in the skeleton of Blue cod that feed on crabs, shrimp, shell-fish, etc. The
data suggest that, for the stages mentioned above, the food-chain concentration
factor is at least 10:1. As noted by Groth (1975b),
“we have very little knowledge of the sublethal effects of fluoride on
behaviour or reproductive processes, or of the potential accumulation of the
pollutant in aquatic foodchains. Yet such effects, should they occur, would
probably be more important ecologically than the mortality which might result
from very high, but short lived, pollution episodes”.

2.1.2 Terrestrial Vegetation

Dochinger (1971) dates the awareness of fluoride-induced damage caused in terrestrial
vegetation back to German reports of the 1880’s and states that “For the
last 30 years, the injury to agriculture by fluorine compounds has intensified
because of the expansion of industries ….” Bossavy (1971) has summarized
estimates of the damage occurring primarily to forests, in European countries.

Literature on the biochemical and morphological changes caused by exposure of
terrestrial vegetation to fluoride has been reviewed by McCune and Weinstein
(1971), Chang (1975) and Weinstein (1977). For consideration of some other aspects
of the effects of fluoride on plants, such as the uptake of fluoride from soils,
the influence of environmental factors on the uptake of airborne fluoride, etc.,
the reader is referred to Marier and Rose (1971), NAS (1971), Treshow (1971),
Miller and McBride (1975) and Weinstein (1977).

Terrestrial plants exposed to airborne fluoride frequently display foliar damage,
sometimes grow less vigorously, and almost invariably accumulate significant
amounts of fluoride in their foliage. These effects are all of aesthetic, economic
or environmental significance. The interrelations among, and criteria for, each
of these factors will therefore be considered with regard to airborne fluoride
concentrations. Ecological Effects

As noted above, it can be assumed that many of the fluoride-induced changes
occurring in vegetation will decrease the plant’s ability to maintain its ecological
. However, studies of the actual ecological effects have rarely
been undertaken. Coniferous trees seem to be the most seriously affected forest
species in many situations (Gordon 1976; Tourangeau et al. 1977; Carlson and
Dewey 1971). Moreover, fluoride-induced changes in relative species dominance
have been confirmed by Sidhu (1977b), who also commented that:

“Preliminary results of a recent long-term study of the effects of fluoride on forest vegetation in Newfoundland showed that the softwood tree canopy (balsam fir, black spruce, larch) was being replaced by undergrowth of hardwoods (white birch, American ash). As the mortality of the original tree cover (softwoods) continued, the shrub layer showed a significant increase in raspberry, skunkcurrant, calamagrostis, and fireweed, underneath the hardwood tree species. Hardwood species (which defoliate every year) tend to accumulate fluorides at higher concentrations and at a faster rate than the softwoods. Therefore, it is suspected that the change from soft- to hardwood tree cover will result in the addition of higher amounts of fluoride to the soil. Also, the wildlife of the area feeding on the hardwoods will experience fluoride toxicity within a shorter period and over larger affected areas.”

Although epiphytes and bryophytes are considered more tolerant
to fluoride than conifer species (Sidhu 1977b), alterations in species frequency
among these organisms have been observed (LeBlanc et al. 1971, 1972; Gilbert
1971). LeBlanc et al. (1972) report that a few species such as Frullania
, Lecanora impudens, and Physcia ciliata could
not be found within a 12 km (7.5 miles) distance from a fluoride source, although
they were prevalent in the surrounding territory. Even the species that were
able to maintain themselves close to the source were up to five times more plentiful
beyond the pollution zone. Gilbert (1971 also reported complete absence of a
Lecanora species in a fluoride-polluted area.

It should also be noted that if fluoride is injurious to pollinating insects
(see Section 2.2.2), this could result in an indirect, but potentially extensive,
effect on some ecological communities. Fluoride-induced Effects on Agricultural and Forest Crops

Information on fluoride-induced injury to vegetation,
which is often not directly applicable to the development of criteria, has appeared
in numerous reports. This information is nevertheless important for an overall
concept of fluoride phytotoxicity, and is therefore briefly reviewed.

Bennett and Hill (1973) exposed 4- to 8-week-old barley and alfalfa plants to
hydrogen fluoride in fumigation chambers for a single 2-hour period, and measured
carbon dioxide uptake as an index of photosynthesis. With this short exposure
period, 50 ppb HF were required “before clearly measurable inhibition of
the net carbon dioxide rates occurred”. The percentage inhibition of apparent
photosynthesis was linearly related to HF concentration throughout the range
from 0 to 250 ppb, with no evidence of a no-effect threshold within the accuracy
of the measurements. Poovaiah and Wiebe (1973) noted that fumigation of soybean
plants with low (15 to 20 ppb) concentrations of HF for 1 to 4 hours caused
stomatal closing, reduced transpiration, and increased leaf temperature. McLaughlin
and Barnes (1975) report that trees which had accumulated 10 to 60 ug fluoride
per g dry weight of leaf tissue (from sodium fluoride in an aqueous spray) had
a reduced rate of photosynthesis and an increased rate of dark respiration.

Bale and Hart (1973a, b) exposed seedling roots of barley (Hordeum vulgare)
to solutions containing 1 x 10-2, 1 x 10-4 and 1 x 10-6 M (190, 1.9 and 0.019
ppm) fluoride (as NaF or HF) for 12 to 72 hours and examined dividing cells
for chromosomal aberrations. They concluded that “it is clear that each
of the concentrations of sodium fluoride and hydrofluoric acid used in these
experiments is capable of inducing chromosomal abnormalities and of producing
mitotic inhibition in the meristematic region of roots”.

Pack (1971a, b) grew beans from seedling to maturity in the presence of 0.58
to 10.5 ug fluoride/m3 air, then grew a second generation, without exposure
to fluoride, from the seeds of these plants. Exposure of the first generation
to as little as 2.1 ug F/m3 caused the development of less vigorous second-generation
seedlings. Vins and Mrkva (1973) report a decrease of 30 to 70% in the diameter
growth of pine trees at pollution levels that caused no otherwise-visible injury.
These authors relate the decreased growth primarily to sulfur dioxide pollution,
but there is an interesting relation between increasing fluoride emissions between
1960 and 1967 (their Fig. 1) and the rapid decline in annual diameter increments
(their Fig. 5) during this same period. (NOTE: See also Carlson, C.E., and Hammar,
W.P. 1976. Impact of fluorides and insects on radial growth of lodgepole pine.
Proc. Montana Acad. Sci. 35: 39.)

Facteau et al. (1973) reported that the growth of pollen tubes in the styles
of cherry blossoms was decreased by fumigation with hydrogen fluoride either
before or after pollination. Pollen tube length, expressed as a percent of style
length 72 hours after fumigation, decreased linearly as a function of the product
of exposure-time and atmospheric fluoride level. A somewhat similar result was
obtained in studies with apricot flowers (Facteau and Rowe 1977). Fluoride-induced
reduction in pollen germination and tube growth has also been observed in tomato
and cucumber plants (Sulzbach and Pack 1972) while inhibited seed production
or fruiting has been reported, with soybean, bell-pepper, sweet corn, and cucumber
being more susceptible than pea, grain sorghum, or wheat (Pack and Sulzbach

Conover and Poole (1971) found that cuttings of Cordyline terminalis
var “Baby Doll”, a horticultural foliage plant, suffered serious (approaching
50%) leaf necrosis when set for rooting in water containing 0.5 ppm fluoride.

“Soft-suture” of peaches is the “best known example of fluoride
injury to fruit” (NAS 1971). Facteau and Rowe (1976) were able to induce
this injury in Elberta peaches by spraying the trees at weekly intervals with
0.025% ammonium fluoride solution.

Maclean et ae. (1976) conclude that hydrogen fluoride (0, 5.0 and 9.7 ug F/m3
for 7 days) was more phytotoxic to tomato plants grown in magnesium-deficient
media than to those grown in complete media. Similarly, Pack and Sulzback (1976)
have demonstrated how calcium nutrition can influence the response of plants
to airborne gaseous fluoride. Pilet and Bejaoui (1975) report that fluoride
added to the culture medium for Rubus hispidus tissues markedly reduced
oxygen absorption by the tissues, particularly in media deficient in calcium
and magnesium. Increased levels of calcium and magnesium had a protective action,
in that they lessened the degree of fluoride inhibition of oxygen absorption.

The concept that vegetation may be stressed by pollutants present at levels
that induce relatively minor injury, or even at levels that do not induce detectable
injury in normal healthy plants, requires further study. The importance of this
concept relates to the possible summing of stresses from various sources, with
the total stress inducing an injury that cannot be easily related to any one
cause. Evidence of such stress-induced injuries resulting from multiple causes
is difficult to establish experimentally, but the effect of magnesium deficiency
discussed above (MacLean et al. 1976) is probably a dual-stress phenomenon.
However, in foliage, there can also be an in situ effect of fluoride on magnesium.
In a study of air pollution, Garrec et al. (1977) observed that fluoride accumulation
led to a depletion in the magnesium content of pine needles; in addition, there
was a similar depletion in foliar manganese content.

Probably the most striking example of multi-stress effects is to be found in
studies of the relation between atmospheric pollution and insect
of forest species. The host-parasite relation is complex, and
as noted by Heagle (1973), the presence of an atmospheric pollutant may act
to either the advantage or disadvantage of the insect.

However, studies of forest species under field conditions have demonstrated
that the stress placed on trees by pollutants can increase the degree of infestation
by insects. Heagle (1973) states that “A common finding is that trees injured
and weakened by pollutants are more likely to be attacked by insects that normally
require weakened trees for successful reproduction”. Jensen (1975) notes
that “Some evidence has been provided that air-pollution stress can initiate
and/or aggravate insect infestation and microbial infection of woody plants”.
Hay (1975) states that “Insects and mites have been implicated as a stress
factor on trees being influenced by pollutant emissions”.

Most of the above statements have been made relative to pollutants in general,
but the work of Carlson et al. (1971, 1974) and of Carlson and Hammer (1974)
shows that atmospheric fluoride induces an insect-favouring stress in forest
trees. Failure to recognize this stress-factor led prior investigators to incorrectly
diagnose a combined fluoride injury and insect attack in the “death-band
area” near Kitimat, B.C. (Gordon 1976).

When pollutants act in combination, each exerts its own stress, and each can
influence a different metabolic function. Thus, the effects of exposure to sulfur
dioxide and gaseous fluoride mixtures induced additive effects on citrus species,
but may have induced greater-than-additive effects on Zea mays and Hordium
(Reinert et al. 1975).

In apricot orchards, trees that were stressed by competition from weeds showed
more leaf damage from airborne fluoride than did trees from well-tended plots
(Oelschlager and Moser 1969).

Fluoride-induced stresses undoubtedly affect vegetation in diverse ways, depending
on the species and conditions. One possible mode of action in coniferous trees
has been noted by Bligny et al. (1973) who report that exposure to fluoride
delayed the formation of epicuticular waxes on the lower surfaces of Abus
needles. This could increase water loss from the needles, and also
increase their susceptibility to invasion by parasitic organisms. Keller and
Schwager (1971) have attempted to relate fluoride-induced stress to an increased
activity of an enzyme (peroxidase). Yee-Meiler (1974) has shown an increased
phenolic content of Norway spruce subjected to “physiologischen Schadigungen”
by fluoride (0.257 mg F/dm2 per 30 days).

Fluoride concentrations expressed in ug/dm2 per day or month
refer to data collected by exposing lime-filter papers to ambient air for the
started time. Marier and Rose (1971), using the greenhouse data of Adams (1961),
suggested conversion by the equation:

Airborne F(ug/m3) = 0.006 x lime-paper (ug F/dm2 per month).

Israel (1974a) has compared results based on field trials,
and suggested the equation:

Airborne F(ug/m3) = 0.003 x lime-paper (ug F/dm2
per month).

Israel’s estimate of the accuracy of the conversion is +

The difference in conversion factors (0.006 vs 0.003) may
relate to differences in air velocity across the lime-paper (Israel 1974a).

More recent, Sidhu (1977a) has also conducted a field-study
intercomparison, and has proposed an equation that can be expressed as follows:

Airborne F(ug/m3) = 0.0076 x lime-paper (ug F/dm2 per

which yields values 2 1/2 times higher than Israel’s, but
only about 25% higher than the Adams equation proposed by Marier and Rose (1971).
Furthermore, Sidhu (1977a) concludes that “In the absence of a more reliable
and accurate regression equation, Adams’ equation can be used to convert the
fluoridation plate data to the ug F/m3.”

2.1.3 Criteria for Crop Injury

Regression equations calculated from data presented in recent papers and which
indicate mathematical relations between yield and airborne fluoride; or yield
and foliar fluoride; or foliar fluoride, airborne fluoride and exposure time,
are presented in Table 9. The yield vs airborne fluoride data are also presented
in Fig. 2. Because of the small number of data-points
available, some of these regressions do not achieve statistical significance.
However, taken as a group, they reveal a consistent pattern of increasingly
harmful effects with increasing exposure of vegetation to fluoride.

Significant correlations between airborne fluoride levels and foliar fluoride
concentrations (“C” and ‘T’ in Table 9) are difficult to attain under
field conditions. However, if close attention is paid to the location of sampling
sites and to the selection of foliage of uniform age from specific species and
varieties, such correlations can be achieved. The correlation coefficient for
Linzon’s (1971) data (Table 9) is 0.48. In controlled greenhouse studies, the
concentration of fluoride and the age of foliage are usually controllable factors;
under these conditions, the time of exposure can be included as a component
of the regression equation [data of McCune and Hitchcock (1971) and MacLean
and Schneider (1973) in Table 9].

The data on the yield of oranqes (Leonard and Graves 1970, 1972) shown in Table
9 and Fig. 2, were for trees exposed for 28 months in field shelters with low
ambient levels of fluoride pollution (i.e. 0.1 to 0.4 ug/m3). Data for beans
(Pack 1971a) were from a 70-day greenhouse study at high levels of airborne
fluoride. Both indicate a severe loss of yield with increasing fluoride levels,
i.e. approximately 19% per 0.1 ug/m3 for oranges, and 3% per ug/m 3 (approximately
1.2 ppb) for beans. The data on strawberries (Pack 1972) are from a 5-month
greenhouse study, and indicate about a 5% loss in weight of individual fruits
per ug/m3 increase in airborne fluoride. This loss in yield (fruit set does
not appear to have been affected) was accompanied by a statistically significant
decline in fruit quality, as indicated by the “development rating”
assigned by the original author.

Yield of oranges in field shelters was also related to the fluoride content
of the foliage (Leonard and Graves 1972), declining by about 5% for each 100
ppm increase in fluoride in 10-month-old leaves.

In field tests, Israel (1974b) observed a highly significant multiple regression
between the fluoride content of forage and of air-plus-soil. This was expressed

F(foliage) = 11.4 F(air) + 0.0085 F(soil),

where fluoride content of foliage and soil is expressed in ppm and of air in
ug/dm2 per day absorbed by lime-paper.

In a recent survey of a Newfoundland region in Canada, Sidhu (1977a) concluded
that “The safe levels of fluoride in air for forest species appear to be
between 0.17 to 0.23 ug/m3.” This is very close to the lower limit given
by the estimates of Marier and Rose (1971), i.e. “The average gaseous fluoride
level in ambient air should be below 0.4 ug/m3 and might have to be as low as
0.2 ug/m3.”


Table 9. Regression equations relating fluoride concentrations to plant response.

Plant studied Range of fluoride concentrations Regression equation (1) Note Reference
Citrus 0.14 – 0.45 ppb Y (%) = 99.7 – 176 F (2) Leonard and Graves 1970
Citrus Not stated Y = 381.91 – 1.3132 CY = 417.25 – 0.8797 C (3)(4) Leonard and Graves 1972
Pine Not stated BI = 0.06 + 0.1607 FBI = 0.93 + 0.0027 F (5)(6) Hortvedt 1971
Bean 2.2 – 13.9 ug/m3 C = 14 + 102 FY (%) = 102.2 – 3.45 F (7)(8) Pack 1971a
Orchard grassAlfalfa up to 11 ug/m3up to 11 ug/m3 C = 1.13 FT – 1.17C = 1.89 FT + 0.74 (9)(9) McCune and Hitchcock 1971
Vegetation 20 – 128 ug/m3 C = 8.50 + 0.314 F (10) Linzon 1971
Strawberry 0.55 – 10.4 ug/m3 Y (%) = 99.5 – 5.1 F (11) Pack 1972
Timothy and red clover mix 2.3 ug/m35.0 ug/m3 C = 2.555 + 4.120 FTC = 30.288 + 3.820 FT (12)(13) MacLean and Schneider 1973
(1) Y = yield, BI = Tip burn index, F = airborne fluoride
concentration, C = concentration of fluoride in foliage, T = time; all expressed
in units used by the original authors except where (%) indicates our calculations
as a percentage of the control value.(2) Data of author’s Table 4, average value for 6 varieties, converted to
% of control (Pot 6); “F” values from Table 1, “Mean”.(3) Author’s Figure 1, p. 158, 10-month old leaves.(4) Author’s text, p. 158, “old” leaves.(5) Author’s Figure 5, p. 300, 1-year old needles.(6) Author’s Figure 5, p. 300, 2-year old needles.(7) Data from author’s Table 1, 70-day exposure. We calculated a single
average “control” value of “F” for each variety.(8) Data from author’s Table 3, yields calculated as a percent of the individual
controls.(9) Author’s equations from p. 291.(10) Regression equation calculated by us from all complete data, except
for three with foliage levels above 200 ppm fluoride, dry weight basis,
in author’s Tables 1, 3, 5 and 6.(11) Data on “Weight per fruit” from author’s Table 3, calculated
as a percentage of the weight of fruit from control plants. Calculations
based on the author’s data suggest that the number of fruit set ranged from
399 to 558 for the control plants and from 348 to 576 for the treated plants
and was not significantly affected by fluoride concentration.(12) Author’s Table 1, “F” = 2.3 ug/m3.(13) Author’s Table 1, “F” = 5.0 ug/m3.


2.2.1 Aquatic Species

The ecological significance of the exposure of aquatic animals to fluoride has
been studied to a limited extent, but much more research is required before
broad conclusions can be drawn. The following paragraphs summarize the available
recent information.

A large number of species have been shown to suffer injury from exposure to
fluoride (Groth 1975a). The response of fish to fluoride is influenced by a
number of factors such as species and strain, concentration of calcium and chloride
in the water, temperature, and the size or age of fish used in the study (Sigler
and Neuhold 1972). The response of other species to fluoride is probably influenced
by at least some of these factors, but few data are available.

Fish and other aquatic species tend to accumulate fluoride from the environment,
primarily in the skeleton (including the gills) and exoskeleton. Groth (1975a)
has tabulated the accumulation of fluoride by a number of species. Stewart et
al. (1974) analyzed specimens from an uncontaminated estuarine-coastal area
of New Zealand. They report fluoride levels from 509 to 2885 ppm (ash basis)
in the skeleton, and from 31 to 209 ppm in the exoskeletons, of different species.
Wright and Davison (1975) also report “background” fluoride levels
for a number of species, but give the data only as ug/g fresh weight. These
authors also report data from controlled experiments that clearly demonstrate
accumulation of fluoride in the exoskeleton of shore crab (Carcinus maenas
). Blue crab (Callinectus sapiduz), exposed to 20 ppm fluoride in water,
accumulated fluoride in the exoskeleton and suffered a 4.5% reduction in growth
increment per molt (Moore 1971). Moore estimates that this would result in a
52% reduction in the final size of an average crab.

Wright (1977) reported a whole-body fluoride concentration of 10 ppm, wet weight
basis, in fry of Brown trout exposed to 5 ppm fluoride in tap water for 200
hours. These fry suffered increased mortality, as compared to fry in a control
tank, but mortality appeared to occur only in a susceptible portion of the total

Hemens and Warwick (1972) and Hemens, et al. (1975) studied the potential environmental
effects of the “scrub water” from an aluminum smelter in South Africa.
Brown mussels (Perma perma) were the most sensitive of the organisms
tested. In this species, mortality occurred at fluoride levels from 1.4 to 7.2
mg/l in sea water after exposure for 15 days, but lack of food during the test-period
may have enhanced toxicity. All species tested had accumulated fluoride extensively
(wholebody fluoride to water-borne external fluoride ratios varied from 25:1
to 149:1) after exposure for 72 days at 52 ppm fluoride. The authors interpret
some of their data as being indicative of a greater fluoride accumulation during
deposition of new skeletal material.

Lubinski and Sparks (1975) attempted to assess the total toxicity to Bluegills
of several pollutants present in the Illinois river by expressing the contribution
of each pollutant as “Bluegill toxicity units”. They found that fluoride
was one of six major contributors to the total toxicity of the river water.
Taft and Martin (1974) have reported
the absence of all living organisms in a fluoride-polluted zone of Tampa Bay.

2.2.2 Insects

Data on the effects of exposure of insects to fluoride are limited. Lillie (1970)
has reviewed the literature on the toxicity of fluoride to honeybees and concluded
that 4 to 5 ug of accumulated fluoride per bee may be the lethal level. Assuming
an average dry weight per bee of 30 m , this corresponds to 130 to 170 ppm (dry
weight). Trautwein et al. (1972) reported that the average total-body fluoride
content of winter-killed bees ranged from 0.63 to 4.81 ug per bee (21 to 160
ppm dry weight basis). The highest levels were found in bees from hives located
near sources of fluoride pollution.

Carlson and Dewey (1971) report data from the analysis of a number of insect
species captured in non-polluted and polluted areas of Montana (Table 10). All
insects were presumably collected live, but the honeybees with an average fluoride
content of 221 ppm probably would not overwinter successfully. Other insects,
such as bumblebees at 406 ppm and sphinx moth at 394 ppm fluoride, must be considered
endangered, even in the absence of further evidence.

Mohamed (1971) reported evidence that exposure to fluoride caused chromosome
damage and mutagenesis in fruit flies (Drosophila metanogaster). In a
continuation of these studies, Gerdes et al. (1971a, exposed four strains of
fruit flies at airborne gaseous HF levels of 0, 1.3, 2.9, 4.2 and 5.5 ppm for
periods of up to 6 weeks. All flies were killed within 3 days at 5.5 ppm. All
strains suffered at least 25% mortality in 6 weeks, even at the lowest level
(1.3 ppm) of exposure, but the relation between mortality and fluoride concentration
was non-linear, especially for the two “wild type” strains. In these
two strains, about 65% of the population appeared to be resistant to fluoride,
even at the 4.2 ppm level.

The offspring of the surviving flies from the 0, 1.3, and 2.9 ppm fluoride levels
of the above experiment were also studied (Gerdes et 1971b). Statistically significant
declines in fecundity and egg hatchability with increasing parental exposures
were observed. Gerdes et al. concluded that the exposure to fluoride caused
genetic damage. The dose-response plots for vegetation (Section 2.l.3) and swine
(see Section 2.2.4) do not appear to indicate the existence of a no-effect threshold
for fluoride. If the genetic effects in insects respond to dose in a similar
manner, the cumulative genetic, evolutionary, and ecological effects of exposure
to low levels of environmental fluoride could become manifest with continued
exposure of successive generations.

Table 10. Fluoride content of insects from polluted and non-polluted areas of Montana (Carlsson and Dewey 1971).  

Type of insect

Range of fluoride contents,
ppm, dry weight basis

Non-polluted area

8 species, all types

3.5 – 16.5
Polluted area


58 – 406

Foliage feeders

21.3 – 48.6

Cambium feeders

8.5 – 52.5


6.1 – 170


2.2.3 Wildlife

Considerable data on the accumulation of fluoride in the skeleton of wild animals
have become-available recently (Kay 1975; Kay et al. 1975a, b; 1976; Stewart
et al. 1974), but data on actual injury to wild species remain sparse. Wild
animals accumulate some fluoride from natural sources, and early field studies
were handicapped by a lack of data on this “background” level. This
lacuna has now been partially filled.

The fluoride concentration of femurs from over 30 species collected in non-polluted
areas of Montana have been reported (Kay et al. 1975a). Data for species from
which bones of 5 or more individuals were analyzed are summarized in Table 11.
This tabulation indicates that the bones of carnivorous species contained more
fluoride (dry fat-free basis) than did those of herbivorous species, but the
data are insufficient to permit firm conclusions regarding food-chain build-up
of fluoride.

In general, data on the accumulation and distribution of fluoride in the bones
of wild species confirm the observations made on laboratory and domestic animals.
Fluoride concentrations were lower in “the less metabolically active diaphyseal
portion of the long bones” than in the distal portions which are “composed
largely of cancellous bone” (Kay 1975). Bone fluoride content appeared
to increase linearly with the age of the animal for 6 years or longer (Kay et
al. 1976). Geographic variations were observed (e.g. means of 72.5 and 248.4
ppm fluoride in bones from different populations of deer mice), but these were
small relative to changes known to result from environmental contaminations
(Kay et al. 1975a).

Table 11. Fluoride content of bones of animals collected in non-polluted areas of Montana (Kay et al.1975a).  

No. of animals
F content of femur ppm,
dry fat-free


103.1 + 16.2*

Columbian ground squirrel

112.5 + 10.2

Deer mouse

143.8 + 7.8


266.4 + 59.8

Northern flying squirrel

141.8 + 30.7


161.0 + 37.1

Red squirrel

151.9 + 29.5

Redback vole

258.0 + 25.3

Whitetail jackrabbit

258.6 + 27.7

Shortail weasel

363.6 + 97.1

Vagrant shrew

474.8 + 98.1
* Mean and standard error of mean.

Stewart et al. (1974) have provided data on the “background”
fluoride levels in bones of various species in New Zealand, Tibia or entire
skeletons were analyzed; data are reported as ppm in bone ash. (NOTE: Bones
contain 50 to 70% ash. Thus a rough conversion of “ppm, ash” to “ppm,
dry fat-free” can be made by multiplying the former by 0.6.) The mean value
for two opossums was 247 ppm, and that for a single rabbit was 184 ppm fluoride.
These values thus agree with those reported for Montana.

When compared with these background levels, bone fluoride concentrations ranging
up to and above 5000 ppm dry fat-free basis, (Kay et al. 1975b; Newman and Yu
1976; Harris 1974) are clearly indicative of environmental contamination by
fluoride and its ingestion by wild animals. Gordon (1970a) recorded extreme
values of 12,700 ppm fluoride (ash basis) in the femur of Mus musculus,
and of 16,000 ppm in a rabbit femur. A relation between bone fluoride levels
in small rodents and the distance from a fluoride source has been demonstrated
(Gordon 1970a).

The data currently available are not sufficient to indicate the environmental
significance of fluoride pollution for wildlife, but there are indications of
serious effects. Lameness induced by fluorosis has been observed in wild ungulates
by Kay et al. (1975b), who note that it appeared to be more severe than the
lameness observed in cattle at similar bone fluoride levels. In a predator-prey
situation, even a minor loss of mobility can lead to rapid elimination of the
individual affected. An apparent population age-shift was also observed (Kay
et al. 1975b), and this “suggests that fluorosis was so severe that older,
most susceptible, deer had been removed from that (the
Teakettle mountain) herd”.

In view of the data discussed above, we feel obligated to disagree with the
statement made by Suttie (1977), to the effect that “There seems to be
no real basis for assuming that these animals (wildlife) are any more susceptible
to the adverse effects of fluoride ingestion than other herbivores, and it is
generally felt that if the most sensitive domestic species, cattle, are protected
the area will be safe for wildlife”. In Section 2.2.4 on “Livestock”,
we discuss a number of factors that influence the severity of skeletal fluorosis.
In a comparison of domestic to wild animals, nearly all of these factors [e.g.
nutritional status (particularly in winter), physical exertion, variability
of fluoride exposure-level, age when exposure begins, degree of individual variability,
etc.] can be unfavorable to the wild animal. Suttie’s statement ignores all
of these factors, and also ignores the increased vulnerability that even mild
fluorosis can create in a predator-prey situation. Kay et al. (1975b) observed
that, for a given level of fluoride in the bones, deer appear to suffer more
severe lameness than cattle. This confirms that the factors listed above do
influence the severity of fluorosis, and also increase the wild animals’ susceptibility
to fluoride toxicity.

Data on wild birds are very limited. Stewart et al. (1974) and Kay et al. (1975a)
have reported some background data (Table 12). In general, short-lived, seed-eating
birds had lower bone fluoride levels than the longer-lived omnivorous species.

The mobility of birds largely precludes sampling of individuals who have remained
in a fluoride contaminated area for long periods. However, the high “background”
fluoride levels observed in some members of the omnivorous species suggests
that there may be some danger of developing skeletal fluorosis. [Note: Fluoride
levels exceeding 4,500 to 5,500 ppm, dry fat-free basis in the long bones, are
considered indicative of marginal fluorosis in cattle (NAS 1971)].

House martins (Dilichon virbica) may be sensitive to fluoride, as few nests
were found in heavily polluted areas (Newman 1977).

Table 12. Fluoride levels in the bones of wild birds from non-polluted areas.

Reference and bird species Number of species
F content of bones
ppm, ash basis
Stewart et al. 1974
Carnivorous or omnivorous

Red-billed gull


White-faced heron


Mallard duck


Black-backed gull


Harrier hawk


Hedge sparrow





Kay et al. 1975a
(ppm, dry fat-free)
Standard Error

Blackbilled magpie


Blue grouse


Ruffed grouse


Sage grouse


Sharptail grouse

Spruce grouse



2.2.4 Livestock

Aschbacher (1973) has stated that “Of all airborne pollutants which may
affect farm animals, fluorine has caused the most serious and widespread damage”.
Research on fluorosis in livestock has been extensive and a number of reviews
have been published (Shupe 1970; Obel 1971; Shupe et al. 1972; Trautwein et
al. 1972; NAS 1974; Fleischer et al 1974; Suttie 1977).

In brief, studies on skeletal fluorosis in livestock have led to the following

(1) Fluorosis results from chronic ingestion of fluoride at levels above those
usually arising from natural sources over a prolonged period; thus, it is more
commonly observed in older animals.

(2) If exposure occurs during the period of tooth formation, tooth
may occur. This can increase tooth wear and contribute to a decline
in the nutritional status and well-being of the animal.

(3) In severe cases, animals become intermittently or permanently lame, and
bone exostoses become radiologically or even visually apparent, especially near
the leg joints.

(4) The severity of the fluorosis is influenced by a number of factors in addition
to total fluoride intake and duration of exposure. Absorption of ingested fluoride
is influenced by the chemical form and solubility of the fluoride and by other
components of the diet (e.g. calcium, aluminum, etc., NAS 1974). Fluoride toxicity
is enhanced by a low nutritional status of the animal (Suttie and Faltin 1973).
The schedule of exposure also influences fluoride toxicity, with alternating
periods of high and low exposure being more harmful than uniform exposures (Suttie
et at. 1972). Physical activity also tends to increase the severity of bone
lesions caused by excessive fluoride (Shupe and Olson
; Shupe et at. 1972).

The age of the animal when exposure begins also affects the development of fluorosis.
This is especially important as regards dental effects caused by exposure during
the period of tooth formation, although age also influences the receptivity
(i.e. affinity) of bone for fluoride. Evidence for a declining rate of fluoride
accumulation with age does not seem to have been presented for large domestic
species, but has been shown with rats, rabbits, and dogs (WHO 1970; NAS 1971).
Because bone lesions appear to be related to bone fluoride levels (NAS 1971),
exposure to fluoride from weaning onward may be more harmful than exposures
later in life.

(5) There are distinct differences among domestic species in their tolerance
to fluoride. Cattle seem to be the most sensitive of the common North American
domestic species, whereas swine are less sensitive and poultry are comparatively

Although there is general agreement on the above points throughout the industrialized
countries, there is a diversity of opinion as to the levels of fluoride that
can be permitted in animal forages and feeds. There are a number of reasons
for this diversity, not the least of which has been the emphasis placed on osteosclerosis
by many researchers in the field, and the difficulty of quantitatively expressing
the degree of osteofluorotic injury. Particular attention should be given to
the more insidious forms of osteofluorosis, such as the marked arthritic changes
observed in dairy cattle fed fluoride-contaminated phosphate supplements (Griffith-Jones

Three indices of fluoride exposure have been proposed for use with livestock.
The most widely accepted index in the U.S. is the fluoride content of fodder
(and of feed supplements); however the fluoride content of bone is a more useful
diagnostic index, and the fluoride content of urine may also have some diagnostic

Suttie (1969a) has proposed that standards for the fluoride content of forage
should be set at:

not over 40 ppm, dry weight basis, as a yearly

not over 60 ppm, for more than 2 consecutive months;

not over 80 ppm, for more than one month.

Various U.S. State regulations make it unlawful for an industry
to emit fluoride at a level that will cause the fluoride content of locally-grown
forage to exceed 30 (Kay 1971) or 40 (Gordon and Tourangeau 1977) ppm, dry weight
basis. These levels have been selected largely on the basis of data obtained
in controlled animal studies (Suttie 1969a) which are not always relevant to
actual farm conditions.

In controlled tests, conditions are selected to minimize the effects of many
of the factors, discussed on p. 46, that are known to influence the severity
of fluorosis. For example, the exposure level is kept constant or varied on
a simple controlled schedule; animals of uniform age are selected; adequate
nutrition is provided; physical exertion is restricted; and individual responses
are largely eliminated by randomization and averaging. Obviously, the severity
of fluorosis observed in such tests will be less than those to be expected in
some individual animals in a range herd. In general, it can be concluded that
the toxicity of a substance having a crippling effect will be underestimated
by studies done on penned animals (i.e., those having restricted mobility) rather
than on grazing animals whose nutritional needs cannot be met without mobility.

Bourbon et al. (1971) and Gordon and Tourangeau (1977) have suggested a single
standard of 20 ppm fluoride, air-dried basis, for all fodder. However, it must
be noted that soils and fertilizers also contribute to the fluoride content
of fodders. Suttie (1969b) reported that “some rather high fluoride forages
(112 ppm) can be found in areas with no known source of industrial fluorides
…” Thus, regulations that attempt to control the level of fluoride in
fodders by restricting airborne industrial emissions may prove inadequate.

Standards controlling the fluoride content of fodders also fail to provide protection
against high fluoride levels in mineral supplements and other types of feed
(Suttie 1969b; Marier 1971; Obel 197 Griffith-Jones 1977; Hillman 1977).

The fluoride content of bone has also been suggested as a quantitative index
of exposure to fluoride (NAS 1971). For monitoring of live animals, this would
require an inconvenient biopsy; but, in the case of farm herds, post-mortem
samples from slaughtered animals are often available. Results of feeding trials
at the University of Wisconsin (summarized in NAS 1971) indicated that bone
fluoride levels in “the range of 4,500 to 5,500 ppm (dry fat-free basis
in long bones such as metacarpal or metatarsal) might be considered as the marginal
zone of toxicosis, and that lower concentrations were not indicative of damage”.
This conclusion, however, appears to be specific to the experimental conditions
used. Another NAS (1974) report has stated:

“Cancellous bone such as the frontal ribs, vertebrae, and those of the pelvis, have a higher fluoride content than the more compact metatarsal and metacarpal bones … There is also a marked variation in the fluoride content of such different anatomical areas (within) bone (such) as the metatarsal or metacarpal; the diaphyseal portion has a lower fluoride content than the metaphyseal portion”

Obel and Erne (1971) observed serious fluorosis in calves
with 500 to 2,400 ppm, and in cows with 900 to 2,800 ppm fluoride in metacarpal
bone ash (assuming 60% ash, these figures correspond to 300, 1,440, 540 and
1,680 ppm, dry fat-free basis, respectively). Obel and Erne suggest that a phosphate
deficiency may have contributed to the severity of fluorosis in some of the
cattle examined. Zumpt (1975) observed fluorosis in sheep at femur bone fluoride
levels of 2,400 to 3,200 ppm dry fat-free basis.

The fluoride content of urine has also been suggested (Burns 1970) as an index
of fluoride ingestion by cattle. This index might be advantageous because of
the ease of sample collection, but the relation between fluoride intake and
urinary fluoride is not well established.

Although Burns (1970) reports a reasonably close relation between urinary fluoride
and the fluoride concentration in samples from the pasture vegetation, Huber
and Schurch (1970) report much less agreement. Israel (1974b) reports a correlation
coefficient of 0.87 between annual average urinary fluoride levels from cattle
and annual average feed and forage samples. Annual averages of urinary fluoride
were based on 3 samples per year from each of 10 to 13 animals per herd. Thus,
under practical conditions, it appears that extensive sampling is required for
urine analyses to provide a reliable indication of fluoride ingestion. Burns
(1970) suggests that 10 ppm would be “a suitable figure to use as a threshold
level” for urinary fluoride. Based on the equation given by Israel (1974b),
this would correspond to a fluoride content in the fodder of less than 20 ppm.

There are continuing difficulties in answering the question of whether or not
fluorine is an essential element of diet (NAS 1974). The criteria for essentiality
and the difficulties of proving it by animal experimentation have been discussed
(Underwood 1962; NAS 1971). One of the greatest difficulties is that practically
every natural water supply and foodstuff contains traces of fluoride and it
is almost impossible to prepare fluorine-free (e.g. < 0.005 ppm) control
diets adequate in other respects (NAS 1974). In view of the conflicting results
and conclusions from experiments with mice (Underwood 1977) it is not yet possible
to assign an essential role to traces of dietary fluoride.

We have found only one set of research data from which a mathematical relation
between fluoride intake and the response of a livestock species can be calculated.
Forsyth et al. (1972c) fed diets containing 0, 30, 150 and 450 ppm fluoride,
as sodium fluoride, to young swine, and recorded average daily weight gains
for up to 18 weeks. No data are given on the fluoride content of the basal diet.
The data (reproduced in Fig. 3) indicate a linear
decline in growth rate with increasing dietary fluoride. In the 18-week experiment,
the values at 150 and 450 ppm fluoride are significantly (p < 0.01) different
from the control values. The regression equations (our calculations) indicate
a loss of about 4% in the average daily weight gain, over the 18-week period,
for each 100 ppm increment in dietary fluoride. Said et al. (1974) have reported
that “retarded liveweight gain was the first significant sign of fluorosis”
in a 25-month study of Wether sheep fed from 53 to 70 ppm fluoride in the total

In the absence of additional quantitative criteria, and in view of the fact
that the indices of fluorosis discussed above (i.e. bone and urine fluoride
concentrations) do not appear to be satisfactory relative to actual farm experience,
and do not appear to give adequate protection to wild species, the present authors
cannot suggest criteria that would be of use in setting or revising Canadian
standards for exposure of animals to fluoride. However, two suggestions can
be made.

1. It should be emphasized that total fluoride intake is the only reliable index
of chronic exposure for fluoride. The use of maximum “safe-levels”
of fluoride in fodders is based on the assumption that intake of fluoride from
all other sources, including water, will be low and relatively constant. Reported
instances in which fluoride from sources other than fodder detrimentally affected
the health of animals (Obel 1971; Griffith-Jones 1972, 1977; Parsonson et al.
1975; Hillman 1977) testify to the need to assess the contribution from all
sources. Oelschlager (1974) has stated that “there appears to be a lack
of full appreciation of the extraordinary amounts of fluoride which reach the
feed rations through mineral supplement mixtures…”

2. Further research aimed at developing criteria relating fluoride intake, preferably
in mg/kg body weight per day, to tissue fluoride contents and injury should
be stressed. This research should include studies on animals at less-than-optimum
nutritional status. Attention should be paid to the less obvious effects of
fluoride, such as the reduced growth of swine (and sheep) discussed above, rather
than to osteofluorosis. Blood plasma F- should be assessed, as a possible indicator
of fluoride exposure and the likelihood of fluoride intoxication. Nutritional
factors are of extreme importance in chronic fluoride intake (see Section 5.6).
Chronic dietary deficiencies can aggravate the effects of a given fluoride dosage,
and such factors should be considered in the assessment of dose:response interrelations.


Terrestrial species have evolved in contact with small but variable amounts
of fluoride, and can be assumed to have a degree of tolerance for trace amounts
of ingested fluoride. Nevertheless, ingestion of even small amounts seems to
induce physiological responses, and many of these responses are dose-dependent.
At some level of intake and duration of exposure, therefore, the effects will
cease to be “tolerable” and will begin to exert a stress on the organism.
The level of bioaccumulation of fluoride that will exceed the animal’s tolerance
is not necessarily constant but may vary with the efficiency of various bodily
functions and with the stresses being imposed concurrently by environmental,
nutritional, pathological, and other factors. The various physiological responses
of animals to fluoride are thus of interest. In the following section, we attempt
to review and summarize recent information relative to these responses.



3.1.1 Fluoride Content of Blood

The presence of fluoride in animal and human blood has been recognized for many
years (WHO 1970). About three-fourths of the fluoride in blood is contained
in the plasma; the remainder is in the erythrocytes, which make up 40 to 50%
of the blood volume. For various reasons, analysis of serum provides the most
satisfactory information for the assessment of interrelations between fluoride
and other factors (WHO 1970). Cowell (1975) noted that some anticoagulants used
in the preparation of serum may contain fluoride as a contaminant; caution is
therefore necessary in their use.

The forms in which the fluoride in blood exists are still the subject of research.
Taves (1968) presented evidence for the existence of two forms of fluoride in
human blood, i.e. organically-bound (4.6 umol/1) and free or ionic fluoride
(0.7 umol/l; data for people in a non-fluoridated community). The ionic fluoride
(F-) moiety was shown to be the one that responded to changes in fluoride ingestion;
total serum fluoride is a less sensitive indicator of such changes (Taves 1968,
1970). Some doubt as to the significance of the organically-bound fluoride fraction
was raised when Taves (1971) reported that it was not present in the blood of
dogs and rats. Taves had suggested that the bound fluoride was associated with
serum albumin, but neither Jardillier and Desmet (1973) nor Ekstrand et al.
(1977) could find any evidence for protein-bound fluoride in human blood. Jardillier
and Desmet (1973) suggested that the bound fluoride reported by Taves was “en
fait du fluor lie par covalence a des petites molecules organiques”. This
suggestion has now been partially confirmed (Taves et al. 1976) by the isolation
of a major component of the bound fluoride, showing “an nmr pattern consistent
with a derivative of perfluorinated octanoic acid”. Taves et al. suggest
that the presence of this compound in plasma is “at least partially the
result of contamination from industrial sources”.

Final identification of the bound fluoride in human plasma, and clarification
of its source and physiological significance, must await further research. Its
absence from the plasma of dogs and rats (Taves 1971) and from bovine plasma
(Taves et al. 1976) may indicate a specificity of human metabolism or a peculiarity
of the human environment.

As recently as 1970 (WHO 1970), it could be stated that “regulatory mechanisms
operate within the body to maintain the plasma fluoride content… within narrow
limits”. However, improved analytical procedures, greater attention to
the ionic fluoride fraction in blood, and more critical studies have now shown
that the plasma inorganic fluoride ion concentration responds rapidly and systematically
to varying fluoride ingestion and to various physiological factors. Thus, Taves
(1970) observed that “serum ion concentrations (of fluoride) were 0.6 and
0.7 umol/l in two individuals after an overnight fasting and increased to nearly
twice this level in about 50 minutes after drinking 500 ml of water” containing
52 umoles fluoride per litre (i.e. 1 ppm).

Intraperitoneal injection of 0.1 mg F/kg body weight into rats caused a rapid
rise in plasma F- values, which reached a peak of about 10 umol/l in less than
10 minutes (Angmar-Manson et al. 1970). Plasma F- levels then declined progressively
to levels of less than 2 umol/l 45 minutes after injection.

When rats were provided with drinking water containing 50 ppm fluoride, the
plasma F- level increased from 0.17 ± .007 mg/kg to 0.26 ± 0.013 mg/kg in 4
weeks (Singer et al. 1976). After replacement of the fluoridated water with
distilled water, plasma F- dropped to the original level in 3 days. Suketa et
al. (1976) found that a single oral dose of 50 mg F/kg body weight given to
rats raised plasma F- levels to above 2.0 ppm (105 pmol/1) in one hour. This
peak was followed by a somewhat irregular decline, and the plasma F- concentration
approached the original level 10 hours after fluoride administration.

The plasma F- level of 4- to 5-month-old humans also varied with fluoride intake
(Hellstrom 1976). Fourteen infants on breast feeding (low fluoride ingestion)
had a mean plasma F- level of 0.027 ppm, while 16 infants on formulae prepared
with fluoridated water (1.2 ppm) had a mean plasma F- level of 0.045 ppm (1.42
and 2.37 umol/l, respectively). Kuznetso (1969) reported that total fluoride
in the “biological media” of human fetuses during the 5th to 12th
week of pregnancy was 1.56 ppm in unexposed mothers, and 2.72 ppm in mothers
working in a superphosphate factory.

Posen et al. (1971) reported that plasma F increased with increasing bone fluoride
(iliac crest, dry fat-free basis) in human patients undergoing hemodialysis
treatment with fluoridated water l ppm. Ericsson et al. (1973) examined humans
in a community with 10 ppm fluoride in the drinking water, and reported a positive
correlation between bone fluoride (iliac crest biopsy) and plasma F-. Urine
fluoride, on the other hand, showed little relation to either bone or plasma
fluoride. An interrelation between bone fluoride and plasma F- is also suggested
by the data of Parkins et al. (1974), which indicate that both bone fluoride
(iliac crest biopsy) and plasma F- increased with increasing age in humans.

Equations relating plasma F- level to age in humans have been presented by several
authors (Table 13). Some of the researchers found no significant correlation
for some of the sub-groups studied [e.g. men under 45 (Husdan et al. 1976)],
but the trend towards increasing plasma F- with increasing age is clear. In
non-fluoridated communities, young adults tend to have plasma F- levels below
0.7 umol/l, and this increases by about 0.01 to 0.02 pmol/l per year (Table

As illustrated by Hanhijarvi’s data (Table 13), plasma F levels vary with the
amount of fluoride in the drinking water. However, the range of values reported
by various workers (Table 14) suggests that interlaboratory standardization
of the methodology might be advantageous.

Table 13. Regression equations relating plasma ionic fluoride levels to age in adult humans.

Community water Subjects Regression equation (1) Reference
Fluoridated Both sexes F = 0.631 + 0.0368A Parkins et al. 1974
Non-fluoridatedFluoridated Both sexesBoth sexes F = 0.761 + 0.00259AF = 0.975 + 0.00929A Hanhijarvi 1975
FluoridatedFluoridated WomenMen over 45 (2) F = 0.375 + 0.022AF = 0.906 + 0.0101A Husdan et al. 1976
Non-fluoridatedNon-fluoridated MenWomen F = 0.683 + 0.016ANo correlation Kuo and Stamm 1975
(1) F = plasma F-, umol/l; A =
age in years.(2) Husdan et al. (1976) report a constant plasma F- level of 0.906 +
0.306 umol/l for men under 45 years of age.


Table 14. Mean plasma ionic fluoride levels for humans residing in non-fluoridated and fluoridated communities.

No. of individuals
Plasma F- umol/l
0.88 +0.0091.3 + 0.0005
Hanhijarvi 1975
Non-fluoridatedFluoridated, 3.8 ppm
3.3 +0.66.7 + 1.2
Jardillier and Desmets 1973
0.46 +0.10.41 + 0.02
Bierenbaum et al. 1974
Non-fluoridated0.15 ppm
Desmet et al. 1975

Administration of 27.15 mg fluoride (as 60 mg NaF) to patients
suffering from osteoporosis or Paget’s disease gave rise to plasma Flevels of
3.26 to as high as 14.41 umol/l in the individual patients (Cowell 1975). Cowell
states that the plasma F- levels “appear to be directly proportional to
the dose”, but examination of his Figures 8 and 9 suggests that the relation
was non-linear. The range of plasma F levels reported by Cowell (1975) for normal
adults was 0.31 to 2.21 umol/l.

Posen et al. (1971) reported arterial blood plasma F- levels ranging from 8.9
to 23 umol/l in 10 patients receiving hemodialysis treatment with fluoridated
(1 ppm) water. The plasma F- tended to increase with increasing time on the
hemodialysis program, and the mean F- value increased during dialysis (pre-dialysis
mean, 16 ± 4 umol/l; postdialysis mean, 28 ± 3). Cordy et al. (1974) reported
a mean plasma F value of 3.35 ± 0.28 umol/l for 34 patients using a non-fluoridated
water for hemodialysis, and of 12.30 ± 1.98 umol/l for 7 patients using fluoridated

Renal insufficiency in humans can result in high plasma F levels. Seidenberg
et al. (1976) reported a mean value of 2.40 umol/l in 10 patients suffering
from a renal insufficiency in a non-fluoridated area, whereas thirteen “controls”
from the same area had a mean plasma F concentration of 0.67 umol/l. Hosking
and Chamberlain (1972) observed a slower decline in the plasma F level in anuric
than in normal patients following single-dose intravenous injection of 18F;
however, in anuric patients with secondary hyperparathyroidism, the uptake of
fluoride by bone offset the lack of renal excretion, and plasma F- declined
even more rapidly than in control patients. Kuo and Stamm (1975) report that
an impaired ability to excrete creatinine was apparently not related to plasma
F- levels. Hanhijarvi (1975) has shown that some of the persons with renal insufficiency
show no change in plasma F- level, while others have a level 312 to 5 times
higher than seen in the similarly-exposed general population. In the same survey,
Hanhijarvi has also shown that diabetics have abnormally high plasma F- levels.


3.1.2 Effect of Fluoride on Blood Components

In addition to increasing the plasma F- levels, ingestion of fluoride influences
several blood constituents. Recent reports on such changes are summarized in
Table 15 for experimental animals, and in Table 16 for humans. Because of the
multiplicity of experimental procedures used and the various blood components
studied, comparisons are difficult. The most obvious onsistent observation in
animals (Table 15) has been a decrease in red blood cells; this was observed
in both the rabbit and rat, and by various methods of measurement (blood iron,
blood hemoglobin, erythrocyte count). This observation is of interest because
of the reports of anemia in humans residing near sources of fluoride emissions
(see Section 5.4).

The relations between some of the changes reported in Table 16 and fluorosis
in humans will be discussed in Sections 5.4 to 5.9. For the present, we conclude,
in agreement with Manocha et al. (1975), that “More elaborate studies on
non-human primates …. are needed to clarify the effects of different doses
of fluoride on the biological system”.

When a single intraperitoneal dose of 20 mg fluoride/kg body weight was given
to rats (Baker 1974), total serum fluoride increased by over 100-fold in 5 minutes,
serum calcium declined 20% in 30 minutes, and serum phosphorus increased 40%
in 1 hour. For all three components, there was a gradual reversal from these
peak values, and concentrations were approaching normal 4 hours after injection.

For more discussion of this aspect of fluoride interactions, the reader is referred
to the reports by Guminska and Sterkowicz (1975) and Manocha et al. (1975),
who also discuss various enzymatic processes.

Table 15. Effect of fluoride on the levels of various blood components in experimental animals.

Species Fluoride source Duration of study Blood component studied Response Reference
Rat Diet, 450 or 600 ppm 3 days2 weeks CitrateCitrate IncreasedReturned to normal Shearer et al. 1971.
Rabbit Oral, 1 mg/kg 30 days Total blood iron Decreased Soldatovic and Nadeljkovic-Tomic
Oral, 20 mg/kg body weight
30 days ErythrocytesLeucocytesHemoglobinBlood iron Decreased”””
Rat Water, 150 ppm 75 days Red cell count Decreased Kahl et al. 1973
Rat Air, 0.05, 0.47 and 4.98 mg/m3 Blood hemoglobinLeucocyte countErythrocyte countCholinesterase activity Decreased””” Danilov and Kas’yanova 1975
Rabbit Water, 10 ppm 12 weeks Serum glutamate-oxalate transaminaseSerum alkaline phosphataseSerum malate dehydrogenaseSerum lactate dehydrogenaseSerum isocitrate dehydrogenase Decreased, total periodDecreased, 2nd 6-week periodDecreased, 1st 6-week periodDecreased, total periodDecreased, 2nd 6-week period Ferguson 1976
Rat Single oral dose, 50 mg/kg
body weight
3 hours12 hours Erythrocyte fluid potassium” ” calciumPlasma acid phosphatasePlasma alkaline phosphataseMg-activated ATPaseNa + K-activated ATPaseMg-activated ATPaseNa + K-activated ATPase IncreasedDecreasedDecreasedNo changeIncreasedDecreasedReturning to normal” ” “ Suketa et al. 1976
Rat Water 57 ppm 70 days Whole body hematocritErythrocyte volumePlasma volumeTrue blood volumePlasma vol./unit body weightTrue blood vol./unit body weight DecreasedSlightly decreasedIncreased””” Kahl and Ewy-Dura 1976
Guinea pig Water; 5, 25 ppmHigh-fat dietLow-fat dietHigh-fat dietLow-fat diet 20 weeks Serum triglycerides” “Serum phospholipids” “ DecreasedIncreasedDecreasedNo change Townsend and Singer 1977



Table 16. Effect of fluoride on the levels of various blood components in humans.

Fluoride source Duration of study Blood component studied Response Reference
Industrial exp. Not stated Manganese Nikolaev and Kas’Yanova 1971
Oral, 5 mg/dayWater 1 ppm 6 weeks3,4 weeks22 weeks Alkaline phosphatase” “” “ Ferguson 1971
Industrial exp. Long-term ManganeseAluminumCobaltZincIron DecreasedDecreasedIncreasedIncreasedIncreased Nikolaev and Sidorkin 1972
Industrial exp. 18 years Erythrocyte pyruvate kinaseSerum pyruv. kinaseSerum lactate dehydrogenaseErythrocyte ATP IncreasedIncreasedIncreasedDecreased Guminska and Sterkowicz 1975
Infants on formula, 1.2 ppm in water 4-5 months Serum alkaline phosphatase Increased Hellstrom 1976



3.2.1 Fluoride Content of Urine

The fluoride content of urine has been suggested as an index of animal exposure
(cf. Section 2), and as a diagnostic test for humans during chronic exposure to
fluoride (e.g. Pantucek 1975). A number of reports dealing with the fluoride content
of urine under various conditions have been published during the last seven years,
and these are reviewed below.

Toth and Sugar (1975) reported data from a survey of individuals in an area of
Hungary in which the water was low in fluoride and there was no industrial source
of fluoride emissions. The average fluoride content of 24-hour urine samples was
0.26 ± 0.01 mg/l; the highest value recorded was 0.57 mg/l. Variations between
days, and with time of day, were apparent but rarely exceeded 0.2 mg/l.

Mose et a. (1969) reported a relation between average urinary fluoride levels,
population density, and degree of industrialization, indicating that urinary fluoride
was increased by domestic and industrial pollution. Archer et al. (1975) were
unable to detect an increase in urinary fluoride among grade 5 students residing
at various distances from an aluminum smelter. However, Archer et al. did not
take 24-hour urine samples, and their study was conducted against a background
intake from fluoridated (0.8 ppm) water and a mean excretion of 0.97 ± 0.42 mg
fluoride per litre urine. This is in contrast with the Mose et al. (1969), study,
where the average urinary fluoride level for the “control” rural population
was only 0.27 ± 0.03 mg/l (our estimate from authors’ bar-gravph).

Balazova et al. (1970), and Tsunoda et al. (1973) also report increased urinary
excretion of fluoride by individuals residing near sources of airborne fluoride
emissions. Mean values from Tsunoda et ai. are (i.e., as mg F excreted in 24 hr):

Non-polluted area Polluted Area
Men 0.79 + 0.07 2.05 + 0.30
Women 0.73 + 0.07 1.77 + 0.43

Tsunoda et al. (1973) emphasize the inadequacy of spot-testing
of urines as a means of detecting exposure to fluoride. Hodge and Smith (1977)
also note that it is “impossible to assess general working conditions from
a single spot urine sample from a single individual”. The limited usefulness
of spot-urine samples is also indicated by the data of Ericsson et al. (1973),
who reported that the fluoride concentration of “night urine” from
individuals in a community with 10 ppm fluoride in the drinking water was not
related to either plasma F- or iliac crest fluoride concentrations. Plasma F-
itself was positively related to iliac crest fluoride.

Relative to the absorption and excretion of fluoride by workers and by residents
in areas subjected to fluoride emissions. it appears that particulate fluoride
is excreted in the urine “as promptly and quantitatively” as gaseous
fluoride (Hodge and Smith 1977).

Polakoff et ae. (1974) observed small but statistically significant increases
in urinary inorganic fluoride by some workers in a polytetrafluoroethylene factory.

Pantucek (1975) studied the excretion of fluoride in urine by welders over a
one month period. The average urinary fluoride level in groups of unexposed
workers was 0.70 ± 0.03 mg1l or less. Exposed workers excreted from 1.4 to 1.8
mg fluoride/l urine before work on Monday, and the level rose progressively
to attain 2.0 to 2.3 mg/l on Friday morning. Afternoon samples (between 1300
and 1400 hours) ranged from 2.5 to 3.3 mg/l on Mondays, and from 3.6 to 4.4
mg/l on Fridays.

Urinary fluoride excretion in 24-hour samples taken at least 6 days after any
period of exposure was markedly higher in long-term aluminum smelter workers
(average of 27 years’ exposure) than in a control group (Boillat et al. 1976).
Values estimated from the authors’ graphs are:

Controls, av., 0.7 mg/24 hrs; range, 0.2 to 1.1 mg

Exposed, av., 2.4 mg/24 hrs; range, 1.2 to 5.3 mg.

Dinman et al. (1976a) report a linear relation between 24-hour fluoride excretion
in urine and the atmospheric fluoride level to which potline
in an aluminum smelter were exposed (11 workers over a single 24-hour
period). Dinman et al. (1976b) also present curves showing a relation between
the concentration of fluoride in post-shift urine and the number of days worked
after a 2- or 3-day break, but these curves must be accepted with caution. No
probability limits are shown, and the curves are not carried beyond the third
working day, because “the variation after the third day was so great between
and within job categories, that it was impossible to fit a statistically significant
regression line”. Calculations from the tabular data (Dinman et a. 1976b,
Table 2) give coefficients of variation between 5.8 and 22.9 for days 0 to 3,
and between 6.7 and 10.6 for days 4 to 7. Also, the 2.5- to 3.4-fold increases
represented by these curves are excessive even in terms of the assumptions made
by Dinman et al., that workers in a “non-steady state” excrete 50%
of the ingested fluorine, while for workers in a “steady-state” at
the end of the week, excretion “approaches 100%” of the dose.

Fluoride that has accumulated in the skeleton of humans is not readily excreted
in urine, following a reduction in fluoride intake. Thus, when Spencer, Osis
and Wiatrowski (1974) administered 10 mg fluoride daily for 32 days to hospitalized
patients, a total retention of 114 mg (36%) occurred. Retention was relatively
constant throughout the 32-day period, with no evidence that the proportion
excreted increased with exposure-time. Of the 114 mg retained, only 9.8 mg was
excreted in 12 days following cessation of treatment. In a later paper, Spencer
et al. (1975) note that a negative fluoride balance occurred only during the
first six-day period following cessation of fluoride administration. The fluoride
ingestion subsequent to treatment was 4.36 mg/day from foods and beverages.

Jolly (1976) conducted fluoride-balance studies on control and fluorotic patients
(10 in each group) having fluoride intakes from dietary sources of 3.74 ± 0.30,
and 3.44 ± 0.25 mg/day, respectively, during the test period. Fluoride excretion,
over three consecutive 24-hour periods, was 3.34 ± 0:23 mg/day (urine plus feces;
feces contributed 8 to 12% of the total) in control patients and was not related
to age. In the fluorotic group, excretion averaged 6.55 + 1.52 mg/day,
and tended to decline with increasing age. The length of the hospitalization
period before the tests is not reported. Presumably, the excess fluoride excretion
over intake involves loss of non-lattice bound (i.e. surface) skeleton fluoride
from these fluorotic patients (WHO 1970).

Hanhijarvi (1975) also reported a variation in renal fluoride excretion with
age. Fluoride clearance increased with age “Until about age 50, whereafter
a slight decline was found” in communities with either low (0.02 ppm) or
high (1 ppm) fluoride in the water. Hanhijarvi interprets the increase before
age 50 to “a possible slow saturation of the bones with fluoride”,
and the subsequent decrease to “diminishing renal function which is characteristic
for older people”.

Urinary excretion of fluoride is affected by a number of factors besides age.
Kuo and Stamm (1975) studied groups of men and women classified on the basis
of creatinine clearance tests, and recorded the following 24-hour urinary fluoride

Creatinine clearance: Normal Impaired
Men 0.87 + 0.70 mg 0.30 + 0.35 mg
Women 0.70 + 0.58 mg 0.24 + 0.12 mg

Hanhijarvi (1975) reported the following renal fluoride
clearances for patients with various conditions:


Controls (28 patients) 1.10 + 0.10 mg/day
Pregnancy (11 patients) 0.84 + 0.08 mg/day
Diabetes mellitus (2 patients) 0.59 + 0.19 mg/day
Renal insufficiency (5 patients) 0.59 + 0.12 mg/day

Buttner and Karle (1974) observed greater fluoride retention,
which implies lower urinary excretion, in unilaterally nephrectomized rats given
25 to 100 ppm fluoride in drinking water.


3.2.2 Effect of Fluoride on Urine Components

Fluoride ingestion also influences the concentration of some other urine components.
Polyuria occurred in rats following addition of 0.1% sodium fluoride to the
diet (Hamuro 1972b). Urinary excretion of calcium and magnesium by rats increased
significantly with increasing fluoride in the diet (Benetato et al. 1970); for
the final three months of the test period, the recorded values were:


Fluoride levels: 0 2.26 4.25 mg/day
Calcium excretion 0.16 + 0.010 0.19 + 0.003 0.26 + 0.017 mg/day
Magnesium excretion 0.24 + 0.010 0.24 + 0.25 0.41 + 0.025 mg/day

Marier (1968) reviewed the importance of dietary magnesium
and its interrelations with fluoride. In a more recent study, Hamuro (1972a)
observed that, in a strain of mice which are prone to kidney calcification induced
by magnesium deficiency, the normally-observed increase in urinary phosphorus,
in response to magnesium deficiency, was largely prevented by increased dietary
fluoride. Added dietary fluoride had no effect on urinary phosphorus in normal
mice. In a study of kidney calcification, Ophaug and Singer (1976) found that
fluoride “exerted an initial protective effect”, but that the longer
term effect was “to promote calcification of kidneys”. They also noted
that fluoride significantly retarded the mobilization of skeletal magnesium.

Speirs and Adams (1971) reported that ingestion of 2 or 4 mg fluoride per day
by healthy men increased the 24-hour urinary excretion of hydroxyproline and
of citrate (3-week control period vs 3-week exposure period). These authors
conclude that “low doses of fluoride seem to have some (additional) systemic
effects, but relatively large daily variations in the urinary composition ….
mask the significance of these small effects”. Additional observations
on the urinary excretion of hydroxyproline (and other metabolites) are referred
to in our discussion of fluorosis in humans (see Sections 5.3 to 5.8).



Recent reports in which data on fluoride-induced changes in liver, kidney, bone-marrow,
spleen, and some neural tissues of experimental animals have been presented,
are summarized in Table 17. Only one corresponding study on humans has been
noted: Franke et al. (1973) discussed structural changes in anterior brain cells
and muscle cells of fluorotic patients.

The involvement of neural and muscle cells in the pathology of fluorosis, as
discussed by Franke et al. (1973), accords with observations of Czechowicz et
at, (1974) on brain cells, and of Benetato et al. (1970) on neuro-muscular excitability
(Table 17). The loss of auditory response in Guinea pigs (Kowalewska 1974) may
also be indicative of neural injury.

Although a single high dose of fluoride caused an increase in the calcium and
magnesium content of rat kidneys, (Suketa et al. 1976), the longer-term response
to low doses was a decrease in calcium and magnesium in kidneys (Benetato et
al. 1970). Somewhat contrasting changes in enzyme activities have been reported
for high (Zhavaronkov and Dubynin 1971) and low (Manocha et al. 1975) doses
of fluoride in squirrel monkeys (Table 17).

The changes in liver enzymes related to carbohydrate metabolism observed by
Shearer et al. (1970, 1971, 1972) appear to have been transient in normal rats,
but observations were not carried beyond 17 days. Related non-transient changes
have been reported in guinea pigs (Czechowicz et al. 1974) and squirrel monkeys
(Manocha et al. 1975

The changes in iron incorporation by spleen and whole blood (Kahl et al. 1972)
are probably related to the erythrocyte changes noted in Table 16.

Table 17. Effect of fluoride on levels of metabolites in, and physiological activities of, animal soft-tissue organs.

Species Fluoride source Duration of study Organ or tissue studied Response Reference
Rat Sub-cut. inject..3, 1.0 mg/kg body weight daily 206 days Liver”””” Marked impairment of ATPase act.Decreased ability to absorb 02 and discharge CO2Decreased respiratory coefficientDecreased alk. phosphatase act.Cytological changes in cell nuclei Zharvoronkov et al. 1969
Rat Oral in milk,2.26, 4.52 ,g F/l 4 months Cerveau, kidney, musclesKidney, liverNeuro-muscular Decreased calciumDecreased magnesiumExcitability, changes characteristic of latent tetany Benetato et al. 1970
Rat Inter-perit. inject.2 mg/kg body weight 15 min. Liver” Apparent inhibition of pyruvate
kinasePossible inhibition of enolase
Shearer and Suttie 1970
Diet, 450, 600 ppmControlled feed 3 day17 day Liver” Increased citrateLittle effect on citric acid cycle intermediates
Rat Diet, 450, 600 ppmControlled feed 3 day14 day Liver” Citrate increasedCitrate returned to normal Shearer et al. 1971
Rat Sub-cut. inject. 12 mg/kg body weight daily 12 weeks Kidney Morphological alterationsIncreased enzyme activitiesDecreased alk. and acid phosphatase activity Zhavoronkov and Dubynin 1971
Rat Diet, 600 ppm 3 day14 day3 day14 day Liver, normal animal” ” “Liver, parathyroidect. rat” ” “ Citrate decreasedCitrate returned to normalCitrate increasedCitrate remained high Shearer 1972
Rat Water, 150 ppm 75 day LiverBone marrowSpleenWhole blood Increased incorporation of iron” ” “Decreased ” ” “Decreased ” ” “ Kahl et al. 1973
Guinea pig Sub-cut. inject.4 mg/kg 90 day Corte organAuditory nerve Microphonic potential decreasedNerve potential decreased Kowalewska 1974
Guinea pig Inter. musc. inject.4 mg/kg b.w. daily 3 months Purkinje’s cells,(Cerebral cortex) Intensified reaction for succinic
dehydrogenase, glucose-6-phosphatase, ATPase, alk. phosphatase, and non-specific
Czechowicz et al. 1974
Squirrel monkey Water, 1, 5 ppm 18 months Kidneys Cytological changes, increased
acid phosphatase activityIncreased activity of enzymes of citric acid pentose shunt
Manocha et al. 1975
Rat Int. perit.16 mg/kg b.w. 6 to 96 hours Kidney medullaKidney cortex Increased calcium, increased magnesiumIncreased calcium, increased magnesium, increased Ca++-ATPase Suketa et al. 1977

3.4 BONE

3.4.1 Fluoride Content of Bone

We have already discussed some aspects of bone fluoride in Section 2.2.4. In
the following paragraphs, we attempt to summarize recent studies dealing with
the physicochemical effects of fluoride on bones. The health aspects of this
subject, and the hormonal etc. interrelations, will be considered in Sections
5.1 to 5.8. For additional information on the skeletal effects of fluoride,
the reader is referred to recent discussion papers (Spencer et al. 1971; FCT
1973; Franke et al. 1975; Rao and Friedman 1975), as well as to the WHO (1970)
monograph, Marier and Rose (1971), Groth (1973), and Section 5 of the present

Accumulation of fluoride in the mammalian skeleton begins during gestation.
Female mice given water containing 50 ppm fluoride gave birth to offspring with
a skeletal fluoride content of 900 ppm, as compared to 4300 ppm in the dam’s
skeleton (Messer et at 1974 . The fluoride level in the newborn mice decreased
slightly during breastfeeding, because the fluoride content of milk is low.
In swine (Forsyth et al. 1972b), the fluoride content of the bones of newborn
piglets appeared to be a linear function of the fluoride fed to the dams, which
were given fluoride supplements ranging from 0 to 450 ppm in the diet. Increasing
the calcium and phosphorus content of the dam’s diet decreased the fluoride
concentration in the piglets’ bones. Inorganic fluoride that has been metabolically
released from methoxyflurane (a fluorinated anesthetic also used as an analgesic
during human childbirth) is “preferentially deposited in the fetal skeleton”
in rats after 15 days of gestation (Fiserova-Bergerova 1976).

Shen and Taves (1974) showed that, in humans, blood from the fetal cord contained
about 75% as much fluoride ion as found in the mothers’ blood, indicating that
the fetal bones are exposed to fluoride during development. Hanhijarvi (1975)
reported a decrease in plasma F- concentration in the mothers’ blood during
human pregnancy; this decrease presumably resulted from a rapid transfer of
fluoride to the developing fetal skeleton. Hellstrom (1976) observed that the
fluoride content of the bones of newborn humans were higher, by 50% or more,
when the mothers drank water containing 1.1 ppm F, than when the mothers drank
low-fluoride (about 0.25 ppm) water.

Although a number of ancillary factors have been identified, the accumulation
of fluoride in the skeleton of animals during growth and maturity is primarily
controlled by three factors: the amount of fluoride absorbed via the digestive
system and lungs; the reactivity or receptivity of the skeletal surfaces; and
the efficiency of fluoride excretion by the kidneys. Receptivity of the skeletal
surfaces is a function of bone age and type, with young bone and cancellous
bone being more receptive than old bone or cortical bone (WHO 1970). The efficiency
of renal excretion is a function of kidney health, and may decline with increasing
age (cf. discussion by Husdan et al. 1976).

If no large variations in fluoride ingestion have occurred, analyses of a selected
bone sample such as the iliac crest (cf. Franke and Auermann 1972) usually show
a progressive increase in bone fluoride with age in humans. Theoretically, fluoride
accumulation in bone might occur less rapidly as the fluoride content of bone
increases, and a “plateau” effect after about age 50 has been discussed
(cf. Jackson and Weidmann 1958; WHO 1970). However, this effect, if it occurs,
may be offset by declining kidney efficiency with age, and the net effect in
humans seems to approximate a linear relation between bone fluoride content
and age. For a population (100 individuals) using a low-fluoride water (0.2
ppm), the regression equation for bone fluoride (iliac crest) and age found
by Schellmann and Zober (1975) was:


Bone F(ug/g ash) = 136.5 + 6.1 x Age (years) or,

Bone F(ug/g fresh weight) = 37.5 + 1.38 x Age (years).

For 20 individuals currently using fluoridated water (1
ppm, residence time not reported), Parkins et al. (1974) also analyzed iliac
crest bone, and found the equation to be:


Bone F(ug/g fresh weight) = 441.9 + 48.3 x Age (years).

Comparison of this equation with that of Schellman and Zober
(1975) clearly demonstrates the cumulative effect of even 1 ppm fluoride in
water on skeletal fluoride of the aging human. Persons with renal failure can
have a skeletal fluoride content 4-fold greater than that of similarly exposed
persons who have healthy kidneys (Mernagh et al. 1977; Marier 1977; see also
Section 5.8).


3.4.2 Fluoride-Induced Changes in Bone

The changes induced in bone as a result of fluoride accumulation can include
both direct and indirect (e.g. hormonal, as discussed in Section 5) effects.
Most studies on the direct effects of fluoride on bone, such as those on bone
density, attempt to relate the bone properties directly to dietary or drinking-water
fluoride (cf. Section 2.2 . 4). However, it appears probable that analysis of
blood plasma F- concentrations in exposed individuals and populations would
provide a more meaningful index of the chronic exposed and saturation-status
of bones to fluoride (cf. Section 3.1.1).

Recent reports on some effects of fluoride on physical properties of animal
bone are summarized in Table 18. The data in this Table adequately present the
complexity of the interrelations involving fluoride, calcium, age, and species.
They clearly show, however, that exposure to fluoride, at dose-levels and durations
that induce no recognizable symptoms of bone-fluorosis, is nevertheless a potent
cumulative agent affecting structure of bone and its response to other stress
factors. For more extensive discussion, see Spencer et al. (1971), Cohn et al.
(1971), Franke et al. (1976), and Inkovaara et al. (1975).

Recent papers have also confirmed that ingested fluoride influences the chemical
composition of bones. Wolinsky et al. (1972) analyzed pooled samples of femurs
and tibias from control rats and rats given 200 ppm fluoride in drinking water
for 2 weeks. Fluoride ingestion decreased the concentration of citrate and of
lipids in the bone, and decreased the in vitro formation of lipid from 14C-acetate
or citrate.

Chan et al. (1973) reported that, in quail, fluoride (750 ppm in diet for 35
days) increased bone ash and the magnesium content of the ash. Rosenquist (1974)
observed a higher magnesium level in fluorotic bone from rabbits given 10 mg
fluoride/kg body weight per day for 14 weeks. Fluoride ingestion also increased
the magnesium content of bone ash of rats, roosters, and quail which had been
made osteopenic by control of the calcium and phosphorus content of the diets
(Riggins et al. 1976).

Miller et al. (1977) report a higher calcium and somewhat lower phosphorus content
in bones of cows suffering from osteoporosis as a result of environmental exposure
to fluoride. They also observed an increased bone alkaline phosphatase activity,
but no change in the citrate content of these bones.

Henrikson et al. (1970) found that, in osteoporotic bones of dogs, fluoride
caused a slight decrease in calcium content, and an increase in phosphorus content
of bone ash. The mineral mass also increased with increasing dietary fluoride,
but there was no fluoride-related improvement (i.e. radiographic etc.) in the
degree of osteoporosis. The effects of fluoride on human bone are discussed
in Sections 5.3 to 5.6.

Table 18. Effects of fluoride on some physical properties of animal bones.

Species Fluoride source and duration Observations Reference
Rat Diet, 3.4, 10, 45 ppm,15 weeks With adequate calcium, increase
in flexability, no decrease in strengthWith deficient calcium, increase in flexibility, decrease in strength
Beary 1969
Dog Diet, 1, 3, 9, 27 ppm, 287 days With low calcium – high phosphorous,
no radiologic effects, decreased mineral mass in mandibles, no effect on
bending and tension tests
Henrikson et al. 1970
Mice Water, 10 ppmfrom birth Slight reduction in age-related
decline in breaking strength
Rao et al. 1972
Rat, young 45, then 135 ppm in water during
Increased radiographic density
if calcium adequateDecreased radiographic density if calcium deficient
Reddy et al. 1972c
Rat, adult Same as above Slightly increased radiographic
Quail Diet, 750 ppm, 35 days Increased radiographic density
if calcium deficientReduced breaking strength of femurs
Chan et al. 1973
Rat Increased radiographic density Erickson and Ekberg 1975
Rat Water, 50, 150 ppm Reduced breaking strengthReduced cross-sectional areaReduced modulus of elasticityCalcium, phosphorous deficiencyaggravated these effects Guggenheim et al. 1976
RatRoosterQuail Water, 100 ppm, 3 monthsDiet, 600 ppm, 4 monthsDiet, 750 ppm, 35 days Reduced breaking strength in osteogenic
boneReduced breaking strength in osteogenic boneReduced breaking strength in osteogenic bone
Riggins et al. 1976



During the past seven years, a number of research papers presented evidence that
some inorganic fluorides are mutagenic to plant and animal cells. These papers,
exclusive of those dealing with humans, are discussed in the following paragraphs.

Reference has been made in Section 2 to the studies of Bale and Hart (1973a, b)
on fluoride mutagenicity in Hordium vulgare (barley), and to the studies
of Gerdes et al. (1971b) on mutagenicity in Drosophila melanogaster (fruit
fly). As little as 10-6 M sodium fluoride (0.019 ppm F) caused chromosomal-bridges,
fragments, and gaps to develop during mitosis of cells in the root tips of barley
seedlings (Bale and Hart 1973a). In fruit flies (Gerdes et al. 1971b), 1.3 and
2.6 ppm of airborne fluoride, as hydrofluoric acid, caused genetic damage during
a six-week exposure.

Mohamed and Kemner (1970) and Mohamed (1971) have also demonstrated chromosomal
damage in D. melanogaster after exposing males of a specific genotype to
an unspecified concentration of hydrogen fluoride. The degree of chromosomal damage
increased with increasing exposure times from 6 to 12 hours. Mitchell and Gerdes
(1973) exposed adult D. melatogaster flies to fluoride by allowing them
to feed on filter paper strips saturated with a 7% sucrose solution containing
up to 6% sodium fluoride. Gene changes in chromosome X were then detected by a
cross-breeding technique. The percentage of sex-linked recessive lethals observed
was linearly related to the concentration of fluoride in the food (our plot),
and was significantly different from the control value at the two highest fluoride
concentrations (original authors’ calculations). Tests with stannous fluoride
gave similar results, except that the stannous salt was about one-third as mutagenic
(on a fluoride-equivalent basis) in comparison to the sodium salt.

Jagiello and Lin (1974) examined meiotic division stages in mammalian oocytes
that were exposed to fluoride, as NaF, in vitro. Chromosomal aberrations were
observed at the following concentrations of fluoride in the medium:


Mouse oocytes: 91 and 181 ppm

Ewe oocytes: 11, 23 and 91 ppm

Cow oocytes: 4.5, 11, 23 and 9 1ppm

The percentage of ewe oocytes undergoing division was sharply
reduced by fluoride concentrations above 23 ppm, and the percentage divisions
of cow oocytes was reduced at 4.5 ppm fluoride and above.

Mohamed and Chandler (1976) examined cells from the bone marrow and testes of
mice after exposure of the animals to fluoride in drinking water at levels of
0, 1, 5, 50, 100, and 200 ppm for up to six weeks. The number of chromosomal
breaks and abnormalities increased with the fluoride content of the drinking
water and with duration of exposure. The authors concluded that, with mice,
even 1 ppm fluoride in the drinking water (the diet contained 0.263 ppm) caused
genetic damage.

Russian researchers (Voroshilin et al. 1973, 1975; Gileva e,t al. 1972, 1975)
have presented data on the mutagenic effects of inorganic fluoride on bone marrow
cells of white rats. Inhalation of cryolite or of a cryolite + HF mixture (e.g.
271, 543 and 1628 ug F/m3 as cryolite, 6 hours/day, 6 days/week for up to 5
months) induced chromosomal aberrations and hyperploidy. The extent of damage
increased with increasing fluoride concentration. Damage was greater in 17-months
old than in one- to two-months-old rats.

Voroshilin et al. (1975) observed no increase in dominant lethals (total embryonic
deaths) when male mice were exposed to 1.0 mg HF/m3 for 2 or 4 weeks before
mating. On the other hand, Danilov and Kasyanova (1975) reported increases in
embryonic deaths resulting from exposure of female rats to 0.05, 0.5 and 5.0
mg HF/m3 (conditions of exposure are poorly specified).


The WHO monograph (1970) on “Fluorides and Human Health” repeatedly
states that the fluorine-carbon bond is not cleaved by biological processes.
However, evidence for biological cleavage of the C-F bond had been presented
in 1956 and confirmed in 1961 (cf. Ward and Huskisson 1972). Between 1965 and
1968, Goldman and his colleagues presented a series of papers on the bacterial
cleavage of C-F bonds, including that in 2-fluorobenzoic acid (cf. Harper and
Blakley 1971). During recent years, a considerable volume of additional data
on the biological cleavage of C-F bonds has accumulated. It is now apparent
that few, if any, organo-fluorine compounds are biologically stable.

Much of the recent interest in the biological breakdown of organic fluorine
compounds has arisen because there has been widespread use of some organohalides
as anesthetics. Under operating-room conditions, anesthetics must be considered
as atmospheric environmental pollutants. It has been shown that operating-room
pollution with fluorine-containing organohalides can give rise to an increased
urinary output of inorganic fluoride by operating-room personnel (Spierdijk
1972). Spierdijk (1972) also reviewed data indicating an increased “incidence
of abortion occurring amongst the wives of male anaesthetists, female anaesthetists
and nurse-anaesthetists” during employment. Exacerbation of subclinical
myasthenia gravis has been attributed to occupational exposure to methoxyflurane
(Elder et al. 1971).


Methoxyflurane (Penthrane), CH3-0-CF2-CC12H, appears to be one of the most biologically
unstable of the organohalide anesthetics. Research on methoxyflurane and related
anesthetics has been reviewed by Holaday (1972), Schuh (1974), Conn (1974),
and Gottlieb and Tray (1974). Biochemical pathways for the breakdown of several
of the organofluoride anesthetics have been reviewed by Loew et al. (1974).

Kidney damage can appear within a few days following methoxyflurane anesthesia.
This phenomenon was studied by Cousins and Mazze (1973), who reported that peak
(i.e. transient) post-anesthesia plasma F- levels in afflicted humans exceeded
90 umol/l. The nephrotoxicity was accompanied by an increased urine volume of
low osmolarity, and increased thirst, with the syndrome tending to obey a short-term
dose-response pattern in man. Mazze et al. (1972) and Cousins et al. (1974)
have shown that kidney damage in rats exposed to methoxyflurane was caused by
high inorganic fluoride concentrations and not by oxalic acid, which is also
a metabolic breakdown product of methoxyflurane. Taves et al. (1972) also related
the nephrotoxicity and polyuria to the metabolicallyreleased inorganic fluoride.
Mazze et al. (1972) showed that the degree of kidney damage was related to methoxyflurane
dose in rats.

Mazze and Cousins (1974) state that

“The predominant factor in the production of nephrotoxicity
following methoxyflurane administration appears to be the dosage; however, additional important factors are the rate of metabolism of the drug, renal sensitivity to inorganic fluoride, presence of enzyme induction, and interaction with other nephrotoxins. A high rate of methoxyflurane biotransformation to inorganic fluoride in a patient overly susceptible to fluoride’s renal toxic effects could result in the occurrence of marked nephrotoxicity from a relatively low dose of methoxyflurane”.

A high degree of individual variability has also been noted,
relative to other organofluorine anesthetics (see later) such as halothane (Cascorbi
et al. 1970). Also, Eichhorn et al. (1976) have postulated that “the threshold
for fluoride-induced nephrotoxicity (following enflurane anesthesia) may be
lower in diseased kidney”.

Hagood et al. (1973) reported that nephrotoxicity induced by methoxyflurane
anesthesia in clinical practice gave a calculated mortality rate of 50%.

The fluoride-mediated toxicity of methoxyflurane is influenced by the presence
of other drugs (Churchill et al. 1976; Cousins et al. 1974). Also, metabolic
breakdown of methoxyflurane is enhanced in obese patients (Young et al. 1975;
Samuelson et al. 1976) and is probably related to the retention of organofluorine
compounds in fatty tissue, as noted in studies with halothane (Bhoopathi et
al. 1974) as well as methoxyflurane (Bell et al. 1975).

In rats, low dose levels of methoxyflurane over long periods were more injurious
than were equivalent high-dose short-term treatments (Arthaud and Loomis 1975).
In rats treated with either methoxyflurane or inorganic fluoride, nephrotoxicity
was associated with plasma F levels of 20 umol/l (Roman et at,1977).

Creaser et al. (1974) and Clark et al. (1976) determined the plasma F- levels
of mothers and their newborn children when methoxyflurane was used as an analgesic
during labor. Measured values in the mother’s blood were somewhat higher than
in the neonate’s blood at birth (e.g. 23.1 ± 1.6 and 16.3 ± 1.4 umol/l, respectively
(Clark et al. 1976)). The plasma F level of the mothers’ blood declined progressively
after delivery, but was still markedly above normal 48 hours postpartum. The
infants’ blood lost fluoride less rapidly than the mothers’ blood, and the plasma
F level in infants was not significantly different from that of the mother’s
blood either 24 or 48 hours postpartum. Fiserova-Bergerova (1976) observed deposition
of fluoride in the bones of fetal rats after exposure of pregnant female rats
to enflurane or methoxyflurane.


Enflurane (Ethrane), CFzH-0-CF2-CFCIH, and sevoflurane, CFH2-0-CH(CH3)-CF3,
are also metabolized with consequent release of inorganic fluoride in the human,
but the extent of fluoride release is less than with methoxyflurane (Schuh 1974;
Loew et al. 1974; Cook et al. 1975; Carter et al. 1976). However, nephrotoxicity
was observed within 5 days following administration of enflurane in humans (Mazze
et al. 1.977). In these patients, the average (24 hr) plasma F- level was 15
umol/l, and there was no evidence of a no-effect threshold.

Other organohalides used to induce anesthesia,’such as fluoroxene (CF3-CH2-0-CH=CH2),
isoflurane (CF3-0-CClH-CF3) and halothane (CF3-CBrClH), release little inorganic
fluoride during oxidative metabolism (Loew et al. 1974; Gion et al. 1974.) (NOTE:
Metabolites other than inorganic fluoride compounds may, however, be toxic,
e.g. trifluoroethanol (Gion et al. 1974; Tucker et al. 1973; Fiserova-Bergerova
1977).) It appears that this relatively greater stability is attributable to
the bonding position of fluorine in these compounds (i.e. entirely on CF3 groups).
However, Hitt et al. (1974) noted that isoflurane is approximately one-tenth
as soluble as methoxyflurane, and suggested that the substrate concentration
in vivo may limit its metabolic degradation to inorganic fluoride. Hitt et al.
(1974) observed a release of inorganic fluoride from isoflurane by preparations
of rat-liver mitochondria in vitro, especially if the live rats had been preconditioned
(enzyme induction) by exposure to phenobarbital.

The fluoride of the CF3 groups in these compounds is released during metabolism
under hypoxic conditions in vitro (van Dyke and Gandolfi 1976) and in vivo in
rats (Widger et al.1976).


As mentioned above, inorganic fluoride is released from 2-fluorobenzoic acid
by bacterial metabolism. Inorganic fluoride is also released by bacteria from
p-fluorobenzoic acid (Harper and Blakley 1971). However, release of inorganic
fluoride from fenfluramine, which has a CF group in the ortho-position on the
benzene ring, has not been detected (Arvela et al. 1973; Macrae 1975).

Gerhards et al. (1971) reported that fluorine in the 6a-position of fluocortolone
(a corticoid drug) is released during metabolism of this compound in the human
body. Schellman and Zober (1975) observed an “abnormally-high” iliac
crest bone fluoride level (2,640 pg/g ash) in a patient following prolonged
treatment with fluorine-containing corticosteroids (see Sections 3.4.1 and 5.3).

Excretion of inorganic fluoride by rats after inhalation of various fluorinated
ethylene compounds has also been reported (Dilley et al. 1973).

Another class of organo-fluorine compounds that has aroused considerable environmental
interest involves conversion of inorganic to organic fluoride (i.e. formation
of C-F bonds). This conversion leads to the formation of monofluoro-organic
acids (e.g. fluroacetate) in some species of vegetation (NAS 1971; Ward and
Huskisson 1972). Studies published since our previous review (Marier and Rose
1971) have shown that both fluoroacetate and fluorocitrate are formed by cultured
soybean (Glycine max) cells (Peters and Shorthouse 1972b), and fluorocitrate
by cultured tea (Thea sinensis) cells (Peters and Shorthouse 1972a) when
10-3 M sodium fluoride is added to the medium.

No correlation between soil fluoride and organic fluoride in plant leaves has
been observed (Hall and Cain 1972); however, studies on vegetation subjected
to airborne fluoride still appear to be lacking for follow-up assessment of
biotransformation phenomena (cf. Marier and Rose 1971). Analysis for organo-fluoride
compounds should include plant tissues other than foliage, because Hall (1972)
detected high levels of organic fluorides in the seeds of some plant species.
Also, Vickery and Vickery (1972, 1975) studied the locale of synthesis and translocation
of fluoroacetate in Dichapetalum species. Conversion to long-chain fluoro-fatty
acids appeared to occur in the developing seeds.

It has been reported that the toxicity of fluorocitrate to rats is much lower
when it is administered orally than when injected intraperitoneally (Peters
et al. 1972). When 14C-labelled fluorocitrate was administered orally to rats,
most of the citrate moiety (73% of the 14C-label) was excreted in the urine
within 24 hours, along with considerable quantities of inorganic fluoride. The
toxicity of fluoroacetate, which is subject to in situ conversion to fluorocitrate,
was essentially unchanged whether it was administered orally or by interperitoneal
injection (Peters and Shorthouse 1971). One of the features of this poisoning
is the inhibition of aconitase, which results in an elevated blood citrate level
(Peters 1957).

In recent toxicological studies on the industrial use of BF3 catalysis, Bedford
et al. (1977) have isolated “two fluoroacetate precursors” (i.e. 2-fluoroethanol
and 2-fluoroethoxy-ethanol) which are apparently found as by-products, subsequent
to fluoride-ion release during catalysis.


There has been an increasing utilization of fluorine compounds by our technologically-oriented
society (EST 1972; Farkas and Parsons 1974). There has also been evidence of
an increased fluoride content in the human food-chain of several North American
localities (Prival and Fisher 1974; Kramer et al. 1974).

Studies of the interrelation between human illness and fluoride exposure are
largely dependent on uncontrolled exposure of industrial workers to variable
concentrations of fluoride, or on epidemiological studies in polluted neighborhoods.
Total fluoride exposure from all sources is usually not known. Accurate definition
of the dose-response relation is thus rarely possible, even under the specific
conditions of a particular study. Because factors that influence the nutritional
and health status of some individuals also affect their response to fluoride,
meaningful dose-response criteria having broad applicability to all humans are
presently unattainable.

Nevertheless, in the following discussion, we attempt to assemble and review
recent reports relating to fluoride-induced illness in humans. This includes
data on total exposure from all identifiable sources, and information on the
clinical and subclinical aspects of fluoride-induced injuries and on their detection.
Also, where possible, emphasis is put on the identification and fluoride-response
of segments of the total population who, for one reason or another, may be more
“at risk” than other segments of the population. In the absence of
criteria from which “safe levels” having a defined margin of safety
can be determined, observed injuries in such “at risk” groups may
provide an indication that exposure levels are approaching those that might
adversely affect an “average” individual exposed to various co-existing
sources of fluoride in man’s everyday environment.


5.1.1 Intake From Foods and Beverages

The amount of fluoride ingested daily from foods and beverages by humans has
been a subject of controversy during the period under review. The history of
one widely-quoted Table, which appears in the National Academy of Sciences (NAS
1971) and WHO (1970) monographs, has been reviewed by Farkas (1975a). Farkas
concludes that composite tables on fluoride intakes published prior to and during
the early 1970s were based on insufficient data and included misquoted data.
Having examined the original sources, we conclude that these Tables require
major revision.

Recent reviews on the intake of fluoride by humans include those by Jerard and
Patrick (1973), Prival and Fisher (1974), and Marier (1977). Two quotations
from these reviews illustrate the concern being caused by increased human exposure
to fluoride:

“Considering the paucity of recent data available on total fluoride intake, there is a clear need for more accurate, detailed information concerning the distribution of fluoride intake levels in the population” (Prival and Fisher 1974).

“Careful study is needed of the upward shift in environmental fluoride
and an effort should be made to appraise total exposure from all sources in order to protect the environment and people of varying vulnerability” (Jerard and Patrick 1973).

One of the major factors thought to be contributing to the
increase in human exposure to fluoride is the increasing fluoride content of
foods. Such an increase can arise from three main sources, namely, the use
of fluoridated water in food and beverage processing
, the exposure of crops
to airborne fluoride [and to water-borne fluoride in areas irrigated with fluoridated
water (Auermann 1973)], and the use of fluoride-containing fertilizers. Literature
on the contributions of fluoride from air, irrigation waters, and fertilizers,
to man’s total intake has been reviewed by Jerard and Patrick (1973). Bittel
and Vaubert (1971) state that:

“Il faut noter qu’en general la teneur en fluor des aliments vegetaux de l’homme ne cesse de croitre: il y a a ce fait plusieurs raisons, d’abord l’utilisation, de plus en plus grande, d’engrais et amendements calciques ou phosphocalciques de moins en moins purifies en vue de contenir suffisamment d’oligoelements (bore en particulier) et, d’autre part, l’influence de plus en plus etendue de la pollution industrielle”.

Not enough work has been done on food-chain infiltration
of fluoride, to assess the extent to which this contributes to total intake
from all sources. The data in Table 19 are presented to illustrate the influence
of the various environmental factors discussed above. The data are restricted
to those that have appeared since 1970, and are representative of the effects
of these environmental factors. The effect of fluoridated water used for food
processing is apparent from the data on Gouda cheese (Elgersma and Klomp 1975)
and beer (Tamacas et al. 1974) and is probably also a factor in the data on
baby formula (Farkas and Farkas 1974). The influence of airborne fluoride is
obvious in the high values for leafy vegetables reported by Gordon (1970a),
Jones et ciL (1971) and Vouilloz. The influence of fertilizer-borne fluoride
is less obvious, but the high values for “control” lettuce samples
reported by Gordon (1970b) are thought to have resulted from this source. The
high fluoride content of wheat, spinach, cabbage, carrots, and other Indian
foods (Lakdawala and Puneka1973) presumably results from uptake of soil- or
fertilizer-borne fluoride. The high value for Cola soft-drink in Bombay was
not caused by a high fluoride level in the city’s own water (Lakdawala and Punekar


Table 19. Recent data illustrating the effects of environmental factors on the range of fluoride content in some foods.

Food Explanation Fluoride content *(ppm, or mg/kg, or mg/l) Reference
Gouda cheese” “ NormalFluoridated, 1 ppm in processing water 0.27up to 2.16 Elgersma and Klomp 1975
BeersWines Fluoridated areas” “ up to 1.0up to 0.7 Tamacas et al. 1974
PablumBaby formulaOrange juice Ready to use” ” “” ” “ 4 to 120.9 to 1.00.9 Farkas and Farkas 1974
CabbageLettuce Exposed, washed (1)” “ 2.8 to 3.24 (2)12.0 to 19.6 (2) Jones et al. 1971
Fruits and vegetables Exposed up to 100 Vouilloz 1975
Lettuce” Control samplesExposed 24 to 8039 to 226 Gordon 1970b
Whole wheatWheat flourSpinachCabbageCarrotsCola drinks As used” “” “” “” “” “ 2.6 to 3.34.8 to 6.40.8 to 4.11.3 to 2.31.9 to 4.91.3 to 1.4 (4) Lakdawala and Punekar 1973 (3)
Cow’s milk” “ NormalExposed pastures 0.087 to 0.1320.287 Dirks et al. 1974
PorkCow beefChoice beef Mechanically deboned” “” “ 8.8 to 13.5 (5)30.4 to 41.7 (5)13.6 to 23.3 (5) Field et al. 1976
BeefPork Mechanically deboned” “ 9.8 to 16.27.6 Kruggel and Field 1977
* As reported by the various authors.Notes: (1) “Exposed” signifies exposed to airborne fluoride
pollution.(2) Mean values for various locations and times.(3) Selected items from an extensive table of foods from Bombay, India.(4) Bombay city water contained only 0.08 ppm fluoride.(5) Values for hand deboned products; pork – 1.7 to 3.2; cow beef – 3.1
to 3.5; choice beef – 1.6 to 3.2ADDENDUM: For a comprehensive review, see
Kumpulainen, J., and Koivistoinen, P. 1977. Fluorine in foods. Residue
Rev. 68: 37-57.

The data on mechanically-deboned
(Table 19) indicate how a change in processing procedure can alter the
trace element composition of a food. The high fluoride content of the mechanically-deboned
meats results from the inclusion of bone chips, and hence is related to the fluoride
intake of the animals (Kruggel and Field 1977). Contamination of forages by fluoride
increased the fluoride content of milk by 0.15 to 0.19 ppm (Dirks et al. 1974);
however, it must be remembered that reconstitution of powdered milk with fluoridated
water will cause a much larger increase in the fluoride content of the milk beverage.
The same applies to reconstitution of frozen concentrated fruit
and other such products.

A survey conducted by Farkas and Parsons (1974) indicated that the use of fluoridated
water for food and beverage processing in Canada is extensive. Specific production
data are not available, but 50% of the cities in which breweries were located,
and 43% of the cities where carbonated beverages were processed in cans, had fluoridated
water. Similarly, 51% of the factories processing vegetables, and of those processing
pasta products and soups, were located in cities with fluoridated water. Thus,
it is probable that about 50% of the beverages, pasta products, and canned foods
consumed by Canadians, contain about 0.5 ppm more fluoride (Marier and Rose 1966;
SanFillipo and Battistone 1971) than the same products contained before the cities’
waters were fluoridated.

Cigarettes may be another significant source of fluoride intake by humans. Okamura
and Matsuhisa (1965) reported the following results for fluoride content of cigarettes:


Type ofCigarette
No. of BrandsAnalyzed
ppm F in Cigarettes
ug F perCigarette(Average)
42 to 640
35 to 420


This is the only published report we have seen on the fluoride content of cigarettes.
Although Cecilioni (1974) mentions cigarettes as a source of fluoride, we are
not aware of any published North American study in which the contribution of
cigarette smoking has been seriously considered relative to total fluoride intake.

A study by Full and Parkins (1975) raises the possibility that Teflon-lined
cookware may contribute to the fluoride ingested by humans. Full and Parkins
boiled fluoridated (1 ppm) water at a moderate rate until a one-third or one-half
reduction in volume was attained, then determined the fluoride content of the
residual water by ion specific electrode. In aluminum ware, waterborne fluoride
concentration was decreased. In stainless steel and Pyrex ware, fluoride ion
concentration increased, but to a lesser degree than expected on the basis of
volume reduction. In Teflon-coated ware, the concentration of fluoride ion increased
to nearly 3 ppm. This result requires confirmation; but, if it is correct, then
the release of fluoride into foods during cooking in plastic-coated wares requires

In their study of English teenagers, Hardwick and Ramsey (1976) estimated that
the mean daily intake of fluoride from dentrifice was 0.32 mg, with an extreme
high of 5.0 mg. Three children out of 274 “usually swallowed all of the
dentifrice used”. Fluoride tablets (1 mgF/ tablet) are also available in
the U.K., and Hamilton (1974) notes that, in spite of a cautionary statement
on the package, “the sale of fluoride tablets does not appear to be related
to the fluorine content of the drinking water”.

Recent data on the amount of fluoride ingested by children and adults are summarized
in Tables 20 and 21, respectively. However, few, if any, of these data are all-inclusive
estimates of total fluoride intake. None of the data include fluoride intake
from dentifrice, or from cigarettes. Kramer et al. (1974) state that their data
are “exclusive of drinking water”. The “market basket” approach
would appear to underestimate the effect of fluoridated water used for domestic
and commercial food preparation. In describing this approach, Cummings (1966)
listed only five canned foods (fish, corn, pork-and-beans, tomatoes, and peaches).
SanFillipo and Battistone (1971) list only one canned food (pork-and-beans).
Neither author lists any canned soups. This does not appear to be realistic;
Kramer et al. (1974) state that, in their survey the hospital diets studied
“in great part consist of processed, canned foods …” Cummings (1966)
indicates a “market basket” allowance of only 1800 g (about 63 oz)
of soft drinks for a two-week period, which probably does not allow for between-meal
consumption of such beverages. Lakdawala and Punekar (1973) discuss the fluoride
content of carbonated beverages, but it is not clear whether between-meal beverages
were included in their calculated fluoride intakes, as quoted in Table 20.

The intakes for adults reported by Jenkins (1973) may be somewhat atypical,
because they are estimated on the basis of urinary excretions of 4 to 5 mg/day
by people who drink large amounts of tea.

Probably the most all-inclusive data in Table 21 are those of Spencer et al.
(1970), which were obtained under hospital conditions. There is no reason to
assume that the hospitals diets were selectively high in fluoride.

In Table 22, data on the percentage contribution of low-fluoride water and various
foods to the fluoride intake are compared for India and North America. The marked
differences shown by these data reflect the low estimates of fluoride intake
from food that were commonly accepted in North America during the 1960s, especially
relating to “baseline” estimates in unfluoridated areas.

Until further data become available we recommend that statements relating to
fluoride intakes by adults in North America should assume a “from foods”
fluoride intake of 1.5 to 2.75 mg/day, and an intake from “foods and beverages”
(in areas with water fluoridated at 1 ppm) of 3.5 to 5.5 mg/day. Such estimates
should include the caution that these intakes may be exceeded by persons exposed
to hot environments, by copious tea drinkers, and by individuals with polydipsia
(excessive thirst).

Table 20. Recent data on the daily intake of fluoride by children.

Fluoride intake, mg/day
Age Locality with1 ppm F in water Locality withlow F in water Reference
School 1.0 to 2.15 (1) Balazova et al. 1970
Infants 1.48 to 1.90 Ericsson et al. 1972
0-1 year9-11 year15-18 year 0.37 to 1.291.602.25 0.11 to 0.450.891.07 Auermann 1973
2-8 year 2.7 1.6 Jenkins 1973
under 12over 12 0.74 to 2.01.21 to 2.71 Lakdawala and Punekar1973
5-6 year14 year 0.85 to 1.11.06 to 2.10 Lee 1975
1-4 week6-8 week3-4 month4-6 month 0.320.571.021.23 Wiatrowski et al. 1975
(1) The high value refers to children living in an area exposed to airborne
fluoride. “Food” contributed 1.4 mg in this area as compared to
0.8 mg in a “control” area.



Table 21. Recent data on the daily intake of fluoride by adults.

Fluoride intake, mg/day

1 ppm F in water
Low F in water
3.57 to 5.37
1.45 to 2.74
Spencer et al. 1970
2.1 to 2.4
0.8 to 0.9 (“food stuff” only)
San-Fillippo and Battistone 1971 (1)
0.81 (without exercise)1.20 (moderate exercise)1.98 (strenuous excercise)
Auermann 1973
7 to 10 (heavy tea drinkers)
Jenkins 1973
1.73 to 3.44
0.78 to 1.03 (exclusive of beverages)
Kramer et al. 1974
1.23 to 2.41
0.73 to 0.94 (three meals only)
Osis et al. 1974.
(1) A second report by San-Fillippo et alo. (1972) has been omitted from
this table as it refers exclusively to military meals.


Table 22. The percentage contribution of water and various foods to the fluoride ingested by humans.

% contribution of F
India (Lakdawala and Punekar 1973)

Water (0.083 ppm)

2.0 – 10.8


0.7 – 21.0


16.0 – 52.0


9.0 – 45.0


6.0 – 22.0


0.3 – 15.0
U.S. (NAS 1971, Table 9-4)

Water (< 0.1 ppm)


All foods



5.1.2 Intake From Air

It is generally assumed that an “average” man doing moderately strenuous
work inhales approximately 20 m3 of air in a 24-hour period (NAS 1971). In view
of the known relations between fluoride ingestion and urinary excretion (NAS 1971),
the rapid response of urinary fluoride to inhaled gaseous or particulate fluoride
(Hodge and Smith 1977) is indicative of efficient and probably essentially-complete
absorption of inhaled fluoride into the body. Inhalation of air containing 0.1
ug fluoride/m3 [a level that is rarely encountered in non-industrial urban areas
(Thompson et al. 1971)] would thus result in an intake of only 2.0 ug fluoride
per day. The contribution of airborne fluoride to the daily intake is therefore
considered to be negligible (however, see below) by most authorities (e.g. NAS

However, at the other end of the scale, Hodge and Smith (1977) appear to consider
it acceptable to expose workmen, during an 8-hour shift, to a fluoride concentration
of 2.5 Mg/m3. Assuming the respiration of 10 m3 of air during a working shift
(Dinman et al. 1976a), fluoride absorption from the air by workers exposed to
this concentration could approach 25 mg. In summer, potroom workers perspire an
average of 6 kg sweat per day and may thus excrete 25 to 50% of the ingested fluoride
in sweat (Dinman et al. 1976a). Nevertheless, post-shift urinary fluoride concentrations
of 7.01 ± 0.47 to 8.65 ± 0.69 mg/l (daily averages for 25 “anode-men”
exposed to an average airborne fluoride concentration of 2.19 ± 0.16 Mg/M3) have
been reported (Dinman et al. 1976b). These data clearly indicate that, under some
circumstances, humans can receive a considerable amount of fluoride from airborne
sources (see Section 5.9.2).



Fluorides are known to cause chromosome damage and mutations in plant and animal
cells (Section 3.5) and might therefore be considered as possible carcinogens.
The majority of studies on correlations between fluoride exposure and deaths from
various causes, including cancer, have focussed on fluoride exposure via drinking
water. Studies of differences in the “crude cancer death-rate” between
cities with nonfluoridated and fluoridated water supplies have led to conflicting
results (Nixon and Carpenter 1974; Bierenbaum et al. 1974; Kinlen 1974, 1975;
Burk 1975; Yiamouyiannis and Burk 1976, 1977; Hoover et al. 1976).

An excess of respiratory-tract cancers was reported in fluorspar miners in Newfoundland
and attributed to airborne radon and radon daughters (deVilliers and Windish 1964;
deVilliers et al. 1971). Since the rate of incidence in these miners was five
times greater (per-unit of radiation exposure) than in Colorado uranium miners,
Little et al, (1965) postulated a co-carcinogenic role for fluorspar.

Cecilioni (1972a, b) has drawn attention to a three-fold increase in the death
rate from respiratory cancer in the steel city of Hamilton, Ontario, in comparison
with the Canadian average.

Studies of steelworkers (Lloyd et al. 1970) and aluminum workers (Discher et al.
1976; Discher and Breitenstein 1976; Milham 1976) have led to inconclusive results
concerning the effect of the work environment on respiratory health. In all the
studies involving industrial exposures, no distinction can be discerned among
various toxic and possibly carcinogenic factors in the work environment. Such
factors can act independently or synergistically. Until definitive studies involving
specific exposures to fluorides have been made, we can draw no conclusions about
the carcinogenic or co-carcinogenic activity of fluoride.


Hodge and Smith (1977) have recently reviewed occupational exposure to fluoride,
as it relates to aluminum or phosphate fertilizer production, with emphasis on
clinical osteosclerosis. Other metabolic irregularities (e.g. those affecting
respiratory function, arthritis, kidney, blood, etc.) are considered, but Hodge
and Smith (1971) state that, in general, “their relation to fluoride exposure
is doubtful, unless the exposure conditions exceed those typical of U.S. operations”
However, it is important to recognize that there is usually a preemployment selection
of workmen on a health basis. Thus Franke et al. (1975) have reported that

“At the examination for employment, persons with the following diseases are considered totally unsuitable: liver and kidney changes, blood and thyroid gland diseases, post-traumatic or congenital skeletal
damage, infections and para-infectious diseases of the apparatus of locomotion (rheumatism; Bechterev’s disease); also, workers with distinct degenerative changes of the spine and of the large joints are unsuitable”.


In all probability, there is also a continuing selection
for health among exposed workers (cf. Lloyd et al. 1970). Thus, at least some
smelters exercise a “selection of the fittest” policy, thereby ensuring
that the workmen are in good health and, as such, more likely to tolerate exposure
to fluoride. In spite of this, Guminska and Sterkowicz (1975) and Schellmann
and Zober (1975) have recently emphasized that fluoride intoxication is a problem
of increasing importance in technological countries.

Although emphasis in North America has been on the osteosclerotic manifestations
of occupational fluorosis, earlier stages of fluoride-induced
changes in bone are now being utilized as diagnostic aids by some researchers.
Franke and Auermann (1972) have described this procedure, and Horn and Franke
(1976) have demonstrated how microscopic scanning techniques can be used to
recognize graduated bone changes associated with mild to severe fluorosis. Franke
and Auermann (1972), and Schlegel (1974) emphasized that muscular-skeletal complaints
can be related to the histological bone changes of mild fluorosis. Moreover,
Popov et cit. (1974) observed neurological symptoms in 63 of 80 workmen examined,
and noted that the incidence of neurological symptoms was not related to the
skeletal stage (whether pre-osteal or definite osteal) of fluorosis. Hiszek
et al. (1971) emphasized that the high incidence of locomotor ailments caused
by fluoride occur in the absence of obvious radiological evidence of fluorosis.

Other recent observations indicate that fluoride-induced bone changes are not
necessarily symmetrical or bilateral. Thus, Herbert and Francon (1971) describe
the case of a Potroom worker who had left-hip sciatica and nephritis, with diffuse
lumbar arthralgias. The fluoride content of the iliac crest bone was between
5,100 and 5,800 ppm, ash basis. Harbo (1973) describes the case of a workman
with sensory loss in the upper left extremity, with muscular wasting and pain.
The highest fluoride content found in vertebral bone samples was 2,700 ppm,
ash basis.

In a discussion of bone fluoride levels, Riggins et al. (1974) report that some
researchers feel that 2,000 ppm fluoride in dry fat-free bone “should be
considered toxic” and that “skeletal fluorosis in humans can be seen
when the concentration of fluoride in bone ash exceeds 3,000 ppm”. These
two values for bone fluoride are compatible with each other, assuming that bone
contains about 60% ash. Franke and Auermann (1972) have concluded that “In
cases of genuine violent complaints, clear histological changes, and fluorine
values above 4,000 ppm in the bone ash of the iliac crest cylinder, the disease
should be classified as an occupational one, even with few clinical or radiological
findings”. Boillat et al. (1976) advocated bone-biopsy fluoride analysis
as a diagnostic aid in the case of workmen with “articular pain and limitation
of motion”; these cases had concomitant hypocalcaemia, hypocalciuria, and
hyper-hydroxyprolinuria, and Boillat et al. concluded that nutritional factors
play an important role in such afflictions.

These various observations indicate that the diagnosis of fluoride-related ailments
is in a state of evolution, and is approaching a hitherto unknown degree of
thoroughness and sensitivity.


Respiratory ailmentsmay also be related to occupational exposure to fluoride.
Mangold and Beckett (1971) observed an “immediate upper respiratory irritation”
by fluorides, as contrasted to a “delayed pulmonary response to cadmium
oxide fumes and nitrogen dioxide” among silver brazers exposed to the mixed
fumes. This suggests that respiratory ailments may not reflect the total body-burden
of fluoride, but might accrue from repeated localized contacts of fluoride (especially
HF) with respiratory tissues. It is interesting to note that, after intravenous
infusion of Na 18F into rats, the “lungs contained the greatest amount
of fluoride” of any of the soft tissues (Knaus et al. 1976).

Golusinski et al. (1973) reported that, of 130 potroom workers in an aluminum
smelter, 30% had the characteristic changes of rhinitis, with hypertrophic and
atrophic lesions. Similarly, Fesenko et al. (1972) found that, of 1,141 workmen
examined, 36% had skeletal fluorosis, and 10% of these also had rhinopharyngolarynqltis.
The rhinopharyngolaryngitis topic is one that Hodge and Smith (1977) consider
worthy of consideration for future research on fluoride effects. As for atrophic
rhinitis, Brown et al. (1966) have studied its etiology, pathogenesis, and prophylaxis
in swine; among their conclusions, the authors emphasized that an inadequate
calcium ingestion (or a low dietary Ca/P ratio) leads to nutritionally-induced
secondary hyperparathyroidism and consequent generalized osteitis fibrosa. Thus,
we again note the contribution of nutritional inadequacies (or imbalances) in
such syndromes.

The diagnostic value of plasma inorganic fluoride determinations has been discussed
elsewhere in this report (Section 3) and by Marier and Rose (1971), Ericsson
and Ekberg (1975), and Inkovaara et al. (1975). Guminska and Sterkowicz (1975)
have reported a significant decrease in blood erythrocyte ATP in workmen exposed
to fluorides. Nikolaev et al. (1971, 1973) have drawn attention to a 16-to-30%
reduction in blood manganese among workers exposed to fluorides. Although Furlanetto
et al. (1973) concluded that “manganese seemed not to affect the proportional
fixation of fluoride” in bones and teeth, researchers must not lose sight
of the fact that a dietary lack of manganese can induce skeletal abnormalities,
including generalized rarefaction of bone (Tal and Guggenheim 1965), thickened
leg joints with stunted growth, and impaired reproductive function (NAS 1973).
Rao and Friedman (1975) have reported that “A further toxic effect of fluoride
on bone formation may relate to the fluoride bonding with manganese, a cation
necessary for glycosylation, an intermediary step in the formation of collagen”.

The foregoing examples, along with the blood and tissue changes noted in Tables
15, 16, and 17, attest to the need for consideration of a multiplicity of factors
in the assessment of injuries accruing from long-term exposure to fluorides.
The metabolic changes discussed cannot be assumed to be of no biological significance,
as regards chronic fluoride intoxication.

Hodge and Smith (1977) concluded that a workplace airborne fluoride concentration
below 2.5 mg/m3 “will be tolerated without injuring human health during
a working lifetime”. In contrast, Vishnevski (1969) questioned the U.S.S.R.
airborne limit of 0.5 mg/m3 because workmen were found to exhibit increased
sensitivity to light, increased toxic symptoms and increased skeletal incorporation
of fluoride. Vishnevski concluded that the occupational fluoride level in ambient
air should not exceed 0.1 mg/m3. Guminska and Sterkowicz (1975) expressed concern
that occupational airborne fluoride concentrations of 0.22 mg/M3 would be deleterious
to workers’ health (see Section 5.9.2).


As discussed in Section 5.3 on occupational fluorosis, screening of workmen
to assure a reasonable health status (Franke et al. 1975) undoubtedly reduces
the incidence of overt fluoride-related complaints. No such protection is provided
for people residing in areas adjacent to fluoride-emitting industries. Hodge
and Smith (1977) considered this aspect of fluoride pollution, but additional
discussion is warranted. It is indisputable that persons exposed to fluorides
in the workplace constitute an “at risk” group because of the high
concentrations of fluoride to which they are exposed during their working shift.
It is less widely recognized that persons residing
in polluted areas
also constitute an “at risk” segment of the
population because of their more continuous exposure to moderate concentrations
(cf. Hunter 1969). It must not be forgotten that children, the elderly, and
the chronically ill and infirm, all form part of populations residing adjacent
to sources of fluoride emissions. Table 23 summarizes some of the metabolic
abnormalities that have been observed in such persons.

In Table 23, note that the “joint pains” alluded to by Murray
and Wilson (1946)
and the “neuromuscular arthritis” described
by Waldbott and Cecilioni (1969) are
similar to symptoms that were discussed in our preceding comments about occupational
fluorosis (Hiszek et az. 1971; Franke and Auermann 1972; Harbo 1973; Popov et
al. 1974; Schlegel 1974; Boillat et al. 1976). Riggs and Jowsey (1972), in their
studies of fluoride therapy for osteoporosis in humans, observed that some patients
developed “transient arthralgias and stiffness of the joints. These symptoms
are dose-dependent and promptly disappear on discontinuation of the drug (i.e.
fluoride)”. The occurrence of anemia in “neighborhood fluorosis”
accords with earlier observations (cf. Marier and Rose 1971) and with the related
data in Table 16, as discussed in Section 3.1.2.

In a preceding Section of this report, Tables 15, 16 and 17 presented summaries
of biochemical changes induced by fluoride in blood and in soft tissues. It
remains to be determined whether some of these changes are consistent features
of subclinical or mild forms of neighborhood fluorosis in humans.

Table 23. Health problems among residents near fluoride-emitting sources.

Place Source Population Symptoms Reference
England Iron 5 adults, 4 children Murray and Wilson 1946.
Czechoslovakia Aluminum 78 children surveyed Low hemoglobin withhigh erythrocyte Rippel et al. 1967 Balazova et al. 1970
Canada-U.S. Phosphate Fert. 31 adults Neuromuscular arthritis etc. Waldbott and Cecilioni 1969.
U.S. Iron 1 adult Neuromuscular arthritis etc. ” ” “
Hungary Aluminum 227 children surveyed Low hemoglobin Leloczky 1971
E. Germany HF plant 27 adults Early skeletal fluorosis Schmidt 1976a
E. Germany Aluminum 16 adults Bone changes Schmidt 1976b



This form of fluorosis has been linked to chronic ingestion of naturally-fluoridated
waters (WHO 1970). In a recent Algerian study of hydrofluorosis,
Poey et al. (1976) have reported that the early stages of chronic fluoride intoxication
are associated with changes in blood and urine components, and that these precede
radiologically-detectable bone abnormalities. In the early phase, there was an
increase in blood urea and acid phosphatase, with a concomitant increase in urinary
output of phosphorus and urea. As the fluoride intoxication progressed, there
was a gradual impairment of urinary creatinine clearance, leading to renal insufficiency
(see Sections 3.2.1, 3.2.2, and 5.8).

Much of the information relating to endemic fluorosis has originated from India,
where skeletal fluorosis has been associated with water-borne fluoride concentrations
of 2 to 3 ppm or lower (Jolly et al. 1968; Krishnamachari 1976). Although osteosclerosis
seems to be the only fluoride-related bone abnormality recognized in North America
(see Hodge and Smith 1977), the skeletal abnormalities observed in India are not
confined to the osteosclerotic form (Teotia et al. 1976).

Even at comparable degrees of fluoride exposure, the epidemiological studies in
India have provided some striking contrasts. In the Punjab area, Jolly et al,
(1968, 1974) have invariably observed the osteosclerotic type of bone disease
in fluorotic patients. In contrast, Teotia et al. (1974, 1976) have encountered
osteoporosis, rachitis, and the osteomalacia type of bone disease associated with
a fluoride-induced compensating secondary hyperparathyroidism. The bone rarefaction
phenomena observed by Teotia et al. were not confined to adults, but were common
in children 11 to 14 years of age (Teotia et a. 1971). Faccini and Teotia (1974)
described the histopathological features of the osteomalacia-like fluorotic bone.
This abnormality can resemble the osteitis fibrosa cystica of the “wine fluorosis”
described by Soriano (1966). It also resembles the condition reported in fluoridated
hemodialysis patients by several researchers (Posen et al. 1971; Jowsey et al.
1972a; Johnson and Taves 1974; Riggs et al. 1976).

In Teotia’s surveys, the serum immuno-reactive parathyroid hormone levels correlated
positively with serum alkaline phosphatase and with urinary excretion of total
hydroxyproline (Teotia et al. 1974; Faccini and Teotia 1974). In studies of hydrofluorosis
in Italy, Frada et al. (1974) observed increases in bone alkaline phosphatase
and urinary hydroxyproline in fluorotic patients. These observations are in accord
with those of Boillat et al. (1976) who reported hydroxyprolinuria in patients
with occupational fluorosis. Boillat et al. concluded that nutritional factors
play an important role relative to fluorosis-related hydroxyprolinuria.

Jolly et al. (1974) discussed nutritional factors relative to the different clinical
patterns seen in different regions of India. They state that “In Punjab,
where the (daily) dietary intake of calcium averages 1 gram, osteomalacia and
rickets are not encountered in cases of fluorosis. However, in Andhra Pradesh
and Rajasthan, a low calcium intake coupled with intake of fluoride produces changes
of rickets and osteomalacia”. This conclusion is supported by other reports.
Thus, in the Rai Bareli district, where osteomalacia is the commonly-seen form
of bone fluorosis, Teotia et al. (1974) reported a daily calcium intake averaging
645 mg, and a phosphorus intake averaging 1738 mg/day (Ca/P ratio = 0.37). Krishnamachari
and Krishnaswamy (1974) reported that the adult male in the Nalgonda district,
where Genu VaIgum (see next paragraph) is the prevalent form of fluorosis, has
an average daily intake of 297 mg calcium and 2096 mg phosphorus (Ca/P ratio =

An extremely severe form of fluorosis observed in India is termed “genu valgum”,
and is characterized by a crippling “knock knees” syndrome with osteosclerosis
of the spine and concomitant osteoporosis of the limb bones, and by very high
serum parathyroid hormone levels suggestive of hyperparathyroidism (Krishnamachari
and Krishnaswamy 1973). Dietary studies indicated no vitamin D deficiency, but
low dietary calcium, a low Ca/P ratio, and a high molybdenum content in some locally-grown
foods; a high urinary excretion of copper was also noted (Krishnamachari and Krishnaswamy
1974; Agarwal,1975; Krishnamachari 1976). The crucial role of copper was recognized
by Krishnamachari (1976) who states that “None of the villagers whose water
contained more than 0.1 ppm of copper had Genu Valqum, although their water contained
high levels of fluoride”.

These studies on Genu Valqum indicate that water-borne elements other than fluoride
can influence the skeletal abnormalities encountered in endemic fluorosis. Several
reports have shown a relation between the development of fluorosis and the calcium
and magnesium content of the drinking water. Jolly et al. (1968) reported on the
situation in two villages whose water contained an average of 3.3 ppm fluoride,
but where the two populations had a markedly different (10% as compared to 45.6%)
incidence of skeletal fluorosis. The lower incidence of fluorosis was associated
with a higher total hardness of the water. Because the nutritional status, climate,
duration of fluoride exposure etc., did not differ between the two villages, Jolly
et al. (1968) concluded that the calcium and magnesium components of hard water
had a “protective influence”. Such a protective effect has been discussed
by Marier, Rose and Boulet (1963).

Similar results have been reported from the Rajasthan area of southern India,
where Thergaonkar and Bhargava (1974) compared fluoride intoxication in 16 villages
with waters of different degrees of hardness and fluoride contents ranging from
0.3 to 2.7 ppm. They concluded that “incidence of fluorosis is probably reduced
by (waterborne) calcium … ard the severity … is (directly) related to bicarbonates
in the water, apart from the fluoride concentration”. These conclusions are
supported by the data of Kathuria et al. (1974). Teotia and Teotia (1975), in
a study in the Uttar Pradesh area of India, did not find a reduced incidence of
fluoride intoxication in hard-water areas, and suggested that “the simultaneous
intake of excessive amounts of (waterborne) magnesium … interferes with calcium’s
(protective) action”.

No clear relation between waterborne magnesium and fluorosis was observed in the
studies conducted by Thergaonkar and Bhargava (1974) or by Kathuria et al. (1974),
but the assessment may have been complicated by “low nutritional levels and
lack of a balanced diet” (Thergaonkar and Bhargava 1974). A similar survey
of human population groups in Czechoslovakia (Vejrosta et al. 1975) attributed
a beneficial effect to waterborne magnesium, in terms of ensuring the integrity
of mineralized tissues. In Rumania, Benetato et al. (1970) studied calcium and
magnesium metabolism in hospitalized patients who had neurological symptoms of
early (i.e. pre-skeletal) chronic fluorosis, associated with ingestion of drinking-waters
containing 2.85 to 3.6 ppm fluoride. The study included parallel observations
in rats, and led the authors to conclude that chronic fluoride intake can induce
latent calcium and magnesium deficiency in which the electrolyte changes (especially
of magnesium) contribute to the serious metabolic derangements. A report of the
Royal College of Physicians of London (1976) concluded that “There is no
evidence that the consumption of water containing approximately 1 mg/litre of
fluoride in a temperate climate is associated with any harmful effect irrespective
of the hardness of the water.”

The role of drinking water components should not be underestimated. Hankin et
al. (1970) found that hard waters can contribute significantly to the total dietary
intake, i.e. from 3.5 to 15.9% of the daily intake of calcium, and from 8.9 to
27.3%0 of the daily magnesium. During recent years, the World Health Organization
has been emphasizing this area of research (Masironi 1975). Sundstrom (1972) observed
bone-resorption cavities indicative of “mild fluorosis” in some rats
given 1 ppm fluoride in distilled drinking-water during a 2-year period, and therefore
recommend “A special long-term study, in which the effects of (fluoride in)
distilled and artificially-fluoridated waters are compared with those of naturally
fluoride-containing waters”.



In the preceding discussion, we have considered the influence of waterborne calcium
and magnesium, and how this factor may help to protect against the onset and severity
of fluorosis.

The beneficial effects of calcium and magnesium in alleviating fluorosis has been
confirmed in animal studies. Low-calcium diets increased bone fluoride in rats
(Guggenheim et al. 1976), increased the severity of bone fluorosis, with “exostosis”
lumps, in rabbits (Reddy and Rao 1972b), and increased bone fluoride and the severity
of its effects in monkeys (Reddy and Srikantia 1971). Conversely, high dietary
calcium and phosphorus lowered bone fluoride in swine (Forsyth et al. 1972), and
calcium supplementation decreased the lesions of fluorosis in cows, horses,
and swine (Spencer, Cohen and Garner 1974).

Marier (1968) reviewed the metabolic interrelation of magnesium and fluoride;
Table 24 represents his summary comparison. Marier notes that, in magnesium-deficient
dogs, fluoride supplements prevented soft-tissue calcification, but not the muscle
weakness or convulsions; in magnesium-deficient rats, fluoride aggravated the
hypomagnesaemia, thereby intensifying the convulsive seizures.

Rapidly growing chicks appear to present a particular problem, because they develop
a “leg weakness” syndrome when fed diets that contain high levels of
both magnesium and fluoride (cf. Marier 1968). Also, Rogler and Parker (1972)
have observed that a diet high in calcium could partially prevent the onset of
toxicity with the high-magnesium high fluoride diet. An underlying imbalance among
the various mineral nutrients is thereby suggested. Hakansson and Svensson (1977)
have reported that rapidly growing chicks, especially when given highly concentrated
feed, seem to have difficulties in utilizing dietary magnesium, and retaining
it in their leg bones. It appears probable that this magnesium-related problem
is aggravated by concomitant high-fluoride supplementation, even though the toxic
symptoms can be partially alleviated by prior increases in dietary calcium.

Hamuro (1972a, b) studied the effects of fluoride on magnesium-deficient mice
and concluded, on the basis of 6-day studies, that fluoride supplements prevented
renal calcinosis. However, a longer-term study with magnesium-deficient rats indicated
(Ophaug and Singer 1976) that fluoride exerted only an initial protective effect
on kidney calcinosis, and that the long-term effect was to promote kidney calcification.
Suketa et al. (1977) observed that, in rats, a single large dose of fluoride increased
kidney calcium 10 times more than it increased kidney magnesium; this would favor
in situ calcification (Marier 1968).

Pita et al. (1972) have shown that fluoride supplements increased the magnesium
content of mineralized tissues in rats. Ophaug and Singer (1976) reported that
fluoride exerted a significant effect in retarding the mobilization of skeletal
magnesium in rats. O’Dell et al. (1973) observed that fluoride had a “magnesium-sparing”
effect in Guinea pigs, but found that high fluoride supplements were toxic when
magnesium was severely limiting. O’Dell et al. concluded that “a high intake
of magnesium should be highly beneficial in areas where fluorosis prevails”.

Thus, there is evidence that fluoride intake can increase the long-term metabolic
requirement for magnesium. The same may be true for manganese. Note that we have
previously discussed depletion of manganese in fluoride-polluted pine needles
(Garrec et al. 1977), and the reduction in blood manganese among workers exposed
to fluorides (Mikolaev et al. 1971, 1973). These nutritional interrelations have
not yet been adequately quantified.

The same considerations apply to vitamin C. Unlike most species, primates cannot
synthesize their own vitamin C, and are entirely dependent on their food-chain
to supply an adequate intake. In a study of fluoride supplementation in monkeys,
Reddy and Srikanti (1971) showed that a diet low in vitamin C enhanced the onset
of skeletal fluorosis, and that a low protein intake accelerated rarefaction of
bones. Earlier, Gabovich and Maistruk (1963) had shown that vitamin C supplementation
reduced the toxic effects of fluoride in industrial workers and in Guinea pigs.
Marier and Rose (1971) discussed Russian studies in which fluorosis was found
to be most severe in children who had a vitamin C deficiency. Marier and Rose
also discussed Australian work, which showed that vitamin C supplementation alleviated
fluorosis in Guinea pigs.

It appears possible that chronic exposure to fluoride increases the metabolic
requirement for vitamin C; but again, such nutritional interrelations have not
yet been quantified.

There is, however, definite evidence that fluoride supplementation creates a greater
metabolic requirement for calcium in humans. Much of this evidence has accrued
from attempts to treat human osteoporosis by means of high doses of fluoride.

Some researchers (e.g. Franke et al. 1974) have reported success in the treatment
of human osteoporosis, using 20 to 60 mg NaF per day (i.e. 9 to 27 mg F/day).
However, experiments with several species of animals have shown that administration
of fluoride alone does not reverse or improve osteoporosis (Henrikson et al. 1970;
Spencer et al. 1971; Reddy and Rao 1972a; Kuo and Wuthier 1975; Griffiths et al.
1975, 1976). Similarly, several researchers have concluded that administration
of fluoride alone does not improve human bone rarefaction (Albright and Grunt
1971; Cohn et al. 1971; Inkovaara et al. 1975). Jowsey et al.(1972b) have emphasized
that administration of less than 20.5 mg F/day did not consistently increase bone
formation, whereas 27 or more mg F/day produced abnormal bone. This form of high-fluoride
therapy has been termed “an experimental drug for osteoporosis” (Gordan

Using 25 and 20.5 mg F/day, respectively, Inkovaara et al. (1975) and Zanzi et
al. (1975) observed spontaneous bone
during the course of treatment. Merz et al. (1970), using dosages
usually ranging between 22 and 34 mg F/day, discontinued fluoride administration
to their patients, to avoid the eventual development of osteomalacia. (Note: Osteomalacia
and spontaneous bone fractures have also been encountered in patients on hemodialysis
with fluoridated water; see Section 5.8).

Studies with rats, swine, dogs, and monkeys, have shown that, in the absence of
fluoride supplementations, a calcium deficiency (or too low a dietary Ca/P ratio)
is likely to lead to “nutritional osteoporosis” (Henrikson 1968; Reddy
and Rao 1972a; Rantanen et al. 1972; Kuo and Wuthier 1975; Griffiths et al.. 1975
and 1976). The osteoporotic condition will not be reversed or improved by supplementation
with fluoride alone (Henrikson et al. 1970; Reddy and Rao 1972a; Spencer et al.
1971 and 1974; Kuo and Wuthier 1975). If the calcium insufficiency is not corrected,
fluoride supplementation can induce osteomalacia (Rantanen et al. 1972; Kuo and
Wuthier 1975; Griffiths et al. 1975 and 1976).

The most positive results in the treatment of human osteoporosis with fluoride
have been obtained by use of concomitant calcium supplements. In the studies by
Cohn et al. (1971), high calcium supplements reduced bone pain in osteoporotic
patients, whereas fluoride administration did not achieve this effect.

Jowsey et al. (1972b) have stated that “osteomalacia and secondary hyperparathyroidism
observed in previous studies were caused by fluoride and a calcium intake insufficient
to mineralize the new bone … Fluoride might be expected to aggravate any tendency
toward increased parathyroid hormone secretion in osteoporosis”. Kyle et
al. (1975) commented that “in the absence of additional calcium, the bone
is incompletely mineralized. If fluoride administration continues … the net
result will be osteomalacia and increased bone resorption”.

To prevent osteomalacia, the calcium supplement must be “administered concurrently”
with fluoride (Riggs and Jowsey 1972). Jowsey et al. (1972b) and Kyle et al. (1975)
recommend that, in high-fluoride therapy, the calcium supplements, given concomitantly,
should be 35 to 40 times the fluoride supplement, by weight. Marier (1977) noted
that the calcium supplement is given in addition to the “adequate” calcium
levels ingested in a normal diet, and this is thus indicative of a fluoride-induced
increase in the metabolic requirement for calcium. If this same fluoride-to-calcium
proportionality applies to chronic daily intake of fluoride, then the ingestion
of 5 mg of fluoride per day would require a supplemental intake of 200 mg calcium
per day. This extrapolation may not be justified, but it serves to emphasize the
need for an adequate intake of dietary calcium during long-term exposure to fluoride.

A vitamin D supplement of 50,000 units twice weekly has been recommended during
high-fluoride treatment of osteoporosis (Jowsey et ae. 1972b; Riggs and Jowsey
1972; Kyle et al. 1975). However, Takizawa et al. (1975) did not obtain improvement
in geriatric patients with this high vitamin D dosage. Riggs et al. (1976) recently
compared two dosages of vitamin D (50,000 units, twice weekly, vs 400 units daily)
given in conjunction with the calcium and fluoride. They concluded that “we
do not recommend the large doses of vitamin D”. (The vitamin D topic is also
discussed in connection with fluoridated hemodialysis; see Section 5.8).

In relating the significance of the various nutrient-versus fluoride interrelations
discussed above to low-dose long-term daily exposure of humans to fluoride, it
is pertinent to note that ingestion of fluoride has increased over the past few
decades and is probably still escalating (Marier 1977; also Section 5.1). When
one considers that nutritional surveys have shown that sizeable proportions of
the North American population have an inadequate dietary intake of calcium and
of vitamin C (see “U.S. 1969”; “Canada 1973”), the need for
vigilance is apparent.

Table 24. Symptoms common to both fluoride intoxication and magnesium deficiency (Marier 1968).

Symptom Fluoride intoxication Magnesium deficiency
Leg cramps, or “pins and needles” Sauerbrunn et al. 1965(human) Fourman and Morgan 1962(human)
Muscular twitching Kretchner et al. 1963;Taves et al. 1965(human) Hanna et al. 1960;Suter and Klingmann 1955(human)
Tetaniform convulsions(with normal serum Ca) Ibid. Fourman and Morgan 1962(human)
2 to 3-fold increase in serum P(at time of convulsion) Ibid. Martindale and Heaton 1964(rats)
Cataracts (optical neuritis) Geall and Beilin 1964(human) Fourman and Morgan 1962(human)
Bone exostoses and/orsoft tissue calcification Weatherall and Weidmann 1959(rabbits, cats, and rats) O-Dell et al. 1960(rats)



Day and Powell-Jackson (1972)
reported that water-borne fluoride appeared to increase the prevalence of goitre
in an area where goitre was already endemic. Teotia and Teotia (1975) observed
a high incidence of goitre (up to 18% in the total population) in areas of endemic
hydrofluorosis. Crawford (1972) has commented on such interactions as follows:

“A Medical Research Council (U.K.) memorandum reported that, in some areas, even moderate concentrations of fluoride in drinking-waters
could block iodine absorption. It is known that the iodine concentrations are lower in soft than in hard waters …. If fluoride is added to soft waters …. a proportion of the population may come to have suboptimal iodine intake. The effects might be subtle and slow to develop, and would certainly not be picked up by the crude screening used at present”.

In studies with rats, Back (1970) found that the thyroid
preferentially retained increased amounts of fluoride for two weeks following
fluoride administration. Zucas and Lajolo (1975) reported that removal of the
pituitary gland caused increased deposition of skeletal fluoride, an effect
that seemed to be “related to thyroid hypofunction”. Bobek et al.
(1976) observed that fluoride supplementation for a two-month period caused
a slight decrease in blood thyroxine, a phenomenon thought to be caused by a
fluoride-mediated alteration in the functioning of thyroxine-binding proteins.

Day and Powell-Jackson (1972) recommended further research on the amino-acid
precursors of thyroxine, particularly tyrosine and its metabolites, because
increased urinary loss of tyrosine has been reported in fluorosis; also, because
tyrosine deficiency is a known cause of thyroid hypofunction.



In the human body, the kidneys are probably the most crucial organ during the
course of low-dose lonq-term exposure to fluoride. Healthy kidneys excrete 50
to 60% of the ingested dose (Marier and Rose 1971). Kidney malfunction can impede
this excretion, thereby causing an increased deposition of fluoride into bone.
Marier (1977) has reviewed data showing that, in persons with advanced bilateral
pyelonephritis, the skeletal fluoride content can be 4-fold that of similarly-exposed
persons with normal kidneys. Similarly, Mernagh et al. (1977) have reported
a 4-fold higher skeletal fluoride content in persons with the renal failure
of osteodystrophy. It has also been shown (Seidenberg et al. 1976; Hanhijarvi
1975) that plasma F- levels can be 3 1/2 to 5 times higher than normal in persons
with renal insufficiency. It is thus apparent that persons afflicted with some
types of kidney malfunction constitute another group that is more “at risk”
than is the general population. (Note: Some kidney-related problems have already
been discussed in Sections 4.1 and 4.2).

Understandably, people who have little, or no, kidney function constitute a
particular “at risk” group. This includes persons exposed to long-term
hemodialysis with fluoridated
(1 ppm) water, which aggravates the bone lesions of uremic renal osteodystrophy,
by increasing the severity of bone osteomalacia and the incidence of spontaneous
bone fractures (Posen et al. 1971; Johnson and Taves 1974). These effects parallel
those observed in high-dose fluoride therapy of osteoporotic patients, i.e.
osteomalacia (Merz et al. 1970; Jowsey et al. 1972b; Kyle et al. 1975) and spontaneous
fractures (Inkovaara et al. 1975; Zanzi et al. 1975).

Inkovaara et al. (1975) recommended that the plasma inorganic fluoride ion (plasma
F-) concentration should not exceed 3 umol/l if spontaneous fractures are to
be avoided. As a basis for comparison, Nielsen et al. (1973) report predialysis
plasma F- levels of 9 umol/l, and Posen et al. (1971) extrapolate their observations
to an initial (i.e. before the first dialysis treatment) level of 7 umol/l.
Still higher plasma F- levels have been observed during the course of fluoridated
dialysis (Posen et al. 1971; Jowsey et al. 1972a; Nielsen et al. 1973; Cordy
et al. 1974). The plasma F- can attain a concentration of 36 pmol/l during long-term
fluoridated hemodialysis treatment (Fournier et a. 1971), and this level is
about 50 times higher than normally found in residents of unfluoridated communities
(Taves 1968; Hanhijarvi 1975).

During maintenance of patients on fluoridated hemodialysis, the increased body-burden
of fluoride is reflected by high levels of fluoride in bone (Posen et al. 1971),
and a high molar F/Ca ratio in bone (Jowsey et al. 1972a; Cordy et al. 1974).
Posen et al. (1971) administered high doses (as high as 200,000 units per day)
of vitamin D concurrently, and it was felt that this contributed to the severity
of the bone changes. The patients studied by Cordy et al. (1974) had a daily
vitamin D intake of only 460 units (Note: Riggs et al. (1976) now recommend
400 units daily during high fluoride therapy). Cordy et al. (1974) observed
lower plasma F- levels and less severe bone disease than previously reportedly
Posen et al. (1971).

In a study of fluoridated-hemodialysis patients, Nielsen et al. (1973) observed
that 86% of their patients showed evidence of secondary hyperparathyroidism,
along with a significant increase in serum alkaline phosphatase. These manifestations
have also been seen in endemic fluorosis patients (Krishnamachari and Krishnaswami
1973; Teotia et al. 1974; Faccini and Teotia 1974; Sivakumar and Krishnamachari

As discussed by Rao and Friedman (1973), some dialysis clinics have not encountered
problems with fluoridated hemodialysis. Nevertheless, several researchers consider
it prudent to use non-fluoridated water, so as to reduce the risk of osteomalacia
(Stewart 1969; Posen et al. 1971; Jowsey et al. 1972a; Cordy et al. 1974; Lough
et al. 1975; Rao and Friedman 1975).

Persons suffering from nephropathic Diabetes Insinidus make up another subgroup
that is more “at risk” than the general Population. Table 25 summarizes
the observations on 10 such cases about whom we have found reports. A striking
feature of this tabulation is the young age at which skeletal fluorosis has
become evident in some of these patients. Thus, Juncos and Donadio (1972) diagnosed
skeletal fluorosis in an 18-year-old boy and a 17-year-old girl. The only other
report of skeletal fluorosis in children appears to be that of Teotia et al.
(1971), which dealt with endemic hydrofluorosis in India.


Table 25. Fluorosis in persons who have the Diabetes Insipidus syndrome.

F intake from Drinking-watermg/day
Authors’ Comments
F, mg/l
Intake, l/day
64 M 2.56 4 to 10 10.24 to 25.6 Skeletal fluorosispolydipsiapolyuriapyelonephritisDiabetes Insipidus “Drinking-water seems to have been his
only source of fluoride intake…Prolonged polydipsia may be hazardous to
persons who live in areas where the levels of fluoride in drinking-water
are not those usually associated with significant fluorosis.”
Sauerbrunn et al. 1965
1817 MF 2.61.7 “about 2 gal.””large amounts” approx. 20? Skeletal fluorosispolydipsiapolyuriarenal insufficiency “It is postulated that the renal insufficiency,
which resulted in the large intake of fluoride-containing water and reduced
excretion of fluoride, combinedto produce systemic fluorosis.”
Juncos and Donadio 1972
1011 MM 1.01.0 1.25 to 3.01.25 to 3.0 1.25 to 3.01.25 to 3.0 Dental fluorosispolydipsiapolyurianephropathyDiabetes Insipidus “We are reporting two children with nephrogenic
diabetes insipidus and fluorosis, and suggest looking for evidence of fluoride
toxicity in individuals with polydipsia. Substituting non-fluoridated water
as part of the fluid intake is recommended.”
Greenberg et al. 1974
351413108 FFFFM 10 to 1510 to 1510 to 154 to 54 to 5 5 to 7.55 to 7.55 to 7.52 to 2.52 to 2.5 Dental fluorosispolydipsiapolyuriaDiabetes Insipidus “Drinking of large amounts of water,
even with lower-than accepted fluoride content, can produce fluorosis of
the teeth.”
Klein 1975

Comparison of the “Diagnosis”, “F intake”,
and “Age” columns in Table 25 suggests that there is a progression from
nephropathy, to renal insufficiency, to pyelonephritis,with increasing age and/or
sustained fluoride intake. This is corroborated in the Algerian studies of various
stages of human hydrofluorosis, as conducted by Poey et al. (1976). Gradual impairment
of urinary creatinine clearance was indicative of progressive inhibition of kidney
glomerular filtration, affecting tubular reabsorption of water. In the final stage,
urinary excretion of fluoride accounted for only 10 to 20% of the intake (i.e.
about 1/6 to 1/3 of normal), and was thus indicative of high fluoride retention.
Based on histologically-detectable glomerular degeneration, Poey et al. concluded
that fluoride can complicate, or can actually induce, nephropathy. The WHO (1970)
report had recognized that “the remote possibility that fluoride may aggravate
renal disease has not been conclusively ruled out”.

In all the cases tabulated in Table 25, the patients had the excessive thirst
of polydipsia. As noted by Klein (1975), even if the beverages consumed to satisfy
this thirst have a “lower-than-accepted” fluoride content, this can
result in an excessive intake of fluoride. Experinmental confirmation of fluoride-induced
polydipsia is found in the 18-month study by Manocha et al. (1975), who observed-that
fluoride caused a considerable increase in the water intake of monkeys.

Another aspect of the Diabetes Insipidus syndrome is the abnormally-high output
of urine (polyuria). In a report of a study with rats, Hamuro (1972b) has stated:

“The Polyuria induced by fluoride was accompanied by an enhanced sodium excretion and a decrease in osmolality. These results were consistent with previous findings that the administration of fluoride caused polyuria in laboratory animals”.

In a study with humans, Taves et al. (1972) remarked that

“The data …. is consistent with the hypothesis that F (i.e. fluoride) is the cause of the polyuria”

Singer and Forrest (1976), writing about drug-induced states of nephrogenic
Diabetes Insipidus in humans, state that

“these studies (with sodium fluoride) have been interpreted to suggest that sodium fluoride does not grossly impair (kidney) ascending-limb sodium chloride transport, but may cause nephrogenic Diabetes Insipidus, either by washing-out the medullary solute (since fluoride is a vasodilator in many vascular beds), or by reducing the collecting-duct permeability”.

Manocha et ai. (1975) reported “significant cytochemical
changes” in the kidneys of monkeys which had consumed water containing
1 or 5 ppm fluoride for an 18-month period. Thus, there is suggestive evidence
that fluoride may cause nephrogenic Diabetes Insipidus.

Corroborative evidence for the above statementsis found in the results of studies
on the effect of the metabolically-unstable anesthetic, methoxyflurane, which
releases inorganic fluoride to body fluids. Thus, Gottlieb and Trey (1974) have

“This (methoxyflurane) syndrome is …. similar to that of nephrogenic Diabetes Insipidus …. these patients were unable to concentrate urine, despite fluid deprivation or administration of vasopressin …. the (kidney) changes indicated that the lesion was of the distal nephron …. None of the control patients developed nephrogenic Diabetes Insipidus, while methoxyflurane patients developed this syndrome. (There was) a relationship among the dose of methoxyflurane …. serum peak inorganic fluoride levels, and renal effects”.

Cousins and Mazze (1973) also commented on the methoxyflurane
syndrome as follows:

“Serum hyperosmolality occurring simultaneously with polyuria, and decreased urine osmolality is evidence of water-losing nephropathy, although primary antidiuretic hormone deficiency (Diabetes Insipidus) could produce similar abnormalities. However, unlike Diabetes Insipidus, polyuria following methoxyflurane administration was vasopressin-resistant, indicating a renal lesion …. (Also, excessive) thirst and polyuria added difficulty to post-operative management”.

With reference to the foregoing quotation, it is pertinent
to note (see Table 25) that the patient treated by Sauerbrunn et al. (1965),
and the two treated by Greenberg et at, (1974) were vasopressin-resistant; this
suggests that they were suffering from fluoride-mediated renal lesions. Conversely,
the two youngest patients in Klein’s (1975) study responded to vasopressin treatment,
i.e. indicative of hereditary Diabetes Insipidus (Cousins and Mazze 1973; Klein

Thus, as noted by Marier (1977), there are identifiable individuals among the
general population who are more “at risk” relative to fluoride intoxication,
because they are afflicted with the polyuria-polydipsia syndrome of Diabetes
Insipidus. Such people ingest abnormal amounts of fluoride in beverages, and
retain an abnormally high proportion of the total ingested fluoride, as reflected
in Hanhijarvi’s (1975) observation that they have a low urinary fluoride clearance
and a high plasma F- level. It is therefore appropriate to note that, recently
(see JAMA 1976), diabetes (including cases with nephropathy) has been ranked
third as a cause-of-death factor, accounting for an estimated 300,000 deaths
per year; also, the incidence of diabetes increased by 6% per year during the
period 1965-1975. If an escalation in the incidence of nephropathic diabetes
has occurred, it should be carefully considered in relation to the evidence
discussed in the present report.



Precise data on the total daily intake of fluoride by humans are scarce. Most
studies of fluoride as a possible hazard to human health have reported only
the fluoride exposure from a single source (e.g. water, or polluted air in the
workplace). This preoccupation with fluoride from one source, coupled with the
difficulty of quantifying or even identifying, with certainty, the early stages
of a fluoride-induced injury, has inhibited attempts to establish criteria that
would define an acceptable level of intake. However, the need for such criteria
is apparent, and two recent publications have reported attempts to set a value
for an acceptable daily intake.

Farkas (1975) attempted to estimate a “safe” level of fluoride intake
by means of a questionnaire directed to “authorities” in the fields
of dentistry, medicine, nutrition, and biological research. The questionnaire
requested a definite estimate in absolute terms (mg/kg body weight per day),
but no consensus developed. Many respondents merely indicated that various levels
of ingested fluoride, expressed in parts-per-million (ppm), were acceptable
or recommended. However, five of the respondents agreed that 0.05 to 0.07 mg/kg
body weight per day was a reasonable estimate of the acceptable daily intake
of fluoride.

Toth (1975) contended that the amount of fluoride “which is ingested with
drinking-water” should be considered optimal. Toth estimated that this
amount is 0.045 mg/kg body weight per day for infants, and declines to 0.023
mg/kg for adults. By considering various factors, Toth also estimated a “tolerable
dose” (which presumably approximates an acceptable daily intake) of 0.073
mg/kg body weight per day for infants, and 0.033 mg/kg for adults. Reference
Man has a body mass of 70 kg (ICRP 1975). Man’s total water intake has been
estimated to be 2400 ml/day (Spencer et al. 1970). Therefore, ingestion of 2400
ml of water with a fluoride concentration of 1 mg/l would provide a daily fluoride
intake of 0.034 mq/kg from water.

A more direct, criteria-based, approach to the estimation of an acceptable daily
intake is urgently required, and the present authors therefore reluctantly present
the following calculations, in spite of the obvious limits to their accuracy.


5.9.1 Criteria Based on Bone Fluoride and Plasma F- 

Kramer et al. (1974) presented data on the fluoride content of drinking water,
in relation to the total fluoride intake from three daily meals by adults residing
in 16 locations in the U.S. The regression equation calculated (by us) from
these data is:

Daily fluoride intake, mg = (0.98 ± 0.41) + (ppm waterborne F x 1.63 ± 0.50)

This equation (hereafter referred to as “Equation A”) has an r2 value
of 0.44. Although the error factors are large, this equation enables us to convert
waterborne fluoride levels to daily dietary intakes. It must be emphasized that
the Kramer et al. (1974) data do not include between-meal ingestion of drinking-water
or other fluoridated beverages, and therefore underestimate the total daily
fluoride intake.

Hodge (1952) and Jackson and Weidmann (1958) reported the fluoride content of
drinking-water as it relates to the fluoride content of dry fat-free rib bone
of lifetime residents in four communities whose drinking-waters differ in fluoride
content (i.e. 0.06, 0.5, 0.8, and 1.9 ppm). In Fig.
, we have plotted dietary fluoride intake (calculated from waterborne fluoride
by Equation A, above) in relation to the fluoride content of rib bone from 55-year-old
individuals, as reported by Hodge (1952) and by Jackson and Weidmann (1958).

If a fluoride content of 2500 to 4000 ppm in dry fat-free rib (a cancellous-type
bone) can be taken as an upper range of tolerable limits (cf Section 5.3; also
Jackson and Weidmann 1958), this would reflect an intake of 2.5 to 4.1 mg fluoride
from the three daily meals. For a 70 kg adult, this calculation leads to an
estimated acceptable intake, from the three daily meals, of 0.036 to 0.059 mg/kg
body weight.

Another approach to criteria is based on blood plasma ionic fluoride. Hanhijarvi
(1975) has determined the equations relating blood plasma F- levels to age of
humans residing in areas with non-fluoridated (0.2 ppm) or fluoridated (1.0
ppm) water. The conversion of waterborne fluoride content to dietary fluoride
intake was again calculated (by us) from Equation A, above. Blood plasma F-
levels for 55-year-old humans were calculated by using Hanhijarvi’s equations.
The results are plotted in Fig. 4, where they
are extrapolated linearly to allow an interpretive assessment.

Inkovaara et al. (1975) have provided the best available estimate of a tolerable
maximum level for plasma F-; during a 10-month study of geriatric patients,
they observed spontaneous bone fractures at blood plasma F- levels as low as
2.1 umol/l, and recommended that it should not rise in excess of 3.0 pmol/l.
In terms of Fig. 4, these levels reflect a range
of fluoride intakes of 3.7 to 5.3 mg fluoride from the three daily meals, i.e.
equivalent to a range of 0.053 to 0.076 mg/kg for a 70 kg human.

In summary (and including the estimates by Farkas and by Toth, cited in Section
5.9), the available estimates of an acceptable daily intake of fluoride for
humans are:

0.05 to 0.07 mg/kg (Farkas; questionnaire)

0.033 to 0.073 mg/kg (Toth; “tolerable” level)

0.036 to 0.059 mg/kg (our calculation based on rib bone)

0.053 to 0.076 mg/kg (our calculation based on plasma F-)

The agreement among these values is sufficiently close to
suggest that the various approaches are meaningful. Serious attempts to increase
the data-base for such calculations should be encouraged. However, it must not
be overlooked that the acceptable intake derived from such calculations applies
only to an “average” individual, and that some “safety factor”
must be applied to ensure protection of the less resistant individuals in the
general population.

In Sections 3.1.1 and 5.8, we discussed the fact that, in persons with renal
insufficiency, the bone and plasma F levels can be 4 times higher than in similarly-exposed
persons with normal kidneys. Until there is evidence to the contrary, we must
therefore conclude that some persons with renal insufficiency have only one-quarter
the fluoride tolerance reported in the above tabulation.

In Section 5.1.1, we stated that the current total fluoride intake from foods
and beverages, in areas with fluoridated (1 ppm) water, probably ranges from
3.5 to 5.5 mg/day, i.e. equivalent to 0.05 to 0.08 mg/kg/day for a 70 kg “average”
human. This range is almost within the ranges for long-term “acceptable
daily intake” tabulated above.


5.9.2 Assessment of Fluoride Intake From Air

The information discussed in Sections 5.9 and 5.9.1 allows an evaluation of
human intake of fluoride from air, especially in the workplace where there has
been disagreement about what constitutes an acceptable concentration of airborne
fluoride. Inhaled fluorides, whether in gaseous or in particulate form, are
almost completely absorbed into the bloodstream (WHO 1970; Hodge and Smith 1977).
Therefore, assuming an inhalation of 10 m3 of air during an 8-hour working shift
(Dinman et al. 1976a), the various airborne fluoride levels discussed for workplace
exposure can be converted into daily uptakes of fluoride by a 70 kg human.

Because workplace exposure is on a 5 day/week basis, a correction factor of
5/7 was used to express intake on an equalized “per day” basis. [In
summary, the airborne fluoride concentration is multiplied by 10 (M3), divided
by 70 (kq), then multiplied by 5/7; the overall conversion factor is thus 0.102].
This calculation leads to the following tabulation:


Airborne Fluoride mg/m3

Discussed by

Fluoride Uptake FromWorkplace Airmg/kg/day
Hodge and Smith (1977)
Vishnevski (1969)
Guminska and Sterkowicz
Vishnevski (1969)

These calculated intakes are from air only and must be considered
additional to those discussed in Section 5.9.1. it is apparent that the long-term
implications of the occupational exposures require careful study, taking into
consideration both the non-occupational exposures to fluoride and the period
(i.e. portion of lifetime) of additional exposure to workplace airborne fluorides.


1. Despite improvements in, and more extensive use of, emission control equipment,
large quantities of fluoride continue to be discharged into the atmosphere from
industrial sources. In 1972, at least 14,236 metric tons of fluoride (calculated
as fluorine) were discharged into the air in Canada, and some 150,000 metric
tons were discharged in all of North America.

2. Large quantities of fluoride are also discharged into streams, rivers, lakes
and oceans, as a component of industrial waste-waters. It appears probable that
the amounts thus discharged are several-fold larger than the amounts discharged
into the atmosphere. Many systems utilized for airborne emission-control contribute
extensively to the amount of fluoride discharged in wastewaters.

3. Much of the fluoride discharged into the atmosphere arises from “point
sources” such as smelters. Dispersion of the pollutant in the surrounding
area is not uniform; therefore, the siting of monitoring devices, and the selection
of sites for sampling of vegetation, must consider:

a) Stratification of the pollutant, with the higher levels
of atmospheric fluoride found at the greater heights. This has significance
to ecological damage to vegetation on exposed hilltops or mountainsides, even
at considerable distances from the emitting source;

b) The “shielding” effects of vegetation and other obstacles, which
results in lower fluoride exposure (and uptake by) vegetation growing on the
downwind side of such obstacles.

4. The ecological impact of airborne fluoride emissions
is known to be serious with regard to coniferous forests and to epiphytes and
bryophytes. However, lichens and mosses are less susceptible to fluoride injury
than coniferous trees. Relatively little is known about the effects of fluoride
on aquatic life, even though large amounts of fluoride are known to be released
into some waterways. More attention should be paid to fluoride effects on pollinating
insects and on plankton.

5. Airborne fluoride has had a serious impact on agricultural
and silvicultural species. With airborne gaseous fluoride there is no evidence
for a no-effect threshold level below which no reduction in crop yield occurs,
especially over the long term. For exposure of Canadian forest species, the
average (30-day) airborne gaseous fluoride concentration should not exceed 0.2
ug/M3. There is an urgent need for long-term Cause/Effect studies of species
known to be sensitive to fluoride injury.

6. There have been episodes where the impact
of fluoride pollution on livestock and on wild ungulates has been severe. To
date, regulations limiting the fluoride content of fodders have provided neither
adequate protection against economic loss to the farmer nor adequate control
of airborne fluoride. For young growing swine, an 18-week exposure to dietary
fluoride (whether in forages, feeds, or mineral supplements) can be expected
to decrease daily weight-gain by about 4% for each 100 ppm increment of dietary
fluoride. The need for further data on which Cause/Effect equations can be based
is apparent. Loss of weight-gain may be a suitable measure of sub-clinical (pre-skeletal)

7. There is clear evidence that wildlife species are
more vulnerable to fluoride toxicosis than are livestock species. The impact
seems to be most severe on predator species, because they must capture their
prey and because they are more susceptible to the bioaccumulation of fluoride
through their food chain. Cause/Effect studies of these species should include
consideration of the multiple stresses imposed by the ecosystem (e.g. malnutrition).

8. Researchers in various regions of the world have reported that human hydrofluorosis
is less severe when the waterborne fluoride is ingested from hard waters, than
from soft waters. There is evidence that chronic intake of fluoride increases
the long-term metabolic requirement for both calcium and magnesium. Other studies
have indicated that fluoride may increase the metabolic requirement for vitamin
C and manganese. The Cause/Effect aspects of these dietary/nutritional factors
require urgent attention, with regard to chronic intake of fluoride. There is
no doubt that inadequate nutrition increases the severity of fluoride toxicosis.

9. Fluoride has displayed mutagenic activity in studies of vegetation, insects,
and mammalian oocytes. There is a high correlation between carcinogenicity and
mutagenicity of pollutants, and fluoride has been one of the major pollutants
in several situations where a high incidence of respiratory
has been observed. For these reasons, the relation between airborne
fluoride and incidence of lung cancer needs to be investigated.

10. Long-term ingestion, with accumulation of fluoride in animals and man, induces
metabolic and biochemical changes, the significance of which has not yet been
fully assessed. It cannot be assumed that such changes are of no significance
to human health. There is evidence that neurological complaints are related
to the early histological changes that precede overt skeletal fluorosis. There
is also evidence that the early bone changes can reflect an entire gamut of
abnormalities, depending on factors such as nutritional
and metabolic status. Further studies on the early subtle changes of fluoride
toxicosis in humans, in terms of both diagnostic aids and Cause/Effect interrelations,
should have a high priority.

11. Fluoride is a persistent bioaccumulator, and is entering into human food-and-beverage
chains in increasing amounts. Careful consideration of all available data indicates
that the amount of fluoride ingested daily in foods and beverages by adult humans
living in fluoridated communities currently ranges between 3.5 and 5.5 mg.
For a 70 kg human adult, this range is close to the 0.03 to 0.07 mg/kg/day estimated
for “an acceptable daily intake”. In addition to the food-chain, dentifrices
and pharmaceuticals can contribute siqnificantly to the fluoride intake of some

12. Inhalation of airborne fluoride may contribute several milligrams to the
total daily intake of industrial workers, and
may be significant for persons residing near sources of fluoride emissions.
However, the effect of airborne fluoride on human respiratory tissue is not
necessarily related to total bodyburden, but may relate to the direct impact
of fluoride on respiratory tissues. The contribution of cigarette-smoking to
fluoride intake also requires study.

13. In the assessment of the impact of fluoride on animals and man, more attention
should be focussed on the concentration of inorganic fluoride in blood plasma.
Available evidence indicates that accurate assessment of the plasma F concentration
can provide valuable information about the body-burden during chronic fluoride

14. In addition to industrial workers, there are several sub-groups of the population
who may be more affected by environmental fluoride than the population at large.
These are persons who:


a) Have a sub-optimal nutritional status, especially
with regard to calcium, magnesium, vitamin C, manganese, or a low dietary Ca/P
ratio (Note: This also applies to animals);

b) Live in the proximity of fluoride-emitting industries;

c) Live in regions where goitre is endemic, because
there is suggestive evidence that fluoride may increase the incidence of goitre
in such regions;

d) Have kidney impairments, particularly those with bilateral pyelonephritis
or nephropathic Diabetes Insipidus;

e) Have the excessive-thirst polydipsia associated with diabetes, because they
consume large quantities of fluids.

These may be called “critical groups” (ICRP 1977)
either because they accumulate more fluoride or suffer toxic effects more readily.

15. Standards limiting emissions or environmental concentrations of fluoride
should be based on criteria which include those derived from studies of these
“critical groups”.

16. In addition to the research recommendations we have
made, we would like to acknowledge those presented in a recent U.S. National
Academy of Sciences report (see Fleischer et al. 1974):


a) Additional detailed studies are needed of the health
of human and animal populations exposed to high concentrations of airborne

b) The gross effects of fluoride on plants and animals have been studied,
but much needs to be done on the basic biochemical lesions induced by fluoride,
and on dietary factors affecting fluoride uptake by man;

c) The very large emission of fluorocarbons (freons), and their rapidly increasing
use, require study of their distribution, rate of degradation, and possible
effects on plants, animals and humans;

d) Waste waters of high fluoride content have been released from phosphate
processing and from the aluminum industry, with detrimental effects to such
marine organisms as oysters and crabs. Possible chronic effects from exposure
of such organisms to lower levels of fluoride need study;

e) In view of the high fluoride content reported to exist in some fish-protein
concentrates used as food supplements, the possible impact of this added source
of fluoride in the diet should be further investigated;

f) Methods of sampling and separating gaseous and particulate forms of airborne
fluoride need study and standardization;

g) Further work is needed on the relation of the uptake of fluorine by plants
to its concentration in the air;

h) Study of the form of fluorine in plants is highly desirable, especially
the nature of fluorine bonding in plant tissue and its solubility in aqueous

i) More data are needed on the relation of the fluoride content of groundwaters
to the mineralogical and chemical composition of the source rocks.

These NAS recommendations are fully compatible with the
information that we have presented in this report or in our previous review
(Marier and Rose 1971).


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National Research Council of Canada – 1977