Chemosphere 50 (2003) 251–264
www.elsevier.com/locate/chemosphere
Fluoride toxicity to aquatic organisms: a review
*
Julio A. Camargo
Departamento Interuniversitario de Ecologıa, Edificio de Ciencias, Universidad de Alcal
a, Alcal
a de Henares, Madrid E-28871, Spain
Received 8 March 2002; received in revised form 22 July 2002; accepted 23 August 2002
Abstract
Published data on the toxicity of fluoride (F ) to algae, aquatic plants, invertebrates and fishes are reviewed. Aquatic
organisms living in soft waters may be more adversely affected by fluoride pollution than those living in hard or
seawaters because the bioavailability of fluoride ions is reduced with increasing water hardness. Fluoride can either
inhibit or enhance the population growth of algae, depending upon fluoride concentration, exposure time and algal
species. Aquatic plants seem to be effective in removing fluoride from contaminated water under laboratory and field
conditions. In aquatic animals, fluoride tends to be accumulated in the exoskeleton of invertebrates and in the bone
tissue of fishes. The toxic action of fluoride resides in the fact that fluoride ions act as enzymatic poisons, inhibiting
enzyme activity and, ultimately, interrupting metabolic processes such as glycolysis and synthesis of proteins. Fluoride
toxicity to aquatic invertebrates and fishes increases with increasing fluoride concentration, exposure time and water
temperature, and decreases with increasing intraspecific body size and water content of calcium and chloride. Freshwater invertebrates and fishes, especially net-spinning caddisfly larvae and upstream-migrating adult salmons, appear to
be more sensitive to fluoride toxicity than estuarine and marine animals. Because, in soft waters with low ionic content,
a fluoride concentration as low as 0.5 mg F /l can adversely affect invertebrates and fishes, safe levels below this fluoride
concentration are recommended in order to protect freshwater animals from fluoride pollution.
Ó 2002 Elsevier Science Ltd. All rights reserved.
Keywords: Fluoride toxicity; Aquatic organisms; Review
Contents
1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . .
2. Toxicity to algae and aquatic plants . . . . . . . . . . .
3. Bioaccumulation in aquatic invertebrates and fishes
4. Toxicity to aquatic invertebrates . . . . . . . . . . . . . .
5. Toxicity to fishes . . . . . . . . . . . . . . . . . . . . . . . . .
6. Ecological risk assessment . . . . . . . . . . . . . . . . . .
Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . .
References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
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*
Fax: +34-91-8854-929.
E-mail address: [email protected] (J.A. Camargo).
0045-6535/03/$ - see front matter Ó 2002 Elsevier Science Ltd. All rights reserved.
PII: S 0 0 4 5 - 6 5 3 5 ( 0 2 ) 0 0 4 9 8 - 8
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255
257
259
261
261
261
252
J.A. Camargo / Chemosphere 50 (2003) 251–264
1. Introduction
Of all chemical elements in the Periodic Table, fluorine is the most electronegative and the most reactive
(Greenwood and Earnshaw, 1984; Gillespie et al., 1989).
Because of its great reactivity, fluorine cannot be found
in nature in its elemental state. It exists either as inorganic fluorides (including the free anion F ) or as
organic fluoride compounds (e.g., freons), always exhibiting an oxidation number of )1. In the global environment, inorganic fluorides are much more abundant
than organic fluoride compounds (Greenwood and
Earnshaw, 1984; Gillespie et al., 1989).
The first major natural source of inorganic fluorides
is the weathering of fluoride minerals (CEPA, 1994).
Most important inorganic fluoride minerals in the
earthÕs crust are fluorapatite (Ca5 (PO4 )3 F), fluorite
(CaF2 ) and cryolite (Na3 ALF6 ). Volcanoes are the second major natural source through the release of gases
with hydrogen fluoride (HF) into the atmosphere
(CEPA, 1994). Global release estimates suggest annual
emissions of inorganic fluorides between 60 and 6000
kilotonnes (Symonds et al., 1988; CEPA, 1994). Marine
aerosols are the third major natural source (CEPA,
1994). It has been estimated that marine aerosols can
globally contribute with about 20 kilotonnes of inorganic fluorides each year (Symonds et al., 1988; CEPA,
1994).
In water, inorganic fluorides usually remain in solution (as fluoride ions) under conditions of relatively low
pH and hardness and in the presence of ion-exchange
materials such as bentonite clays and humic acids (Pickering et al., 1988; CEPA, 1994). Inorganic fluorides in
solution may however be removed from the aquatic
phase by the precipitation of calcium carbonate, calcium
phosphate, calcium fluoride and, even, magnesium fluoride (Stumm and Morgan, 1996). Aquatic organisms
living in soft waters may thus be more affected by fluoride pollution than those living in hard or seawaters
because the bioavailability of fluoride ions (and, consequently, their toxic action) is reduced with increasing
water hardness.
Levels of fluoride in surface waters vary according
to location and proximity to emission sources. Fluoride
concentrations in unpolluted fresh waters generally
range from 0.01 to 0.3 mg F /l whilst in unpolluted
seawaters generally range from 1.2 to 1.5 mg F /l
(Dobbs, 1974; Martin and Salvadori, 1983; Nriagu, 1986;
Fuge and Andrews, 1988; Camargo et al., 1992a; CEPA,
1994; Camargo, 1996a; Datta et al., 2000). Nevertheless,
natural elevated levels of inorganic fluoride may be
found in regions where there is geothermal or volcanic
activity. Field surveys (Neuhold and Sigler, 1960) have
shown that hot springs and geysers of Yellowstone
National Park contain from 25 to 50 mg F /l, while
Madison and Firehole rivers, also in Yellowstone
National Park, can contain from 1 to 14 mg F /l. The
natural thermal waters of New Zealand contain from
1 to 12 mg F /l (Mahon, 1964). Water samples from
Walker and Pyramid lakes, in Nevada, contain up to 13
mg F /l (Sigler and Neuhold, 1972). Fluoride concentrations may also be elevated in ground waters from
regions where there is no geothermal or volcanic activity
(Rajagopal and Tabin, 1991; CEPA, 1994; Sarma and
Rao, 1997).
Human activities can unfortunately lead to increased
local levels of fluoride in surface waters. Skjelk
avle
(1994) found that some Norwegian streams close to aluminium smelters might contain more than 10 times the
fluoride natural background level of about 0.05 mg F /l.
Similar results have been reported by Warrington (1992)
for an aluminium smelter on the Kitimat River (BC,
Canada) and by Roy et al. (2000) for an aluminium
smelter on the Saguenay River (Que., Canada). Other
important human activities causing significant increases
(more than 100 times the natural background level) in
the fluoride concentration of surface waters are phosphate fertiliser plants (Somashekar and Ramaswamy,
1983; Van Craenenbroeck and Marivoet, 1987), plants
producing fluoride chemicals such as hydrogen fluoride,
calcium fluoride, sodium fluoride and sulphur hexafluoride (Zingde and Mandalia, 1988; Karunagaran and
Subramanian, 1992; CEPA, 1994), plants manufacturing brick, ceramics and glass (CEPA, 1994; Camargo,
1996a), and the use of fluoride containing pesticides
(Fuge and Andrews, 1988; CEPA, 1994). Discharges of
fluoridated municipal waters also cause significant increases (about five times the natural background level)
in the fluoride concentration of recipient rivers (Sparks
et al., 1983; Camargo et al., 1992a). In addition, though
with a lower risk for aquatic life, aluminium smelters
and other industries located along the marine coast can
increase the fluoride concentration of seawaters (Wright
and Davison, 1975; Oliveira et al., 1978; Pankhurst et al.,
1980; Ares et al., 1983; CEPA, 1994).
In spite of the fact that fluoride must be considered
as a serious pollutant since its concentration in many
aquatic ecosystems is significantly increasing as a consequence of manÕs activities, relatively little is known
about fluoride toxicity to aquatic life. In this study,
published data on the toxicity of fluoride (F ) to algae,
aquatic plants, invertebrates and fishes are reviewed.
2. Toxicity to algae and aquatic plants
Fluoride can either inhibit or enhance the population
growth of algae, depending upon fluoride concentration,
exposure time and algal species (Table 1). Some algae
are able to tolerate inorganic fluoride levels as high as
200 mg F /l. The toxic action of fluoride on algal growth
may reside in the fact that fluoride ions can affect nu-
J.A. Camargo / Chemosphere 50 (2003) 251–264
253
Table 1
Effects of fluoride on population growth of freshwater and marine algae
Species
Aquatic medium
Fluoride level
(mg F /l)
Effect
Reference
Chlorella pyrenoidosa
Selenastrum capricornutum
Ankistrodesmus braunii
Cyclotella meneghiniana
Oscillatoria limnetica
Scenedesmus quadricauda
Stephanodiscus minutus
Synechococcus leopoliensis
Nitzschia palea
Chlorella vulgaris
Amphora coffeaeformis
Agmenellum quadruplicatum
Amphidinium carteri
Freshwater
Freshwater
Freshwater
Freshwater
Freshwater
Freshwater
Freshwater
Freshwater
Freshwater
Freshwater
Brackish
Seawater
Seawater
Dunaliella tertiolecta
Seawater
Nannochloris oculata
Pavlova lutheri
Seawater
Seawater
Prasinocladus marimus
Chroomonas salina
Rhodomonas lens
Seawater
Seawater
Seawater
190
123
6 50
6 50
6 50
6 50
6 50
25–50
10–100
266–380
70
6 100
100
150–200
6 100
50–200
6 100
100
150–200
6 100
6 100
25–100
Smith and Woodson, 1965
LeBlanc, 1984
Hekman et al., 1984
Hekman et al., 1984
Hekman et al., 1984
Hekman et al., 1984
Hekman et al., 1984
Hekman et al., 1984
Joy and Balakrishnan, 1990
Rai et al., 1998
Joy and Balakrishnan, 1990
Oliveira et al., 1978
Oliveira et al., 1978
Antia and Klut, 1981
Oliveira et al., 1978
Antia and Klut, 1981
Oliveira et al., 1978
Oliveira et al., 1978
Antia and Klut, 1981
Oliveira et al., 1978
Oliveira et al., 1978
Oliveira et al., 1978
Bellerochea polymorpha
Chaetoceros gracilis
Seawater
Seawater
6 100
6 100
50–200
Nitzschia angularis var.
affinis
Thalassiosira weissflogii
Seawater
100
58–82% growth inhibition (72 h)
Growth inhibition (96 h EC50 )
Unaffected (175 h)
Unaffected (175 h)
Unaffected (175 h)
Unaffected (175 h)
Unaffected (175 h)
Growth inhibition (175 h)
Growth enhancement (96 h)
Growth inhibition (15 d EC50 )
Growth enhancement (96 h)
Unaffected (23 d)
27% growth inhibition (23 d)
25–90% growth inhibition (36 d)
Unaffected (18 d)
Unaffected (16 d)
Unaffected (18 d)
30% growth inhibition (25 d)
50–65% growth inhibition (33 d)
Unaffected (18 d)
Unaffected (18 d)
20–30% growth enhancement
(23 d)
Unaffected (18 d)
Unaffected (18 d)
20–70% growth enhancement
(34 d)
25% growth inhibition (23 d)
Oliveira et al., 1978
Seawater
Skeletonema costatum
Seawater
6 100
50–200
82
Unaffected (18 d)
Unaffected (33 d)
Growth inhibition (96 h EC50 )
Oliveira et al., 1978
Antia and Klut, 1981
LeBlanc, 1984
Oliveira et al., 1978
Oliveira et al., 1978
Antia and Klut, 1981
Exposure times are presented in parenthesis.
cleotide and nucleic acid metabolism governing processes of algal cell division (Antia and Klut, 1981). On
the contrary, the growth stimulation in certain algal
species could be due to a fluorine requirement for optimal growth (Oliveira et al., 1978). Aquatic plants seem
to be effective in removing fluoride from contaminated
water under laboratory and field conditions. The fluoride content in aquatic macrophytes increases with increasing fluoride concentration and exposure time
(Hocking et al., 1980; Shirke and Chandra, 1991; Sinha
et al., 2000).
McNulty and Lords (1960), studying the effect of
sodium fluoride on the metabolism of the green alga
Chlorella pyrenoidosa in laboratory experiments, observed that low concentrations of NaF significantly
increased oxygen consumption and total phosphorylated
nucleotides in the respiration of C. pyrenoidosa. The
observed metabolic stimulation was related to the concentration of undissociated hydrogen fluoride in the
suspending media. In contrast, Holzer (1954) reported
that the rate of respiration in this algal species was reduced by 50% with 570 mg F /l, and Kandler (1955)
stated that 1900 mg F /l might inhibit respiration in C.
pyrenoidosa by 25%. Smith and Woodson (1965), following laboratory experiments with this freshwater algal
species, found that the growth rate of C. pyrenoidosa
was inhibited 58–82% at a fluoride concentration of 190
mg F /l after 72 h of exposure.
Oliveira et al. (1978) exposed, in seawater of 26‰
salinity, 12 species of marine phytoplankters (Agmenellum quadruplicatum, Amphidinium carteri, Dunaliella
tertiolecta, Nannochloris oculata, Prasinocladus marimus,
Pavlova lutheri, Chroomonas salina, Rhodomonas lens,
Bellerochea polymorpha, Chaetoceros gracilis, Nitzschia
angularis var. affinis and Thalassiosira weissflogii) to
fluoride concentrations ranging from 0 to 100 mg F /l
for exposure times of up to 25 days. They found
that, while the population growth of the cryptomonad
R. lens was stimulated by 20–30% at fluoride concentrations of 25, 50 and 100 mg F /l, the growth rates of
254
J.A. Camargo / Chemosphere 50 (2003) 251–264
the haptophyte P. lutheri, the dinoflagellate A. carteri
and the diatom N. angularis var. affinis were inhibited
25–30% at a fluoride concentration of 100 mg F /l. This
inhibitory effect on algal growth was not accompanied
by lethal signs of cell lysis. Oliveira et al. (1978) suggested that the growth stimulation in R. lens could be
due to a fluorine requirement for optimal growth of
certain microalgae. The other eight test algal species
showed good growth without indication of significant
(either inhibition or enhancement) effects.
Antia and Klut (1981) exposed, in seawater of 14–
15‰ salinity, five euryhaline phytoplankton algae
(Dunaliella tertiolecta, Thalassiosira weissflogii, Pavlova
lutheri, Chaetoceros gracilis and Amphidinium carteri) to
fluoride concentrations ranging from 50 to 200 mg F /l
for exposure times of up to 36 days. They found that,
while the population growth of the diatom C. gracilis
was considerably stimulated (averaging 45% in rate) by
the presence of inorganic fluoride (50–200 mg F /l), the
growth rates of the dinoflagellate A. carteri and the
haptophyte P. lutheri were inhibited 25–90% at fluoride
concentrations P 150 mg F /l. Antia and Klut (1981)
suggested that this inhibitory effect on algal growth
would be due to the fact that fluoride ions affect nucleotide and nucleic acid metabolism governing processes of algal cell division. The population growth of
the chlorophyte D. tertiolecta and the diatom T.
weissflogii was virtually unaffected at all fluoride concentrations.
Klut et al. (1981) found that, although the marine
dinoflagellate Amphidinium carteri was unable to grow
on nutrient-enriched seawater containing 200 mg F /l,
this marine phytoplankter could be adapted to grow
under such water conditions after repeatedly being cultured with stepwise increases in sub-inhibitory fluoride
concentrations. The process of adaptation to inorganic
fluoride by A. carteri may have involved some form of
genetic change thereby resulting presumably in the development of a mutant phenotype (Klut et al., 1981). In
inorganic fluoride-adapted dinoflagellate cells, electron
microscopic observations revealed abnormal ultrastructural features in the chloroplast (especially the pyrenoid), mitochondria and nucleus.
Hekman et al. (1984) examined the toxic effects of
inorganic fluoride on six freshwater phytoplankton
algae (Synechococcus leopoliensis, Oscillatoria limnetica,
Ankistrodesmus braunii, Scenedesmus quadricauda, Cyclotella meneghiniana and Stephanodiscus minutus) after
175 h of exposure to different F concentrations. They
found that, except in the case of S. leopoliensis, where
growth and photosynthesis inhibition was induced at
fluoride concentrations ranging from 25 to 50 mg F /l,
there was no significant effect for the other test algal
species at fluoride concentrations of up to 50 mg F /l.
Nichol et al. (1987) studied inorganic fluoride-resistant
cells of the freshwater alga Synechococcus leopoliensis
using the ratio isotope 18 F, and found that resistant cells
had a different, but unknown, process for controlling
the afflux of inorganic fluoride compared to sensitive
cells. Joy and Balakrishnan (1990) studied the effect of
inorganic fluoride on the freshwater diatom Nitzschia
palea and the brackish water diatom Amphora coffeaeformis during 96 h exposures in laboratory experiments. Significant enhancement of population growth
occurred in N. palea at fluoride concentrations between 10 and 100 mg F /l; in A. coffeaeformis significant
stimulation occurred at 70 mg F /l, and above 90 mg
F /l the growth declined. This growth enhancement
would have been brought about the action of F
through increasing cell multiplication (Joy and Balakrishnan, 1990).
LeBlanc (1984) estimated 96 h EC50 values, on the
basis of growth inhibition, to be 123 mg F /l for the
freshwater blue–green alga Selenastrum capricornutum
and 82 mg F /l for the marine alga Skeletonema costatum. This surprising result (i.e., the marine alga was
more sensitive than the freshwater alga) could be due to
the fact that, for both algal species, 96 h EC50 values are
related to soft water (<100 mg CaCO3 /l). Rai et al.
(1998) estimated 15 day EC50 values, based on algal
growth inhibition, for the freshwater green alga Chlorella vulgaris at different pH values: 266 mg F /l at pH
6.0 and 380 mg F /l at pH 6.8.
Regarding aquatic plants, Wang (1986), on the basis
of reduction in frond growth from the common duckweed (Lemna minor) toxicity bioassays, presented a
water quality standard (for the State of Illinois, USA) to
be 1.4 mg F /l. Wang (1986) also estimated a 96 h EC50
value to be higher than 60 mg F /l. Hocking et al. (1980)
reported that the fluoride content of the seaweeds
Fuchus distichus and Ectocarpus sp. decreased with increasing distance to an aluminum smelter operating on
the Kitimat River (BC, Canada). Shirke and Chandra
(1991), after studying the fluoride uptake by the duckweed Spirodela polyrrhiza under field and laboratory
conditions, found that the fronds of this aquatic plant
might contain high levels of fluoride: the maximum
fluoride level was 918 lg F /g dry weight after an exposure of 7 days to 20 mg F /l. Shirke and Chandra
(1991) also reported that there was no significant effect
on chlorophyll and protein contents in S. polyrrhiza.
Similarly, Sinha et al. (2000) studied the fluoride removal from contaminated water by the submerged
macrophyte Hydrilla verticillata under field and laboratory conditions. They found that the fluoride content
in this aquatic plant increased with increasing fluoride
concentration and exposure time: the fronds of H. verticillata could contain up to 1892 lg F /g dry weight
after an exposure of 7 days to 20 mg F /l. Sinha et al.
(2000) also reported that the chlorophyll and protein
contents in H. verticillata decreased (in contrast with the
findings for S. polyrrhiza).
J.A. Camargo / Chemosphere 50 (2003) 251–264
3. Bioaccumulation in aquatic invertebrates and fishes
Aquatic animals such as fish and invertebrates can
take up fluoride directly from the water or to a much
lesser extent via food (Neuhold and Sigler, 1960;
Hemens and Warwick, 1972; Nell and Livanos, 1988).
This fluoride uptake is a function of fluoride concentration in the aquatic medium, exposure time and water
temperature (Neuhold and Sigler, 1960; Moore, 1971;
Hemens and Warwick, 1972; Wright and Davison, 1975;
Milhaud et al., 1981; Pillai and Mane, 1985; Nell and
Livanos, 1988). Subsequently, although it may be eliminated (as F ions) via excretory systems (Kessabi,
1984), fluoride tends to be accumulated in the exoskeleton of invertebrates and in the bone tissue of fishes
(Tables 2 and 3). Fluoride accumulation in hard tissues
may be viewed as a defense mechanism against fluoride
intoxication because of the removal of fluoride from
body circulation (Sigler and Neuhold, 1972; Kessabi,
1984). In addition, fluoride accumulation can play an
important role in the hardening of hard tissues (the
exoskeleton of marine crustaceans, mainly) due to the
combination of fluoride with calcium and phosphorous
to give fluorapatite (Zhang et al., 1993; Sands et al.,
1998).
Hemens and Warwick (1972) found that estuarine
invertebrates and fish could accumulate relatively high
amounts of fluoride in their bodies after 72 days of exposure to a water mean fluoride concentration of 52 mg
F /l. Mean whole-body fluoride levels (as lg F /g ash)
255
were 1414 for the mud crab Tylodiplax blephariskios,
3116 for the sand shrimp Palaemon pacificus, 3248 for
the penaeid prawn Penaeus indicus and 7743 for the
striped mullet Mugil cephalus. Soevik and Braekkan
(1979), analysing fluoride levels in two krill species, estimated whole-body fluoride concentrations of 1500
and 2300 lg F /g dry weight for the Antarctic krill
Euphausia superba and the North Atlantic krill Meganyctiphanes norvegica, respectively. Oehlenschlager and
Manthey (1982), examining the fluoride content of three
benthic marine crustaceans, reported whole-body fluoride levels (as lg F /g dry weight) of 2200 and 2500
for two species of Paracerodocus and 900 for Serolis
cornuta. Pillai and Mane (1985), conducting laboratory
studies on the toxic effects of a fluoride effluent on fry of
the freshwater fish Catla catla, reported that fluoride
accumulation was relatively high after 96 h of exposure
to a fluoride concentration of 13.2 mg F /l. These fry
had a mean whole-body fluoride level of 766.7 lg F /g
ash, whereas control fry (exposed to a fluoride concentration < 0:6 mg F /l) had a mean whole-body fluoride
level of 89.5 lg F /g ash. Virtue et al. (1995), analysing
the biochemical composition of the marine crustacean
Nyctiphanes australis, found that this aquatic invertebrate could contain a whole-body fluoride level of up to
3507 lg F /g dry weight.
Aquatic invertebrates can however contain relatively
low levels of fluoride in their whole bodies. Spaargaren
(1988), studying halogen accumulation in the brown
shrimp Crangon crangon, found that its whole-body
Table 2
Fluoride levels (lg F /g dry weight) in marine crustaceans
Species
Whole-body
Exoskeleton
Muscle
Reference
Callinectes sapidus
Crangon vulgaris (wet weight)
Leander serratus (wet weight)
Portunus depurator (wet weight)
Euphausia superba
–
7.0
6.7
3.9
1500
1058
–
2300
2153
2200–2500
900
2.6
3507
1.44
0.87
–
510
171
2.9
118
–
–
298
17
18
17
–
2594
2232
–
3343
–
–
–
–
–
–
5977
1254
1351
362
4602
271
5104
10
2.7
3.3
2.2
–
4.5
74
–
5.7
–
–
–
–
–
–
–
24
–
–
–
0.78
257
Moore, 1971
Wright and Davison, 1975
Wright and Davison, 1975
Wright and Davison, 1975
Soevik and Braekkan, 1979
Adelung et al., 1987
Sands et al., 1998
Soevik and Braekkan, 1979
Adelung et al., 1987
Oehlenschlager and Manthey, 1982
Oehlenschlager and Manthey, 1982
Spaargaren, 1988
Virtue et al., 1995
Sands et al., 1998
Sands et al., 1998
Sands et al., 1998
Sands et al., 1998
Sands et al., 1998
Sands et al., 1998
Sands et al., 1998
Sands et al., 1998
Sands et al., 1998
Meganyctiphanes norvegica
Paracerodocus sp
Serolis cornuta
Crangon crangon
Nyctiphanes australis
Calanus propinquus
Euchaeta antarctica
Euphausia crystallorophias
Antarctomysis maxima
Cyllopus lucasii
Hyperia macrocephala
Cyphocaris micronyx
Notocrangon antarcticus
Thysanoessa macrura
256
J.A. Camargo / Chemosphere 50 (2003) 251–264
Table 3
Fluoride levels (mg F /kg wet weight) in freshwater and marine fishes
Species
Aquatic medium
Skeletal bone
Muscle
Reference
Cyprinus carpio
Oncorhynchus mykiss
Oreochromis leucostictus
Mugil auratus
Mugil cephalus
Mugil labrosus
Micromesistius australis
Notothenia gibberifons
Notothenia rossii
Notothenia neglecta
Chaenocephalus aceratus
Freshwater
Freshwater
Freshwater
Seawater
Seawater
Seawater
Seawater
Seawater
Seawater
Seawater
Seawater
1100
450
210.6
105–140
45–630
52–940
1207
1156
964
865
1143
20.8
2.9
1.97
2.8–3.3
1.3–26
1.5–15.3
1.4
1.3
2.2
3.7
1.8
Neuhold and Sigler, 1960
Neuhold and Sigler, 1960
Gikunju, 1992
Milhaud et al., 1981
Milhaud et al., 1981
Milhaud et al., 1981
Oehlenschlager and Manthey,
Oehlenschlager and Manthey,
Oehlenschlager and Manthey,
Oehlenschlager and Manthey,
Oehlenschlager and Manthey,
fluoride concentration was 2.6 lg F /g dry weight.
Sands et al. (1998), examining fluoride levels in a range
of Antarctic marine crustaceans, reported that copepods
had the lowest fluoride concentrations: whole-body
levels of 1.44 and 0.87 lg F /g dry weight were measured for Calanus propinquus and Euchaeta antarctica,
respectively.
Other studies have focused on the differential bioaccumulation of fluoride in several tissues of aquatic animals. Moore (1971), studying the uptake and content
of fluoride by the marine blue crab Callinectes sapidus,
found that fluoride was mainly accumulated in the
exoskeleton (298 lg F /g dry weight) and less in the
muscle (10 lg F /g dry weight). Wright and Davison
(1975), examining the accumulation of fluoride by several marine crustaceans inhabiting an area (seawater
with a fluoride concentration of 3.4 mg F /l) close to an
aluminium smelter, reported that fluoride was largely
accumulated in the exoskeleton. Higher exoskeleton
fluoride levels (as lg F /g wet weight) for the prawn
Leander serratus, the swimming crab Portunus depurator
and the shrimp Crangon vulgaris were 18, 17 and 17,
respectively. In contrast, higher muscle fluoride levels (as
lg F /g wet weight) were 3.3, 2.2 and 2.7, respectively.
Adelung et al. (1987) analysed fluoride content in two
krill species and found that, although internal organs
(muscle) contained less than 10 lg F /g dry weight, the
concentration of inorganic fluoride in the cuticle (exoskeleton) was much higher: 2594 lg F /g dry weight for
Euphausia superba and 3343 lg F /g dry weight for
Meganyctiphanes norvegica. Sands et al. (1998) examined the concentration of fluoride in several Antarctic
marine crustaceans (Table 2). They reported that the
exoskeleton of euphausiids had the highest overall fluoride concentrations: levels of up to 5977 lg F /g dry
weight were found in the exoskeleton of Euphausia
crystallorophias; in the mouthparts of E. superba a level
of 12 876 lg F /g dry weight was quantified. According
to Sands et al. (1998), the higher fluoride concentration
found in the mouthparts of Antarctic krill would indicate that inorganic fluoride plays an important role in
1982
1982
1982
1982
1982
the hardening of the exoskeleton since those body parts
are, by necessity, the hardest parts of the exoskeleton.
The major inorganic constituents of the krill exoskeleton, Ca and P, would combine with fluoride to give
Ca5 (PO4 )3 F, a known hardener (Zhang et al., 1993).
In fish, Neuhold and Sigler (1960), conducting laboratory experiments to examine the effects of sodium
fluoride on the carp Cyprinus carpio and the rainbow
trout Oncorhynchus mykiss (Salmo gairdneri, previously), found that the skeletal bone accumulated much
higher fluoride levels (range 450–1100 mg F /kg wet
weight) than the muscle (range 2.9–20.8 mg F /kg wet
weight). They concluded that the incorporation of fluoride into skeletal bone was by an apparent secondorder reaction, suggesting active transport. Milhaud
et al. (1981), comparing fluoride levels in mullet species
(Mugil auratus, M. cephalus and M. labrosus) from
Gabes Gulf (South Tunisia), where inorganic fluoriderich effluents were discharged (fluoride concentration in
seawater was about 2–3 mg F /l), with fluoride levels in
mullets from the Tunis Bay, which is remote from point
sources (fluoride concentration in seawater was about
1.4 mg F /l), found that the mullets from Gabes Gulf
had mean fluoride levels (as mg F /kg wet weight) of 320
for bone and 9.6 for muscle, whereas the mullets from
Tunis Bay had mean fluoride levels (as mg F /kg wet
weight) of 73 for bone and 1.8 for muscle. Oehlenschlager and Manthey (1982), studying tissue levels of
inorganic fluoride in marine fishes (Micromesistius australis, Notothenia gibberifons, N. rossii, N. neglecta and
Chaenocephalus aceratus), found that vertebra fluoride
levels were in the range 865–1207 mg F /kg wet weight
and muscle fluoride levels were in the range 1.3–3.7 mg
F /kg wet weight. Gikunju (1992), examining the fluoride concentration in tilapia fish (Oreochromis leucostictus) from Naivasha lake (mean freshwater fluoride
concentration of 2.5 mg F /l), in Kenya, found fluoride
levels (as mg F /kg wet weight) of 210.6, 143.1, 4.96 and
1.97 for bone, gills, skin and muscle, respectively.
Some aquatic invertebrates can accumulate relatively
high concentrations of fluoride in their soft tissues.
J.A. Camargo / Chemosphere 50 (2003) 251–264
Barbaro et al. (1981), working in the Lagoon of Venice,
found that soft tissues of the barnacle Balanus amphitrite
and the mussel Mytilus galloprovincialis could contain
up to 85 lg F /g dry weight. Nell and Livanos (1988),
conducting laboratory experiments with juvenile individuals of two oyster species, Saccostrea commercialis
and Ostrea angasi, reported a linear relationship between fluoride tissue levels and fluoride seawater levels:
soft tissues could contain from 45 to 204 lg F /g dry
weight with increasing seawater fluoride additions from
0 to 30 mg/l. Furthermore, fluoride accumulation in S.
commercialis also increased with increasing water temperature as an expected consequence of increased metabolism.
4. Toxicity to aquatic invertebrates
The toxic action of fluoride on the health of aquatic
(and terrestrial) animals resides in the fact that fluoride
ions act as enzymatic poisons, inhibiting enzyme activity
(e.g., phosphatases, hexokinase, enolase, succinic dehydrogenase, pyruvic oxidase) and, ultimately, interrupting metabolic processes such as glycolysis and synthesis
of proteins (Kessabi, 1984). Nevertheless, the actual
mechanism at molecular level by which inorganic fluoride inhibits enzyme activity and interrupts metabolic
processes remains a matter of speculation. The combination of fluoride with calcium and magnesium ions,
which in many instances are needed as cofactors by
different enzymes, might be a primary mechanism
(Kessabi, 1984).
Fluoride toxicity to aquatic invertebrates increases
with increasing fluoride concentration, exposure time
and water temperature (LeBlanc, 1980; Dave, 1984;
Fieser et al., 1986; K€
uhn et al., 1989; Camargo and
Tarazona, 1990; Camargo, 1991b; Camargo et al.,
1992a,b). Although, as far as I know, there is no empirical evidence (as there is for fish) demonstrating that
fluoride toxicity decreases with increasing the water
content of calcium, it is acceptable to assume that such a
decreased toxicity may also occur in aquatic invertebrates. Estuarine and marine invertebrates may indeed
be more tolerant to fluoride toxicity than freshwater
invertebrates as a probable consequence of the elevated
calcium content in estuarine and seawaters (Hemens and
Warwick, 1972; Hemens et al., 1975; Pankhurst et al.,
1980; Connell and Airey, 1982; McClurg, 1984). In addition, recent laboratory studies (Camargo, 2002) have
shown that the toxicity of fluoride ions to aquatic invertebrates (insect larvae) may decrease with increasing
intraspecific body size and water content of chloride.
The water flea Daphnia magna has been the most
widely used invertebrate species to study the toxicity
of inorganic fluoride in acute laboratory experiments
(Table 4). 24 and 48 h LC50 values (for immobilisation)
257
vary between 205–352 and 98–304 mg F /l, respectively,
48 h LC50 values decreasing with increasing water temperature (Fieser et al., 1986). Furthermore, LeBlanc
(1980) reported a 48 h NOEC of 50 mg F /l, and K€
uhn
et al. (1989) reported a 24 h NOEC of 231 mg F /l. In
chronic (21 days) laboratory experiments, Dave (1984)
estimated a fluoride safe concentration of 4.40 mg F /l
for parthenogenetic reproduction in D. magna (water
temperature of 20 °C and water hardness of 250 mg
CaCO3 /l). Fluoride concentrations below this safe level
could stimulate reproduction due probably to hormetic
effects (Dave, 1984).
Nevertheless, from a simple comparison of LC50
values (Table 4), it is evident that caddisfly larvae (Trichoptera; Insecta) may be more sensitive to fluoride
toxicity than neonates of D. magna. In general, these
caddisfly larvae migrated from their retreat and capture
nets and protruded their anal papillae before dying
(Camargo et al., 1992a,b). Protruded anal papillae were
darkened in larvae of Hydropsyche species exposed to
the highest fluoride concentrations (Camargo et al.,
1992a,b). On the basis of acute mortality data and using
the multifactor probit analysis software (USEPA, 1991;
Lee et al., 1995), safe concentrations (infinite hour LC0:01
values) of fluoride in soft water (15.6–40.2 mg CaCO3 /l)
were estimated for Palearctic (Camargo and La Point,
1995) and Nearctic (Camargo, 1996b) caddisfly larvae:
1.79 (0.88–2.94) mg F /l for Chimarra marginata, 1.18
(0.47–2.18) mg F /l for Hydropsyche lobata, 0.73 (0.32–
1.28) mg F /l for H. bulbifera, 0.56 (0.22–1.06) mg F /l
for H. exocellata, 0.39 (0.13–0.84) mg F /l for H. pellucidula, 0.70 (0.40–1.20) mg F /l for H. occidentalis,
0.20 (0.10–0.40) mg F /l for H. bronta, and 0.70 (0.30–
1.30) mg F /l for Cheumatopsyche occidentalis.
Camargo (2002), carrying out laboratory studies with
net-spinning caddisfly larvae of Hydropsyche tibialis, has
recently found that the intraspecific tolerance of these
larvae to fluoride toxicity increases with increasing body
size and water content of chloride. The ameliorating
effect of chloride would be primarily due to a constraint
in the incorporation of fluoride ions into the cytosolic
side of cell membrane because of the presence of chloride ions at the external side of cell membrane. If we
assume that fluoride ions can enter ion-uptake cells in
chloride epithelia by active transport, which is accomplished by carriers that have affinity for chloride ions at
the external side of cell membrane, then an elevated Cl
level in the aquatic environment would induce an increased competition between Cl and F ions at the
external side and, consequently, a decreased fluoride
influx at the cytosolic side. Chloride epithelia, as ionuptake sites, are common in many aquatic insect species
(Komnick, 1977). The ameliorating effect of body size
might be due to a better developed osmoregulatory
ability in last instar larvae and, thereby, a better ability
to eliminate fluoride ions from the body.
258
J.A. Camargo / Chemosphere 50 (2003) 251–264
Table 4
LC50 values of fluoride at different exposure times for freshwater and marine invertebrates
Species
Developmental stage
Temperature
(°C)
Total hardness
(mg CaCO3 /l)
LC50 value
(mg F /l)
Reference
Daphnia magna
Neonates
Neonates
Neonates
Neonates
Neonates
Neonates
Neonates
Neonates
h)
h)
h)
h)
h)
h)
h)
h)
23.2
23.2
20
20
15
20
25
20
173
173
250
250
169.3
169.3
169.3
160
308 (24 h)
154 (48 h)
205 (24 h)
98 (48 h)
304 (48 h)
251 (48 h)
200 (48 h)
352 (24 h)
LeBlanc, 1980
LeBlanc, 1980
Dave, 1984
Dave, 1984
Fieser et al., 1986
Fieser et al., 1986
Fieser et al., 1986
K€
uhn et al., 1989
Chimarra marginata
Last instar larvae
Last instar larvae
Last instar larvae
15.2
15.2
15.2
15.6
15.6
15.6
120 (48 h)
79.7 (72 h)
44.9 (96 h)
Camargo, 1991b
Camargo, 1991b
Camargo and Tarazona,
1990
Cheumatopsyche
pettiti
Last
Last
Last
Last
Last
18
18
18
18
18
40.2
40.2
40.2
40.2
40.2
128 (48 h)
73.2 (72 h)
42.5 (96 h)
31.9 (120 h)
24.2 (144 h)
Camargo
Camargo
Camargo
Camargo
Camargo
Hydropsyche
bulbifera
Last instar larvae
Last instar larvae
Last instar larvae
15.2
15.2
15.2
15.6
15.6
15.6
79.2 (48 h)
44.9 (72 h)
26.3 (96 h)
Camargo, 1991b
Camargo, 1991b
Camargo and Tarazona,
1990
Hydropsyche
exocellata
Last instar larvae
Last instar larvae
Last instar larvae
15.2
15.2
15.2
15.6
15.6
15.6
86.6 (48 h)
43.7 (72 h)
26.5 (96 h)
Camargo, 1991b
Camargo, 1991b
Camargo and Tarazona,
1990
Hydropsyche lobata
Last instar larvae
Last instar larvae
15.2
15.2
15.6
15.6
78.2 (72 h)
48.2 (96 h)
Camargo, 1991b
Camargo and Tarazona,
1990
Hydropsyche
pellucidula
Last instar larvae
Last instar larvae
Last instar larvae
15.2
15.2
15.2
15.6
15.6
15.6
112 (48 h)
59.1 (72 h)
38.5 (96 h)
Camargo, 1991b
Camargo, 1991b
Camargo and Tarazona,
1990
Hydropsyche bronta
Last
Last
Last
Last
Last
instar
instar
instar
instar
instar
larvae
larvae
larvae
larvae
larvae
18
18
18
18
18
40.2
40.2
40.2
40.2
40.2
52.6
25.8
17.0
13.4
11.5
(48 h)
(72 h)
(96 h)
(120 h)
(144 h)
Camargo
Camargo
Camargo
Camargo
Camargo
et
et
et
et
et
al.,
al.,
al.,
al.,
al.,
1992a
1992a
1992a
1992a
1992a
Hydropsyche
occidentalis
Last
Last
Last
Last
Last
instar
instar
instar
instar
instar
larvae
larvae
larvae
larvae
larvae
18
18
18
18
18
40.2
40.2
40.2
40.2
40.2
102 (48 h)
53.5 (72 h)
34.7 (96 h)
27.0 (120 h)
21.4 (144 h)
Camargo
Camargo
Camargo
Camargo
Camargo
et
et
et
et
et
al.,
al.,
al.,
al.,
al.,
1992a
1992a
1992a
1992a
1992a
Mysidopsis bahia
Penaeus indicus
–
–
–
–
Seawater
Seawater
10.5 (96 h)
1118 (96 h)
LeBlanc, 1984
McClurg, 1984
(<24
(<24
(<24
(<24
(<24
(<24
(<24
(<24
instar
instar
instar
instar
instar
larvae
larvae
larvae
larvae
larvae
et
et
et
et
et
al.,
al.,
al.,
al.,
al.,
1992a
1992a
1992a
1992a
1992a
In all cases, test fluoride solutions were made using sodium fluoride (NaF).
On the other hand, estuarine and marine invertebrates appear to be more tolerant to fluoride toxicity
than freshwater invertebrates. Hemens and Warwick
(1972) reported that no toxic effects were observed in
two species of estuarine prawns (Penaeus indicus and P.
monodon) after exposures of 96 h to a maximum fluoride
concentration of 100 mg F /l (water temperature of 20.5
°C). Hemens et al. (1975) found that exposures to 5.5 mg
J.A. Camargo / Chemosphere 50 (2003) 251–264
F /l for 113 days did not cause mortality or obvious
physical deterioration in two estuarine crustaceans, the
mud crab Tylodiplax blephariskios and the penaeid
prawn Penaeus indicus (water temperature in the
range 23.5–27.0 °C). Pankhurst et al. (1980) found that
fluoride concentrations of up to 100 mg F /l caused
negligible mortality on the anemone Anthopleura aureoradiata (after 144 h), the bivalve Mytilus edulis (after
160 h) and the red krill Munida gregaria (after 259 h).
Connell and Airey (1982), studying the chronic effects of
fluoride on two species of estuarine amphipods, Grandidierella lutosa and G. lignorum, in a 90 day life-cycle
bioassay (water temperature in the range 23.5–25.0 °C),
reported that female fecundity was the most sensitive
biological parameter to fluoride toxicity. On the basis of
chronic data on female fecundity, Connell and Airey
(1982) proposed a maximum acceptable toxicant concentration of 4.15 mg F /l for both amphipod species.
Fluoride concentrations below this MATC value might
stimulate female fecundity (Connell and Airey, 1982).
McClurg (1984) estimated a 96 h LC50 of 1118 mg F /l
for the estuarine prawn Penaeus indicus (Table 4), and
found that fluoride concentrations of up to 11 mg F /l
had no significant effects on the growth of this penaeid
prawn over a 20 day exposure period.
Some estuarine and marine invertebrates can however exhibit a relatively higher sensitivity to fluoride
toxicity. The brown mussel Perna perna showed up to
30% mortality after an exposure of 120 h to an initial
fluoride concentration of 7.2 mg F /l and a water temperature of 20.5 °C (Hemens and Warwick, 1972).
Larvae of the brine shrimp Artemia salina, exposed to a
fluoride concentration of 5.0 mg F /l for 288 h, showed
significant growth impairment (Pankhurst et al., 1980).
LeBlanc (1984) reported a 96 h LC50 as low as 10.5 mg
F /l for the saltwater mysidacean Mysidopsis bahia
(Table 4).
5. Toxicity to fishes
Fluoride toxicity to fishes increases with increasing
fluoride concentration in the aquatic medium, exposure
time and water temperature (Neuhold and Sigler, 1960;
Angelovic et al., 1961; Hemens and Warwick, 1972;
Wright, 1977; Heitmuller et al., 1981; Pimentel and
Bulkley, 1983; Smith et al., 1985; Camargo, 1991a;
Camargo and Tarazona, 1991). Conversely, fluoride
toxicity decreases with increasing intraspecific fish size
and water content of calcium and chloride (Neuhold and
Sigler, 1960, 1962; Pimentel and Bulkley, 1983; Smith
et al., 1985). Lethal concentrations of fluoride, causing
acute intoxication in fish, bring on symptoms of fluorosis before death (Sigler and Neuhold, 1972; Smith
et al., 1985; Camargo and Tarazona, 1991): initial lethargy and apathetic behaviour accompanied by anorexia,
259
hypoexcitability, decreased respiratory rates, increased
fluoride levels in the blood, a darkening of the skin
(dorsal side) and an increased mucus secretion (from the
respiratory and integumentary epithelium); then violent
aimless movements with a loss of equilibrium; and finally death in a state of partial or complete muscular
contraction/paralysis (tetany). In addition, elevated fluoride concentrations can cause a delay in the hatching of
fertilised eggs of freshwater fish (Ellis et al., 1948; Pillai
and Mane, 1984).
The rainbow trout Oncorhynchus mykiss (Salmo
gairdneri, previously) has been the most widely used fish
species to study the toxicity of inorganic fluoride in
laboratory experiments. From a simple comparison of
LC50 values (Table 5) it is evident that O. mykiss seems
to be more sensitive to fluoride toxicity than other
freshwater fish species (the brown trout Salmo trutta,
the carp Cyprinus carpio, the threespine strickleback
Gasterosteus aculeatus and the fathead minnow Pimephales promelas). The ameliorating effect of water hardness on the tolerance of O. mykiss and G. aculeatus to
fluoride toxicity (Table 5) is mainly due to the formation/precipitation of innocuous complexes such as
Ca5 (PO4 )3 F, CaF2 and MgF2 (Pimentel and Bulkley,
1983; Smith et al., 1985). Hard waters can also serve as
calcium and magnesium reservoirs for fish whose systemic calcium and magnesium levels are depleted as a
result of fluoride intoxication (Pimentel and Bulkley,
1983). A higher tolerance (within the same species) in
larger individuals of O. mykiss and C. carpio to fluoride
toxicity (Table 5) would be due to a higher capability in
larger fishes to either eliminate or immobilise/accumulate fluoride ions (Neuhold and Sigler, 1960; Sigler and
Neuhold, 1972).
Not only water hardness and fish size, but also water
temperature and chloride content, can affect fish tolerance to fluoride toxicity. Angelovic et al. (1961) found
that the mean time period to the initial mortality of
rainbow trout (7.6–12.7 cm), exposed to a maximum
fluoride concentration of 25 mg F /l for 10 days, decreased with increasing water temperature (range 7.2–
23.9 °C). This fact was likely due to an increase in the
trout metabolic rate which would increase the uptake of
fluoride ions from the aquatic medium. Neuhold and
Sigler (1962) reported that mortality of rainbow trout
(10–18 cm), exposed to a maximum fluoride concentration of 25 mg F /l for 5 days, decreased with increasing
water content of chloride (range 0–9 mg Cl /l). They
concluded that a moderate increase in the concentration
of chloride ions might elicit a response in trout for fluoride excretion. In this respect, it is likely that the cellular mechanism responsible for the ameliorating effect
of chloride on the tolerance of trouts to fluoride toxicity
could be similar to that for the ameliorating effect of
chloride on the tolerance of net-spinning caddisfly larvae
to fluoride toxicity (as suggested above).
260
J.A. Camargo / Chemosphere 50 (2003) 251–264
Table 5
LC50 values of fluoride at different exposure times for freshwater fishes
Species
Body size
Temperature
(°C)
Total
hardness
(mg CaCO3 /l)
LC50 value
(mg F /l)
Reference
Oncorhynchus mykiss
10–20 cm
5.4–6.3 cm
5.4–6.3 cm
5.4–6.3 cm
5.4–6.3 cm
<3 g
8–10 cm
8–10 cm
8–10 cm
8–10 cm
8–10 cm
8–10 cm
12.8
12
12
12
12
15
15.3
15.3
15.3
15.3
15.3
15.3
<3
17
49
182
385
23–62
22.4
22.4
22.4
22.4
22.4
22.4
2.7–4.7 (480 h)
51 (96 h)
128 (96 h)
140 (96 h)
193 (96 h)
200 (96 h)
138.5 (72 h)
107.5 (96 h)
92.4 (120 h)
85.1 (144 h)
73.4 (168 h)
64.1 (192 h)
Neuhold and Sigler, 1960
Pimentel and Bulkley, 1983
Pimentel and Bulkley, 1983
Pimentel and Bulkley, 1983
Pimentel and Bulkley, 1983
Smith et al., 1985
Camargo, 1991a
Camargo, 1991a
Camargo and Tarazona, 1991
Camargo and Tarazona, 1991
Camargo and Tarazona, 1991
Camargo and Tarazona, 1991
Salmo trutta
8–10
8–10
8–10
8–10
8–10
8–10
16.1
16.1
16.1
16.1
16.1
16.1
21.2
21.2
21.2
21.2
21.2
21.2
223.0 (72 h)
164.5 (96 h)
135.6 (120 h)
118.5 (144 h)
105.1 (168 h)
97.5 (192 h)
Camargo, 1991a
Camargo, 1991a
Camargo and Tarazona,
Camargo and Tarazona,
Camargo and Tarazona,
Camargo and Tarazona,
cm
cm
cm
cm
cm
cm
1991
1991
1991
1991
Cyprinus carpio
10–36 cm
18.3
<3
75–91 (480 h)
Neuhold and Sigler, 1960
Gasterosteus aculeatus
<1 g
<1 g
<1 g
20
20
20
78
146
300
340 (96 h)
380 (96 h)
460 (96 h)
Smith et al., 1985
Smith et al., 1985
Smith et al., 1985
Pimephales promelas
<1 g
16–20
20–48
315 (96 h)
Smith et al., 1985
In all cases, test fluoride solutions were made using sodium fluoride (NaF).
With regard to the effect of fluoride on egg hatching,
Ellis et al. (1948) indicated that fertilised eggs of rainbow trout, exposed to a fluoride concentration of 1.5 mg
F /l, were delayed from 7 to 10 days in hatching beyond
comparable eggs in water containing no fluorides. Pillai
and Mane (1984) reported that the hatching of fertilised
eggs of the freshwater fish C. catla was delayed by 1 h at
a fluoride concentration of 3.56 mg F /l, and by 2 h at
fluoride concentrations of 7.34, 10.01 and 16.68 mg
F /l. The eggs of C. catla showed significant decreases
in water and protein and an increase in fluoride content.
Safe concentrations of fluoride have been proposed
for trout species. Pimentel and Bulkley (1983), using a
general application factor of 0.05 times the 96 h LC50
value, recommended maximum chronic exposure levels
of fluoride for fry (5.4–6.3 cm) of rainbow trout: 2.5 mg
F /l at a water hardness of 17 mg CaCO3 /l and 9.6 mg
F /l at a hardness of 385 mg CaCO3 /l. Camargo (1996a),
on the basis of acute mortality data and using the
multifactor probit analysis software (USEPA, 1991; Lee
et al., 1995), estimated safe concentrations (infinite hour
LC0:01 values) of fluoride for fry (8–10 cm) of rainbow
trout and brown trout in soft water (21.2–22.4 mg
CaCO3 /l): 5.14 (3.10–7.53) mg F /l for O. mykiss and
7.49 (4.42–10-96) mg F /l for S. trutta.
These estimated safe concentrations of fluoride for
trout species would however be insufficient to protect
other salmonid species. Damkaer and Dey (1989), conducting bioassay experiments on the behaviour of upstream-migrating adult salmons exposed to elevated
fluoride concentrations in the Columbia river (very soft
water with low ionic content), found that fluoride concentrations of about 0.5 mg F /l could adversely affect
the upstream migration of adult individuals of the chinook salmon Oncorhynchus tshawytscha and the coho
salmon Oncorhynchus kisutch. They concluded that a
fluoride concentration as low as 0.2 mg F /l may be near
or below the threshold for fluoride sensitivity in these
two salmon species. In contrast with this relatively high
sensitivity to fluoride, field studies have shown that
healthy, growing populations of rainbow trout exist in
the Firehole river (up to 14 mg F /l) in Yellowstone
National Park and in Walker and Pyramid lakes (up to
13 mg F /l) in Nevada (Sigler and Neuhold, 1972). It
should be evident that physiological and genetic adaptation to high fluoride concentrations can occur in wild
fish populations.
Estuarine and marine fishes, on the other hand,
appear to be more tolerant to fluoride toxicity than
freshwater fishes as a probable consequence of the ele-
J.A. Camargo / Chemosphere 50 (2003) 251–264
vated content of calcium and chloride in estuarine and
seawaters. Hemens and Warwick (1972) reported that
no mortality was observed in three estuarine fish species (the ambassid Ambassis safgha, the crescent perch
Therapon jarbua and the striped mullet Mugil cephalus)
after exposures of 96 h to a maximum fluoride concentration of 100 mg F /l (water temperature of 20.5 °C).
Hemens et al. (1975) found that a exposure to 5.5 mg
F /l for 113 days did not cause mortality or obvious
physical deterioration in juvenile individuals of the
striped mullet Mugil cephalus (water temperature in the
range 23.5–27.0 °C). Heitmuller et al. (1981), performing
marine short-term (4 days) toxicity bioassays, reported a
NOEC of 226 mg F /l for juvenile individuals of the
sheepshead minnow Cyprinodon variegatus.
6. Ecological risk assessment
Human activities, such as aluminium smelters, discharges of fluoridated municipal waters, and plants
manufacturing brick, ceramics, glass and fluoride
chemicals, may cause significant increases in the fluoride
concentration of surface waters (Wright and Davison,
1975; Oliveira et al., 1978; Pankhurst et al., 1980; Ares
et al., 1983; Somashekar and Ramaswamy, 1983; Sparks
et al., 1983; Van Craenenbroeck and Marivoet, 1987;
Zingde and Mandalia, 1988; Camargo et al., 1992a;
Karunagaran and Subramanian, 1992; Warrington,
1992; CEPA, 1994; Skjelk
avle, 1994; Camargo, 1996a;
Roy et al., 2000). In some cases, increased fluoride
concentrations higher than 10 mg F /l have been measured in recipient rivers (Van Craenenbroeck and
Marivoet, 1987; CEPA, 1994; Camargo, 1996a). Consequently, and even though safe levels of fluoride for
aquatic life have not been yet determined (USEPA,
1986), it should be evident from the data presented in
this review that discharges from anthropogenic sources
(i.e., fluoride pollution) may result in a serious ecological
risk for aquatic organisms.
Aquatic organisms living in soft waters with low
ionic content may be more adversely affected by fluoride
pollution than those living in hard or seawaters because
the bioavailability of fluoride ions is reduced with increasing water content of calcium and chloride (Neuhold
and Sigler, 1962; Sigler and Neuhold, 1972; Pimentel
and Bulkley, 1983; Smith et al., 1985; Camargo, 2002).
Furthermore, freshwater animals are, in general, more
sensitive to fluoride toxicity than freshwater algae and
macrophytes (Tables 1, 4 and 5). And among freshwater
animals, net-spinning caddisfly larvae (Camargo and La
Point, 1995; Camargo, 1996b) and upstream-migrating
adult salmons (Damkaer and Dey, 1989) seem to be the
most sensitive. In consequence, safe levels of fluoride for
aquatic life must be primarily based on the tolerance of
261
net-spinning caddisfly larvae and upstream-migrating
adult salmons.
Because a fluoride concentration as low as 0.5 mg
F /l can adversely affect net-spinning caddisfly larvae
(Camargo and La Point, 1995; Camargo, 1996b) and
upstream-migrating adult salmons (Damkaer and Dey,
1989) inhabiting soft waters with low ionic content, safe
levels below this fluoride concentration are recommended in order to protect these freshwater animals
from fluoride pollution. In the case of hard waters with
high ionic content, safe levels of fluoride could be increased up to 2–3 times the fluoride concentration of
0.5 mg F /l. However, in spite of this proposal of safe
levels, it is necessary to point out certain future research
needs on the toxicity of fluoride to aquatic organisms.
There is a real need for studies on the toxicity of fluoride to aquatic decomposers such as saprophytic fungi
and bacteria. These microorganisms play an important
role in the structure and function of aquatic ecosystems
through the recycling of nutrients (Wetzel, 1983; Allan,
1995). It should also be evident that more studies on the
(acute and chronic) toxicity of fluoride to aquatic producers, specially macrophytes, are needed. At present, the
actual mechanisms at molecular level by which inorganic
fluoride can either inhibit or enhance algal population
growth are still unknown. It would indeed be interesting
to examine if fluoride toxicity to algae and macrophytes
decreases with increasing water content of calcium and
chloride (as it occurs in aquatic animals). Regarding
aquatic animals, most studies have concerned with acute/
short-term laboratory experiments and, consequently,
more chronic/long-term studies would be needed to verify
and improve safe levels of fluoride. Lastly, because inorganic fluoride bioaccumulates in aquatic producers
(algae and macrophytes) as well as in aquatic consumers
(invertebrates and fishes), studies on the biotransference
and biomagnification of fluoride through food chains
and food webs would have to be carried out.
Acknowledgements
I am particularly grateful to University of Alcala for
providing funds and logistical support to carry out this
study. I am also grateful to the World Health Organization for inviting me to participate as an adviser in the
Task Group on Environmental Health Criteria for
Fluorides.
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Julio A. Camargo teaches ecotoxicology, limnology and evolutionary ecology at the University of Alcala. He has research
priorities in aquatic ecology and toxicology, with the primary
goal of preserving freshwater ecosystems.