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Human and Ecological Risk Assessment: An International Journal
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Environmental Risk Assessment of Trifluoroacetic Acid
Jean Charles Boutonnet
a
; Pauline Bingham
b
; Davide Calamari
c
; Christ de Rooij
d
; James Franklin
d
;
Toshihiko Kawano
e
; Jean-Marie Libre
f
; Archie McCul-loch
g
; Giuseppe Malinverno
h
; J. Martin Odom
i
; George M. Rusch
j
; Katie Smythe
k
; Igor Sobolev
l
; Roy Thompson
m
; James M. Tiedje
n
a
Elf Atochem S.A., Centre d'Application de Levallois, 95 rue Danton, BP 108, 92303 Levallois-Perret
Cedex, France.
b
Rhodia, Ltd., St. Andrews Road, Avon-mouth, Bristol BS11 9YF U.K..
c
Department of
Structural and Functional Biology, University of Milan, Via Ravasi 2, 21100 Varese VA Italy.
d
Solvay
S.A., rue de Ransbeek 310, B-1120 Brussels, Belgium.
e
Daikin Industries Ltd., 1-1 Nishi-Hitotsuya,
Settsu-shi, Osaka 566 Japan.
f
Elf Atochem S.A., 4 Cours Michelet, La Défense 10, Cédex 42 92091 Paris
la Défense, France.
g
ICI Chemicals & Polymers Ltd., Environment Department, P.O. Box 13, The
Heath, Runcorn, Cheshire WA7 4QF U.K..
h
Ausimont S.p.A, via Lombardia 20, 20021 Bollate (Milan),
Italy.
i
E.I. DuPont de Nemours & Company, Experimental Station (E328-B47), Wilmington, DE 19898.
j
1Allied Signal Inc., P.O. Box 1139, Morristown, NJ 07962-1139.
k
AFEAS Program Office, 1333 H Street
NW, Washington, DC 20005.
l
Chemical & Polymer Technology Inc., 5 Rita Way, Orinda, CA 94563.
m
Zeneca Ltd., Brixham Environmental Laboratory, Freshwater Quarry, Brix-ham, Devon TQ5 8BA U.K..
n
Center for Microbial Ecology, Michigan State University, East Lansing, MI 48824-1326.
To cite this Article Boutonnet, Jean Charles, Bingham, Pauline, Calamari, Davide, de Rooij, Christ, Franklin, James,
Kawano, Toshihiko, Libre, Jean-Marie, McCul-loch, Archie, Malinverno, Giuseppe, Odom, J. Martin, Rusch, George M.,
Smythe, Katie, Sobolev, Igor, Thompson, Roy and Tiedje, James M.(1999) 'Environmental Risk Assessment of
Trifluoroacetic Acid', Human and Ecological Risk Assessment: An International Journal, 5: 1, 59 — 124
To link to this Article: DOI: 10.1080/10807039991289644
URL: http://dx.doi.org/10.1080/10807039991289644
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Human and Ecological Risk Assessment: Vol. 5, No. 1, pp. 59-124 (1999)
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Reproduction of this material without the consent of the publisher is prohibited.
Environmental Risk Assessment of Trifluoroacetic
Acid
Jean Charles Boutonnet (Ed.),
1
Pauline Bingham,
2
Davide Calamari,
3
Christ de
Rooij,
4
James Franklin,
5
Toshihiko Kawano,
6
Jean-Marie Libre,
7
Archie McCul-
loch,
8
Giuseppe Malinverno,
9
J. Martin Odom,
10
George M. Rusch,
11
Katie
Smythe,
12,
* Igor Sobolev,
13
Roy Thompson,
14
and James M. Tiedje
15
1
Elf Atochem S.A., Centre d’Application de Levallois, 95 rue Danton, BP 108,
92303 Levallois-Perret Cedex, France;
2
Rhodia, Ltd., St. Andrews Road, Avon-
mouth, Bristol BS11 9YF U.K.;
3
Department of Structural and Functional Biol-
ogy, University of Milan, Via Ravasi 2, 21100 Varese VA Italy;
4
Solvay S.A., rue
de Ransbeek 310, B-1120 Brussels, Belgium;
5
Solvay S.A., rue de Ransbeek 310,
B-1120 Brussels, Belgium;
6
Daikin Industries Ltd., 1-1 Nishi-Hitotsuya, Settsu-
shi, Osaka 566 Japan;
7
Elf Atochem S.A., 4 Cours Michelet, La Défense 10, Cé-
dex 42 92091 Paris la Défense, France;
8
ICI Chemicals & Polymers Ltd., Envi-
ronment Department, P.O. Box 13, The Heath, Runcorn, Cheshire WA7 4QF
U.K.;
9
Ausimont S.p.A, via Lombardia 20, 20021 Bollate (Milan), Italy;
10
E.I.
DuPont de Nemours & Company, Experimental Station (E328-B47), Wilming-
ton, DE 19898;
11
Allied Signal Inc., P.O. Box 1139, Morristown, NJ 07962–
1139;
12
AFEAS Program Office, 1333 H Street NW, Washington, DC 20005;
13
Chemical & Polymer Technology Inc., 5 Rita Way, Orinda, CA 94563;
14
Zeneca Ltd., Brixham Environmental Laboratory, Freshwater Quarry, Brix-
ham, Devon TQ5 8BA U.K.;
15
Center for Microbial Ecology, Michigan State
University, East Lansing, MI 48824–1326
ABSTRACT
The Montreal Protocol was developed in 1987 in response to concerns that
the chlorofluorocarbons (CFCs) were releasing chlorine into the stratosphere
and that this chlorine was causing a depletion of stratospheric ozone over Ant-
arctica. This international agreement called for a phase out of these CFCs. In-
dustry initiated a major effort to find replacements that are safe when properly
* Primary contact.
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Hum. Ecol. Risk Assess. Vol. 5, No. 1, 1999
60
used and safe to the environment. The toxicology and environmental fate of
these first generation replacements has been studied extensively.
It was determined that the new substances break down in the environment
to give predominantly carbon dioxide, water and inorganic salts of chlorine
and fluorine. The only exception is that some substances also break down to
yield trifluoroacetic acid (HTFA), a substance resistant to further degradation.
Recognizing this, industry embarked on a research and assessment program to
study the potential effects of trifluoroacetate (TFA) on the environment and
to investigate possible degradation pathways. The results of these recently com-
pleted studies are summarized below and described in further detail in this pa-
per.
Trifluoroacetic acid is a strong organic acid with a pKa of 0.23. It is miscible
with water and its low octanol/water partition coefficient (log P
ow
= –2.1) indi-
cates no potential to bioaccumulate.
Industrial use is limited and environmental releases are very low. Some ad-
ditional TFA will be formed from the breakdown of a few halogenated hydro-
carbons, most notably HFC-134a (CF
3
CH
2
F), HCFC-124 (CF
3
CHFCl), and
HCFC-123 (CF
3
CHCl
2
). As these substances have only been produced in limit-
ed commercial quantities, their contribution to environmental levels has been
minimal. Surprisingly, environmental measurements in many of diverse loca-
tions show existing levels of 100 to 300 ng·l
–1
in water with one site (Dead Sea)
having a level of 6400 ng·l
–1
. These levels cannot be accounted for based on
current atmospheric sources and imply a long-term, possibly pre-industrial
source. Generally, soil retention of TFA is poor although soils with high levels
of organic matter have been shown to have a greater affinity for TFA when con-
trasted to soils with low levels of organic matter. This appears to be an adsorp-
tion phenomenon, not irreversible binding. Therefore, TFA will not be
retained in soil, but will ultimately enter the aqueous compartment.
Modeling of emission rates and subsequent conversion rates for precursors
has led to estimates of maximum levels of TFA in rain water in the region of
0.1
µ
g·l
–1
in the year 2020.
TFA is resistant to both oxidative and reductive degradation. While there
had been speculation regarding the possibility of TFA being degraded into
monofluoroacetic acid (MFA), the rate of breakdown of MFA is so much high-
er than for TFA that any MFA formed would rapidly degrade. Therefore, there
would be no buildup of MFA regardless of the levels of TFA present in the en-
vironment.
Although highly resistant to microbial degradation, there have been two re-
ports of TFA degradation under anaerobic conditions. In the first study, natu-
ral sediments reduced TFA. However, even though this work was done in
replicate, the investigators and others were unable to reproduce it in subse-
quent studies. In the second study, radiolabeled TFA was removed from a
mixed anaerobic
in vitro
microcosm. Limited evidence of decarboxylation has
also been reported for two strains of bacteria grown under highly specific con-
ditions. TFA was not biodegraded in a semi-continuous activated sludge test
even with prolonged incubation (up to 84 days).
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61
TFA does not accumulate significantly in lower aquatic life forms such as
bacteria, small invertebrates, oligochaete worms and some aquatic plants in-
cluding
Lemna gibba
(duckweed). Some bioaccumulation was observed in ter-
restrial higher plants, such as sunflower and wheat. This result appeared to be
related to uptake with water and then concentration due to transpiration water
loss. When transferred to clean hydroponic media, some elimination of TFA
was seen. Also, more than 80% of the TFA in leaves was found to be water ex-
tractable, suggesting that no significant metabolism of TFA had occurred.
At an exposure level of 1200 mg·l
–1
of sodium trifluoroacetate (NaTFA) —
corresponding to 1000 mg·l
–1
HTFA — no effects were seen on either
Brachy-
danio rerio
(a
fish) or
Daphnia magna
(a water flea). With duckweed, mild effects
were seen on frond increase and weight increase at the same exposure level.
At a concentration of 300 mg·l
–1
no effects were observed. Toxicity tests were
conducted with 11 species of algae. For ten of these species the EC
50
was great-
er than 100 mg·l
–1
. In
Selenastrum capricornutum
the no-effect level was
0.12 mg·l
–1
. At higher levels the effect was reversible. The reason for the
unique sensitivity of this strain is unknown, but a recovery of the growth rate
was seen when citric acid was added. This could imply a competitive inhibition
of the citric acid cycle.
The effect of TFA on seed germination and plant growth has been evaluated
with a wide variety of plants. Application of NaTFA at 1000 mg·l
–1
to seeds of
sunflower, cabbage, lettuce, tomato, mung bean, soy bean, wheat, corn, oats
and rice did not affect germination. Foliar application of a solution of 100
mg·l
–1
of NaTFA to field grown plants did not affect growth of sunflower, soya,
wheat, maize, oilseed rape, rice and plantain. When plantain, wheat (varieties
Katepwa and Hanno) and soya were grown in hydroponic systems containing
NaTFA, no effects were seen on plantain at 32 mg·l
–1
, on wheat (Katepwa) and
soya at 1 mg·l
–1
, or on wheat (Hanno) at 10 mg·l
–1
; some effects on growth were
seen at, respectively, 100 mg·l
–1
, 5 mg·l
–1
, 5 mg·l
–1
, and 10 mg·l
–1
and above.
TFA is not metabolized in mammalian systems to any great extent. It is the
major final metabolite of halothane, HCFC-123 and HCFC-124. The half-life
of TFA in humans is 16 hours. As expected, the acute oral toxicity of the free
acid is higher than the one of the sodium salt. The inhalation LC
50
(2 hour ex-
posure) for mice was 13.5 mg·l
–1
(2900 ppm) and for rats it was 10 mg·l
–1
(2140 ppm). Thus, TFA is considered to have low inhalation toxicity. The irri-
tation threshold for humans was 54 ppm.
As one would expect of a strong acid, it is a severe irritant to the skin and
eye. When conjugated with protein, it has been shown to elicit an immunolog-
ical reaction; however, it is unlikely that TFA itself would elicit a sensitization
response. Repeat administration of aqueous solutions have shown that TFA
can cause increased liver weight and induction of peroxisomes. Relative to the
doses (0.5% in diet or 150 mg·kg
–1
·day
–1
by gavage) the effects are mild.
In a series of Ames assays, TFA was reported to be non-mutagenic. Its carci-
nogenic potential has not been evaluated. Although TFA was shown to accu-
mulate in amniotic fluid following exposure of pregnant animals to high levels
of halothane (1200 ppm), no fetal effects were seen. Likewise, a reproduction
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62
study that involved exposure of animals to halothane at levels up to 4000 ppm
for 4 hours per day, 7 days per week, resulted in no adverse effects. Given the
high levels of halothane exposure, it is unlikely that environmental TFA is a re-
productive or developmental hazard.
Overall the toxicity of TFA has been evaluated in stream mesocosms, algae,
higher plants, fish, animals and humans. It has been found to be of very low
toxicity in all of these systems. The lowest threshold for any effects was the re-
versible effect on growth of one strain of algae,
Selenastrum capricornutum
,
which was seen at 0.12 mg·l
–1
. There is a 1000-fold difference between the no-
effect concentration and the projected environmental levels of TFA from
HFCs and HCFCs (0.0001 mg·l
–1
). Based on available data, one can conclude
that environmental levels of TFA resulting from the breakdown of alternative
fluorocarbons do not pose a threat to the environment.
Key Words:
hydrochlorofluorocarbons (HCFCs), hydrofluorocarbons (HFCs),
CFC alternatives, trifluoroacetate (TFA), environmental fate, eco-
toxicity.
INTRODUCTION AND PERSPECTIVE
The Montreal Protocol was developed in 1987 in response to concerns that
the chlorofluorocarbons (CFCs) were releasing chlorine into the stratosphere
and that this chlorine was causing a depletion of stratospheric ozone over Ant-
arctica. This international agreement called for a phase out of these CFCs. In-
dustry initiated a major effort to find safe replacements. The Alternative
Fluorocarbons Environmental Acceptability Study (AFEAS) was formed in
1989 by several chemical companies from Europe, Japan, and the United
States with the aim of assessing the environmental acceptability of a range of
hydrochlorofluorocarbons (HCFCs) and hydrofluorocarbons (HFCs). Studies
by independent researchers over the next six years were coordinated with work
supported by U.S. and European government agencies. The results showed
that, unlike CFCs, alternative fluorocarbons will break down readily in the low-
er atmosphere to form simple inorganic species already present in the environ-
ment (NASA, NOAA and AFEAS, 1995 and references cited therein).
However, a few of the HCFCs and HFCs can be expected to form trifluoro-
acetyl halides that will dissolve in environmental water to give trifluoroacetate
(TFA) salts.
In 1991, in response to concerns that this would introduce a new and poten-
tially hazardous material into the environment, AFEAS initiated a research
program to determine the environmental fate of TFA and to provide a timely
and accurate forecast of potential ecological impact. The initial concern had
been that HFC-134a (1,1,1,2-tetrafluoroethane, an alternative to CFC-12)
would form TFA, which has relatively low mammalian toxicity, and that the
TFA would subsequently react in the environment to give monofluoroacetic
acid (MFA), which is acutely poisonous at low doses. Theoretical and experi-
mental data (Emptage, 1994) quickly removed these concerns about MFA ac-
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63
cumulation in the environment, but there was still a clear need to examine the
ecological impact of TFA.
A panel of advisors was convened in 1992 to outline research that would pro-
vide a valid and systematic approach to assessing the impact of TFA. The ex-
perts suggested a strategy to identify key “pressure” points in the major
biogeochemical cycles — carbon fixation, mineralization, nitrogen fixation
and bioaccumulation in plants — where TFA could conceivably exert a major
influence. Furthermore, plans were devised for assessing whether TFA can be
biodegraded in the environment. This approach was then used to guide an
evolutionary research program that was executed over the last 4 years through
collaborative efforts between academic and industrial laboratories. The mini-
mum levels of quality control were compliance with “Good Laboratory Prac-
tice” guidelines or publication of the results in peer-reviewed journals.
A number of physico-chemical and biodegradation studies have been com-
pleted, in addition to biological investigations. During the course of the work
it became clear that TFA is not a new environmental contaminant. It is present
in contemporary air, precipitation and surface waters from around the world,
and before significant amounts of HFC or HCFC precursors have been pro-
duced and released.
The program described here represents one of the most comprehensive ef-
forts to assess the environmental impact of a chemical ever undertaken by in-
dustry. The risk assessment is a synthesis of the research results and an
assessment based on best available data. It consists of a comparison of the an-
ticipated future levels of TFA, arising from decomposition of HFCs and
HCFCs, with concentrations that would exert environmental effects. More spe-
cifically, the risk assessment covers: general substance information, sources,
distribution and releases into the environment, exposure measurements, deg-
radation, accumulation, effects assessment, and risk characterization.
In this article, the following abbreviations will be used:
TFA: trifluoroacetate
HTFA: trifluoroacetic acid
NaTFA: sodium trifluoroacetate
GENERAL SUBSTANCE INFORMATION
Identification of the Substance
Name: Trifluoroacetic acid
CAS number: 76–05–1
Synonym: HTFA
Formula: CF
3
CO
2
H
Molecular weight: 114
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64
Name: Sodium trifluoroacetate
CAS number: 2923–18–4
Synonym: NaTFA
Formula: CF
3
CO
2
Na
Molecular weight: 136
Physico-Chemical Properties
HTFA
Physical state: Colorless fuming (hygroscopic) liquid
Specific gravity: 1484 kg·m
–3
at 25°C
Melting point: –15.3°C
Boiling point: 72°C
Vapor pressure: 105.7 hPa at 20°C
Water solubility: Miscible in all proportions
(Feenstra-Bieders and Olthof, 1992)
Partition coefficient
n-Octanol/water (log):
–2.1 (Feenstra-Bieders and Olthof, 1992;
Thus, 1997)
–0.2 (calculated according to Rekker)
0.325 (ClogP for Windows V 1.0.0)
0.5 (SRC’s KOWWIN v1.52)
pKa: 0.23
pH (1% solution): 1
pH (100 mg·l
–1
distilled water): 3.1
pH (10 mg·l
–1
distilled water): 7.4 – 7.75
pH (100 mg·l
–1
buffered
ISO water): 4
pH (10 mg·l
–1
buffered
ISO water): 7.4 (Thus and van Dijk, 1996)
Henry’s Law Constant K
H
′
= 1.1
×
10
–2
Pa·m
–3
·mol
–1
at 25°C
(Bowden
et al.,
1996)
UV/vis absorption: No absorption at
λ
> 250 nm
(van Dijk, 1992a)
Conversion Factors
(air, 1 atm, 25°C): 1 mg·l
–1
= 214 ppm; 1 ppm = 4.66 mg·m
–3
NaTFA
Physical state: White (hygroscopic) powder
Melting point: 207°C (decomposition)
Water solubility: 625 g·l
–1
(25°C)
pH (2% solution): 7
Partition coefficient
n-Octanol/water (log): –3.31 (SRC’s KOWWIN v1.52)
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Hum. Ecol. Risk Assess. Vol. 5, No. 1, 1999
65
SOURCES
Production and Known Sources
According to Elliott (1994), trifluoroacetic acid is manufactured by only
three companies: Halocarbon Products Corporation in the United States, and
Rhône-Poulenc and Solvay in Europe. The quantities involved are relatively
small — of the order of 1000 metric tons per year (J. Franklin, personal com-
munication), an estimate that is consistent with the quoted revenue of the ma-
jor U.S. producer (McConville, 1996). The processes are enclosed, with
effluent being subject to local requirements for disposal of chemical waste.
Similar chemical waste streams can also arise from processes in which trifluo-
roacetate is produced as a byproduct. From one such plant (DuPont – Fay-
etteville, North Carolina), the site effluent contained about 3 ppm of
trifluoroacetate (Chumley, 1992).
Environmental trifluoroacetate can also be produced during the oxidation
of several organofluorine compounds released to the atmosphere by human
activities. Halothane and isoflurane anaesthetics, which have been in use since
the 1970s, yield trifluoroacetate (R. Atkinson, personal communication). Also,
some of the fluorocarbon alternatives to CFCs decompose in the atmosphere
to form trifluoroacetate (NASA, NOAA, and AFEAS, 1995) and, while these
sources are currently small, they could become significant in the future and
merit the closer scrutiny given below.
Uses
Trifluoroacetic acid is widely used in the fine chemicals industry and as a
laboratory reagent: for derivatizing carbohydrates, amino acids and peptides;
as a catalyst in esterification reactions and the Beckmann rearrangement of
oximes to amides, and for protein synthesis (Elliott, 1994). In all cases, the ma-
terial is either consumed or becomes part of a chemical waste stream. The
quantities released into the atmosphere are very small indeed.
Emission Pattern
The known sources can be expected to yield fluxes of trifluoroacetate into
the environment that differ significantly in both geographical distribution and
the first compartment into which they are released. For example, only if there
are fugitive emissions of vapor will material that is used as a chemical interme-
diate give rise to a point source of atmospheric contamination, otherwise re-
leases can be expected to be to the aqueous compartment.
The Henry’s Law constant (K
H
′
) for trifluoroacetic acid estimated from mea-
surements by Bowden
et al.
(1996) is 8.95 ± 0.1
×
10
–3
mol·kg
–1
·atm
–1
, which
translates into a value for K
H
of 1.1
×
10
–2
Pa·m
–3
·mol
–1
. With such a value, and
because trifluoroacetic acid is totally miscible with water and has a log K
OW
val-
ue of –0.2, the preferred environmental compartment will be water, rather
than air, ground, or biota. The rules set out by Ballschmiter (1992) state that,
if K
H
is less than 2 Pa·m
–3
·mol
–1
and log K
OW
is less than 4, the material should
preferentially distribute into the aqueous compartment of the environment.
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Thus, trifluoroacetic acid that is emitted from processes as vapor, or trifluoro-
acetate that is formed from other materials by reaction in the atmosphere will
partition rapidly into cloud, rain, or surface water.
Substances that decompose in the atmosphere by oxidation or photolysis
will give rise to diffuse fluxes of decomposition products into the atmosphere
as the first compartment. These substances have atmospheric lifetimes ranging
from about 1 to 40 years (WMO, 1994) so that most decomposition occurs at
or near the background concentration of the substance after it has become
well mixed in the atmosphere. The decomposition mechanisms depend on
physical and chemical properties and processes that are relatively well under-
stood (NASA, NOAA, and AFEAS, 1995) so that it is possible to calculate accu-
rately the future atmospheric concentrations of precursors to trifluoroacetate,
given a particular set of scenarios for emissions. Furthermore, the flux of tri-
fluoroacetic acid, into the atmosphere and thence into rain and surface water,
can be estimated from the precursor concentrations with confidence and pro-
vides the possibility of verifying the magnitude of known fluxes against envi-
ronmental observations.
Concentrations of trifluoroacetate observed in the air, in precipitation and
in surface waters in Europe are several orders of magnitude larger than the
known source fluxes would allow (Frank
et al.,
1996). Similar concentrations
have been observed in contemporary water and air samples from Nevada, USA;
Canada; Australia and South Africa (Zehavi and Seiber, 1996; Frank and Klein,
1997; Grimvall
et al.,
1997), suggesting that there is one or more large un-
known source of environmental trifluoroacetate. The sizes of the fluxes of the
known source gases are not in doubt; their concentrations calculated from es-
timated emissions are consistent with observations, and additional contempo-
rary sources, substantially larger than those known, must be invoked to explain
the observed environmental concentrations of trifluoroacetate.
The current environmental burden from the decomposition of man-made
fluorocarbons, and the anticipated future burden resulting from future emis-
sions of such precursors, should be evaluated in the context of the background
concentrations that have been observed.
DISTRIBUTION AND RELEASES INTO ENVIRONMENT
From Trifluoroacetic Acid Production and Use
With a pKa of 0.23, the strength of trifluoroacetic acid approaches that of
mineral acids. Unlike the mineral acids, it is miscible not just with water but
with fluorocarbons and most common organic solvents including methanol,
benzene, carbon tetrachloride, acetone, ether, and hexane. It is a good solvent
for proteins, leading to its use in protein synthesis. It is useful for making de-
rivatives of carbohydrates, amino acids and peptides from which the trifluoro-
acetyl protective group can be removed relatively easily (Elliott, 1994).
Trifluoroacetic acid is a useful catalyst for esterifications of alcohols, acylations
of aromatics (Halocarbon, 1967) and in the Beckmann rearrangement of
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Hum. Ecol. Risk Assess. Vol. 5, No. 1, 1999
67
oximes to amides (Huber, 1955). These uses are typical of the fine chemicals
industry and of laboratories, and the total global production of trifluoroacetic
acid is expected to reflect the small scale of the uses. Although there are no
audited data for production and use, informal estimates suggest that the quan-
tities involved are of the order of 1000 metric tons per year.
Trifluoroacetic acid is stable to oxidation and so can be prepared by a num-
ber of routes that involve the oxidation of compounds that have α-trifluorom-
ethyl groups. Thus, alkaline permanganate oxidation of 2,3-dichloro-
hexafluorobut-2-ene gives an 87% yield of the acid (Henne and Trott, 1947)
and, in a manner similar to atmospheric degradation processes, photochemi-
cal oxidation of 1-chloro-2,2,2-trifluoroethane or 1,1-dichloro-2,2,2-trifluoro-
ethane with oxygen gives high yields of trifluoroacetyl chloride that can be
hydrolyzed to trifluoroacetic acid (Haszeldine and Nyman, 1959; Dittmann,
1975).
Like the use processes, the production processes involve reactions in con-
tained and/or aqueous systems from which fugitive losses of trifluoroacetic
acid to the atmosphere are small. In many cases the use is as a raw material so
that most of it is wholly converted and cannot be released as trifluoroacetic
acid or its salts. From both production and use, any persistent loss is likely to
be in the form of solution in chemical waste streams, the disposal and treat-
ment of which will be subject to local controls.
Deliberate production and use of trifluoroacetic acid does not appear to be
a significant contributor to current global environmental levels.
By-Product from Chemical Syntheses
Production of hexafluoropropylene oxide can result in an aqueous waste
stream containing trifluoroacetate. The concentration of trifluoroacetate mea-
sured in one plant outfall was 3 ppm or less (Chumley, 1992) and, while this
may or may not be typical of such processes, the compartment and the concen-
tration are consistent with the properties discussed in a previous section.
Atmospheric Oxidation of Fluorinated Hydrocarbons
In recent years the atmospheric decomposition of halocarbons, particularly
those fluorocarbons that could replace CFCs, has been the subject of intensive
study. It is now agreed that the mechanism involves an initial abstraction of the
hydrogen atom in the molecule by atmospheric hydroxyl radicals to yield water
and a haloalkyl radical. The latter reacts rapidly with oxygen to give a haloalkyl-
peroxy radical that undergoes further reactions, generally with nitric oxide or
hydroperoxyradicals. The product in both cases is a haloalkoxy radical, which
can react further in one of three ways, depending on its molecular composi-
tion: C-Cl bond cleavage, C-C bond cleavage or hydrogen abstraction (Cox et
al., 1995; McCulloch and Sidebottom, 1993; Midgley, 1995). Thus, the general
reactions of those haloalkoxy radicals that could form precursors to trifluoro-
acetate are:
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68
CF
3
CYlO· → CF
3
CYO + Cl· (1)
CF
3
CYZO· → ·CF
3
+ CYZO (2)
CF
3
CYHO· + O
2
→ CF
3
CYO + HO
3
· (3)
where Y and Z can be Br, Cl, F, or H.
The ultimate products of many of these reactions are fluoride and chloride
ions and carbon dioxide. These are neither new to the environment nor ex-
pected to be introduced at rates that would significantly enhance the back-
ground levels (WMO, 1989). However, if Y is either Br, Cl, CF
3
, or F then the
products of reactions (1) and (3) are trifluoroacetyl halides that can, on hy-
drolysis with atmospheric or surface water, form trifluoroacetic acid. Thus, tri-
fluoroacetic acid is by no means a universal product of the degradation of
fluorinated hydrocarbons; the parent compound requires a CF
3
group and a
halogen atom on the α carbon adjacent to it. Substances that fulfill these cri-
teria are:
CF
3
CHClBr 1-bromo-1-chloro-2,2,2-trifluoroethane ‘Halothane’
CF
3
CHClOCHF
2
1-chloro-2,2,2-trifluoroethyl ‘Isoflurane’
difluoromethyl ether
CF
3
CHCl
2
1,1-dichloro-2,2,2-trifluoroethane HCFC-123
CF
3
CHFCl 1-chloro-1,2,2,2-tetrafluoroethane HCFC-124
CF
3
CH
2
F 1,1,1,2-tetrafluoroethane HFC-134a
CF
3
CHFCF
3
1,1,1,2,3,3,3-heptafluoropropane HFC-227ea
HFC-125 (1,1,1,2,2-pentafluoroethane) matches the criteria but undergoes
solely C-C bond cleavage according to reaction (2); reaction (1) would require
the cleavage of a C-F bond which is not possible under atmospheric conditions.
Halothane and isoflurane anaesthetics are expected to form totally trifluoro-
acetyl chloride or bromide (probably the former) (R. Atkinson, personal com-
munication). HCFC-123 is converted 98% to trifluoroacetyl chloride, HCFC-
124 100% to trifluoroacetyl fluoride and HFC-227ea yields equal amounts of
carbonyl difluoride and trifluoroacetyl fluoride (Zellner et al., 1994). HFC-
134a, however, can undergo both reactions (2) and (3) so that it forms a mix-
ture of trifluoroacetyl fluoride, formyl fluoride and carbonyl difluoride (Tua-
zon and Atkinson, 1995; Cox et al., 1995). The extent to which each reaction
will proceed in the atmosphere depends on the local temperature where de-
composition is occurring and the local pressure, particularly partial pressure
of oxygen for reaction (3). At sea-level, only some 18% of the HFC-134a de-
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69
composing is converted to trifluoroacetyl fluoride but, with increasing altitude
(and consequently reducing temperature and pressure), the yield of trifluoro-
acetyl fluoride increases and it becomes the major product at the tropopause.
An average yield for the troposphere is calculated to be in the region of 40%
(Cox et al., 1995; Franklin, 1993; Midgley, 1995). Recently, it has been suggest-
ed that these conversions to trifluoroacetyl fluoride might be overestimates
and that the true value could be as little as 12% or one third of the value now
accepted. This view is based on work done in a laboratory system that con-
tained NO (Wallington et al., 1996), which is thought to enhance the scission
reaction via a relatively stable CF
3
CFHOONO intermediate. If this is truly rep-
resentative of the atmospheric behavior of HFC-134a, it would reduce the ex-
pected flux of TFA from this source by a factor of about three. A yield of 40%
is, however, used throughout the calculations of source strengths in this assess-
ment on the grounds that this is the value used in most of the atmospheric
modeling studies and that it represents a maximum from the HFC-134a
source.
Trifluoracetyl fluoride is stable towards photolysis in the troposphere but
trifluoroacetyl chloride (and bromide) will photolyze there, reducing the ap-
parent yields of these acid halides from the parent compounds to about 60%
of stoichiometry in the case of the acid chloride (Cox et al., 1995).
The acid halides will be taken up by atmospheric or surface water relatively
rapidly; the rates are governed by the rate constants of the hydrolysis reactions
forming trifluoroacetic acid and their Henry’s constants for aqueous solution.
In all cases, the maximum lifetime for the acid halide would be of the order of
30 days and the minimum lifetime, which is governed more by cloud formation
probabilities, is about 10 days (Kolb et al., 1995a,b).
These lifetimes are long compared with regional mixing times so that, hav-
ing been formed in the atmosphere, the trifluoroacetyl halide would be dis-
persed on a continental scale before being deposited as trifluoroacetic acid.
Nevertheless, the lifetimes are short compared to those of the parent com-
pounds which are governed by the rates of reaction with hydroxyl radicals. The
atmospheric lifetimes of the parent compounds range from about 1 year, in
the case of the anaesthetics, halothane and isoflurane (Brown et al., 1990),
through 1.4 years for HCFC-123 (WMO, 1994), 6.1 years for HCFC-124, and
14.6 years for HFC-134a (IPCC, 1996a). In all cases, such lifetimes mean that
the compounds will be hemispherically, if not globally, dispersed and so they
will react at concentrations that are relatively constant geographically. Local
variations in hydroxyl radical concentrations are liable to have more profound
effects on the rates and extent of reactions (Madronich and Dentener, 1995;
Kanakidou et al., 1995).
Current Releases
The atmospheric concentrations of the HCFCs and HFC-134a are deter-
mined. On the contrary, there are no concentration measurements and only
informal estimates for releases of the anaesthetics (halothane and isoflurane)
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70
to the atmosphere. It has been assumed that halothane remains the most used
of the fluorocarbon anaesthetics at a maximum of 1500 metric tons per year
globally, and that all the material used is emitted to atmosphere (McCulloch,
1995b). The figure is consistent with the productive capacity and market share
of the major manufacturer (C&P News, 1995 and 1996). The same references
suggest that production has been at or near this level for several years so that
the fluxes of these anaesthetics and their atmospheric decomposition can be
assumed to be in equilibrium. If the use of isoflurane were half as much as that
of halothane, the total quantity of trifluoroacetyl chloride formed by the an-
aesthetics would be in the region of 1540 metric tons per year; given that this
material photolyzes, the effective flux would be 930 metric tons per year, equiv-
alent to a global deposition rate of 800 metric tons per year of trifluoroacetic
acid.
The potential HCFC precursors to trifluoroacetic acid are at the point of
their limits of detection in the atmosphere (P.G. Simmonds, personal commu-
nication), inferring that their concentrations are, at most, 0.1 parts per trillion
by volume (pptv). The upper limits for the atmospheric burdens of these com-
pounds are 2500 metric tons of HCFC-123 and 2300 metric tons of HCFC-124,
consistent with the values given in McCulloch (1995b) for lower limits of de-
tection. The value for HCFC-123, coupled with its atmospheric lifetime of 1.5
years, equates to a decomposition rate of 1690 metric tons per year which, after
allowing for the photolysis of the trifluoroacetyl chloride product, gives a dep-
osition rate for trifluoroacetic acid of 760 metric tons per year. Similarly, the
deposition rate for trifluoroacetic acid from HCFC-124 is calculated to be 320
metric tons per year.
Contemporary concentrations of HFC-134a in the atmosphere are 2.5 pptv
in the Northern Hemisphere and 1.2 pptv in the Southern Hemisphere
(Montzka et al., 1996; Oram et al., 1996). This corresponds to an atmospheric
burden of approximately 31,000 metric tons. Using the atmospheric lifetime
of 14.6 years, the decomposition rate is calculated to be 2140 metric tons per
year. At a 40% yield, this would amount to 960 metric tons per year of trifluo-
roacetic acid deposited globally from this source.
HFC-227ea has yet to be detected in the atmosphere and so the potential
contribution from it has been assumed to be zero now. Therefore, the maxi-
mum estimate for the total contemporary deposition rate of trifluoroacetate
from fluorinated hydrocarbons is 800 + 760 + 320 + 960, or a total of 2800 met-
ric tons per year.
Scenarios for Future Releases of Precursors
Alternative Scenarios
With the exception of the anaesthetics, demand for which is static or even
falling, the fluorocarbon precursors to trifluoroacetic acid are expected to be
released in progressively larger amounts throughout the coming decades. Pro-
duction and use of HCFCs is controlled under the Montreal Protocol, so that
the emissions of HCFC-123 and HCFC-124 will become limited within the next
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71
25 years and will eventually fall to zero. There is no specific regulatory limit on
HFCs and so, theoretically, growth of HFC-134a emissions could continue un-
abated. In order to judge the potential future effects of these materials on the
environment, various scenarios have been proposed, based on several method-
ologies.
1. For the Enquete Commission of the German Parliament (DB, 1994) on
life cycle analysis applied to HFC-134a and other substitute refrigerants,
emissions of HFC-134a were calculated to reach 160 kT·yr
–1
by 2020, with
production in that year running at 310 kT·yr
–1
. The calculations were
based on projection of underlying demand, using projected regional
GDPs, and views on the extent of substitution and integrity of contain-
ment.
2. The demand for alternative fluorocarbons over the period to 2020 was es-
timated by McCulloch (1994, 1995a) from a detailed analysis of the his-
torical market for CFCs and estimates for their replacement by HCFCs
and HFCs. The values for HFC-134a production and emissions in 2020
were almost identical to those in DB (1994). No data for HCFCs 123 and
124 were provided by this work.
3. In order to gauge the potential climate change effect from future emis-
sions of fluorocarbons, the Intergovernmental Panel on Climate Change
(IPCC) has provided scenarios for HCFCs and HFCs. These were devel-
oped by projecting CFC demand and proposing almost total substitution
of that demand using HCFCs and HFCs. The scenarios, which cover all
greenhouse gases, were originally proposed in 1992 to replace the 1990
scenarios that were considered unacceptable by the parties to the Rio
Convention. They have been updated in the light of changing regula-
tions and the latest version is contained in the “Second Assessment Re-
port” (IPCC, 1996b). Only for HCFC-123 is the emission specified in
mass units; it is proposed that it reaches a maximum of 130 kT·yr
–1
at the
end of this century, falling to 13 kT·yr
–1
by 2020. Emissions of HCFC-124
are not mentioned; presumably they are expected to be very low. For
HFC-134a, the scenario output is in terms of atmospheric concentrations
and values of 200–230 pptv are proposed for the year 2020, the variability
arising from different assumptions about overall economic growth. The
equivalent emissions can be back-calculated from these concentrations
and, for HFC-134a, amount to 400–470 kT·yr
-1
. HFC-227ea is not includ-
ed specifically in this scenario.
The much larger values proposed in the IPCC scenarios reflect the under-
lying assumption that the materials are used to substitute in all historic markets
for CFCs, including immediately dispersive uses such as aerosol propellants.
They may be unrealistically high. For example, the IPCC scenarios anticipated
that the current atmospheric concentration of HFC-134a would be 13 to
14 pptv. The actual measurements, as described above, show a global average
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72
of about 2 pptv (Montzka et al., 1996; Oram et al., 1996). For reasons such as
this the IPCC scenarios are not used in this assessment.
Development of a Standard Concentration of TFA
The possible future deposition rate of trifluoroacetic acid has been calculat-
ed using mathematical models of the lower atmosphere. In addition to the
chemical reactions themselves, these models take account of the geographical
distributions of:
• The concentration of hydroxyl radicals with which the parent com-
pounds react first of all
• The emissions of the parent compounds
• The rainfall that scavenges trifluoroacetic acid from the atmosphere
(Kanakidou et al., 1995; Rodriguez et al., 1994).
Because most of the rainfall and the highest regional concentrations of hy-
droxyl occur in the tropics, the mass deposition rate of trifluoroacetic acid is
highest in this region. Starting from the scenario described in McCulloch
(1994), Kanakidou et al. (1995) calculate that the tropical concentration would
be up to 2 mol·m
–2
·yr
–1
and that this would be equivalent to 1 nmol·l
–1
of rain-
water. The latter value would correspond to 100 ng·l
–1
. Over Europe the dep-
osition rate in 2020 would be about 0.5 µmol·m
–2
·yr
–1
and, over the United
States, levels would be up to double this.
These values are of the same order of magnitude as those calculated by Ro-
driguez et al. (1994), on the basis of a scenario from the U.S. Environmental
Protection Agency; the latitudinal variation (up to a factor of 40) and longitu-
dinal variation (a factor of 3) are also similar. In the year 2020, the modeled
decomposition flux of HFC-134a was calculated to be 115 kT·yr
–1
, with a fur-
ther 20 kT·yr
–1
each of HCFCs 123 and 124. The implied global deposition of
trifluoroacetic acid then would be 155 kT·yr
–1
.
With such close agreement on potential future deposition rates, this quan-
tity (rounded to 160,000 metric tons per year), and a corresponding rainwater
concentration of 0.1 µg·l
–1
(0.0001 mg·l
–1
) are adopted as standards for this
risk characterization.
Accumulation of TFA in Aquatic Ecosystems
The objective of this section is to review whether conditions could occur in
the environment which could lead to a significant accumulation of TFA in nat-
ural water bodies. It appears to be possible to accumulate a riverborne flux of
TFA from “Normal” concentrations of a few hundreds of nanograms per liter
to several thousands of nanograms per liter in receiving waters from which
there is no outflow, only evaporation. In the case of the River Jordan/Dead Sea
system (Table 3), consideration of the volumes and flows indicates accumula-
tion over many hundreds of years.
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Recently, the assumption has been made by Tromp et al. (1995) that TFA
can accumulate to very high concentrations in seasonal wetlands, which are
special aquatic ecosystems that dry out periodically and are replenished only
by rainfall. It is necessary to examine if the hypothesis developed in that work
could occur in real conditions.
Tromp et al. (1995) conducted a sensitivity study of the potential for enhance-
ment of TFA concentrations in such ecosystems. In the study, several assump-
tions were used concerning both the conditions that might lead to higher
concentrations of TFA in rainwater (in comparison with expected average con-
centrations calculated by mathematical models; see previous sections) and the
characteristics of the seasonal wetlands that would favor accumulation. A calcu-
lation presented in the paper shows that a seasonal wetland, receiving rainfall for
a total period of 30 years at a local (enhanced) TFA concentration of 1 mg·l
–1
,
with an evaporative loss rate of 5 yr
–1
, and a loss of TFA out of the wetland by
seepage (or other physical or biological process) at a rate of 0.1 yr
–1
, would reach
a concentration of 100 mg·l
–1
at the end of that period. However, this is simply a
mathematical calculation that has been conducted by adding together the con-
sequences of extreme conditions without regard for the probability that they
could occur simultaneously. In the real environment it is highly improbable that
all of the necessary conditions for accumulation will happen together and be
maintained for decades. The step-by-step review of those conditions carried out
below indicates that the probability that such accumulation will take place over
decades is close to zero.
First, the seasonal wetland would need to be located near a large urban area,
where the assumption is made that local conditions could lead to enhanced
TFA concentration in rainwater. These local conditions would have to meet
several criteria.
1. The concentrations of precursors to TFA should be higher than average,
due to local emissions. A factor of 10 to 20 is suggested in Tromp et al.
(1995), but observations in the Los Angeles area (notorious for local pol-
lution) show an historic factor of about 3 for CFC-11 (Bastable et al.,
1990). CFC-11 was used in the blowing of polyurethane foams and in
whole building air conditioning. The most significant TFA precursor is
expected to be HFC-134a, which is used in refrigeration and air condi-
tioning applications where smaller emissions factors are expected and
further improvements in containment are likely in the coming decades.
The appropriate enhancement factor therefore is less than 3.
2. The oxidizing capacity of the local atmosphere (including the OH radi-
cal concentrations) would be higher due to local pollution, thus increas-
ing the rate of conversion of the precursors. A possible factor of 10 is
quoted in Tromp et al. (1995), but calculations based on actual observa-
tions of tropospheric ozone, NOx, water vapor, carbon monoxide, meth-
ane and non-methane hydrocarbons at Riverside, California in June
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74
suggest that the OH field enhancement in that portion of the Los Ange-
les Basin does not exceed a factor of 2 (Ko et al., 1995).
3. It would be necessary to have an appropriate combination of pollution
event and rainfall, the most effective rainout of TFA occurring when rain
immediately followed a pollution event. The implicit assumption in
Tromp et al. (1995) is that only the most effective rainout occurs and that
it is repeated systematically throughout the year for decades. The real sit-
uation, for example, in the Los Angeles Basin, is that rainfall happens
preferentially in the winter months, when photochemical activity in the
atmosphere is lowest (Ko et al., 1995). The same authors pointed out that
no rainfall event was observed during June 1990 and, in that case, TFA
would be dispersed (on a continental scale) before being rained out. It is
expected that pollution events would last more than one day, allowing
precursor concentrations to increase; this would be possible under an in-
version layer that trapped the gases within the “boundary” layer. Obser-
vations show that such highly polluted conditions are associated with low
wind speed. On the other hand, rainout of the TFA formed within the
pollution event needs a change in meteorological conditions bringing in
a new air mass, with clouds and rain. For effective rainout of the TFA, this
change in meteorology must not affect the stability of the boundary layer
and must not disperse the TFA before rainout. It is not clear that such a
mechanism could exist in the real atmosphere and, even if it did, it would
be an exceptional event. An attempt to calculate the resulting increase of
the average TFA concentration from uncorrelated rain and pollution
events gives a value of the order 15% increase at most (Ko et al., 1995).
The probability of maintaining conditions that give total local rainout
over a period of decades is therefore not far from zero.
Secondly, the long-term accumulation in seasonal wetlands requires:
4. A closed system, maintained for decades, where seepage or loss by other
physical processes such as aeolian transport would be small. Vernal pools,
examples of seasonal wetlands suggested to be most at risk, have been
shown to have significant seepage (Sefchick et al., 1994).
5. An evaporation rate high enough to have a severalfold increase in TFA
concentration each year.
Tromp et al. (1995) chose a value of 5 yr
–1
for this parameter, but the largest
evapoconcentration factor for vernal pools is 2.2 as reported by Sefchick et al.
(1994) and inferred from chloride ion concentration that is indicated as the
best conservative solute with sodium ion. It is also important to point out that
such evapoconcentration factors reflect concentrations variations but not ac-
cumulation from year to year. Evapoconcentration factors vary with the type
and location of the wetland and with the ion concentrated and this observation
yields the most appropriate value.
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Each condition listed above, even taken alone, has little probability of oc-
curring. For accumulation to take place, not only do the conditions have to oc-
cur together but this combination of conditions has to be maintained for
decades. In conclusion, although some accumulation of TFA may take place in
aquatic ecosystems like vernal pools, accumulation to high concentrations as
described in Tromp et al. (1995) appears to be highly improbable.
Soil Adsorption
The degree to which TFA adsorbs to soils may affect its bioavailability to
plants and bacteria, the general mobility of this compound through soils and
aquifers, and, of course, the concentration to which TFA could accumulate in
the soil. The soil-water milieu may vary greatly in terms of mineral and organic
compositions, as will the flux of cations and anions through these matrices.
Therefore, if there is an interaction of TFA with specific components of the
soil, this interaction may be site-specific and non-uniform in occurrence.
Clearly, the problem of assessing soil adsorption in a controlled manner and
in a way that is useful for developing a global perspective is very difficult.
Two laboratory studies were conducted to directly assess the partitioning of
TFA between water and a variety of soil types under controlled laboratory con-
ditions. The first study by van Dijk (1992b) showed that TFA did not adsorb to
soil. The second study by Richey et al. (1997) showed that TFA generally inter-
acted only weakly with most soils but was strongly adsorbed by some soils which
contained high levels of organic matter. The data are not necessarily contra-
dictory but may reflect the heterogeneity of soils tested. The studies are dis-
cussed in more detail below.
The first study examined the interaction of sodium TFA with three different
soil types, obtained from locations in Northern Europe. The salient properties
of the soils from van Dijk (1992b) follow.
Sand Sandy Loam Loam
% Clay 2.7 8.3 16.5
% Silt 6.4 27.8 46.8
% Sand 90.9 64.1 36.7
pH 5.5 6.5 7.5
% Organic Carbon 4.3 1.3 2.6
These soils represent a relatively narrow range in terms of organic carbon and
clay content; pH varied from mildly acidic to slightly alkaline. The researchers
tested a soil/water mixture comprising 2 g soil and 10 ml of 10 µM calcium ac-
etate in mixture with 40 µM sodium TFA. The supernatants were assayed for
TFA after 16 hours of incubation. The results of the analyses indicated that
<3% of the added TFA had been retained by the soil. This was equivalent to
partition coefficients (K
d
, soil concn./aqueous concn.) of <0.2 l·kg
–1
.
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A second study was performed on 54 different soils from disperse environ-
ments located predominantly in North America and Europe (Richey et al.,
1997). Rather than list the 54 soils individually, the ranges of important param-
eters are summarized here.
pH 2.76 – 5.85
% Clay 1% – 27%
% Organic Matter 1.4% – 93.3%
In these experiments adsorption isotherms were determined for 1:5 or 1:20
soil:solution ratios (for high organic or high mineral soils, respectively) over a
range of TFA concentrations: 0, 2, 4, 7, 10, 20, 30, 40 µM NaTFA. TFA adsorp-
tion was determined by the difference in supernatant TFA concentrations after
an equilibration period of 24 hours. Of the 54 soils tested, 20 soils showed no
significant retention of TFA. Of the remainder, 23 gave K
d
values <2 (high to
intermediate mobility) and 8 soils exhibited low mobility of TFA (K
d
= 2 to 10).
Only three soils gave K
d
values >10; the maximum K
d
reported was 20 (60% re-
tention) for a peat core containing 93% organic matter. At all sites where TFA
retention was determined for both organic and mineral soils, greater retention
was observed for the organic soil.
TFA adsorption was compared to that of other ions: bromide, chloride, flu-
oride, nitrate, and sulfate. The results indicated that TFA was not the most
strongly retained of the common ions with the relative order for a high
(82.5%) organic soil as follows: fluoride (most strongly retained) >> sulfate >
chloride > TFA > bromide > nitrate. For intermediate (24.8%) organic matter
soil: fluoride = sulfate > chloride = TFA > bromide > nitrate. Anions were also
shown to competitively inhibit TFA adsorption. For example, 150 µM
sulfate — not an excessive environmental sulfate concentration — caused 50%
reduction in TFA adsorption.
In comparing the two studies it is clear that a much larger representation of
soil types was investigated in the work by Richey et al. (1997) and that the or-
ganic content of many of the soils in that work was strikingly higher. Organic
content was determined by Richey et al. to be in positive correlation with TFA
adsorption. The relatively low organic content of the soils in the van Dijk study
may have led to the accurate conclusion for that study that TFA did not adsorb.
One disturbing aspect of the van Dijk study was the use of 10 mM calcium ac-
etate in the TFA equilibration mixture, use of such high concentration of a
TFA analogue might have suppressed TFA binding. It should be noted that the
TFA concentration used in both studies was 3 to 4 orders of magnitude higher
than that anticipated in the environment in the year 2020.
Field manipulation experiments were conducted at two sites at the Hub-
bard Brook Experimental Forest (New Hampshire, USA). Field results gener-
ally agreed with laboratory studies, indicating that there is considerable
variability in soil retention of TFA. TFA adsorption occurred mainly in the sur-
face soil, where organic content was highest, and diminished with soil depth.
Additions of TFA to an upland forest soil were largely transported with drain-
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age water (>70%). In contrast, additions of TFA were strongly retained in a
wetland with <5% of the added TFA exiting the system in water. The remain-
der of the TFA in the wetland was found in soil and plant tissues (Likens et al.,
1997). As in the laboratory studies of Richey et al. (1997), TFA mobility in field
plots (Berger et al., 1997) was found to be comparable with that of bromide,
which is widely used as a conservative hydrologic tracer.
It can be concluded that typical soils do not exhibit significant retention of
TFA, but that some partitioning to organic-rich soils is observed. Although
none of the investigators examined the reversibility of TFA binding, based on
the maximum observed partition coefficient (K
d
) of 20 (Richey et al., 1997),
the maximum concentration bound to soil — in equilibrium with the project-
ed rainfall concentration from CFC alternatives of 0.0001 mg·l
–1
— would be
0.002 mg·kg
–1
.
EXPOSURE MEASUREMENTS
Aquatic Compartment
A number of groups have determined TFA in the aquatic environment at
concentrations ranging upwards of a few nanograms per liter. Air, rain, and
surface waters in Europe were extensively sampled and analyzed for trifluoro-
acetate during 1995 by Frank et al. (1996) and Frank and Klein (1997). This
built on previous work by the same authors. The analytical methods were stan-
dardized and measurements made in air and rainwater samples during 1993
and 1994 (Frank et al., 1995). The results of their determinations of the TFA
concentrations in the Roter Main river in 1995 are given in Table 1; Figure 1
shows the analyses for both 1995 and 1996. There is no discernible seasonal
trend.
Table 2 shows the results of TFA determinations in river and lake waters in
Europe and the approximate locations of the sample points. The locations of
European seawater, air, and precipitation sample points are shown in Figure 2.
Table 1. TFA content of the Roter Main
River at Bayreuth, Germany
(Frank et al., 1996).
Date of Sample TFA (ng·l
–1
)
March 1995 160 ± 27
April 1995 80
May 1995 140
June 1995 110
October 1995 280
November 1995 110
December 1995 60
Average 140
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Surface water in and around Reno, Nevada, showed far higher concentra-
tions of trifluoroacetate using a comparable, but different, analytical method
(Zehavi and Seiber, 1996). The highest concentration that was recorded ex-
ceeded that in the Dead Sea (Frank et al., 1996) by a factor of 6. In both cases,
if the samples of the receiving waters and the rivers feeding them are represen-
tative, accumulation over many hundreds of years would be necessary to ac-
count for the observations.
Table 3 shows the concentrations reported for surface waters in Israel, Rus-
sia, Brazil, South Africa, and the United States. Table 4 shows the seawater con-
centrations recorded by Frank and co-workers for samples from the Baltic and
North Seas and the Atlantic and Pacific Oceans. Figure 3 shows the sampling
locations for all of the non-European surface water determinations, together
with the locations of precipitation samples. Table 5 lists the results of determi-
nations of trifluoroacetate in contemporary and ancient spring, mineral, and
tap waters.
Figure 1. TFA in the river Roter Main at Bayreuth.
.
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Table 2. Present TFA concentration in lakes and rivers in Europe (Frank
et al., 1996; Frank and Klein. 1997)
TFA determination
Location Sample date ng.1
–1
n S.D.
a
R. Elbe (Wittenberg, 24 Feb 95 200 3 10
Germany)
R. Danube (Regensburg, 25 Mar 95 200 3 2
Germany)
R. Pegnitz (Nürnberg, 26 Mar 95 150 3 3
Germany)
R. Mistelbach (Bayreuth, 24 Mar 95 215 3 9
Germany)
R. Main (Mainleus, 3 Apr 95 40 3 16
Germany)
R. Neckar (Tübingen, 23 Apr 95 260 3 5
Germany)
R. Elbe (Hamburg-Altona, 7 Jul 95 100 5 14
Germany)
R. Weiser (Bremerhaven, 6 Jul 95 100 4 7
Germany)
R. Rhine (Duisburg-Ruhort, 6 Jul 95 630 4 15
Germany)
R. Warnow (Rostock, 9 Jul 95 60 4 25
Germany)
R. Nabb (Schwandorf, 9 Jul 95 130 3 14
Germany)
R. Seine (Paris, France) Jul 95 40
R. Loire (Nantes, France) Jul 95 330
R. Rhine (Bregenz, Austria) 14 Sep 95 55 3 9
R. Kemiojki (Tärytie, Finland) 15 Jul 96 210 5 20
Lake Fichtel (Fichtelberg, 17 May 95 70 3 25
Germany)
Lake Weiβenstadt 17 May 95 115 2 100
(Weiβenstadt, Germany)
Lake Constance (Constance, 14 Sep 95 60 3 8
Germany)
Lough Skannive (Ireland) 1 Nov 95 20 3 21
Lough Ahalia (Conemara, 1 Nov 95 <10 3 —
Ireland)
a
S.D. is the Standard Deviation of n analyses.
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While the whole of the world has not been covered by the sampling pro-
gram, it is apparent that trifluoroacetic acid is a ubiquitous contaminant of sur-
face waters over a large part of the earth’s surface. It is also present in air and
rain samples (see subsequent section) and in samples of old ice. The determi-
nations in enclosed lakes in Nevada and Israel that could indicate long-term
accumulation, over several hundred years, would point to a source that is not
just a present day industrial activity.
Figure 2. Approximate locations and concentrations of environment TFA
measured in Europe Air and precipitation samples: B, Bayreuth; G,
Gdansk; MG, Mårmo Glacier; and MH, Mace Head. Seawater
determination are shown in italic; rivers and lakes are in normal print.
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Figure 3. Approximate locations and concentrations of environment TFA other
than in Europe. Air and precipitation samples: D, Davis, California; MC,
Mount Cook, New Zealand; QML, Queen Maud Land, Antarctica; R,
Resolute, Canada; and Reno, Nevada. Seawater measurements are shown
in italic; rivers and lakes are in normal print.
Terrestrial Compartment
There are no data on background concentrations of TFA in soil or on soil
surfaces.
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Atmosphere
Over the same period of time that they were determining trifluoroacetate
concentrations in surface waters, Frank and co-workers conducted an extend-
ed campaign to sample and analyze air and precipitation in Bayreuth, Germa-
ny. The results are shown in Table 6 for 1995. Figures 4 and 5 show the air and
precipitation analyses for both 1995 and 1996.
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Values determined during the earlier campaign, in 1993 and 1994 when the
method was being developed, were of a similar order, ranging from 18 to 620
pg·m
–3
for air (with one outlier at 3230 pg·m
–3
) and <3 to 120 pg·g
–1
(ng·l
–1
)
for rainwater (Frank et al., 1995). The determinations are consistent with the
levels observed in surface waters (see previous section).
The observation that concentrations in air under the tree canopy in a forest
are consistently greater than concentrations in a clearing (reproduced in Fig-
ure 6) was interpreted by Frank and Klein (1997) to indicate that the forest
canopy was involved in the deposition mechanism.
Analyses have also been performed on air and precipitation samples from
the western USA (Zehavi and Seiber, 1996) and on samples of precipitation
from a wide range of remote locations in the Northern and Southern Hemi-
spheres (Grimvall et al., 1997). The data are recorded in Table 7.
Similarly to their surface water determinations, Zehavi and Seiber (1996) re-
ported levels of TFA in rain and fog that were high but consistent with the mea-
surements of Frank et al. (1996). Grimvall and co-workers have shown not only
that TFA is widely distributed — from Arctic to Antarctic — at low concentra-
tions, but that it was present in precipitation in the Swedish Arctic some 400
years ago. The levels in old ice from the Mårmo Glacier are similar to present
day concentrations. Given the lower limit of detection of Grimvall’s method,
the results in Table 7 are consistent with those in Table 5.
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Figure 4. TFA in air near Bayreuth from March 1995 to October 1996. The width of
each column represents the duration of the sampling; from November
1996 the length of a typical sampling period was 5 to 7 days.
The significance of the discovery of low concentrations of TFA in precipita-
tion in remote areas is that it indicates a geographically dispersed source. With
an atmospheric lifetime (towards rainout) of a few weeks, it would not be pos-
sible for TFA that is generated in the northern hemisphere to be present in
southern hemispherical precipitation.
There are too few results to map the deposition of TFA definitively, so that
a global mass balance is not possible. It is, however, instructive to examine the
actual deposition in a well characterized location, such as Bayreuth, and com-
pare it with the deposition from known sources predicted by atmospheric
models.
Global Measurements and Their Significance for Future Emissions
As described in a previous section, the current potential sources of trifluo-
roacetic acid would provide, at most, 2800 metric tons per year distributed
throughout the world. It is also apparent, from the atmospheric modelling
that has been carried out to assess the possible future deposition of trifluoro-
acetic acid, that this quantity will not be deposited evenly over the globe but
will fall preferentially on tropical regions. Models of future deposition can be
used to geographically distribute the current contribution. Kanakidou et al.
(1995) calculated that, from a global flux of 160,000 metric tons per year of tri-
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fluoroacetic acid, the deposition rate in Europe would be 0.5 µmol·m–
2
·yr
–1
;
scaling the current contribution to this would imply a deposition rate of 0.5 ×
2800/160,000 or 8.8 nmol·m
–2
·yr
–1
, equivalent to 1 µg·m
–2
·yr
–1
. The annual
rainfall at Nurnberg, near Bayreuth, is 620 mm·yr
–1
(van der Leeden, 1975)
so that the expected concentration in rainwater there from current sources is
1/0.62 µg·m
–3
or 1.6 µg·m
–3
, which equates to 1.6 ng·l
–1
.
The measured concentration at Bayreuth averages 100 ng·l
–1
so the expect-
ed concentration of trifluoroacetic acid in rain from known sources is less
than
1
/
60
th of that observed.
DEGRADATION
Overview
The ionic character of TFA precludes simple partitioning of the molecule
into fatty tissues of animals. As mentioned previously, it is expected that the
compound will be found in aqueous compartments such as lakes, rivers,
oceans, estuaries, and associated sediments. The steady state levels of TFA with-
in these compartments will depend on the ambient input concentrations of
Figure 5. TFA in precipitation near Bayreuth from March 1995 to October 1996.
Expanded scale from May to September 1996.
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TFA in precipitation and TFA removal processes. The ultimate fate of TFA will
largely depend on four processes:
A. Biodegradation
B. Abiotic mineralization
C. Accumulation or uptake of TFA by organisms (see Accumulation Sec-
tion)
D. Adsorption (see earlier section)
Processes A and B are truly degradative and contribute to a net loss of TFA
from the system; they will be the focus of discussion in this section. Process C
may result in transformation of TFA into a biological constituent molecule, a
fluorinated metabolite of TFA, or simply temporary sequestration. Process D is
not true removal of the TFA from the environment but rather sequestration of
the compound onto, for example, soil particles, resulting in less mobility or
bioavailability.
Biodegradation
Reductive dehalogenation is well known to be an effective natural mecha-
nism for destruction of chlorinated hydrocarbons. However, defluorination by
this mechanism is thought to be much more difficult due to the lower reduc-
Figure 6. TFA in precipitation at a forest site and nearby clearing (1-km south of the
University of Bayreuth). The higher column of each pair always represents
the TFA contents in the precipitation collected under the spruce tree
canopy, including tree canopy runoff.
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tion potential of the C-F bond (–2 volts). Monofluoroacetate (MFA) is known
to be degraded hydrolytically by enzymes generically known as fluoroacetate
dehalogenases. In fact, microorganisms containing these enzymes have been
isolated and shown to hydrolyze MFA with release of fluoride. However, these
organisms do not degrade difluoroacetate (DFA) or trifluoroacetate. This is
consistent with the relative stability of these compounds (TFA>>DFA>MFA)
(Emptage, 1994).
TFA may be considered to be an analogue of a central biological
metabolite — acetate. Transformation of TFA by enzymatic systems specialized
for acetate metabolisms might be another route for TFA transformation result-
ing either in mineralization or incorporation into cellular material. However,
rates for TFA transformation by this route have been estimated using purified
enzyme systems and found to be less than 0.5% of the rate of acetate (Peijnen-
burg et al., 1994), which suggests that TFA will not compete effectively with ac-
etate in biological systems.
Demonstration of TFA Biodegradation
The studies summarized below represent a variety of empirical approaches
to demonstrate TFA biodegradation in field- or laboratory-derived microbial
cultures.
Measurements
Determination of the biodegradative potential of microorganisms for a spe-
cific compound is typically carried out using one transformation, or lack of
transformation (of TFA), which can be inferred from one or more of the fol-
lowing analyses: TFA-dependent microbial growth, TFA-dependent oxygen
consumption, appearance of defluorination or decarboxylation products such
as methane or carbon dioxide using [1-
14
C]-labeled TFA, fluoride anion, and
direct demonstration of TFA loss.
Microbiological Approaches
1. Enrichment for degradative bacteria using the compound in question
(TFA) as the sole carbon and energy source in the presence of added nu-
trients such as ammonium or phosphate. In this case bacteria, which can
utilize the compound, will grow preferentially.
2. Microbial growth on standard carbon and energy sources is independent
of the test compound; however, biodegradation of the test compound is
dependent on growth and metabolism of the bacteria on natural or stan-
dard carbon sources. Standard biodegradation tests employing organic-
rich inocula such as digestor sludge is one example of this approach.
This is a co-metabolic mechanism where transformation of the test com-
pound does not result in any specific advantage for the organism or does
it result in any enrichment of bacteria that specifically degrade the test
compound.
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3. Intrinsic metabolic capability of a naturally occurring sample. Transfor-
mation of the test compound is assessed in a naturally occurring sample
without any additional nutrient or carbon source.
Anaerobic Degradation of TFA: Reductive Defluorination
Reductive processes are expected to be operative under anaerobic or highly
reducing conditions such as those referred to as “methanogenic” or “sulfate-
reducing.” Anticipated products from the reductive defluorination of TFA are
ultimately methane, fluoride, and carbon dioxide.
Early studies using unamended marine sediment (San Francisco Bay) or
freshwater sediments incubated with
14
C-labeled TFA clearly demonstrated a
reductive defluorination process resulting in sequential formation of DFA,
MFA, acetate and methane (Visscher et al., 1994). This sequence was observed
under methanogenic and to a lesser degree sulfate-reducing conditions. Other
respiratory acceptors such as nitrate, ferric iron or oxygen failed to produce
defluorination although some fluoroform was reported under aerobic condi-
tions. This phenomenon has only been observed in certain field samples by the
one laboratory. The results also showed that TFA inhibited methanogenesis at
low concentrations (10 µM) and that at higher concentrations of TFA, the
compound could inhibit its own biodegradation.
Reinvestigation of samples from the same field sites (by the same laborato-
ry) failed to confirm the initial results (Matheson et al., 1996). Another labora-
tory reported that marine sediments from the sites that had previously been
shown to be active by Visscher et al. (San Francisco Bay) for methanogenic TFA
biodegradation were unable to degrade TFA (Emptage et al., 1997). The initial
results of Visscher et al. have not been replicated despite numerous attempts
by the original investigators as well as others.
Chauhan et al. (1995) carried out anaerobic closed microcosm bottle incu-
bations inoculated with soils obtained from globally diverse sites as well as
anaerobic digestor sludge. These were assessed for transformation of [2-
14
C]-
labeled TFA to
14
CO
2
. These tests showed no evidence for biotransformation
with either nutritionally amended or unamended bottles.
An assessment of anaerobic biodegradation by using the “sequential col-
umn microcosm” was also conducted by the same group. This approach in-
volves a succession of compartmentalized but interdependent microbial
communities that are distinguished on the basis of mode of respiration. The
sequence starts with the most reducing conditions (i.e., methanogenic) and
progresses toward more oxidizing physiologies by separate additions of sulfate,
iron, nitrate, and ultimately oxygen into the sequence of chambers down-
stream of the initial methanogenic chamber. In this way a broad cross-section
of microbial metabolism may be individually assessed in one experiment. [1-]
or [2-
14
C]-labeled TFA input and
14
C-labeled TFA output as well as total radio-
active counts were monitored for each chamber and inferred from the loss in
steady-state levels of TFA. In addition, fluoroform and
14
CO
2
were assayed in
the methanogenic chamber. Results from these experiments showed sustained
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and long-term loss of 25 to 30% of the incoming TFA in the methanogenic
chamber. Only very minor amounts of
14
CO
2
were detected in the headspace
of the culture vessel. However, total counts of radioactivity exiting the chamber
were only slightly diminished. The results suggest that either the TFA was com-
plexed such that it did not migrate on the HPLC column and was thus invisible
by this method, or that it was transformed to another soluble aqueous phase
metabolite that was not recovered in the HPLC analysis. No losses were noted
for subsequent chambers involving sulfate reduction, iron reduction, nitrate
respiration, photosynthetic sulfide respiration, or aerobic dark respirations.
No definite product was ever identified from the putative methanogenic trans-
formation; however, a conversion of TFA to a nonvolatile product was hypoth-
esized to account for the losses. The inference that some biotransformation
occurred based on loss alone must be viewed cautiously until direct evidence
for the “nonvolatile product” can be obtained.
Aerobic Degradation of TFA: Decarboxylation
Compelling evidence for decarboxylation of TFA was obtained using pure
cultures of Azoarcus tolulyticus tol-4, or Pseudomonas putida strain KZ6R. The re-
searchers demonstrated release of labeled carbon dioxide from TFA labeled in
the 1-position (Chauhan et al., 1995). Decarboxylation activity was strictly de-
pendent on the substrates on which the cells were grown. A. tolulyticus had to
be pre-grown on toluene as the carbon source with nitrate as the electron ac-
ceptor, whereas P. putida had to be aerobically pre-grown on 4-chloroben-
zoate. Neither organism exhibited the activity during its normal growth on
these substrates but rather only after harvesting the cells, concentrating them
and resuspending them in fresh medium. The TFA decarboxylation activity of
the cells was dependent on the age of the culture and the metabolic state of
the cells. The resting cells rapidly lost the ability to decarboxylate TFA. Howev-
er, the cells were capable of transforming up to 16% of millimolar levels of TFA
during the optimal period of activity. The exacting nutritional requirements,
and the instability of the activity argue against extrapolating this observation
to the natural environment outside the laboratory. It is not known at this time
how effective the transformation is at very low levels of TFA anticipated in the
environment. It is also unclear as to what the biochemical mechanism is other
than the correlation with aromatic catabolism.
Aerobic Closed Bottle Tests using Diverse Inocula
The transformation of [1-
14
C] TFA was assessed in aerobic microcosms in-
oculated with composite samples consisting of a blend of soils from geograph-
ically diverse regions of the globe, digestor sludges, or organic rich soil
fractions from sites in Michigan (Chauhan et al., 1995). No evidence for aero-
bic mineralization was obtained from analysis for
14
CO
2
in these incubations.
A semicontinuous activated sludge test was conducted by van Ginkel and
Kroon (1992a,b) according to EEC, OECD, and ISO test guidelines. The pro-
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cedures assessed fluoride release from aerobic digestor sludge incubations
that included co-metabolic substrates yeast extract, acetate or peptone and vi-
tamin B
12
. The semi-continuous sludge test demonstrated that TFA had no ef-
fect on the efficacy of the sludge as a biodegradation agent. No fluoride was
detected in the effluent from the SCAS units even though there was an approx-
imately 20% reduction of TFA in the effluent (van Ginkel and Kroon, 1992a).
Closed bottle tests were also performed. These incubations typically lasted for
periods of up to 84 days after which fluoride was measured. No detectable flu-
oride was observed in the bottle tests. The tests were considered negative for
TFA biodegradation (van Ginkel and Kroon, 1992b).
Aerobic Biodegradation of TFA by Monooxygenase-Containing Bacteria
Certain bacteria contain enzymes (monooxygenases) that are capable of in-
serting oxygen into aliphatic and aromatic hydrocarbons. By virtue of this,
these bacteria are capable of aerobic degradation of a wide variety of chlorinat-
ed hydrocarbons such as trichloroethylene (TCE), dichloroethylene and vinyl
chloride. To test for the ability of monooxygenases to degrade TFA, a set of
methylotrophic and propanotrophic bacteria were grown and tested against
compounds known to be degradable (i.e., TCE) as well as TFA. DeFlaun
(1996) tested nine strains comprising Pseudomonas mendocino KR1, Pseudomonas
putida F1, Methylosinus trichosporium, Mycobacterium vaccae, as well as five isolates
(designated ENV2C, ENV2D, ENV2R, ENV2W, and ENVOB) that were known
to degrade HFCs or HCFCs. These strains were tested against [1-
14
C] TFA with
TCE as a positive control (Deflaun, 1996). Assays were performed for
14
CO
2
,
fluoride, MFA, and TFA. The results showed that even with cultures which vig-
orously degraded TCE within 24 hours there was no detectable decarboxyla-
tion or dehalogenation of TFA after 13 days. No MFA or fluoride was detected
in any incubation.
Abiotic Mineralization
Irradiation of natural waters by sunlight can produce reactive species such
as HO·, RO
2
·, and O
2
·H and solvated electrons. The only ones capable of oxi-
dizing or reducing TFA are HO and solvated electrons (eAq
–
). However, the
concentrations of these species in natural waters are likely to be very low and
their rates of reaction with TFA, together, do not support a role for these
agents in the destruction of TFA (Maruthamuthu et al., 1995; Maruthamuthu
and Huie, 1995).
One possible photocatalytic mechanism involving ferric- or manganese-
assisted photodecarboxylation (photo-Kolbe) reactions was shown to be ef-
fective for degradation of chloroacetates and mono- or difluoroacetates but
not for trifluoroacetate (Maruthamuthu and Huie, 1994).
Trifluoroacetic acid has been shown to undergo photooxidation on a vari-
ety of iron oxyhydroxide surfaces (Pehkonen et al., 1995). Similarly with the
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other fully halogenated acetic acids, this reaction proceeds by a photo-Kolbe
mechanism to yield HF and CO
2
, but the rate is so low that it is unlikely to have
much environmental relevance.
TFA was shown to be inert to photocatalytically generated oxidizing species
on semiconductor surfaces. TFA was also shown to be inert to redox- and rad-
ical-induced degradation except at pHs above 10.5 (Asmus et al., 1994).
In a study of the rates of reductive dehalogenation versus structure, t
1/2
times for reductive transformation of a series of chlorocarbons as well as TFA
(based on structure-activity relationships) were obtained under anaerobic con-
ditions in the presence of sediment. The reductive dehalogenation reactions
were presumably microbial in origin, but this was not directly demonstrated in
the study. In contrast to the fully chlorinated ethanes and ethenes, which are
readily dehalogenated, the t
1/2
for TFA — obtained by extrapolation — was
found to be over 200 years in these biologically active sediments (Peijnenburg
et al., 1994). These results are consistent with those obtained in other labora-
tories investigating the reductive dehalogenation of TFA but contrast with the
report by Visscher et al. (1994).
Due to the recycling of TFA between water and air and the rapid rainout of
TFA from clouds (10 days), it is unlikely that there will be any aqueous phase
degradative process that could compete with rainout kinetically. Oxidation of
TFA by NO
3
·, SO
4
–
·, Cl
2
–
·, OH· followed by decarboxylation may occur with a
t
1/2
on the order of 80 years. Ferric-assisted photodecarboxylation was shown
to be ineffective for TFA (Wine, 1994).
Conclusion
TFA is a highly stable molecule to known abiotic or biological degradation
mechanisms. The microbiological studies summarized here involved both em-
pirical screens of naturally occurring soils as well as focused tests of known deg-
radative mechanisms. TFA biodegradation was investigated in aerobic and
anaerobic bacterial cultures with pure isolates as well as crude environmental
samples. These studies have not identified evidence for a widespread, environ-
mentally significant, biological mechanism for defluorination of TFA. There-
fore, the assessment is that TFA will be very long-lived in the environment. This
prediction suggests that renewed emphasis be placed on identifying the back-
ground levels of TFA in the environment as well as how this TFA may compart-
mentalize.
Despite published data claiming rapid reductive defluorination of TFA
(Visscher et al., 1994), the evidence for this mechanism of TFA removal must
be considered hypothetical until the original result can be further confirmed
or supported. The weight of theoretical considerations as well as the combined
experience of the five laboratories involved in these projects suggests that con-
firmation is unlikely. This does not rule out the possibility that the result was
correct but due to a rare concatenation of physical and biological factors.
The result reported by Chauhan et al. (1995) that clearly demonstrated de-
carboxylation, but not defluorination, of TFA may be of limited environmental
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94
significance as a mechanism for destruction of TFA in the environment. Expe-
rience with a wide variety of real environmental samples by that laboratory, as
well as other laboratories, suggest that decarboxylation of TFA is not wide-
spread in nature or is it easily induced where it is present. Indeed, even the lab-
oratory strains only produce the activity under very controlled conditions. The
fate of the CF
3
–
moiety in the decarboxylation reaction is unknown.
Therefore, for the purpose of this risk assessment, we will assume, as a rea-
sonable worst case, that no degradation of TFA is likely to occur in the environ-
ment.
ACCUMULATION
The physico-chemical properties of TFA (high solubility and low K
ow
) indi-
cate that the molecule would not partition significantly to the lipid phase of
cells and therefore significant bioconcentration by this mechanism would not
be expected. However, other mechanisms that might lead to a bioaccumula-
tion of TFA have been investigated.
Due to the structural similarity of TFA to acetate, a major biosynthetic pre-
cursor molecule, studies were performed to assess the level of cellular incorpo-
ration of TFA into biomass. Bott and Standley (1994) investigated the
incorporation of [2-
14
C] TFA into microbial cell mass, small invertebrates, and
plants. The cellular components were assayed and comprised the lipid, carbo-
hydrate, and proteinaceous fractions.
Aquatic communities, including bacteria, aquatic invertebrates, and plants,
were set up in flowing water mesocosms to provide a source of biological sam-
ples for experimentation. One of these communities was exposed to NaTFA
(30 µg·l
–l
) for nearly 2.5 years; the other was an unexposed control run for a
comparable period of time. Results showed that sediment microbial commu-
nities exposed to TFA in mesocosms for approximately 2.5 years incorporated
more radiolabeled TFA than communities exposed for <1 year, with rates in-
creasing from –1 × 10
–l3
to 22 × 10
–13
µg TFA·cell
–l
·day
–l
(Bott and Standley, in
review). The cause of the change, whether due to a microbial adaptation to the
TFA exposure or a change in some other environmental variable(s) is unclear.
Incorporation was correlated positively with the amount of TFA added to sam-
ples and exposure time.
Microbial incorporation of radiolabeled TFA by cells cultured from those
sediments (at a test concentration of 10 µg·l
–l
) demonstrated incorporation
into biomolecule fractions such as lipids, proteins and residual (e.g., cell
wall) materials. Incorporation into these fractions in bacteria increased from
0.005% of total counts to 0.024% over this 2.5 year period (Bott and Stand-
ley, 1994; Standley and Bott, in review). Oligochaete worms exposed to 40 µg
NaTFA·l
–l
also showed low uptake of TFA into biomolecules. The oligocha-
etes contained radiolabel that was not extractable as TFA; it was primarily as-
sociated with the protein fraction. The presumptive incorporation product
from TFA has not been identified. Jewelweed also contained radiolabel that
was not extractable as TFA, but the levels of label incorporated ranged from
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0.0069% of the total exposure for roots to 0.012% of the total for leaves after
a 32-day exposure. Apparent incorporation factors were low for other macro-
invertebrates and plants. These studies suggest that microbial biomass is not
a major sink for TFA, but that there is an association of TFA with biomass that
could be due to biochemical incorporation.
As part of a study to determine the possible defluorination of TFA by the
freshwater green alga, Selenastrum capricornutum, van Dijk (1996) reported a
bioaccumulation factor (BCF) of approximately 10, based on radiolabeled res-
idues in cells filtered from the medium. However, the study did not correct for
any radioactivity from the medium that may have been retained on the filter
due to evaporation or adsorption and therefore may have overestimated the
radioactivity associated with the cells.
In a toxicity study with Lemna gibba (duckweed, a floating aquatic higher
plant) described in another section,
14
C-labeled residues of TFA were deter-
mined after 7 days exposure to a range of concentrations in the culture medi-
um (Smyth et al., 1993). The bioconcentration factors were low, ranging from
1.0 to 1.6, reflecting the hydrophilic properties of the substance and the lack,
in this species, of a differentiated water transport (transpiration) system.
In terrestrial higher plants, the uptake and transpiration of water, to replace
that lost by the leaves as a consequence of photosynthetic gas exchange, pro-
vides a mechanism for the bioaccumulation of TFA. Prior to the AFEAS pro-
gram, root uptake and transport of TFA in the xylem tissues had been
demonstrated in tomato seedlings (Rollins et al., 1989), using nuclear magnet-
ic resonance imaging. However, the exposure concentration was very high (ap-
proximately 51,000 mg HTFA·l
–1
as a buffered solution) and the exposure
duration was too short (17 hours) to evaluate the bioaccumulation potential.
Subsequently, sunflower (Helianthus annuus) seedlings were exposed to a sin-
gle concentration (2 µg HTFA·l
–1
) of
14
C-radiolabeled trifluoroacetic acid in
the aqueous (hydroponic) medium surrounding the roots (Thompson et al.,
1994). Plants were removed at intervals and the concentration of
14
C residues
determined in root, stem, and leaf tissue. After 12 days the remaining plants
were transferred to clean medium and sampled for a further 4 days.
14
C resi-
dues in the leaf tissues increased continuously over the period of exposure,
with a BCF of approximately 22 after 12 days. The stem tissue behaved similarly
but with a lower rate of accumulation (12-day BCF of approximately 5). Root
tissue reached apparent equilibrium after 5 days exposure with a BCF of ap-
proximately 3. The accumulation rate was somewhat less than would be expect-
ed from passive influx in the transpiration stream without efflux. The
concentrations measured in the root medium also suggested either a concur-
rent efflux or some partial barrier to influx.
All tissues showed a decline in
14
C-residue concentrations on transfer to
clean medium. Although this can be attributed largely to dilution by growth, a
significant quantity of radiolabel was lost to the medium surrounding the roots
and it was concluded that the plants showed some excretion (depuration) of
accumulated radiolabel to the root medium. More than 80% of the
14
C resi-
dues in the leaves were found to be extractable in water after tissue maceration
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96
(Thompson and Gillings, 1996). Fractionation of the extract using ion chro-
matography showed the residues co-eluted with a trifluoroacetate standard
spiked into leaf extract, suggesting that no significant metabolism of TFA had
occurred.
As part of a 5-week hydroponic toxicity study with wheat (Thompson et al.,
1995) (described in a subsequent section), the accumulation of
14
C residues of
trifluoroacetate was also monitored. At 1 mg NaTFA·l
–1
, the bioconcentration
factor in the total aerial tissue (shoots), based on fresh weight, increased con-
tinuously over the exposure period, to a final value of 27. The concentrations
in the tips of the shoots were approximately four times greater than in the re-
maining shoot tissues. This factor was used to estimate a tissue concentration
of approximately 1000 mg·kg
–1
(fresh weight) corresponding with tissue ne-
crosis observed in the shoot tips at 10 mg NaTFA·l
–1
, after 28 days. After 35
days, a BCF of 43 was determined for the whole shoots.
Davison and Pearson (1997) analyzed tissues of soya and wheat exposed to
NaTFA by determination of inorganic fluoride, after fusion with sodium car-
bonate to cleave any C-F bonds. Concentrations were expressed as NaTFA in
dry tissue, after background fluoride subtraction. For soya exposed to 5 mg·l
–1
under hydroponic conditions for 44 days, the concentration in the tissue of the
oldest (first trifoliate) leaves reached a plateau after approximately 20 days.
Each successive set of leaves attained higher tissue concentrations, except for
the fifth trifoliates that were still expanding at final harvest but already con-
tained a higher concentration than had been reached by the older, second tri-
foliates. As symptoms first appeared on the third trifoliate leaves, the tissue
concentration was approximately 150 mg NaTFA·kg
–1
dry weight. It was con-
cluded that the stage of development of the leaf is important and that TFA is
most toxic to young, expanding leaves and has little effect on mature leaves
with a similar tissue concentration.
In a further study with soya at 1 mg NaTFA·l
–1
, Davison and Pearson (1997)
found that the slight symptoms observed were associated with tissue concentra-
tions of 160 to 190 mg NaTFA·kg
-1
dry weight. Tissue concentration in the
whole shoots of wheat exposed to the same concentration (which caused no
effects) reached a plateau at approximately 55 mg NaTFA·kg
–1
dry weight after
25 days and had decreased slightly after 43 days. At 5 mg NaTFA·l
–1
, the tissue
concentration was approximately 190 mg TFA·kg
–1
dry weight when growth
was beginning to be inhibited.
The final leaf/shoot tissue concentrations (on a dry weight basis) for the dif-
ferent hydroponic studies are summarized in Table 8. For comparison, these
are also expressed as bioconcentration factors based on tissue fresh weights, al-
though it should be noted that fresh weights were not determined (and are
therefore estimated) for some of the studies. Different analytical methods
were also employed (see above). At concentrations at or below the no effect
level of 1 mg·l
–1
, bioconcentration factors ranged from 5.4 to 27. Frank (1994)
reported concentrations of TFA in spruce needles sampled from Tübingen,
Germany, of 98 to 195 ng.g
–1
(presumed to be based on fresh weight). Recent
measurements (Frank et al., 1996) suggest an average concentration of approx-
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imately 100 ng·l
–1
in rainwater, suggesting a bioconcentration factor of 1 to 2.
Although conifer needles are relatively long lived, this may indicate that the
relatively low water usage of such leaves may enable excretory translocation to
become more significant.
There is good evidence to suggest that the toxicity of TFA to plants is related
to the tissue concentration of the accumulated material that is dependent on
the transpiration flow from the roots. However, because the mode of action ap-
pears to involve specific processes in the leaf, possibly related to leaf expan-
sion, it is not possible to specify precisely the critical tissue concentrations in
terms of the whole leaf. However, the lowest tissue concentration associated
with adverse effects was 150 mg·kg
–1
dry weight, in expanding soya leaves (Davi-
son and Pearson, 1997). It would appear that when exposure is at or below the
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threshold for effects, the tissue concentrations tend to plateau and accumulat-
ed TFA is reduced by growth and by excretion via the roots. Leaf-fall represents
an ultimate removal mechanism for any residual TFA in mature leaves.
EFFECTS ASSESSMENT
Microorganisms
Effect of TFA on Methanogenic Communities
A vital part of the carbon cycle involves the formation and degradation of
acetate during anaerobic methanogenesis. This process occurs globally but
can be differentiated in terms of specific environments. Commercially or envi-
ronmentally important methanogenic populations are found in anaerobic di-
gestors, sediments, and rumen of ruminant animals. The principal
microorganisms are generically referred to as methanogens, many species of
which specialize in the transformation of acetate to methane and carbon diox-
ide. Methanogenesis was assessed for any toxic effects of TFA by measuring the
rates of methane formation from anaerobic digestor samples, sediment sam-
ples, and rumen samples in the presence of increasing concentrations of TFA
up to 10 mM (1114 ppm) according to Emptage et al. (1997). The results indi-
cated that there was no significant effect of TFA, even at the 10 mM concentra-
tion, on any incubation. Monofluoroacetate did however exert toxicity, as
expected, at levels as low as 10 mM. These results contrast with those reported
by Visscher et al. (1994), which suggested inhibition of methanogenesis at low
concentrations (0.1 mM TFA). The discrepancy in the results cannot be ac-
counted for at this time except to note that the toxicity was observed in the
same sediments which exhibited rapid TFA biodegradation (see a previous sec-
tion). TFA biodegradation was not observed in the study by Emptage et al.
(1997). Overall, the results suggest that TFA is inert in these systems and that
endogenous methanogenesis was neither stimulated nor inhibited.
Effect of TFA on Activated Sludge
A semicontinuous activated sludge test was conducted by van Ginkel and
Kroon (1992a) according to EEC, OECD, and ISO test guidelines. This test
also indicated that TFA had no discernible effect on the performance of the
sludge for catalyzing biodegradation of organic carbon.
Effect of TFA on Acetate Metabolism by Microbial Communities
Acetate is a key intermediate in most living organisms. Ecologically, the
mineralization of acetate to carbon dioxide is a key link in the biogeochemical
carbon cycle. Therefore, it is essential to know whether TFA, which is structur-
ally close to acetate, could interfere with acetate metabolism.
Experiments were conducted to assess the interaction of TFA with the me-
tabolism of [1-
14
C] acetate by freshwater microbial communities which had not
been previously exposed to TFA (Bott and Standley, in review). The approach
comprised additions of non-labeled TFA or non-labeled acetate to parallel in-
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cubations containing [1-
14
C] acetate. Evidence for interaction with metabo-
lism of acetate would then be inferred from the effects observed on
incorporation of radiolabel into biomass or evolution of
14
C carbon dioxide.
According to Bott and Standley (1994), none of the differences in
14
C acetate
metabolism between the unamended control and any of the TFA or acetate
treatments were statistically significant (Dunnetts’s test, p > 0.05). Between 85
and 90% of the [1-
14
C] acetate radioactivity was found in the biomass and there
was no significant difference between controls and TFA-treated incubations.
These results suggest that TFA at concentrations several orders of magnitude
higher than those anticipated in the environment did not impact acetate min-
eralization to carbon dioxide or did it affect incorporation of acetate into cel-
lular material.
Similar experiments were performed using samples derived from TFA-
treated (30 µg·l
–1
3-months exposure) or control mesocosms. Experimental
TFA additions were 200, 20, and 2.4 µg·l
–1
. Cell-specific rates of acetate min-
eralization rates were similar for samples from the TFA-treated and control
mesocosms. With samples from each mesocosm there were never statistically
significant differences in [1-
14
C] acetate metabolism over the range of TFA
concentrations tested. The percentage of radioactivity in the biomass ranged
between 80 and 90% and there were no significant differences between un-
amended controls and TFA-treated samples.
At extraordinarily high nonradioactive acetate additions of 4 to 450 mg·l
–1
,
there was suppression of
14
C acetate metabolism, as would be anticipated from
expected isotopic dilution of
14
C acetate. At 440 mg·l
–1
, TFA also suppressed
14
C acetate metabolism, although to a lesser extent than did cold acetate sug-
gesting some weak interaction of TFA with acetate metabolism at these con-
centrations.
These competition experiments using benthic microbial communities also
clearly demonstrated that TFA and acetate were not strongly competitive com-
pounds in terms of mineralization or incorporation into biomass.
The effect of TFA has been investigated in three species of free-living nitro-
gen-fixing bacteria (Nagel and Odom, 1997). The species were selected on the
basis of phylogenetic diversity and because they have been used in many labo-
ratory studies on the biochemistry of nitrogen fixation. A common aerobic soil
microorganism (Azotobacter vinelandii), a freshwater photosynthetic bacterium
(Rhodobacter capsulatus), and a common anaerobe (Clostridium pasteurianum)
were the test species. The effect of TFA on these species was determined dur-
ing nitrogen-dependent growth, growth on fixed nitrogen (ammonium ion),
and on nitrogenase activity itself. The experiments were designed to deter-
mine whether TFA was specifically toxic to nitrogen fixation versus a more gen-
eral physiological aspect. No effect of TFA on growth either by N
2
fixation or
with ammonium ion as nitrogen source was noted even at concentrations as
high as 1 mM (∼100 ppm) TFA with either the R. capsulatus or A. vinelandii
strain. C. pasteurianum was only tested for TFA toxicity during growth on fixed
nitrogen due to difficulties in obtaining significant rates of N
2
fixation with this
strain. Increases in production of molecular hydrogen were noted at very high
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TFA concentration (>50 ppm) in Rhodobacter capsulatus. This effect could be
due to either stimulation of nitrogenase activity or inhibition of hydrogen re-
utilization at these very high TFA concentrations.
Aquatic Organisms
Only acute toxicity studies are available; unless otherwise stated these were
carried out with NaTFA. A number of studies employed radiochemical analysis
of the test solutions, which confirmed the nominal concentrations and dem-
onstrated no significant decline during the test period. In view of this and the
known stability of the compound and low potential for adsorption, the remain-
ing studies were carried out without analysis of water concentrations. All end-
points reported here are based on nominal concentrations.
Acute toxicity in fish, determined in Brachydanio rerio, did not reveal any ef-
fect at 1200 mg·l
–1
during 96 hours of exposure. For water flea Daphnia magna,
the same result was obtained after 48 hours of exposure. The acute NOEC for
these two species, representing primary and secondary consumers, is 1200
mg·l
–1
NaTFA (corresponds to 1000 mg·l
–1
of TFA).
A study using HTFA with Daphnia magna showed a 24-hour EC
50
of 55 mg·l
–1
(Rhône-Poulenc, 1995); this was attributed to a pH effect.
A toxicity study on duckweed (Lemna gibba) showed EC
50
values for frond in-
crease and weight increase of respectively: 1100 and 1200 mg·l
–1
. For both end-
points, a NOEC of 300 mg·l
–1
has been recorded (Smyth et al., 1993).
Toxicity tests with sodium trifluoroacetate in algae were conducted on 11
different species: Selenastrum capricornutum, Chlorella vulgaris, Scenedesmus sub-
spicatus, Chlamydomonas reinhardtii, Dunaliella tertiolecta, Euglena gracilis, Phaeo-
dactylum tricornutum, Navicula pelliculosa (Smyth et al., 1994a), Skeletonema
costatum (Smyth et al., 1994b), Anabaena flos-aquae (Smyth et al., 1994c) and Mi-
crocystis aeruginosa. These 11 species belong to 4 different classes: Chlorophyce-
ae (4 freshwater and 1 marine species), Euglenophyceae (1 freshwater
species), Cyanophyceae (2 freshwater species) and Bacillariophyceae (1 fresh-
water and 2 marine species).
Selenastrum capricornutum (alternative names: Raphidocelis subcapitata or
Pseudokirchneriella subcapitata) was the most sensitive species for sodium trifluo-
roacetate. This unicellular alga is the most frequently used for ecotoxicity deter-
mination under laboratory conditions. Based on the results of five toxicity tests
with Selenastrum capricornutum, a concentration of 0.12 mg·l
–1
(120 µg·l
–1
) can
be considered a toxicity threshold concentration. Adverse effects on the growth
of this species were not found at this concentration. For the remaining 10 algal
species the EC
50
values were all higher than 100 mg·l
–1
(see Table 9).
Algal studies were also conducted with potential metabolites of trifluoroac-
etate, like difluoroacetate and monofluoroacetate. Monofluoroacetate is a tox-
ic substance which mode of action is the inhibition of the citric acid cycle.
Studies with Selenastrum capricornutum and Scenedesmus subspicatus revealed ef-
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fects on growth at sodium monofluoroacetate concentrations which were sev-
eral orders of magnitude lower than effect concentrations of sodium
trifluoroacetate (see Table 10).
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There is some evidence that exposure to trifluoroacetate results also in the
inhibition of the citric acid cycle. Selenastrum capricornutum exposed to sodium
trifluoroacetate showed a recovery of the growth when citric acid was added.
Furthermore, there seems to be a correlation between the sensitivity of an algal
species for trifluoroacetate and monofluoroacetate. Both compounds are not
toxic for Chlorella vulgaris (see Table 10).
One semi-field study with mesocosm streams has been conducted with sodi-
um trifluoroacetate (Bott and Standley, 1995) to study the potential effects of
trifluoroacetate on freshwater algal communities and primary productivity.
The long-term exposure to a mean sodium trifluoroacetate concentration of
31 to 32 µg·l
–1
had no effect on the primary productivity of the diatom domi-
nated algal flora. Effects on organic carbon excretion which were related to
high levels of TFA were noted in some experiments. TFA did not alter the algal
species composition in the stream mesocosm.
In conclusion, based on the results of laboratory toxicity tests in fish, Daph-
nia, duckweed, and in a large number of algal and based on the results of the
semi-field study with stream mesocosms, an exposure of an aquatic ecosystem
to a sodium trifluoroacetate concentration of 0.12 mg·l
–1
did not show adverse
effect.
Terrestrial Organisms
Terrestrial Plants
Inputs of TFA to the terrestrial environment will be exclusively via wet and
dry deposition from the atmosphere. Due to the high water solubility of TFA
and its salts, the assessment of terrestrial effects should be based primarily on
precipitation and soil-water concentrations.
Prior to 1993, there was little information available on the toxicity of TFA to
higher plants. Poignant (1957) had observed phytotoxicity to Triticum vulgare,
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by pre-emergence application, at and above 1500 mg NaTFA·l
–1
. He also re-
ported the same level of activity to maize (Zea mays), meadow grass (Poa an-
nua), and sweetcorn (Alopecurus agrestis). Ingle (1968) reported significant
inhibition of shoot growth of germinating wheat seeds and root growth of ger-
minating tomato seeds at 114 and 285 mg HTFA·l
–1
(buffered solutions), re-
spectively; TFA showed the weakest phytotoxicity of the seven halogenated
aliphatic acids tested. Atmospheric exposure of a conifer, Taxus baccata var.
Repandens, to HTFA had been investigated by gradually increasing the vapor
concentration from 100 to 2100 ppm over a period of 21 days (Freist, 1986).
No effects were observed after 19 days (up to 1900 ppm). Slight necrosis of the
tips of the young needles was observed at 2100 ppm, which was attributed to
the acidity of condensation water which formed on the needles.
Because of these indications of phytotoxic potential, the effect of TFA has
been investigated for a large number of terrestrial higher plant species. Unless
otherwise stated, the experiments employed sodium trifluoroacetate and the
concentrations are expressed as the sodium salt.
In preliminary studies with sunflower, wheat, and mung bean, using OECD
protocols, the exposure concentrations (1, 10, 100, and 1000 mg·kg
–1
) were ex-
pressed by dry weight of soil (Windeatt and Thompson, 1993a,b). As explained
above, the results from such tests are of limited value since the soil moisture
levels, and thus the soil water concentrations of TFA, fluctuate widely between
plant watering occasions. However, they provided evidence that further studies
on higher plants were necessary. Significant effects on germination and emer-
gence (14 days from seed sowing) were observed at 1000 mg·kg
–1
dry soil for
wheat and mung bean and at 10 mg·kg
–1
dry soil for sunflower. Subsequent
vegetative growth was more sensitive; after a further 14 days, a significant re-
duction in the weight of aerial tissue was observed at 10 mg·kg
–1
dry soil for
wheat and mung bean and at 1 mg·kg
–1
dry soil for sunflower.
The effects of aqueous exposure to TFA on seed germination was investigat-
ed (Thompson and Windeatt, 1994) for 10 species of terrestrial plant, includ-
ing dicotyledons (sunflower, cabbage, lettuce, tomato), leguminous
dicotyledons (mung bean, soybean), and monocotyledons (wheat, corn, oats,
rice). The seeds were added to filter papers soaked with solutions of TFA; ex-
cept for small seed species, the seeds were also presoaked in the same solu-
tions. No effects were observed for any species at the maximum concentration
tested that was 1000 mg·l
–1
.
Seedlings of plantain (Plantago major) and wheat (Triticum aestivum variety
Katepwa) were exposed to a range of TFA concentrations in the aqueous (hy-
droponic) medium surrounding the roots. After 14 days, growth of plantain
was affected at 100 mg·l
–1
of sodium trifluoroacetate, but there was no effect at
32 or 10 mg·l
–1
(Thompson, 1995). Growth of wheat was affected at the lowest
concentration tested (32 mg·l
–1
) in the preliminary 14-day study and this spe-
cies was tested again in a longer study at lower concentrations (Thompson et
al., 1995). After 5 weeks, wheat growth was inhibited at 10 mg·l
–1
with tissue ne-
crosis evident first in the shoot and leaf tips; there was no effect at 1 mg·l
–1
of
TFA.
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In a similar hydroponic study (Davison and Pearson, 1997), growth of the
same variety of wheat was inhibited at a concentration of 5 mg·l
–1
; the effects
became apparent after 30 to 35 days, with more severe effects at 10 and
100 mg·l
–1
. As in the earlier study, there was no significant effect at 1 mg·l
–1
at
the end of the exposure (43 days). Using the same system, another variety of
wheat (Hanno) showed lower sensitivity; no effects were found at 5 mg·l
-1
and
only slight growth inhibition at 10 mg·l
–1
after 52 days exposure.
Under these hydroponic conditions, soya bean plants, which had proven
more sensitive than several other species to foliar application of TFA (see be-
low), gave results essentially similar to those of the sensitive wheat variety. In
a preliminary experiment, symptoms of toxicity were observed at 5 and
10 mg·l
–1
but were more severe at the higher concentration, becoming appar-
ent after 20 days. Subsequently, exposure concentrations of 5 mg·l
–1
and
1 mg·l
–1
were tested, in separate experiments, for 43 and 33 days, respectively.
At 5 mg·l
–1
, effects on growth (dry weight) were observed after approximately
30 days, with symptoms appearing more severe in the younger, expanding
leaves. At 1 mg·l
–1
, there was no significant effect on final weight (33 days).
However, in some plants approximately 25% of the 4th and 5th trifoliate leaves
(the youngest) appeared to show slight symptoms (slight rounding of the tip
and barely discernible necrosis of the margin), but this was variable between
plants.
Other studies (Davison and Pearson, 1997) have investigated the effects of
foliar application of solutions of NaTFA. Seedlings (7 days old) of seven spe-
cies of terrestrial plant (sunflower, soya, wheat, maize, oilseed rape, rice, and
plantain) were field-grown, to ensure that the leaves were fully hardened,
sprayed (60 microns droplet size) with solutions of NaTFA for 7.5 hours and
kept wet with the mist for a further 16.5 hours. The soil and plant roots were
protected from the application. After 3 weeks, there was no effect of NaTFA on
the height, leaf number, stomatal conductance, final harvest weight or chloro-
phyll and carotenoid concentrations of any of the species at the maximum con-
centration tested which was 100 mg·l
–1
. There was a significant decrease in
specific leaf area only for wheat at 100 mg·l
–1
.
A subsequent study, at only 100 mg·l
–1
, used a similar system but with labo-
ratory grown plants (wheat, maize, sunflower and soya), both with and without
protection of the soil and roots from the spray (Davison and Pearson, 1997).
Only soya showed any symptoms of toxicity. These were apparent by both
routes of exposure, suggesting that laboratory grown plants were more sensi-
tive to foliar application than those field-hardened; however, the symptoms
were much more severe when the NaTFA solution was able to reach the roots.
In a further experiment, soya seedlings were exposed to a range of concen-
trations of NaTFA applied to the soil surface in a volume equivalent to 10 mm
of rainfall every 3 days for 44 days. There was no effect at 1, 5, and 10 mg·l
–1
on
dry weight, leaf size, or stomatal conductance (transpiration) and no visible in-
jury. At 100 mg·l
–1
, symptoms of toxicity were apparent and leaf size (but not
dry weight) was reduced, suggesting TFA affected leaf expansion. Although
this regime simulated relatively high rainfall (100 mm per month), the effec-
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105
tive concentration was more than one order of magnitude higher than by hy-
droponic exposure. This suggests that the hydroponic exposures represent the
worst case situation, possibly because transpiration rates are maximized.
Studies by Emerich (1997) have shown that TFA at a concentration of
1 mg·kg
–1
of soil had no effect on the germination or growth of soybean seed-
lings. Also, 1 mg TFA·kg
–1
had no discernible effect on the nitrogen-fixing ca-
pacity of these soybean plants as determined by plant nodule development or
by direct assay of nitrogenase activity of these soybean nodules. Toxic effects
were, however, observed in the development of the plants and in the nodula-
tion patterns at 10 and 100 mg TFA·kg
–1
. Comparable results were obtained
when the studies were carried out under hydroponic conditions, at concentra-
tions of 1, 10, and 100 mg TFA·l
–1
. There were no significant effects on any pa-
rameter at 1 mg·l
–1
, with pronounced effects on plant development at 10 and
100 mg·l
–1
.
It can be concluded that the most significant route of exposure for terres-
trial plants is uptake from soil-water via the roots. TFA enters the transpiration
stream and is taken to the shoots and leaves where there is a tendency for TFA
to bioaccumulate as the water evaporates. For the most sensitive species, with
long-term, continuous exposure under conditions allowing high transpiration
rates, no effect on the growth of plants occurs at 1 mg·l
–1
(see Table 11). This
concentration is four orders of magnitude higher than the projected rainwater
concentration of TFA derived from CFC alternatives.
Terrestrial Invertebrates
Because of the low toxicity of TFA to aquatic invertebrates (see a previous
section) and other higher organisms, it was not considered necessary to test
terrestrial, soil-dwelling organisms.
Mammals
Distribution, Toxico-Kinetics, and Metabolism
The distribution half-life of TFA in rabbits was about 20 minutes (Kinoshita,
1989). After intravenous administration in rats, approximately 58% of the TFA
was distributed in the total body water, whereas plasma proteins were binding
10% (Holaday and Cunnah, 1976).
In human blood, TFA binds mainly to albumin (44 to 54%) and only little
(4 to 14%) to erythrocytes (Dallmeier and Henschler, 1981; Holaday and Cun-
nah, 1976).
After NaTFA administration via drinking water to male Wistar rats the
amount of organically bound fluorine in the plasma and the liver reached a
steady level within 3 days of treatment with a 1:1 ratio of liver to plasma con-
centrations (Stier et al., 1972).
In pregnant mice, the TFA concentration ratio between amniotic fluid/ma-
ternal plasma became >1 at 24 hours after an intravenous infusion. Approxi-
mately 20 to 30% of the TFA was bound to amniotic fluid and plasma
macromolecules (Ghantous et al., 1986).
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(
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TFA has been shown to be the ultimate biotransformation product of HCFC-
123 (2,2-chloro-1,1,1-trifluoroethane) (Harris et al., 1991; Olson et al., 1991). In
a study with HCFC-123 in lactating Sprague-Dawley rats, TFA was determined
in milk. The measured concentration in milk increased 5 to 23% after saponi-
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108
fication, indicating that a part of the TFA was covalently bound to peptides or
proteins (Buschmann, 1996).
TFA was also identified as a major metabolite of halothane. A study (Kinosh-
ita, 1989) using rabbits exposed to halothane demonstrated that the biliary
route is the most important excretory pathway for biotransformed TFA. A
large quantity of the acid is excreted into the bile, reabsorbed in the gas-
trointestinal tract, and appears in the circulating blood. The elimination half-
life after intravenous injection or oral administration of TFA was 34 and 48
hours, respectively. Enterohepatic circulation is one of the probable mecha-
nisms of the long lasting excretion of TFA.
Kawahara et al. (1988) have reported salivary excretion of TFA in surgical
patients and Mirkov et al. (1988) have reported excretion of TFA in the diges-
tive juice after halothane inhalation in guinea pigs.
The majority of metabolized HCFC-123 is excreted as TFA in the urine (Vin-
egar et al., 1994). Following a 6-hour exposure of lactating Sprague-Dawley rats
to 1000 ppm HCFC-123 and after 1 hour of nursing (15 ml milk per litter con-
taining 50 µg·ml
–1
TFA), each litter excreted over 16 hours an average quantity
of 79 µg TFA in 9.1 ml urine. This study illustrates the relative high concentra-
tion in the mothers milk and the prolonged retention of TFA by the rat pup
(Buschmann, 1996). In adult rats the plasma half-life after intravenous injec-
tion was 30 hours (Holaday and Cunnah, 1976).
Intravenous-injected TFA in two human volunteers was completely (95 to
100%) excreted via the urine over 63 to 72 hours (plasma half-life 25 to 32
hours) (Holaday and Cunnah, 1976).
In a 2-week (5-day week) exposure study with 22 ppm halothane the TFA
concentration in the blood in the second week was only slightly higher com-
pared to that of the first week. TFA level in human blood and urine are corre-
lated linearly with levels in blood (0.2 to 9 mg·l
–1
) being approximately three
times higher (Dallmeier and Henschler, 1981).
TFA has not been found to be metabolized to any appreciable extent in rats
(Fraser and Kaminsky, 1988).
Acute Toxicity
Acutely, trifluoroacetic acid would be categorized as being moderately toxic
by the oral route of administration having a LD
50
of a 2 to 5% solution in the
range of 200 to 400 mg·kg
–1
body weight in rats and mice (Patty, 1963). Oral
LD
100
, LD
10
, and LD
0
values were reported by another laboratory to be 1000,
500, and 250 mg·kg
–1
body weight, respectively (Kheilo and Kremneva, 1966).
The sodium salt of trifluoroacetic acid is much less toxic than the free acid.
No deaths were observed when mice were administered intraperitoneally up to
5000 mg·kg
–1
of NaTFA; whereas, a dose of 150 mg·kg
–1
HTFA caused death in
2 out of 5 mice, a result which is comparable to that obtained with an equimo-
lar dose of hydrochloric acid (Blake et al., 1969).
Other reported LD
50
values obtained with the sodium salt using the intrap-
eritoneal route of administration include: >4000 mg·kg
–1
(Rosenberg, 1971)
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109
and >2000 mg·kg
–1
(Airaksinen and Tammisto, 1968; Airaksinen et al., 1970).
An intravenous LD
50
for NaTFA was reported to be 1200 mg·kg
–1
for mice (Air-
aksinen and Tammisto, 1968).
Information on the acute inhalation toxicity of HTFA is limited to a single
Russian report (Kheilo and Kremneva, 1966). Values were reported for LC
50
for both mice and rats. The LC
50
for a 2-hour exposure in mice was reported
to be 13.5 mg·l
–1
(∼2900 ppm); additional values for LC
100
and LC
10
(2 hour)
were 20.4 mg·l
–1
(4366 ppm) and 9.2 mg·l
–1
(1968 ppm), respectively. The
2-hour LC
50
for the rat was reported to be 10 mg·l
–1
(2140 ppm) with addi-
tional values for LC
100
and LC
10
(2 hour) of 11.5 mg·l
–1
(2461 ppm) and
8.3 mg·l
–1
(1776 ppm). According to these investigators, for rats the thresh-
old concentrations are 4 mg·l
–1
(856 ppm) based on “the temperature reac-
tion of the body” and 1.5 mg·l
–1
(321 ppm) based on “neuromuscular
excitability.” The threshold concentration for irritation in humans exposed
for one minute is 0.25 mg·l
–1
(∼54 ppm).
Following treatment of Swiss male mice with either a single intraperitoneal
dose of 1000 mg·kg
–1
or 2000 mg·kg
–1
NaTFA, hepatocytes were observed to
have a cloudy swelling with slight accumulation of fat accompanied by an in-
crease in liver glycogen at the lower dose. At the higher dose, vacuolization was
also noted (Rosenberg and Wahlstrom, 1971).
At 2000 mg·kg
–1
NaTFA given to mice intraperitoneally, the glucose-6-phos-
phate dehydrogenase activity in livers and erythrocytes was slightly increased
and a transient decrease in glutathione and NADPH content was observed in
the liver at 12 hours after administration. This decrease was also observed in
erythrocytes at 24 hours (Rosenberg, 1971).
The relative low toxicity of TFA is demonstrated by comparison with two
closely related chemicals: trifluoroethanol (TFE) and trifluoroacetaldehyde
(TFAld). The intraperitoneal LD
50
s of TFE, TFAld and NaTFA were 158 to 195,
650 and >2000 mg·kg
–1
, respectively, in male Swiss mice (Airaksinen and Tam-
misto, 1968; Airaksinen et al., 1990). When these three chemicals were each ad-
ministered to 10-week old, male AlpK/AP strain (Wistar-derived) rats with
single intraperitoneal (Lloyd et al., 1986) or single oral (Lloyd et al., 1988) dos-
es of 10 and 25 mg·kg
–1
body weight, TFE and TFAld caused a dose-related re-
duction in testis weight within 3 days, which was accompanied by
morphological changes. In marked contrast, TFA did not cause any observable
testicular effects on weight or morphology.
Using an in vitro Sertoli/germ cell co-culture system obtained from the rat,
at concentrations of 0.1 to 1 mM, TFAld produced dose-related effects includ-
ing increased germ cell loss of pachytene and dividing spermatocytes. The
germ cell losses were accompanied by leakage of the pachytene spermatocyte
marker enzyme lactic acid dehydrogenase-X (LDHX) and decreased lactate
and pyruvate production. With TFA increased cell loss and increased LDHX
leakage became only apparent at 10 mM. No response was observed with
10 mM TFE (Lloyd et al., 1986). In a similar study (Williams, 1997) in isolated
Leydig cell cultures, Sertoli cell only cultures and Sertoli/germ cell co-cultures
obtained from the Sprague-Dawley rat TFAld had also the most wide ranging
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110
and severe effects on the function of isolated testicular cells, affecting all three
cell systems. TFA and TFE show only a small effect on Leydig cell function and
essentially no effect on the Sertoli/germ cell cultures. The apparent difference
of activity of TFE between in vitro and in vivo is explained by the metabolism of
TFE in vivo to TFAld.
TFE, especially when muscular fibrillations were induced by atropine-
neostigmine, caused a drop in the lactate and pyruvate pool in muscle and
liver of mice and guinea pigs, probably by inhibition of glycolysis. In guinea
pig muscle a decrease of ATP was also seen. NaTFA had little, if any, effect
on ATP, lactate and pyruvate. Both NaTFA and TFE had no effect on citrate
values, showing that they do not block the Krebs’ cycle (Airaksinen and
Tammisto, 1968).
In a perfusion experiment with rat liver, Stier et al. (1972) showed that
NaTFA enhanced lactate and pyruvate turnover and uptake without appar-
ent stimulation of the tricarboxylic acid cycle or glucose synthesis.
Single intraperitoneal doses of 2.1 mmol·kg
–1
of TFE and TFAld produced
significant bone marrow and intestinal toxicity, which is characterized by leu-
copenia and loss of intestinal dry weight. This eventually leads to a lethal sep-
ticemia in male Wistar rats. HTFA, when administered at 240 mg·kg
–1
(a dose
equimolar to TFE and TFAld) did not produce any similar toxic effects (Fraser
and Kaminsky, 1988).
Repeated Dose Toxicity
When HTFA or NaTFA were added to drinking water to a concentration
of 114 g·l
–1
(1 N·l
–1
), male Sprague-Dawley rats rejected the solutions which
resulted in dehydration and body weight loss (Blake et al., 1970). As a conse-
quence, an increase (<30%) in the liver-to-body weight ratio was observed
within 10 days. In contrast, daily intragastric administration of 1 ml·kg
–1
·day
–
1
of 1N NaTFA (114 mg·kg
–1
·day
–1
) to rats, for 8 days, did not cause dehydra-
tion or body weight loss and did not significantly affect the liver-to-body
weight ratio. Furthermore, 1 N NaTFA did not affect the duration of hexa-
barbital-induced hypnosis suggesting no change in the metabolic activity of
the liver.
In one other report (Stier et al., 1972), NaTFA was administered via drink-
ing water to male Wistar rats such that the animals were provided with approx-
imately 130 µmol NaTFA per 100 g body weight per 24 hours (equivalent to
150 mg·kg
–1
·day
–1
, based on an assumed normal consumption of water) for 5
to 6 days. After 5 days of treatment, the relative liver weight was increased by
about 43%. Furthermore, the glycogen content of the liver was decreased.
Male Wistar rats were fed trifluoroacetate, perfluorobutyrate, and perfluo-
rooctanoate in their diets for 5 to 14 days at concentrations of 5000 ppm
(500 mg·kg
–1
·day
–1
) for the TFA, and 2500 ppm for the latter two compounds
(Just et al., 1989). All three compounds induced hepatomegaly, anorexia, and
peroxisome proliferation. The relative liver weight increased by 20 to 30%.
With TFA, the body weight gain was similar to controls. Perfluorobutyrate and
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perfluorooctanoate were found to be more active than TFA. The rise in activity
of peroxisomal fatty acid β-oxidation was only slight for TFA, amounting to
about a 2-fold increase, whereas the increase in activity was 9- to 10-fold after
perfluorobutyrate and perfluorooctanoate treatment.
Male rats and male guinea pigs were fed diets containing 7500 ppm (rat:
750 mg·kg
–1
·day
–1
) for 25 days. A decrease of body weight without concomitant
changes in food consumption was observed for rats. Decreases in serum cho-
lesterol, triglycerides, glucose, and insuline levels were reported in rats and to
a lesser extent in guinea pigs. Increased liver weights and diffuse liver hyper-
trophy were observed in treated rats, but in guinea pig the liver weights were
reduced. The rate of hepatic peroxisomal β-oxidation was increased in rats but
not in guinea pigs supporting the conclusion that peroxisome induction is a
rodent-specific phenomena (Warheit, 1993).
Studies on repeated inhalation exposures of animals to HTFA are limited to
one report (Kheilo and Kremneva, 1966). Rats were exposed for 4-hours daily
6 times a week at concentrations of 0.4 to 0.7 mg·l
–1
(86 to 150 ppm) HTFA for
5 months. Effects observed included: irritation of mucous membranes, in-
creased proteinuria, and “altered neuromuscular excitability.” These investiga-
tors noted that chronic exposure to 0.025 to 0.05 mg·l
–1
(∼5 to 11 ppm) caused
only very slight symptoms and suggested that they can be considered to be
close to the threshold concentrations for chronic exposures.
In conclusion, the major target organ in TFA-exposed rats is the liver, show-
ing mild effects (increased weight, hypertrophy and induction of peroxi-
somes) at 150 mg·kg
–1
·day
–1
by gavage for 5 days.
Irritation
Application of solutions of HTFA to the skin of rats or guinea pigs, at con-
centrations of 20% or greater, caused marked coagulation, necrosis; 10% con-
centration was moderately severely irritating and concentrations of from 2 to
5% were moderately irritating (Patty, 1963). Kheilo and Kremneva (1966) re-
ported HTFA to be corrosive to rabbit skin.
Corrosivity
As with any strong acid, severe eye irritation would be expected from con-
tact with HTFA (Patty, 1963). Ocular irritation was observed during the Rus-
sian acute inhalation studies in rats and mice (Kheilo and Kremneva, 1966).
Sensitization
TFA, when conjugated to human and guinea pig albumin, was shown to in-
duce a specific delayed-type hypersensitivity in guinea pigs (Mathieu et al.,
1974). Erythema and induration were observed 24 to 48 hours after intrader-
mal challenge with antigen, as evidence for cutaneous hypersensitivity. Perivas-
cular mononuclear infiltration was observed, histologically. This prompted the
investigators to conclude that TFA may serve as a hapten able to elicit specific
cellular immune reponses.
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Several reports relating to cellular-mediated immune response have ap-
peared in the literature. These have included demonstrations of trifluoro-
acetylated neoantigens such as: TFA-albumin (Reves and McCracken,
1976a,b), TFA-ovalalbumin (Ford et al., 1984), TFA-neoantigen (present in mi-
crosomal carboxylesterase) (Satoh et al., 1984), and trifluoroacetylated hepa-
tocytes (Satoh et al., 1985) that have resulted from exposure to halothane or
TFA.
However, the role of TFA in cell-mediated immunity is questionable be-
cause TFA, as well as halothane, did not stimulate DNA formation in cultured
lymphocytes obtained from the blood of human subjects previously exposed to
halothane (Waldron and Ratra, 1972), and the delayed hypersensitivity re-
sponse to TFA had no effect on halothane-induced liver damage in mice and
rats (Ford et al., 1984). Furthermore, skin test hypersensitivity reactions with
TFA, prepared as a complex with autologous serum protein, did not correlate
with hepatic necrosis induced by halothane exposure in guinea pigs (Reves
and McCracken, 1976b).
Mutagenicity
NaTFA was tested for mutagenicity in the standard Ames assay (Blake et al.,
1981). NaTFA was found to be nonmutagenic to Salmonella typhimurium strains
TA98, TA100, and TA1535, both in the presence and absence of Aroclor 1254-
induced rat liver or rat testes S-9 activation systems. HTFA was tested in anoth-
er laboratory (Waskell, 1978) using Salmonella typhimurium strains TA98 and
TA100, both with and without rat liver S-9 activation, and was also found to be
nonmutagenic. Additionally, HTFA was tested using the repair-deficient
strains TS24, TA2322, and TA1950 of Salmonella typhimurium. HTFA did not in-
hibit the growth of any of the above DNA repair-deficient strains relative to a
normal repair-proficient hisG, thus indicating a lack of genetic activity. In yet
another laboratory (Baden et al., 1976), HTFA or urine from patients anesthe-
tized with halothane were all found to be nonmutagenic to Salmonella typhimu-
rium strains TA98 and TA100, either in the presence or absence of Aroclor
1254-induced rat liver S-9 activating medium.
Carcinogenicity
No information was found with regard to the carcinogenic potential of TFA.
Toxicity for Reproduction
TFA has not been tested for teratogenicity or fetotoxicity.
RISK CHARACTERIZATION
Present and Future Environmental Concentrations
As described in detail in previous sections, the present contributions to en-
vironmental levels of trifluoroacetic acid can be summarized as follows.
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113
Production and Use of Trifluoroacetic Acid
Trifluoroacetic acid is manufactured at about 1000 metric tons per year and
is widely used in the fine chemicals industry and as a laboratory reagent. It is
either consumed or becomes part of a chemical waste stream. The quantities
released into the atmosphere from this source are very small indeed and do
not appear to be a significant contribution to global environmental levels.
Use of Fluorocarbon Anaesthetics
Halothane and isoflurane anaesthetics, which have been in use since the
1970s, both yield trifluoroacetate when degraded. Halothane is the most used
of the fluorocarbon anaesthetics at a maximum of 1500 metric tons per year
globally. All the material used is assumed to be emitted to the atmosphere. The
global deposition rate of trifluoroacetic acid from this source is estimated to
be 800 metric tons per year.
Use of Certain CFC Substitutes
CFCs substitutes — HCFCs and HFCs — break down in the environment to
give predominantly carbon dioxide, water, and inorganic salts of chlorine and
fluorine. However, some substances, most notably HFC-134a, HCFC-124, and
HCFC-123, also break down to yield trifluoroacetic acid (HTFA), which is re-
sistant to further atmospheric degradation. Because the HCFCs and HFCs
have until now been produced only in limited commercial quantities, their
contribution to present environmental levels of TFA has been estimated to be
small — around 2000 metric tons per year.
In total, the present known emission sources of all fluorocarbon precursors
(anaesthetics + CFC substitutes) will yield trifluoroacetic acid at a rate of 2800
metric tons per year.
To derive future environmental concentrations, predicted production and
releases of the relevant HCFCs and HFCs, along with their rates of transforma-
tion into TFA, have been modeled.
Substances that decompose in the atmosphere by oxidation or photolysis
will give rise to diffuse fluxes of decomposition products in the atmosphere as
the first compartment. HCFCs and HFCs that break down to trifluoroacetic
acid have atmospheric lifetimes ranging from about 1.4 to 14.6 years, so that
most decomposition occurs at or near the background concentration of the
substance after it has become well mixed in the atmosphere. The decomposi-
tion mechanisms depend on physical and chemical properties and processes
that are relatively well understood so that it is possible to calculate accurately
the future atmospheric concentrations of precursors to trifluoroacetate, given
a particular set of scenarios for emissions.
It has been calculated (see earlier section for details) that in the year 2020
the global deposition of trifluoroacetic acid as a breakdown product of the
HCFCs and HFCs will be about 160,000 metric tons per year.
Based on the physico-chemical properties of TFA, the preferred environ-
mental compartment will be water, rather than air, ground, or biota. Thus, tri-
fluoroacetic acid that is emitted from processes as a vapor, or trifluoroacetate
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114
that is formed from other materials by reaction in the atmosphere will parti-
tion rapidly into cloud, rain, or surface water.
In rain water a range of concentrations has been calculated, taking into ac-
count geographical variations (OH radical concentrations, amount of rain, re-
gional releases of parent compound). Because most of the rainfall and the
highest regional concentrations of hydroxyl occur in the tropics, the mass dep-
osition rate of trifluoroacetic acid is highest in this region. It was calculated that
in 2020 the tropical concentration would be 2 µmol·m
–2
·yr
–1
or 100 ng·l
–1
in
rainwater. Over Europe, the deposition rate would be about 0.5 µmol·m
–2
·yr
–1
and over the United States levels would be up to double this.
For this risk assessment, a maximum level of TFA in rain water in the region
of 0.1 µg·l
–1
(0.0001 mg·l
–1
) in the year 2020 was adopted as the standard for
the predicted environmental concentration.
In surface water, it appears to be possible to accumulate a riverborne flux of
TFA in lakes or seas from which there is no outflow, only evaporation. How-
ever, consideration of the volumes and flows indicates significant local accu-
mulation over global levels takes many hundreds of years.
A possible buildup in special ecosystems, such as seasonal wetlands (special
aquatic ecosystems that dry out periodically and are replenished only by rain-
fall), has been postulated but significant buildup over background levels can
only be anticipated if a number of low probability events occur at the same
time and endure for a long period of time. Therefore, although some accumu-
lation of TFA may take place in special ecosystems like vernal pools, accumu-
lation over several orders of magnitude appears to be highly improbable.
Generally, soil retention of TFA is poor although soils with high levels of or-
ganic matter have been shown to have a greater affinity for TFA when contrast-
ed with soils with low levels of organic matter. This appears to be an adsorption
phenomenon, not irreversible binding. Based on the maximum observed par-
tition coefficient (K
d
) of 20, the maximum concentration bound to soil — in
equilibrium with the projected rainfall concentration (from CFC alternatives)
in 2020 of 0.0001 mg·l
–1
— would be 0.002 mg·kg
–1
.
Degradation mechanisms of TFA that could mitigate future concentrations
of this compound have been studied quite extensively. Given the high thermo-
dynamical stability of the CF
3
group, no significant abiotic mineralization pro-
cess has been observed or is forecast in environmentally relevant conditions.
TFA is highly resistant to both microbial oxidative and reductive degrada-
tion. Although biodegradation has been observed under specific conditions,
the relevance of these results to the real world are considered to be doubtful.
For the purpose of this risk assessment, we will assume, as a reasonable worst
case, that no degradation of TFA is likely to occur in the environment.
TFA’s low octanol/water partition coefficient (log P
ow
= –2.1) indicates no
potential to bioaccumulate. TFA does not accumulate significantly in lower
aquatic life forms such as bacteria, small invertebrates, oligochaete worms and
some aquatic plants including Lemna gibba (duckweed). In terrestrial higher
plants such as sunflower and wheat, some bioaccumulation was seen (biocon-
centration factors up to 43 based on fresh weight). This appeared to be related
to uptake with water and then concentration due to transpiration water loss.
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In the event that TFA is degraded, it may be transformed into monofluoro-
acetic acid (MFA). However, as the rate of breakdown of MFA (hydrolytically
by fluoroacetate dehalogenases) is so much higher than for TFA, any MFA
formed would rapidly degrade. Therefore, there would be no buildup of MFA
regardless of the levels of TFA present in the environment.
Effect Assessment
Given the fact that deposited TFA would remain in water, a number of
studies have been conducted aiming to derive a no-effect concentration in
the aquatic compartment. Standard acute tests on fish and Daphnia carried
out with NaTFA show that they are insensitive to large concentrations (up to
1 g·l
–1
TFA). This low toxicity of TFA is reinforced by toxicological data on
mammals developed previously, namely, in the context of TFA being a me-
tabolite of several anaesthetics (see earlier section ). On the other hand, the
no-effect concentration for the standard algae species Selenastrum capricornu-
tum is around 0.10 mg·l
–1
(as TFA). To see if this sensitivity was general
among algae, 10 other species belonging to 4 different classes have been test-
ed; these other species were not sensitive to NaTFA (no-effect concentration
>100 mg·l
–1
). Even among the chlorophyceae, Selenastrum capricornutum
seems to be unique; therefore, even if this species was affected above
0.1 mg·l
–1
, the ecology of the system (i.e., the functionality) where it lives
would not be affected. Consequently, it is considered that for the protection
of the aquatic environment the no-effect concentration of 0.1 mg·l
–1
of this
alga can be used.
Extensive research has been devoted to effects on higher plants, as they
could be exposed to TFA in rainwater through leaves and stems and TFA in
pore water through roots. A number of species were tested, with particular
emphasis on those having an important role in feeding people and cattle. Ex-
posure of the leaves was shown to be less important than exposure of the
roots. By application to the soil, the most sensitive species was sunflower, with
an effects threshold of 1 mg NaTFA·kg
–1
(dry soil). The same soil concentra-
tion had no effect on the germination, growth and nitrogen-fixing capacity
of soya seedlings. However, because of the limited binding of TFA to soil, ef-
fects are better defined as concentrations in the soil water. By exposure of
the roots to aqueous solutions, the no-effect concentration was 1 mg NaT-
FA·l
–1
for the long-term growth of wheat and soya. Therefore, as a conserva-
tive figure, the no-effect concentration in soil-water is considered to be 0.1
mg·l
–1
.
Risk Assessment
In comparing the present and future environmental concentrations of TFA
(maximum of 0.0001 mg·l
–1
in rainwater in 2020) with the no-effect concentra-
tions in both surface water and soil (0.1 mg·l
–1
), it is concluded that no notice-
able risk for the environment can be anticipated as there is a 1000-fold
difference. The local enrichments in certain aquatic systems and organic rich
soils do not influence this conclusion.
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For the time being, an important question remains concerning the origin
of the large present levels of TFA that have been measured in the environment
(fresh and marine surface waters, rain, and air) and cannot be explained by
the known industrial sources. These levels, in the range of 0.1 µg·l
–1
for water
and 50 ng·l
–1
for air, are roughly 60 times what is estimated to arise from known
sources today and more or less equal to what is anticipated in 25 years from
now.
ACKNOWLEDGMENTS
The authors would like to thank the principal investigators who contributed
to the AFEAS TFA research program: A.G. Berends, A.H.C. Groeneveld and
N.R.M. van Dijk (Solvay Duphar), Thomas Bott and Laurel Standley (Stroud
Water Research Center), Craig Criddle (Michigan State University), Alan Davi-
son (University of Newcastle), Charles Driscoll (Syracuse University), David
Emerich (University of Missouri), Hartmut Frank (University of Bayreuth),
Anders Grimvall (Linköping University), Malcolm Ko (Atmospheric and Envi-
ronmental Research, Inc.), Gene Likens (Institute of Ecosystem Studies), Ron
Oremland (U.S. Geological Survey), Steve Schwarzbach (U.S. Fish and Wild-
life Service), James Seiber (University of Nevada – Reno), and C.G. van Ginkel
(formerly Akzo). In addition, we wish to acknowledge Karlheinz Ballschmiter
(University of Ulm), Gérard Blake (University of Savoie) and Paul Falkowski
(Brookhaven National Laboratory), who served as advisors or reviewers.
The TFA research program was sponsored by the Alternative Fluorocarbons
Environmental Acceptability Study (AFEAS), a consortium of international
chemical manufacturers. AFEAS members include: AlliedSignal, Inc. (USA),
Asahi Glass Co., Ltd. (Japan), Ausimont S.p.A. (Italy), Daikin Industries, Ltd.
(Japan), E.I. DuPont de Nemours & Co., Inc. (USA), Elf Atochem S.A.
(France), Hoechst AG (Germany), ICI Chemicals & Polymers Ltd. (UK),
LaRoche Industries Inc. (USA), Rhodia Ltd. (UK), and Solvay S.A. (Belgium).
The program is managed by RAND Environmental Science & Policy Center
(Washington, DC).
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