PESTICIDE RESIDUES IN FOOD - 1997
Sponsored jointly by FAO and WHO
with the support of the International Programme
on Chemical Safety (IPCS)
TOXICOLOGICAL AND ENVIRONMENTAL
EVALUATIONS 1994 1997
Joint meeting of the
FAO Panel of Experts on Pesticide Residues
in Food and the Environment
and the
WHO Core Assessment Group
Lyon 22 September - 1 October 1997
The summaries and evaluations contained in this book are, in most
cases, based on unpublished proprietary data submitted for the purpose
of the JMPR assessment. A registration authority should not grant a
registration on the basis of an evaluation unless it has first
received authorization for such use from the owner who submitted the
data for JMPR review or has received the data on which the summaries
are based, either from the owner of the data or from a second party
that has obtained permission from the owner of the data for this
purpose.
2,4-Dichlorophenoxyacetic acid (2,4-D), salts and esters
First draft prepared by
Dr P.J. Campbell
Ministry of Agriculture, Fisheries, and Food
York, United Kingdom
Environmental transport, distribution, and transformation
Volatilization
Water
Hydrolysis
Photolysis
Biodegradation
Bioaccumulation and biomagnification
Soil
Hydrolysis
Photolysis
Adsorption and desorption
Mobility and leaching
Degradation in soil under laboratory conditions
Degradation in soil under field conditions
Uptake by plants
Environmental levels
Air
Water
Soil
Plants
Effects on organisms in the laboratory and the field
Microorganisms
Aquatic organisms
Plants
Toxicity
Other effects on plants
Invertebrates
Toxicity
Other effects on invertebrates
Vertebrates
Toxicity
Other effects on vertebrates
Terrestrial organisms
Plants
Invertebrates
Toxicity to arthropods
Toxicity to earthworms
Other effects on invertebrates
Vertebrates
Toxicity to birds
Toxicity to birds' eggs
Effects on mammals
Effects on amphibia
Risk assessment based on agricultural use
Microorganisms
Aquatic organisms
Acute risk to freshwater pelagic organisms
Long-term risk to freshwater pelagic organisms
Risk to sediment-dwelling invertebrates
Risk to amphibia
Bioaccumulation
Terrestrial organisms
Plants
Invertebrates
Vertebrates
Birds
Mammals
Evaluation of effects on the environment
Risk assessment
Aquatic environment
Terrestrial environment
References
1. Environmental transport, distribution, and transformation
1.1 Volatilization
Certain ester formulations (butyl, ethyl, isopropyl) of 2,4-D are more
volatile than amine salt formulations, such that vapour drift can be
virtually eliminated by use of salts of 2,4-D (Que Hee & Sutherland,
1974), and use of the more volatile esters is being discontinued in
most countries owing to potential drift of either droplets or vapour.
Use of the highly volatile esters may result in drifting of as much as
25-30% of the applied 2,4-D off target crops (Grover et al., 1972;
Maas & Kerssen, 1973; Grover, 1974, 1976; Maybank et al., 1978).
In an evaluation of the use of silica gel and XAD-4 as adsorbents for
the removal of the n-butyl and isooctyl esters of 2,4-D, their
overall efficiencies for both trapping and ease of extraction were 85-
100% (Grover & Kerr, 1978). The amount of the dimethylamine salt of
2,4-D that volatilized after 48 h at 38°C under light emitted by a
100-watt incandescent bulb was < 10%, irrespective of the surface
(Que Hee & Sutherland, 1974; Que Hee et al., 1975).
The volatilization of 14C-2,4-D ethylhexyl ester (as Esteron 99TM
concentrate) under laboratory conditions, according to the guidelines
of the US Environmental Protection Agency, was characterized by
applying it to Galstown sandy loam at an equivalent rate of 13.5-17.7
lb acid equivalent per acre (15-20 kg/ha). The volatility was
maintained by passing air over the soil at 25°C and 95-100% relative
humidity at air flow rates of 100-300 ml/min for 15 days; the pressure
in the chamber was held at 1 atm (101.3 kPa). The volatile gases were
then trapped. The losses from the soil were only 0.05-0.23%.
Volatilization was most rapid immediately after application, declined
with time up to day 7, and then remained essentially constant up to
day 15. The volatility increased directly with the air flow rate
(Doyle, 1991).
In an extensive review of the literature on the volatility of salts
and esters of 2,4-D, Que Hee and Sutherland (1981) concluded that the
effective partial pressure for deposits on plants, in water, and on
soils may be different from that expected in comparison with glass
surfaces. The most important factor in assessing the risk for vapour
drift is the volatility of the compound itself. Thus, the predicted
rates of loss from 1 ha were 0.2 kg/h for the n-butyl ester of
2,4-D and 0.25 kg/day for the isooctyl ester (Elliott & Wilson, 1983).
In a test to rate the volatility of the methyl, ethyl, isobutyl,
butyl, 2-butoxyethyl, and 2-ethylhexyl esters of 2,4-D in tomato
plants, the esters and formulations with a vapour pressure > 3.3 mPa
(at 25°C) were rated as highly volatile while those with a vapour
pressure <: 0.6 mPa were considered to have low volatility (Noble &
Hamilton, 1990).
In a study of the rate of dissipation of the propylene glycolbutyl
ether ester of 2,4-D and vapour loss in air from soil in chambers at
temperatures up to 35°C, 87-97% had dissipated by degradation and
0.1-2.6% by volatilization in air at all temperatures after 154 days
(Nash, 1989a). An empirical model was used to estimate dissipation
when volatilization is the predominant pathway for loss (Nash, 1989b).
1.2 Water
1.2.1 Hydrolysis
A comparison of the kinetics of hydrolysis and the half-lives for many
salts and esters of 2,4-D, based on an extensive review of the
literature, showed that the hydrocarbon esters hydrolyse much more
slowly than the alkyl ether esters, with the exception of the methyl
ester (Que Hee & Sutherland, 1981). In a comparison of the hydrolysis
of the methyl and butoxyethyl esters and the butyl and octyl esters of
2,4-D in aqueous solution, esters with ether linkages near the carboxy
group were generally hydrolysed more rapidly than the hydrocarbon-
chain esters. The rapid rates observed in basic water, considerably
longer than those in acidic or neutral water, suggest that hydrolysis
is the major pathway for degradation of 2,4-D esters in natural water.
The half-lives for hydrolysis of 2,4-D esters in aqueous solution at
28°C and pH 9 were from a low of 0.6 h for the 2-butoxyethyl ester to
37 h for the 2-octyl ester (Zepp et al., 1975, 1976).
In studies conducted according to the guidelines of the US
Environmental Protection Agency, 14C-2,4-D acid (at 10-4mol/L) was
not hydrolysed at pH 5, 7, or 9 after 30 days at 24.9°C in sterile
aqueous solution in the dark (Creeger, 1989a). The hydrolysis of
14C-2,4-D ethylhexyl ester in sterile aqueous solutions containing 1%
acetonitrile, in the dark, at 25°C for up to 30 days was slow at pH 5
and moderate at pH 7. At the end of the study, 77.7 % of the applied
radiolabel was present as the parent ester at pH 5, while only 59.3%
remained at pH 7. At pH 9, 2,4-D ester degraded rapidly and
represented 14.2% of the applied dose after 144 h. The half-life of
14C-2,4-D ethylhexyl ester at pH 5 was found by extrapolation to be
99.7 days, with a half-life at pH 7 calculated to be 48.3 days; the
half-life at pH 9 was calculated to be 52.2 h. 14C-2,4-D represented
81.4% of the applied radiolabel at the end of the study (Concha et
al., 1993a). Hydrolysis of 14C-2,4-D ethylhexyl ester was evaluated
in natural river water containing 1% acetonitrile, for up to 24 h in
the dark at 25°C. The ester degraded rapidly, only 7.2% of the applied
radiolabel remaining as parent compound after 24 h. 14C-2,4-D
represented 93.8% of the applied dose. The half-life was calculated to
be 6.2 h (Concha et al., 1993b).
1.2.2 Photolysis
The final oxidation product of 2,4-D in aqueous solutions irradiated
with a polychromatic lamp was carbon dioxide (Boval & Smith, 1973).
The kinetics of photodegradation and the photolysis products of 2,4-D
at various pH values have been described after photolysis by
ultraviolet radiation (Chamarro & Esplugas, 1993) and ultraviolet
radiation with ozone (Prado et al., 1994). The photoproducts of the
ethyl, butyl, and 2-methylheptyl esters of 2,4-D under Pyrex-filtered
light included hydrochloric acid and esters of 2- and 4-
chlorophenoxyacetic acid (Binkley & Oakes, 1974). The photoproducts of
the butyl ester of 2,4-D under ultraviolet light included 2,4-
dichloro-phenol, the methyl ester of 2,4-D, the n-butyl ester of
5-chloro-2-hydroxyphenylacetic acid, and hydrochloric acid (Que Hee &
Sutherland, 1979).
Two studies were conducted according to the guidelines of the US
Environmental Protection Agency. In one, 14C-2,4-D was irradiated for
30 days under simulated sunlight (one-half day light, one-half day
dark) at 24.8°C in a filter-sterilized aqueous solution buffered at pH
7. The decay was first-order over more than two half-lives, with a
half-life of 12.98 days (7.57 days of constant exposure to light).
Only 1,2,4-benzenetriol was present at > 10% of the initial
concentration of 14C-2,4-D. Production of 14C-carbon dioxide had
reached 25% by the end of the study. No photodegradation occurred in
the dark (Creeger, 1989b). In the second study, 14C-2,4-D ethylhexyl
ester was irradiated with natural sunlight for up to 31 days in
aqueous solution at 25.3°C and pH 5; control samples were tested
concurrently in the dark. The ester degraded slowly when exposed to
natural sunlight, accounting for 80.3% of the applied dose after 31
days of exposure, with 84.8% in the controls. The three major
degradates were 2,4-D (5.0%), 2,4-dichlorophenol (2.1%), and
2-ethylhexyl-4-chlorophenoxyacetate (1.1%). The ester degraded more
slowly in the dark, only 6.5% 2,4-D being observed after 31 days. The
half-life for degradation of 14C-2,4-D ethylhexyl ester was
calculated to be 128.2 days in light and 252.5 days in the dark
(Concha & Shepler, 1993a). The same photoproducts were reported by
Zepp et al. (1975). The quantum yields for photoalteration and the
photolysis half-lives were also reported for several other esters of
2,4-D (Zepp et al., 1976).
In a study of the photodynamics of the methyl and butoxyethyl esters
of 2,4-D in surface waters, the major photoproducts at concentrations
exceeding the solubilities of the esters in water (> 300 ppm) were
the dehalogenated 2-and 4-chlorophenoxy acetic acid esters (Zepp et
al., 1975). No significant phototransformation of 2,4-D was seen in
water after irradiation for up to 168 h at 304 nm. The upper limit of
the quantum efficiency, F, was found to be < 0.014 (Klopffer,
1991).
After irradiation with ultraviolet light of the pure n-butyl, mixed
n-butyl, and isobutyl esters of 2,4-D in aqueous and hydrocarbon
solutions, reductive dechlorination was preferential at the ortho
position at 300 nm, irrespective of the solvent, wheres
photodecomposition was negligible at 350 nm (Que Hee et al., 1979).
Addition of 0.85% (v/v) acetophenone to aqueous solutions of 2,4-D
containing 0.15% (v/v) oxysorbic strongly sensitized photodegradation
of 2,4-D (Harrison & Wax, 1985). A review of the photodecomposition of
2,4-D includes the photoproducts of 2,4-D irradiated in alcohol, the
rates of degradation of 2,4-D esters irradiated in different solvents,
and a discussion of the photodegradation factors for esters and salts
(Que Hee & Sutherland, 1981).
1.2.3 Biodegradation
The dissociation of 2,4-D and its dimethylamine salt in an aqueous
solution were studied by conductometry according to the guidelines of
the US Environmental Protection Agency. 2,4-D dimethylamine
dissociated within 1 min when added to stirred water, whereas 2,4-D
acid dissociated within > 120 min. These results confirm the
complete, rapid dissociation of 2,4-D dimethylamine in aqueous
solution, forming dimethylammonium ion and the conjugate base of 2,4-D
(Reim, 1989).
In a study of anaerobic aquatic metabolism conducted according to the
guidelines of the US Environmental Protection Agency, 14C-2,4-D was
tested in viable pond sediment and water at 25.1°C for up to one year.
Most of the applied radiolabel (24.1-85.5%) was found in the aqueous
phase and a moderate amount (11.2-42.3%) in organic sediment. The
amount of bound radiolabel increased with increasing degradation, so
that after 240 days 34.7% of the applied radiolabel remained
unextracted. Partitioning of the extracted residue on day 240 showed
2.4% in the humic acid fraction and 14.9% in the fulvic acid fraction,
23.5% remaining unextracted. The level of 14C-carbon dioxide reached
22.1% after 365 days. 14C-2,4-D accounted for 25.9% of the applied
dose in the aqueous phase and 13.2% in the sediment after one year.
The two maj or metabolites were 2,4-dichlorophenol, accounting for
21.6% at day 30 and 4.2% after one year, and carbon dioxide. The
volatile compounds seen after one year were 4-chlorophenol (1.9%),
2,4-dichloroanisole (0.7%), and 2,4-dichlorophenol (0.7%). The
half-life for anaerobic degradation of 14C-2,4-D in water was
calculated to be 312 days (Concha & Shepler, 1994a).
In a similar study of aerobic metabolism, conducted at 25°C for up to
46 days, 14C-2,4-D degraded slowly for the first 25 days,
representing > 75% of the applied dose, but degraded rapidly within
the next 10 days, accounting for 0.5% of the applied dose at the end
of the study (Concha & Shepler, 1993b). An initial lag in the
degradation rate had been observed previously (McCall et al., 1981)
and was postulated to be the lag time necessary to degrade 2,4-D, to
mutate in order to produce this specific enzyme (Loos, 1975), or to
allow enough time for a small microbial population to degrade enough
2,4-D to be detectable (Chen & Alexander, 1989). In a plot of time of
exposure against percent 14C-2,4-D, the concentration reached 50% at
about 29 days after application. The half-life was determined to be
4.5 days. The major metabolite under aerobic conditions was carbon
dioxide, which comprised 63.9% of the applied radiolabel at the end of
the study. The three metabolites found were 2,4-dichlorophenol
(representing 1.1% of the applied dose on day 35 and 0.1% at the end
of the study), 4-chlorophenoxyacetic acid (1.1% on day 14 but
insignificant amounts at the end of the study), and 4-chlorophenol
(1.4% radiolabel after 20 days and undetectable at the end of the
study). The unextractable residue increased with time and comprised
15.6% of the applied dose at the end of the study. During the lag time
of 25 days, > 64% of the radiolabel was found in the aqueous phase.
About 10-14% could be extracted from the sediment with an alkaline
solvent, and about 4% could be extracted with acidic acetone. After
the lag time, the distribution of radiolabel shifted markedly as the
2,4-D acid degraded, and by the end of the study, 3% of the applied
dose was in the aqueous phase, 1% in basic solvents, and 0.6% in
acidic solvents (Concha & Shepler, 1993a).
Three trials were conducted in the United States according to the
guidelines of the US Environmental Protection Agency. In Louisiana,
the formulated dimethylamine salt of 2,4-D was applied aerially to
rice in the green-ring stage of growth with a canopy height of 25-30
in (64-76 cm) at a target application rate of 1.5 lb/acre (1.68
kg/ha). The concentration of 2,4-D acid residue in water was highest
on day 0 (mean, 1.372 ppm) and then declined to a mean of 0.194 ppm
three days after application. No 2,4-D acid was detected (limit of
quantification, 0.01 ppm) in water samples on day 7, 15, or 30 of
sampling. The half-life in water was calculated to be 1.1 days.
Residues of 2,4-dichlorophenol and 4-chlorophenoxyacetic acid were
detected in samples taken on days 0, 1, and 3, but were below the
limit of quantification (Barney, 1994). Two further trials were
conducted during 1994 in small ponds in North Carolina and North
Dakota, with two applications of the dimethylamine salt (Hatfield,
1995a,b). In North Carolina, a subsurface application of 40.38 lb/acre
(45 kg/ha) was followed by an application of 45.01 lb/acre (50 kg/ha)
30 days later. In North Dakota, 41.8 lb/acre (47 kg/ha) were applied
twice. In North Carolina, 2,4-dichlorophenol and 2,4-dichloroanisole
were found immediately after the first application, but no residues
were detected 21 days after the second application. The half-lives in
pond water were 19.7 days after the first application and 2.7 days
after the second. In North Dakota, 2,4-dichlorophenol and
4-chlorophenol were detected in water samples immediately after the
first application, but none were detected 60 days after the last
application. The half-lives were 13.9 days after the first application
and 6.5 days after the second.
Mineralization of 14C-2,4-D by microbial communities in two streams
in southwestern Ohio, USA, was very limited (Palmisano et al., 1991).
In an investigation of the degradation rate of 2,4-D in river water in
relation to the nutrient levels, sediment load, and dissolved organic
carbon content of the water, the limiting factor was found to be not
the numbers of organisms capable of degrading 2,4-D but rather the
nutrient status of the river (Nesbitt & Watson, 1980a).
The dissipation of the isooctyl ester of 2,4-D (60% acid equivalent)
at nominal rates of 1.0 or 2.5 kg/ha was measured in outdoor
enclosures constructed in a typical bog lake in sandy soil in
northeastern Ontario, Canada. The rate of dissipation of 50% of the
initial concentration of 2,4-D (DT50) from the lake water was 4.5-7.8
days, depending on the application rate. Within 15 days, < 5% of the
applied dose remained in the water, but up to 25% was adsorbed on the
sides of the enclosure (Solomon et al., 1988). The average half-lives
of 2,4-D in samples of groundwater from three locations in eastern
Arkansas, USA, were > 800 days, although the absence of aquifer
materials may have resulted in unrealistically slow degradation rates
(Cavalier et al., 1991).
1.2.4 Bioaccumulation and biomagnification
The dissipation of residues of 2,4-D w as evaluated after application
to ponds in Missouri, Georgia, and Florida, USA. When the
dimethylamine salt was applied at 2.24, 4.48, or 8.96 kg/ha, the
residues in water were negligible (< 5 µg/L) within two weeks, and
those in silt and fish were < 200 µg/kg. Of 307 fish samples
analysed, only 45 contained detectable residues. The highest level in
pond water was 692 µg/L three days after application of 8.96 kg/ha,
the highest in silt was 170 µg/kg three days after application of 8.86
kg/ha, and the highest in fish was 102 µg/kg four days after
application of 8.86 kg/ha (Schultz & Harman, 1974; Schultz & Gangstad,
1976). In a study of the fate of the butoxyethanol ester of 2,4-D
applied at 23 kg ai/ha, as granules, to outdoor, artificial,
polyethylene ponds infested with Eurasian watermilfoil, the water
temperature dropped from 25 to 0°C after 56 days. The levels of the
ester were low and fell to < 1.0 µg/L within 15 days, the rapid
disappearance being attributed to hydrolysis in the alkaline pond
water, which had a pH of 9.5 at the time of initiation, the low
aqueous solubility of the ester, its uptake by the water weeds, and to
its sorption to the plastic lining of the tanks (Birmingham & Colman,
1985).
Four 10-ha areas of Lake Seminole, Georgia, USA, with dense beds of
water milfoil were treated with either 2,4-D dimethylamine salt at
22.5 kg/ha or 2,4-D butoxyethanol ester at 45 kg/ha. Both formulations
were converted to free 2,4-D acid within 24 h after application. The
maximum water concentrations found were 3.6 mg/L for the dimethylamine
salt and 0.68 mg/L for the butoxyethanol ester. No fish sampled 13
days after application contained detectable levels of 2,4-D (Hoeppel &
Westerdahl, 1983).
After application of the dodecyl-tetradecyl amine salt of 2,4-D at 4.5
kg/ha followed by spot treatments of this amine and/or dimethylamine
salt to 7000 acres of canal in Florida, USA, the highest level of
2,4-D found in water was 27 µg/L after one day and the highest level
in silt was 5 µg/kg after 3-15 days. Three of 60 fish analysed
contained > 10 µg/kg 2,4-D, while 16 of 60 had levels < 10 µg/kg
(Schultz & Whitney, 1974).
The mean residues of 2,4-D in 353 individual fish of eight species
from lakes in Canada after treatment with 2,4-D in 1977-80 ranged from
< 5 to 60 µg/kg (Frank et al., 1987).
1.3 Soil
1.3.1 Hydrolysis
No reliable data exist on the hydrolysis of salts or esters in sterile
soils.
1.3.2 Photolysis
In a study conducted according to the guidelines of the US
Environmental Protection Agency, 14C-2,4-D was irradiated for 30 days
under simulated sunlight (one-half day light, one-half day dark) at
24.9°C on air-dried, sieved, autoclave-sterilized loam. Aliquots were
also treated in the dark. 2,4-D was not photodegraded substantially.
The photolysis half-life was found by extrapolation to be 68 days. The
levels of metabolites other than 2,4-D found in the were not > 10%
of the initially applied dose. 14C-Carbon dioxide production had
reached 5.05% by the end of the study (Creeger, 1989c).
1.3.3 Adsorption and desorption
Organic matter, soil pH, and exchangeable aluminium are the major
factors that determine adsorption of 2,4-D acid and its dimethylamine
salt in soils (Liu & Cibes-Viade, 1973; Grover & Smith, 1974). In a
study of the sorption of [1-14C]-2,4-D on 42 samples of topsoil and
subsoil from 21 sites in Belgium (pH, 3.25-6.91), most sorption
occurred on the soil with the lowest pH (3.25) and least on a subsoil
with pH of 6.4. The soil adsorption coefficients based on organic
carbon (KOC) were 31.2-470.9 for 18 of the soils. After application
of 2.0 kg/ha of 14C-2,4-D, about 70% of the initial amount was
recovered in the first 5 cm of a loamy sand and 1.4% at 10-15 cm after
48 mm of percolating water, whereas in a Zolder sand 73% of the 2,4-D
was found at 10-15 cm. The mobility of 2,4-D was inversely related to
the amount adsorbed. After application of 2,4-D at 4.4 kg/ha to a 1-m
Podzol soil column (Zolder sand) with 800 mm of 0.025 mol/L CaCl2, no
2,4-D was found in the soil solution at the bottom of the column. The
maximum depth of penetration was 50 cm, and only 7% of the applied
2,4-D was recovered in the entire soil profile after 62 days (Moreale
& Van Bladel, 1980). These findings are consistent with those of
Grover (1977), who found that the adsorption of 2,4-D correlated with
the organic content of seven Canadian soils and not with the clay
content. When the rate constants, activation energy, heats of
activation, and entropies of activation for the adsorption of
analytically pure 2,4-D on humic acid from a black chernozem soil from
Canada were calculated, the adsorption was found to follow the
Freundlich-type isotherm. The rate of adsorption of 2,4-D on humic
acid at both 5 and 25°C was initially rapid but was slower
subsequently (Khan, 1973, 1975a).
The Freundlich coefficients for the sorption of 2,4-D on seven
Canadian prairie soils ranged from 0.09 to 1.3, with KOC values of
8.7-21. Sorption of 1-butyl and 2-octyl esters could not be measured
because of rapid hydrolysis but is apparently similar to that of 2,4-D
acid (Grover, 1973). Significantly more 2,4-D was adsorbed by the top
0-5 cm than in the lower layers, which correlates well with the
organic content of silt loam. The Freundlich sorption coefficients
were 0.67 µg/g of soil (or 20.9 µg/g of organic matter) for k and
0.82 for 1/n. As the adsorption-desorption process is the main
mechanism that affects the availability, mobility, and degradation of
the herbicide in soils, it is not surprising that 2,4-D binds more
strongly to soils with a high content of organic matter than to those
with a low content. In a study of the degradation of [carboxyl-14C]-2
4-D and the dimethylamine and isooctyl ester formulations in runoff,
over 80% of the applied compound decomposed to 14C-carbon dioxide
within five weeks, during the active degradation phase, but only 3%
more decomposed during an additional five weeks. An average of 0.26%
of the applied radiolabel could be identified as 2,4-D, with 0.05% in
sediment and 0.21% in water. The amount of 2,4-D applied was thus
affected by the concentration in the runoff (Wilson & Cheng, 1978).
The adsorption properties of 14C-radiolabelled 2,4-D and its
2-ethylhexyl and butyl esters were evaluated at 24°C in a silt loam, a
sandy loam, a loam, and a clay (organic carbon, 0.22-3.08; pH 5.9-7.5)
at initial concentrations of 0.2-5.0 mg/L. The supernatant was by
analysed by radio-high-performance liquid chromatography (HPLC). No
degradation of 2,4-D was observed. The soil adsorption coefficient
(Kd) was 0.08-1.11 ml/g, with an average of 0.78 ml/g. The KOC was
34-79 ml/g, with an average of 48 ml/g. The KOC calculated from the
Freundlich equation was 31-74 ml/g, with an average of 45 ml/g. The
weakest sorption was that in a sandy loam, which had the highest pH
(7.5) and the lowest organic carbon content (0.22%). Sorption of the
esters was not measured owing to rapid hydrolysis (half-life = 79 min
for the 2-ethylhexyl ester and 26 min for the butyl ester) (McCoy &
Lehmann, 1988).
The role of chemical structure, the nature of soil constituents, and
physico-chemical factors such as pH were analysed to evaluate the
adsorption of 2,4-D in soils, expressed as Freundlich isotherms. Nine
soils with organic contents varying from 0.9 to 14.1% were
investigated. The sorption (KOC) varied from 21-33 in a calcic
cambisol to 196-767 in an acidic ferralsol. The Kd values for 2,4-D
ranged from 0.3 for the calcic cambisol to 26.6 for an andosol
(Barriuso & Calvet, 1992). The adsorption of 2,4-D was examined in two
Brazilian oxysols, with plots under natural vegetation and with plots
that had borne crops for up to 56 years after clearing. The adsorption
of 2,4-D was always greater in soils with natural vegetation and
increased strongly with decreasing soil pH (Barriuso et al., 1992).
Sorption coefficients were developed in a study of degradation in
silty clay loam, silty clay, loamy sand, and a clay. The soil-water
distribution ranged from 0 in clay to 9.05 mg/L in the silty clay
(Ogram et al., 1985). 2,4-D was not removed from domestic water by
activated sludge treatment with conventional methods for treating
potable water (Hill et al., 1986).
The adsorption of 2,4-D by 19 soil-sediment materials, in which the
clay, sand, silt, and organic matter contents varied greatly, was
investigated with a batch equilibrium technique. The organic content
was the single most important factor. The adsorption coefficient (K)
ranged from a low of 0.38 to a high of 39.1 after 24-h equilibrium,
and the KOC ranged from 9 to 330 for 2-h adsorption and from 40 to
415 for 24-h adsorption (Reddy & Gambrell, 1987). The soil adsorption
coefficient (log KOC) based on HPLC capacity factors was 2.59, which
compared well with values in the literature ranging from 1.70 to 2.73
(Hodson & Williams, 1988). The adsorption coefficient K for three
soils, determined by HPLC, was 0.6-4.3 and the KOC was 112-145, in
comparison with the KOC in model adsorbents for 2,4-D based on OECD
guidelines, which was 20 (Rippen et al., 1982). In a similar study of
10 Danish soils, adsorption correlated significantly with the organic
content of the soils (Lokke, 1984).
In batch and column experiments with a sandy loam under saturated and
unsaturated conditions, the sorption of 2,4-D had a slight but
significant effect on its transport under either saturated
(retardation factor, 1.8) or unsaturated conditions (retardation
factor, 3.4). Biodegradation was extensive. In batch experiments,
2,4-D (100 mg/kg) was completely mineralized under either saturated or
unsaturated conditions over a four-day period after a three-day lag
phase (Estrella et al., 1993). In a model of the competitive
adsorption and desorption of 2,4-D on a volcanic soil with a high
organic content (8.7%), the agreement with experimental data was
excellent (Susarla et al., 1992).
Freundlich constants and equilibrium desorption values were determined
for the adsorption of 2,4-D and biodegradation of the adsorbed 2,4-D
in soil columns containing a New Zealand silt loam soil. The
equilibrium desorption values ranged from 80% at a soil concentration
of 1% (w/v) to 67.8% at a concentration of 50% (Bhamidimarri & Petrie,
1992). In a study of the effects of long-term conventional and
long-term low-input farming on the adsorption of 14C-2,4-D and the
soil properties that control adsorption at various slope positions,
samples were characterized for soil organic carbon and clay content,
soil pH, and linear adsorption partition coefficients (K d) The K d
values were higher for the low-input farms ( 1.27-0.54, bottom to top
slope positon) than for the conventional farms (0.77-0.31)
(Mallawatantri & Mulla, 1992).
In a study conducted according to the guidelines of the US
Environmental Protction Agency, the adsorption and desorption of
unaged 14C-2,4-D was evaluated in non-sterilized Louisiana rice paddy
sediment (clay soil), at concentrations of 0.10, 0.51, 1.00, 2.47, and
5.02 ppm at 22°C. Partitioning between the sediment and 0.01 mol/L
CaCl2 was determined. The Kd value for was 1.22 (KOC = 58.1), which
suggests that the acid was adsorbed onto sediment from the water
within 24 h. In the desorptive phase, the K1value was 1.64 (K1OC=
78.1), which suggests that the acid is moderately to highly mobile in
rice paddy sediment (Cohen, 1991).
It has been suggested that subsoil horizons be included in models
designed to predict the mobility of herbicides in soils. The sorption
distribution coefficient (Kdm) for surface and subsoil horizons was
determined for 14C-2,4-D in six Atlantic coastal plain soils in the
USA. The mean distribution coefficient, averaged over all soils and
horizons, was 0.65 L/kg. A model was developed involving only organic
matter and exchangeable acidity, which predicted the sorption of 2,4-D
(R2 = 0.68) (Johnson & Sims, 1993).
Supercritical carbon dioxide, with and without methanol as a modifier,
was used to extract bound residues of 2,4-D from soil, plant, and
grain samples. Three months after application of 14C-2,4-D, at .226
µCi/g (10ppm) to soil, 14% was found to be bound in the soil. The
recovery of 2,4-D on mineral soil was only 57.4%. After addition of
methanol to the process, the recovery improved significantly, to 80.7%
(Khan, 1995). 1995).
1.3.4 Mobility and leaching
In a study conducted according to the guidelines of the US
Environmental Protection Agency, the mobility of soil-aged residues of
14C-2,4-D was determined in a sandy soil. The soil was first treated
with 1.4 mg/kg 14C-2,4-D and aged aerobically for 28 days at 20°C, at
a 263 moisture content of 75% field capacity. After 28 days, an
aliquot of the treated soil was placed on 28 cm of untreated soil of
the same type in a column. In order to evaluate the leaching of 2,4-D
in this soil, it was percolated with 0.01 N CaCL. After 28 days, 69.1%
of the applied radioactivity was released as 14C-carbon dioxide,
5.39% being extractable. Of the extractable residue, 23.9% was
recovered as 2,4-D. Fully 98.8% of the radiolabel applied to the top
of the column was retained after addition of the CaCl2, 96.73% being
retained in the top 5 cm. Analysis of the radiolabel in the top 5 cm
showed that 8.14% was on 2,4-D and 72.98% on unidentified, highly
polar metabolites (Zohner, 1990a). In a similar study, conducted
according to a different guideline, the same soil was treated with
14C-2,4-D at a rate of 2.19 mg/kg and aged aerobically for 30 days at
20°C, when 100 gm dry weight of the aged soil was placed on top of an
untreated 28-cm column of the same soil type. It was then percolated
with distilled water. After 30 days, the total radiolabel in the soil
had decreased to 16.27% of the applied dose, of which 4.08% was
extractable. 14C-Carbon dioxide released during aging represented
71.57% of the applied radiolabel; and the extractable residues were
unidentified, highly polar metabolites. A total of 2.15% of the
radiolabel applied to the column was recovered in the leachate, of
which 0.015 µg was identified as 2,4-D (Zohner, 1990b).
The mobility and degradation of 14C-2,4-D dimethylamine salt on a
sandy soil was evaluated in two lysimeters for two years. A total of
2025 mm water was added to the soil during this period from rainfall
and irrigation. The results were consistent with those of studies in
the laboratory, indicating rapid degradation of 2,4-D. Extremely small
amounts of residues of 14C-2,4-D (0.133 and 0.070 µg/L of parent
equivalents) were found. Apart from 14C-carbon dioxide, no typical
soil metabolites were detected in the leachate. Only 0.26-0.29% of the
applied radiolabel was found on 2,4-D in extracts from the upper 17 cm
of soil. Thus, although some sorption experiments in the laboratory
indicate that 2,4-D is mobile, its rapid degradation in soil prevents
significant downward movement under normal agricultural conditions
(Burgener, 1993).
In studies of the mobility of 2,4-D on thin-layer plates and its
dissipation in tobacco soils in Ontario, Canada, 2,4-D moved with the
water front 'as expected' from its solubility in water (Sharom &
Edgington, 1986). The mobility of 14C-2,4-D acid was compared with
that of triclopyr and picloram on thin-layer chromatographic plates
prepared from several agricultural and forestry soils in Ontario (a
luvisol, a gleysol, a podsol, and an organic soil). The mobilities
were similar for herbicides of similar structure. In a comparison of
the relative adsorptions of concentrations of 5, 10, 25, and 50 mg/L
of the three herbicides on the basis of adsorption coefficients (Kd)
in at least one horizon, the adsorption coefficients for triclopyr and
2,4-D were similar in all soils examined (Jotcham et al., 1989).
The mobility of 14C-2,4-D and its hydrolysis product (by both
chemical and microbial degradation), 14C-2,4-dichlorophenol, was
studied in six soils with clay contents of 8.0-34.6% and organic
matter contents of 0.7-6.1% by thin-layer chromatography. The Rf
value for the parent compound was 0.56-1.00, and that for the
hydrolysis product was 0.11-0.58, indicating that the hydrolysis
product was less mobile than the parent. The mobility of the parent
acid was pH dependent (Somasundaram et al., 1991).
The dimethylamine and isooctyl ester formulations of 2,4-D were
applied at a high rate (8.9 or 0.9 kg/ha) to a silt loam field plot
under both winter wheat and fallow cropping schemes, and simulated
rainfall (16 mm) was applied one day later. 2,4-D was found at 24 cm
three days after application, at 40 cm after five days, and moved
downwards for 30 days; there was some evidence of movement below the
40-cm sampling depth during the growing season. After 191 days, the
concentrations at 40 cm were similar (0.02-0.04 ppm) for the winter
wheat and fallow cropping plots at high and low application rates;
however, the concentration of residues was extremely low (0.05 ppm)
after 45 days at the highest rate of application (8.9 kg/ha) (Wilson &
Cheng, 1976). In two scenarios of the chemical transport and fate of
2,4-D in soil and groundwater--no water flowing and steady-state water
flow on a sandy soil in which 2,4-D was located in a 10-cm layer--2,4-
D degraded to the same extent (36.5 and 37%) with and without water,
and the same amount remained in the soil (58.5 and 63%, respectively).
It was concluded that the model assumptions should be modified on a
site-specific basis (Jury, 1992).
Polyethylene columns, 4.8 cm in diameter and 50 cm high, were filled
with sandy loam, and 2,4-D was applied at 8 kg/ha. The columns were
watered with 55.8 ml/day for 30 days, equivalent to four years of
rainfall in the United Kingdom. The sorption and degradation were KOC
= 108 and half-life = 73 h at 1 µg/g and 213 h at 10 µg/g. 2,4-D was
virtually completely eluted from the column, and 39-47% was recovered
in the eluate. The compound began to leave the column on day 9, whe
there was an equivalent of 28 cm water, and the maximum concentration
left the column on days 10-12 (Lopez-Avila et al., 1986).
The leaching of 2,4-D from home lawns was monitored on ceramic plates
placed under a sandy loam after application at 3.3 or 1.1 kg/ha per
year, with an irrigation scheme. The mean concentrations of 2,4-D in
the soil water percolate were only 0.87 - 0.55 µg/L, suggesting
excellent degradation in the root zone and no threat to groundwater
from these application rates (Gold et al., 1988). The annual geometric
mean concentration of 2,4-D in leachate from home lawns characterized
by silt loam or a sandy loam mantle was less than 1 µg/L (Gold &
Groffman, 1993). The maximum concentration of 2,4-D in drainage water
after application to two clay soils at a rate of 0.40 kg ai/ha
intermittently, between spring 1985 and autumn 1990, was only 0.235
µg/L at one site at one sampling interval (Felding, 1995).
1.3.5 Degradation in soil under laboratory conditions
The hydrolysis of 14C-2,4-D ethylhexyl ester was evaluated in one
silty clay and one sandy loam soil slurry in the dark in an incubator
for 4 h at 25°C. The ester degraded rapidly in both slurries, only
9.6% remaining in the silty clay and 12.6% in the sandy loam at the
end of the study. The half-life of the ester was calculated to be 1.25
h in the silty clay and 1.45 h in the sandy loam (Concha et al.,
1993a).
The half-life for hydrolysis of the isopropyl, n-butyl, and isooctyl
esters of 2,4-D was < 1 min in aqueous 0.1 N NaOH solution. After 24
h in a sandy loam, a heavy clay, and a loam from Saskatchewan, Canada,
the isopropyl and n-butyl esters had completely hydrolysed, while
20-30% of the isooctyl ester remained (Smith, 1972a).
The rate constant, k, for the linear regression plot of
de-esterification of 14C-2,4-D isopropyl ester in a sandy loam
maintained at 25°C for 8 h was 0.752 h-1 and the half-life was 0.9 h
The application rate was based on a dry weight in soil of 9.66 ppm.
The k value for de-esterification in a water/sediment mixture
maintained at 25°C for 24 h was 0.053 h-1 after application at a rate
of 9.67 ppm based on the total weight of water and sediment, and the
half-life was 13.1 h.
In an extensive review of the literature on the degradation, fate, and
persistence of phenoxyalkanoic acid herbicides in soil, 2,4-D was
reported to degrade rapidly in the presence of microorganisms. In the
tables presented, the half-life for degradation of 2,4-D ranged from
two days in a silt loam to about 40 days in a sandy clay, and the
percent carbon dioxide evolution on several soils was 10-95% (Smith,
1989). High concentrations of 2,4-D break down slowly, with an initial
lag phase followed by a greatly increased rate; the first-order rate
constants for the fast phases of degradation were independent of the
concentration, but the duration of the slow phase increased linearly
from 11 to 28 days as the concentration of 2,4-D increased from 1.3 to
25 µg/gm of soil. Soils treated with large amounts of 2,4-D retained
their ability to degrade additional applications of 2,4-D rapidly
(Parker & Doxtader, 1982). Hydrolysis of the isooctyl ester of 2,4-D
in a Naff silt loam and aqueous solution at either 10 or 30°C for up
to 192 h was rapid, with > 80% degraded within 72 h at 30°C. A lag
phase for breakdown was again seen (Wilson & Cheng, 1978). It was
reported that the kinetics of the degradation of 2,4-D in soil are
affected by the effects of temperature and moisture on the microbial
population. An initial lag phase was seen only at temperatures above
the optimum (27°C), and no fast phase was observed. Further, the rate
of decomposition of 2,4-D decreased with soil moisture tension at
temperatures between 20 and 35°C, which corresponded to reduced
activity of 2,4-D-degrading microorganisms. The soils used in this
study had been stored for three years after air drying (Parker &
Doxtader, 1983). When a flooded Finnish sandy clay (pH 4.7) was
incubated after application at a rate of 1000 ppm, only 159 ppm were
recovered after 72 weeks. At an application rate of 10 ppm, 1.7 ppm
2,4-D were recovered (Sattar & Paasivirta, 1980).
The effects of soil properties and soil-degrading microorganisms on
the degradation of 2,4-D have been studied widely (Torstensson, 1978;
Loos et al., 1979; Fournier et al., 1981; Ou, 1984; Sinton et al.,
1986; Scheunert et al., 1987; Kuwatsuka & Miwa, 1989; Rothmel &
Chakrabarty, 1990; Somasundaram & Coats, 1990; Oh, 1991; Masson et
al., 1993; Han & New, 1994; Ka et al., 1994; Myers et al., 1994;
Robertson & Alexander, 1994; Smith et al., 1994; Veeh et al., 1996).
The environmental factors that are significant in determining
degradation rates are pH, temperature, moisture, supplemental
nutrients, substrate concentration, and aeration. Rapid degradation
correlated well with increased microbial concentrations.
Pseudomonas, Arthrobacter, Mycoplana, and Xanthobacter spp. and
Flavobacterium peregilum have been shown to cleave the ether linkage
between the oxygen and the aliphatic side chain of 2,4-D to form
2,4-dichlorophenol.
After an investigation of the effect of inoculum preparation and
density on the efficiency of remediation of 2,4-D in a Pseudomonas
cepacia strain (BRI6001) after full-scale application, it was
recommended that diffusion first be minimized by ensuring proper soil
fragmentation, water content, and aeration, and then use of a
bioaugmentation level of 106 to 108 bacteria/g. At the highest level
of bioaugmentation, the rate of degradation of 2,4-D was less than two
days (Greer et al., 1980; Comeau et al., 1993).
Stoichiometric equations were found to be useful in predicting oxygen
consumption and heat production during the aerobic degradation of
14C-ring- and carboxy-labelled 2,4-D by Pseudomonas cepacia
(Fradette et al., 1994a). Thermograms of the rate of heat evolved
versus time derived by microcalorimetry indicated the time required
for primary biodegradation of 2,4-D by Pseudomonas cepacia in a
liquid medium and sterilized sand under aerobic conditions (Fradette
et al., 1994b).
Another soil bacterium, Pseudomonas testosteroni, isolated from
field soils, used 2,4-D amine residues from farm operations and
herbicide containers as a carbon energy source and degraded the
herbicide over a range of temperature and concentrations. The
water-soluble amine formulations were found to degrade rapidly but the
ester formulations degraded very slowly (Smith & Mortenson, 1991).
Pemberton and Fisher (1977) and Fisher et al. (1978) isolated plasmids
that code for degradation of 2,4-D from a strain of Alcaligenes
paradoxus. Subsequently, the ability of 70 environmental strains of
Pseudomonas to use 2,4-D as a sole source of carbon and energy was
shown to be controlled by plasmids (Pierce et al., 1981, 1982).
Of six microroganisms isolated from herbicide-treated soils
( Flavobacterium peregrinum, Pseudomonas fluorescens, Arthrobacter
globiformis, Brevibacterium sp., Streptomyces viridochromogenes,
and an unidentified Streptomyces sp.), Flavobacterium was the most
active in degrading 2,4-D, 20 mg/kg being completely degraded within
20-30 days. In a liquid medium, Flavobacterium degraded 93.5% of
added 2,4-D within 80 h (Le Van To, 1984).
More microorganisms capable of degrading 2,4-D were found in soils in
Natal under sugar cane (Saccharum officinarum) with rhizospheres
than under African clover (Trifolium africanum), suggesting that the
rapid degradation in the rhizospheres is an additional mechanism for
the protection of certain plants against herbicides applied to soils
(Sandmann & Loos, 1984). Thermophilic microorganisms mineralized 18%
of 14C-2,4-D in composting yard trimmings after 50 days at a
temperature of 55°C (Michel et al., 1995).
Melanic fungi incorporate significant amounts of the ring portion of
2,4-D into the relatively resistant humic acid-type polymers they
form, while the side-chain portion of 2,4-D is used for synthesis of
general cell-wall components and products (Wolf & Martin, 1976). In a
study of the biodegradation, incorporation into biomass, and
stabilization in humus of labelled 2,4-D carbons and decomposition
products from four soils, the side-chain carbon in particular was
rapidly biodegraded at rates comparable to that of glucose at low
concentrations. After one year, up to 8% of the residual activity was
present in the biomass in an unextractable form (Stott et al., 1983).
Concentrations of dissolved oxygen below 1.0 mg/L were rate limiting
for the biodegradation of 2,4-D at 25°C by a bacterial culture from
sewage sludge, suggesting that 2,4-D requires dissolved oxygen as a
cosubstrate for metabolism (Shaler & Klecka, 1986).
Degradation of 2,4-D ethyl ester on a wetland farm soil containing 42%
clay was evaluated for up to 14 days. Degradation was rapid, less than
0.22 ppm remaining after 14 days at various moisture contents
(Bhanumurthy et al., 1989).
In a laboratory experiment with technical-grade 2,4-D (purity, 97%)
and its dimethylamine salt, 14C-2,4-D was introduced into a Webster
silty clay loam, a sandy loam, and a muck soil. Degradation was
measured by the evolution of carbon dioxide. All three soils degraded
2,4-D at application rates of 500 ppm, but at 5000 ppm degradation
ranged from 59% of formulated 2,4-D in the organic soil to a very low
percentage of the technical-grade product in Cecil soil (Ou et al.,
1978).
Samples of chernozem soil were enriched with various nutrients and
2,4-D to determine whether the presence of available metabolic
substrates alters the decomposition of 2,4-D. Overall proliferation of
bacteria and an increased relative proportion of bacterial strains
capable of mineralizing 2,4-D were seen in the enriched samples (Kunc
& Rybarova, 1984). Samples of a Cecil sandy loam were analysed for
2,4-D for up to 14 days after the addition of lime and sulfur. Neither
compound altered the degradation of 2,4-D in sterile soil; the most
rapid degradation occurred in nonsterile soils that had been limed to
adjust the pH to 7-7.4 (Smith, 1972b).
2,4-D was added to municipal sewage processed such that 2,4-D was the
sole carbon source, and biological oxygen depletion was monitored as a
measure of degradation. Less than 5% of the available oxygen was
depleted, indicating poor biodegradation due to the low numbers of
degrading microorganisms (Lieberman & Alexander, 1981). Degradation of
2,4-D was enhanced in clay, sandy loam, and fine sandy loam from
southwestern USA that was either freshly amended or preconditioned for
two months with sewage sludge. Sludge additions had no effect on 2,4-D
degradation in a soil that had been treated previously with 2,4-D
(O'Connor et al., 1981). After incubation of 2,4-D with 52 bacteria
isolated from soil and sewage, 41 of the isolates metabolized 2,4-D
but none grew by using 2,4-D as a carbon source. Nearly all of the
2,4-D had been metabolized within seven days (Rosenberg & Alexander,
1980).
In a study of aerobic soil metabolism conducted according to the
guidelines of the US Environmental Protection Agency, 14C-2,4-D was
added to Catlin silty clay at 25°C for up to 16 days. 14C-2,4-D
degraded rapidly, representing 0.5% of the applied dose after 16 days
of exposure. The calculated half-life was 1.7 days, with pseudo-first-
order kinetics. The major metabolite under aerobic conditions was
carbon dioxide, which accounted for 51.2% of the applied dose at the
end of the study. Two other metabolites were found:
2,4-dichlorophenol, accounting for 3.5% at day 2 and 0.4% at the end
of the study, and 2,4-dichloroanisole, accounting for 2.5% at day 9
and 1.5% at the end of the study. After partitioning of fulvic acid
and humic acid in an extracted residue sample from day 5, 16.1% of the
applied dose was in the fulvic acid fraction and 11.1% in the humic
fraction. HPLC of the fulvic fraction showed that 6.1% of the applied
dose was 2,4-D (Concha & Shepler, 1994b).
Half-lives of 10.8-31.4 days were observed for [1- or 2-14C]-2,4-D in
six soils from Saskatchewan, Canada, under aerobic conditions at field
capacity moisture at 26°C, in an aerobic soil metabolism study.
Initial degradation occurred by cleavage of the ether linkage and not
by decarboxylation (Foster & McKercher, 1973). The half-life for
mineralization of [1- or 2-14C]-2,4-D to carbon dioxide in a
chernozem soil incubated at 28°C was about 90 h, with a lag phase of
less than three days (Kunc & Rybarova, 1983). In a study of the
mineralization of [1-14C]-2,4-D in nine Belgian soils at 22°C, the
more alkaline soils (pH > 6) mineralized 2,4-D rapidly, 80-95% being
degraded after 30 days, with a lag phase of 10-15 days in the
evolution of 14C-carbon dioxide. Mineralization was slow in acidic
soils (pH < 6), < 10% of the applied dose being degraded after 30
days (Moreale & Van Bladel, 1980). Aerobic soil degradation kinetics
were determined for 14C-2,4-D on a loam, a silty clay loam, clay, clay
loam, silt loam, and a sandy loam in the dark at 25°C for up to 230
days. The average half-life was four days, and the main metabolite was
carbon dioxide (McCall et al., 1981).
In three sandy soils at 18°C in the laboratory, the average half-life
for 2,4-D was again four days, and degradation closely followed
first-order kinetics for five half-lives (Altom & Stritzke, 1973).
After application of 2,4-D to a heavy clay, clay loam, and sandy loam
from the prairies in Saskatchewan, Canada, at 20°C for up to 35 days,
the half-life was less than seven days (Smith, 1978a). The half-life
of 14C-2,4-D was not affected by the addition of other herbicides in
the laboratory at 20°C for up to 35 days, and the half-life in a heavy
clay and a sandy loam was again less than seven days (Smith, 1979). In
investigations on the persistence of 14C-2,4-D in heavy clay, sandy
loam, and clay loam, the half-lives were similar regardless of whether
the soils had been pretreated with other herbicides or insecticides,
being three to six days in the clay loam, three to eight days in the
sandy loam, and three to 11 days in the heavy clay soil (Smith, 1980).
The aerobic metabolism of 14C-2,4-D was evaluated in four Canadian
soils, in the dark, under laboratory conditions at 20°C for up to 24
days. In soils that had not received herbicides recently, > 89% of
the applied 14C-2,4-D was metabolized, 25-31% being released as carbon
dioxide, 2-10% recovered as 2,4-dichloroanisole, and 39-43% being
unextractable. In soils that had received treatment with 2,4-D, about
50% of the applied 14C-2,4-D was metabolized to carbon dioxide, 1-4%
to 2,4-dichlorophenol, 2-5% to 2,4-dichloroanisole, and 22-30% being
unextractable (Smith & Aubin, 1991a). Degradation of 14C-2,4-D under
laboratory conditions was faster in Canadian soil that had received 43
annual applications than in soils from untreated control plots (Smith
& Aubin, 1991b).
As the pKa of 2,4-dichlorophenol is 7.89 ( Dictionary of Organic
Compounds, 1996), it is present in most soils in the volatile
phenolic form rather than in the phenolate anion form. Smith (1985)
postulated that failure to isolate 2,4-dichlorophenol and
2,4-dichloroanisole from soils treated with 2,4-D may be due to prior
volatilization or losses during sample preparation. 14C-2,4-D was
broken down rapidly in a clay, a clay loam, and a sandy loam, with
> 70% degradation within 10 days. Carbon dioxide was the main
degradation product, representing 30-42% of the applied radiolabel.
The presence of 2,4-dichlorophenol and 2,4-dichloroanisole was
confirmed by gas chromatography. These metabolites are generated
initially by cleavage of the oxygen-carbon bond at the 2 position on
the side-chain. Methylation in the soils then accounts for the
formation of 2,4-dichloroanisole. Several other studies have indicated
that 2,4-dichlorophenol in soil is rapidly dissipated by both
biological and nonbiological mechanisms (Baker & Mayfield, 1980;
Bollag et al., 1980; Soulas & Fournier, 1987), so that such residues
are not likely to accumulate under field conditions.
In the report of a study on the effect of sulfate on the anaerobic
dechlorination of 27 ppm 2,4-dichlorophenol in freshwater sediments,
few experimental details were provided, but the compound was
completely dechlorinated to 4-chlorophenol when samples from the
sulfate-amended sediments were incubated at 19-40°C (Kohring et al.,
1989). Identical dechlorination of 2,4-D to 4-chlorophenol by cleavage
of the ether linkage and loss of the 2-chloro atom was found in
anaerobically digested municipal sewage (Mikesell & Boyd, 1985).
Estuarine sediment treated with 2,4-D at 24 µmol/L at 13.5°C under
anaerobic conditions formed 2,4-dichlorophenol, at 0.5% of the applied
dose, after 33 days (Eder, 1980).
In a stable, sulfate-reducing model ecosystem capable of reproducing
biological, chemical, and hydraulic conditions under anaerobic
conditions, 2,4-D was almost resistant to biodegradation (Kuhlmann &
Kaczmarzcyk, 1995).
The phytotoxicity and detoxification of 2,4-D in 11 soils from spruce
and pine forests in Sweden with pH values of 3.5-5.1 were compared
with its degradation in three agricultural soils from Sweden with pH
values of 6.3-7.6. In one experiment in which the soils were
inoculated with a 2% salt solution containing 100 ppm 2,4-D, the
degradation time corresponded to a residual amount of 3 ppm of the
2,4-D left in the flasks and ranged from one to four weeks in the
forestry soils and only one week in the agricultural soils
(Torstensson, 1975).
The distribution and effects of 14C-2,4-D butyl ester were evaluated
in a ryegrass ecosystem which consisted of sandy loam soil, annual
ryegrass, numerous invertebrates, and a vole, all housed in a
terrestrial microcosm chamber. Thirty days after application at 1.0
kg/ha, all of the radiolabel detected in the soil was in unextractable
residues in the top 1.0 cm, while the plant material contained an
average of 8.9 mg/kg, which was identified primarily as
2,5-dichloro-4-hydroxyphenoxyacetic acid (Gile, 1983).
A vehicle-portable analytical system for on-site analyses developed by
the Emergencies Science Division of Environment Canada was used to
analyse soil contaminated from burst barrels of chemicals at the scene
of a fire. 2,4-D was found on the site. The recoveries with the
microwave-assisted process and with Soxhlet extraction were acceptable
(Li et al., 1995).
1.3.6 Degradation in soil under field conditions
The dissipation of 2,4-D has been evaluated in many soils in many
areas of the world, and analytical methods have been developed for a
wide range of formulations and soils, typically involving hydrolysis
of the 2,4-D ester to the acid, extraction of the samples with various
solvents, chemical derivatization, and identification by gas
chromatography or capillary column gas chromatography and mass
spectroscopy (Khan, 1975b; Smith, 1976, 1978b; Kan et al., 1981, 1982;
Ahmed et al., 1989; Bruns et al., 1991). A quantitative method has
been developed for the determination of 2,4-D in soils, involving use
of anion-exchange membranes and detection by gas chromatography, which
was used successfully to determine 2,4-D in a commercial formulation
on the surfaces of a loam and a heavy clay. A linear relationship was
observed between the amount of 2,4-D amine detected and the level
spiked onto the soil surface (Szmigielska & Schonenau, 1995).
After application of an n-butyl-isobutyl ester formulation of 2,4-D
to wheat at the three-leaf stage planted in a Canadian black chernozem
soil of unknown composition at a rate of 420g acid equivalent per
hectare, no residues of 2,4-D acid were found in the straw or grain
collected after 80 days. No butyl ester residues were found in the
soil, and < 0.01 ppm 2,4-D was found within 22 days of application
(Cochrane & Russell, 1975). Field experiments were conducted on a Naff
silt loam soil with a split-plot design and two replicates in
Washington State, USA, for up to 175 or 191 days in 1973 and 1974; one
plot was planted with winter wheat and one was under fallow cropping
conditions. The applications consisted of formulated dimethylamine
salt or the isooctyl ester of 2,4-D at 1.1 and 11.2 kg/ha in 1973 and
0.9 and 8.9 kg/ha in 1974, respectively. Considerably more 2,4-D was
recovered from the plots treated with the amine than from those
treated with the ester at the high rate of application for both years.
The average residue concentrations of amine and ester were very low,
regardless of the application rate, after 175 days (0.02-0.12 ppm in
winter wheat and 0.01-0.11 ppm in fallow) in 1973 or 191 days
(0.024).04 ppm in winter wheat or fallow) in 1974 (Wilson & Cheng,
1976). The residues of 2,4-D dimethylamine salt in wheat 35-45 days
after application as a tank mixture during two consecutive growing
seasons had dissipated to < 0.1 mg/kg. No residues were detected in
straw or seeds at maturity in either year (Cessna & Hunter, 1993).
Applications of the propylene glycol butyl ether ester of 2,4-D at 3.4
kg acid equivalent per hectare under conditions of low rainfall on
moderately coarse soils with relatively low organic matter disappeared
rapidly. A maximum concentration of 95.2 mg/kg was found in samples of
vegetation collected within 15 min of application. After 12 months,
the mean residue level had declined to 3.8 mg/kg (Plumb et al., 1977).
The persistence of 2,4-D was also evaluated in a chaparral
environment. Chamise (Adenostoma fasciculatum), grass and forbs,
soil surface litter, and soil were sampled for up to 360 days after
treatment at 4.5 kg/ha. The residues on foliage and litter decreased
rapidly up to 93% within 30 days after treatment and remained constant
until the winter rainfall. No residues were found below 0-5 cm of the
soil surface (Radosevich & Winterlin, 1977). Liquid and granular
formulations of the dimethylamine salt of 2,4-D were applied to turf
at 1.0 or 2.24 kg acid equivalent per hectare in both laboratory and
field experiments. In all of the field experiments, there was a rapid,
significant decrease in dislodgeable residues over time, and by day 7
the residue levels were about 0.02% of the total applied (Thompson et
al., 1984a).
Dissipation of the isooctyl ester of 2,4-D and its acid metabolite
were measured in air, wheat, and soil components in Canada for up to
35 days after application. The acid was detected immediately after
application of a rate of 450 g/ha as the isooctyl ester. Its
hydrolysis to the acid and subsequent degradation of the acid in the
soil were dependent on the availability of water. The acid levels
increased from 156 g/ha on day 1 to 185-210 g/ha over the next 14 days
(Grover et al., 1985).
The dissipation of 2,4-D in soil and water associated with rice
production in three cultural systems--paddy rice, rainfed lowland
rice, and bare ground--in Arkansas, USA, was rapid up to 49 days after
application, with a DT50 in all three systems of 10 days or less
(Johnson et al., 1995).
The degradation of 14C-2,4-D was evaluated after application at a
rate of 1.0 kg/ha on small sandy loam plots for up to 95 weeks. Less
than 2% of the applied radiolabel was recovered from 2,4-D after 45
weeks and < 1% after 95 weeks (Smith & Muir, 1984).
The persistence of the dimethylamine salt was evaluated in four
agricultural soils (loam, clay, fine sand, and sandy clay loam planted
with barley) in southern Ontario and two forestry soils (sand and clay
containing a variety of grasses and conifers) in northern Ontario,
Canada. The application rates were 0.56 kg/ha in the agricultural
fields and 2.24 kg/ha in the forestry areas. The levels of residues on
the agricultural soils were close to the levels of detection (about 5
µg/kg) at the end of the growing season. The degradation patterns in
forest soils were similar to those in the agricultural soils but
higher at the end of the growing season, probably due to the fact that
four applications were made to the forest soils (Thompson et al.,
1984b). Persistence of 2,4-D in forest soil was also evaluated for
brash control in areas of conifer reforestation in five sites in
southern Sweden and six in northern Sweden. 2,4-D disappeared rapidly
from all sites, but low concentrations remained on leaves and branches
for up to two years in the northern sites, ranging from 0.2 to 0.7 µg
2,4-D per soil core (Torstensson et al., 1989).
Soils treated with 2,4-D in the laboratory can develop enhanced
ability to degrade the chemical, seemingly by metabolism by increased
numbers of microorganisms (Roeth, 1986; Smith & Lafond, 1990; Smith et
al., 1991). Increased rates of 2,4-D degradation after repeated use
have also been reported in field studies (Torstensson et al., 1975;
Smith et al., 1989). Significantly more 2,4-D-degrading microorganisms
were found in soils in Saskatchewan, Canada, after 32 years of 2,4-D
treatment than in soils from control plots (Cullimore, 1981). When the
same field plots had received 40-43 annual applications of 2,4-D
formulations, they retained greater degradation of 2,4-D than
untreated control plots for at least 48 weeks after the last
treatment. The levels of residues were below the level of detection
(< 0.02 mg/kg), with no evidence of leaching from the topsoil. The
residues degraded rapidly, 6-9% of the applied 2,4-D remaining after
eight days (Smith et al., 1989; Smith & Aubin, 1991 a). No
agronomically significant effects of these annual treatments were seen
on soil biochemical processes, such as microbial biomass, respiration,
and soil enzyme activity, or on soil fertility (Biederbeck et al.,
1987). After cessation of 2,4-D applications to a Canadian clay soil,
the microflora maintained their ability to degrade 2,4-D rapidly for
at least 204 weeks (Smith & Aubin, 1994). In a study of the effects of
2,4-D and glyphosate on microbial activity in a chernozem soil under
zero-tillage chemical fallow conditions, 2,4-D was applied as the
amine salt at field rates and at 10 times the field rate. No effect
was seen on microbial biomass, mineralization, or nitrification in the
field; however, in the laboratory, application at rates of 2 and 100
times the field rate reduced nitrification by 1 and 79%, respectively
(Olson & Lindwall, 1991).
2,4-D was applied as the propylene glycol and butyl ether esters by
helicopter to brushfields on shallow, rocky, clay loam in three
mountainous areas of southwestern Oregon, USA. The estimated
half-times (overestimates according to the authors) ranged from 18.9
days in the litter, at an application rate of 2.2 kg/ha, to 234.7 days
in the crown, with an application rate of 3.3 kg/ha. The
concentrations of 2,4-D in the surface soil layer decreased rapidly
initially and then more slowly throughout the winter, rarely exceeding
0.02 mg/kg at 45-60 cm, the lowest depth sampled. Less than 30% of the
applied 2,4-D was recovered 37 days after application, and no residues
was recovered 15 cm below the soil surface (Newton et al., 1990). The
soil persistence and lateral movement of 2,4-D after application as a
stem-foliage spray for brush control on two power line rights-of-way
were measured by analysing runoff water and soil samples for up to 48
weeks after treatment. No residue of 2,4-D was found in soil or water
15 weeks after spraying (Meru et al., 1990). After application of
2,4-D amine salt to a power line right-of-way in Ontario, Canada,
which had recently been cleared, < 0.1% of the applied salt was lost
laterally during seven rainfalls (the average rainfall during one day
was > 30 mm), indicating that rain falling even 24 h after spraying
does not cause the release of significant concentrations of 2,4-D
(Suffling et al., 1974). Residues of 2,4-D were detected in soil
samples from a forest watershed adjacent to a treated right-of-way in
eastern Ontario, Canada, at least 36 m from the right-of-way 12 days
and two and eight months after application. It was not clear whether
the 2,4-D found was actually derived from the application site or was
from another source (McKinley & Arron, 1988).
Water, sediment, and samples from the wall of an outdoor enclosure
located in a typical bog lake in a sandy soil area of northeastern
Ontario, Canada, were analysed for residues of the isooctyl ester of
2,4-D for more than 100 days. Less 2,4-D was adsorbed to the sediments
than to the sides of the enclosure, where < 25% was found (Solomon
et al., 1988).
Thirty-five studies on terrestrial, aquatic, and forestry field
dissipation were conducted in the USA in 1993 and 1994 according to
the guidelines of the US Environmental Protection Agency. The
half-lives and other data are given in Table 1. Fourteen studies of
terrestrial dissipation were conducted in 1993 to compare the
disappearance of the dimethylamine salt and the 2-ethylhexyl ester of
2,4-D when applied in a spray solution to pasture and turf in
Colorado, North Carolina, and Texas with grass growing on the plots
and to bare soil in the same area. Four trials on cultured wheat were
established in North Carolina and four in Colorado. Twelve similar
studies were conducted in 1994 in California, North Dakota, Nebraska,
and Ohio on pasture and turf with grass present and on bare soil
according to the timing for use on turf, wheat, and corn. Four further
studies were conducted in 1994 to compare the disappearance of
granules of the dimethylamine salt and 2-ethylhexyl ester, the former
in North Dakota and the latter in Ohio. The materials were applied as
for the treatment of turf, with turf present and on bare soil.
An aquatic trial was conducted in 1993 in Louisiana on rice, in which
the dimethylamine salt was applied aerially to a water-filled rice
paddy at a target application rate of 1.5 lb acid equivalent per acre
(1.68 kg/ha). The highest 2,4-D acid residue in soil (mean total of
three depths) was found one day after application, the residues
declined to 0.013 ppm three days after application. Metabolites were
Table 1. Dissipation of 2,4-D after application at various sites in the USA
Site Type 2,4-D No. of Soil type Crop Rate Metabolite Half-life Reference
form applications (kg ae/ba) (days)
CA Terrestrial DMA 2, ground Sandy loam Pasture 2.24 DCA 5.7 (1st) Hatfield (1995c)
DCP 30.5 (2nd)
CA Terrestrial 2-EHE 2, ground Sandy loam Pasture 2.24 DCA 1.8 ester (1st) Hatfield (1995d)
DCP 4.7 ester (2nd)
10.6 acid (1st)
27.3 acid (2nd)
CA Terrestrial DMA 2, ground Sandy loam Turf 2.24 DCA 29.1 (1st) Hatfield (1995e)
7.5 (2nd)
CA Terrestrial 2-EHE 2, ground Loamy sand Turf, soil 2.24 DCP < 2 ester (both) Hatfield (1995f)
Soil 7.9 acid (1st)
9.7 acid (2nd)
Grass 5.7 acid (1st)
9.3 acid (2nd)
Thatch 9.8 acid (1st)
12.9 acid (2nd)
CA Terrestrial DMA 2, ground Sandy loam Bare soil 2.24 DCA 7.3 (1st) Hatfield (1995g)
7.6 (2nd)
CA Terrestrial 2-EHE 2, ground Loamy sand Bare soil 2.24 DCA < 3 ester (both) Hatfield (1995h)
DCP 4.4 acid (1st)
15.0 acid (2nd)
CO Terrestrial 2-EHE 2, broadcast Sandy clay loam Wheat 1.4 DCP 2.2 ester (1st) Silvoy (1995a)
2.1 ester (2nd)
4.8 acid (1st)
2.5 acid (2nd)
Table 1. (continued)
Site Type 2,4-D No. of Soil type Crop Rate Metabolite Half-life Reference
form applications (kg ae/ba) (days)
CO Terrestrial 2-EHE 2, broadcast Sandy clay loam Bare soil 1.4 DCP 1.4-1.7 ester(both) Silvoy (1995b)
6.0 acid (1st)
2.0 acid (2nd)
CO Terrestrial DMA 2 Sandy clay loam Bare soil 1.4 DCA 5.1 (1st) Silvoy (1994a)
CO Terrestrial DMA 2 Sandy clay loam Wheat 1.4 DCA 5.1 (1st) Silvoy (1994b)
NC Terrestrial DMA 2, broadcast Sand Wheat 1.4 DCP 5.5 (1st) Barney (1995a)
DCA 2.7 (2nd)
NC Terrestrial 2-EHE 2, broadcast Sand Wheat 1.4 DCP 4.0 ester (1st) Barney (1995b)
DCA 1.8 ester (2nd)
9.3 acid (1st)
6.2 acid (2nd)
NC Tenestrial DMA 2, broadcast Sand Turf, soil 2.24 DCP 3.3 (1st) Barney (1995c)
DCA 2.3 (2nd)
Grass 6.4 (1st)
7.7 (2nd)
NC Terrestrial DMA 2, broadcast Sand Bare soil 2.24 DCP 3.4 (1st) Barney (1995d)
(turf rates) DCA 2.5 (2nd)
NC Terrestrial 2-EHE 2, broadcast Sand Turf 2.24 DCP 0.34 ester (1st) Barney (1995e)
DCA 4.4 acid (1st)
2.2 acid (2nd)
NC Terrestrial DMA 2, broadcast Sand Bare soil 1.4 DCP 2.9 (1st) Barney (1995f)
(wheat rates) DCA 2.6 (2nd)
Table 1. (continued)
Site Type 2,4-D No. of Soil type Crop Rate Metabolite Half-life Reference
form applications (kg ae/ba) (days)
NC Terrestrial 2-EHE 2, broadcast Sand Bare soil 1.4 DCA 12.9 ester (1st) Barney (1995g)
(wheat rates) DCP 5.2 ester (2nd)
6.0 acid (1st)
3.9 acid (2nd)
NC Terrestrial 2-EHE 2, broadcast Sand Bare soil 2.24 Barney (1995h)
(turf rates)
ND Terrestrial DMA 2, ground Sandy loam Turf, soil 2.24 DCP 10.3 (1st) Hatfield (1995i)
DCA 5.1 (2nd)
Grass and thatch 2.5-6.4 (both)
ND Terrestrial DMA 2, ground Loam Bare soil 2.24 DCP 19.6 (1st) Hatfield (1995j)
DCA 18.4 (2nd)
ND Terrestrial DMA 2, ground Bare soil DCP 3.9 (1st) Hatfield (1995k)
DCA 4.5 (2nd)
ND Terrestrial 2-EHE 2, ground Bare soil DCP 4.4 ester (1st) Hatfield (1995l)
DCA 3.6 ester (2nd)
6.1 acid (1st)
6.4 acid (2nd)
NE Terrestrial DMA 4, ground Silt loam Bare soil 2.24 DCA 8.6 (1st) Hatfield (1995m)
1.12 DCP 3.9 (2nd)
0.86 1.1 (3rd)
1.68 2.8 (4th)
NE Terrestrial 2-EHE 4, ground Silt loam Bare soil 2.24 DCP 2,5-4.1 ester Hatfield (1995n)
1.12 DCA All applications
0.86 46.5 acid (1st)
1.68 3.5 acid (2nd)
4.4 acid (4th)
Table 1. (continued)
Site Type 2,4-D No. of Soil type Crop Rate Metabolite Half-life Reference
form applications (kg ae/ba) (days)
OH Terrestrial DMA 4, ground Silt clay loam Bare soil 2.24 DCA 23.5 (1st) Hatfield (1995o)
(corn site) 1.12 DCP 5.9 (2nd)
0.56 0.9 (3rd)
1.68 10.4 (4th)
OH Terrestrial 2-EHE 4, ground Clay loam Bare soil 2.24 DCP 10.9 ester (1st) Hatfield (1995p)
1.12 4.6 ester (2nd)
0.68 1.2 ester (3rd)
1.68 2.9 ester (3rd)
5.5 acid (2nd)
1.2 acid (3rd)
7.0 acid (4th)
OH Terrestrial 2-EHE 2, ground Silt loam Turf 2.24 DCP 14.3 ester (1st) Hatfield (1995q)
(granule) DCA 10.9 ester (2nd)
Soil 13 acid (2nd)
Grass and thatch 9.3-4.0 acid
OH Terrestrial 2-EHE 2, ground Silt loam Bare soil 2.24 DCP 7.4 ester (1st) Hatfield (1995a)
(granule) (Turf site) DCA 6.3 ester (2nd)
-10.7 acid (1st)
6.6 acid (2nd)
TX Terrestrial DMA 2, broadcast Sandy loam Pasture 2.24 DCP 7.9 (1st) Barney (1995i)
10.2 (2nd)
TX Terrestrial 2-EHE 2, broadcast Silt loam Pasture 2.24 DCP 1.4 ester (1st) Barney (1995j)
0.5 ester (2nd)
4.2 acid (1st)
13.1 acid (2nd)
Table 1. (continued)
Site Type 2,4-D No. of Soil type Crop Rate Metabolite Half-life Reference
form applications (kg ae/ba) (days)
LA Aquatic DMA 1, aerial Silt loam Rice (water) 1.68 DCP 1.1 Barney (1994)
Rice (soil) 4-CPA 1.5
NC Aquatic DMA 2, subsurface Pond (water) 46.81 DCP 19.7 (1st) Hatfield (1995b)
injection Pond (soil) DCA 2.7 (2nd)
DCP 7.6 (1st)
2.0 (2nd)
ND Aquatic DMA 2, subsurface Pond (water) 46.81 DCP 13.9 (1st) Hatfield (1995r)
injection Pond (soil) 4-CPA 6.5 (2nd)
DCP -17.4 (1st)
29.5 (2nd)
GA Forest 2-EHE 1, aerial Sandy clay loam Exposed soil 4.48 DCP 1.0 ester Barney (1996)
broadcast Protected soil DCA 1.7 ester
Exposed soil 4.0 acid
Protected soil 3.6 acid
Foliage 7.2 2-EHE
23.5 2,4-D
44.0 DCP
Leaf litter 51.0 2-EHE
52.2 2,4-D
68.3 DCP
84.7 DCA
OR Forest DMA 1, aerial Loam Exposed soil 4.48 DCP 38.7 Barney (1995k)
broadcast Protected soil DCA 50.8
Foliage 37.4 2,4-D
80.8 DCP
Leaf litter 65.7 2,4-D
111.3 DCP
CA, California; CO, Colorado; NC, North Carolina; ND, North Dakota; NE, Nebraska; OH, Ohio; TX, Texas; LA, Louisiana; GA, Georgia;
OR, Oregon;
DMA, dimethylamine samt; 2-EHE, 2-ethylhexyl ester; DCA, 2,4-dichloroanisole; DCP, 2,4-dichlorophenol, 4-CPA, 4-chlorophenoxyacetate
not detected above the limit of quantification at any sampling
interval. Two aquatic trials were conducted during 1994, in North
Carolina and in North Dakota, in which two applications of the
dimethylamine salt at approximately 4 ppm were made. At both sites,
the half-life was much shorter for the second application than the
first, especially in water, due perhaps to the build-up of
microorganisms capable of degrading 2,4-D. The half-life of 2,4-D at
the North Carolina site was shorter than that in North Dakota. The
effects of 2,4-D in the water phase in these studies are discussed in
section 4.2.4.
Two studies of forestry dissipation were conducted in 1993, one in
Georgia with a 2-ethylhexyl ester formulation on a sandy clay loam and
one in Oregon with a dimethylamine salt formulation on a loam. Each
site received one broadcast aerial application at 4.0 lb acid
equivalent per acre (4.48 kg/ha). Soil, leaf litter, foliage,
sediment, and water were collected and analysed. The half-lives are
given in Table 1. The dimethylamine salt of 2,4-D disassociates
rapidly, leaving free 2,4-D, which can undergo further degradation.
The 2-ethylhexyl ester undergoes de-esterification on contact with a
microbially active substrate. Since the ester first undergoes
de-esterification to yield 2,4-D, the amount of 2,4-D increases during
the first few days, while the amount of the ester decreases. The
conversion from ester to acid is not complete before some of the 2,4-D
has degraded. Degradation of the ester is determined by adding the
amount of ester (converted to 2,4-D equivalent) and the amount of
2,4-D found. The data for the 1994 study of field spraying indicate
average half-lives of 4.3 days for the dimethylamine salt and 5.3 days
for the 2-ethylhexyl ester, after removal of a few outliers. The data
for 1993 indicate average half-lives of 4.5 days for the dimethylamine
salt and 5.1 days for the ester. These results do not indicate any
difference between the two forms of 2,4-D in the rate of degradation
or dissipation. The data are also consistent with those for field
dissipation over 40 years.
1.3.7 Uptake by plants
The effects of 2,4-D acid, dimethylamine salt, diethanolamine salt,
and 2-ethylhexyl ester on the vegetative vigour, seed germination, and
seedling emergence of 10 non-target crops (buckwheat, corn, cucumber,
mustard, oats, onion, radish, sorghum, soya bean, and tomato) was
determined in studies conducted according to the guidelines of the US
Environmental Protection Agency.
When 2,4-D acid was applied at concentrations equivalent to 4.2, 2.1,
1.05, 0.53, 0.26, 0.13, 0.065, 0.03, 0.015, and 0.0075 lb ai/acre
(4.7, 2.4, 1.2, 0.59, 0.29, 0.15, 0.07, 0.03, 0.017, and 0.0084
kg/ha), onion was the most sensitive monocot, with an NOEL and an
EC25 value of < 0.0075 lb ai/acre (< 0.0084 kg/ha); the EC50
was 0.089 lb ai/acre (0.1 kg/ha). Neither corn nor oats was sensitive
to 2,4-D acid. Cucumber, mustard, and tomato were the most sensitive
dicot species, with calculated NOELs lower than the lowest dose
tested. The response of radish at 0.0075 lb acid equivalent/acre
(0.0084 kg/ha) was 93% that of the controls and appeared to be the
NOEL (Bakus, 1992a). In a continuation of the study, cucumber,
mustard, onion, and tomato were tested at concentrations of 0.0075,
0.004, 0.002, and 0.001 lb ai/acre (0.0084, 0.004, 0.002, and 0.001
kg/ha). The NOEL for cucumber and tomato was 0.002 and the that for
mustard and onion, 0.0075 lb ai/acre (0.0084 kg/ha) (Bakus, 1993a). In
studies in which 2,4-D acid was applied in petri dish assays at
0.75-3.8 lb ai/acre (0.84-4.2 kg/ha) and to seeds planted in field
soil at 0.75-4.2 lb ai/acre (0.84-4.2 kg/ha), 2,4-D had little effect
on seed germination in petri dishes. Tomato was the most sensitive
species, with NOEL, EC25 and EC50 values of < 0.12 lb ai/acre
(0.84 - 4.2 kg/ha). Slight distortion was observed on tomatoes in the
seedling emergence phase, and stunting was noted on onion, cucumber,
sorghum, soya bean, and buckwheat (slight). Buckwheat, corn, cucumber,
oats, sorghum, and tomato had NOEL, EC25, and EC50 values > 4.2 lb
ai/acre (< 0.13 kg/ha), while the NOEL values for mustard, onion,
radish, and soya bean were > 1.05 lb ai/acre (> 1.18 kg/ha)
(Bakus, 1992b). When tomato was treated at concentrations of 0.0075,
0.004, 0.002, or 0.001 lb acid equivalent/acre (0.0084, 0.004, 0.002,
and 0.001 kg/ha), no unusual growth or inhibition of radicle, roots,
hypocotyl, coleoptile, cotyledon, or leaf was observed, as compared
with controls. For all the species tested except radish, the percent
emergence NOEL in the soil was higher than the percent germination
NOEL in the petri dishes. The plant system is not likely to have
direct exposure to the test material in a soil substrate (Bakus,
1993b).
The effect of the dimethylamine salt on vegetative vigour was studied
at concentrations of 0.96, 0.48, 0.24, 0.12, 0.06, 0.03, and 0.015 lb
acid equivalent/acre 1.08, 0.54, 0.27, 0.13, 0.07, 0.03, and 0.0017
kg/ha). Onion was the most sensitive species, with an EC50 of 0.142
lb/acre (0.16 kg/ha). The NOEL for sorgham was 0.06 (0.07), and that
for oats, the least sensitive monocot, was > 0.96 lb acid
equivalent/acre (> 1.08 kg/ha). The most sensitive dicot was
mustard, with an EC50 of 0.061 lb acid equivalent/acre (0.07 kg/ha).
Cucumber was the least sensitive dicot, with an EC50 of 0.96 lb acid
equivalent/acre (1.08 kg/ha) (Bakus & Crosby, 1992a). When oats were
treated at 3.8 and 1.9 lb acid equivalent/acre (4.3 and 2.1 kg/ha),
the symptoms were slight, with a projected NOEL of 1.9 lb acid
equivalent/acre (2.1 kg/ha) (Bakus, 1993c).
The effect of 2,4-D dimethylamine salt on seed germination and
seedling emergence studies was evaluated in petri dishes and in field
soil. The doses evaluated were 0.96, 0.48, 0.24, 0.12, 0.06, 0.03, and
0.015 lb acid equivalent/acre (1.08, 0.54, 0.27, 0.13, 0.07, 0.03, and
0.017 kg/ha) in the soil phase and 0.0075, 0.004, 0.002, 0.001, and
0.0005 lb acid equivalent/acre (0.0084, 0.004, 0.002, 0.001, and
0.0006 kg/ha) in the petri dishes. In the petri dish studies, the
NOEL, EC25, and EC50 values for corn, cucumber, oats, onion, sorghum,
and soya bean were > 0.96 lb acid equivalent/acre (> 1.08
kg/ha). A dose-response relationship was seen for tomato, with
calculated NOEL, EC25 and EC50 values of 0.24, 0.26, and 0.56 lb/acre
(0.27, 0.29, and 0.63 kg/ha), respectively. No effect was seen on the
seedling emergence phase of corn and oats at 0.96 lb acid
equivalent/acre (1.08 kg/ha). The remaining eight species showed some
stunting, but only cucumber showed symptoms at the lowest rate.
Mustard, onion, and radish showed significant effects at emergence; no
significant effect of treatment was seen on buckwheat, corn, cucumber,
oats, sorghum, soya bean, or tomato. The most sensitive species was
mustard; radish and onion were less sensitive, with an NOEL of 0.48 lb
acid equivalent/acre (0.54 kg/ha). All species except tomato that
showed no effect in the soil bioassay also showed no effect in the
petri dish assay. The fresh weight of emerging seedlings in soil
differed from the percent seedling emergence, the NOEL, EC25 and EC50
values for inhibition of fresh weight of buckwheat, cucumber, mustard,
radish, and sorghum being lower than the percent emergence. This
indicates that 2,4-D dimethylamine salt can inhibit the growth of
seedlings that emerge from treated soil (Bakus & Crosby, 1992b). In
nine of the same plant species (no buckwheat), concentrations of 3.8,
1.9, 0.96, and 0.48 lb acid equivalent/acre (4.3, 2.1, 1.08, and 0.54
kg/ha) were evaluated in both the petri dish assay and seeds planted
in field soil. The NOEL for corn and soya bean was 3.8 lb acid
equivalent/acre (4.3 kg/ha), but that for cucumber and oats was 1.9
(2.1) and that for onions and sorghum was < 1.9 lb acid
equivalent/acre (< 2.1 kg/ha). The NOEL for mustard and radish was
reported to be > 0.06 but < 0.48 lb acid equivalent/acre (> 0.07
but < 0.54 kg/ha). The NOEL for tomato was > 0.96 but < 1.9 lb acid
equivalent/acre (> 1.08 but < 2.1 kg/ha) (Bakus, 1993d).
2,4-D 2-ethylhexyl ester was tested at concentrations of 0.96, 0.48,
0.24, 0.12, 0.06, 0.03, 0.015, and 0.0075 lb acid equivalent/acre
(1.08, 0.54, 0.27, 0.13, 0.07, 0.03, 0.017, and 0.0084 kg/ha). Onion
was the most sensitive species, but the data were variable, with an
NOEL of 0.24 lb acid equivalent/acre (0.27 kg/ha). Corn and oats were
not affected. The most sensitive dicot was tomato. Cucumber was the
least sensitive species, with NOEL, EC25, and EC50 values of 0.015,
0.192, and 0.78 lb acid equivalent/acre (0.017, 0.22, and 0.87 kg/ha),
respectively (Bakus & Crosby, 1992c). The NOEL in a further study was
3.8 acid equivalent/acre (4.3 kg/ha) for corn and > 0.96 but < 1.9
lb acid equivalent/acre (> 1.08 but < 2.1 kg/ha) for oats (Bakus,
1993e).
In studies of the effect of 2,4-D 2-ethylhexyl ester on seed
germination and seedling emergence in petri dishes and in field soil,
concentrations of 0.96, 0.48, 0.24, 0.12, 0.06, 0.03, 0.015, 0.0075,
and 0.004 lb acid equivalent/acre (1.08, 0.54, 0.27, 0.13, 0.07, 0.03,
0.017, 0.0084, and 0.0045 kg/ha) were evaluated, with additional rates
of 0.002 and 0.001 lb acid equivalent/acre (0.002 and 0.004 kg/ha) in
the petri dishes. None of the species examined was very sensitive to
the ester during germination, and no effects were noted on buckwheat,
corn, oats, sorghum, soya bean, or tomato in the emergence phase. When
mustard, onion, radish, and tomato were tested at the lower rates in
petri dishes, only mustard was more sensitive in the soil bioassay.
The most sensitive monocots were onion and sorghum, and radish was the
most sensitive dicot. The percent emergence of cucumber, mustard,
radish, and tomato was affected (Bakus & Crosby, 1992d). In a similar
study, concentrations of 3.8, 1.9, 0.96, 0.48, and 0.24 lb acid
equivalent/acre (4.3, 2.1, 1.08, 0.54, and 0.27 kg/ha) of the ethyl
ester were evaluated. The NOEL was 3.8 lb acid equivalent/acre (4.3
kg/ha) for corn, cucumber, oats, onion, sorghum, soya bean, and tomato
in petri-dish assays. In the emergence bioassay, the NOEL value for
onion was 0.24 lb acid equivalent/acre (0.27 kg/ha) (Bakus, 1993f). In
another evaluation of cucumber and onion in the seedling emergence
bioassay at doses of 0.96, 0.48, 0.24, 0.12, 0.06, 0.03, 0.015,
0.0075, 0.00375, 0.001875, and 0.0009375 lb acid equivalent/acre
(1.08, 0.54, 0.27, 0.13, 0.07, 0.03, 0.017, 0.0084, 0.0042, 0.0021,
and 0.0011 kg/ha), onions were the most sensitive species, with an
NOEL of 0.24 lb acid equivalent/acre (0.27 kg/ha); the EC25 and EC50
values were 0.445 and > 0.96 lb acid equivalent/acre (0.5 and > 1.08
kg/ha), respectively. The NOEL, EC25, and EC50 values for cucumber
were > 0.96 lb acid equivalent/acre (> 1.08 kg/ha) (Bakus, 1995).
The effects of 2,4-D diethanolamine salt on vegetative vigour were
studied at concentrations of 1.5, 0.75, 0.375, 0.1875, 0.09, 0.045,
0.0225, 0.01, 0.005, and 0.0025 lb acid equivalent/acre (1.7, 0.84,
0.42, 0.21, 0.1, 0.05, 0.03, 0.01, 0.006, and 0.0028 kg/ha). At the
two highest rates, only sorghum appeared to be unaffected four days
after treatment; all of the other species showed stunting, distortion,
or both. At the two lowest rates, effects were seen only in tomato
seven days after treatment, the uppermost leaves being distorted.
Tomato, cucumber, buckwheat, and radish were the most sensitive dicot
species: the statistically generated NOEL values for tomato and
cucumber were < 0.0025 lb acid equivalent/acre (< 0.0028 kg/ha),
the lowest dose tested. The EC25 and the EC50 values were 0.00462 and
0.1545 lb acid equivalent/acre (0.005 and 0.17 kg/ha for tomato and
0.00801 and 0.07971 lb acid equivalent/acre (0.009 and 0.089 kg/ha)
for cucumber, respectively. Onion was the most sensitive monocot
species tested, with a statistically generated NOEL of 0.0225 lb acid
equivalent/acre (0.025 kg/ha) and EC25 and the EC50 values of 0.03668
and 0.12868 lb acid equivalent/acre (0.04 and 0.14 kg/ha). Oats were
the least sensitive species, with NOEL, EC25, and EC50 all > 1.5
lb acid equivalent/acre (> 1.7 kg/ha) (Bakus, 1992c).
The effect of 2,4-D diethanolamine salt on seed germination and
seedling emergence was studied in petri dishes and field soil at
concentrations of 1.5, 0.75, 0.375, 0.1875, 0.09, and 0.045 lb acid
equivalent/acre (1.7, 0.84, 0.42, 0.21, 0.1, and 0.05 kg/ha). In the
petri dishes, the least sensitive species was cucumber, with an NOEL
of > 1.5 lb acid equivalent/acre (> 1.7 kg/ha). Buckwheat, corn,
onion, and radish had an NOEL value of 0.75 lb acid equivalent/acre
(0.84 kg/ha). Emergence after 14 days in the soil bioassay was lowest
for onion, being, 30% of that in controls. Buckwheat, corn, oats, and
soya beans had NOEL, EC25, and EC50 values > 1.5 lb acid
equivalent/acre (>. 1.7 kg/ha). Mustard and radish were fairly
sensitive, but onion was the most sensitive, with NOEL, EC25, and
EC50 values < 0.045 lb acid equivalent/acre (< 0.05 kg/ha). Corn
was the least sensitive species in fresh weight determinations, with
an NOEL of 0.75 lb acid equivalent/acre (0.84 kg/ha). Sorghum was
highly sensitive to the test material, with an NOEL and EC25 value
< 0.045 lb acid equivalent/acre (< 0.05 kg/ha). The NOEL values
in the petri dish assays showed that the diethanolamine salt had
little effect on the germination of seeds, except for tomato and soya
bean. The percent results for emergence in the soil bioassay generally
correlate with the results for germination, with two exceptions: the
percent emergence of tomato and soya bean was higher in this phase of
the study than the percent germination in the petri dish test (Bakus,
1992d).
The effect of 2,4-D diethanolamine salt on seedling emergence and
early seedling growth of onion was tested at concentrations of 1.5,
0.75, 0.375, 0.1875, 0.09, 0.045, 0.0225, 0.011, 0.0056, and 0.0028 lb
acid equivalent/acre (1.7, 0.84, 0.42, 0.21, 0.1, 0.05, 0.03, 0.01,
0.006, and 0.003 kg/ha). Plant dry weight at the end of the test, 32
days after treatment, was the most sensitive indicator of effects. The
lowest values 32 days after planting were 0.1875 lb dry weight of acid
equivalent/acre (0.21 kg/ha) for the no-observed-effect concentration
(NOEC), 0.296 lb (0.33 kg/ha) dry weight for the EC25 value, and
0.713 lb (0.8 kg/ha) dry weight for the EC50 value (Crosby, 1996).
The effects of 2,4-D isopropyl ester on vegetative vigour, seed
germination, and seedling emergence were studied in 10 plant species
(cabbage, corn, cucumber, lettuce, oat, onion, perennial ryegrass,
soya bean, tomato, turnip) at several concentrations. The measured
concentrations varied from a high of 0.12 lb ai/acre (0.13 kg/ha) to a
low of 0.0000073 lb ai/acre (0.0000081 kg/ha) (nominal rate for
turnip). The vegetative vigor of lettuce and cabbage was most heavily
affected, with NOEC values of 0.0073 and 0.015 lb ai/acre (0.0082 and
0.017 kg/ha), respectively. Corn, cucumber, oat, onion, perennial
ryegrass, and soya bean were the least affected, with NOEC, EC25, and
EC50 values > 0.12 lb ai/acre (> 0.13 kg/ha). Seed germination
was most affected for lettuce, with NOEC, EC25, and EC50 values of
0.00050, 0.0022, and 0.0065lb ai/acre (0.00056, 0.0025, and 0.0073
kg/ha), respectively. Soya bean, oat, perennial ryegrass, and corn
were the least affected, with NOEC, EC25, and EC50 values of >
0.11-0.12 lb ai/acre (> 0.12-0.13 kg/ha). Lettuce seedling
emergence was the most heavily affected, with NOEC, EC25, and EC50
values of 0.00056, 0.0011, and 0.0024 lb ai/acre (0.00063, 0.0012, and
0.0027 kg/ha); the emergence of cabbage, turnip, onion, and soya bean
seedlings was also affected, with NOEC values ranging from 0.0017
(0.0019) for cabbage to 0.0070 lb ai/acre (0.008 kg/ha) for turnip.
Corn, oat, and perennial ryegrass were the least affected, with NOEC,
EC25, and EC50 values of > 0.11 lb ai/acre (> 0.12 kg/ha) (Hoberg,
1996).
An extensive review of the chemical and environmental factors that
affect the foliar absorption and translocation of 2,4-D is available
(Richardson, 1977), and another summarizes more than 10 studies of the
processes, correlations, and models of the mechanisms of uptake of
organic chemicals, including 2,4-D, by citrus trees and barley, soya
beans, and vegetables from soil and the atmosphere (Paterson & Mackay,
1990). A steady-state model for equilibrium partitioning of organic
chemicals, including 2,4-D, in multi-compartment plants has been
described, and a dynamic model of the time course of chemical uptake
and distribution was proposed (Paterson et al., 1991). This was
expanded into a three-compartment (root, stem, and foliage) mass
balance model of plants, in which the uptake of organic chemicals,
including 2,4-D, from soil and the atmosphere was quantified. The
results for 2,4-D show that an appreciable fraction of the chemical in
soil is transpired into plant foliage within 72 h, and the compartment
response times are short. Most of the chemical reaches the foliage
(Paterson et al., 1994).
A series of experiments was performed on the accumulation, permeation,
and desorption of [2-14C]-2,4-D in and across adaxial cuticles on the
leaves of rubber plants, bitter oranges, tomatoes, and green peppers.
Accumulation of radiolabel from the cuticles of the four species was
predicted by determining the partition coefficients for the plant
cuticle:water system. The agreement between the measured partition
coefficients ( Kexp) and the coefficients ( Kcal) predicted from log
octanol:water partition coefficient ( Ko:w) were good, with log
Kexp = 2.606 and log Kcal = 2.478 (Kerler & Schönherr, 1988a). A
study was conducted to determine whether the permeation of [2-14C]-
2,4-D into astomatous circular membranes of Citrus auranthium L.
leaves can be predicted from the Ko:w. Various equations were used
for comparisons with measured values. Lipid solubility was the most
important determinant of the permeation of chemicals from leaf
surfaces. The experimental permeability ( Pexp) in the [2-14C]-2,4-D
system was 2.8, which agreed with the calculated permeability ( Pcal)
of 2.7 when the partition coefficient and the molar volume of the
solute were considered (Kerler & Schönherr, 1988b). The desorption of
[2-14C]-2,4-D from cuticular and polymer matrix membranes isolated
from four plant species ( Citrus, Ficus, Lycopersicon, and
Capsicum) was studied by preloading the cuticles and membranes with
radiolabelled 2,4-D by sorption from solutions and then subjecting
them to simultaneous bilateral desorption. Only 2-3% of the 2,4-D
initially contained in the cuticles was desorbed from the outer
surface, while 86-92% was desorbed from the inner surfaces within 6 h.
The initial rates of desorption from the inner surfaces were 50-80
times greater than those from the outer surfaces. The authors
maintained that asymmetrical desorption is due mainly to the presence
of apolar and crystalline soluble lipids (waxes) on the surface and in
the outer layers of the cuticles, which drastically decrease the
mobility of 2,4-D in the outer layers of the leaf. In fruit cuticles,
extensive cutinization of the anticlinal and periclinal walls
increases the inner surface area and thus contributes significantly to
asymmetry. The fate of chemicals sorbed in cuticles depends on the
presence or absence of a sink (Schönherr & Riederer, 1988).
The rate of uptake of 14C-2,4-D in aqueous solution was studied in
needles from five conifer species. The uptake was biphasic in all
species, with a rapid first phase, which was complete within about 30
min, and a slower, constant second phase lasting 30-360 min. The rates
of uptake were proportional to specific surface areas between the
solutions and the needles and the partition coefficients. The mean
permeance, P, for 2,4-D was 3.07 × 10-11 (Schreiber & Schönherr,
1992).
An investigation was carried out to determine the metabolites formed
when soya bean cotyledon callus cultures were grown with
[1-14C]-2,4-D as an auxin. ß-Glucosidase treatment of the
water-soluble fractions from the tissue yielded eight aglycones after
eight days. The metabolite 4-hydroxy-2,5-dichlorophenoxyacetic acid
was the most abundant aglycone produced during the 32-day growth
period, 4-hydroxy-2,3-dichlorophenoxyacetic acid being a minor
metabolite. Seven ether-soluble components were detected; 2,4-glutamic
acid was detected in large amounts after 24 h, while 2,4-D aspartic
acid was the most abundant metabolite after a longer period. It was
concluded that 2,4-D amino-acid conjugates were actively metabolized
by the tissue to free 2,4-D and water-soluble metabolites (Feung et
al., 1972). Similar metabolic patterns were noted in callus tissue
from soya bean cultures and several other plants, with the addition of
several new amino-acid conjugates (Feung et al., 1973, 1975), and the
presence of the amino-acid conjugates was confirmed in a similar study
(Davidonis et al., 1980).
An 11-fold increase in uptake of 2,4-D by the roots of barley plants
was seen as the pH of the nutrient solution dropped from 6.5 to 4.0.
At pH 4, the uptake of 2,4-D seemed to be influenced by metabolism
(Shone & Wood, 1974). A similar association with pH was seen in
excised wheat roots (Zsoldos & Haunold, 1979). The uptake of
14C-2,4-D in solution into barley roots and subsequent translocation
into shoots increased as the pH of the solution decreased from 4.0 to
8.0 (Briggs et al., 1987). In a study of the absorption,
translocation, and metabolism of 14C-2,4-D in hemp dogbane seedlings
grown in nutrient solution, the plants were removed from the solution
at various intervals up to 12 days after treatment and sampled for
analysis. After 12 days, 34-55% of the radiolabel was found in 2,4-D,
and negligible amounts were lost as volatile compounds. A temperature
of 30°C instead of 25°C did not affect translocation of 2,4-D (Schultz
& Burnside, 1980).
Experiments were performed to determine whether the absorption of
14C-2,4-D from culture solutions by excised barley ( Hordeum
vulgate L.) roots was passive or active. Absorption was followed at
0.5°C and at 21°C, over 30 min under anaerobic conditions and in the
presence of metabolic inhibitors. 2,4-D was concentrated within the
roots to several times its external concentration, indicating that
2,4-D is taken up by roots by an adsorption mechanism and that energy
is required to maintain the integrity of the adsorbing surfaces of the
cell (Donaldson et al., 1973). Uptake of 14C-2,4-D was similar in
heterotrophic cell suspensions of soya bean ( Glycine max. L.) and
four perennial Glycine accessions, but the metabolism differed
considerably: soya bean metabolized only 7% of the absorbed radiolabel
at concentrations of 2 and 10 µmol/L, while the four perennials
metabolized at least 79% at 10 µmol/L and 64% at 50 µmol/L of 2,4-D.
The main metabolite identified by HPLC was the glycoside conjugate of
4-hydroxy-2,5-dichlorophenoxyacetic acid (White et al., 1990).
In a study of the effects of spray application parameters on foliar
uptake and translocation of the triethanolamine salt of 2,4-D in the
second leaf of Vicia faba after 24 h, the 2,4-D salt was applied to
provide doses of 4-420 g acid equivalent/ha. The quantity of the
triethanolamine salt taken up increased with increasing dose, with an
average uptake of 16 ± 5.5% and 70 ± 8.0% being translocated (Stevens
& Bukovac, 1987). The absorption, translocation, and metabolism of
14C-2,4-D was studied after application alone and with 14C-picloram
to leafy spurge in a greenhouse. Leafy spurge absorbed 34% of the
14C-2,4-D alone or with 14C-picloram. Eight minor metabolites were
observed, but only two were present at > 0.1% of the total radiolabel
recovered, and no attempt was made to identify them (Lym & Moxness,
1989). In the presence of 0.01 mmol/L 2,4-D, the uptake of potassium
by maize was reduced, and the effect increased at lower pH (Haunold &
Zsoldos, 1984).
The persistence of 2,4-D, as a formulation of mixed amine salts
containing 500 g acid equivalent/L formulation, as 400 g/L 2,4-D butyl
ester, and in other formulations, was examined after spraying at rates
of 1.1-112 g/ha on soya beans, tomatoes, and turnips in a growth
chamber for up to 35 days. When 2,4-D was applied to tomatoes at a
simulated drift level of 11.2 g/ha, the residues declined from 0.51 to
0.05 µg/g within 21 days. Similar results were seen for soya beans
when 2,4-D was applied at a rate of 0.05 g/ha, but at the highest
application rate (50 g/ha) the residue level was 0.01 µg/g after 35
days (Sirons et al., 1982). In similar experiments on Silene
vulgaris with [2-14C] -2,4-D, 65% of the dose, applied by
microsyringe to the youngest fully expanded leaf as eight to 10
droplets on either side of the midvein, was absorbed within 72h. More
than 30% of the absorbed radiolabel had translocated from the treated
leaf within 24 h after treatment. Metabolism did not appear to be an
important mechanism in conferring tolerance to 2,4-D (Wall et al.,
1991).
The uptake and phytotoxicity of vapours of 2,4-D butyl ester labelled
at the 2 position and in the ring, and also as a formulated product,
was studied in tomato, lettuce, and barley leaves. The relationship
between uptake, measured as the amount of radiolabel in the plant, and
vapour concentration was linear and independent of the duration of
exposure for both species. Twenty-four hours after exposure, the
leaves of tomatoes contained 63-93% of the total 2,4-D. The results
indicate that about 30% of the dose of 2,4-D butyl ester sprayed on
barley leaves would evaporate, while 70% would remain in the plant
(van Rensburg & Breeze, 1990; Breeze, 1990; Breeze et al., 1992).
Formulated 2,4-D ester markedly increased the penetration of labelled
difenzoquat into wild oat, while an amine formulation had no effect
(Sharma et al., 1976). A gas chromatographic method was developed for
analysis of 2,4-D in triticale over a growing season at two sites.
2,4-D dimethylamine was applied at 0.56 kg/ha after emergence, and
2,4-D was derivatized to its methyl ester. On the day after
application, residues on the order of 30 mg/kg were seen, which
decreased to undetectable levels in mature straw and seed (Cessna,
1990).
Leaf washings were used to determine the relative rate of uptake of
2,4-D dimethylamine salt from wheat in the field at a post-emergence
application rate of 487 g acid equivalent/ha (119 g acid equivalent/ha
as dicamba and 375 g acid equivalent/ha as 2,4-D) at the four- to
five-leaf growth stage. Samples for residue analysis were taken on the
day of application and two and seven days after. Leaf washings on the
day of application indicated that the canopy had intercepted 51% of
the application, while 65% was taken up by the crop (Cessna, 1993).
Simultaneous application of 2,4-D sodium salt at 0.33 kg ai/ha with
asulam had no significant effect on the absorption, translocation, and
biochemical action of asulam on bracken fern in the field or in the
laboratory with respect to frond density or the reduction of folate
levels, although there was noted antagonism with respect to protein
synthesis (Hinshalwood & Kirkwood, 1988).
Incubation of cultures of Daucus carota or Lactuca sativa cells
with 2,4-D showed that the toxicity of methyl mercury is partly
hormone-mediated and light-sensitive (Czuba, 1987, 1991).
The properties of leaf surfaces and their interactions with spray
droplets affect the foliar absorption and redistribution of 2,4-D.
Eleven plant species were treated with 14C-2,4-D in aqueous solutions
with and without surfactant. Uptake of 2,4-D did not correlate with
the presence of specialized leaf surface structures, cuticular
morphology, or distribution within dried deposits. Regression analysis
indicated that the epicuticular wax and cuticular membranes were the
major sinks for 2,4-D. Uptake of 2,4-D by apple, field bean, and maize
was significantly reduced in the presence of surfactant (Stevens &
Baker, 1987). In another study of the effect of physicochemical
properties and the role of surfactant on the uptake into and local
translocation within leaves sprayed with 14C-2,4-D, its movement into
rape and strawberry leaves was reduced with the addition of
surfactant; however, uptake was markedly increased in rape but little
changed in strawberry leaves. It was speculated that a higher
proportion of the chemical remains associated with the surfactant in
the leaf cuticle 24 h after application and would not be available for
movement. The surfactant may also alter the distribution of 2,4-D
between leaf cells, phloem tissue, and transpiration stream (Stevens
et al., 1988).
Seven days after application of 14C-2,4-D to three-week-old seedling
peas, alfalfa, and grapes as an aqueous solution or in treated soil,
significant amounts of 2,4-D had been absorbed into alfalfa (63.6%),
grapes (69.5%), and peas (56.7%) from the aqueous solutions, but the
comparable rates of uptake from treated soil were very much lower,
being 3.1% for alfalfa, 1.4% for grapes, and 2.9% for peas. 2,4-D was
absorbed at a higher rate by grapes grown in a greenhouse (61.4%) than
in the field (45.4%) (Al-Khatib et al., 1992).
In a study of the effects of different adjuvants on the foliar uptake
of 2,4-D sodium salt in wild oat and field bean, 14C-2,4-D was
applied as approximately 0.2-µl droplets containing about 0.5 g/L ai
with adjuvants at concentrations of 0.05-5.0 g/L. Concentrations of
surfactants > 0.05 g/L were necessary to increase the uptake of 2,4-D
into either plant substantially (Holloway & Edgerton, 1992).
The leaves of three-year old aspen (Populus tremens) in the
greenhouse were treated with the butoxyethanol ester of 2,4-D at 0.5
kg acid equivalent/L, a very high rate for this application, and the
residues of 2,4-D were measured for one year. The average residue one
day after treatment was 2300 mg/kg fresh weight but had fallen to 1300
mg/kg after 37 days; the average residue after one year was 870 mg/kg
(Eliasson, 1973). The 2,4-D amine salt and picloram potassium salt
were applied to the base of aspen saplings with a roller applicator as
a 7.3:1 w/w mixture (acid equivalent) at concentrations of 0.72, 2.14,
6.42, or 19.25 g/L. Measurements were made 39 and 84 days after
treatment. About 86% of the 2,4-D and picloram residues in the average
poplar sapling were located in the leaves, with 4.5% in the twigs and
9.1% in the stems (Cessna et al., 1989).
The effect of simulated rainfall on the phytotoxicity of foliar
applications of 0.2-0.6 kg/ha of the alkanolamine salt or the
butyoxyethanol ester of 2,4-D, depending upon the plant being treated,
was studied in the greenhouse. Simulated rainfall at 1, 5, 10, and 15
mm within 1 min of 2,4-D treatment reduced the phytotoxicity of the
alkanolamine salt of 2,4-D to a much greater extent than that of the
butoxyethanol ester. The effects ranged from complete elimination of
the phytotoxicity of the alkanolamine salt to soya beans to no
reduction in the phytotoxicity of the butoxyethanol ester to common
lambsquarters (Behrens & Elakkad, 1981).
In a study of the effect of 2,4-D on the cell number, fresh weight,
dry weight, and stored starch content of three species of
heterotrophic algae in vitro, increased concentrations of 2,4-D
resulted in decreased numbers and decreased starch content of
Polytoma uvella and Polytomella papillata cells, but
Prototheca chlorelloides was less sensitive (Pelekis et al., 1987).
Application of 2,4-D dimethylamine salt and 2,4-D butoxyethyl ester to
man-made freshwater ponds did not affect the numbers of moulds,
yeasts, or total fungi of Myriophyllum spicatum in the water column.
The mean differences between the control and treated ponds were
erratic and variable (Sherry, 1994). A formulated product of 2,4-D
used in silvicide practices significantly reduced the radial growth of
each of three species of ectomycorrhizal fungus that infect forest
trees in Canada at concentrations > 1000 ppm, and growth was
completely inhibited at concentrations < 5000 ppm (Estok et al.,
1989).
2. ENVIRONMENTAL LEVELS
2.1 Air
2,4-D acid has extremely low volatility, with a vapour pressure
ranging between 1.05 × 10-2 mm Hg (Grover & Kerr, 1978) and 1.4 ×
10-7 mm Hg (Chakrabarti & Gennrich, 1987) and with a Henry's law
constant of 1.3 × l0-10 atm.m3/mol at 25°C. Therefore, the likelihood
of 2,4-D occurring in air is remote.
In a study of the dissipation of the isooctyl ester of 2,4-D and its
acid metabolite in air, wheat, and soil components for up to 35 days
after application in Canada, the cumulative loss of the isooctyl ester
over the first five days was 20.8% of the amount applied, apparently
by volatilization. Measurements of airborne drift after ground
application demonstrated that only 3-8% of the applied herbicide
drifts as spray droplets when preparations with low volatility are
applied as large droplets; however, ultra-low-volume aerial
applications or use of more volatile esters may result in as much as
25-30% of the 2,4-D drifting off target. Drift can be reduced by
accurately following manufacturers' directions under proper
environmental conditions (Grover et al., 1985).
The levels of applied liquid and granular 2,4-D taken up by home
gardeners and household members who did not apply 2,4-D were monitored
in air samples inside homes and downwind of the application site. The
'unprotected' group had the highest exposures, but these were
consistently associated with spills and other contact. Residues of
2,4-D were detected in five of 76 air samples. A level of 0.006
mg/m3 was found inside a house after use of liquid 2,4-D and 0.01
mg/m3 in outside air after use of the granular preparation (Harris et
al.,1992).
Atmospheric deposition samples were collected weekly from early May to
early September 1984-87 in a small agricultural watershed near Regina,
Saskatchewan, Canada. As expected, maximum deposits occurred during
application, followed by a rapid decline. The amount of 2,4-D
deposited during the period of investigation ranged from 0.08 to 0.28%
of the amount applied (Waite et al. (1995).
2.2 Water
In a study reported as an abstract, 2,4-D was applied at 2.24 or 4.48
kg/ha to a watershed with an average declivity of 40% in North
Carolina, USA. The largest proportion of the applied dose that moved
from plots in surface runoff over eight months was 0.049%, and the
maximum possible losses of herbicide through the flume of the
watershed during the 160-day period was < 0.12% (Sheets et al.,
1972). 2,4-D was not detected in water associated with rice production
in Arkansas, USA, 28 days after application of 1.1 kg/ha (Johnson et
al., 1995). In a study of the soil persistence and lateral movement of
2,4-D after application as a stem-foliage spray for brush control on
two power-line rights-of-way, runoff water was analysed for up to 48
weeks after treatment. By week 4, residues of 0.3-1.9 mg/m3 2,4-D
were found in the treated areas, but no residues were found in water
8-11 weeks after spraying. Little lateral movement was detected (Meru
et al., 1990).
A review of the movement, persistence, and fate of phenoxy herbicides
in forests included the results of analyses in streams that drain
forest areas. It was suggested that the concentration of 2,4-D in
streams would be low after application in forest and rangeland,
consistent with the short persistence of 2,4-D (Norris, 1981). In a
study of the ability of several rivers in Western Australia to degrade
2,4-D, clear seasonal differences in both the concentrations of 2,4-D
and the degrading capacity of the water were seen, which correlated
with the amount of agricultural runoff, the sediment content of the
water, river flow, and the temperature of the water. Rivers receiving
agricultural runoff degraded 2,4-D faster than those receiving runoff
principally from forests (Watson, 1977).
The levels of 2,4-D in well water, watersheds, and streams draining
watersheds in Ontario, Canada, have been reported (Frank et al., 1979;
Frank & Sirons, 1980; Frank et al., 1982, 1987), and similar
evaluations of groundwater, surface water, and spring runoff were
reported in a watershed in Saskatchewan (Waite et al., 1992a,b).
Several sources of well-water contamination were investigated,
including spills from mixing, loading, and application and from
back-siphoning of spray solutions in 1979-84. The highest level of
2,4-D found after back-siphoning of spray solutions was 29 µg/L, which
decreased to < 0.1 µg/L after 17 days. 2,4-D was found in 19 of 255
wells investigated (Frank et al., 1987). The levels of 2,4-D were
measured in water samples taken from the mouth of the Grand, Saugeen,
and Thames rivers in Ontario, Canada, in 1981-90. The mean levels
never exceeded 0.7 µg/L in any given year in 1981-85 in either of the
three rivers. During the period 1986-90, no 2,4-D was found in the
Thames River, while an annual loading of 181 kg (said to be from an
unusual discharge during 1988) was found in the Grand River (Frank &
Logan, 1988; Frank et al., 1991). When 2,4-D was incubated in the
waters of four rivers in China for 56 days, an average of 76.1%
remained in the waters after application of the equivalent of 100 mg/L
of 14C-2,4-D, while 83.5% remained after application of 100 µg/L
(Wang et al., 1994).
In a small watershed in Saskatchewan, Canada, in 1985-87, 2,4-D was
found in pond water at a maximum level of 0.51 ppb, with a mean of
0.08 ppb over the three-year sampling period. The mean levels of 2,4-D
in spring runoff were 0.17 ppb during 1985 and 0.15 ppb in 1987 (Waite
et al., 1992b).
Water and bottom sediment were collected from eight stations in the
wetlands of the delta of the Axios, Loudias, and Aliakmon rivers in
Greece during 1992 and 1993. Samples were reported to have been
analysed within two to three days after sampling after storage at 4°C
before extraction. The levels of 2,4-D were 0.02-0.46 µg/L (Albanis et
al., 1994). In an extension of this work to include other products
from 1985-89, 2,4-D was found in about 1% of the samples of both
surface (modal range, < 0.2 - 0.3 mg/L; maximum, 2.1 mg/L) and
groundwater (0.11 - 0.2 mg/L) (Croll, 1991). Extensive analyses of
groundwater samples in the European Union showed that the levels
rarely exceeded 0.1 mg/L. The rare levels > 0.1 mg/L are likely to be
due to point sources such as spills or direct drainage from surface
water, rather than leaching through the soil profile. Further, point
sources are major contributors of residues due to poor handling
(mixing, loading, and clean-up) and poor application practices;
non-agricultural uses of 2,4-D are primarily responsible for residues
> 0.1 mg/L. In England, the National Rivers Authority has established
a maximum allowable concentration of 200 mg/L for non-ester forms of
2,4-D in freshwater and saltwater and an annual average concentration
of 40 mg/L. For esters of 2,4-D, the maximum allowable concentration
is 10 mg/L and the annual average concentration 1.0 mg/L. The vast
majority of the documented level of residues of 2,4-D salts and esters
are far below these values (Lewis et al., 1996; Isenbeck-Schroter et
al., 1997).
The maximum and mean levels of 2,4-D detected in 10 permanent and nine
semipermanent lakes in Saskatchewan, Canada, after the severe drought
of 1988 were 0.43 and 0.10 µg/L. 2,4-D was detected in < 10% of the
samples of sediment analysed. The frequency of detection was
significantly greater in brackish than in saline lakes (Donald &
Syrgiannis, 1995).
In a literature review on losses in runoff waters from agricultural
fields, the loss of 2,4-D from runoff from corn fields in Georgia,
USA, was 0.007-1.0% (Wauchope, 1978). The various processes for
removing 2,4-D from wastewaters have been reviewed (Cloutier, 1983).
One means that has been investigated is absorption onto peat. The
factors that affect the results include the concentration of 2,4-D,
pH, peat particle size, temperature, and the concentration of peat.
More than 90% of the 2,4-D was removed by peat under optimal
conditions, and the adsorption of 2,4-D was adequately described by
the Freundlich isotherm (Cloutier et al., 1985).
2.3 Soil
Samples of dykeland (silt loam) in Canada were collected from depths
of 20-30 cm and analysed for up to 385 days after surface application
of mixed amines or mixed esters of 2,4-D over 55 weeks. 2,4-D was
rapidly degraded from the amine formulation within 14-42 days after
application, and < 5% remained after 70 days. No residues of 2,4-D
from the amine application were found below 20-30 cm. At the maximum
recommended application rate (5.6 kg/ha), the concentration of 2,4-D
peaked at 1.84 mg/kg soil (top 10 cm) after 14 days and fell to 0.04
mg/kg after 70 days (Stewart & Gaul, 1977).
The dissipation of 2,4-D after application at 1 or 2 mg/kg of soil at
15, 40 and 90 cm was determined for up to 41 months in soya beans in
the field. Degradation of 2,4-D was rapid under aerobic conditions,
virtually all having dissipated within the first five months (Lavy et
al., 1973).
Residue were determined in soil and shallow groundwater after
long-term application of 2,4-D in southern Alberta, Canada, at rates
of 104-915g ai/ha to one site between 1976 to 1989 and to another from
1980 to 1989. Even with this extensive use, no residues of 2,4-D were
detected in soil sampled in 1991 (Miller et al., 1995). This finding
is consistent with the low persistence of 2,4-D in soil (e.g. Foster &
McKercher, 1973; Stewart & Gaul, 1977).
2.4 Plants
Studies conducted in North Carolina and California, USA, in which the
ethylhexyl ester and dimethylamine salt of 2,4-D were applied to grass
at a rate of 2.24 kg acid equivalent/ha, showed residues of up to 120
mg equivalent/kg wet weight for the ester and 153 mg equivalent/kg wet
weight for the salt. The levels declined after seven days to 28 and 60
mg acid equivalent/kg wet weight for the salt and ester, respectively
(Barney, 1995c,e; Hatfield, 1995e,f).
3. EFFECTS ON ORGANISMS IN THE LABORATORY AND THE FIELD
3.1 Microorganisms
The effect of 2,4-D on microorganisms has been evaluated in several
studies (Torstensson, 1978; Kuwatsuka & Miwa, 1989; Narain Rai, 1992;
Masson et al., 1993). Generally, the numbers of aerobic bacteria,
actinomycetes, and fungi in soils were not affected by 2,4-D applied
at rates corresponding to 25 ppm. The population of 2,4-D-degrading
microrganisms increased during the observed lag period in most soils
(Kuwatsuka & Miwa, 1989). After application of the dimethylamine salt
and the isooctyl ester at 0.95 kg/ha, the populations of fungi,
bacteria, or actinomycetes were not significantly affected by sampling
times but were affected by the form of 2,4-D applied: the ester
reduced bacterial population by 26.3%, the fungal population by 19.5%,
and and that of actinomycetes by 30%, while the dimethylamine salt
reduced the populations by only 10.1, 11.4, and 16%, respectively
(Narain Rai, 1992).
In studies of the effects of 2,4-D on the activities of invertase and
amylase and the respiration of various microorganisms in a sandy soil,
the extent of oxygen consumption increased with the concentration of
2,4-D (Tu, 1988). In a sandy loam in southwestern Ontario, Canada,
2,4-D did not affect microbial ammonification of organic nitrogen
indigenous to soil and was not toxic to various denitrifying
microorganisms (Tu, 1994).
3.2 Aquatic organisms
3.2.1 Plants
3.2.1.1 Toxicity
The effect of 2,4-D acid on various aquatic plants was evaluated in
screening studies conducted according to the guidelines of the US
Environmental Protection Agency. The concentrations evaluated were
1.91-2.13 mg/L, which approximated the maximum aquatic application
rate of 38 lb acid equivalent/acre (42.6 kg/ha). Exposure of
Skeletonema costatum (Hughes et al., 1994a) to 2.08 mg/L resulted in
10% stimulation relative to controls, while exposure of Navicula
pelliculosa (Hughes et al., 1994b), Anabaena flos-aquae (Hughes et
al., 1994c), and Lemna gibba (Hughes et al., 1994d) resulted in
inhibition of 24.3% (2.13 mg/L), 0.488% (2.02 mg/L), and 75% (1.91
mg/L), respectively. The isopropyl ester at 0.13 mg/L stimulated
Selenastrum costatum by 11% (Hughes et al., 1995).
In most of the studies summarized in Table 2, the acute toxicity of
2,4-D to aquatic plants was determined according to the guidelines of
the US Environmental Protection Agency for studies of nontarget
plants. The five-day EC50, EC25, and NOEC values were determined for
the ethylhexyl, butoxyethyl, and isopropyl esters ( Selenastrum only)
and the dimethylamine, diethanolamine, and triisopropanolamine salts
of 2,4-D. The EC50 for the ethylhexyl ester (chemically identical to
the isooctyl ester) ranged from a low of 0.23 mg/L for Skeletonema
costatum to a high of > 30 mg/L for Anabaena flos-aquae (Hughes,
1990a,b), the EC50 for the butoxyethyl ester was 1.66 mg/L for
Skeletonema costatum to 24.9 mg/L for Selenastrum capricornutum
(Hughes, 1989, 1990c), the EC50 for the dimethylamine salt ranged
from 5.28 mg/L for Navicula pelliculosa to 153 mg/L for Anabaena
flos-aquae (Hughes, 1990b,d), the EC50 for the diethanolamine salt
ranged from 11 mg/L for Selenastrum capricornutum to > 97 mg/L for
Navicula pelliculosa (Thompson & Swigert, 1993a,b), and the EC50
for the triisopropanolamine salt ranged from 82.4 mg/L for a marine
diatom to 133 mg/L for Anabaena flow-aquae (Hughes et al., 1994e,f).
The EC50 values for the higher aquatic plant Lemna gibba varied
with the form of 2,4-D, ranging from 0.5 mg/L for the ethylhexyl ester
to 3.3 mg/L for 2,4-D acid (Hughes 1990e, Hughes et. al., 1997).
3.2.1.2 Other effects on plants
Outdoor artificial ponds planted with Myriophyllum spicatum were
treated with the butoxyethanol ester of 2,4-D at a rate of 25 kg
ai/ha. M. spicatum had completely collapsed five days after
application. Significant decreases in the level of dissolved oxygen
and pH and an increased level of dissolved organic carbon was seen
seven days after treatment (Birmingham et al., 1983). The effect of
2,4-D was investigated on total chlorophyll production by Chlorella
vulgaris, Chlorococcum hypnosporum, Stigeoclonium tenue, Tribonoma
sp., Vaucheria geminata, and Oscillatoria lutea. At concentrations
of 0.001, 0.01, 0.05, 1, 10, and 100 mg/L, 2,4-D had no effect on
Table 2. Toxicity of 2,4-D to aquatic plants
Organism Flow/ Temperature Active ingredient Exposure 2,4-D conc. Reference
Static (°C) parameter (mg ai/L)
Green algae (Chlorella fusca) Static Acid 24-h EC50 88.9 Faust et al. (1994)
Algae (Skeletonema costatum) Static Ethylhexyl ester 5-d EC25 0.10a Hughes (1990a)b
5-d EC50 0.23
5-d NOE 0.1875
Butaxyethyl ester 5-d EC25 1.09 Hughes (1990c)b
5-d EC50 1.66
5-d NOEC 0.785
Marine diatom Static Dimethylamine salt 5-d EC25 15.79a Hughes (1990f)b
5-d EC50 36.60
5-d NOEC 96.25
Static Diethanolamine salt 5-d EC50 > 95a Thompson & Swigert
5-d NOEC 95 (1993c)
Static Triisopropanolamine salt 5-d EC25 60.1a Hughes et. al. (1994e)b
5-d EC50 82.4
5-d NOEC 50.4
Algae (Navicula pelliculosa) Static Ethylhexyl ester 5-d EC25 1.9a Hughes (1990g)b
5-d EC50 4.1
5-d NOEC 3.75
Freshwater diatom Static Butoxyethyl ester 5-d EC25 0.957 Hughes (1990h)b
5-d EC50 1.86
5-d NOEC 1.76
Table 2. (continued)
Organism Flow/ Temperature Active ingredient Exposure 2,4-D conc. Reference
Static (°C) parameter (mg ai/L)
Dimethylamine salt 5-d EC25 2.21a Hughes (1990d)b
5-d EC50 5.28
5-d NOEC 1.70
Static Diathanolamine salt 5-d EC50 > 97a Thompson & Swigert
5-d NOEC 97 (1993b)
Static Triisopropanolamine salt 5-d EC25 86.5a Hughes et al. (1994g)b
5-d EC50 124.0
5-d NOEC < 36.6
Algae (Anabaena flos-aquae) Static Ethylhexyl ester 5-d EC25 > 30a Hughes (1990b)b
5-d EC50 > 30
5-d NOEC > 30
Butoxyethyl ester 5-d EC25 5.96 Hughes (1990i)b
5-d EC50 6.37
5-d NOEC 3.14
Blue-green algae Static Dimethylamine salt 5-d EC25 38.5a Hughes (1990j)b
5-d EC50 153.0
5-d NOEC 67.86
Static Diethanolamine salt 5-d EC50 > 96a Thompson & Swigert
5-d NOEC 96 (1993d)
Static Triisopropanolamine salt 5-d EC25 77.4a Hughes et al. (1994h)b
5-d EC50 133
5-d NOEC < 99.6
Table 2. (continued)
Organism Flow/ Temperature Active ingredient Exposure 2,4-D conc. Reference
Static (°C) parameter (mg ai/L)
Lemna gibba G3 Static Acid 14-d EC25 1.721 Hughes et al. (1997)b
14-d EC50 3.30
14-d NOEC 2.029
Ethylhexyl ester 14-d EC25 0.15a Hughes (1990d)b
14-d EC50 0.50
14-d NOEC 0.1875
Butoxyethyl ester 14-d EC25 0.169 Hughes (1990e)b
14-d EC50 0.576
14-d NOEC 0.204
Duckweed Static Dimethylamine salt 14-d EC25 0.19a Hughes (19901)b
14-d EC50 0.58
14-d NOEC 0.27
Static Diethanolamine salt 14-d LC50 0.60a Thompson & Swigert
14-d NOEC < 0.079 (1993e)
Static Triisopropanolamine salt 14-d EC25 0.794a Hughes et al. (1994f)b
14-d EC50 2.37
14-d NOEC 0.354
Algae (Selenastrum Static 22-26 Acid 5-d EC25 29.0a Hughes (1990m)b
capricornutum) 5-d EC50 33.2
5-d NOEC 26.4
Acid 96-h EC50 25.9 St Laurent et al.
96-h EC50 24.2 (1992)
Table 2. (continued)
Organism Flow/ Temperature Active ingredient Exposure 2,4-D conc. Reference
Static (°C) parameter (mg ai/L)
Green algae Static Ethylhexyl ester 5-d EC25 > 22.7a Hughes (1990n)b
5-d EC50 > 22.7
5-d NOEC 15.0
Butoxyethyl ester 5-d EC25 10.5 Hughes (1989)b
5-d EC50 24.9
5-d NOEC 12.5
Static Dimethylamine salt 5-d EC25 25.9a Hughes (1990o)b
5-d EC50 66.5
5-d NOEC 19.2
Static Isopropyl ester 5-d EC50 > 0.13a Hughes et al. (1995)
(tier I)
Static Diethanolamine salt 5-d EC50 11.0a Thompson & Swigert
5-d NOEC 0.5 (1993a)
Static Triisopropanolamine salt 5-d EC25 66.8a Hughes (1994)b
5-d EC50 103.0
5-d NOEC 55.4
a Reliable data based on measured concentrations
b Study conducted for compliance with FIFRA registration by the US Environmental Protection Agency
Table 3. Toxicity of 2,4-D to freshwater invertebrates
Organism Flow/ Temp Alkalinity Hardness pH Active Exposure 2,4-D conc. Reference
Static (°C) ingredient parameter (mg ai/L)
Oligochaete worm Flow 20 30 30 7.8 Free acid 48-h LC50 122.2 Bailey & Liu (1980)
(Lumbriculus 96-h LC50 122.2
variegatus) 96-h LC0 86.7
Water flea (Daphnia Static 19-21 51 78 8.4 Isooctyl ester 48-h LC50 5.2 Alexander et al.
magna) (1983a)a
Static 20 7.0-8.2 Acid 48-h LC50 36.4 Alexander et al.
(1983b)a
Static 20 77-84 90-108 7.8-8.0 Dimethylamine salt 48-h LC50 184 Alexander et al.
(1983c)a
Static 25 85 100 7.6 Acid 48-h LC50 247.2 McCarty & Batchelder
(1977)a
Flow 19-21 170-175 160-180 4.0-8.5 Acid 21-d NOEL 79b Ward & Boeri (1991a)a
21-d LOEL 151
21-d MATC 109
21-d EC50 235
Flow 18.9-24.5 22-29 65-68 6.8-7.6 Dimethylamine salt 21-d NOEL 27.5b Ward (1991a)a
21-d LOEL 59.6
21-d MATC 40.5
3-d LC50 130-243
Flow 19-21 160-185 180 7.8-8.6 Ethylhexyl ester 21-d NOEL 0.015b Ward & Boeri (1991b)a
21-d LOEL 0.027
21-d MATC 0.020
21-d EC50 0.13
2-d LC50 > 0.2
Table 3. (continued)
Organism Flow/ Temp Alkalinity Hardness pH Active Exposure 2,4-D conc. Reference
Static (°C) ingredient parameter (mg ai/L)
Flow 20 7.6 Acid 24-h EC50 1.124 Lilius et al. (1995)
Static 17 39 7.2 Dimethylamine salt 48-h LC50 > 100 Mayer & Ellersieck
(1986)
Static 20 8.4-8.6 Free acid 96-h LC50 417.8 Presing (1981)
Static 20 8.4-8.6 Sodium salt 96-h LC50 932.1 Presing (1981)
Flow Diethanolamine salt 48-h EC50 > 100b Graves & Peters
21-d NOEC 23.6 (1991a)
Flow Isopropylamine salt 48-h LC50 583 Alexander et al.
(1983d)
Flow Triisopropanolamine 48-h LC50 748 Mayes (1989)
Flow Isopropyl esterc 48-h EC50 2.6b Drottar & Swigert
(1996a)
Flow Butoxyethyl ester 48-h LC50 7.2 Alexander et al.
(1983e)
Water flea Flow 25 81 57.07 8.18 Acid 48-h LC50 236 Oris et al. (1991)
(Ceriodaphnia dubia)
Water flea (Daphnia Flow 20 7.6 Acid 24-h EC50 1.47 Lilius et al. (1995)
pulex)
Copepod (nauplius Static 20 31.6 70 6.7 Free acid 96-h LC50 8.72 Robertson (1975)
larva)
(Cyclops varnalis) Static 20 31.6 70 6.7 Alkanolamine 96-h LC50 54.8 Robertson (1975)
Scud (Gammarus Static 15 272 7.4 Dimethylamine salt 24-h LC50 > 100 Mayer & Ellersieck
fasciatas) Static 15 272 7.4 Dimethylamine salt 96-h LC50 > 100 (1986)
Table 3. (continued)
Organism Flow/ Temp Alkalinity Hardness pH Active Exposure 2,4-D conc. Reference
Static (°C) ingredient parameter (mg ai/L)
Freshwater prawn Static 27 113.9 7.5 Sodium salt 24-h LC50 2342 Omkar & Shukla (1984)
(Macrobranchium Static 27 113.9 7.5 Sodium salt 48-h LC50 2309
lamerrei) Static 27 113.9 7.5 Sodium salt 72-h LC50 2267
Static 27 113.9 7.5 Sodium salt 96-h LC50 2224
Freshwater prawn Static 27 112.7 7.5 Sodium salt 24-h LC50 2644 Omkar & Shukla (1984)
(Macrobranchium Static 27 112.7 7.5 Sodium salt 48-h LC50 2536
naso) Static 27 112.7 7.5 Sodium salt 72-h LC50 2435
Static 27 112.7 7.5 Sodium salt 96-h LC50 2397
Freshwater prawn Static 28 112.7 7.5 Sodium salt 24-h LC50 2474 Omkar & Shukla (1984)
(Macrobranchium Static 28 112.7 7.5 Sodium salt 48-h LC50 2381
dayanum) Static 28 112.7 7.5 Sodium salt 72-h LC50 2333
Static 28 112.7 7.5 Sodium salt 96-h LC50 2275
Red swamp crayfish Static 20 100 8.4 Alkanolamine salt 96-h LC50 1389 Cheah et al. (1980)
(immature)
(Procambarus clarki)
Midge (larva) 15 78-95 55 7.3-7.8 Dimethylamine salt 24-h IC50 1490 Bunting & Robertson
(Chaoborus 15 78-95 55 7.3-7.8 Dimethylamine salt 96-h IC50 890 (1975)
punctipennis)
a Study conducted in accordance with international guidelines and good laboratory practice
b Reliable data based on measured concentrations
c Formulated product
chlorophyll production by these algae (Ramirez Torres & O'Flaherty,
1976).
A concentration of 10 µg/ml of the sodium salt of 2,4-D (80% ai) in
liquid culture medium stimulated the growth and nitrogen fixation of
the heterocystous bloom-forming blue-green alga Anabaenopsis
raciborskii. The alga tolerated up to 800 µg/ml (Das & Singh, 1977).
Technical-grade 2,4-D increased the growth and nitrogen assimilation
of Azolla mexicana and its phycobiont Anabaena azollae when 2,4-D
was added at 1 ppm to the nitrate-containing medium, whereas all of
the plants died within 10 days at 10 ppm 2,4-D (Holst et al., 1982).
The phytotoxicity of 2,4-D to Selenastrum capricornutum was compared
by the microplate and flask bioassay methods. The EC50 values (95%
confidence interval) found with the two methods were similar, being
25.9 (23.8-28.3), r2 = 0.60 and 24.2 (23.7-24.7), r2= 0.93;
respectively (St Laurent et al., 1992). The phytotoxicity of the
'expected environmental concentration' (2.197 mg/L) of 14C-2,4-D acid
to 10 kinds of algae was low, and it caused < 50% inhibition of the
growth of Lemna (Peterson et al., 1994).
3.2.2 Invertebrates
3.2.2.1 Toxicity
The toxicity of 2,4-D and its salts and esters to freshwater
invertebrates and marine organisms is summarized in tables 3 and 4.
Numerous studies have been conducted of the toxicity in freshwater and
estuarine aquatic invertebrates. Studies conducted according to the
guidelines of the US Environmental Protection Agency by the 'Industry
Task Force II on 2,4-D Research Data' are also summarized in the
tables.
The 21-day EC50 of 2,4-D in the water flea, Daphnia magna, in a
flow-through system ranged from 0.13 mg/E for the isooctyl ester (Ward
& Boeri,1991b) to 235 mg/L for the acid (Ward & Boeri, 1991 a). No
EC50 was available for the dimethylamine salt. Comparable 21-day NOEL
values were found for the 2-ethylhexyl ester (0.015 mg/L; Ward &
Boeri, 1991b), the dimethylamine salt (27.5 mg/L; Ward, 1991a), and
the acid (79 mg/L; Ward & Boeri, 1991a).
The 48-h LC50 ranged from > 100 mg/L for the dimethylamine salt
(Mayer & Ellersieck, 1986) to 932.1 mg/L for the sodium salt (Presing,
1981 ). The 48-h LC50 values for the isopropylamine and
triisopropanolamine salts fall within this range; the 48-h EC50 for
the isopropyl ester was 2.6 mg/L (Drottar & Swigert, 1996b) and that
for the butoxyethyl ester was 7.2 mg/L (Alexander et al., 1983e).
Clearly, the isooctyl ester, which is chemically identical to the
ethylhexyl ester, is more toxic to invertebrates than the acid or the
amine salt. The results of the studies of the Industry Task Force II
on 2,4-D Research are consistent with those in the literature.
The toxicity of 2,4-D to the freshwater invertebrates Oligochaete
worms, Ceriodaphnia dubia, Daphnia pulex, copepods, scuds, and other
species was generally similar to that in Daphnia magna, i.e. 10->
100 mg/L, except for the sodium salt. The 24-, 48- and 72-h LC50
values for the sodium salt of 2,4-D in static systems were 2200-2700
mg/L for the freshwater prawn Macrobranchium lamerrei. The
dimethylamine salt was also relatively non-toxic to midge larvae
(Chaoborus punctipennis), with values ranging ranging from 890 mg/L
for the 96-h LC50 to 1490 mg/L for the 24-h LC50 (Omkar & Shukla,
1984).
A considerable body of data is available on the toxicity (shell
deposition for oysters) of 2,4-D in estuarine aquatic invertebrates
(Table 4). The 96-h EC50 for eastern oysters (Crassostrea
virginica) ranged from > 0.21 mg/L for the 2-ethylhexyl ester (Ward
& Boeri, 1991f) to 179 mg/L for the triisopropanolamine salt (Dionne,
1990b), and the values for the acid, diethanolamine salt,
dimethylamine salt, and isopropylamine salt fall within this range
(Wade & Overman, 1991; Graves & Peters, 1991c; Ward, 1991c; Ward et
al., 1993); the EC50 for formulated Esteron 99 containing 66% 2-
ethylhexyl ester is also within this range (Ward & Boeri, 1991e). The
48-h EC50 for the dimethylamine salt, > 210-< 320 mg/L, indicates
considerably lower toxicity (Heitmuller, 1975).
The esters of 2,4-D are clearly more toxic to invertebrate species
such as the tidewater silverside (Menidia beryllina), Atlantic
silverside (Menidia menidia), grass shrimp (Palaemonetes puqio),
pink shrimp (Panaeus duorarum), and Dungeness crab (Cancer
magister) than is the dimethylamine salt or the acid (Table 4). The
same is true for formulated 2-ethylhexyl ester (Ward & Boeri,
1991c,h).
3.2.2.2 Other effects on invertebrates
Grass shrimp (Palaemonetes pugio) showed avoidance reactions to
water containing 2,4-D butoxyethanol ester at 1 or 10 mg/L (Hansen et
al., 1973). The phototactic behaviour of larval estuarine grass
shrimps exposed to 49% 2,4-D amine was reduced at all stages
evaluated. Adult shrimp were not affected by doses up to 5000 mg/L. No
correlation with dose was seen for egg hatchability (Moyer, 1975).
The toxicity of the diethanolamine salt of 2,4-D to Daphnia magna
was examined in a life-cycle toxicity test under flow-through
conditions conducted according to the guidelines of the US
Enviromnental Protection Agency. The daphnia were less than 24h old at
the start. The NOEC was 23.6 mg ai/L (Holmes & Peters, 1991). In a
similar study, the toxicity of the butoxyethyl ester of 2,4-D to
Daphnia magna was tested under flow-through conditions for 21 days.
The maximum acceptable toxicant concentration was 0.70-0.29 mg/L, and
the NOEC was 0.29 mg/L (Gersich et al., 1989). The effect of 2,4-D and
its dimethylamine salt on the reproduction of Daphnia magna was
tested in a study conducted according to the European guidelines.
Parental survival was not affected at 21.5, 46.2, or 100 mg/L, but
reproduction, expressed as the number of young per surviving adult,
was significantly reduced at 100 and 215 mg/L and in the control
dimethylamine group (Mark & Hantink-de Rooy, 1989).
3.2.3 Vertebrates
3.2.3.1 Toxicity
Studies of the effects of the potassium or sodium salt on early
life-stages, embryos, larvae, and fish four days after hatching were
conducted on several species in flow-through systems (Table 5). The
12-, 24-, 36-, and 48-h LC50 values for the sodium salt in embryonic
bleak (Alburnus alburnus) were 12.9-159.4 mg/L (Biro, 1979). The
LC50 values for the potassium salt of 2,4-D in embryos and four-day-
old fish of various species in flow-through systems ranged from 4.2
mg/L for the 27-day LC50 to > 201 mg/L at four days (Birge et al.,
1979).
In studies conducted according to the guidelines of the US
Environmental Protection Agency, 2,4-D acid and ethylhexyl ester had
no effect on the early life stages, embryo hatch, larval weight, or
larval length of the fathead minnow (Pimephales promelas) at
concentrations of 12.6-102 mg/L for up to 32 days (acid). The 32-day
NOEC for the acid was 63.4 mg/L (Mayes et al., 1990a,b), comparable to
the 33-day NOEC for the diethanolamine salt of 29.1 mg/L (Graves &
Peters, 1991e). The ethylhexyl ester was more toxic, with a 32-day
NOEC of 0.12 mg/L (Mayes et al., 1990a), and the 96-h LC50 of the
butoxyethyl ester was 0.93 mg/L (Mayes et al., 1989a).
Many data are available on the acute toxicity of 2,4-D and its various
formulations to fish. Rainbow trout (Oncorhynchus mykiss) and
bluegills (Lepomis machrochirus) are the species most often tested
for the US Environmental Protection Agency, and the results are
summarized in Table 6 and by Mayer and Ellersieck (1986).
In static systems, the dimethylamine salt and the acid had similar
96-h LC50 values in rainbow trout of 250 and 358 mg/L, respectively
(Alexander et al., 1983b,c). The 24-, 48-, and 96-h LC50 values for
the amine salt to grass carp (Ctenopharyngodon idella) ranged from
1313 to 3080 mg/L (Tooby et al., 1980) and those for mosquito fish
(Gambusia affinis) from 405 to 500 mg/L (Johnson, 1978).
The esters of 2,4-D are generally more toxic to fish than are the
various salts or the acid. In flow-through systems, the 96-h LC50
values for rainbow trout ranged from 0.69 mg/L for the isopropyl ester
(Drottar & Swigert, 1996a) to 7.2 mg/L for the ethylhexyl ester (Mayes
et al., 1990a). The 96-h LC50 values for the various salts ranged
from > 120 mg/L for the diethanolamine salt (Graves & Peters, 1991f)
to 2840 mg/L for the isopropanolamine salt (Alexander et al., 1983d).
The values for the dimethylamine salt (Alexander et al., 1983c),
triisopropanolamine salt (Mayes et al., 1989b), and acid (Alexander et
al., 1983b) also fall within this range. The 96-h LC50 values for
rainbow trout fry and smolts were very low, irrespective of the
Table 4. Toxicity of 2,4-D to estuarine and marine organisms
Oxganism Flow/ Temperature Salinity pH Active ingredient Exposure 2,4-D conc. Reference
Static (°C) parameter (mg ai/L)
Bay mussel (Mytilus 17.2-18.6 22.9-24.5 6.4-7.8 Free acid 96-h LC50 259 Liu & Lee (1975)
edulis) 96-h EC50 262
(attachment)
(trocophore larva) 17.2-18.6 22.9-24.5 6.4-7.8 Free acid 48-h EC50 211.7 Liu & Lee (1975)
(normal
development)
Tidewater silverside Flow 21.2-22.8 6.9-8.0 Ethylhexyl estera 96-h LC50 > 1.1b Ward & Boeri
(Menidia beryllina) 96-h NOEC 1.1 (1991c)c
Flow 21.7-22.8 7.8-8.4 Isooctyl ester 96-h LC50 > 0.24b Ward & Boeri
96-h NOEC 0.24 (1991d)c
Flow 21.5-22.7 8.2-8.5 Dimethylamine salta 96-h LC50 469b Ward (1991b)c
Flow 21.3-22.7 5.98.3 Acid 96-h LC50 175b Vaishnav et al.
(1990a)c
Flow Triisopropanolamine 96-h LC50 376b Sousa (1990a)
96-h NOEC 60.7
Atlantic silverside Flow Diethanolamine salt 96-h LC50 > 118b Graves & Peters
(Menidia menidia) (1991b)
Flow Isopropylamine salt 96-h LC50 298b Sousa (1990b)
96-h NOEC < 26.9
Eastern oyster Static 19-21 7.5-8.5 Dimethylamine salta 48-h EC50 > 210-< 320 Heitmuller (1975)c
(Crassostrea
virginica) (larva) Flow 21.2-23.2 7.7-8.0 Dimethylamine salta 96-h EC50 136b Ward (1991c)c
(shell deposition) 96-h NOEC 40.6
Flow 18.1-19.6 7.3-7.8 Ethylhexyl estera 96-h EC50 >0.71b Ward & Boeri
96-h NOEC 0.71 (1991e)c
Flow 15.3-17.7 7.3-7.4 Ethylhexyl ester 96-h EC50 > 0.21b Ward & Boeri
96-h NOEC 0.21 (1991f)c
Table 4. (continued)
Oxganism Flow/ Temperature Salinity pH Active ingredient Exposure 2,4-D conc. Reference
Static (°C) parameter (mg ai/L)
Flow 20-21 3.7 8.1 Acid 96-h EC50 146b Wade & Overman
(1991)c
Flow 21.1-22.1 7.0-8.0 Acid 96-h EC50 57b Ward et al.
96-h NOEC 30 (1993)c
Flow Diethanolamine salt 96-h EC50 > 112b Graves & Peters
96-h NOEC < 6.9 (1991c)
Flow Isopropylamine salt 96-h EC50 63.9b Dionne (1990a)
96-h NOEC 31.7
Flow Triisopropanolamine 96-h EC50 179b Dionne (1990b)
96-h NOEC 77.2
Flow 29 25 Isooctyl ester 96-h EC50 1.0 Mayer (1987)
(shell growth)
Flow 28 25 Propylene glycol 96-h EC50 0.055 Mayer (1987)
butyl ethyl ester (shell growth)
Copepod (Nitocra 21 7 7.8 Butoxyethanol 96-h LC50 3.1 Linden et al.
spinipes) (1979)
Grass shrimp Flow 21.2-22.8 7.6-7.8 Ethylhexyl estera 96-h LC50 > 1.4b Ward & Boeri
(Palaemonetes pugio) 96-h NOEC > 1.4 (1991g)c
Flow 21.2-22.9 7.3-8.1 Ethylhexyl ester 96-h LC50 > 0.14b Ward & Boeri
96-h NOEC 0.14 (1991h)c
Pink shrimp Flow 3.6-8.3 21.1-22.6 Acid 96-h LC50 554b Vaishnav et al.
(Panaeus duorarum) (1990b)c
Static 19-21 7.5-8.5 Dimethylamine salta 96-h NOEC > 1000 Heitmuller
(1975)c
Flow 20.7-22.6 8.2-8.5 Dimethylamine salta 96-h LC50 181b Ward (1991d)c
96-h NOEC 65
Flow Diethanolamine salt 96 h LC50 > 99.6b Graves & Peters
96-h NOEC (1991d)
Table 4. (continued)
Oxganism Flow/ Temperature Salinity pH Active ingredient Exposure 2,4-D conc. Reference
Static (°C) parameter (mg ai/L)
Flow Isopropylamine salt 96-h LC50 623b Sousa (1990c)
96-h NOEC 140
Flow Triisopropanolamine 96-h LC50 744b Sousa (1990d)
96-h NOEC 410
Brown shrimp
(Panaeus aztecus)
(juvenile) Static 26 30 Butoxyethanol 48-h LC50 5.6 Mayer (1987)
(adult) Flow 29 26 Isooctyl 48-h LC50 0.48
Dungeness crab
(Cancer magister)
(first zoel) Static 13 25 Acid (technical) 96-h LC50 > 10 Caldwell (1977)
(first instar
juvenile) Static 13 25 Acid (technical) 96-h LC50 > 100
Fiddler crab Static 19-21 7.5-8.5 Dimethylamine salta 96-h NOEC >1000 Heitmuller (1975)c
(Uca pugilator)
Blue crab (juvenile)
(Callinectes
sapidus) Static 24 29 Propylene glycol 48-h LC50 2.8 Mayer (1987)
butyl ethyl ester
Estuarine crab
(Chasmagnathus
granulata)
(first zoeI) Static 6.5-7.5 Isobutoxyethanol 24-h LC50 4.5-13.5 Rodriguez & Amin
ester (1991)
48-h LC50 1.06
72-h LC50 0.43
96-h LC50 0.30
Table 4. (continued)
Oxganism Flow/ Temperature Salinity pH Active ingredient Exposure 2,4-D conc. Reference
Static (°C) parameter (mg ai/L)
(juvenile) 24-h LC50 > 6.4
48-h LC50 > 6.4
72-h LC50 5.55
96 h LC50 2.89
12 6.5-7.5 Isobutoxyethanol 24-h LC50 > 10 Rodriguez &
ester 48 h LC50 > 10 Lombardo (1991)
72-h LC50 6.73
96-h LC50 3.37
(adult) 22-24 12 6.5-7.5 Isobutoxyethanol 4-week LC50 > 50 Rodriguez et al.
(juvenile) ester 30.36 (1992)
Estuarine crab 12 6.5-7.5 Isobutoxyethanol 24-h LC50 > 400 Rodriguez &
(Uca uruguayensis) ester 48-h LC50 > 400 Lombardo (1991)
72-h LC50 213
96-h LC50 130
22-24 12 6.5-7.5 Isobutoxyethanol 4-week LC50 > 30 Rodriguez et al.
ester (1992)
a Formulated product
b Reliable data based on measured concentrations
c Study conducted according to international guidelines and good laboratory practice
Table 5. Toxicity of 2,4-D to fish in early life stages
Organism Flow/ Temp Alkalinity Hardness pH Active ingredient Exposure 2,4-D conc. Reference
Static (°C) parameter (mg ai/L)
Fathead minnow Flow 24.9-25.7 41-50 70-75 7.1-7.6 2-Ethylhexyl ester 32-d NOEC 0.16 Mayes et al.
(Pimephales promelas) 32-d MATC 0.12m (1990a)a
Flow 25.0-25.6 23-49 68-74 6.5-7.7 Acid 32-d NOEC 63.4m Mayes et al.
32-d MATC 80.4 (1990b)a
Flow 25.3-25.8 47-52 70-76 7.1-7.8 Dimethylamine salt 28-d NOEC 17.1b Dill et al.
28-d MATC 22 (1990)a
Flow Diethanolamine salt 33-d NOEC 29.1b Graves & Peters
(1991e)
Flow Butoxyethyl ester 96-h LC50 0.95b Mayes et al.
(1989a)
Bleak (Alburnus Sodium salt 12-h LC50 159.4 Biro (1979)
alburnus) (embryo) 24-h LC50 129.0
36-h LC50 63.9
48-h LC50 12.9
(larvae) 12-h LC50 111.2
24-h LC50 70.6
36-h LC50 62.1
48-h LC50 51.6
Goldfish (Carassius Flow 18.2-25.8 66.7 53.3 7.84 Potassium salt 4-day LC50 > 187 Birge et al.
auratus) (embryo) Flow 18.2-25.8 65.3 197.5 7.78 Potassium salt 4-day LC50 > 201 (1979)
(4-day post-hatch) Flow 18.2-25.8 66.7 53.3 7.84 Potassium salt 8-day LC50 133.1
Flow 18.2-25.8 65.3 197.5 7.78 Potassium salt 8-day LC50 119.1
Largemouth bass
(Micropterus salmoides)
(embryo) Flow 18.2-25.8 66.7 53.3 7.84 Potassium salt 3.5-day LC50 165.4 Birge et al.
Flow 18.2-25.8 65.3 197.5 7.78 Potassium salt 3.5-day LC50 160.7 (1979)
(4-day post-hatch) Flow 18.2-25.8 66.7 53.3 7.84 Potassium salt 7.5-day LC50 108.6
Flow 18.2-25.8 65.3 197.5 7.78 Potassium salt 7.5-day LC50 81.6
Table 5. Toxicity of 2,4-D to fish in early life stages
Organism Flow/ Temp Alkalinity Hardness pH Active ingredient Exposure 2,4-D conc. Reference
Static (°C) parameter (mg ai/L)
Rainbow trout Flow 18.2-25.8 66.7 53.3 7.84 Potassium salt 23-day LC50 11.0 Birge et al.
(Oncorhynchus mykiss) (1979)
(embryo) Flow 18.2-25.8 65.3 197.5 7.78 Potassium salt 23-day LC50 4.2
(4-day post-hatch) Flow 18.2-25.8 66.7 53.3 7.84 Potassium salt 27-day LC50 11.0
Flow 18.2-25.8 65.3 197.5 7.78 Potassium salt 27-day LC50 4.2
a Results generated in accordance with international guidelines and good laboratory practice
b Reliable data based on measured concentrations
Table 6. Toxicity of 2,4-D to fish
Organism Flow/ Temp Alkalinity Hardness pH Active ingredient Exposure 2,4-D conc. Reference
Static (°C) pmameter (mg ai/L)
Rainbow trout Static 19-21 51 78 8.4 Isooctyl ester 96-h LC50 > 5 Alexander
(Oncorhynchus mykiss) et al. (1983a)a
Flow 11.7-12.7 46 70 7.4-7.7 Ethylhexyl esterb 96-h LC50 7.2c Mayes et al.
192-h LC50 4.6 (1990c)a
Static 12-20 7.0-8.2 Acid 96-h LC50 358 Alexander et
al. (1983b)a
Static 19-21 47 7.1 Dimethylamine salta 96-h TL50 377 Bentley (1974)a
96-h NOEL 210
Static 12-20 77-84 90-108 7.8-8.0 Dimethylamine salt 96-h LC50 250 Alexander et al.
(1983e)a
Flow Dimethylamine salt 48-h LC50 240 Bogers & Enninger
(1990a)
72-h LC59 240
96-h LC50 240
Static 14 Soft 6.3 Diethanolamine 96-h LC50 409 Wan et al. (1991)
6.3 Isooctyl ester 96-h LC50 167
Inter- 7.5 Diethanolamine 96-h LC50 511
mediate 7.5 Isooctyl ester 96-h LC50 164
Hard 8 Diethanolamine 96-h LC50 744
8 Isooctyl ester 96-h LC50 79
Flow 18.2-25.8 66.7 53.5 7.84 Potassium salt 27-d LC50 0.032c Birge et al.(1979)
Flow 18.2-25.8 65.3 197.5 7.78 Potassium salt 27-d LC50 0.022c Birge et al.(1979)
Flow Diethanolamine salt 96-h LC50 > 120c Graves & Peters
(1991f)
Flow Isopropylamine salt 96-h LC50 2840 Alexander et al.
(1983d)
Flow Triisopropanolamine 96-h LC50 317 Mayes et al.
(1989c)
Flow Butoxyethyl ester 96-h LC50 2.0 Alexander et al.
(1983d)
Flow Isopropyl ester 96-h LC50 0.69c Drottar & Swigert
(1996a)
Table 6. (continued)
Organism Flow/ Temp Alkalinity Hardness pH Active ingredient Exposure 2,4-D conc. Reference
Static (°C) pmameter (mg ai/L)
Flow Isopropyl esterb 96-h LC50 0.78c Drottar & Swigert
(1996e)
(fingerlings) Flow 15 7.6 21.6 4.54 Acid 96-h LC50 < 100c Doe et al. (1988)
5.6 96-h LC50 < 400
6.8 96-h LC50 < 1000
8.48 96-h LC50 > 1000
(fry) Flow 15 18 17 7.1 Butoxyethanol esterb 96-h LC50 0.518c Finlayson & Verrue
Total 2,4-D 96-h LC50 0.642c (1985)
Propylene glycol 96-h LC50 0.329c
butyl ethyl esterb
Total 2,4-D 96-h LC50 0.514c
(smolts) Flow 15 18 17 7.1 Butoxyethanol esterb 96-h LC50 0.468c Finlayson & Verrue
(1985)
Total 2,4-D 96-h LC50 1.338c
Flow 15 18 17 7.1 Propylene glycol 96-h LC50 0.342c Finlayson & Verrue
butyl ethyl esterb (1985)
Total 2,4-D 96-h LC50 1.555c
Loading factor Static 14 18 17 7.1 Butoxyethanol esterb 96-h LC50 1.206c Finlayson & Verrue
4.2 g fish/L Total 2,4-D 96-h LC50 1.422c (1985)
Loading factor Static 15 18 17 7.1 Butoxyethanol esterb 96-h LC50 3.689c Flnlayson & Verrue
9.8 g fish/L Total 2,4-D 96-h LC50 4.487c (1985)
Bluegill (Lepomis Static 19-21 51 78 8.4 Isooctyl ester 96-h LC50 > 5 Alexander et al.
machrochirus) (1983a)a
Static 12-20 7.0-8.2 Acid 96-h LC50 263 Alexander et al.
(1983b)a
Static 19-21 47 7.1 Dimethylamine saltb 96-h TL50 387 Bentley (1974)a
96-h NOEL 280
Static 12-20 77-84 91-108 7.8-8.0 Dimethylamine salt 96-h LC50 524 Alexander et al.
(1983e)a
Flow Diethanolamine salt 96-h LC50 > 121c Graves & Peters
(1991g)
Table 6. (continued)
Organism Flow/ Temp Alkalinity Hardness pH Active ingredient Exposure 2,4-D conc. Reference
Static (°C) pmameter (mg ai/L)
Flow Isopropylamine salt 96-h LC50 1700 Alexander et al.
(1983d)
Flow Triisopropanolamine 96-h LC50 432 Mayes et al.
(1989a)
Flow Butoxyethyl ester 96-h LC50 0.61 Alexander et al.
(1983d)
Flow Isopropyl ester 96-h LC50 0.31c Drottar & Swigert
(1996c)
Flow Isopropyl esterb 96-h LC50 0.31c Drottar & Swigert
(1996d)
Fathead minnow Static 19-21 51 78 8.4 Isooctyl ester 96-h LC50 >5 Alexander et al.
(Pimephales promelas) (1983a)a
Static 12-20 7.0-8.2 Acid 96-h LC50 320 Alexander et al.
(1983b)a
Static 12-20 77-84 90-108 7.8-8.0 Dimethylamine salt 96-h LC50 344 Alexander et al.
(1983e)a
Isopropylamine salt 96-h LC50 2180 Alexander et al.
(1983d)
Butoxyethyl ester 96-h LC50 2.5 Alexander et al.
(1983d)
Striped bass Static 20 50 7.2 Free acid 24-h LC50 85.6 Rehwoldt et al.
(Morone saxatillis) 96-h LC50 70.1 (1977)
Banded killifish Static 20 50 7.2 Free acid 24-h LC50 306.2 Rehwoldt et al.
(Fundulus diaphanus) 96-h LC50 26.7 (1977)
Pumpkinseed sunfish Static 20 50 7.2 Free acid 24-h LC50 120 Rehwoldt et al.
(Lepomis gibbosus) 96-h LC50 94.6 (1977)
White perch Static 20 50 7.2 Free acid 24-h LC50 55.5 Rehwoldt et al.
(Roccus americanus) 96-h LC50 40 (1977)
Table 6. (continued)
Organism Flow/ Temp Alkalinity Hardness pH Active ingredient Exposure 2,4-D conc. Reference
Static (°C) pmameter (mg ai/L)
American eel Static 20 50 7.2 Free acid 24-h LC50 427.2 Rehwoldt et al.
(Anguilla rostrata) 96-h LC50 300.6 (1977)
Carp (Cyprinus carpio) Static 20 50 7.2 Free acid 24-h LC50 175.2 Rehwoldt et al.
(1977)
22 141-223 7.0-7.5 Acid 24-h LC50 310 Neskovic et al.
48-h LC50 295 (1994)
96-h LC50 270
Static Dimethylamine salt 48-h LC50 >560-<1000 Bogen & Enninger
72-h LC50 >560-<1000 (1990b)
96-h LC50 >560-<1000
Static 20 50 7.2 Free acid 96-h LC50 96.5 Rehwoldt et al.
(1977)
Guppy (Lebistes Static 20 50 7.2 Free acid 24-h LC50 76.7 Rehwoldt et al.
reticulata) 96-h LC50 70.7 (1977)
Carp (Cirrhina Dimethylamine salt 96-h LD50 > 100 Singh & Yadav
mrigla hamilton) (1978)
fingerlings
Grass carp Flow 13 270 8.1 Amine salt 24-h LC50 3080 Tooby et al.
(Ctenopharyngodon 48-h LC50 2540 (1980)
idella) 96-h LC50 1313
Bleak (Alburnus Static 10 15 7.8 Butoxyethanol ester 96-h LC50 3.2-3.7 Linden et al.
alburnus) (1979)
Mosquito fish Static 21-22 Amine salt 24-h LC50 500 Johnson (1978)
(Gambusia affinis) 48-h LC50 445
96-h LC50 405
Table 6. (continued)
Organism Flow/ Temp Alkalinity Hardness pH Active ingredient Exposure 2,4-D conc. Reference
Static (°C) pmameter (mg ai/L)
Mullet (Mugil Static Sodium salt 24-h LC50 68.0 Tag El-Din et al.
cephalus) 96-h LC50 32.0 (1981)
Punti (Puntius ticto) Static 23.5 Ethyl ester 24-h LC50 1.6 Verma et al.
(1984)
Medaka (Oryzias Sodium salt 48-h LC50 > 40 Hashimoto &
latipes) Nishiuchi (1978)
in Hidaka et al.
(1984)
Pink salmon Static 14 Soft 6.3 Diethanolamine 96-h LC50 291 Wan et al. (1991)
(Oncorhynchus Isooctyl ester 96-h LC50 30
gorbuscha) (fry) Inter- 7.5 Diethanolamine 96-h LC50 363
mediate Isooctyl ester 96-h LC50 70
Hard 8.0 Diethanolamine 96-h LC50 438
Isooctyl ester 96-h LC50 21
Alaska coho salmon Static 14 Soft 6.3 Diethanolamine 96-h LC50 472 Wan et al. (1991)
(Oncorhynchus kisutch) Isooctyl ester 96-h LC50 156
(fingerling) Inter- 7.5 Diethanolamine 96-h LC50 493
mediate Isooctyl ester 96-h LC50 158
Hard 8.0 Diethanolamine 96-h LC50 662
Isooctyl ester 96-h LC50 63
Chinook salmon Flow 9 18 17 7.1 Butoxyethanol esterb 96-h LC50 0.315c Finlayson & Verrue
(Oncorhynchus Total 2,4-D 96-h LC50 0.373c (1985)
tshawytscha) (fry)
(smolts) Flow 15 18 17 7.1 Butoxyethanol esterb 96-h LC50 0.375c
Total 2,4-D 96-h LC50 1.250c
Propylene glycol 96-h LC50 0.246c
butyl ethyl esterb
Total 2,4-D 96-h LC50 1.117c
Table 6. (continued)
Organism Flow/ Temp Alkalinity Hardness pH Active ingredient Exposure 2,4-D conc. Reference
Static (°C) pmameter (mg ai/L)
Goldfish (Carassius Flow 18.2-25.8 66.7 53.3 7.84 Potassium salt 8-d LC 18.2c Birge et al.(1979)
auratus) Flow 18.2-25.8 65.3 197.5 7.78 Potassium salt 8-d LC1 8.9c Birge et al.(1979)
Largemouth bass Flow 18.2-25.8 66.7 53.3 7.84 Potassium salt 7.5-d LC1 13.1c Birge et al.(1979)
(Micropterus Flow 18.2-25.8 65.3 197.5 7.78 Potassium salt 7.5-d LC1 3.2c Birge et al.(1979)
salmoides)
Cutthroat trout (Salmo Butyl ester 96-h LC50 0.78 Woodward (1982)
clarki) (juvenile) Propylene glycol 96-h LC50 0.77 Woodward (1982)
butyl ethyl ester
Isooctyl ester 96-h LC59 > 50 Woodward (1982)
5 40 7.2 Butyl ester 96-h LC50 490 Woodward & Mayer
10 40 7.2 96-h LC50 540 (1978)
15 40 7.2 96-h LC50 770
5 40 7.2 Propylene glycol 96-h LC50 490
butyl ethyl ester
10 40 7.2 96-h LC50 1030
15 40 7.2 96-h LC50 780
10 40 6.5 Butyl ester 96-h LC50 750 Woodward & Mayer
10 40 7.5 96-h LC50 740 (1978)
10 40 8.5 96-h LC50 835
10 40 6.5 Propylene glycol 96-h LC50 930
butyl ethyl ester
10 40 7.5 96-h LC50 1220
10 40 8.5 96-h LC50 930
10 40 8.5 96-h LC50 1075
10 40 7.8 Butyl ester 96-h LC50 860 Woodward & Mayer
10 170 7.8 96-h LC50 860 (1978)
10 300 7.8 96-h LC50 860
10 40 7.8 Propylene glycol 96-h LC50 1000
butyl ethyl ester
Table 6. (continued)
Organism Flow/ Temp Alkalinity Hardness pH Active ingredient Exposure 2,4-D conc. Reference
Static (°C) pmameter (mg ai/L)
10 170 7.8 96-h LC50 860
10 300 7.8 96-h LC50 1000
Lake trout (Salvelinus 5 40 7.2 Butyl ester 96-h LC50 600 Woodward & Mayer
namaycush) 10 40 7.2 96-h LC50 640 (1978)
15 40 7.2 96-h LC50 820
5 40 7.2 Propylene glycol 96-h LC50 700
butyl ethyl ester
10 40 7.2 96-h LC50 630
15 40 7.2 96-h LC50 1000
10 40 6.5 Butyl ester 96-h LC50 820 Woodward & Mayer
10 40 7.5 96-h LC50 840 (1978)
10 40 8.5 96-h LC50 1170
10 40 6.5 Propylene glycol 96-h LC50 840
butyl ethyl ester
10 40 7.5 96-h LC50 1125
10 40 7.8 Butyl ester 96-h LC50 880 Woodward & Mayer
10 170 7.8 96-h LC50 930 (1978)
10 300 7.8 96-h LC50 1075
10 40 7.8 PGBEE ester 96-h LC50 1050
10 170 7.8 96-h LC50 1200
10 300 7.8 96 h LC50 1150
a Results generated in accordance with international guidelines and good laboratory practice
b Formulated product
c Reliable data based on measured concentrations
formulation. As the pH increased, the toxicity of 2,4-D acid to
fingerling Salmo gairdneri decreased drastically. At pH 8.5, the
96-h LC50 value was more than an order of magnitude higher than at pH
4.5 (> 1000 mg/L and < 90 mg/L, respectively) (Doe et al., 1988).
Similar 96-h LC50 values were seen for bluegills, the values for
esters ranging from 0.31 mg/L for the isopropyl ester and formulated
isopropyl ester (Drottar & Swigert, 1996c,d) to 0.61 mg/L for the
butoxyethyl ester (Mayes et al., 1989c). The 96-h values for the salts
of 2,4-D ranged from > 121 mg/L for the diethanolamine salt (Graves &
Peters, 1991g) to 1700 mg/L for the isopropylamine salt (Alexander et
al., 1983d). The 96-h LCd, values for the esters in fathead minnow
ranged from 2.5 mg/L for the butoxyethyl ester (Alexander et al.,
1983e) to > 5 mg/L for the isooctyl ester (Alexander et al., 1983a),
whereas the values for the salts ranged from 344 mg/L for the
dimethylamine salt (Alexander et al., 1983c) to 2180 mg/L for the
isopropylamine salt (Alexander et al., 1983d). The LC50, values for
seven species of fish in static systems were 55.5-427.2 mg/L for 24 h
and 26.7-300.6 mg/L for 96 h (Rehwoldt et al., 1977). The 96-h LC50
values for the free acid, the butyl ester, and the isooctyl ester in
various species of salmon ranged from < 1 to 50 mg/L, most values
being < 1 mg/L (Meehan et al., 1974).
6.2.3.2 Other effects on vertebrates
Embryos and larvae of the fathead minnow, Pimephales promelas, were
exposed to the butoxyethyl ester of 2,4-D in a flow-through system at
concentrations as high as 416.1 µg/L for 32 days. The NOEC was 80.5
µg/L, and the maximum acceptable toxicant concentration was 96.0 µg/L
(Mayes et al., 1989a).
In a study of the inhibitory effect of 2,4-D sodium salt (80%) on the
hatching of carp eggs, the average hatch was 92% with none deformed at
25 mg/dm, whereas none hatched at 100 mg/dm (Kapur & Yaduv, 1982).
Carp embryos and larvae were exposed to the sodium salt of 2,4-D (both
pure and as an 85% formulation) and maintained at 23°C for 34 days.
The formulated product at 50 mg ai/L was not harmful to embryos, but
it induced behavioural changes, disturbances in feeding, some
morphopathological changes, and ultimately death in larvae (Kamler et
al., 1974).
3.3 Terrestrial organisms
3.3.1 Plants
As 2,4-D is a herbicide and plant growth regulator, it is well
established that it can affect a number of species of terrestrial
plants, particularly non-grass species.
In field trials with paraquat in combination with other herbicides,
including 2,4-D, for the control of barley, wheat, and oats conducted
in 1982 and 1983, the phytotoxicity of paraquat was slightly reduced
when applied in combination with commercially formulated mixtures that
included 2,4-D as the dimethylamine salt and acid (O'Donovon &
O'Sullivan, 1986). Repeated applications of 2,4-D (as the 50% EC
formulation, 0.5 kg/ha) had no significant effect on either alfalfa
yields or weed suppression between treatments applied in the spring
and fall (Waddington, 1987). Applications of the iso-octyl ester of
2,4-D at 0.56 kg/ha post-emergence in combination with atrazine
resulted in the most effective control of Russian thistle and kochia
in dryland corn during the 1987, 1988, and 1989 growing seasons. The
corn yields were highest with these treatments in 1987 and 1989
(Blackshaw, 1990). In a review of management systems for conservation
fallow on southern Canadian prairies, repeated use of 2,4-D was
reported to effectively control most weeds (Blackshaw & Lindwall,
1995). As a component of a study on the effect of soil heat treatment
and microflora on the efficacy of glyphosate in bean seedlings, the
diethanolamine salt of 2,4-D was also evaluated. The quantity of
sprayed 2,4-D required to kill 50% of the bean seedlings was similar
in autoclaved and raw soil (Levesque et al., 1992).
Application of 6 µg/µl of 2,4-D every second day for a total of five
applications to the furled fourth leaf of the main shoot apex of wild
oats (Avena fatua L.) in greenhouses reduced imazamethabenz-induced
tillering but did not increase the efficacy of imazamethabenz (Chao et
al., 1994). In a study of the effect of 2,4-D on spore germination and
appressorial formation on Colletotrichum gloeosporioides f. sp.
malvae, a bioherbicide, various concentrations of the isooctyl ester
and amine salt had no significant effect (Grant et al., 1990).
An investigation was carried out on the efficacy of and residues from
the use of the butoxyethanol ester (granule) or dimethylamine salt of
2,4-D in plants, water, and sediment in Buckhorn Lake, Ontario,
Canada, over a three-year period. More than 20 indigenous species of
plants were examined (Carpentier et al., 1988). The results confirmed
the finding that the persistence of 2,4-D is controlled in part by the
organic content of the sediments and that it moves laterally through
water. The two forms of 2,4-D had essentially the same efficacy
(Bothwell & Daley, 1981). 2,4-D was among the least toxic herbicides
with regard to the rates of phosphate and ammonium assimilation in
lake phytoplankton when compared with inhibition of photosynthesis by
the 14C-bicarbonate assimilation technique. (Brown & Lean, 1995).
Eurasian watermilfoil was exposed to 2,4-D at 0.5 mg acid equivalent/L
for 12, 24, 36, 48, 60, or 72 h or to 1.0 and 2.0 mg acid equivalent/L
for 12, 24, 36, and 48 h; two untreated controls were available. The
plants were treated when the shoot apices reached to within 5-10 cm of
the water surface, after two weeks. As expected, plant damage
increased with increasing concentrations and exposure times. Control
should be achieved by exposure to a minimum 2,4-D concentration of 0.5
mg acid equivalent/L for more than 72 h and 1.0 mg acid equivalent/L
for more than 24-36 h (Green & Westerdahl, 1990).
3.3.2 Invertebrates
3.3.2.1 Toxicity to arthropods
(a) Bees
In studies conducted according to the guidelines of the Council of
Europe (CoE)/European Plant Protection Organization (EPPO) and US
guidelines, the 72-h LD50, values of the dimethylamine salt and the
ethylhexyl ester of 2,4-D to for honeybees exposed orally and by
contact were > 100 mg/bee (Palmer & Krueger, 1997a,b,c,d).
Four nuclei of bees (Apis mellifera L.) were moved into four areas
of a cage containing a heavy stand of London rocket (Sisymbrium
irio) for seven days of acclimatization before spraying. Dead bee
traps were placed for each hive. Immediately before spraying, five
wire cages each containing about 75 adult bees were placed in each
area, and dead bees were counted for 24 and 48 h after spraying at 2.2
kg ai/ha. No significant difference was observed between test and
control values (Moffet, 1972).
Eight hectares of flowering clover pasture, which was the sole source
of nectar for field bees from an apiary of 35 hives, received an
application of 3.5 kg/ha of the sodium salt of 2,4-D. Half of the area
was dusted under ideal conditions at 8.00 h and half the following day
at 6.00 h. Sufficient drift occurred to affect clover over an area of
42 ha. No bees were dusteci, as application took place before they
began to fly. There were 3500 bees/ha before dusting, 3000 on the day
after application, and fewer as the plants withered. No evidence of
repellence was observed. Nectar bees were picked off clover flowers
while they were collecting and placed in observation cages with
feeders of sugar syrup. The mortality within two days of application
among 243 field bees collected was 22%, and behavioural changes were
seen, including agitation, quivering, and attempts to sting each
other. No deaths occurred among control bees collected from a site two
miles away. No adverse effect was observed on brood or hive activity.
In a further test, while the nectar and pollen were being brought into
the hive and the queen was laying heavily, 200 foragers were heavily
dusted with 2,4-D mixture as they entered the hive. No deaths among
adult bees and no disorganization or effect on broods were observed
(Palmer-Jones, 1966).
The sodium salt of 2,4-D was considered to be non-toxic to bees
(Apis mellifera) exposed for up to 96 h at 11 µg/bee, the highest
dose tested (Atkins et al., 1975).
The 24-h oral LD50 value for the sodium salt in bees (Apis
mellifera carnica) was 62.1-92.0 µg/bee, depending on the location
of the colony and diet (Wahl & Ulm, 1983). When 2,4-D was given in 60%
sucrose syrup at concentrations of 100-1000 mg/kg to newly emerged
worker bees, no deaths were observed (Morton & Moffet, 1972).
As reported by WHO (1989), feeding of worker honey bees (Apis
mellifera) with 2,4-D salts in sucrose syrup gave 24-h LC50 values
of 104 and 115 µg/bee (Jones & Connell, 1954; Beran & Neururer, 1955).
2,4-D acid was fed to honey bees in 60% sucrose syrup at 10, 100, or
1000 mg/L and the half-time, i.e. the time for 50% of the bees in a
cage to die, was monitored. The half-time at the two lowest doses was
significantly longer than that of controls (37.2 days at 10 mg/L and
40.4 days at 00 mg/L, with a control value of 33.4 days) but was
significantly reduced at 1000 mg/L (18.6 days). The butoxyethanol and
isooctyl (commercial formulation) esters of 2,4-D had no effect on
survival times when fed at the same doses. The dimethylamine salt of
2,4-D (commercial formulation) had no effect at 10 or 100 mg/L but
shortened the half-time at 1000 mg/L (Morton et al., 1972).
(b) Other non-target arthropods
The herbicide U46 Combi fluid, containing 350 g/L 2,4-D acid and 300
g/L (4-chloro-2-methylphenoxy) acetic acid (MCPA), was tested in the
laboratory by applying 1.8 mg of a 1.5% formulation in water per cm2
on the microhymenopterous egg parasite Trichogramma cacoeciae. The
effect was determined by counting the reduction in parasitism, as
compared with the control, after eight days. Three replicates of ± 350
females each and three control replicates with a similar number of
animals were used. The adults were exposed for seven days, and the
parasitic capacity was evaluated after 12 days. The average numbers of
parasitized host eggs per adult wasp at the end of the study were
13.47, 9.42, and 15.6 (average, 12.83) in the three replicates treated
with U46 Combi fluid and 19.33, 21.37, and 22.29 (average, 20.99) in
the control cages. The average reduction in parasitic capacity was
calculated to be 38.9%. As the reduction was < 50%, U46 Combi fluid
was classified as harmless to T. cacoeciae on the basis of the
classification of the International Organization for Biological
Control of Noxious Animals and Plants (Konig, 1989).
BAS 009011H, containing 350 g/L 2,4-D dimethylamine salt and 300 g/L
MCPA dimethylamine salt, was tested in Poecilus cupreus aged five to
six weeks. Thirty insects in five replicates of six at an equal sex
ratio were used to test BAS 009011H and for the control and reference
(Afugan) groups. The test concentrations were equivalent to the
maximum recommended rate of 8 L formulation/ha in 400 L of water. No
deaths occurred among the beetles, and no behavioural effects were
seen; 73% of the insects in the reference group died. In the first
week, the feeding rate of treated beetles was similar to the controls,
but in the second week feeding fell to 55% of control values. Exposure
was therefore extended to 28 days, in order to observe any delayed
effect. Feeding increased in the third week, so that the total
reduction in feeding during the study was < 30%. The number of pupae
eaten by beetles was 3.95 for controls and 2.78 for beetles treated
with BAS 009011H (Dohmen, 1990).
The LC50 of the formulated amine salt of 2,4-D at 4 lb/gal (0.4 kg/L)
to two mite species was 0.22 kg/ha for Neoseiulus fallacis and 0.98
kg ai/ha for Tetranychus urticae (Rock & Yeargan, 1973).
Application of 2,4-D at a rate of 2.2 kg ai/ha to leafy spurge
(Euphorbia esula) and cyprus spurge (E. cyparissias) had no effect
on unsprayed larvae of the leafy spurge hawk moth (Hyles
euphoribiae) which fed on treated leaves or on the number of adults
that emerged from pupae obtained from the mating of treated or
untreated adults (Rees & Fay, 1989).
As reported by WHO (1989), beneficial coccinellid larvae were sprayed
with a preparation of mixed amine salts of 2,4-D at a rate equivalent
to 0.56 kg acid equivalent/ha 1, 3, 6, 9, or 12 days after hatching.
The development period was lengthened when the larvae were treated on
days 3, 6, 9, or 12, but there was no effect when they were sprayed on
the first day after hatching. The rate of mortality before pupation
was more than doubled in all treated groups, but that during pupation
was no different from that of controls (Adams, 1960).
Adult thistle-rosette weevils (Ceuthorhynchidius horridus), which
are used for the biological control of musk thistle, were dosed with
2,4-D amine salt at five concentrations of 0.17-147.8 kg/ha. No
significant change in mortality rate was observed up to 175 days after
treatment at doses <- 1.68 kg/ha, but the rate was significantly
increased at 16.8 and 84 kg/ha three days after treatment, and at
147.8 kg/ha mortality was increased both on day 3 and subsequently.
The five-day LC50 values were 70.2 kg/ha for males and 61.4 kg/ha for
females, which are 41.8 times the recommended application rate of 2,4-
D for males and 36.6 times that for females (Trumble & Kok, 1980).
European cockroaches (Blatella germanica) reared on food containing
1000 mg/kg showed negligible effects on reproduction (Riviere, 1976).
Wheat plants were wetted with a 0.3% solution of mixed isopropyl and
butyl esters of 2,4-D and exposed to females of the wheat stem sawfly
(Cephus cinctus). The plants were sprayed seven days before
oviposition, at the time of oviposition, or 7, 14, or 21 days after
oviposition. The eggs took about seven days to hatch. The highest rate
of mortality among larvae (96.4%) occurred after spraying at the time
of egg laying, and the effectiveness of 2,4-D in killing larvae
decreased with time of exposure, larval mortality being 68.1% 7 days,
60.8% 14 days, and 37% 21 days after oviposition, with mortality in
controls of 30%. When plants were sprayed seven days before egg
laying, larval mortality was 46.9%. Adult flies were not affected
(Gall & Dogger, 1967).
Two species of beetle (Carabidae), Bembidion femoratum and
B. ustulatum, exposed to sand dosed with 2,4-D at 1 g/m2 had
mortality rates > 50% within four days of exposure. B. ustulatum
showed 100% mortality within 10 days of exposure to 1 g/m2 and within
four days of exposure to 2 g/m2. About 20% of the individuals of
B. femoratum survived 14-day exposures to 1 or 2 g/m2 (Muller
(1971).
3.3.2.2 Toxicity to earthworms
The toxicity of the dimethylamine salt of 2,4-D was tested in the
earthworm (Eisenia foetida) at concentrations of 10-1000 mg ai/kg of
soil for up to 14 days. The 14-day LC50 was 350 mg/kg. No deaths or
sublethal effects were seen at doses < 100 mg ai/kg, equal to 200
mg formulation/kg, after 14 days. All earthworms at the highest
concentration died (Adema & Roza, 1989). When earthworms were exposed
to 2,4-D acid sprayed onto filter paper in glass vials, the calculated
48-h LC50 value was 61.6 (95% confidence interval, 41-92.4) µg/cm2
(Roberts & Dorough, 1984).
3.3.2.3 Other effects on invertebrates
Above-ground application of 2,4-D on field crops at 1250 g ai/ha for
over 20 months did not reduce the epigeal predator fauna
(staphylinids, carabids, and spiders). Field processing for planting
was reported to be more harmful (Everts et al., 1989). 'Relatively
insoluble' 2,4-D applied to a forest floor at a rate of 33.6 kg/ha
tended to remain in the litter layer when decomposition was initiated.
Fully treated substrate resulted in 50% mortality of adult millipeds
(Scytonotus simplex) by day 7 (Hoy, 1985).
No effect on soil microarthropods was reported when turf was treated
with a mixture of MCPA at the recommended rate and the butyl ester of
2,4-D at 10 times the recommended rate (Rapoport & Cangioli, 1963).
Adult millipeds (Scytonotus simplex) were exposed to 2,4-D at a rate
equivalent to 33.6 kg/ha, either uniformly to filter paper substrate,
to half the substrate, or added to half the food consisting of
air-dried red alder leaves. The mortality rate was highest for
millipeds on fully treated substrate, lower when they were exposed
only to contaminated food, and lowest when they were exposed to
substrate half of which was treated with twice the dose of the fully
treated substrate. Those on fully treated substrate began to die on
the first day and more than 50% were dead by day 7 (Hoy, 1985).
After a single application of 2,4-D (Rapoport & Cangioli, 1963) or
repeated applications for 10 and 11 years (Davies, 1965; Bieringer,
1969), none or very little effect was seen on collembolan and mite
populations. Population increases seen after use of 2,4-D were
ascribed to increased microbiological and bacteriological activity in
the soil, and a population decrease 11 months after 2,4-D application
may have been related to lower plant residues in the soil (Prasse,
1975).
3.3.3 Vertebrates
3.3.3.1 Toxicity to birds
The US Environmental Protection Agency requires data on the toxicity
of pesticides to mallard ducks (Anas platyrhynchos) and bobwhite
quail (Colinus virginianus). Those data are presented in Table 7. In
Table 7. Toxicity of 2,4-D to birds
Species Sex Age Route Active ingredient Exposure 2,4-D conc.a Reference
parameter
Mallard duck Mature Oral Isooctyl ester Acute LD50 663 Beavers (1984a)
Anas platyrhynchos 9 days Diet Isooctyl ester Acute LC50 > 5620 Beavers (1984b)
10 days Diet Acid Acute LC50 > 5620 Culotta et al. (1990a)
14 days Diet Dimethylamine salt 8-day LC50 > 4640 Fink (1974a)
10 days Diet Dimethylamine salt Acute LC50 > 5620 Long et al. (1990a)
NOEC 562
M 4 months Oral Acid Acute LD50 > 2000 Hudson et al. (1984)
M 3-5 months Oral Sodium salt Acute LD50 > 2025 Hudson et al. (1984)
M 7 months Oral Amine salt Acute LD50 < 2000 Hudson et al. (1984)
F 3-5 months Oral Acid (technical grade) Acute LD50 > 1000 Hudson et al. (1984)
Oral Isopropylamine salt Acute LD50 > 398 Beavers (1983a)
10 days Diet Isopropyl ester Acute LC50 > 5930 Palmer &Beavers (1996a)
10 days Diet Butoxyethyl ester Acute LC50 > 5620 Grimes et al. (1990a)
23 days Diet Butoxyethanol ester 5-day LC50 > 5000 Hill et al. (1975)
17 days Diet Dimethylamine salt 5-day LC50 > 5000 Hill et al. (1975)
10 days Diet Diethanolamine salt Acute LC50 > 5620 Hoxter et al. (1991)
9 days Diet Isopropylamine salt Acute LC50 > 5620 Beavers (1983b)
10 days Diet Triisopropanolamine salt Acute LC50 > 5620 Driscoll et al. (1990a)
Bobwhite quail 21 weeks Oral Dimethylamine salt Acute LD50 500 Hoxter et al. (1990a)
(Colinus virginianus) No mortality 250
11 days Diet Isooctylester Acute LC50 > 5620 Beavers (1984c)
10 days Diet Acid Acute LC50 > 5620 Culotta et al. (1990b)
No mortality 3160
14 days Diet Dimethylamine salt 8-day LC50 > 4640 Fink (1974b)
10 days Diet Dimethylamine salt Acute LC50 > 5620 Long et al. (1990b)
No mortality 3160
23 weeks Oral Isopropyl ester Acute LD50 1879 Palmer & Beavers (1996b)
No mortality 308
19 weeks Oral Butoxyethyl ester Acute LD50 > 2000 Lloyd et al. (1990)
18 weeks Oral Diethanolamine salt Acute LD50 595 Campbell et al. (1991)
27 weeks Oral Triisopropanolamine Acute LD50 405 Culotta et al. (1990c)
10 days Diet Isopropyl ester Acute LC50 > 5930 Palmer & Beavers (1996c)
Table 7. (continued)
Species Sex Age Route Active ingredient Exposure 2,4-D conc.a Reference
parameter
10 days Diet Butoxyethyl ester Acute LC50 > 5620 Grimes et al. (1990b)
23 days Diet Butexyethanol ester 5-day LC50 > 5000 Hill et al. (1975)
23 days Diet Dimethylamine salt 5-day LC50 > 5000 Hill et al. (1975)
12 days Diet Isopropylamine salt Acute LC50 > 5620 Beavers (1983c)
10 days Diet Diethanolamine salt Acute LC50 > 5620 Hoxter et al. (1991)
10 days Diet Triisopropanolamine salt Acute LC50 > 5620 Driscoll et al. (1990b)
Japanese quail M 2 months Oral Acid (technical grade) Acute LD50 668 (530-842) Hudson et al. (1984)
(Coturnix japonica) 14 days Diet Acetamide salt 5-day LC50 > 5000 Hill et al. (1975)
12 days Diet Butoxyethanol ester 5-day LC50 > 5000 Hill et al. (1975)
20 days Diet Dimethylamine salt 5-day LC50 > 5000 Hill et al. (1975)
Pheasant F 3-4 months Oral Acid (technical grade) Acute LD50 472 (340-654) Hudson et al. (1984)
(Phesianus colchicus) 10 days Diet Butoxyethanol ester 5-day LC50 > 5000 Hill et al. (1975)
10 days Diet Dimethylamine salt 5-day LC50 > 5000 Hill et al. (1975)
Chukar partridge M,F 4 months Oral Acid (technical grade) Acute LD50 200-400 Hudson et al. (1984)
(Alectoris chukar)
Rock dove (Columba M,F Oral Acid (technical grade) Acute LD50 668 (530-842) Hudson et al. (1984)
livia)
a Oral doses, mg/kg body weight; dietary doses, ppm
studies conducted according to the guidelines of the US Environmental
Protection Agency, the oral LD50 values for mallard ducks are
generally high, ranging from > 398 mg/kg for the isopropylamine salt
(Beavers, 1983a) to > 2000 mg/kg for the sodium salt and acid (Hudson
et al., 1984). In similar studies, the oral LD50 values for northern
bobwhite quail ranged from 405 mg/kg for the triisopropanolamine salt
(Culotta et al., 1990c) to > 2000 mg/kg for the butoxyethyl ester
(Lloyd et al., 1990). The dietary LC50 values in mallard ducks and
northern bobwhite quail were all > 5000 ppm, for the isooctyl ester,
isopropyl ester, butoxyethyl ester, dimethylamine salt, diethanolamine
salt, isopropylamine salt, triisopropanolamine salt, and acid
(Beavers, 1983b, 1984b; Driscoll et al., 1990a; Grimes et al., 1990b;
Long et al., 1990a; Hoxter et al., 1991; Palmer & Beavers, 1996a), and
the dietary LC50 values for the isopropylamine salt, diethanolamine
salt, and triisopropanolamine salt in northern bobwhite quail all
exceeded 5620 ppm, even in a five-day study in 23-day-old birds.
Similar toxicity was seen for Japanese quail (Coturnix japonica) and
pheasants (Phesianu colchicus) with the acetamide, dimethylamine
salt, and butoxyethanol ester (Hill et al., 1975).
2,4-D is less toxic to vertebrates treated in the diet than dosed
orally. Furthermore, the toxicity of all of the forms tested was low.
Data submitted by the Industry Task Force II on 2,4-D Research Data
was consistent with values in the literature. Owing to its good
solubility in water (23 180 ppm at pH 7), its low log P value (-0.95
to-0.75 at pH 7), and its rapid excretion after uptake, 2,4-D does not
tend to accumulate to any significant extent in the environment; owing
to its low toxicity to birds when given by the dietary route,
secondary poisoning is unlikely to occur.
3.3.3.2 Toxicity to birds' eggs
As reported by WHO (1989,), several studies have been conducted on the
toxicity of various 2,4-D formulations to birds eggs. No adverse
effects were found on the hatchability of eggs or on the incidence of
deformities or mortality of hatched chicks after the eggs of pheasants
or chickens were sprayed with an isooctyl ester formulation of 2,4-D
on day 13 of incubation at a dose equivalent to 0.28 kg/ha (Kopischke,
1972). Similarly, spraying of chicken's eggs before incubation with an
amine salt of 2,4-D at concentrations of up to 15 times the
recommended field application rate of 3 kg/ha had no effect on
hatching success or on the survival of chicks three to four weeks
after hatching (Somers et al., 1974). Spraying chicken's eggs on day
0, 4, or 18 of incubation with 2,4-D propylene glycol butyl ethyl
ester formulation at up to 10 times the field application rate also
had no effect on the hatchability or on survival or growth of chicks
after hatching (Somers et al., 1978a). Birds hatched from eggs
similarly treated by spraying showed no significant adverse effects in
reproductive performance, i.e. egg laying performance of females and
testicular weight and sperm count in males (Somers et al., 1978b). No
effect on egg hatching rate or on body weight or malformation rate was
seen in chicks after the eggs of Japanese quail, pheasants, and
chickens were sprayed at 20 kg/ha before incubation or three days
after the start of incubation (Hilbig et al., 1976a). In a follow-up
study on the reproductive performance of birds hatched from the
treated eggs, no effects were seen on laying capacity, fertility, or
hatchability of the eggs (Hilbig et al., 1976b).
The effects of 2,4-D dimethylamine salt on the eggs of Japanese quail,
grey partridge, and red-legged partridge were studied by spraying eggs
at doses equivalent to the recommended application rate of 1.2 kg/ha
and at two higher doses equivalent to 2.4 and 6 kg/ha. No effects were
seen on hatching rate or embryonic or chick mortality during the first
month after hatching or on embryonic or chick malformations.
Histopathological examination of partridge thyroids revealed no
effects. When the residues of 2,4-D were measured in the partridge
eggs that had received the highest dose, very little 2,4-D was found
to have penetrated the eggshell; the highest level of residue measured
was a total egg content of 19.3 µg in an 11-g egg, 15 days after
treatment. The lack of effect of 2,4-D on sprayed eggs was attributed
to its poor penetration (Grolleau et al., 1974). No adverse effect of
2,4-D was seen on hatchability and no increase was seen in
abnormalities in pheasant or quail chicks from eggs sprayed 24 h
before hatching with a dose 12 times higher than that recommended for
application. Only at a dose 30 times higher than the recommended rate
did hatchability fall by 10-15%, relative to controls. No increased
incidence of abnormalities was reported at this dose (Spittler, 1976).
When mallard eggs were immersed for 30 s in aqueous emulsions of
2,4-D, the calculated LC50 was equivalent to a field application rate
of 216 (155-300) kg/ha, which is 32 times the recommended field
application rate (Hoffman & Albers, 1984). Chicken eggs injected with
10, 100, or 200 mg/kg 2,4-D, equivalent to 0.5, 5, or 10 mg/egg, had
hatching rates of 80-90%, 70%, and 50% of the solvent control hatch
rate at the three doses, respectively (Dunachie & Fletcher, 1967). In
a similar study, an injection of < 1 mg/egg of the dimethylamine
salt of 2,4-D had no effect, whereas injections of 2 mg/egg reduced
both hatchability and survival of hatched chicks; 5 mg/egg reduced the
hatching rate to 15% of control levels, and there were no surviving
chicks after one week. No hatching occurred after an injection of 10
mg/egg. When eggs were treated by immersion in solutions of 2,4-D for
10 s, no effect was seen in a solution of 10 g/L and only a slight
effect in 50 g/L. Hatching success and survival of chicks up to four
weeks post-hatch was more than 80% of control values after immersion
in 50 g/L (Gyrd-Hansen & Dalgaard-Mikkelsen, 1974)œ
3.3.3.3 Effects on mammals
The acute toxicity of 2,4-D, its diethanolamine, dimethylamine,
isopropylamine, and triisopropanolamine salts, and the 2-butoxyethyl
and 2-ethylhexyl) esters was determined in rats and rabbits. After
oral administration, the isopropylamine salt was the least toxic
(LD50 = 2322 mg ai/kg for male rats and 1646 for female rats; Carreon
et al., 1983), and the LD50 for the acid was 699 mg ai/kg (Myer,
1981a). The LD50 values for all of the other salts and esters fell
between these ranges; those for the dimethylamine salt and ethyhexyl
esters were 863 and 896 mg/kg bw, respectively (Myer, 1981 b; Jeffrey,
1987a; Berdasco et al., 1989a; Schults et al., 1990a). After dermal
administration to rabbits, none of the salts or esters was toxic, with
LD50 values > 2000 mg ai/kg (Myer, 1981c,d,e; Carreon et al., 1983;
Jeffrey, 1987b; Berdasco et al., 1989b; Schults et al., 1990b). After
administration by inhalation, the triisopropanolamine salt was the
least toxic, with an LC50 > 10.7 mg ai/L (Nitschke & Stebbins,
1991); all of the other salts and esters had LC50 values of
> 1.79-5.39 mg ai/L (Streeter & Young, 1983; Auletta & Daly, 1986;
Streeter et al., 1987; Jackson & Hardy, 1990; Cieszlak, 1992).
3.3.3.4 Effects on amphibia
The results of studies conducted according to the CoE/EPPO and US
guidelines on the toxicity of 2,4-D to amphibia are shown in Table 8
(Palmer & Krueger, 1997e,f). LC50 values for 2,4-D ethythexyl ester
could not be determined owing to its rapid hydrolysis and very low
solubility in water. The 96-h acute LC50 values based on measured
concentrations were 359 mg/L for the 2,4-D acid and 337 mg/L for the
dimethylamine salt. The 96-h LC50 for the acid in Indian toads
(Bufo melanostictus) was 8.05 mg/L, and the 24-h LC50 was 13.77
mg/L (Vardia et al., 1984); the 48-h value in common frogs (Rana
temporaria) was 50 mg/L (Cooke, 1972), while the amine salt was
relatively nontoxic, with 96-h LC50 values of 200-288 mg ai/L
(Johnson, 1976). Although the studies on amphibia are limited to only
a few species, 2,4-D acid appears to be more toxic in these species
than the dimethylamine salt.
4. Risk assessment based on agricultural use
The information on use and application rates used in this risk
assessment refers to the agricultural use of 2,4-D within the European
Union and the United States. It should be possible to extrapolate the
assessment to other agricultural uses at similar application rates
elsewhere in the world. 2,4-D can be formulated in a variety of
different salts (e.g. dimethylamine, sodium, diethanolamine,
triisopropanolamine, and isopropylamine salts) and esters (e.g. 2-
ethylhexyl and butoxyethyl), but the dimethylamine salt and ethylhexyl
ester account for about 95% of the global use of 2,4-D (personal
communicaton from the Industry Task Force II on 2,4-D Research Data).
The following risk assessment is therefore restricted to these forms
of 2,4-D. As all forms are rapidly transformed to the acid, data for
the dimethylamine salt and ethylhexyl ester can be converted to the
acid equivalent by multiplying the value by the following molecular
mass-based correction factor:
molecular mass of 2,4-D acid (221)
Correction factor =
molecular mass of dimethylamine (265) or
ethylhexylester (333)
Table 8. Toxicity of 2,4-D to amphibia (tadpoles)
Organism Flow/ Temperature Alkalinity Hardness pH Active ingredient Exposure 2,4-D conc. Reference
static (°C) parameter (mg ai/L)
Leopard frog 22 Acid 48-h LC50 462a Palmer & Krueger
(Ranapipiens) 72-h LC50 445a (1997e)
96-h LC50 359a
22 Dimethylamine salt 48-h LC50 480a Palmer & Krueger
72-h LC50 376a (1997f)
96-h LC50 337a
Indian toad (Bufo 25 210 220 8.3 Free acid 24-h LC50 13.77 Vardia et al.
melanostictus) 48-h LC50 9.03 (1984)
96-h LC50 8.05
Frog Static 21-22 Amine salt 24-h LC50 321 Johnson (1976)
(Limnodynastes 48-h LC50 300
peroni) 96-h LC50 287
Toad (Bufo Static 21-22 Amine salt 24-h LC50 346 Johnson (1976)
marinus) 48-h LC50 333
96-h LC50 288
Common frog (Rana 17-29 Free acid 48-h LC50 > 50 Cooke (1972)
temporaria)
a Reliable data based on measured concentrations
In order to convert data for the acid to the salt or ester equivalent,
the acid value should be multiplied by the inverse of the above
correction factor.
The main uses of 2,4-D are shown in Table 9. It can be applied by
either conventional tractor-mounted or -drawn hydraulic sprayers or by
air, for instance in forestry use. This risk assessment is based on
the following applications (converted to the dimethylamine salt or
ethylhexyl ester equivalent), which are representative of real maximum
application rates:
Single ground-based 2.69 kg/ha dimethylamine salt
hydraulic application
(maximum rate = 2.24 kg 3.37/ha kg ethylexyl ester
2,4-D acid/ha)
Single aerial application 5.37 kg/ha dimethylamine salt
(maximum rate = 4.48 kg 6.75 kg/ha ethylhexyl ester
2,4-D acid/ha)
Aquatic weed control water concentration, 1.13-4 mg/L
dimethylamine salt
The risk assessment is based on the principle of calculating
toxicity:exposure ratios (TERs; see Figure 1) from the CoE/EPPO
Environmental Risk Assessment models and trigger values.
4.1 Microorganisms
The most important sources of exposure of soil microorganisms to 2,4-D
are likely to be ground or aerial applications at maximal individual
or seasonal rates of up to 4.28 kg/ha dimethylamine salt or 5.4 kg/ha
ethylhexyl ester. In studies in the laboratory, concentrations of
2,4-D (form unstated) of < 10 µg/g soil had no effect on soil
respiration or nitrification and were not toxic to various
denitrifying microorganisms, and concentrations up to 25 µg/g soil had
no effect on soil bacteria, fungi, or actinomyces. These
concentrations are equal to application rates of 7.5 and 18.75 kg
2,4-D/ha, respectively, based on a soil depth of 5 cm and a soil
density of 1.5 g/cm3 and correspond to about 1.4 and 4.7 times the
maximum recommended application rate (both single and seasonal). The
risk to soil microorganisms from use of 2,4-D should thus be low. In
another study, applications of the dimethylamine salt and isooctyl
ester at rates corresponding to 0.95 kg 2,4-D/ha, resulted in 10-15%
and 27-29% reductions in populations of soil bacteria, fungi, and
actinomyces, respectively. As the trigger for concern in the CoE/EPPO
microorganism risk assessment scheme is > 30%, the risk to soil
microorganisms from use of 2,4-D is low. This is reinforced by the
reports in section 5.3 that little or no residue of 2,4-D is found in
soils in the field.
Table 9. Main uses of 2,4-D, in kg/ha of acid
Crop Labelled Typical Maximum seasonal
application rate application rate application rate
USAa European Unionb
Cereals 0.28-1.4 0.49 0.41-0.9 2.0
Corn 0.28-1.4 0.50 1.2 3.4
Sorghum 0.28 1.1 0.50 1.1
Soya beans 0.56-1.1 0.53 1.1
Sugar cane 1.1-2.2 0.90 4.5
Rice 0.56-1.7 1.1 1.7
Pasture 0.56-2.2 0.63 0.9-1.5 4.5
Top fruit 0.56-2.2 1.2 4.5
Turf 1.1-2.2 - 0.9-1.5 4.5
Non crop land 0.56-4.5 1.2 4.5
Fallow/stubble 0.56-3.4 - 0.9-2.9 2.2
Forests 0.56-4.5 2.5 4.5
Aquatic weeds 2-4 mg/L in 1.13 mg/L
treated water
a typical average rate
b European usage
4.2 Aquatic organisms
The main potential sources of risk to aquatic organisms from the use
of 2,4-D are overspray during aerial use at 5.37 kg dimethylamine
salt/ha or 6.75 kg ethylhexyl ester/ha (4.48 kg 2,4-D acid
equivalents/ha), spray drift from ground-based hydraulic applications
at 2.69 kg dimethylamine salt/ha or 3.37 kg ethylhexyl ester/ha (2.24
kg 2,4-D acid equivalents/ha), or use to control aquatic weeds at UP
to 5 mg dimethylamine salt/L treated water. For each of these
situations, the predicted environmental concentration (PEC) in a 30-cm
static surface-water body arising from ground-based spray drift 1 m
from the edge of a spray boom (from Ganzelmeier et al., 1995) or
aerial overspray, was calculated from the European Union model, as
follows:
maximum application rate (kg 2,4-D/ha) × A
(% spray drift)
PEC (mg/L) =
300
where A = 5 for ground-based hydraulic spray applications 1 m from
edge of boom and 100 for overspray from aerial applications.
4.2.1 Acute risk to freshwater pelagic organisms
The results of toxicity tests reported in section 6.2 shows that the
ethylhexyl ester is significantly more toxic to aquatic organisms than
the acid. Section 4.2.3 indicates that the ethylhexyl ester degrades
rapidly to 2,4-D acid in water, with a reported DT50 of 6.2 h
(section 4.2.1). Therefore, the risk to aquatic organisms from the
ethylhexyl ester is somewhere between that posed by the ester and that
of the less toxic acid. In order to take into consideration both of
these risks, the inital assessment is based on the worst-case scenario
for the ethylhexyl ester; if a risk is identified, a further
assessment is carried out for the acid. Similarly for algae, 2,4-D
acid is about an order of magnitude less toxic than either the
dimethylamine salt or the ethylhexyl ester. Therefore, when the data
on the ester or salt indicate a risk to algae, it is re-assessed on
the basis of data for the acid. In contrast, 2,4-D acid is an order of
magnitude more toxic to fish than the dimethylamine salt, which,
however, degrades rapidly to the acid in water. Therefore, the risk to
fish from the dimethylamine salt is re-assessed with data for the acid
when the TER for the dimethylamine is less than an order of magnitude
higher than the EPPO trigger value.
The acute LC50 and EC50 values for the most sensitive fish (Tables 5
and 6) were 250 mg/L dimethylamine salt and 7.2 mg/L ethylhexyl ester,
with a 96-h LC50 as low as 1 mg/L for the ester. The values for the
most sensitive aquatic invertebrate (Tables 3 and 4) were 184 mg/L for
the dimethylamine salt and 5.2 mg/L for the ester. Although an LC50
of 4.0 mg/L for the dimethylamine salt to Daphnia has been reported,
the more recent results of studies conducted according to guidelines
and good laboratory practice were used. The values for the most
sensitive algal species (Table 2) were 5.28 mg/L dimethylamine salt
and 4.1 mg/L ethylhexyl ester, and those for the most sensitive
aquatic plant species were 0.58 mg/L dimethylamine salt and 0.5 mg/L
ethylhexyl ester.
The acute LC50 and EC50 values for the most sensitive aquatic
organisms to 2 4-D acid (Tables 3-6) were 26.7 mg/L for fish, 79 mg/L
for aquatic invertebrates, and 29 mg/L for algae.
4.2.1.1 Spray drift from ground-based applications
The acute PEC for spray drift (1 m from the edge of a spray boom,
based on the assumptions listed in section 7.2) into a 30-cm static
water body at the maximum application rate (see above) is 0.045 mg/L
for the dimethylamine salt and 0.056 mg/L for the ethylhexyl ester.
Therefore, the TERs based on such PECs and the LC50 and EC50 values
for the dimethylamine and ethylhexyl forms of 2,4-D, respectively are:
fish, 5555 and 128 (22.2 from published data); aquatic invertebrates,
4088 and 93; algae, 117 and 73; and higher aquatic plants, 12.9 and
8.9. On the basis of the CoE/EPPO risk assessment scheme for aquatic
organisms, these TERs (> 10) indicate a low acute risk to fish,
aquatic invertebrates, and algae; however, for higher aquatic plants,
the TER for the dimethylamine salt indicates a low risk but that for
the ethylhexyl ester indicates a potential risk. The ethylhexyl ester
is, however, rapidly degraded to the acid in water. The acid was less
toxic to Lemna, with a 14-day EC50 of 3.3 mg/L, which gives a TER
of 88.4 on the basis of a PEC of 0.037 mg/L, and indicates a low risk
to higher aquatic plants.
4.2.1.2 Overspray from aerial spray applications
The acute PEC for aerial overspray in a 30-cm static water body at the
maximum application rate on the basis of the assumptions listed in
section 7.2, is 1.79 mg/L dimethylamine salt or 2.25 mg/L ethylhexyl
ester. The TERs for the dimethylamine and ethylhexyl forms of 2,4-D,
respectively, based on these PECs and the LC50 and EC50 values given
above are: fish, 140 and 3.2 (1.8 in published data); aquatic
invertebrates, 103 and 2.3; algae, 3.0 and 1.8; and higher aquatic
plants, 0.32 and 0.22. On the basis of the CoE/EPPO risk assessment
scheme for aquatic organisms, the TERs for the dimethylamine indicate
a low acute risk to both fish and aquatic invertebrates (i.e. TERs
>10); however, the TERs for the dimethylamine to algae and higher
aquatic plants and those for the ethylhexyl ester to all aquatic
organisms indicate a high acute risk. As both the esters and salts of
2,4-D are rapidly degraded to the acid, the risk was re-assessed on
the basis of data for 2,4-D acid, giving a PEC of 1.49 mg/L and TERs
of 17.9 for fish, 53 for aquatic invertebrates, 19.4 for algae, and
2.2 for higher aquatic plants (Lemna), indicating low acute risks to
fish, aquatic invertebrates, and algae but a potential risk to higher
aquatic plants from aerial use of 2,4-D.
4.2.1.3 Aquatic weed control
For the control of aquatic weeds, water can be treated with up to 4
mg/L 2,4-D. Use of the ethylhexyl ester is not recommended for aquatic
weed control. On the basis of the PEC and the LC50 and EC50 values
for the dimethylamine, the TERs are: fish, 50; aquatic invertebrates,
36.8; algae, 1.06; and higher aquatic plants, 0.12. On the basis of
the CoE/EPPO risk assessment scheme for aquatic organisms, these TERs
indicate a low acute risk to fish and aquatic invertebrates; however,
the TERs for the dimethylamine salt to algae and aquatic plants and
the TERs for the ethylhexyl ester to all aquatic organisms indicate a
high acute risk. As both the esters and salts of 2,4-D are rapidly
degraded to the acid, the risk was re-assessed on the basis of data
for 2,4-D acid, giving a PEC of 4 mg/L and TERs of 5.8 for algae and
0.8 for higher aquatic plants. These TERs still indicate a potential
risk; however, since the application is for the control of aquatic
weeds, the risk can be ignored. The low TER for fish indicates that
the use may pose an acute risk; however, that risk should be balanced
against the risks of other alternatives, such as not controlling weeds
(e.g. algal bloom leading to water deoxygenation) and the potential
damage caused by manual weed control, which may both pose higher risks
to fish and other aquatic organisms.
4.2.2 Long-term risk to freshwater pelagic organisms
The long-term NOEC values reported in Table 5 for the most sensitive
aquatic species tested were 17.1 mg/L for dimethylamine salt and 0.12
mg/L for the ethylhexyl ester for fish at early life stages and 27.5
mg/L dimethylamine salt and 0.015 mg/L ethylhexyl ester for aquatic
invertebrates. As both the ester and the salt degrade rapidly in water
to 2,4-D acid (e.g. the DT50 for the ester is 6.2 h), the assessment
of long-term risk was based on data on the long-term NOECs of the
acid, i.e. 63.4 mg/L for the most sensitive fish (Table 5) and 79 mg/L
for aquatic invertebrates (Table 3).
4.2.2.1 Levels of 2,4-D acid in surface water
Instead of using PECs in this risk assessment, actual field levels of
2,4-D (assumed to be the acid) in various surface waters (as reported
in section 5.2) were used. The levels of the acid in surface waters
generally ranged from 0.00008 mg/L (in a small watershed in
Saskatchewan, Canada) to 0.0021 mg/L (in ground- and surface waters in
the United Kingdom). Although a higher level, 0.029 mg/L, was reported
in Saskatchewan, it was attributed to back-siphoning of spray
solution. As this source of contamination is considered to be
accidental, it was not used in this risk assessment, which addresses
the risk from normal or good agricultural use. The TERs based on the
measured levels of 2,4-D in surface water and the NOEC values are all
well in excess of the CoE/EPPO trigger of concern (i.e. TERs > 10 for
fish and aquatic invertebrates) and indicate a low risk to these
organisms.
4.2.2.2 Aquatic weed control
Water can be treated with up to 4 mg/L of 2,4-D for the control of
aquatic weeds. On the basis of this PEC, the TERs based on the NOEC
for 2,4-D acid are 15.4 for fish and 19.7 for aquatic invertebrates,
indicating low risks.
4.2.3 Risk to sediment-dwelling invertebrates
2,4-D partitions and can persist in aquatic sediments, particularly
under anaerobic conditions. Levels up to 0.17 mg/kg sediment (assumed
to be 2,4-D acid) were found in a pond three days after application of
8.86 kg/ha of the dimethylamine salt (section 4.2.4). As the highest
seasonal application rate is 4.48 kg 2,4-D acid equivalents/ha, which
is about half the rate used in the above study, the level of 2,4-D
would be 0.085 mg/kg sediment. On the basis of this PEC and the LC50
of 86.7 mg 2,4-D acid/L for the most sensitive sediment-dwelling
invertebrate species tested, the oligochaete worm Lumbriculus
variegatus, the TER for sediment-dwelling invertebrates would be
1020, indicating a low risk.
Risk to amphibia
The LC50 values shown in Table 8 for tadpoles of a variety of frog
species range from 8.05 to 359 mg 2,4-D acid equivalents/L. On the
basis of the lowest LC50 value of 8.05 mg/L and PEC values of 0.037,
1.49, and 5 mg/L 2,4-D acid for spray drift from ground-based
applications, aerial applications, and aquatic weed control,
respectively, the TERs are 214, 5.36, and 1.1, indicating a low risk
to amphibia; however, contamination of surface waters arising from
overspray during aerial application and aquatic weed control might
pose a risk to amphibia, as indicated by TERs < 10. The risk from
aquatic weed control should be balanced against the risks from other
alternatives, however, such as not controlling weeds (e.g. algal bloom
leading to water deoxygenation) and potential damage caused by manual
weed control.
4.2.5 Bioaccumulation
No data were submitted on bioaccumulation in fish. Residues of < 5-
102 µg/kg of 2,4-D have been monitored, indicating a mimimal risk of
bioaccumulation.
Terrestrial organisms
Plants
2,4-D is a translocated, selective herbicide used to control a variety
of broad-leaved weeds (section 6.3.1). Consequently, any risk to
broad-leaved non-target plants is to be expected from its mode of
action and consequent area of use. 2,4-D has been reported to cause
phytotoxic damage to neighbouring crops, attributed to vapour drift;
however, such drift was associated mainly with use of the volatile
ester forms of 2,4-D. Use of the butyl ester is now restricted and
continues to decline globally, in part due to the campaigning by the
registrants (personal communication from Industry Task Force II on
2,4-D Research Data).
4.3.2 Invertebrates
4.3.2.1 Bees
Bees may be exposed to 2,4-D while they forage on flowering weeds in
treated crops; the main uses of 2,4-D are not on flowering crops. The
LD50 values for dimethylamine salt and the ethylhexyl ester in
honeybees were > 100 µg/bee for both oral and contact exposure. On
the basis of the maximum individual application rate of 4.48 kg 2,4-D
acid equivalents/ha, the hazard quotient would be < 45. As the
CoE/EPPO trigger for concern is a hazard quotient > 50, the acute
risk to honeybees from use of 2,4-D at high application rates should
be low. Studies in which these forms of 2,4-D were fed to bees in
syrup showed no effect on survival at 100 mg/L. In field studies,
application of 2,4-D at 2.24 kg/ha had no significant effect on caged
foraging bees. In view of the restricted exposure of bees, the risk
should be low. Furthermore, 2,4-D has not been implicated in any
incidents of poisoning of honeybees in the Wildlife Incident
Investigation scheme in the United Kingdom (personal communication
from the Industry Task Force II on 2,4-D Research Data).
4.3.2.2 Other non-target arthropods
Few of the studies used in the risk assessment were conducted
according to internationally recognized guidelines or good laboratory
practices, and most predate the use of such standards by a
considerable time. Additionally, the test substances were usually not
adequately characterized or identified and in some cases were
mixtures. Non-target arthropods may be exposed to 2,4-D during any of
its various agricultural and non-agricultural uses. Standard
laboratory tests conducted with a mixture of 2,4-D acid and
dimethylamine salt (2.8 kg 2,4-D/ha) and MCPA reduced the fecundity of
the parasitic wasp Trichogramma cacoeciae by 38.9%, caused a
transient 55% drop (not significant over the whole test period) in the
feeding rate of the carabid beetle Poecilus cupreus, and had no
significant effect on the staphylinnid beetle Aleochara bilineata.
Doubled pre-pupal mortality and increased development time were seen
in larval coccinellids sprayed directly with 2,4-D acid at an
application rate of 0.56 kg/ha. Similarly, mortality rates of 36-96.4%
were seen among wheat-stem sawflies at a concentration of 0.3% of the
mixed isopropyl and butyl esters of 2,4-D; a mortality rate of 50% was
reported for Bembidion femoratum and B. ustulatum at a rate equal
to 10 kg/ha, i.e. 2.5 times the maximum recommended rate, for the
predatory mites Neoseiulus fallacis and Tetranychus utricae after
applications of 0.22 and 0.98 kg/ha of the amine salt, respectively,
and for adult millipedes exposed to 33.6 kg/ha (8.4 times the maximum
recommended application rate). No significant effect was reported on
thistle-rosette weevils treated with various amine salts of 2,4-D at a
rate of 1.68 kg/ha or on leafy spurge hawk moth larvae treated at 2.2
kg/ha; and in field studies, no significant effects on epigeal
predator fauna (staphylinids, carabids, and spiders) were seen with
application of 2,4-D at 1.25 kg/ha. Similarly, single and repeated
field applications of 2,4-D (rate unstated) resulted in little or no
effect on soil collembolans or mites.
The risk assessment is limited by the standard of the available data.
The CoE/EPPO and International Organization for Biological Control of
Noxious Animals and Plants trigger for concern with regard to
non-target arthropods, is an effect > 30%. The limited laboratory
data indicate that 2,4-D may have insecticidal activity, particularly
at high application rates, and hence may pose a risk to this
compartment of the terrestrial environment. Limited field data at
lower, more typical rates of application (< 1.25 kg/ha) indicate,
however, that this risk may not be realized in the field.
4.3.2.3 Earthworms
Earthworms may be exposed after single or multiple applications of
2,4-D to a wide variety of crops but particularly after its use on
grass, fallow land, and stubble. On the basis of a maximum application
rate of 5.37 kg 2,4-D dimethylamine salt/ha, a soil depth of 5 cm, and
a soil density of 1.5 g/cm3, the PEC for the dimethylamine salt is
7.2 mg/kg soil. A study in Eisenia foetida with the dimethylamine
salt gave a 14-day LC50 of 350 mg/kg soil. The TER for earthworms
would thus be 49, which is above the CoE/EPPO trigger value of 10. The
acute risk to earthworms of the use of the dimethylamine should
therefore be low. This conclusion is reinforced by the reported levels
of 2,4-D in soil (section 5.3), which are well below the PEC used in
the above calculation.
4.3.3 Vertebrates
Vertebrates are likely to be exposed to 2,4-D while grazing on treated
or contaminated vegetation or eating contaminated insects. The
estimated residues on food items represent the maximum values for
pesticides immediately after application and do not take into
consideration the degradation of 2,4-D in the environment (which is
rapid). Furthermore, it is assumed that all food consumed contains the
maximum residue levels. For the purpose of this risk assessment, the
worst-case maximum application rates of 5.37 kg dimethylamine salt/ha
and 6.75 kg ethylhexyl ester/ha (4.48 kg 2,4-D acid equivalents/ha)
from aerial application and 2.69 kg dimethylamine salt/ha and 3.37 kg
ethylhexyl ester/ha (2.24 kg 2,4-D acid equivalents/ha) for
ground-based and aerial spraying of pastures were used.
4.3.3.1 Birds
Table 7 gives LD50 values of 500 mg/kg bw for the dimethylamine salt
and 663 mg/kg bw for the isooctyl ester for bobwhite quail and mallard
ducks, respectively, the most sensitive species tested. The dietary
LC50 values for the mallard duck, the most sensitive species tested,
were > 4640 ppm dimethylamine salt and > 5620 ppm isooctyl ester.
The indicator birds used in the risk assessment were a 3-kg greylag
goose (Anser anser), with a total daily food consumption of 900 g
vegetation, as the grazing bird (Owen, 1975) and an 11-g blue tit
(Parus caeruleus), with a total daily food consumption of 8.23 g, as
the small insectivorous bird (Kenaga, 1973).
(a) Aerial applications (4.48 kg 2,4-D acid equivalents/ha for
insectivorous birds and 2.24 kg 2,4-D acid equivalents/ha for
grazing birds)
Grazing birds: The residues on short grass measured initially were
either 290 mg/kg of grass for 2,4-D dimethylamine salt or 459 mg/kg
for the ethylhexyl ester. If a 3 -kg greylag goose fed exclusively on
contaminated grass, it would ingest 261 mg dimethylamine salt or 413
mg ethylhexyl ester, equal to 87 mg/kg bw of the salt and 138 mg/kg bw
of the ester. Consequently, the acute oral TERs would be 5.7 for the
salt and 4.8 for the ester. These TERs are below the CoE/EPPO trigger
value of 10, indicating a high risk to grazing birds from aerial
applications of both 2,4-D dimethylamine salt and ethylhexyl ester.
The short-term results of feeding studies, which represent more
realistic exposure, would give TERs, based on the LC50 values, of
> 16 and > 12 for the salt and ester forms, respectively. These TERs
are above the CoE/EPPO trigger for high risk but may be below the
trigger for medium risk (< 100). As the residues in grass on day 7
had declined to between one-half and one-third of those on day 0 and
as 2,4-D has not been implicated in incidents of poisoning in birds
when used normally (UK Wildlife Incident Investigation Scheme), the
risk to grazing birds from 2,4-D is unlikely to be high (personal
communication from the Industry Task Force II on 2,4-D Research Data).
Insectivorous birds: The initial residues on small insects after
aerial overspray (based on 29 times the application rate in kg/ha as
mg/kg; CoE/EPPO vertebrate risk assessment scheme) would be 156 mg/kg
insects for the dimethylamine salt or 196 mg/kg for the ethylhexyl
ester. On the basis of the values for acute oral toxicity, if an 11-g
blue tit fed exclusively on contaminated insects, it would ingest 1.3
mg of the salt or 1.6 mg of the ester, equal to 118 and 145 mg/kg bw,
respectively. Consequently, the acute oral TERs would be 4.2 for the
salt and 4.6 for the ester. These TERs are below the CoE/EPPO trigger
value of 10, indicating a high risk. The short-term results of feeding
studies, which represent more realistic exposure, would give TERs,
based on the LC50 values, of > 30 and > 29 for the salt and the
ester, respectively. These TERs are just below the CoE/EPPO trigger
for medium risk, but it should be noted that they are given as
'greater than'. As 2,4-D is more likely to be used in early growth
stages, it is likely that large insects dominate the diet of these
birds. As the initial residues on large insects are an order of
magnitude lower than on small ones, the acute oral TERs would rise to
46 for the salt and 42 for the ester, and the short-term dietary TERs
would rise to > 300 and > 290, respectively. Therefore, the acute
risk to small insectivorous birds from aerial applications of 2,4-D
dimethylamine salt or ethylhexyl ester is considered to be low. It
should be noted that 2,4-D has not been implicated in incidents of
poisoning in birds when used normally (UK Wildlife Incident
Investigation Scheme), which confirms that the risk to insectivorous
birds 2,4-D is low (personal communication from the Industry Task
Force II on 2,4-D Research Data).
(b) Ground-based application of 2.24 kg 2,4-D acid equivalents/ha
Grazing birds: The residues measured initially on short grass after
ground-based applications were 290 mg/kg grass of 2,4-D dimethylamine
salt and 459 mg/kg ethylhexyl ester. If a 3-kg greylag goose fed
exclusively on such contaminated grass, it would ingest 261 mg of the
salt or 413 mg of the ester, equal to 87 and 138 mg/kg bw,
respectively. Consequently, the acute oral TERs would be 5.7 for the
salt and 4.8 for the ester. These TERs are below the CoE/EPPO trigger
value of 10, indicating a high risk to grazing birds from aerial
applications of both 2,4-D dimethylamine salt and ethylhexyl ester.
The short-term results of feeding studies, which represent more
realistic exposure, would give TERs, based on the LC50 values, of
> 16 and > 12 for the salt and ester forms, respectively. These TERs
are above the CoE/EPPO trigger for high risk but may be below the
trigger for medium risk (< 100). As the residues in grass on day 7
had declined to between one-half and one-third of those on day 0 and
as 2,4-D has not been implicated in incidents of poisoning in birds
when used normally (UK Wildlife Incident Investigation Scheme), the
risk to grazing birds from 2,4-D is unlikely to be high (personal
communication from the Industry Task Force II on 2,4-D Research Data).
Insectivorous birds: The initial residues on small insects after
aerial overspray (based on 29 times the application rate in kg/ha as
mg/kg; CoE/EPPO vertebrate risk assessment scheme) would be 78 mg/kg
insects of the dimethylamine salt or 97.7 mg/kg ethylhexyl ester. On
the basis of the values for acute oral toxicity, if an 11-g blue tit
fed exclusively on contaminated insects, it would ingest 0.64 mg of
the salt or 0.8 mg of the ester, equal to 58.2 and 72.7 mg/kg bw,
respectively. Consequently, the acute oral TERs would be 9.1 for the
salt and 8.6 for the ester. These TERs are below the CoE/EPPO trigger
value of 10, indicating a potential risk to insect-eating birds. The
short-term results of feeding studies, which represent more realistic
exposure, would give TERs, based on the LC50 values, of > 59.5 and
> 57.5 for the salt and the ester, respectively. These TERs may be
below the CoE/EPPO trigger for medium risk (< 100), but it should be
noted that they are given as 'greater than'. As 2,4-D is more likely
to be used in early growth stages, it is likely that large insects
dominate the diet of these birds. As the initial residues on large
insects are an order of magnitude lower than on small ones, the acute
oral TERs would rise to 86 for the salt and 91 for the ester, and the
short-term dietary TERs would rise to > 300 and > 290, respectively.
Similarly, the short-term TERs would rise to > 595 for the salt and
> 755 for the ester. As these revised dietary TERs are > 100, the
acute risk to small insectivorous birds from ground-based application
of both 2,4-D dimethylamine salt or ethylhexyl ester is considered to
be low (personal communication from the Industry Task Force II on
2,4-D Research Data).
4.3.3.2 Mammals
LD50 values of 863 mg/kg bw for the dimethylamine salt and 896 mg/kg
bw for the ethylhexyl ester were reported for the rat, the most
sensitive species tested. The indicator mammals used in the risk
assessment were a 1200-g rabbit (Oryctolagus cuniculus), with a
total dally food consumption of 500 g vegetation (Ross, personal
communication), as the grazing mammal and an 18-g shrew (Sorex
araneus), with a total daily food consumption of 18 g (Churchfield,
1986), as the small insectivorous mammal.
Aerial applications (4.48 kg 2,4-D acid equivalents/ha for
assessing the risk to insectivorous mammals and 2.24 kg/ha for
assessing the risk to grazing mammals)
Grazing mammals: The residues measured initially on short grass
after aerial applications were 290 mg/kg grass for 2,4-D dimethylamine
salt and 459 mg/kg for the ethylhexyl ester. If a 1200-g rabbit fed
exclusively on such contaminated grass, it would ingest 145 mg of the
salt or 230 mg of the ester, equal to 121 and 192 mg/kg bw,
respectively. Consequently, the acute oral TERs would be 7.1 for the
salt and 4.7 for the ester. These TERs are below the CoE/EPPO trigger
value of 10, indicating a high risk to grazing mammals from aerial
applications of both 2,4-D dimethylamine salt and ethylhexyl ester. As
the residues in grass on day 7 had declined to between one-half and
one-third of those on day 0 and as 2,4-D has not been implicated in
incidents of poisoning in birds when used normally (UK Wildlife
Incident Investigation Scheme), the risk to grazing mammals from 2,4-D
is unlikely to be high (personal communication from the Industry Task
Force II on 2,4-D Research Data).
Insectivorous mammals: The initial residues on large insects after
aerial overspray (based on 2.7 times the application rate in kg/ha as
mg/kg; CoE/EPPO vertebrate risk assessment scheme) would be 14.5 mg/kg
insects for the dimethylamine salt or 18.2 mg/kg for the ethylhexyl
ester. On the basis of the values for acute oral toxicity, if an 18-g
shrew fed exclusively on contaminated insects, it would ingest 0.26 mg
of the salt or 0.33 mg of the ester, equal to 14.4 and 18.3 mg/kg bw,
respectively. Consequently, the acute oral TERs would be 60 for the
salt and 49 for the ester. These TERs are below the CoE/EPPO trigger
value of 100, indicating a medium risk to insect-eating mammals from
aerial application of 2,4-D, but it should be noted that 2,4-D has not
been implicated in incidents of poisoning in mammals when used
normally (UK Wildlife Incident Investigation Scheme), suggesting that
the risk to insectivorous mammals from 2,4-D is unlikely to be high.
(personal communication from the Industry Task Force II on 2,4-D
Research Data)
(b) Ground-based application of 2.24 kg 2,4-D acid equivalents/ha
Grazing mammals: The residues measured initially on short grass
after ground-based applications were 290 mg/kg grass for 2,4-D
dimethylamine salt and 459 mg/kg for the ethylhexyl ester. If a 1200-g
rabbit fed exclusively on such contaminated grass, it would ingest 145
mg of the salt or 230 mg of the ester, equal to 121 and 192 mg/kg bw,
respectively. Consequently, the acute oral TERs would be 7.1 for the
salt and 4.7 for the ester. These TERs are below the CoE/EPPO trigger
value of 10, indicating, a high risk to grazing mammals from
ground-based applications of both 2,4-D dimethylamine salt and
ethylhexyl ester. As the residues in grass on day 7 had declined to
between one-half and one-third of those on day 0 and as 2,4-D has not
been implicated in incidents of poisoning in birds when used normally
(UK Wildlife Incident Investigation Scheme), the risk to grazing
mammals from 2,4-D is unlikely to be high (personal communication from
the Industry Task Force II on 2,4-D Research Data).
Insectivorous mammals: The initial residues on large insects after
aerial overspray (based on 2.7 times the application rate in kg/ha as
mg/kg; CoE/EPPO vertebrate risk assessment scheme) would be 7.26 mg/kg
insects for the dimethylamine salt or 9.1 mg/kg for the ethylhexyl
ester. On the basis of the values for acute oral toxicity, if an 18-g
shrew fed exclusively on contaminated insects, it would ingest 0.13 mg
of the salt or 0.16 mg of the ester, equal to 7.2 and 8.9 mg/kg bw,
respectively. Consequently, the acute oral TERs would be 120 for the
salt and 101 for the ester. These TERs are above the CoE/EPPO trigger
value of 100, indicating a low risk to small insect-eating mammals
from ground-based application of 2,4-D (personal communication from
the Industry Task Force II on 2,4-D Research Data).
5. Evaluation of effects on the environment
2,4-Dichlorophenoxyacetic acid (2,4-D) is a selective herbicide, which
is available as the free acid, salts, and esters. It has low
volatility and should not partition to air after application. Amine
salt formulations of 2,4-D are less volatile than butyl, ethyl, or
isopropyl ester formulations.
The 2-ethylhexyl ester is hydrolysed under alkaline conditions
(half-life, 48 days at pH 7 and 2.2 days at pH 9). 2,4-D may be
degraded slowly by photolysis. Its half-life in aqueous solution was
4.5 days under aerobic conditions and 312 days under anaerobic
conditions. The major breakdown product was carbon dioxide;
2,4-dichlorophenol, 2,4-dichloroanisole, and 4-chlorophenoxyacetic
acid were formed as intermediates. The half-lives of 2,4-D in natural
waters after aerial application of its dimethylamine salt ranged from
1.1 to 20 days. 2,4-D formulations were found to be rapidly hydrolysed
or biodegraded in ponds and lakes. There was no evidence that 2,4-D
bioaccumulates in aquatic organisms.
The behaviour of 2,4-D salts and esters in soils is greatly influenced
by the organic matter content and pH: 2,4-D is more strongly adsorbed
in soils with a higher organic matter content and/or lower pH values.
The rapid biodegradation of 2,4-D in soil prevents significant
downward movement under normal field conditions. Run-off from treated
soil has been estimated at 0.01 - 1% of the applied 2,4-D; the maximum
recorded concentration after run-off was about 0.2 µg/L. In
non-sterile soils, various esters of 2,4-D are hydrolysed very rapidly
(> 72% within 72 h). A number of microbial organisms rapidly degrade
2,4-D, with half-lives of 1.25 h to 40 days, usually between 3 and 10
days. The dimethylamine salt dissociates rapidly, leaving 2,4-D that
then undergoes further degradation.
Field trials in the United States with the dimethylamine salt or the
ethylhexyl ester at 2.24 kg acid equivalents/ha on grass resulted in
maximum initial residues at day 0 of 120 or 153 mg acid
equivalents/kg, respectively. These initial residues had decreased by
one-half to one-third by day 7.
In general, populations of aerobic bacteria, actinomycetes, and fungi
in soils were not affected by 2,4-D at 25 ppm. Application of the
isooctyl ester at a rate of 0.95 kg/ha, however, reduced the
populations of bacteria by 26.3%, those of fungi by 19.5%, and those
of actinomycetes by 30%; the dimethylamine salt reduced these
populations by about half as much.
2,4-D applied at the maximum recommended rate stimulated (by 10%) the
growth of Skeletonema costatum, whereas it inhibited that of
Navicula pelliculosa (24%) and Lemna gibba (75%). The five-day
EC50 values for the toxicity of 2,4-D and its salts and esters differ
widely with respect to both algal species and compounds, ranging from
0.23 mg ai/L for the ethylhexyl ester in Skeletonema costatum to 153
mg ai/L for the dimethylamine salt in Anabaena flosaquae. The acute
toxicity of 2,4-D to the aquatic higher plant Lemna gibba also
depended upon the salt or ester used, with 14-day EC50 values of 3.3
mg/L for 2,4-D, 0.58 mg/L for the dimethylamine salt, and 0.5 mg/L for
the ethylhexyl ester. At concentrations of 0.001-100 mg/L, 2,4-D had
no effect on chlorophyll production in several algal species.
Anabaenopsis raciborskii tolerated up to 800 µg/ml of 2,4-D in
liquid culture.
Many studies have been performed on invertebrate freshwater,
estuarine, and marine species, including Daphnia, Gammarus,
Macrobranchium, Crassostrea, Palaernonetes, Panaeus, and Uca, and
many forms of 2,4-D have been evaluated, including 2,4-D itself, the
dimethylamine, diethanolamine, isopropylamine, tri-isopropanolamine,
and sodium salts, and the ethylhexyl, butoxyethyl, and isopropyl
esters. 2,4-D and its salts are generally less toxic to these
organisms than the ester forms. The 48-h toxicity (LC50) values to
Daphnia magna ranged from 5.2 mg ai/L for the isooctyl ester to 184
mg ai/L for the dimethylamine salt, and the 21-day NOEC values ranged
from 0.0015 mg ai/L for the ethylhexyl ester to 27.5 mg ai/L for the
dimethylamine salt and 79 mg ai/L for the acid.
Grass shrimp avoided water containing concentrations of 1 or 10 mg/L
of the butoxyethyl ester. In a study covering the life cycle of
Daphnia magna, the NOEC was 23.6 mg/L diethanolamine salt. In a
study of the long-term toxicity study of the butoxyethyl ester in
Daphnia magna, the maximum acceptable toxicant concentration was
0.70-0.29 mg/L.
The 96-h LC50 values for frog and toad tadpoles ranged from 8 mg/L
for the free acid to 477 mg/L for the dimethylamine saltœ
The effects of 2,4-D and its salts and esters have been studied on
various growth stages of fish species such as Oncorhynchus,
Lepomis, Pimephales, Gambusia, Micropterus, and Salmo. Generally,
2,4-D and its salts are less toxic to fish than are the esters Typical
96-h LC50 values for adult fish were 5-10 mg ai/L for the isooctyl
ester, 200-400 mg ai/L for 2,4-D, and 250-500 mg ai/L for the
diethanolamine salt, although lower figures have been reported. Fish
in early life stages appear to be more sensitive, with 32-day NOEC
values ranging from 0.12 mg ai/L for the ethylhexyl ester to 17.1 mg
ai/L for the diethanolamine salt and 63.4 mg ai/L for the acid.
The NOEC for embryos and larvae of the fathead minnow, Pimephales
promelas, exposed to < 416.1 µg/L of the butoxyethyl ester for 32
days was 80.5 µg/L, and the maximum acceptable toxicant concentration
was estimated to be 96 µg/L. The sodium salt of 2,4-D did not inhibit
the hatching of carp eggs at 25 mg/L, but at a concentration of 100
mg/L none hatched. At 50 mg ai/L, the sodium salt was not harmful to
carp embryos but induced behavioural changes, some morphopathological
alterations, and, ultimately, death in carp larvae.
The LD50 values for the dimethylamine salt and ethylhexyl ester for
the honeybee after oral exposure and contact were > 100 µg/bee. No
toxic effects have been seen in bees in the field.
2,4-D (in combination with MCPA) did not harm. Trichogramma
cacoeciae at 1.5% in water or Aleochara bilineata at the
recommended rate. Mixed amine salts and mixed isopropyl esters of
2,4-D were toxic to coccinellid larvae and to sawflies. No
reproductive effects were observed in European cockroaches reared on
food containing 1000 mg/kg 2,4-D (unspecified).
Application of 2,4-D at 1250 g ai/L in field crops did not affect
staphylinids, carabids, or spiders during a 20-month observation
period. Some adult millipedes exposed to 2,4-D at a rate of 33.6 kg/ha
died on the first day, and the rate was 50% higher than that of
controls by day 7.
The 14-day LC50 for earthworms exposed to the dimethylamine salt was
350 mg/kg soil, but no deaths occurred at concentrations < 100 mg
ai/kg. A 48-h LC50 of 61.6 µg/cm2 was reported for earthworms
exposed on filter paper.
The LD50 values for birds are 200 to > 2000 mg/kg bw for mallards,
bobwhite quail, Japanese quail, pheasants, chukar partridges, and rock
doves The dietary LC50 values for the acid, salt, and ester exceeded
4640 mg/kg diet for mallards, bobwhite quail, Japanese quaff, and
pheasants. Application of formulations at doses greater than the
recommended rate did not adversely affect the reproductive performance
of pheasants, quail, partridges, or chickens.
The oral LD50 values for the acid and its salts and esters in rats
and rabbits were 699-2322 mg/kg bw. The dermal LD50 value for rabbits
was > 2000 mg/kg bw, and the LC50 value after inhalation was
1.8-10.7 mg/L.
5.1 Risk assessment
The information on use and application rates used in this risk
assessment is based on the agricultural use of 2,4-D within the
European Union and the United States. 2,4-D can be formulated as a
variety of salts (e.g. dimethylamine, sodium, diethanolamine,
tri-isopropanolamine, and isopropylamine) and esters (ethylhexyl,
isooctyl, and butoxyethyl); however, the dimethylamine salt and
ethylhexyl ester account for about 95% of the global use of 2,4-D.
This risk assessment is therefore restricted to the use of those
compounds. Both are rapidly hydrolysed to 2,4-D. The main uses of
2,4-D include application to cereals, corn, sorghum, soya beans, sugar
cane, rice, pasture, top fruit, turf, non-cropland, fallow land,
forests, and aquatic weeds. Applications can be made by either
conventional tractor-mounted or -drawn hydraulic sprayers or by aerial
application (e.g. in forests) at rates of 0.25-4.48 kg acid
equivalent/ha.
This risk assessment is based on the principle of calculating
toxicity:exposure ratios (TERs) and on the CoE/EPPO Environmental Risk
Assessment models and trigger values.
5.1.1 Aquatic environment
The main risk to aquatic organisms from the use of 2,4-D is due to
overspray during aerial use, spray drift from ground-based hydraulic
applications, or use to control aquatic weeds. The CoE/EPPO risk
assessment scheme for aquatic organisms showed low acute risk (TERs
> 10) to fish, aquatic invertebrates, and algae due to spray drift
arising from ground-based hydraulic applications and to overspray
during aerial applications. A potential acute risk (TER < 10) to
higher aquatic plants and amphibia due to overspray during aerial
application was identified. The use of 2,4-D to control aquatic weeds
also presents a potential acute risk (TER < 10) to algae and higher
aquatic plants, but this risk can be ignored as these organisms are
the targets of 2,4-D used in this way. A potential acute risk to
amphibia remains from use to control aquatic weeds, but this risk must
be balanced against those associated with alternative means of aquatic
weed control, such as no control (e.g. algal bloom leading to water
deoxygenation) and manual weed control, both of which may pose a
higher risk to fish and other aquatic organisms. The ethylhexyl ester
is not recommended for this use.
Owing to the very rapid degradation of the salts and esters of 2,4-D
in water, the long-term risk to aquatic organisms from these compounds
was considered to be low. The primary breakdown product, 2,4-D acid,
is, however, more persistent in water; therefore, the long-term risk
assessment was based on its use. The levels of 2,4-D measured in
surface waters after approved uses range from 0.00008 mg/L in small
watersheds in Saskatchewan, Canada, to 0.0021 mg/L in ground- and
surface waters in the United Kingdom. These values indicate that the
long-term risk to fish and invertebrates living in water columns and
sediments is low.
Terrestrial environment
(a) Microorganisms
The most important sources of exposure of soil microorganisms to 2,4-D
are likely to be ground and aerial applications. The results of
laboratory studies indicate that application of 2,4-D at rates of 7.5
kg/ha aerially and 18.75 kg/ha on the ground should pose a low risk to
soil microorganisms; these rates correspond to 1.4 and 4.7 times the
maximum recommended single or seasonal application rate. Application
of the dimethylamine salt and the isooctyl ester at rates
corresponding to 0.95 kg 2,4-D/ha resulted in 10-30% reductions in
populations of soil bacteria, fungi, and actinomyces; the ester caused
greater reductions than the salt. As the trigger for concern in the
CoE/EPPO microorganism risk assessment scheme is an effect > 30%, the
risk to soil microorganisms from the use of 2,4-D should be low.
(b) Plants
2,4-D is a translocated, selective herbicide used to control a variety
of broad-leaved weeds. Consequently, although it may pose a risk to
broad-leaved non-target plants, this is to be expected from its mode
of action and consequent use.
(c) Invertebrates
Bees may be exposed to 2,4-D while foraging flowering weeds present in
treated crops. At the maximum individual application rate of 4.48 kg
acid equivalent/ha, the hazard quotients for toxicity after oral
exposure and contact were > 45 for both the dimethylamine salt and
the ethylhexyl ester. As the CoE/EPPO trigger for concern is a hazard
quotient > 50, the acute risk to honeybees from the use of 2,4-D at
this rate of application should be low. This conclusion is supported
by the fact that 2,4-D has not been implicated in any incident of
poisoning of honeybees in the UK Wildlife Incident Investigation
Scheme.
Arthropods may be exposed to 2,4-D during its many agricultural and
non-agricultural uses. On the basis of the CoE/EPPO triggers for
concern with regard to effects on non-target arthropods in laboratory
studies (i.e. effects > 30%), 2,4-D may pose a risk to arthropods at
high rates of application; however, the data on which this conclusion
is based either related to a joint formulation with MCPA or were old
and perhaps unreliable. Limited data from studies in the field with
the lower, more typical rates of application (< 1.25 kg/ha)
indicate that this risk may not be present in the field.
Earthworms may be exposed after single or multiple applications of
2,4-D to a wide variety, of crops but in particular during its use on
grass, fallow land, and stubble. The TER at a maximum rate of
application of the dimethylamine salt at 5.37 kg/ha is greater than
the CoE/EPPO trigger value of 10, which indicates that the acute risk
to earthworms from the use of 2,4-D should be low.
(d) Vertebrates
Vertebrates are likely to be exposed to 2,4-D while grazing on treated
or contaminated vegetation or consuming contaminated insects.
Estimates of residues on food items represent the maximum values
determined immediately after application and do not take into account
the rapid degradation of 2,4-D in the environment. The risk assessment
further assumes that all food consumed contains 2,4-D at the level of
the MRL.
The short-term dietary TERs, based on measured initial residues on
short grass after application of 2.24 kg acid equivalent/ha, indicate
a potential medium risk (10 < TER < 100) to grazing birds from both
aerial and ground-based applications. The initial residues declined to
one-half or one-third within seven days of application. Furthermore,
2,4-D has not been implicated in any incident of poisoning in birds as
a result of normal use, suggesting that the risk to grazing birds is
unlikely to be high. The short-term dietary TERs based on initial
residues on large insects predicted in the CoE/EPPO vertebrate risk
assessment scheme indicate a low acute risk (TER > 100) to small
insectivorous birds from both aerial and ground applications (4.48 kg
acid equivalent/ha and 2.24 kg acid equivalent/ha, respectively).
Large insects are likely to constitute a higher proportion of both
avian and mammalian diets than small insects during early growth stage
or pm-emergence use.
The acute oral TERs based on measured initial residues on short grass
arising from application at 2.24 kg acid equivalent/ha indicate a
potentially high risk (TER < 10) to grazing mammals from both aerial
and ground-based applications, but the initial residues had declined
to one-half or one-third within seven days of application. The acute
oral TERs based on predicted initial residues on large insects
indicate, however, a medium acute risk (10 < TER < 100) from aerial
applications and a low acute risk (TER > 100) from ground-based
applications, to small insectivorous mammals. It should be noted that
2,4-D has not been implicated in any incidents of poisoning in mammals
as a result of normal use. This suggests that the risk to mammals from
2,4-D is unlikely to be high.
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See Also:
Toxicological Abbreviations