Dichlorophenoxyacetic acid, 2,4- (2,4-D): environmental aspects (EHC 84, 1989)

INTERNATIONAL PROGRAMME ON CHEMICAL SAFETY

ENVIRONMENTAL HEALTH CRITERIA 84

2,4-DICHLOROPHENOXYACETIC ACID (2,4-D) - ENVIRONMENTAL ASPECTS

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policy of the United Nations Environment Programme, the International
Labour Organisation, or the World Health Organization.

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the International Labour Organisation,
and the World Health Organization

World Health Orgnization
Geneva, 1989

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CONTENTS

ENVIRONMENTAL HEALTH CRITERIA FOR 2,4-DICHLOROPHENOXYACETIC ACID
(2,4-D) - ENVIRONMENTAL ASPECTS

1. SUMMARY AND CONCLUSIONS

1.1. Uptake, accumulation, elimination, and biodegradation
1.2. Toxicity to microorganisms
1.3. Toxicity to aquatic organisms
1.4. Toxicity to terrestrial organisms
1.5. Effects of 2,4-D in the field

2. PHYSICAL AND CHEMICAL PROPERTIES

2.1. Synthesis of 2,4-D
2.2. Important chemical reactions of 2,4-D
2.3. Volatility of 2,4-D derivatives

3. SOURCES OF ENVIRONMENTAL POLLUTION

3.1. Production of 2,4-D herbicides
3.2. Uses
3.3. Disposal of wastes

4. UPTAKE, ACCUMULATION, ELIMINATION, AND BIODEGRADATION

4.1. Biodegradation
4.2. Uptake and accumulation by organisms
4.2.1. Laboratory studies
4.2.2. Field studies
4.3. Elimination

5. TOXICITY TO MICROORGANISMS

5.1. Aquatic microorganisms
5.2. Soil microorganisms

6. TOXICITY TO AQUATIC ORGANISMS

6.1. Toxicity to aquatic invertebrates
6.1.1. Short-term toxicity
6.1.2. Behavioural effects
6.2. Toxicity to fish
6.2.1. Effect of formulation on short-term toxicity to fish
6.2.1.1 Tolerance and potentiation
6.2.2. No-observed-effect-levels in short-term tests with fish
6.2.3. Species differences in short-term toxicity to fish
6.2.4. Toxicity to early life-stages of fish
6.2.5. Long-term toxicity to fish
6.2.6. Behavioural effects on fish
6.2.7. Effects of environmental variables on toxicity to fish
6.2.8. Special studies on fish
6.3. Toxicity to amphibians

7. TOXICITY TO TERRESTRIAL ORGANISMS

7.1. Toxicity to terrestrial invertebrates
7.2. Toxicity to birds
7.2.1. Toxicity to birds' eggs
7.2.2. Toxicity to birds after short-term and long-term dosing
7.2.3. Special studies on birds
7.3. Toxicity to non-laboratory mammals

8. ECOLOGICAL EFFECTS FROM FIELD APPLICATION

9. EVALUATION

9.1. Aquatic organisms
9.2. Terrestrial organisms

10. RECOMMENDATIONS FOR FURTHER RESEARCH

REFERENCES

WHO TASK GROUP ON ENVIRONMENTAL HEALTH CRITERIA FOR
2,4-DICHLOROPHENOXYACETIC ACID (2,4-D) - ENVIRONMENTAL ASPECTS

Members

Dr L.A. Albert, Director, Environmental Pollution Programme, National
Institute for Research on Biotic Resources, Veracruz, Mexico
Mr H. Craven, Ecological Effects Branch, Office of Pesticides
Programs, US Environmental Protection Agency, Washington DC, USA
Dr A.H. El-Sebae, Division of Pesticide Toxicology, Faculty of
Agriculture, Alexandria University, Alexandria, Egypt
Dr J.W. Everts, Department of Toxicology, Agricultural University,
Wageningen, Netherlands
Dr W. Fabig, Fraunhofer Institute for Environmental Chemistry and
Ecotoxicology, Schmallenberg-Grafschaft, Federal Republic of
Germany
Dr R. Koch, Division of Toxicology, Research Institute for Hygiene and
Microbiology, Bad Elster, German Democratic Republic (Chairman)
Dr Y. Kurokawa, Division of Toxicology, Biological Safety Research
Centre, National Institute of Hygienic Sciences, Tokyo, Japan
Dr E.D. Magallona, Pesticide Toxicology and Chemistry Laboratory,
University of the Philippines at Los Banos, College of Agriculture,
Laguna, Philippines
Professor P.N. Viswanathan, Ecotoxicology Section, Industrial Toxi-
cology Research Centre, Lucknow, India

Observers

Dr M.A.S. Burton, The Monitoring and Assessment Research Centre,
London, United Kingdom
Dr I. Newton, The Institute of Terrestrial Ecology, Monks Wood
Experimental Station, Huntingdon, United Kingdom

Secretariat

Dr S. Dobson, The Institute of Terrestrial Ecology, Monks Wood
Experimental Station, Huntingdon, United Kingdom (Rapporteur)
Dr M. Gilbert, International Programme on Chemical Safety, World
Health Organization, Geneva, Switzerland (Secretary)
Mr P.D. Howe, The Institute of Terrestrial Ecology, Monks Wood
Experimental Station, Huntingdon, United Kingdom

NOTE TO READERS OF THE CRITERIA DOCUMENTS

Every effort has been made to present information in the criteria
documents as accurately as possible without unduly delaying their
publication. In the interest of all users of the environmental health
criteria documents, readers are kindly requested to communicate any
errors that may have occurred to the Manager of the International
Programme on Chemical Safety, World Health Organization, Geneva,
Switzerland, in order that they may be included in corrigenda, which
will appear in subsequent volumes.

* * *

A detailed data profile and a legal file can be obtained from the
International Register of Potentially Toxic Chemicals, Palais des
Nations, 1211 Geneva 10, Switzerland (Telephone no. 988400 - 985850).

ENVIRONMENTAL HEALTH CRITERIA FOR 2,4-DICHLOROPHENOXYACETIC ACID
(2,4-D) - ENVIRONMENTAL ASPECTS

A WHO Task Group on Environmental Health Criteria for
2,4-Dichlorophenoxyacetic acid (2,4-D) - Environmental Aspects met at
the Institute of Terrestrial Ecology, Monks Wood, United Kingdom, from
14 to 18 December 1987. Dr I. Newton welcomed the participants on
behalf of the host institution, and Dr M. Gilbert opened the meeting on
behalf of the three co-sponsoring organizations of the IPCS
(ILO/UNEP/WHO). The Task Group reviewed and revised the draft criteria
document and made an evaluation of the risks for the environment from
exposure to 2,4-D.

The first draft of this document was prepared by Dr S. Dobson and
Mr P.D. Howe, Institute of Terrestrial Ecology. Dr M. Gilbert and
Dr P.G. Jenkins, both members of the IPCS Central Unit, were respon-
sible for the overall scientific content and editing, respectively.

* * *

Partial financial support for the publication of this criteria
document was kindly provided by the United States Department of Health
and Human Services, through a contract from the National Institute of
Environmental Health Sciences, Research Triangle Park, North Carolina,
USA - a WHO Collaborating Centre for Environmental Health Effects.

INTRODUCTION

There is a fundamental difference in approach between the
toxicologist and the ecotoxicologist concerning the appraisal of the
potential threat posed by chemicals. The toxicologist, because his
concern is with human health and welfare, is preoccupied with any
adverse effects on individuals, whether or not they have ultimate
effects on performance or survival. The ecotoxicologist, in contrast,
is concerned primarily with the maintenance of population levels of
organisms in the environment. In toxicity tests, he is interested in
effects on the performance of individuals - in their reproduction and
survival - only insofar as these might ultimately affect the population
size. To him, minor biochemical and physiological effects of toxicants
are irrelevant if they do not, in turn, affect reproduction, growth, or
survival.

It is the aim of this document to take the ecotoxicologist's point
of view and consider effects on populations of organisms in the
environment. The risk to human health of the use of 2,4-D was
evaluated in Environmental Health Criteria 29: 2,4-Dichlorophenoxy-
acetic acid (WHO, 1984). This document did not consider effects on
organisms in the environment, but did consider environmental levels of
2,4-D likely to arise from recommended uses. No attempt has been made
here to reassess the human health risk; the interested reader should
refer to the original document, which contains the relevant literature
in this area.

This document, although based on a thorough survey of the
literature, is not intended to be exhaustive in the material included.
In order to keep the document concise, only those data which were
considered to be essential in the evaluation of the risk posed by 2,4-D
to the environment have been included.

The term bioaccumulation indicates that organisms take up chemicals
to a greater concentration than that found in their environment or
their food. `Bioconcentration factor' is a quantitative way of
expressing bioaccumulation: the ratio of the concentration of the
chemical in the organism to the concentration of the chemical in the
environment or food. Biomagnification refers, in this document, to the
progressive accumulation of chemicals along a food chain.

1. SUMMARY AND CONCLUSIONS

2,4-D is a selective herbicide which kills broad-leaved plants but
not grasses or conifers. Its chemical structure is a modification of a
naturally occuring plant hormone. 2,4-D is available as the free acid
but is used, in agriculture and forestry, in formulations as a salt or
ester.

1.1. Uptake, Accumulation, Elimination, and Biodegradation

2,4-D does not persist in soil because of its rapid degradation.

The physico-chemical properties of 2,4-D acid and its formulations
have an important effect on its behaviour in environmental
compartments.

The bioavailability to, and uptake by, aquatic and terrestrial
organisms is strongly influenced by the organic matter content of
soils, microbiological activity, and by environmental conditions such
as temperature and pH. Although highly inconsistent, the data on
dissipation and bioavailability in various soils demonstrate a marked
influence of differences in the texture and mineral composition of the
soil. In aerobic soils, with a high content of organic material, and
at high pH values and temperatures, toxic effects are limited because
of rapid degradation of 2,4-D.

Uptake is followed by rapid excretion in most organisms. With the
exception of some algae, the retention of 2,4-D by organisms in the
environment cannot be expected, because of its rapid degradation.

Some microorganisms are capable of utilizing 2,4-D as their sole
carbon source. Repeated application to soil stimulates the number of
organisms capable of degrading the compound.

1.2. Toxicity to Microorganisms

In general 2,4-D is relatively non-toxic to water and soil
microorganisms at recommended field application rates. No effect of
2,4-D was recorded on 17 genera of freshwater and two genera of marine
algae at concentrations up to 222 mg/litre. No effect of 2,4-D on
respiration of either sandy loam or clay loam soils was observed at
concentrations up to 200 mg/kg.

N-fixation by aquatic algae is affected at high concentrations of
2,4-D acid (400 mg/litre). An effect of 2,4-D esters on N-fixation
occurs from a concentration of 36 mg/litre upwards. N-fixing algae in
topsoils appear to be more vulnerable to 2,4-D acid than other algal
species. The Cyanobacteria (blue-green algae) are important as the
major N₂ source in tropical ponds and soils.

In the range of 25.2 to 50.4 mg/litre, 2,4-D was inhibitory to all
types of soil fungi.

Cell division was reduced in a green alga by 2,4-D at 20 mg/litre
and stopped at 50 mg/litre. No effect was observed on a natural
phytoplankton community after exposure to 2,4-D at 1 mg/litre.

However, exposure to esters of 2,4-D reduced productivity in these
organisms.

1.3. Toxicity to Aquatic Organisms

At recommended application rates, the concentration of 2,4-D in
water has been estimated to be a maximum of 50 mg/litre. Most
applications would lead to water concentrations much lower than this
(between 0.1 and 1.0 mg/litre).

The short-term toxicity data on the effects of 2,4-D free acid,
its salts, and esters on aquatic invertebrates is extensive. Ester
formulations are more toxic than the free acids or salts. Sensitivity
variations exist among species in response to the same formulation.
Organisms become more sensitive to 2,4-D when the water temperature
increases. Reproductive impairment occurred at concentrations below
0.1 of the short-term toxic levels determined for these formulations.

LC₅₀ values for fish vary considerably. This variation is partly
due to differences in species sensitivity, chemical structure (esters,
salts, or free acid), and formulation of the herbicide.

Although the free acid is the physiologically toxic entity, the
ester formulations represent a major hazard to fish when used directly
as aquatic herbicides (because they are more readily taken up by fish).
Amine salt formulations used to control aquatic weeds do not affect
adult fish.

The no-observed-effect-level (NOEL) varies with the species and the
formulation: less than 1 mg/litre (coho salmon) to 50 mg/litre (rainbow
trout).

Fish larvae are the most sensitive life stage but are unlikely to
be affected under normal usage of the herbicide.

Long-term adverse effects on fish are observed only at
concentrations higher than those produced after 2,4-D has been applied
at recommended rates.

Few studies are related to the effects of environmental variables,
such as temperature and water hardness, on 2,4-D toxicity to fish.
Higher temperature possibly increases the toxicity. This might be
considered when assessing the safety of 2,4-D to fish during control of
aquatic weeds.

Fish detect and avoid 2,4-D only at higher concentrations than
those obtained under normal conditions of use.

Amphibian larvae are generally tolerant to amine salts of 2,4-D;
the 96-h LC₅₀ values exceed 100 mg/litre. Of the species tested, only
one was sensitive. No information is available on reproductive
development and differentiation or on tissue levels.

1.4. Toxicity to Terrestrial Organisms

Based on the widespread use of 2,4-D and its formulations, insects
of many kinds could be exposed to the material. Although the compounds
are generally classified as non-toxic for beneficial insects, such as
honey bees and natural enemies of pests, some adverse effects have been
reported on the early life-stages and adults of some insects.

Esters are less toxic to insects than are salts or the free acid.

Birds, and particularly the eggs of ground-nesting species, would
be exposed to 2,4-D after spraying. Food items could also be expected
to be contaminated by the herbicide. However, most studies on birds
and their eggs have been conducted at exposures far higher than could
be expected in the field.

LD₅₀ values from acute oral and from short-term dietary dosing
indicate low toxicity of 2,4-D to birds. In longer-term studies,
effects have only been reported at extremely high exposures (for
example, kidney effects after dosing in drinking water with
concentrations in excess of the solubility of the material). There
have been no reported effects on reproductive parameters, even at
excessive exposure levels.

A single study reported adverse effects on the embryos of birds'
eggs sprayed with 2,4-D. Many studies since have shown no effect on
hatchability of eggs and no increased incidence of abnormalities in
chicks even after very high exposure to 2,4-D. Other work indicates a
very poor penetration of the eggshell by the herbicide. It can only be
concluded that after normal, or even after excessive, 2,4-D use, there
would be no effect on birds' eggs.

Based on the available data, no generalization can be made about
the hazard of 2,4-D to mammals in the field. Data on voles indicate
that the herbicide poses no hazard.

1.5. Effects of 2,4-D in the Field

No direct toxic effects, acute or long-term, of 2,4-D applications
under field conditions on any animals species have been observed thus
far.

There are, inevitably, indirect effects resulting from the intended
selective herbicidal properties of the compound. These effects would
result from the use of any herbicide or from other methods of land
management. There will, therefore, be effects for mammals, birds, and
insects because of food deprivation, modification of habitat,
requirements for nesting, shelter, etc.

The application of 2,4-D appears to be less hazardous to the
beneficial epigeal arthropod community than physical cultivation.

2. PHYSICAL AND CHEMICAL PROPERTIES

Details of the physical and chemical properties of 2,4-
dichloropheoxyacetic acid (2,4-D) are given in Environmental Health
Criteria 29: 2,4-D (WHO, 1984). The relevant chapter is summarized
here.

The structures of 2,4-D and of chemically-related phenoxy
herbicides in common use are given in Fig. 1. 2,4-D is a chlorinated
form of a natural plant hormone (auxin).

Some physical properties of 2,4-D and of the 2,4-D derivatives that
are used in agriculture are summarized in Tables 1 & 2.

2,4-D has growth-regulating and herbicidal properties in broad-
leaved plants. Because of its solubility, 2,4-D is rarely used in the
form of the acid; commercial 2,4-D herbicide formulations consist of
the more soluble forms such as alkali salts, amine salts, or esters.
These are combined with solvents, carriers, or surfactants and are
marketed in the form of dusts, granules, emulsions, or oil and water
solutions in a wide range of concentrations.

Table 1. Physical properties of 2,4-D
-------------------------------------------------------
Molecular formula C₈H₆Cl₂O₃

Relative molecular mass 221.0

Melting point 140 - 141 °C

Solubility in water slightly soluble

Solubility in organic solvents soluble

Vapour pressure 52.3 Pa at 160 °C

pKa at 25 °C 2.64 - 3.31
-------------------------------------------------------

2.1. Synthesis of 2,4-D

2,4-D is commonly prepared by the condensation of 2,4-dichloro-
phenol with monochloroacetic acid in a strongly alkaline medium at
moderate temperatures or by the chlorination of phenoxyacetic acid, but
this method leads to a product with a high content of 2,4-dichloro-
phenol and other impurities. Higher reaction temperatures and
alkaline conditions during the manufacture of 2,4-D increase the
formation of polychlorinated dibenzo- p -dioxin (CDD) by-products. One
formulation of 2,4-D was found to contain 6.8 µg/kg of 2,3,7,8-
tetrachlorinated dibenzo- para -dioxin (Hagenmaier, 1986). In other
amine and ester formulations, levels of this dioxin were non-
detectable, i.e., < 1 µg/kg (WHO, 1984). The alkali metal salts of
2,4-D are produced by the reaction of 2,4-D with the appropriate metal
base. Amine salts are obtained by reacting stoichiometric quantities
of amine and 2,4-D in a compatible solvent. Esters are formed by
acid-catalysed esterification with azeotropic distillation of water or

by direct synthesis in which the appropriate ester of monochloroacetic
acid is reacted with dichlorophenol to form the 2,4-D ester.

2.2. Important Chemical Reactions of 2,4-D

Pyrolysis converts various amine salts of 2,4-D to the
corresponding amides. Pyrolysis of 2,4-D and its derivatives is likely
to produce certain CDD isomers. 2,4-D is readily photodegraded.

2.3. Volatility of 2,4-D Derivatives

2,4-D esters with short-chain alcohols are highly volatile. This
influences the effectiveness of their application to target crops,
their effects on neighbouring crops, and the degree of contamination of
the atmosphere. 2,4-D alkali salts or amine salts are much less
volatile than esters, and these products are to be preferred when the
use of 2,4-D esters might lead to evaporative 2,4-D losses and to crop
damage or damage to the surrounding environment.

Details of technical compositions, impurities, and analytical
methods can be found in Environmental Health Criteria 29: 2,4-
Dichlorophenoxyacetic acid (WHO, 1984).

Table 2. Vapour pressure and solubility of 2,4-D salts and esters
--------------------------------------------------------------------------------
Compound Vapour pressure^a Solubility
--------------------------------------------------------------------------------
2,4-D free acid 0.4 mmHg (160 °C) 0.09% in water (25 °C),
85% in acetone (25 °C)

dimethylamine salt 300% in water (20 °C),
soluble in acetone

isopropyl ester 1.4 x 10^-3 mmHg^b insoluble in water, soluble
4.6 x 10^-5 mmHg^b in most organic solvents

butoxyethanol ester 4.5 x 10^-6 mmHg^b insoluble in water, soluble
(butylethyl ester) in most organic solvents

ethylhexyl ester 2.0 x 10^-6 mmHg^b insoluble in water, soluble
in organic solvents

isooctyl ester 2.0 x 10^-6 mmHg^b insoluble in water, soluble
in organic solvents

propyleneglycol butyl 3.0 x 10^-6 mmHg^b insoluble in water, soluble
ether ester in organic solvents

methyl ester 2.3 x 10^-3 mmHg^b

ethyl ester 1.1 x 10^-3 mmHg^b

butyl ester 3.97 x 10^-4 mmHg^b
--------------------------------------------------------------------------------
^a 1 mmHg = 0.133 kPa.
^b Vapour pressures of esters were determined at high temperatures by gas-
liquid chromatography, and these values are the result of extrapolation
to 25 °C. Values vary considerably between authors as a result of this
extrapolation; original values at high temperatures agree. Results are
presented here as an indication of relative vapour pressure at working
temperature. Values from Flint et al. (1968) and Jensen & Schall
(1966).

3. SOURCES OF ENVIRONMENTAL POLLUTION

The following is a summary of the chapter from Environmental Health
Criteria 29: 2,4-Dichlorophenoxyacetic acid (WHO, 1984).

3.1. Production of 2,4-D Herbicides

Comprehensive statistics on 2,4-D herbicide production or use were
not available for review. According to the US Department of
Agriculture, 3 x 10⁸ kg of total herbicides were used in the USA
alone, in 1981. In the past, 10% of the herbicide used was 2,4-D,
which would account for a total use in the USA of about 3 x 10⁷ kg.
In 1975, an estimated 5 x 10⁶ kg were produced in the United Kingdom.
World-wide use of herbicides and annual production, which probably
exceeds 5 x 10⁷ kg/year, are increasing.

3.2. Uses

2,4-D alkali or amine salts or esters are used as agricultural
herbicides against broad-leaved weeds in cereal crops, as well as on
pastures and lawns, in parks, and on golf courses, at rates of about
0.2 to 2.0 kg active ingredient (acid equivalent) per hectare. Esters
are also used at rates of up to 6.0 kg (acid equivalent) per hectare to
suppress weeds, brush, and deciduous trees along rights-of-way and in
conifer plantations and conifer reafforestation areas.

Granular formulations of 2,4-D are used as aquatic herbicides in or
along irrigation and other canals, in ponds, and lakes at rates ranging
from 1 to 122 kg/ha.

2,4-D products can be used at very low application rates as
growth regulators by application of aqueous foliar sprays containing 20
to 40 mg 2,4-D/litre on apple trees to reduce premature fruit-drop, on
potato plants to increase the proportion of medium-size tubers or to
intensify the tuber skin colour of the red varieties, and in citrus
culture to reduce pre-harvest fruit-drop and to increase fruit storage
life.

The highly volatile ethyl, isopropyl, and butyl esters are being
replaced by low-volatile esters or by amine salts to reduce crop damage
resulting from 2,4-D vapour drift, and to decrease atmospheric
pollution.

During recent years, the use of 2,4-D and 2,4,5-T in parks,
forested recreation, and other areas frequently used by the public, has
been reduced in some countries because of increasing concern about
possible toxic effects, especially in relation to CDDs.

3.3. Disposal of Wastes

Environmental pollution with 2,4-D may occur as a result of the
production and disposal of 2,4-D, or of its by-products, and of
industrial effluents. Such pollution will be generally localized to
the production site and to areas of waste dumping, and it is likely to
be more dispersed if disposal or leaching has occurred into water
courses. Disposal of unused 2,4-D in agriculture and washing of
equipment may result in localized land pollution and also pollution of
water supplies through direct contamination or leaching from soil.

4. UPTAKE, ACCUMULATION, ELIMINATION, AND BIODEGRADATION

Appraisal

2,4-D does not persist in soil because of its rapid degradation.

The physico-chemical properties of 2,4-D acid and its formulations
have an important effect on its behaviour in environmental
compartments.

The bioavailability to, and uptake by, aquatic and terrestrial
organisms is strongly influenced by the organic matter content of
soils, microbiological activity, and by environmental conditions such
as temperature and pH. Although highly inconsistent, the data on
dissipation and bioavailability in various soils demonstrate a marked
influence of differences in the texture and mineral composition of the
soil (Graham-Bryce, 1972). In aerobic soils, with a high content of
organic material, and at high pH values and temperatures, toxic effects
are limited because of rapid degradation of 2,4-D.

Uptake is followed by rapid excretion in most organisms. With the
exception of some algae, the retention of 2,4-D by organisms in the
environment cannot be expected, because of its rapid degradation.

Some microorganisms are capable of utilizing 2,4-D as their sole
carbon source. Repeated application to soil stimulates the number of
organisms capable of degrading the compound.

4.1. Biodegradation

2,4-D is readily and rapidly degraded in soil. Warm, moist
conditions and addition of organic matter stimulate degradation.
Autoclaving the soil and inhibiting bacterial metabolism reduce
degradation. The kinetics of 2,4-D disappearance suggest that
microorganisms are responsible. Particular species of microorganisms,
of various types, have been isolated and shown to degrade phenoxyacetic
acid herbicides in pure culture. Degradation of the phenoxyacetic
acids proceeds by two main pathways. These are via a hydroxyphenoxy
acetic acid intermediate or via the corresponding phenol. The
literature has been reviewed by the two workers principally responsible
for this evidence (Audus, 1960, 1964; Loos, 1969). Some microorganisms
are capable of using 2,4-D as their sole carbon source. More often,
2,4-D is co-metabolized with another carbon source. Regular treatment
of soil with 2,4-D stimulates the numbers of organisms which are
capable of degrading the compound. Treatment with other phenoxy
herbicides can also lead to an increase in organisms capable of
degrading 2,4-D.

Butler et al. (1975a) exposed 21 species of freshwater algae
isolated from natural lake water to 2,4-D butoxyethanol ester, at a
concentration of 0.01 mg/litre, and looked for degrading ability. Most
of the cultures fully degraded 2,4-D within 2 weeks. A single culture
retained 64% of the added 2,4-D, while seven isolates reduced 2,4-D to
less than 20% of the amount added. The remaining isolates showed 2,4-D
recoveries ranging from 22% to 53%.

Le Van To (1984) isolated six species of microorganisms
from soil previously treated with herbicides. These were
Flavobacterium peregrinum, Pseudomonas fluorescens, Arthrobacter
globiformis, Brevibacterium sp., Streptomyces viridochromogenes, and
an unidentified Streptomyces species. Flavobacterium was the most
active organism in degrading 2,4-D; degradation of 20 mg/kg of
2,4-D was complete after 20 to 30 days. In a liquid medium,
Flavobacterium degraded 93.5% of added 2,4-D within 80 h. The time
required to degrade half of the 2,4-D added to a sterilized soil along
with nutrient was estimated at 3 days. Li-Tse Ou (1984) investigated
the breakdown of 2,4-D in two types of soil under dry and moist
conditions and at two different temperatures. Numbers of
microorganisms degrading 2,4-D were also estimated. Generally, 2,4-D
disappeared more rapidly from moist soil; after 14 days of a slow rate
of disappearance, however, the removal rate from dry, sandy soil
increased. Numbers of organisms degrading 2,4-D were initially much
lower in sandy than in clay loams. However, numbers increased rapidly
in sandy soils after the addition of the herbicide and, as a result,
2,4-D was eventually degraded more rapidly in sandy than in clay loams.
In moist conditions, at 25 °C, the half-life of 2,4-D was 7 days or
less, whereas in dry conditions, at 35 °C, it could be as long as 250
days. These latter conditions are unlikely to apply in most natural
conditions where 2,4-D is likely to be used.

Rosenberg & Alexander (1980) incubated sewage-sludge bacteria with
2,4-D and found that nearly all of the herbicide had disappeared after
7 days. Subsequent additions of 2,4-D led to destruction of the
compound without a lag period; this suggests selection for organisms
capable of degrading the compound. Similar results were obtained using
bacteria from soil. The time needed for the disappearance of 90% of
the added 2,4-D was 14 days with soil inocula. 2,4-D added
subsequently was reduced by 70% within 3 to 4 days. Various tropical
soils were used in the experiment and all showed a high capacity for
degrading 2,4-D. Thompson et al. (1984) determined the persistence of
2,4-D applied at recommended rates in agricultural soils in Canada. In
all but one soil, a sandy loam, the concentration had declined by 50%
within 7 days. Sattar & Paasivirta (1980) showed slower degradation
of 2,4-D in acid soils. It took 6 weeks for 50% of the 2,4-D to
disappear from the soil and 7% was still left after 24 weeks. In
water-logged soil, there was reduced degradation of the herbicide.

Lewis et al. (1984) studied bacterial breakdown of 2,4-D
butoxyethyl ester and the effects of adding various extra components to
the medium. The addition of unfiltered, spent fungal medium from which
the majority of the fungus had settled out could be either stimulatory
or inhibitory to degradation rates of the herbicide; this depended on
the particular fungus species cultured in the medium. Further
investigation showed that effects were primarily due to differences in
pH. Reduction of the pH below 6 inhibited bacterial transformation of
the compound. Fungi commonly release large amount of organic acids.
The addition of spent fungal medium inhibited the breakdown of 2,4-D
ester. Buffering the added fungal medium reduced this inhibitory
effect; indeed, some stimulation of breakdown occurred after the
addition of buffered, spent medium. The addition of nutrients, or
other bacteria which did not transform 2,4-D, stimulated the
transformation of the herbicide. The authors consider that the most

likely explanation for this phenomenon is induction of other
transforming enzymes. With increasing substrate concentration, further
enzyme systems are induced in bacteria. The presence of other
organisms may stimulate the induction of these other enzymes at lower
substrate concentrations than would normally induce them. Increased
biomass of transforming bacteria in the presence of competing organisms
contributes to increased transformation rates. The nature of the
microbial community can, therefore, greatly change the ability of
degrading bacteria to transform 2,4-D and other xenobiotics.

O'Connor et al. (1981) found that 2,4-D applied at about 1.5 mg/kg
was readily degraded in soil. Adding extra carbon in the form of
dried, digested sewage sludge had a short-term effect in enhancing
degradation of the compound. Torstensson (1975) measured the half-life
of 2,4-D degradation in cultures of soil microorganisms at different
pH. In the pH range of 8.5 to 5.0, the half-life changed little,
ranging from 5 to 8 days. At pH 4.5, the half-life increased to 21
days and, at pH 4.0, increased further to 41 days.

Lieberman & Alexander (1981) added 2,4-D to inocula of municipal
sewage and monitored the biological oxygen depletion (BOD) as a measure
of degradation. The herbicide was added to carbon-depleted inocula
such that the 2,4-D represented the sole carbon source. Less than 5%
of the available oxygen was depleted, indicating poor biodegradation of
2,4-D because of low numbers of organisms capable of degrading the
herbicide as their sole carbon source. A separate study showed that
2,4-D was not toxic to microorganisms in sewage.

Fournier (1980) showed that, while 2,4-D treatment increased the
numbers of soil microorganisms capable of metabolizing 2,4-D as the
sole carbon source and those capable of co-metabolizing the herbicide,
this increase was dependent on the concentration of 2,4-D used. At
concentrations of 2,4-D between 5 and 50 mg/litre, there was a
significant increase in the numbers of organisms metabolizing 2,4-D,
and at 5 mg/litre there was a very pronounced increase in organisms co-
metabolizing the compound. At much higher (500 mg/litre) or much lower
(1.2 µg/litre) 2,4-D concentrations, there was no increase in the
numbers of either metabolizing or co-metabolizing organisms.

Sandmann & Loos (1984) estimated the numbers of microorganisms
capable of degrading 2,4-D in soils with and without the `rhizosphere
effect' of two plants, African clover (Trifolium africanum) and sugar
cane (Saccharum officinarum). The `rhizosphere effect' is a
phenomenon which occurs in close association with the roots of plants,
where material from the root or the metabolic activity of the root
tissue affects the surrounding soil. Particularly high, stimulated
populations were associated with sugar cane. A similar effect, but to
a lesser degree, was found with clover. In the three sugar cane soils
examined, and their corresponding controls, the numbers of organisms
were 46 400, 156 000, and 40 700 per g of soil, with rhizospheres, and
178, 1480, and 6170 per g of soil, without rhizospheres, respectively.
Seibert et al. (1982) failed to demonstrate a rhizosphere effect on
2,4-D degradation in glasshouse studies using soils with and without
maize roots.

Norris & Greiner (1967) investigated the degradation of 2,4-D in
forest leaf litter. Litter from either alder, ceanothus, vine maple,
bigleaf maple or Douglas fir showed comparable ability to degrade
2,4-D, the recovery of 2,4-D being between 60% and 70% after 15 days of
incubation. In a second series of experiments, different formulations
of 2,4-D were added to alder litter. About 50% of the free acid of
2,4-D was degraded within 15 days. Triethanolamine salt and two
commercial formulations (`solubilized acid' and isooctyl ester) were
degraded less than the pure acid. There was between 30% and 40%
degradation of these preparations over 15 days.

Nesbitt & Watson (1980) related the degradation rate of 2,4-D in
river water to the nutrient levels, sediment load, and dissolved
organic carbon content of the water. The addition of sediment or
inorganic nutrients increased the rate of 2,4-D degradation, whereas
the addition of organisms capable of degrading 2,4-D did not increase
the rate of breakdown of the herbicide. This finding indicated that
the limiting factor in breakdown of 2,4-D in river water was not
numbers of organisms but the nutrient status of the river. The authors
noted that in winter, when the river was in peak flow and the water
temperature below that for optimum microbial activity, appreciable
amounts of the herbicide would be washed into the estuary. An earlier
pilot study of seasonal changes in the capacity of river water in
Western Australia to degrade 2,4-D (Watson, 1977) indicated clear
seasonal differences in both river water concentrations of the
herbicide and the degrading capacity of river water. Several rivers
were studied and differences were related to the amount of
agricultural run-off, the sediment content of the water, river flow,
and temperature. Rivers receiving agricultural run-off degraded 2,4-D
better than those receiving run-off principally from forests. This was
presumed to be the result of the preconditioning of organisms to the
herbicide; the investigation corrected for nutrient content of the
water which had been previously shown to affect degradation.

Spain & Van Veld (1983) looked at the degrading ability of
microbial communities taken from sediment cores from freshwater,
estuarine, and marine sites. Some cores were pre-exposed to 2,4-D.
Cores from freshwater sites showed increased degradation of 2,4-D after
pre-exposure to the compound, whereas estuarine and marine cores did
not show this effect. The adaptation of freshwater cores was maximal
after 2 weeks and no longer detectable 6 weeks after pre-exposure.

4.2. Uptake and Accumulation by Organisms

Appraisal

Many studies on the accumulation of 2,4-D have used radioactively
labelled herbicide and have monitored uptake by simple counting of the
label. This fails to take into account that the label could have been
removed from the parent molecule by metabolic breakdown. Values for
uptake should, therefore, be treated as a maximum possible uptake
value for 2,4-D. Such data would not normally be considered
acceptable. However, the accumulation of 2,4-D is so low that these
data serve to illustrate that little of the herbicide is accumulated.

4.2.1. Laboratory studies

Eliasson (1973) sprayed leaves of 3-year-old aspen (Populus
tremens) with the butoxyethanol ester of 2,4-D at 0.5 kg acid
equivalent/litre. The plants were then kept in an open-sided
glasshouse and residues of 2,4-D were monitored. Most of the
herbicide remained in, or on, the sprayed leaves. The average residue
level was 2300 mg/kg fresh weight 1 day after spraying. This level had
fallen to 1300 mg/kg after 37 days and, by day 365, the average residue
level was 870 mg/kg. This was a very high application rate and
indicates that there is no foliar uptake of 2,4-D by plants.

Glynn et al. (1984) exposed coral Pocillopora damicornis to three
concentrations of 2,4-D sodium or amine salts at 0.1, 1.0, or
10.0 mg/litre. The maximum concentration of 2,4-D found in coral
tissue was 0.137 mg/kg after exposure to the amine salt at 10 mg/litre,
but residues were not related to the 2,4-D exposure concentration. The
highest bioconcentration factor (BCF) of 1.33 was found after exposure
to 0.1 mg/litre of the amine salt of 2,4-D, i.e., the coral contained
1.33 times the concentration of 2,4-D in water.

Metcalf & Sanborn (1975) introduced ¹⁴C-labelled 2,4-D into
model ecosystems consisting of an alga Oedogonium, an aquatic plant
Elodea, a snail Physa, and the mosquito fish Gambusia. Total
¹⁴C in the water was equivalent to 0.205 mg 2,4-D/litre. The highest
BCF was in the alga (26.8, based on measurement of radioactivity).
Analysis of all components of the ecosystem for 2,4-D, rather than the
radiolabel, revealed none of the parent compound. The BCF, therefore,
refers to breakdown products rather than 2,4-D itself. Gile (1983)
introduced ¹⁴C-labelled 2,4-D, as the butyl ester, into a simulated
ryegrass ecosystem. The system consisted of a sandy loam soil, annual
ryegrass, several invertebrates, and grey-tailed voles. Voles were
introduced 10 days after spraying 2,4-D as a foliar spray at the
equivalent of 1 kg/ha. The experiment was terminated after 1 month.
Plant material contained an average of 8.9 mg/kg; this was identified
as being mostly 2,5-dichloro-4-hydroxyphenoxyacetic acid. Residue
levels in animals (based on unidentified ¹⁴C residues) ranged from
0.31 mg/kg in snails to 5.28 mg/kg in pillbugs (isopods).

Freitag et al.(1982) measured the bioaccumulation of ¹⁴C-2,4-D in
an alga Chlorella fusca and a fish, the golden orfe. They measured a
24-h static BCF of 6 for the alga, and a 3-day static BCF of <10 for
the fish. This measurement was based on radioactivity and, therefore,
did not distinguish between the parent compound and its breakdown
products.

Schultz (1973) examined uptake and loss of ¹⁴C-2,4-D dimethyl-
amine salt by organs of three species of fish (channel catfish,
bluegill sunfish, and largemouth bass), exposed to 0.5, 1.0, or
2.0 mg/litre of 2,4-D acid equivalent. After exposure to the highest
concentration of 2,4-D dimethylamine salt, there was detectable
radioactivity in all organs examined. Bile showed the highest residues
of ¹⁴C in all three species after 1 week. For the remainder of the
exposure period of 12 weeks, there was an increase of radioactivity in
other organs and a decrease in the bile. At the end of the exposure
period, there was no clear pattern to residue levels of ¹⁴C in

different organs. These levels ranged from 5.04 mg/kg in bile to
35.5 mg/kg in posterior kidney for the channel catfish. For largemouth
bass, the range was from 1.32 mg/kg in muscle to 7.29 mg/kg in liver.
For the sunfish, the lowest residue was 24.75 mg/kg in bile and the
highest 322.7 mg/kg in the pyloric caeca of the gut. After 84 days
exposure to the dimethylamine salt at 2 mg/litre, levels of ¹⁴C in
the muscle of catfish, bass, and sunfish were equivalent to 0.953,
0.035, and 1.065 mg 2,4-D/kg, respectively. No analysis for 2,4-D
itself was carried out. A second study exposed the three fish species
for 2 weeks to ¹⁴C-2,4-D dimethylamine salt at 1 mg/litre and then
for a further 4 weeks to clean water. The disappearance of ¹⁴C was
measured. Loss of ¹⁴C was slow at first but by 4 weeks most tissues
had shown a decline in residues. Samples were analysed for 2,4-D but
none was detectable, suggesting that the ¹⁴C measured was in
breakdown products. The values for 2,4-D residues in this and other
studies using ¹⁴C-labelled material should, therefore, be regarded as
overestimates of retained 2,4-D. Uptake of ¹⁴C-2,4-D was examined at
two different temperatures, 17 °C and 25 °C. The highest residues of
¹⁴C detected in fish were equivalent to 0.122 mg 2,4-D/kg, but no
2,4-D could be found after analysis, except in bluegill sunfish after
14 days. Loss of 2,4-D did not, therefore, seem to change with
differing temperature over this range. A similar study, at two
different water pH values, showed significantly more ¹⁴C uptake in
all three species at the more acidic pH. Analysis of fish tissues for
2,4-D by gas-liquid chromatography showed non-detectable, or trace,
levels in most samples. Only in bluegill sunfish after 7 and 14 days
were residues measurable. These 2,4-D residues showed the opposite
trend to the ¹⁴C results; there was more 2,4-D in fish exposed at the
more alkaline pH. The authors suggest that metabolism of the herbicide
in the fish is suppressed at alkaline pH.

Sigmon (1979) exposed bluegill sunfish to 2,4-D butyl ethyl ester
(3 mg/litre) at three different temperatures, 20, 25, and 30 °C, and
measured the tissue content of 2,4-D after 8 days. None of the groups
differed from the controls, residues being <0.05 mg/kg.

Bluegill sunfish and channel catfish took up <0.5% of the
available ¹⁴C when exposed to ¹⁴C-2,4-D dimethylamine salt at 2
mg/litre (with 1 litre of water per fish) for 7 days (Sikka et al.,
1977). A maximum total ¹⁴C concentration in the fish was
reached after 24 h and did not change significantly over 14 days.
Bluegill sunfish attained a total body concentration of 0.9 mg/kg and
catfish 0.2 mg/kg at 24 h. These values were 2,4-D equivalents of
¹⁴C measured; the compound was not analyzed directly. When bluegill
sunfish were injected intraperitoneally with ¹⁴C-2,4-D dimethylamine
salt, at dose levels of 1 or 2.5 mg/kg body weight, they
excreted 90% of the dose within 6 h of treatment. In a similar
experiment, Stalling & Huckins (1978) exposed bluegill sunfish to
¹⁴C-2,4-D dimethylamine salt at 2 mg/litre and measured both
¹⁴C and 2,4-D in fish and water samples over the following 12 weeks.
Radioactivity was detected in tissues and increased over the
experimental period, but there was no measurable 2,4-D; the detection
limit of the method was 0.1 mg/kg. An in vivo intraperitoneal
injection of 110 µg of ¹⁴C-2,4-D was followed by rapid elimination.

Rodgers & Stalling (1972) measured uptake of ¹⁴C-2,4-D butoxy-
ethanol ester by three species of fish, which were exposed to either
0.3 or 1.0 mg/litre and sampled over the next 168 h. Some fish were
fed and some fasted. Radioactivity in a variety of tissues was
determined; the maximum levels were found within 3 h of exposure in fed
fish. After this, levels declined over the remaining sampling period,
and by the end of the experiment, residues were negligible. The one
exception was the gall bladder, which consistently contained more 2,4-D
than other tissues. Results were different for fasted fish. In almost
all organs of fasted fish, uptake of 2,4-D was slower than for fed
fish, although the levels reached were eventually two to five times
higher than in fed fish. Analysis of the residues showed that only the
liver ever contained the herbicide in the ester form. In all other
tissues, only the acid was present.

Shcherbakov & Poluboyarinova (1973) monitored the accumulation of
2,4-D in carp and Daphnia. The 2,4-D was added as the butyl ester
at concentrations ranging from 0.006 to 5 mg/litre; the recommended
usage rate for this ester leads to water concentrations of about
0.5 mg/litre. Analyses of fish tissues were made for both the ester
and the acid. The highest BCF for the ester, at 395, was found with
fish after a 7-day exposure to 0.5 mg/litre. Acid accumulation was
lower than that of the ester. The experiment lasted for 70 days. At
day 10 and after, only trace amounts of ester were found in fish.
Small amounts of 2,4-D acid were found at day 10, but only trace
amounts after day 70. Residues of 2,4-D ester in Daphnia varied from
23.9 to 518 mg/kg, according to the exposure concentration.

Two experiments have been carried out on the grey slug Derocerus
reticulatum by Haque & Ebing (1983) using ¹⁴C-labelled 2,4-D acid.
The first study, a contact experiment, exposed the slugs to 2,4-D in
contaminated soil at 1.1 mg/kg. The body content of 2,4-D in slugs
reached equilibrium (0.014 mg/kg) after 15 days; this represented a
BCF of 0.013 based on radioactivity. In the second experiment,
slugs were exposed via the food using carrot discs containing 1.1 mg/kg
slug body weight per day over 5 days. Residues of ¹⁴C in the slugs
increased during the feeding period, peaking at 5.5 mg/kg. During the
following 7 days, residues were monitored to investigate loss of
radioactive material. At the end of the experiment, on day 12,
residues were comparable to those at the end of the feeding period.
During the course of feeding 2,4-D-contaminated carrots, more than 80%
of the ingested dose of radioactivity was excreted rapidly; only 20%
was retained. There was no attempt to characterize the ¹⁴C residues;
these may, therefore, represent either 2,4-D or its breakdown
products.

Chickens given a single oral dose of 100, 200, or 300 mg/kg
body weight reached maximum plasma levels of 2,4-D of 90, 130, and
250 µg/ml, respectively. Plasma levels in all groups had fallen to
15 µg/ml or less after 24 h. Continuous dosing of chickens at
300 mg/kg per day led to a faster rate of elimination of the daily dose
of 2,4-D with time (Bjorklund & Erne, 1966).

4.2.2. Field studies

Cope et al. (1970) treated experimental ponds with 2,4-D propylene
glycol butyl ether ester to give water concentrations up to and
including 10 mg/litre. No detectable 2,4-D was found in fish exposed
to 1 mg/litre or less of the herbicide, but residues were found in
bluegill sunfish exposed to 5 or 10 mg/litre. The highest residue
(2 mg/kg) was found 1 day after application. Residues were still
detectable after 3 days but not subsequently. Vegetation
(Potamogeton nodosus) and bottom sediment contained residues of
50.0 and 3.0 mg/kg, respectively, 2 days after treatment with the 2,4-D
ester at 10 mg/litre. The herbicide was still detectable at 0.1 mg/kg
in sediment after 44 days but not thereafter. At 44 days after
treatment, there were residues in the plant of 1.2 mg/kg; this amount
declined to 0.1 mg/kg after 94 days.

Following the field application of 2,4-D butoxyethanol ester at
22.5 kg/ha, Whitney et al. (1973) measured residues of the herbicide in
fish, crustacea, and insect larvae over a 3-week period. The herbicide
had been applied to control eurasian water milfoil. Some 2,4-D was
taken up by these various species; the highest residue concentration
was 0.24 mg/kg in largemouth bass after 8 days. All residues in
organisms were below 0.1 mg/kg after 3 weeks. No 2,4-D could be
detected in water in 33 samples taken after treatment, the detection
limit being 0.10 mg/litre. The highest reported concentration of 2,4-D
in mud was 0.65 mg/kg, 10 days after treatment, but in most samples the
herbicide level in mud was much lower and in several it was
undetectable.

Hoeppel & Westerdahl (1983) treated four areas (10 ha each) of
dense water milfoil beds in Lake Seminole, Georgia, with either 2,4-D
dimethylamine salt or 2,4-D butoxyethanol ester, at each of two
application rates (22.5 or 45 kg/ha). Both formulations were converted
to 2,4-D free acid within 24 h. Maximum water concentrations achieved
in the high rate (45 kg/ha) areas were 3.6 and 0.68 mg/litre for the
dimethylamine salt and butoxyethanol ester, respectively. There was no
detectable uptake of 2,4-D into fish in those areas treated with the
dimethylamine salt. In the ester-treated areas, 4 out of 24 game fish
sampled contained low levels of 2,4-D in muscle (the highest residue
being 0.29 mg/kg) and 18 out of 20 gizzard shad contained detectable
2,4-D in muscle (the highest residue being 6.9 mg/kg). No fish sampled
more than 13 days after treatment contained detectable 2,4-D.

Schultz & Harman (1974) treated nine experimental ponds with 2,4-D
dimethylamine salt at three concentrations: 2.24, 4.48, and 8.96 kg/ha.
Samples of water, bottom sediment, and fish were taken over 147 days.
Maximum water and sediment concentrations of 2,4-D were 0.692 mg/litre
and 0.17 mg/kg, respectively. Of 307 fish sampled, 45 contained
detectable residues of 2,4-D. The highest residue measured was in a
channel catfish at 1.075 mg/kg 1 day after treatment. All residues in
fish after 28 days were less than 0.005 mg/kg; most were undetectable.

Smith & Isom (1967) measured uptake and retention of 2,4-D after
treatment of two field sites for control of watermilfoil with the
butoxyethanol ester. The first site was treated with a granular
formulation at a rate of 112 kg/ha. One bluegill sunfish (Lepomis
macrochirus) contained 0.15 mg 2,4-D/kg on day 50 after treatment. All

other fish, sampled between 72 h and 50 days after treatment,
contained less than 0.14 mg/kg, which was the limit of detection. Two
samples of several species of mussel, held in cages for 96 h following
spraying, showed residues of 0.38 and 0.7 mg/kg. Water levels of 2,4-D
reached a peak of 37 mg/litre within 1 h of application and had fallen
to less than 1 µg/litre within 8 h. Mud samples contained very
variable levels of 2,4-D residues, ranging between 0.14 and 58.8 mg/kg.
The highest residue was found 10 months after application. The second
site was treated at the lower rate of 45 kg/ha. All fish sampled
between 15 days and 9 months after 2,4-D application showed residues
of less than 0.14 mg/kg. Mussels sampled between 1 and 42 days after
application contained residues ranging between <0.14 and 1.12 mg/kg.
Water levels peaked at 157 µg/litre, 1 h after spraying, and mud
residues ranged from <0.14 to 33.6 mg/kg.

Coakley et al. (1964) measured residues in organisms at the center
of a 0.4-ha field plot sprayed with 2,4-D butoxyethanol ester at a rate
of 33.7 kg/ha for watermilfoil control. Two days after application,
oysters (Crassostrea virginica), clams (Mya arenaria), fish (Lepomis
gibbus), and blue crabs (Callinectes sapidus) contained 3.5, 3.7, 0.3,
and <0.8 mg/kg, respectively.

In 1971, over 2800 ha in Loxahatchee National Wildlife Refuge
were sprayed with the dodecyl-tetradecyl amine salts of 2,4-D at a
rate of 4.48 kg/ha. The initial application of 2,4-D was followed by
spot treatments of the same formulation and/or the dimethylamine salt
of 2,4-D. The highest water concentration (0.037 mg/litre of 2,4-D)
was measured 1 day after the initial application. Of 60 fish sampled
in the area, 19 had measurable residues of 2,4-D but only three of
these were greater than 0.1 mg/kg; the highest recorded residue was
0.162 mg/kg. Breast muscle and liver of a bird, the common Florida
gallinule Gallinula chloropus, had residues of 0.3 and 0.675 mg/kg,
respectively, 1 day after spraying. No residues were found in the bird
4 days after spraying (Schultz & Whitney, 1974).

Plumb et al. (1977) treated sprouting chamise (Adenostoma
fasciculatum) with the polyethylene glycol butyl ether ester of 2,4-D
at a rate of 3.4 kg acid equivalent/ha. A maximum concentration of
herbicide (95.2 mg/kg) was found in the plant within 15 min of
application. A residue of 3.8 mg 2,4-D/kg remained in, or on, the
plants (shoots which had been originally sprayed) 1 year after
treatment. When Radosevich & Winterlin (1977) applied the butoxypropyl
ester of 2,4-D to a chaparral area at a rate of 4.5 kg/ha, the
residues measured in chamise were 221 mg/kg and in grass and forbs
269 mg/kg within 2 h of application. After 30 days, these levels had
dropped to 60 mg/kg for chamise and 21 mg/kg for grass and forbs, and,
after 360 days, 0.1 mg/kg was present in chamise. Siltanen et al.
(1981) monitored residues of 2,4-D in the fruit of bilberries 1 year
after the application of 0.25, 0.75, or 2.25 kg/ha acid equivalent. No
residues were detected, the limit of detection being 0.05 mg/kg.

Raatikainen et al. (1979), in a controlled field experiment,
sprayed cowberry and bilberry with an ester formulation of 2,4-D.
Three application rates were used, 0.25, 0.75, and 2.25 kg acid
equivalent/ha, and residues of 2,4-D were measured approximately
1 month after application. Thirty-four days after the application of

0.25 kg/ha, residues in cowberry were 0.3 mg/kg. Cowberries exposed to
0.75 or 2.25 kg/ha were analysed after 35 days and contained residues
of 1.0 and 3.7 mg/kg, respectively. Bilberries treated with 0.25,
0.75, or 2.25 kg/ha were analysed 29 days later; residues were 0.1,
1.3, and 4.8 mg/kg, respectively.

4.3 Elimination

James (1979) studied tissue distribution of ¹⁴C-labelled 2,4-D in
the spiny lobster (Panulirus argus). Labelled herbicide was injected
into the pericardial sinus and animals were sacrificed at regular
intervals. 2,4-D was taken up from the haemolymph, by the green gland,
and excreted unchanged, with an overall half-time of about 8 h. Tuey &
James (1980), in a similar study, found that the clearance of 2,4-D
from haemolymph, via the green gland, was three to five times greater
than the rate of metabolism in the hepatopancreas.

Pritchard & James (1979) studied the renal handling of intra-
venously injected 2,4-D by the winter flounder (Pseudopleuronectes
americanus). 2,4-D, at a concentration of 1 µmol/litre of plasma,
was actively secreted into the glomerular filtrate of the kidney
with clearances of nearly 500 times the glomerular filtration rate.
At higher plasma concentrations of between 10 and 60 µmol/litre, a
transport maximum of 0.85 µmol/g of kidney per h was observed.
Koschier & Pritchard (1980) reported a similar study using
an elasmobranch fish Squalus acanthias. They administered
2.5 µmol ¹⁴C-2,4-D/kg to the fish intramuscularly and monitored
blood and urine ¹⁴C levels. Clearance of total 2,4-D was more than
25 times greater than the glomerular filtration rate, indicating that
2,4-D was being actively secreted by the kidney. 2,4-D was eliminated
in the urine as a taurine conjugate, this representing about 95% of the
excretory products. The plasma contained, primarily, unconjugated
2,4-D (>90%). It seemed, therefore, that 2,4-D was conjugated with
taurine before being excreted in the urine. Guarino et al. (1977), in
a similar study on the dogfish Squalus, also found that 2,4-D was
extensively conjugated to taurine (>90%) and was eliminated
predominantly via the urine; 70% of the administered dose appeared in
the urine within 4 to 6 days. The highest tissue concentration of
2,4-D (14.5 mg/kg) was found in the kidney after 4 h. Plasma
elimination was rapid, with a half-time of 44 min; similarly rapid
clearance was seen from the kidney. Half-time estimates for muscle and
liver were 2 to 3 days and 5 days, respectively.

5. TOXICITY TO MICROORGANISMS

Appraisal

In general 2,4-D is relatively non-toxic to water and soil
microorganisms at recommended field application rates.

No effect of 2,4-D was recorded on 17 genera of freshwater and two
genera of marine algae at concentrations up to 222 mg/litre.

No effect of 2,4-D was observed on respiration of either sandy loam
or clay loam soils at concentrations up to 200 mg/kg.

In the range of 25.2 to 50.4 mg/litre, 2,4-D was inhibitory to all
types of soil fungi.

Cell division was reduced in a green alga by 2,4-D at 20 mg/litre
and stopped at 50 mg/litre. No effect was observed on a natural
phytoplankton community after exposure to 2,4-D at 1 mg/litre.
However, exposure to esters of 2,4-D reduced productivity in these
organisms.

5.1. Aquatic Microorganisms

Hawxby et al. (1977) exposed cultures of three algae
(Chlorella pyrenoidosa, Chlorococcum sp., and Lyngbya sp.,) and
one cyanobacterium (blue-green alga) (Anabaena variabilis) to
concentrations of 2,4-D in the medium of up to 10 µmol /litre
(= 2.21 mg/litre). There was no effect on growth, respiration, or
photosynthetic rate.

Gangawane et al. (1980) studied the effects of 2,4-D on growth and
heterocyst formation in the nitrogen-fixing cyanobacterium (blue-green
alga) Nostoc. The organism was cultured for 30 days in 0, 10, 100,
1000, or 1500 mg 2,4-D/litre. Growth was measured by optical density
and cells forming heterocysts were counted. Growth was inhibited at
both 10 and 100 mg 2,4-D/litre and was eliminated at higher
concentrations. There was also reduced heterocyst formation.

Lembi & Coleridge (1975) demonstrated a marked effect of 2,4-D,
at concentrations of 110 or 220 mg/litre, on cultures of the
green algae Scenedesmus, Ankistrodesmus, and Pediastrum. After 14
days of culture, the three species under control conditions produced
456 x 10², 634 x 10⁴, and 227 cells or colonies per ml of medium,
respectively. Corresponding figures after exposure to 110 mg/litre
were 54 x 10², 41 x 10⁴, and 74 cells or colonies per ml,
respectively. For both Scenedesmus and Ankistrodesmus, these values
were less than the pre-treatment cell concentrations.

Butler et al. (1975b) exposed unialgal cultures of green algae
isolated from Warrior River water to 2,4-D butoxyethanol ester at
0.001, 0.01, 0.1, 1.0, or 4.0 mg/litre. Thirty separate isolates were
used. Concentrations less than or equal to 1 mg/litre of the 2,4-D
ester did not change the growth pattern of the isolates. However, with
a concentration of 4 mg/litre, there was some inhibition of growth, as
indicated by a 10% increase in the number of incubates which showed
poor growth, or no growth, when compared to controls. Some isolates
were unaffected even at this concentration and it can therefore be
assumed that 2,4-D butoxyethanol ester might change the species
composition of green algae populations.

Bednarz (1981) used 12 pure cultures of green algae and
cyanobacteria (blue-green algae) separately and in combination to
investigate the effects of 2,4-D acid. Cultures were exposed to
concentrations of 2,4-D ranging from 0.001 to 10 mg/litre. Low
concentrations of 2,4-D stimulated the growth of most species of algae,
whereas high concentrations inhibited growth. Chlorococcal green algae
were more sensitive to 2,4-D than were filamentous green algae or
cyanobacteria. In further experiments, the authors cultured
combinations of sensitive and tolerant species in the same range of
2,4-D concentrations. Tolerant species used in combinations were
Chlorella pyrenoidosa, Dictyosphaerium pulchellum, and Scenedesmus
quadricaudata. The first two of these tolerant species reduced the
toxicity of 2,4-D to sensitive species in mixed culture. This
protective effect was not seen with Scenedesmus.

Singh (1974) cultured a filamentous, nitrogen-fixing,
cyanobacterium Cylindrospermum sp. in concentrations of 2,4-D acid of
0, 100, 300, 400, 500, 600, 800, 1000, or 1200 mg/litre and examined
growth and heterocyst formation after 8 days. Both parameters were
affected at concentrations higher than 300 mg/litre and cultures were
killed at a concentration of 1000 mg/litre. Kapoor & Sharma (1980)
exposed cultures of the nitrogen-fixing, filamentous cyanobacterium
Anabaena doliolum to 2,4-D ethyl ester (as `Weedone 48' concentrate)
at concentrations of 36, 108, 180, 252, or 324 mg/litre. There was a
dose-related decrease in cell nitrogen over the whole range of 2,4-D
ester exposures. Cell growth was stimulated by lower concentrations of
2,4-D and only inhibited by the highest dose. Tiwari et al. (1984)
exposed cultures of a similar nitrogen-fixing, filamentous
cyanobacterium (Anabaena cylindrica) to 2,4-D acid at concentrations of
0, 100, 500, 700, 1000, or 1500 mg/litre, and examined growth,
heterocyst formation, and nitrogen fixation. For all these parameters,
there was a stimulatory effect of 2,4-D at 100 mg/litre and a
progressive inhibition with higher concentrations. These and similar
algae are considered to be a major source of nitrogen in tropical
ponds and soils. Das & Singh (1977) cultured the nitrogen-fixing
cyanobacterium Anaebaenopsis raciborskii in concentrations of 2,4-D
acid (sodium salt) of 10, 100, 400, 600, 800, and 1000 mg/litre and
measured nitrogen fixation. Control cultures and those exposed at 10
and 100 mg 2,4-D/litre showed no significant differences.
Nitrogen-fixation was inhibited at 400 mg/litre or more and eliminated
at 600 mg/litre.

Butler (1963) reported no effect on a natural phytoplankton
community of exposure to a 1 mg/litre concentration of 2,4-D (as the
acid or dimethylamine salt), or of the dimethylamine salt on pure
cultures of Dunaliella euchlora or Platymonas over 4 h. In a later
study (Butler, 1965), natural phytoplankton communities were exposed to
esters of 2,4-D. Butoxyethanol ester, propylene glycol butyl ether
ester, and ethylhexyl ester reduced productivity (as measured by
carbon fixation) by 16%, 44%, and 49%, respectively, at a concentration
of 1 mg/litre.

Sarma & Tripathi (1980) monitored cell division in the filamentous
green alga Oedogonium acmandrium exposed to 2,4-D acid at 1, 5, 10,
20, or 50 mg/litre of culture medium. At up to 10 mg/litre, 2,4-D
was found to stimulate cell division; a 168 h exposure to 5 mg/litre
increased the incidence of dividing cells by 15% over controls.
However, cell division was reduced at 20 mg/litre and stopped at
50 mg/litre. Abnormalities in chromosomes during cell division
increased with increasing 2,4-D exposure.

Chai & Chung (1975) examined the effects on growth,
photosynthesis, respiration, and chemical composition of exposing
cultures of the green alga Chlorella ellipsoidea to 2,4-D acid at 22 or
88 mg/litre. At 22 mg/litre, 2,4-D increased growth, photosynthesis,
and the cell content of protein and nucleic acids. Carbohydrate
content was unchanged. However, at 88 mg/litre, growth was inhibited,
photosynthesis was no different from controls, and the cell content of
carbohydrate, protein, and nucleic acids was decreased.

Elder et al. (1970) examined the effect of 2,4-D acid on 17 genera
of freshwater and two genera of marine algae exposed at 22, 111, or
222 mg/litre. There was no effect on the growth of any of the
cultures, even at the highest dose of 2,4-D.

Cultures of the flagellate Euglena gracilis were exposed for 24 h
to concentrations of 1, 5, 10, 50, or 100 mg/litre or for 7 days to 10,
50, or 100 mg/litre of 2,4-D acid by Poorman (1973). Cultures in 50
and 100 mg 2,4-D/litre yielded 84% and 74%, respectively, relative to
controls, over 24 h. Lower concentrations of 2,4-D had a slightly
stimulatory effect. After 7 days, there was significant stimulation of
yield with 10 mg/litre; the culture yielded 161% compared to a control.
There was slight stimulation of growth by 50 mg/litre and a reduction
to 78% of control levels with 100 mg/litre.

George et al. (1982) exposed the rotifer Brachionus
calyciflorus to 2,4-D at 5 mg/litre. Median lethal time (LT₅₀) was
24 h and LT₁₀₀ was 31 h.

5.2. Soil Microorganisms

Pachpande & David (1980) isolated the soil alga Chlorococcum
infusionum from paddy fields and cultured the organism in the presence
of 2,4-D acid at concentrations of 0, 1, 2, 3, 4, and 5 mg/litre.
Growth was estimated as dry weight of algal cells filtered out of the
medium. All concentrations of 2,4-D were inhibitory to growth.
At the highest 2,4-D concentration of 5 mg/litre, the culture yield
was reduced from a control level of 720 mg dry wt/litre of medium to
520 mg/litre.

Cullimore & McCann (1977) applied 2,4-D acid to isolated cores
taken from a prairie, loam soil to give approximate concentrations of
1 or 100 mg/kg in the top 2 cm of soil. Soil algal populations were
estimated from subsamples of cores taken before treatment and 1, 5, or
20 days after treatment with herbicide. Thirty-one genera of algae
were identified, of which five were very sensitive to 2,4-D and were
rarely found after treatment. These were Chlamydomonas, Chlorococcum,
Hormidium, Palmella, and Ulothrix. The most resistant genera were
Chlorella, Lyngbya, Nostoc, and Hantzschia; the `percent sensitivity'
of these genera (% of the total number of treatments in which the
genus was absent) was 28%, 6%, 22%, and 44%, respectively. The
reduction in cell numbers of algae in the top layer of the soil after
herbicide treatment was soon offset by an increase in the population of
Chlorella, Stichococcus, Oscillatoria, and Spongiochloris, all of
which recovered very rapidly from the herbicide effects. There was,
however, an overall reduction in cell numbers of nitrogen-fixing algae.

Mukhopadhyay (1980) measured the bacterial, fungal, and
actinomycete populations of soils supporting rice or maize plants which
had been treated with various herbicides for weed control. There was
no effect of 2,4-D, applied at the recommended rate, either on soil
microorganism numbers or on the evolution of carbon dioxide by soil
cultures.

Huber et al. (1980) examined the effect of 2,4-D at 0.3, 0.2, or
0.1 mmol/litre (= 66, 44, and 22 mg/litre, respectively) on seven
cultures of soil microorganisms. There was no effect on the growth of
five of the cultures; these were Nocardia sp., Pseudomonas fluorescens
in both aerobic and anaerobic culture, Bacillus subtilis, and
Ustilago maydis. There was a small reduction in growth at the
highest 2,4-D dose in cultures of Rhizopus japonicus and Aspergillus
niger. 2,4-D had no effect on mycelium growth of three out of four
plant pathogenic fungi in culture; Phytophthora cryptogea showed
reduced mycelial growth at 0.1, 0.2, and 0.3 mmol 2,4-D/litre,
but Fusarium oxysporum, Alternia radicina, and Rhizoctonia solani
were unaffected.

Moubasher et al. (1981) added 2,4-D at three doses (1.9, 7.6, and
15.2 mg/kg) either to soil or to agar medium inoculated with soil
fungi, and the effects on fungal populations were monitored. In soil,
at all three doses, 2,4-D stimulated the fungi. When incorporated in
the agar medium, 2,4-D was stimulatory to overall fungal growth and to
four individual species of fungus at the lowest dose of 6.3 mg/litre,
but inhibitory to two other species. At higher doses of 25.2 or
50.4 mg/litre, the herbicide was inhibitory to all fungi.

2,4-D had a significant inhibitory effect on culture yields of the
bacterium Escherichia coli only at 10-³mol/litre (= 220 mg/litre).
There was no effect at 10-⁴mol/litre (= 22 mg/litre) (Toure & Stenz,
1977).

Prescot & Olson (1972) added 2,4-D, at doses of 0, 0.1, 1.0, 10, or
100 mg/litre, to cultures of the soil amoeba Acanthamoeba castellanii
and monitored growth and reproduction. There was a stimulatory effect
of 2,4-D at all dose levels; this effect was most marked at the lowest
dose and declined with increasing exposure to 2,4-D. The authors

suggest that the amoeba may degrade the 2,4-D and utilize it as a
carbon source. However, Pons & Pussard (1980) found no effect of 2,4-D
(at 28, 54, or 84 mg/litre) on the reproduction of 23 different strains
of free-living soil amoebae.

2,4-D, at 10^-3mol/litre in cultures of the ascomycete Neurospora
crassa, stimulated DNA synthesis but had no effect at lower
concentrations of 10^-4 to 10^-6mol/litre. These concentrations had
no significant effect on either RNA or protein (Schroder et al.,
1970).

Naguib et al. (1980) measured growth, respiration, and absorption
and utilization of sugar and nitrogen in pre-formed fungal mats of
Aspergillus terreus over 72 h in the presence of 200 mg/litre of
2,4-D. The herbicide inhibited sugar inversion and consequently sugar
absorption. It also reduced the incorporation of nitrogen into protein.
Respiration was depressed. Growth of the fungus was suppressed and, on
a dry weight basis, culture mass was reduced to below the initial
level.

Trevors & Starodub (1983) added 2,4-D to sandy loam and clay
loam soils and measured both respiration and electron transport
system (ETS) activity. ETS was assessed by measuring the capacity
of the soil to reduce 2-( p -iodophenyl)-3-( p -nitrophenyl)-5-phenyl
tetrazolium chloride (INT) to iodonitrotetrazolium formazan (INT
formazan). The effects of 2,4-D were tested at concentrations of the
herbicide in soil of 0, 10, 25, 50, 75, 100, or 200 mg/kg. There was
no effect on soil respiration, monitored either as oxygen consumption
or carbon dioxide evolution, at any of the concentrations of 2,4-D in
either soil. There was similarly no effect on ETS in the sandy loam.
However, in the clay loam, there was a progressive inhibition of ETS
over the whole range of concentrations of the herbicide. The control
soil had an ETS activity of 37.3 µg INT formazan production/g soil,
whereas the ETS activity of soil treated with 10 mg 2,4-D/kg was
25.1 µg INT formazan/g, significantly lower than that of the
control. The activity was reduced further with increasing
concentrations of 2,4-D, until an activity of 16.3 µg INT formazan/g
was found at 200 mg 2,4-D/kg.

Deshmukh & Shrikhande (1975) added 2,4-D, at recommended field
rates, and at five times the recommended field rates, to two
types of soil from India. Both doses of 2,4-D inhibited numbers
of Azobacter in both soil types, and the high, but not the low, dose of
2,4-D reduced nitrogen fixation in both soils. The same authors
(Deshmukh & Shrikhande, 1974) monitored the populations of various
microorganisms under the same dosing conditions. 2,4-D stimulated
the numbers of actinomycetes throughout the 6-week incubation period at
both dose levels. Fungal populations were reduced in the first week of
incubation at both dose levels in sandy loam, but only at the higher
dose level in clay loam. This reduction in fungal populations
persisted until the second week with the high dose in the sandy soil
and throughout the incubation period with the high dose in clay soil.
There was a temporary (1 week) reduction in total bacterial numbers
with both 2,4-D dose levels in the sandy soil and with the higher level
in clay soil. Schroder & Pilz (1983) reported that 2,4-D at
approximately 10^-4mol/litre (= 22 mg/kg) had no long-term effect on
soil nitrification.

Welp & Brummer (1985) measured the influence of 2,4-D on the
reducing capacity of soil microorganisms, reduction being monitored as
the capacity to reduce Fe(III) oxides to soluble Fe(II) ions. They
determined no-observed-effect levels (NOEL) of 115 and 95 mg 2,4-D/kg
for two different soil types and corresponding EC₅₀ values on
reduction capacity of 200 and 530 mg 2,4-D/kg soil.

Ruggiero & Radogna (1985) extracted and partially purified soil
diphenolase (laccase) from forest soil. This enzyme, which exists free
in the soil, plays an important role in the metabolism of humic
materials in soil. Oxygen consumption was monitored during the
enzymatic reaction, using either catechol or p -phenylenediamine as
substrate, and the effect of 2,4-D was investigated. The herbicide
inhibited diphenolase activity, and Lineweaver-Burk plots of the
data suggested that 2,4-D acts as a non-competitive inhibitor.
Apparent K values of 28.7 and 6.0 mol/litre were obtained for catechol
and p -phenylenediamine, respectively.

6. TOXICITY TO AQUATIC ORGANISMS

6.1. Toxicity to Aquatic Invertebrates

Appraisal

6.1.1. Short-term toxicity

The short-term toxicity of 2,4-D to aquatic invertebrates is
summarized in tables 3 - 5.

Unfortunately, there are few studies where both the free acid (or
its salts) and ester preparations have been tested on the same organism
under the same conditions. The only organisms for which this applies
are the oyster (Butler, 1963; Butler, 1965), the stonefly (Sanders &
Cope, 1968), and daphnids and shrimp (Sanders, 1970a). These studies
all show that the free acid and its salts are less toxic than ester
formulations; for example the free acid is at least 20 times less toxic
to the water flea Daphnia magna than the least toxic of the esters
tested (Sanders, 1970a). Comparing studies carried out by different
authors and in different systems also suggests a much greater toxicity
of the ester preparations.

Liu & Lee (1975) found that 2,4-D could adversely affect the bay
mussel (Mytilus edulis) at all stages of its life cycle. The attachment
of young mussels to test chamber walls was reduced (data in Table 3).
The authors evaluated, in two duplicate experiments, the effects of
2,4-D acid, at concentrations in sea water of 22.8, 45.7, 91.4, and
182.8 mg/litre, on the growth of larval mussels. After 10 days
exposure, there was a significant reduction in the growth of larvae
exposed to 91.4 mg 2,4-D/litre; larvae were 11.6% smaller than
controls. This reduction was found in only one experimental replicate.
In both experiments, there was reduced growth after 10 days exposure
to 182.8 mg/litre; larvae were 31.9% and 34.9% smaller than controls
in the two experiments. Exposure for 20 days at 91.4 mg/litre
led to reduced growth in both experiments. All larvae exposed to
182.8 mg/litre died within 12 days, but only in one experimental
replicate. Extension of the growth study in the second experiment led
to all larvae dying within 22 days of exposure to 182.8 mg/litre and,
therefore, failing to undergo metamorphosis. The metamorphosis of
larvae exposed from age 30 to 70 days was not affected by 2,4-D at
concentrations up to 176 mg/litre.

Presing (1981) monitored reproduction over four broods in the water
flea Daphnia magna exposed to 0, 5, 10, 25, or 50 mg/litre of
`Dikonirt' (sodium salt of 2,4-D). For the first brood, the only
significant effect was at 50 mg/litre, whereas the fourth brood was
delayed even at 5 or 10 mg/litre. Significant reductions in the
average number of young produced for each female were found with
the two highest concentrations. Young kept until maturity from
each of the tests were themselves exposed to 2,4-D in a repeat
experiment. Again there was a significant effect on young produced at
25 and 50 mg/litre.

Table 3. Toxicity of 2,4-D to estuarine or marine invertebrates
---------------------------------------------------------------------------------------------------------
Organism Flow/ Temp Salinity pH Formulation^c Parameter Water Reference
stat^a (°C) (^o/oo) concentration
(mg/litre)
---------------------------------------------------------------------------------------------------------
Bay mussel 17.2- 22.9- 6.4- free acid 96-h LC₅₀ 259 Liu &
(Mytilus edulis) 18.6 24.5 7.8 (232-289) Lee (1975)
17.2- 22.9- 6.4- free acid 96-h EC₅₀ 262 Liu &
18.6 24.5 7.8 attachment Lee (1975)
(trocophore larva) 17.2- 22.9- 6.4- free acid 48-h EC₅₀ 211.7 Liu &
18.6 24.5 7.8 normal Lee (1975)
development

Eastern oyster flow 18 29 butoxyethanol 96-h EC₅₀ 3.75 Butler
(Crassostrea virginica) shell growth (1963)
flow 29 25 isooctyl 96-h EC₅₀ 1.0 Mayer
shell growth (1987)
flow 28 25 PGBEE 96-h EC₅₀ 0.055 Mayer
shell growth (1987)

Copepod 21 7 7.8 butoxyethanol 96-h LC₅₀ 3.1 Linden
(Nitocra spinipes) (2.4-4.1) et al.
(1979)
Brown shrimp (adult) flow 30 PGBEE 24-h EC₅₀ 0.55 Butler
(Penaeus aztecus) loss of (1963)
equilibrium
(adult) flow 30 PGBEE 48-h EC₅₀ 0.55 Butler
loss of (1963)
equilibrium
(juv.)^b stat 26 30 butoxyethanol 48-h LC₅₀ 5.6 Mayer
(1987)
(adult) flow 29 26 isooctyl 48-h LC₅₀ 0.48 Mayer
(1987)
Dungeness crab (1st zoel) stat 13 25 acid (tech) 96-h LC₅₀ > 10 Caldwell
(Cancer magister) (1977)
(1st instar juv.)^b stat 13 25 acid (tech) 96-h LC₅₀ > 100 Caldwell
(1977)
Blue crab (juv.)^b stat 24 29 PGBEE 48-h LC₅₀ 2.8 Mayer
(Callinectes sapidus) (1987)
---------------------------------------------------------------------------------------------------------
^a Stat = static conditions (water unchanged for duration of test); flow = flow-through conditions
(2,4-D concentration
in water continuously maintained.
^b juv. = juvenile.
^c PGBEE = propylene glycol butyl ethyl ester.

Table 4. Toxicity of 2,4-D to freshwater invertebrates
---------------------------------------------------------------------------------------------------------
Organism Flow/ Temp Alkali- Hard- pH Formulation^d Parameter Water Reference
stat^a (°C) nity^b ness^b concentration
(mg/litre)
---------------------------------------------------------------------------------------------------------
Oligochaete worm flow 20 30 30 7.8 free acid 48-h LC₅₀ 122.2 Bailey &
(Lumbriculus flow 20 30 30 7.8 free acid 96-h LC₅₀ 122.2 Liu (1980)
variegatus)

Water flea stat 21 260 272 7.4 PGBEE 48-h LC₅₀ 0.1 Sanders (1970a)
(Daphnia magna) stat 21 260 272 7.4 dimethylamine 48-h LC₅₀ 4.0 Sanders (1970a)
stat 17 39 7.2 dimethylamine 48-h LC₅₀ > 100.0 Mayer &
Ellersieck(1986)
stat 21 260 272 7.4 butoxyethanol 48-h LC₅₀ 5.6 Sanders (1970a)
stat 21 260 272 7.4 free acid 48-h LC₅₀ > 100.0 Sanders (1970a)
20 8.4- free acid 96-h LC₅₀ 417.8 Presing (1981)
8.6
20 8.4- sodium salt 96-h LC₅₀ 932.1 Presing (1981)
8.6

Water flea 15.6 44 7.4 PGBEE 48-h LC₅₀ 4.9 Sanders &
(Simocephalus (4.0-6.7) Cope (1966)
serrulatus)

Water flea 15.6 PGBEE 48-h LC₅₀ 3.2 Sanders &
(Daphnia pulex) (2.4-4.3) Cope (1966)

Copepod (nauplius larva)
(Cyclops vernalis) stat 20 31.6 70 6.7 free acid 96-h LC₅₀ 8.72 Robertson (1975)
(5.34-11.57)
stat 20 31.6 70 6.7 alkanolamine 96-h LC₅₀ 54.8 Robertson (1975)
(46.45-64.6)

Scud stat 21.1 30 7.1 butoxyethanol 24-h LC₅₀ 1.4 (1.1-1.8) Sanders (1969)
(Gammarus stat 21.1 30 7.1 butoxyethanol 48-h LC₅₀ 0.76 (0.51
lacustris) -1.1) Sanders (1969)
stat 21.1 30 7.1 butoxyethanol 96-h LC₅₀ 0.44 (0.31
-0.62) Sanders (1969)
stat 21.1 30 7.1 PGBEE 24-h LC₅₀ 2.1 (1.7-2.5) Sanders (1969)
stat 21.1 30 7.1 PGBEE 48-h LC₅₀ 1.8 (1.4-2.3) Sanders (1969)
stat 21.1 30 7.1 PGBEE 96-h LC₅₀ 1.6 (1.2-2.1) Sanders (1969)
stat 21.1 30 7.1 isooctyl 24-h LC₅₀ 6.8 (4.8-9.7) Sanders (1969)
stat 21.1 30 7.1 isooctyl 48-h LC₅₀ 4.6 (2.9-7.3) Sanders (1969)
stat 21.1 30 7.1 isooctyl 96-h LC₅₀ 2.4 (1.9-4.8) Sanders (1969)
stat 15.5 260 272 7.4 PGBEE 24-h LC₅₀ 4.1 (2.8-5.8) Sanders (1970a)
stat 15.5 260 272 7.4 PGBEE 48-h LC₅₀ 2.6 (1.7-3.9) Sanders (1970a)
---------------------------------------------------------------------------------------------------------

Table 4. (contd.)
---------------------------------------------------------------------------------------------------------
Organism Flow/ Temp Alkali- Hard- pH Formulation^d Parameter Water Reference
stat^a (°C) nity^b ness^b concentration
(mg/litre)
---------------------------------------------------------------------------------------------------------
Scud stat 15.5 260 272 7.4 PGBEE 96-h LC₅₀ 2.5 (1.7-3.7) Sanders (1970a)
(Gammarus stat 15.5 260 272 7.4 butoxyethanol 24-h LC₅₀ 6.5 (1.0-8.6) Sanders (1970a)
lacustris) (contd.) stat 15.5 260 272 7.4 butoxyethanol 48-h LC₅₀ 5.9 (3.1-11) Sanders (1970a)
stat 15.5 260 272 7.4 butoxyethanol 96-h LC₅₀ 5.9 (3.1-11) Sanders (1970a)

Scud stat 15 272 7.4 dimethylamine 24-h LC₅₀ > 100 Mayer &
(Gammarus fasciatus) stat 15 272 7.4 dimethylamine 96-h LC₅₀ > 100 Ellersieck (1986)

Glass shrimp stat 21 260 272 7.4 PGBEE 48-h LC₅₀ 2.7 Sanders (1970a)
(Palaemonetes stat 21 260 272 7.4 dimethylamine 48-h LC₅₀ > 100 Sanders (1970a)
kadiakensis stat 21 260 272 7.4 butoxyethanol 48-h LC₅₀ 1.4 Sanders (1970a)

Seed shrimp stat 21 260 272 7.4 PGBEE 48-h LC₅₀ 0.32 Sanders (1970a)
(Cypridopsis vidua) stat 21 260 272 7.4 dimethylamine 48-h LC₅₀ 8.0 Sanders (1970a)
stat 21 260 272 7.4 butoxyethanol 48-h LC₅₀ 1.8 Sanders (1970a)

Freshwater prawn stat 27 113.9 7.5 sodium salt 24-h LC₅₀ 2342 Shukla &
(Macrobranchium stat 27 113.9 7.5 sodium salt 48-h LC₅₀ 2309 Omkar (1983)
lamarrei) stat 27 113.9 7.5 sodium salt 72-h LC₅₀ 2267 Shukla &
stat 27 113.9 7.5 sodium salt 96-h LC₅₀ 2224 Omkar (1983)

Freshwater prawn stat 28 112.7 7.5 sodium salt 24-h LC₅₀ 2644 Omkar &
(Macrobranchium stat 28 112.7 7.5 sodium salt 48-h LC₅₀ 2536 Shukla (1984)
naso) stat 28 112.7 7.5 sodium salt 72-h LC₅₀ 2435 Omkar &
stat 28 112.7 7.5 sodium salt 96-h LC₅₀ 2397 Shukla (1984)

Freshwater prawn stat 28 112.7 7.5 sodium salt 24-h LC₅₀ 2474 Omkar &
(Macrobranchium stat 28 112.7 7.5 sodium salt 48-h LC₅₀ 2381 Shukla (1984)
dayanum) stat 28 112.7 7.5 sodium salt 72-h LC₅₀ 2333 Omkar &
stat 28 112.7 7.5 sodium salt 96-h LC₅₀ 2275 Shukla (1984)

Crayfish stat 15.5 260 272 7.4 PGBEE 48-h LC₅₀ > 100 Sanders (1970a)
(Orconectes nais) stat 15.5 260 272 7.4 dimethylamine 48-h LC₅₀ > 100 Sanders (1970a)
stat 15.5 260 272 7.4 butoxyethanol 48-h LC₅₀ > 100 Sanders (1970a)

Red swamp stat 20 100 8.4 alkanolamine 96-h LC₅₀ 1389 Cheah et al.
crayfish (imm.)^c (1174-1681) (1980)
(Procambarus clarki)

---------------------------------------------------------------------------------------------------------

Table 4. (contd.)
---------------------------------------------------------------------------------------------------------
Organism Flow/ Temp Alkali- Hard- pH Formulation^d Parameter Water Reference
stat^a (°C) nity^b ness^b concentration
(mg/litre)
---------------------------------------------------------------------------------------------------------
Sowbug stat 15.5 260 272 7.4 PGBEE 48-h LC₅₀ 2.2 Sanders (1970a)
(Asellus stat 15.5 260 272 7.4 dimethylamine 48-h LC₅₀ > 100 Sanders (1970a)
brevicaudus) stat 15.5 260 272 7.4 butoxyethanol 48-h LC₅₀ 3.2 Sanders (1970a)

Stone fly (naiad) 15.5 35 7.1 butoxyethanol 24-h LC₅₀ 8.5 (5.7-13) Sanders &
(Pteronarcys 15.5 35 7.1 butoxyethanol 48-h LC₅₀ 1.8 (1.5-2.7) Cope (1968)
californica) 15.5 35 7.1 butoxyethanol 96-h LC₅₀ 1.6 (1.3-1.9) Sanders &
15.5 35 7.1 acid (tech) 24-h LC₅₀ 56 (50-63) Cope (1968)
15.5 35 7.1 acid (tech) 48-h LC₅₀ 44 (32-59) Sanders &
15.5 35 7.1 acid (tech) 96-h LC₅₀ 15 (10-22) Cope (1968)

Midge (larva) 15 78-95 55 7.3-7.8 dimethylamine 24-h LC₅₀ 1490 Bunting &
(Chaoborus 15 78-95 55 7.3-7.8 dimethylamine 96-h LC₅₀ 890 (421-1211)Robertson
punctipennis) 20 78-95 55 7.3-7.8 dimethylamine 24-h LC₅₀ 1124 (1975)
---------------------------------------------------------------------------------------------------------
^a Stat = static conditions (water unchanged for duration of test); flow = flow-through conditions
(2,4-D concentration in water continuously maintained.
^b Alkalinity and hardness expressed as mg CaCO₃/litre.
^c imm. = immature.
^d PGBEE = propylene glycol butyl ether ester.

Table 5. Toxicity of 2,4-D to aquatic invertebrates: no observed effect levels
---------------------------------------------------------------------------------------------------------
Flow/ Temp Sali- Alkali- Hard- Water con- Refer-
Organism stat^a (°C) nity nity^b ness^b pH Formulation^c Parameter^d centration ence
(^o/oo) (mg/litre)
---------------------------------------------------------------------------------------------------------
Eastern oyster flow 9 19 free acid 96-h EC₀ 2.0 Butler
(Crassostrea shell growth (1963)
virginica) flow 30 23 free acid 96-h EC₀ 2.0 Butler
shell growth (1963)
flow 25 28 dimethylamine 96-h EC₀ 2.0 Butler
shell growth (1963)

Freshwater oligochaete flow 20 30 30 7.8 free acid 96-h LC₀ 86.7 Bailey
(Lumbriculus & Liu
variegatus) (1980)

Scud stat 21.1 30 7.1 dimethylamine 96-h LC₀ 100 Sanders
(Gammarus lacustris) (1969)

Grass shrimp stat 20 20 butoxyethanol 24-h LC₀ 10 Hansen
(Palaemonetes pugio) et al.
(1973)
Pink shrimp butoxyethanol 48-h LC₀ 1.0 Butler
(Penaeus duorarum) (1965)
PGBEE 48-h LC₀ 1.0 Butler
(1965)
---------------------------------------------------------------------------------------------------------
^a Stat = static conditions (water unchanged for duration of test); flow = flow-through conditions
(2,4-D concentration in water continuously maintained).
^b Alkalinity and hardness expressed as mg CaCO₃/litre.
^c PGBEE = propylene glycol butyl ether ester.
^d LC₀ and EC₀ represent the highest dose used which cause no death or no effect, respectively;
they are not mathematically determined no-effect levels.
George et al. (1982) measured lethal times (LT) after exposure of
the water flea Daphnia lumholtzi to 10 or 20 mg 2,4-D/litre. They
reported, for 10 mg/litre, an LT₅₀ of 38 h and an LT₁₀₀ of 71 h. For
20 mg/litre, the LT₅₀ was 21 h and the LT₁₀₀ was 31 h. Doses of
2,4-D ranging from 0.1 to 50 mg/litre did not affect the behaviour of,
or kill, the copepod Mesocyclops leuckarti within a 30-day exposure
period and so lethal times could not be calculated.

Caldwell (1977) and Caldwell et al. (1979) found the zoeal larva to
be the most sensitive life-cycle stage of the Dungeness crab (Cancer
magister) to the free acid of 2,4-D. Based on the herbicide's toxicity
to this stage, the authors suggest a maximum acceptable toxicant level
(MATC) of <1 mg/litre. At this concentration, there was no mortality,
but there was an effect on moulting.

6.1.2. Behavioural effects

Folmar (1978) tested mayfly nymphs (Ephemerella walkeri) in a `Y'-
shaped avoidance maze. A 2,4-D dimethylamine salt solution was run
into one arm of the maze and clean water was run into a second arm,
both at 400 ml/min. Numbers of nymphs in each arm of the maze were
counted after 1 h. No avoidance of 2,4-D was found at concentrations
of 10 mg/litre and there was no mortality. At 100 mg/litre there was
70% mortality in the test nymphs but still no avoidance of the
herbicide. In a similar experiment using the grass shrimp
(Palaemonetes pugio) exposed to the butoxyethanol ester of 2,4-D,
there was significant avoidance of the herbicide at 1 mg/litre (Hansen
et al., 1973).

6.2. Toxicity to Fish

Appraisal

LC₅₀ values for fish vary considerably. This variation is due to
differences in species sensitivity, chemical structure (esters, salts,
or free acid), and formulation of the herbicide.

The NOEL varies with the species and the formulation: <1 mg/litre
(coho salmon) to 50 mg/litre (rainbow trout).

Fish larvae are the most sensitive life stage but are unlikely to
be affected under normal usage of the herbicide.

Long-term adverse effects on fish are observed only at
concentrations higher than those produced after 2,4-D has been applied
at recommended rates.

Fish detect and avoid 2,4-D only at higher concentrations than
those obtained under normal conditions of use.

6.2.1. Effect of formulation on short-term toxicity to fish

The toxicity of different formulations of 2,4-D to fish is
summarized in Table 6.

The most comprehensive study on the effects of different
formulations of 2,4-D using the same test fish, fingerling bluegill
sunfish (Lepomis macrochirus), was performed by Hughes & Davis (1963)
in static 24-h and 48-h tests. Ester formulations were invariably more
toxic than amine salt formulations. Dimethylamine and alkanolamine
preparations ranged in toxicity from 166 to 900 mg/litre (LC₅₀ in 24-h
tests), depending on the commercial preparation used. Although esters
were always more toxic than amine salts, there was some variation
between different ester formulations (range: 0.9 to 66.3 mg/litre; 24-h
LC₅₀). Most of this variation was between different preparations of
the least toxic of the esters, the isooctyl ester, which ranged in
toxicity from 8.8 to 66.3 mg/litre. All other esters tested produced
LC₅₀ values of 8 mg/litre or less, the most toxic being the isopropyl
with a 24-h LC₅₀ of 0.9 mg/litre. The addition of emulsifiers to acid
preparations increased 2,4-D toxicity; a formulation with emulsifiers
gave an LC₅₀ of 8 mg/litre over 24 h, making it comparable to the
esters in toxicity. All ester formulations were considered by the
authors to present a major hazard to fish when used directly as an
aquatic herbicide, whereas the amine salt formulations could be safely
used to control aquatic weeds without adversely affecting adult fish
(Hughes & Davis, 1963).

A study on a range of ester formulations, using salmonids as test
fish, conducted by Finlayson & Verrue (1985), showed that the toxicity
for salmonids was similar to that for bluegill sunfish. These authors
argue that static tests underestimate the toxicity of 2,4-D esters
because some of the ester is hydrolysed to the less-toxic free acid
during the course of even short-term tests. The presence of test fish
increases the rate of hydrolysis of 2,4-D esters. In a static test,
with two different stocking rates of fish, the apparent toxicity of
2,4-D ester decreased with a greater density of test fish (rainbow
trout) because of this enhanced hydrolysis. Results are given in
Table 6. In their flow-through tests, results were adjusted to take
account of the hydrolysis of ester to 2,4-D acid during the course of
the experiment. Two values are given in Table 6 for each test. The
first is the calculated effect of the non-hydrolysed ester and the
second, entered as `total 2,4-D', is the observed effect of the mixture
of ester and free acid produced by hydrolysis during the course of the
experiment. There is as much as a five-fold difference between the two
values. Alabaster (1969) examined several formulations of 2,4-D in two
species of fish, and found that pelleted herbicide, either as clay-
based or resin-based pellets, was the least toxic to fish of any of the
formulations tested.

Table 6. Toxicity of 2,4-D to fish: effects of different formulations
---------------------------------------------------------------------------------------------------------
Organism Flow/ Temp Alkali- Hard- pH Formulation^c Parameter Water Reference
stat^a (°C) nity^b ness^b concentration
(mg/litre)
--------------------------------------------------------------------------------------------------------
Bluegill sunfish stat 25 40 29 6.9 alkanolamine 24-h LC₅₀ 450-900 Hughes &
(Lepomis macrochirus) stat 25 40 29 6.9 alkanolamine 48-h LC₅₀ 435-840 Davis (1963)
stat 25 40 29 6.9 dimethylamine 24-h LC₅₀ 166-542 Hughes &
stat 25 40 29 6.9 dimethylamine 48-h LC₅₀ 166-458 Davis (1963)
stat 25 40 29 6.9 di-N,N 24-h LC₅₀ 1.5 Hughes &
stat 25 40 29 6.9 di-N,N 48-h LC₅₀ 1.5 Davis
stat 25 40 29 6.9 2,4-D acid + 24-h LC₅₀ 8.0 (1963)
emulsifiers
stat 25 40 29 6.9 2,4-D acid + 48-h LC₅₀ 8.0 Hughes &
emulsifiers Davis (1963)
stat 25 40 29 6.9 isooctyl ester 24-h LC₅₀ 8.8-66.3 Hughes &
stat 25 40 29 6.9 isooctyl ester 48-h LC₅₀ 8.8-59.7 Davis (1963)
stat 25 40 29 6.9 PGBEE 24-h LC₅₀ 2.1 Hughes &
stat 25 40 29 6.9 PGBEE 48-h LC₅₀ 2.1 Davis (1963)
stat 25 40 29 6.9 butoxyethanol 24-h LC₅₀ 2.1 Hughes &
stat 25 40 29 6.9 butoxyethanol 48-h LC₅₀ 2.1 Davis (1963)
stat 25 40 29 6.9 butyl ester 24-h LC₅₀ 1.3 Hughes &
stat 25 40 29 6.9 butyl ester 48-h LC₅₀ 1.3 Davis (1963)
stat 25 40 29 6.9 mixed butyl + 24-h LC₅₀ 1.7 Hughes &
isopropyl esters Davis (1963)
stat 25 40 29 6.9 mixed butyl + 48-h LC₅₀ 1.7 Hughes &
isopropyl esters Davis (1963)
stat 25 40 29 6.9 isopropylester 24-h LC₅₀ 0.9 Hughes &
stat 25 40 29 6.9 isopropylester 48-h LC₅₀ 0.8 Davis (1963)
stat 25 40 29 6.9 ethyl ester 24-h LC₅₀ 1.4 Hughes &
stat 25 40 29 6.9 ethyl ester 48-h LC₅₀ 1.4 Davis (1963)

Cutthroat trout butyl ester 96-h LC₅₀ 0.78 Woodward (1982)
(juvenile) (Salmo clarki) (0.66-0.92)
PGBEE 96-h LC₅₀ 0.77 Woodward (1982)
(0.62-0.96)
isooctyl ester 96-h LC₅₀ > 50 Woodward (1982)

Chinook salmon (fry) flow 9 18 17 7.1 butoxyethanol 96-h LC₅₀ 0.315 Finlayson &
(Oncorhynchus flow 9 18 17 7.1 total 2,4-D 96-h LC₅₀ 0.373 Verrue (1985)
tshawytscha)
(smolts) flow 15 18 17 7.1 butoxyethanol 96-h LC₅₀ 0.375 Finlayson &
flow 15 18 17 7.1 total 2,4-D 96-h LC₅₀ 1.250 Verrue (1985)
flow 15 18 17 7.1 PGBEE 96-h LC₅₀ 0.246 Finlayson &
flow 15 18 17 7.1 total 2,4-D 96-h LC₅₀ 1.117 Verrue (1985)
---------------------------------------------------------------------------------------------------------

Table 6. (contd.)
---------------------------------------------------------------------------------------------------------
Organism Flow/ Temp Alkali- Hard- pH Formulation^c Parameter Water Reference
stat^a (°C) nity^b ness^b concentration
(mg/litre)
---------------------------------------------------------------------------------------------------------
Rainbow trout (fry) flow 15 18 17 7.1 butoxyethanol 96-h LC₅₀ 0.518 Finlayson &
(Salmo gairdneri) flow 15 18 17 7.1 total 2,4-D 96-h LC₅₀ 0.642 Verrue (1985)
flow 15 18 17 7.1 PGBEE 96-h LC₅₀ 0.329 Finlayson &
flow 15 18 17 7.1 total 2,4-D 96-h LC₅₀ 0.514 Verrue (1985)
(smolts) flow 15 18 17 7.1 butoxyethanol 96-h LC₅₀ 0.468 Finlayson &
flow 15 18 17 7.1 total 2,4-D 96-h LC₅₀ 1.338 Verrue (1985)
flow 15 18 17 7.1 PGBEE 96-h LC₅₀ 0.342 Finlayson &
flow 15 18 17 7.1 total 2,4-D 96-h LC₅₀ 1.555 Verrue (1985)
loading factor stat 14 18 17 7.1 butoxyethanol 96-h LC₅₀ 1.206 Finlayson &
4.2 g fish/litre Verrue (1985)
stat 14 18 17 7.1 total 2,4-D 96-h LC₅₀ 1.422 Finlayson &
loading factor stat 15 18 17 7.1 butoxyethanol 96-h LC₅₀ 3.689 Verrue (1985)
8.8 g fish/litre Finlayson &
stat 15 18 17 7.1 total 2,4-D 96-h LC₅₀ 4.487 Verrue (1985)

Harlequin fish flow 20 250 7.2 clay-based 24-h LC₅₀ 7000 Alabaster (1969)
(Rasbora heteromorpha) pellets
flow 20 250 7.2 resin-based 24-h LC₅₀ 3950 Alabaster (1969)
pellets
flow 20 250 7.2 resin-based 48-h LC₅₀ 3100 Alabaster (1969)
pellets
flow 20 20 7.2 sodium salt 24-h LC₅₀ 1160 Alabaster (1969)
flow 20 20 7.2 butoxyethyl 24-h LC₅₀ 1.0 Alabaster (1969)
flow 20 20 7.2 butoxyethyl 48-h LC₅₀ 1.0 Alabaster (1969)

Rainbow trout flow 20 250 7.2 clay-based 24-h LC₅₀ 7000 Alabaster (1969)
(Salmo gairdneri) pellets
flow 20 250 7.2 clay-based 48-h LC₅₀ 4800 Alabaster (1969)
pellets
flow 20 250 7.2 resin-based 24-h LC₅₀ 3400 Alabaster (1969)
pellets
flow 20 250 7.2 resin-based 48-h LC₅₀ 2400 Alabaster (1969)
pellets
flow 20 250 7.2 amine salt 24-h LC₅₀ 250 Alabaster (1969)
flow 20 250 7.2 amine salt 48-h LC₅₀ 210 Alabaster (1969)
---------------------------------------------------------------------------------------------------------
^a Stat = static conditions (water unchanged for duration of test);
flow = flow-through conditions (2,4-D concentration in water
continuously maintained).
^b Alkalinity & hardness expressed as mg CaCO₃/litre.
^c di-N,N = di-N,N-dimethylcocoamine; PGBEE = propylene glycol
butyl ether ester; total 2,4-D = the effect actually observed in the
flow-through test; the value which preceeds each "total 2,4-D" value is
the calculated effect of the ester alone. The authors determined the
degree of hydrolysis of the ester during the course of the test and
subtracted the effect due to the free acid produced by this hydrolysis.
6.2.1.1 Tolerance and potentiation

Chambers et al. (1977) used insecticide-tolerant and insecticide-
susceptible populations of mosquito fish and an esterase inhibitor to
investigate hydrolytic activation and detoxification of 2,4-D esters.
Mosquito fish taken from a wild population which had developed some
tolerance to insecticides also showed some slight tolerance to 2,4-D
ethyl and butyl esters. This tolerance was most pronounced with the
butyl ester, where the 48-h LC₅₀ was raised from 0.98 mg/litre in the
susceptible fish, to 1.70 mg/litre in the tolerant fish. Further
experiments were carried out to find the basis for this tolerance and
for the higher toxicity of 2,4-D esters over that of the free acid.
The addition of DEF (S,S,S-tributyl phosphorotrithioate), a carboxyl
esterase inhibitor, to