Society of Chemical Industryの代理人であるJohn Wiley & Sons Ltd.の許諾を得て転載
Pest Management Science

Pest Manag Sci 58:637±648 (online: 2002)
DOI: 10.1002/ps.479

ReviewHigher-tier laboratory methods for assessing
the aquatic toxicity of pesticides
Alistair BA Boxall,1* Colin D Brown1 and Katie L Barrett2
1

Cranfield Centre for EcoChemistry, Cranfield University, Shardlow Hall, Shardlow, Derby DE72 2GN, UK
Huntingdon Life Sciences, Alconbury, Huntingdon, Cambridgeshire PE17 5HS, UK

2

Abstract: Registration schemes for plant-protection products require applicants to assess the potential
ecological risk of their products using a tiered approach. Standard aquatic ecotoxicity tests are used atlower tiers and clearly de®ned methodologies are available for assessing the potential environmental
risks. Safety factors are incorporated into the assessment process to account for the uncertainties
associated with the use of lower-tier single-species ecotoxicity studies. If lower-tier assessments
indicate that a substance may pose a risk to the environment, impacts can be assessed using moreenvironmentally realistic conditions through the use of either pond mesocosms, arti®cial streams or®eld monitoring studies. Whilst these approaches provide more realistic assessments, the results are
dif®cult to interpret and extrapolation to other systems is problematic. Recently it has been recognisedthat laboratory approaches that are intermediate between standard aquatic toxicity tests and ®eld/
mesocosm studies may provide useful data and help reduce the uncertainties associated with standard
single-species tests. However, limited guidance is available on what tests are available and how they
can be incorporated into the risk-assessment process. This paper reviews a number of these higher-tierlaboratory techniques, including modi®ed exposure studies, species sensitivity studies, population
studies and tests with sensitive life stages. Recommendations are provided on how the approaches canbe incorporated into the risk-assessment process.
# 2002 Society of Chemical Industry

Keywords: higher tier; risk assessment; aquatic; ecotoxicity; pesticides
1 INTRODUCTION

Registration schemes for plant-protection products
require applicants to assess the potential ecological riskof their products using a tiered approach. Standard
tests are used at the lower tiers and clearly de®ned
methodologies are available for assessing the potential
environmental risks (eg toxicity exposure ratios or risk
quotients). For a lower-tier single-species study to be
representative of the wider environment the following
criteria would need to be ful®lled:1 the response of
selected organisms in single-species laboratory tests
should correspond to those of a larger array oforganisms in natural systems and the chosen endpoints
in single-species tests should be more sensitive thanany other at any level of organisation. As standard test
organisms do not necessarily meet these criteria,
uncertainty factors are incorporated into the riskassessment process. In order to re®ne the assessments,
it will be appropriate to evaluate critically the base data
set. However, if these assessments indicate that acompound is likely to pose a risk to the environment,
impacts can be determined using more environmentally realistic conditions.

A number of approaches have been used to addressconcerns arising from the preliminary assessments,
including the use of pond mesocosms,2 arti®cialstreams,3,4 ®eld monitoring studies3 and experimental
ditches.5 These approaches have been applied topesticides and surfactants, and have a number ofadvantages over single-species investigations, including the ability to assess: (1) endpoints at higher levelsof biological organisation, (2) species interactions and(3) indirect effects. The approaches also allow theassessment of population and community recovery.
There are, however, a number of limitations
associated with model ecosystem approaches, mostnotably that the results are dif®cult to interpret andextrapolate to other situations. Reasons for thisinclude the fact that no-effect concentrations are
dependent on the choice of test concentrations; studies
are performed at different times using different
starting conditions; studies vary in duration; sensitivities may vary across experimental systems; andmeasured endpoints are not always consistent orcomparable. The high background variation associated with the studies means that the discriminatory* Correspondence to: Alistair BA Boxall, Cranfield Centre for EcoChemistry, Shardlow Hall, Shardlow, Derby DE72 2GN, UK
E-mail: ecochemistry@cranfield.ac.uk
Contract/grant sponsor: Department of Environment, Transport and the Regions(Received 31 August 2001; revised version received 9 January 2002; accepted 18 January 2002)# 2002 Society of Chemical Industry. Pest Manag Sci 1526±498X/2002/$30.00

101

637

ABA Boxall, CD Brown, KL Barrett

power of the tests can be low. Moreover, in order tosimulate natural conditions and to ensure the maintenance of basic ecological characteristics and functions (ie different trophic levels, possibilities fororganisms to ®nd new habitats, energy input andnutrient cycles), the test systems need to be of asuf®cient size and complexity.6 This means that thestudies can be costly.
Approaches that are intermediate between standard
aquatic toxicity tests and ®eld and microcosm studies
can provide valuable data for use in the risk assessmentof plant-protection products. The objectives of theselaboratory-based higher-tier studies include one ormore of the following: to account for the effects of noncontinuous exposure; to improve the assessment ofspecies sensitivities to pesticides; to assess the effects ofcontaminants on sensitive life stages and populations;to determine the potential for organisms, populations
and systems to recover after exposure to a pesticide;and to determine indirect effects of compounds on
biological communities. There have been signi®cantdevelopments in the design and application of highertier laboratory toxicity tests, primarily over the last
5±10 years. It is thus timely to review the current
position and opportunities for future development.

2 REALISTIC EXPOSURE SCENARIOS

Preliminary acute-risk characterisation methods typically use the initial predicted environmental concentration (PEC) along with the results from standard
ecotoxicity studies. However, the method of entry andrate at which a compound dissipates in the environment may have a signi®cant impact on the relevance ofthe standard ecotoxicity endpoints used. A number ofapproaches have been proposed that account fordifferences in exposure scenarios, dissipation andbioavailability between natural systems and standard
toxicity studies.
2.1 Time-to-event analyses

Both exposure concentration and duration determine
the effect of a toxicant. However, the main approach toquantifying the effects of a toxicant currently focuseson the exposure concentration (eg determination ofLC50). As the tests are performed over a set time
period, they can neglect the effect of exposure
duration. Time-to-event (TTE) approaches characterise the toxicity of a compound by considering boththe intensity and duration of exposure.7 TTE analyses
have been used since the earliest days of toxicity
testing. The approaches draw upon experimental
designs where groups of exposed individuals aremonitored through time and the intervals to someevent, usually death, are recorded for each individual.
The resulting data can then be analysed using a range
of approaches, including: Kaplan±Meier methods, life
table methods, the semi-parametric Cox proportionalmodel and fully parametric models.8 Each of theapproaches can be used to predict time-to-event and totest the signi®cance of a co-variate (eg exposure
concentration) on a time-to-event.
The use of TTE methods explicitly includes
exposure duration in the assessment, but still allows
conventional endpoints (eg LC50) to be estimated. In
addition, TTE information enhances the power ofstatistical tests as more data are extracted per test
treatment, the effects of covariates can be measured
and included in predictive models, and results can beincorporated directly into ecological, epidemiologicaland toxicological models. Whilst summary data fromstandard toxicity tests are usually reported and used in
the risk-assessment process, the use of standard
methods (eg those recommended by OECD) meansthat additional data will have been recorded throughout the duration of the test. It may often be possible touse this supplementary data for TTE assessments
without the need for further experimental study.9
2.2

Variable and pulsed-exposure studies
Pesticide contamination of surface waters typicallyoccurs in pulses as a result of spray drift, overland ¯ow
and drainage inputs,10 and exposure of aquaticorganisms may be short for compounds that dissipate
rapidly. Continuous exposure, as used in standard
tests, may therefore not provide a true estimate of theecotoxicity of compounds and could result in bothover- and under-estimations of ecotoxicity.11,12 Possible reasons for this include: (1) organisms are able todetoxify or depurate any accumulated test compound
during the exposure interval;13,14 (2) induced individual toleranceÐthe ®rst pulse may strengthen survivors through acclimation or induction ofdetoxi®cation enzymes;15,16 (3) individual selectionÐweaker individuals may be removed by the ®rst
pulse, resulting in selection of more robust individuals
and an apparent reduction in responses to future
exposure; and (4) ecological recoveryÐafter exposure,an impacted population may or may not recover,depending on the characteristics of the species that are
affected.
For compounds that dissipate rapidly, the durationof exposure to a pesticide may be very short. Studies
can be performed with varying exposure intervals todetermine the impact, if any, of such short exposures.
For example, Gammarus pulex (L) were exposed tolambda-cyhalothrin for either 1, 3, 6, 12 or 96 h.17
After exposure, organisms were transferred to cleanwater for 96 h and after this time the effects wereobserved. There was a highly signi®cant relationshipbetween effects and duration of exposure. The
numerical value of the effect concentration after
exposure for 1 h was 18 times larger (ie less toxic)
than that for 96 h.17 Similar results have been obtainedfor Hyalella azteca using the pyrethroid cypermethrin.17
Whilst data from standard toxicity studies can beused to provide an initial assessment of the effects ofintermittent exposure to a contaminant,18 the approaches used assume that the toxicity of an inter-

638

Pest Manag Sci 58:637±648 (online: 2002)102

Assessing the aquatic toxicity of pesticides

mittent event is the same as that of a continuous
exposure test of equivalent dose. To overcome thisproblem, the effects of intermittent exposure can also
be assessed using `custom-designed' tests. A numberof studies have investigated the effects of pulsed
exposure on a range of organisms (including daphnids,
amphibian tadpoles, caddis ¯ies, chironomids and®shes) for a range of pesticides (including fenoxycarb,
fenitrothion, tebufenozide, chlorpyrifos and fenvalerate).11,19,20
Two approaches have generally been used, namely
static exposure studies and ¯ow-through studies.
Static renewal studies are particularly useful for
generating `square pulses' (ie constant exposure for a
set duration). This situation is most relevant whenreal-world exposure is expected to be dominated bywater ¯owÐfor example, where a pulse of chemical
travels quickly down a small stream. The use of ¯owthrough exposure systems allows pulsed exposure to
be more realistically simulated with gradual changes inconcentration. This approach has been used tosimulate the effects of degradation of a herbicide on
Selenastrum capricornutum.21 Three compounds wereused in the simulation, namely the parent compoundand two major metabolites, and the pulses of eachmimicked the fate of the compound in the naturalenvironment. Pulsed-exposure experiments can provide information on the potential for organisms torecover, the development of resistance to pesticide
exposure and any latent effects.
As a wide range of possibilities exists for the intensityand duration of exposure to pulsed release ofcontaminants, no testing regime can cover all pulse
discharge scenarios. Mechanistic models have therefore been proposed and tested for predicting thetoxicity of time-varying exposures (Table 1). Investigations into the effects of monochloroamine on rain-

bow trout (Oncorhynchus mykiss Walbaum) andcommon shiners (Notropis cornutus Mitch) demonstrated that concentration  time and Mancini-typemodels can predict an LC50 to within 50%.22 The
Mancini model performed better for longer-term pulse
exposures.
2.3

Inclusion of dissipation processes
Once a compound has entered the aquatic environment, its distribution, and hence potential to cause
adverse effects, will depend on a number of factors,
including the potential for abiotic (eg hydrolysis,oxidation, photolysis) and biotic (aerobic and anaerobic) degradation and the partitioning behaviour of thecompound to sediment and air. A number of studies
have investigated the incorporation of one or more ofthese processes into laboratory ecotoxicity studies.
The processes simulated have included spray drift,effects of suspended sediment, effects of pesticide
entering the environment in run-off and the effects of
bed sediment (Table 2).17,24
By adding sediment to the standard test system, thedegradation and adsorption processes occurring in theenvironment can be simulated.27 Studies using thehydrophobic synthetic pyrethroids, lambda-cyhalothrin (log Kow 6.8), esfenvalerate (log Kow 6.22) andfenvalerate (log Kow 5.0) have demonstrated that thepresence of soil or sediment in the test system results ina signi®cant reduction in the observed effect of apesticide (Table 2). This observation can be explainedby the fact that very little pesticide (ie <1% forlambda-cyhalothrin) is available in the water column
due to partitioning of the compound from the aqueous
phase to sediment.28 The data in Table 2 indicate thatthe addition of sediment had less effect on thehydrophilic compounds isoproturon (log Kow 2.5)
and pirimicarb (log Kow 1.7).

Table 1. Mechanistic models for assessing the effects of intermittent exposure to contaminants

Model
Concentration  time

Generalized model
22

Mancini uptake-depuration
model23dCsoa
 Ku Cwy
dtBreck damage-repair
model14

InputsAm  Cwy  td

ln

1



M
M

Am = a species speci®c constant for m% mortality (mglitre 1 h 1)
Cw = toxicant concentration in exposure water (mg litre 1)
td = time to death (exposure time required for m% mortality; h)y = a power term
kdep  Csoa


1
 d y  ln Cw   lne krep t
krep




 ln

Kdam
DL;50

Pest Manag Sci 58:637±648 (online: 2002)



Cw = toxicant concentration in exposure water (mg litre 1)
Csoa = concentration of toxicant at site of action (mg g 1)
Ku = uptake rate constant (h 1)
Kdep = depuration rate constant (h 1)t = exposure time (h)y = a power term
M = the observed mortality proportion
D = slope of logit response as a function of ln(Cw)Cw = toxicant concentration in exposure water (mg litre 1)
Kdam = damage rate constant (h 1)
Krep = repair rate constant (h 1)
DL,50 = lethal damage at 50% mortalityt = exposure time (h)y = a power term

639

103

ABA Boxall, CD Brown, KL Barrett
Table 2. Changes in observed toxicity caused by a range of modifications to water-only exposure studies17,23,25,26
Reduction factor in toxicity or BCF relative to water-only study

CompoundTest species

lambda-Cyhalothrin
Glyphosate
Esfenvalerate

Pirimicarb
Isoproturon
Herbicide 1

Daphnia magna
Daphnia magna
Daphnia magna
Lepomis macrochirus
Limephilus lunatus
Palaemonetes pugio
Daphnia magna (BCF)
Scenedesmus subspicatus
Selenastrum capricomutum

Herbicide 2

Navicula pelliculosa

Fenvalerate

Simulation ofspray drift

Simulation of spray
drift and suspended
sediment

Application to soil
prior to addition of
water

3
Ð
Ð
Ð
Ð
Ð
Ð
Ð
Ð
Ð
Ð
Ð

40
Ð
Ð
Ð
Ð
Ð
Ð
Ð
Ð
Ð
Ð
Ð

175
Ð
Ð
Ð
Ð
Ð
Ð
Ð
Ð
Ð
Ð
Ð

In the natural environment, water contains dissolved or colloidal organic matter (DOM). These
substances have been shown to interact with organic
compounds, resulting in effects on bioavailability
because only freely dissolved compounds are generally
assumed to be accumulated by organisms. Previous
studies using pesticides, surfactants and polycyclicaromatic hydrocarbons have demonstrated that thepresence of DOM reduces accumulation and toxicity.24,29±31 However, at low concentrations of DOM(ie less than 10 mg litre 1) bioaccumulation may beenhanced.32
In order to simulate the effects of natural concentrations of DOM on ecotoxicity, studies have also been
performed using water samples collected in the ®eld.33
Studies on a range of compounds have demonstrated
that the apparent toxicity of chemicals is only slightly
reduced in tests using natural waters (Table 3).33 Thissuggests that the use of laboratory water does notnecessarily introduce a signi®cant source of uncertainty.
Standard ecotoxicity studies usually require that thetest concentration is maintained at >80% of thestarting concentration throughout the test. For compounds that are volatile or degraded abiotically (eg via
hydrolysis) this can mean that the test system iscovered, that a ¯ow-through test system is required orthat the test solution is replaced regularly. Thus theimpact of abiotic degradation processes on toxicity

Table 3. Mean ratios of EC50 values
obtained using tests on 15 natural
waters to EC50 values obtained using
ISO standard media33

Zinc sulphate
4-Nonylphenol
Phosalone
Pentachlorophenol
2,4,5-Trichloroaniline

Simulation ofsuspended
sediment

Simulation of
bed sediment(not stirred)

Simulation of
sediment(stirred)Ð

120±140Ð

81±280
Ð
Ð
Ð
Ð
9700
Ð
2
Ð
Ð
Ð
Ð

3.3
Ð
Ð
10±100
Ð
Ð
Ð
Ð
Ð
Ð
Ð

3.0
5.0
Ð
Ð

0.7
Ð
>17
>900
>17
>900

could readily be assessed using a static exposuresystem with no replacement of test solution.
The results of modi®ed exposure studies can beused in conjunction with toxico-kinetic models
to assess toxicant effects under non-steady state
conditions. Compartment-based models such as
PULSETOX describe the movement of toxicants
between compartments;34,35 physiologically based
codes describe accumulation, elimination and distribution, and bioenergetically based models such as
FGETS and DEBtox simulate accumulation and loss
in terms of the organism's energy requirements.36,37
To apply the models, information is required on bodyresidue levels and toxic effects. This is a neglected areain ecotoxicology, and further work is required before it
can be incorporated into the risk-assessment process.

3 STUDIES USING ADDITIONAL SPECIESThere is no species that is `most sensitive' to all classesof contaminant.38±40 Species can vary in their sensitivity to chemicals by orders of magnitude.41 Preliminary assessments of pesticide risk to the aquatic
environment therefore incorporate factors of either 10
or 100, in part to cover uncertainties regarding therelative sensitivities of the standard test organisms
compared with the range of species that are likely to bepresent in the ®eld. One approach to reducing these
uncertainties is to test additional organisms in order to

Log Kow

Raphidocelis
subcapitata
microplates

Raphidocelis
subcapitata
erlensBrachionus
calyci¯orus

Daphnia
magna

na
5.92
4.30
5.12
3.45

5.7
0.6
0.7
2.7
1.9

3.0
0.7
1.2
2.1
2.2

2.1
1.2
1.1
1.2
2.1

1.2
0.9
1.7
3.1
1.0

2.3

1.8

1.5

1.6

Mean

640

Pest Manag Sci 58:637±648 (online: 2002)

104

Assessing the aquatic toxicity of pesticides

gain a more precise estimate of the most sensitive
endpoint and/or to derive a distribution of sensitivity.9
The testing of additional organisms also generates
more ecological information for use in site-speci®canalyses.
3.1 Selection of additional test species

A number of approaches have been proposed forselecting additional test species. The preferred approach will depend on a number of factors, including
the nature of the compound being tested and theavailability of data for compounds with a similar modeof action. For compounds that do not have a mode ofaction that is speci®c to a particular taxon, the Aquatic
Dialogue Group42 proposed that the test species
should include at least two species of ®sh, one
invertebrate and one aquatic plant (macrophyte oralgae) plus four other species. The other four test
species should be selected from the group(s) that is/areshown to be the most sensitive after the initial studies
have been completed.42 If chronic studies are to be
performed, it is recommended that one ®sh, one
invertebrate and one plant be used. A further organism
should be selected from the most sensitive group
identi®ed in acute tests.42 Opportunities for extrapolation should not be overlooked for compoundswith a strong similarity to an existing substance. For
example, the number of additional organisms required
to show the range of species sensitivity might justi®ably
be reduced on the basis of data for a similar
compound.
The USEPA43 have recommended that additional
data should include invertebrate acute and chronic
tests, sediment toxicity tests, rooted plant testing andamphibian testing. The increase in the number andtype of invertebrate toxicity tests re¯ects an attempt toprovide some consideration of the range of potential
responses. In assessing impact on aquatic invertebrates, risk assessments generally focus on Daphnia
magna Straus, a parthenogenic, short-lived crustacean.
The following should be considered when selecting
additional species: (1) different and often variable lifespans; (2) unpredictable lengths of life stages anddifferent metamorphic stages; (3) indeterminate
growth in some species; and (4) signi®cant differencesin adipose stores.43 The USEPA suggest that additional species for invertebrate acute toxicity tests could
include stone¯ies and amphipods, whereas chronic
tests should address effects on sexually reproducing
invertebrate species including copepods and chironomids. The addition of an acute amphibian test re¯ects
the need to consider effects on amphibian species.
The sensitivity of individual species is primarily
dependent on the compound's mode of action, andspecies from the same taxonomic group (either
phylum or class) generally respond to contaminants
in a similar manner.44 For chemicals with a non-polar
mode of action, variation in species sensitivity is small.
However, for compounds that are either reactive or
have a speci®c mode of action, the variation can be aslarge as 105±106. Based on an understanding of themode of toxic action of a compound, it may be possible
to identify and group sensitive and less sensitive
organisms. This allows the testing strategy to befocused on groups at high risk.8 For example, Cuppen
et al. 45 recommended the use of non-arthropod
macroinvertebrate species at an advanced stage ofthe risk assessment of fungicides. This is based on theassumption that the spectrum of effects of fungicides
in aquatic systems is similar to results obtained forcarbendazim and pentachlorophenol. Similarly, tests
on insects would be recommended for substances withan insecticidal activity. It should be noted that mode of
toxic action is often not well understood, particularly
for new compounds. Whilst the mode of action of apesticide is an important criterion for grouping oforganisms, habitat, reproductive strategy and life cycle
may also affect the sensitivity of species.
Recent studies46,47 have assessed the suitability of abattery of ecotoxicity tests that evaluate the sub-lethal
and lethal effects of chemicals on algae, protozoans,rotifers, crustaceans and dipterans. The aim of thestudies was to identify tests that could be adopted in
the risk-assessment process. Comparison of the resultsfrom the tests with those from stream and pond
mesocosm studies for lindane, 3,4-dichloroaniline,
atrazine and copper compounds demonstrated thatlowest observed effect concentrations obtained in thelaboratory were generally similar to results obtainedusing pond mesocosms and arti®cial streams (ie withina factor of 6). The exception to this was 3,4dichloroaniline, for which the pond mesocosm was
200 times more sensitive. Similar studies on isoproturon48 demonstrated that single-species tests using a
range of organisms resulted in a more conservative
assessment than a microcosm study.
Campbell et al 9 suggest that, for compounds that do
not have a mode of action speci®c to a particular taxa,
eight species should be tested as a minimum todescribe the distribution of sensitivities of aquaticorganisms. In cases where it is known that a speci®cgroup of organisms is particularly sensitive, theadditional test species should be selected from themost sensitive group. However, if ®sh are the mostsensitive group, fewer test species are required and ®vespecies are probably suf®cient. This is for animal
welfare reasons and because distributions of sensitivity
are usually narrow for ®sh. By using at least eight
additonal test species (®ve for ®sh), there is a highprobability that a good estimate (ie within an order ofmagnitude) of the most sensitive endpoint will beobtained.9 The use of eight additional species also
allows a probabilistic approach to be used for effects
characterisation.42
3.2

Analysis of data
One of the objectives of laboratory ecotoxicity testing
is to assess the potential for impact of a chemical on
natural ecosystems. Rather than using uncertainty
factors, the data from single-species studies can be

Pest Manag Sci 58:637±648 (online: 2002)

641

105

ABA Boxall, CD Brown, KL Barrett

extrapolated to ®eld conditions using either regression
techniques or distribution models.49 A wide range ofapproaches are available that are based on the samegeneral principle (ie they assume that experimental
ecotoxicity data ®t a given distribution), but differ inthe statistical distributions that they use.10,37,39,50±52
Outputs from the methods include: (1) hazardous
concentrations for a speci®ed percentage of species(HCx where x is the speci®ed percentage);51,53 (2)concern levelsÐlevels that can be expected to cause anenvironmental effect at least 95% of the time54 and (3)®nal chronic valuesÐa threshold concentration for
unacceptable effects that is calculated from the lowestof three chronic values (the ®nal chronic value, the®nal plant value and the ®nal residue value). The ®nal
chronic value is calculated from the no observed effectconcentrations (NOECs) for at least eight families.55
There may be problems with availability of internationally accepted methods for this number of chronic
tests (eg for invertebrates).
The results of single-species distribution methods
have been compared with those from multi-species
studies (Table 4).56,57 In most cases, the extrapolation
procedures lead to lower `safe' values than NOECs
from multi-species studies and demonstrate that ®eld
effects can be predicted if uncertainties related tomode of action (eg species groups tested and exposure
regimes) are accounted for. The studies have demonstrated that the larger the number of species, thesmaller the difference between the NOECs derived
using multi-species studies and the `safe' values
derived using statistical techniques.
There is debate over which value from a range of
species sensitivities is most appropriate to protect theaquatic environment. Van den Brink et al 57 suggestthat the 5th percentile value that is based on acute

L(E)C50 data is a generically applicable criterion thatis protective for exposed communities (the higher thenumber of species tested, the lower the difference
between the NOEC for an ecosystem and the 5th
percentile value). Whilst the 5th percentile is therefore
the accepted norm, previous studies with pyrethroids60 and atrazine10 have proposed that the 10th
percentile effect concentration is adequate.
More recently, the use of bootstrapping to derivehazardous concentrations has been investigated.61
The approach involves random sampling (withreplacement) of an ecotoxicity data set for a compound to create a resampling data set (eg of 100
observations). The resulting data set is then ranked
and the 5th percentile value is selected as the HC5.
The resampling is performed a number of times (eg
10 000) to produce a number of HC5 estimates. These
are then ranked, and the value corresponding to 50%
is taken as the best estimate of HC5. Estimates at 2.5%
and 95% are used as the 95% bootstrap con®dence
limits. Large data sets are required to derive reliable
HC5 estimates because knowledge of the mode ofaction of the substance and most sensitive species is
not accounted for.
3.3

Current limitations to the approach
Whilst the use of additional test species provides useful
data, there are a number of limitations to theapproaches compared to standard ecotoxicity studies.
These include: large datasets are required to ensurethat the resulting analyses are as representative as
possible; there is disagreement about the number andtaxonomic distribution of species to be tested; test
results may not be directly comparable; test guidelines
are only available for a few test species; organisms may
not be from the same sensitivity distribution and there

Table 4. Comparison of multi-species NOEC values with results obtained using extrapolation techniques on single-species toxicity data.46,56,57 All values are mg
litre 1

CompoundMS NOEC 58

LindaneDichloroaniline
Atrazine
Copper compounds
Isoproturon
Atrazine
Di¯ubenzuron
Chlorpyrifos
Di¯ubenzuron
Parathion
Azinphos-methyl
Parathion-methyl
Pentachlorophenol
1,2,4-Trichlorobenzene
Trichloroethylene
Tri¯uralin
Dichlobenil
Permethrin
Toxaphene

0.22
1.0±12<3.0±5
1.1
2
20
0.3
0.1
0.1
0.1
0.20±0.25
0.1
20
57
2.8
0.5
5.0
0.023
>1.5

a

Aldenberg andSlob 59 50%0.121200.82±2.2
0.28

Aldenberg andSlob 59 95%

Wagner and
Lokke 53 50%0.13
1200.87±2.1
0.30

0.12

Wagner and
Lokke 53 95%

SSD(method notreported) a 95%

USEPA 54
0.067
0.15

0.17
0.63
13
0.18
0.044

Ð
0.013
0.15
Ð
3.2
44
3100
Ð
Ð
Ð
0.003

Ð
0.00013
0.085
Ð
0.53
1.1
65
Ð
Ð
Ð
<0.0001

Ð
0.011
0.077
Ð
2.6
39
3,100
Ð
Ð
Ð
0.003

Ð
0.00023
0.018
Ð
0.66
1.7
97
Ð
Ð
Ð
<0.0001

0.037

0.001
0.002
0.01
0.024
0.32
19
130
0.2
7.8
0.066
0.0025Species sensitivity distribution.

642

Pest Manag Sci 58:637±648 (online: 2002)

106

Assessing the aquatic toxicity of pesticides

is no agreed way of testing this; and the ecology andphysiology of the test organism may not be fully
understood, making extrapolations to population
effects problematic. Moreover, for many species,
laboratory culturing procedures are not available andthe test species will need to be collected from the ®eld.
This can mean that the age and quality of theorganisms cannot be guaranteed.62 Where possible,organisms collected from the ®eld should be takenfrom an uncontaminated site that has not been
exposed to the test substance or chemicals with asimilar toxic mode of action. In some instances, thetest organism may not be available throughout theyear. Therefore, when designing such studies, either
species should be selected that are available all of theyear or a wider range of possible test animals should beavailable so that a selection can be made at any time.
The use of geographical information systems (GIS) to
identify the proportion of a particular species andhabitat exposed to a given pesticide could also assist in
the identi®cation of additional species for testing.43

4 ORGANISM RECOVERY

Assessment of recovery is often a component incorporated into lower-tier toxicity tests for algae and thehigher plant Lemna sp. An aliquot of previously
exposed cells or fronds is placed in clean medium toestablish whether cell division is reversibly or irreversibly retarded. However, the principle can be extended
to other aquatic plants, invertebrates and ®sh. A few
recovery studies have been reported for pesticides,
including organophosphorus and carbamate pesticides, and for organisms including daphnids, chironomids and black¯y.63,64
The potential for recovery will be determined by thelength of time between exposures and the speci®cmode of action of the compound being considered.
For example, studies into the effects of carbamate
pesticides on chironomids indicated that toxic effects
were reduced if animals had been allowed to recoverfor more than 6 h.63 This observation was probably
due to the reactivation of acetylcholinesterase during
the recovery periodÐcarbamate pesticides have beenshown to be reversible inhibitors of AChE. No suchrecovery was observed for organisms exposed toorganophosphorus insecticides, which are considered
to be irreversible inhibitors.63

5 SYSTEM RECOVERY

Once a system has been impacted by a contaminant,its rate of recovery will be determined by thepersistence of the contaminant and the ecology,
physiology and biochemistry of organisms in thesystem and in proximity to the system.
One approach to predict the likely recovery of asystem is based on the dissipation half-life of thecompound, the initial exposure concentration and thehazardous concentration for 5% of species.65 If

recovery is required within a year after a single
application, then the chemical's half-life should meetthe condition:
T1=2 <

ln 2


C0
ln
HC5

where T1/2 is measured in years, C0 is the initialconcentration and HC5 is the hazardous concentrationfor 5% of species. Although dissipation will benecessary for complete recovery, under some conditions it may not be suf®cient and ecological recovery
may lag behind the disappearance of the chemical.
This lag will depend on a range of factors including
accessibility of shelters, survival of resistant life stages
and presence of untreated areas from which recolonisation can occur. In addition, ability to recolonise will
depend on habitat selection and life history
A number of experimental studies have been
proposed for determining whether an impacted system
is likely to be re-colonised. This is particularly relevant
for mobile organisms such as ®sh, which may beresident in a section of a water body for a short timeand may move from a toxic system. The study design
consists of adding the test organism at the start of thestudy and at regular intervals thereafter. A similar
approach can also be used with those invertebrate
species with a high potential for re-colonisation, withthe re-introduction of individuals into a previously
treated system and monitoring of their subsequent
survival and performance. For example, Crane et al. 66
investigated the toxicity of pirimiphos-methyl toGammarus pulex. Beakers were spiked with the study
compound at day 0 and toxicity to Gammarus pulex
was assessed over 24 h at 1, 4, 8 and 12 days after
application. Generally, mortality decreased with increasing time from spiking, probably as a result of
pesticide degradation.
The factors that in¯uence the recovery of a
population after a signi®cant perturbation are complex. One important factor is the life history of theindividual species.67 Aquatic organisms can vary
widely in their reproductive and dispersal characteristics. For example some species may reproduce
continuously, whereas others may reproduce at
discrete times of the year. Generation times can varyfrom days (eg daphnids) to years (eg dragon¯ies). The
ability for recolonisation can also varyÐmany aquatic
insects have adult life stages with wings whereas othertaxa (eg crustaceans and molluscs) will rely on
recolonisation by more passive forms of dispersal(wind, ¯ooding, transportation by birds).

6 SENSITIVE LIFE STAGE STUDIES

Sensitivity to toxicants is known to vary between
species, but sensitivity can also vary between the
different development stages of an organism.68±71
Standard ecotoxicity studies generally focus on neo-

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ABA Boxall, CD Brown, KL Barrett

nate or juvenile animals, which are likely to be the mostsensitive life stage(s). However, in instances where it is
known that a particular substance is likely to be more
toxic to a life stage not studied in the standard tests (egbased on the results of studies with related compounds
or based on the mode of action of the substance), it
may be appropriate to test this additional life stage
and, as a result, incorporate a smaller uncertainty
factor into the risk-assessment process. For example,
there are instances where older organisms have beenshown to be more sensitive to a particular toxicant
than younger ones. Studies with daphnids20 haveindicated that older organisms are more sensitive thanneonates to chlorpyrifos. One possible reason for thisis that the older animals have an increased ®ltration
rate and receive a higher dose of chemical, as
chlorpyrifos is likely to be associated with food. Berrill
et al 72 showed that 2-week-old tadpoles were moresensitive to endosulfan than newly hatched tadpoles
because endosulfan affected the post-hatch development of the neuromuscular system. By understanding
both the signi®cance of different routes of uptake fordifferent classes of chemical and the physiology of anorganism, it may be possible to predict these types ofdifference in species sensitivity and hence identify themost appropriate life stage to test. The testing ofsensitive life stages may be problematic due to the lackof test and culture methods for certain species.

7 POPULATION LEVEL STUDIESAs standard toxicity studies are generally performedon the most sensitive life stages, results may overestimate effects on populations. Life-cycle and population-level studies can be useful tools for assessing the
likely impact of a pesticide on populations of aquaticorganisms. The structure of a population at the time of
toxicant exposure may be an important factor in
determining toxicant impact.73 Populations in nature
consist of a mixture of life stages, yet toxicologicalstudies usually only start with one life stage. Consequently it may be appropriate to consider age/stage
structure when evaluating toxicant effects.
Both modelling and experimental approaches canbe used to determine population level effects. Laboratory experimental studies have been performed using
plant and invertebrate species9 and have ranged from
simple studies evaluating effects on a limited numberof life stages to complex multi-life-stage studies thatsimulate natural population dynamics.74±77 This typeof test is normally restricted to invertebrates, plantsand ®sh species that have relatively short life histories,
and may not be possible with groups prone tocannibalism unless the individuals are held separately.
Experiments with other ®sh may be possible under®eld conditions.78,79
The data generated from population-level studies
and from simple toxicity studies can be used to model
the effects of compounds on populations in the®eld.80±83 Population models describe the dynamics

of a ®nite group of individuals through time and havebeen used extensively in ecology and ®sheries management. These approaches are well developed for use on®sh84,85 and invertebrates.86±89 By using models, theresults of laboratory studies can be used to evaluate a
range of scenarios9 and can assist in identifying shortand long-term changes in population structure.
Appropriate use of population models requires anunderstanding of the natural history of the species
under consideration, as well as knowledge of how the
toxicant in¯uences its biology. Model inputs can
include somatic growth rates, physiological rates,
fecundity, survival rates of various classes within
populations, and how these change in response to
the study compound.90 In addition, populationdensity effects may be important.

8 HIGHER-TIER LABORATORY STUDIES WITHINENVIRONMENTAL RISK ASSESSMENT FORPESTICIDES

Higher-tier studies are generally carried out in
response to indications of unacceptable risk from alower-tier assessment. The exact nature of the study or
studies to be performed will depend on a number offactors including the results of lower-tier studies andthe physicochemical properties, environmental fateand use pattern of the compound of interest. A
summary of the major higher-tier study types andwhere they ®t within current procedures for risk
assessment of pesticides is provided in Table 5.
Generally the results from the approaches can be usedin three ways to re®ne the lower-tier assessment: (1)the results can be used to justify the use of a lower
uncertainty factor (eg species-sensitivity studies, organism-recovery studies; population-level studies); (2)the results can be used to justify an increase in theeffects concentration used in the risk assessment (egmodi®ed exposure studies); and/or (3) the results canbe used to perform a probabilistic assessment (egspecies-sensitivity studies). The combination of anumber of the approaches (eg use of population
studies coupled to pulsed exposure and dissipation
studies) may also be appropriate.
Higher-tier laboratory approaches have a number ofadvantages over the use of standard ecotoxicity tests
and/or mesocosm/®eld studies. These include: (1) the
endpoints are more pertinent to actual environmental
exposure; (2) they provide scope for the integration of`realistic' measures of fate and exposure into the riskassessment process; (3) they should provide morecon®dence when predicting actual effects in theenvironment; (4) they can be designed to determine
effects on speci®c endpoints, eg mortality, growth,
behaviour; (5) they can be used to identify more clearly
those species that may be at risk, and assist in thetargeting of multi-species tests; (6) the results mayhelp explain observed effects (eg mode/speci®city ofaction); (7) recovery can be assessed for bothindividuals and populations; (8) there are none of644

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Assessing the aquatic toxicity of pesticides
Table 5. Higher-tier toxicity methods
Test type

Rationale for test

Testing ofadditonal
species

To reduce uncertainty factors
and/or develop species
sensitivity distributions

Possible design of test

Data analysis

Test a number of (8 may suf®ce) Lower uncertainty factors by up
additional species:to an order of magnitude or
From a range of groups
use statistical approaches togenerate species-sensitivity

OutputAdvantages/disadvantages

Revised Toxicity Exposure Ratio Identi®cation of most sensitive(TER) or Hazardous
species
Concentration for a speci®ed More relevance than standard
percentage of species (HCx)orspeciesRe®nement of TERs
Allows probabilistic risk

For compounds whose mode ofaction indicates that a
particular group will besensitive, tests should be
performed on species fromassessment
There is disagreement on the
Number of species that need
to be tested
No guidance available on
uncertainty factors to use

distributions

this grouporIf ®sh are likely to be sensitive,

No guidance available on whatlevel of population protectionis acceptable
Lack of standard test methods

then only 5 additional species
should be tested

for additional species
May need to obtain organisms
from the ®eld

Time-to-event
analysis

To assess the toxic effects of
exposure to pesticides overdurations shorter than thestandard test duration

Use data recorded over time for Compare time-to-event data
standard ecotoxicity tests
with expected exposure
duration

If duration < time to effect thenreduce effect estimate

More realistic assessment of
ecotoxicity
Does not address `latency'Ðieeffects brought on by earlier
exposure
Number of observations madeduring the standard test may
not be adequate
No guidance on how effects
estimate can be reduced

Short-term
exposure
studies

To assess effects of short
exposure durations on
pesticide toxicity

Perform test over exposure
Use standard toxicity test
duration predicted for the ®eld
statistical procedures

Pulsedexposurestudies

To simulate pulsed exposure
Static renewal or ¯ow-throughconditions that are likely in thesystem, depending on detail®eldrequired

Use standard toxicity test
statistical procedures and/or
toxico-kinetic models

Revised effect measurement for More realistic assessment of
use in risk characterisation
exposure
Does not consider delayed
effects
Revised effect measurement for More realistic assessment of
use in risk characterisation
exposure
Includes potential for organisms
to recover
Results affected by a number ofvariables (including pulse
duration, time between
pulsed, organism recovery)
so dif®cult to interpret

Dissipation
studies

To assess the impact ofdissipation processes on
toxicity

Generally performed using
Use standard toxicity test
sediment/water systems.
statistical procedures
Studies could also be
performed to assess effects ofphotodegradation (byRevised effect measurement for More realistic assessment of
use in risk characterisation
exposure
Can include effects ofmetabolites
Guidelines not available on

performing studies in thelight), hydrolysis (byperforming a static test rather
than a ¯ow-through study) or
volatilisation.

Studies using
natural

Account for differences
between bioavailability inmatrices

natural and laboratory
systems

Effects on
To determine whether asensitive life
particular life stage isstagessensitive or not
Population
studiesselection of test matrices
Not suitable for all species (egalgae)
Methods not available forassessing photodegradationand volatilisation

Perform toxicity studies using
Use standard toxicity test
water samples collected fromstatistical procedures

Revised effect measurement for Assessment of bioavailability
use in risk characterisation
under natural conditions

the ®eld

No guidance available on
selection of test matrix

A range of approaches areavailable for a number of
species.

Use standard toxicity test
statistical procedures

Toxicity to sensitive life stagesÐ Reduction of uncertainty factors
possible reduction in
uncertainty factor

Perform studies on a population Experimental and modellingof a particular species over aapproaches are available
prolonged time period

Use population modelling
approaches to interpret
results

More ecologically relevant
assessment of effectsÐ
possible reduction in
uncertainty factor

Recoverystudies

To determine whether
Expose organisms and transfer
populations can recover after
to clean media to determine
exposure to a pesticide
time to recovery

Includes effects of recovery
includes selection of tolerantorganisms
May be more ecologically
representative

Assessment of time required for Dif®cult to perform on ®shorganisms to recover
Disagreement over acceptable
recovery period

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ABA Boxall, CD Brown, KL Barrett

the complications that are associated with the interpretation of data from multi-species studies; (9)relative to mesocosm studies, they can be performed
with less regard to the season, the test organisms areusually readily available and the system is more
controlled; and (10) sub-lethal effects can be determined.
However the approaches also have limitations: (1)they cannot provide information on species interactions; (2) results can be more dif®cult to interpret andcompare than those from standard single-species tests;(3) the tests may be more costly than standard singlespecies studies; (4) standardised higher-tier procedures are not available; and (5) the quality ofavailable organisms cannot always be guaranteed.
Moreover, there is currently considerable debate anddisagreement over the relevance of different test
designs and outcomes. For example there is considerable disagreement over (1) the number of species thatshould be studied using the additional-species approach, recommended numbers range from less than
10 to 30±40; (2) the hazardous concentration levelselected for use in probabilistic risk assessment(ranging from 5 to 10%); and (3) the acceptability ofdifferent recovery periods.

9 CONCLUSION

Higher-tier laboratory tests provide a useful intermediate between standard toxicity studies and semi®eld/®eld tests. The review presented here has shownthat there are a wide range of higher-tier studies
available and that a signi®cant body of work has been
generated on most, but not all, of these methods.
There is a need for ¯exible guidelines to be developed.
However these guidelines should not be so ¯exible that
it is impossible to compare results generated indifferent laboratories. The next 5 years are likely tosee signi®cant advances in the development of test
methodologies and in the use of models in support ofdata analysis and risk assessment.
The most important constraint associated withhigher-tier laboratory methods concerns their validation status. There is a need to validate the methods in aconsistent and systematic way by comparing data andrisk assessments generated from standard laboratory
testing, higher-tier laboratory testing and mesocosms.
This would allow evaluation of the advantages andlimitations of laboratory higher-tier testing compared
to standard studies and mesocosm studies.
Data sets on the ecotoxicity of chemicals to species
should be further analysed on the basis of informationon the mode of action of a compound. The development of lists of sensitive species associated with aparticular mode of action would be highly bene®cial
when selecting organisms for testing. Many of theavailable test methods to support species-sensitivity
analyses were developed in North America andadditional tests may be required for application inother regions. Current methods for assessing effects oncommunities from distributions of species sensitivity
assume that all species are equally important. This isclearly not the case and consideration should be given
to identifying important species in aquatic ecosystems.
These species could include economically importantorganisms, keystone species or species that areaesthetically important in a particular water body.
ACKNOWLEDGEMENTS

The authors would like to thank Dr Mike Collins andMr Steve Norman for useful dialogue during thecourse of the project and Professor Allen Burton, Dr
Andy Girling, Dr David Pascoe and Dr Martin
Streloke for their constructive comments on an earlydraft of this manuscript. This study was funded by the
Department of Environment, Transport and theRegions.
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