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Ecotoxicology and Environmental Safety ] (]]]]) ]]]–]]]

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Ecotoxicology and Environmental Safety

0147-65

doi:10.1

� Corr

E-m

PleasCaen

journal homepage: www.elsevier.com/locate/ecoenv

Measuring and modelling mixture toxicity of imidacloprid and thiacloprid onCaenorhabditis elegans and Eisenia fetida

Jose L. Gomez-Eyles a, Claus Svendsen b,�, Lindsay Lister b, Heather Martin c,Mark E. Hodson a, David J. Spurgeon b

a Department of Soil Science, School of Human and Environmental Sciences, University of Reading, Reading RG6 6DW, Berkshire, UKb Centre for Ecology and Hydrology, Monks Wood, Abbots Ripton, Huntingdon PE28 2LS, Cambridgeshire, UKc School of Biosciences, University of Birmingham, Birmingham B15 2TT, UK

a r t i c l e i n f o

Article history:

Received 16 February 2007

Received in revised form

2 June 2008

Accepted 12 July 2008

Keywords:

Concentration addition

Synergism

Antagonism

Dose level dependence

Neonicotinoids

Piperonyl butoxide

13/$ - see front matter Crown Copyright & 20

016/j.ecoenv.2008.07.006

esponding author. Fax: +441487 773 467.

ail address: csv@ceh.ac.uk (C. Svendsen).

e cite this article as: Gomez-Eyles,orhabditis elegans and Eisenia fetida.

a b s t r a c t

While the standard models of concentration addition and independent action predict overall toxicity of

multicomponent mixtures reasonably, interactions may limit the predictive capability when a few

compounds dominate a mixture. This study was conducted to test if statistically significant systematic

deviations from concentration addition (i.e. synergism/antagonism, dose ratio- or dose level-

dependency) occur when two taxonomically unrelated species, the earthworm Eisenia fetida and the

nematode Caenorhabditis elegans were exposed to a full range of mixtures of the similar acting

neonicotinoid pesticides imidacloprid and thiacloprid. The effect of the mixtures on C. elegans was

described significantly better (po0.01) by a dose level-dependent deviation from the concentration

addition model than by the reference model alone, while the reference model description of the effects

on E. fetida could not be significantly improved. These results highlight that deviations from

concentration addition are possible even with similar acting compounds, but that the nature of such

deviations are species dependent. For improving ecological risk assessment of simple mixtures, this

implies that the concentration addition model may need to be used in a probabilistic context, rather

than in its traditional deterministic manner.

Crown Copyright & 2008 Published by Elsevier Inc. All rights reserved.

1. Introduction

In natural environments, organisms are frequently exposed tomixtures of pollutants and it is relatively uncommon to find sitespolluted with one toxicant alone (Altenburger et al., 2003). Thenon-interactive effects of mixtures of chemicals can potentially bepredicted using two existing standard models of concentrationaddition (CA) and independent action (IA). When comparing thesetwo models for joint effect prediction of similar acting com-pounds, the CA model in general yields the best predictions(Altenburger et al., 2000). Equally, it has been shown that the IAmodel predicted the effects of dissimilarly (independently) actingchemicals more closely than the CA model (Backhaus et al., 2000).

Although useful as default models, interaction betweenchemicals can potentially limit the predictive capability of thesetwo models to describe joint effects. In a recent review McCartyand Borgert (2006) identified that most tested mixtures are nearor below simple dose/concentration additivity. Exceptions (bothpositive and negative) tend to occur when tested mixtures have

08 Published by Elsevier Inc. All r

J.L., et al., Measuring andEcotoxicol. Environ. Saf. (20

only a few components or where sensitive whole organism or sub-organismal changes are used as the response metric (McCarty andBorgert, 2006). The basis for such interactions could includechemical and physicochemical interactions with the constituentsof the soil, toxicokinetic interactions affecting uptake andelimination or toxicodynamic interactions affecting compoundmetabolism or associations at the target site (Van Gestel andHensbergen, 1997).

Initial work to validate the CA approach to predict mixturetoxicity effects of similarly acting compounds focussed oncomplex (up to 50 chemical) mixtures initially of narcotic (non-specifically acting) compounds (Hermens et al., 1984a, b), but laterof similarly acting non-narcotic chemicals (Altenburger et al.,2000; Backhaus et al., 2000). Subsequently, some studies havegone on to use multiple dose–response analysis of singlecompounds and their combinations to investigate whetherobserved joint effect data derived from experiments includingdifferent effect levels and ratios of combined exposure showsystematic deviations from CA and IA.

An approach that can be used for testing for such deviationsin the effects of mixture from model predictions was developedby Jonker et al. (2005). This considered that in the absenceof confirmed mechanistic information (i.e. undefined species

ights reserved.

modelling mixture toxicity of imidacloprid and thiacloprid on08), doi:10.1016/j.ecoenv.2008.07.006

ARTICLE IN PRESS

J.L. Gomez-Eyles et al. / Ecotoxicology and Environmental Safety ] (]]]]) ]]]–]]]2

relevant mode of action for one or more of the tested chemicals), CAand IA can be used as equally valid reference models against whichmeasured joint effect data can be statistically compared. Bydeveloping a method to compare model predictions to experimentaldata, Jonker et al. (2005) were able to evaluate whether observedmixture data deviated significantly from CA or IA predictions.

In this paper, the data analysis framework developed by Jonkeret al. (2005) (hereafter referred to as the MIXTOX model) is usedto identify systematic deviations (i.e. synergistic, antagonistic,dose level-dependent and dose ratio-dependent deviations) fromCA for mixtures of two widely used neonicotinoid insecticides,imidacloprid and thiacloprid in two species, the nematodeCaenorhabditis elegans and the earthworm Eisenia fetida. The twopesticides were chosen as they both work through the samemolecular mechanism, namely by impacting on the nicotineacetylcholine receptors of postsynaptic membranes, thereby,acting as agonists to acetylcholine receptors (Tomizawa andCasida, 2005). The known similar mode of action of thesechemicals means that theoretically joint effects should accord topredictions of the CA model—this being the simplest testablescenario. The two species were selected to assess the phylogeneticconservation of response patterns to mixture exposure in speciesfrom different major sub-branches of the animal kingdom(Philippe et al., 2005). Finally for the nematode, a further studywas undertaken that included an analysis of the effect of addingpiperonyl butoxide (PBO) to the imidacloprid and thiaclopridmixture. Since PBO acts in insects as an insecticide synergist byinhibiting the cytochrome P450-mediated metabolism, it wasanticipated that PBO addition may result in synergistic interac-tions with the two pesticide mixture.

2. Materials and methods

2.1. General experimental design

To serve the aim of examining the joint effects of the two known similarly

acting chemicals all studies were designed using the toxic unit (TU) concept

0.000.501.001.502.002.503.003.504.004.50

0.00Toxic Units of Imidacloprid

Toxi

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Imidacloprid mg l-1100 200 300 400 500 600 700 800

Fig. 1. Fixed ratio design used for Caenorhabditis elegans (left) and Eisenia fetida (right)

corresponding ratio label), with their equivalent doses of imidacloprid and thiacloprid

Please cite this article as: Gomez-Eyles, J.L., et al., Measuring andCaenorhabditis elegans and Eisenia fetida. Ecotoxicol. Environ. Saf. (2

(Sprague, 1970): where a TU is defined as the actual exposure concentration of a

chemical (c) divided by its EC50 (the median effect concentration):

TU ¼c

EC50. (1)

Firstly, range-finding experiments were designed for each organism to obtain

EC50 values for each pesticide and organism. Egg and cocoon production were

used as the measurement endpoints for the nematodes and earthworms,

respectively.

The mixture exposures were designed to cover the whole response surface

in the most efficient way possible. This was achieved using a fixed ratio design

(Fig. 1). The data analysis model of Jonker et al. (2005) used here analyses

the experimental data through simultaneous regression fitting of a response

surface to both the single compounds and mixture effect data. To analyse how best

to provide data for such regression analysis a deterministic power analysis

(Svendsen, unpublished) has indicated that as long as the full range of response

is covered within the exposure, then there is little difference in the statistical

power to detect response deviations from CA (or IA) between designs that include

a single replicate for a relatively large number of treatments or ones including

fewer replicated treatments. As sensitivities can drift between experiments, it

is important to ensure that the full range of response (i.e. reduction of

reproduction) is covered. This being the case, then the most robust experimental

designs for delivering data suitable for making statements about the shape of the

full surface are ones that favour maximising the range of exposure levels rather

than those with fewer replicated treatments. For this reason, the designs selected

for the tests with both species used the maximum number of treatments it was

possible to run each as a single replicate, except for the controls, which were

replicated to provide a measure of variability within the test system. Importantly

single compound exposures of each pesticide were carried out simultaneously

with the mixtures for both species, in order to avoid misinterpretations due to

sensitivity drift between experiments. The final design selected was based on a

fixed ratio that was set out to give a series of concentrations (1:0, 0:1) of each

pesticide and three series of binary mixtures of both pesticides at different ratios

(1:1, 1:3 and 3:1) (Fig. 1).

2.2. C. elegans exposures

Experiments were performed with C. elegans var. Bristol, strain N2 originally

obtained from the Caenorhabditis Genetic Centre, Minnesota, USA. Stock cultures

were kept sterile in the dark at 18 1C on nematode growth medium (NGM) agar

(Wood, 1988). Stock cultures were maintained by transferring nematodes weekly

onto fresh plates made up with NGM agar inoculated with 25ml of OP50 (a uracil-

requiring strain of Escherichia coli) and then left to grow overnight at room

temperature so as to provide a surplus of food. At 4 days before the exposure,

0

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mixture exposures. Concentrations in toxic units displayed in the top graphs (with

, in agar or soil, calculated using the TU concept displayed below.

modelling mixture toxicity of imidacloprid and thiacloprid on008), doi:10.1016/j.ecoenv.2008.07.006

ARTICLE IN PRESS

J.L. Gomez-Eyles et al. / Ecotoxicology and Environmental Safety ] (]]]]) ]]]–]]] 3

nematodes were transferred into fresh inoculated agar plates and left to incubate

at 18 1C to allow reproduction and growth of hatched nematodes to young adult

stage.

Fifteen concentration levels ranging from 0 to 4.0 TU were prepared at the five

different ratios, using a log factor of 1.4 between concentrations giving a total of 74

treatment and six controls (Fig. 1). EC50 values to calculate these concentrations

were obtained from a range-finder experiment. The same layout was used for a

second test that included the synergist PBO. In addition to the two pesticides, 6mM

of the synergist was also added to each treatment. This concentration of PBO was

known from range-finding exposures to be well below concentrations where any

direct effects were observed.

Exposures were made up using stock solutions of imidacloprid (Bayer

CropScience, Monheim am Rhein, Germany, 98% purity), thiacloprid (Bayer

CropScience, 98% purity) and PBO (Sigma Chemicals) prepared in ethanol. To

ensure that solvent concentration did not co-vary with pesticide concentration in

the experiment, the same total volume of ethanol was added to all test treatments

including all controls. Ethanol was chosen as the solvent carrier since previous

studies and our experience have shown that this solvent had little measurable

effect on C. elegans reproduction (as indicated through the rate of egg

development) at the concentration of 0.086 mmol ml�1 used in this assay

(Thompson and de Pomerai, 2005).

Exposure treatments were prepared by mixing the required amounts of

stock solutions and ‘‘make up’’ ethanol with 20 ml NGM agar. A volume of 7.5 ml of

each control or spiked NGM agar was then plated onto a 60 mm diameter petri-

plate (for pre-exposures) and multi-well plates were prepared by pipetting 2.5 ml

of each control or spiked NGM agar to labelled wells on 12 well multi-well tissue

culture plates. The 60 mm diameter petri-plates and multi-well plate wells were

then each inoculated with 10ml (in each petri-plate) and 1ml (in each well) of OP50

bacteria previously grown in 2�TY medium (Sigma Chemicals) and incubated for

48 h at room temperature to allow a bacterial lawn to develop.

Total exposure time was 48 h. Reproduction was measured only during the

second 24 h period to avoid counting eggs laid during the early phase of the

exposure when initial body burdens may have been low. Sufficient (424 for

controls and 44 for each test treatment) young adults were transferred into each

60 mm diameter pre-exposure plate for the first 24 h exposure at 18 1C. After the

first 24 h exposure, four worms were transferred into the same exposure

concentration in a well of the multi-well plate at 18 1C for a further 24 h period.

At the end of this second period, the number of worms surviving and the combined

total number of eggs laid and hatched juveniles in each well was counted to give a

measurement of reproduction rate.

2.3. E. fetida exposures

E. fetida were obtained from Blades Biological (Cowden, UK) and placed in

culture on a medium of 33% manure (obtained from a horse grazing on

uncontaminated pasture and not subject to any recent medication), 33%

composted bark (LBS Horticultural, Colne, UK) and 33% peat (LBS Horticultural).

Horse manure was topped up weekly as a food source for the worms.

The soil medium used for the test was prepared by mixing 2 mm sieved clay

loam soil (‘‘Kettering Loam’’ from Broughton Loam, Kettering, UK) with a pH of 7.1

and a 5% organic matter content, with 3% composted bark (on a dry weight basis).

The medium was mixed thoroughly and 1400 g dry weight added into plastic boxes

(170�170�80 mm). To guide in the appropriate selection of concentrations for

the mixture experiment, a preliminary exposure of 14 days was run. These single

compound exposures were designed considering previous experiments have

shown sublethal effects of imidacloprid on earthworms at concentrations of

0.5 mg kg�1 (Capowiez et al., 2005, 2006; Luo et al., 1999). Less data was available

for thiacloprid, so the same concentrations were used for this pesticide. To ensure

coverage across the full expected response range, 12 concentrations were prepared

with a maximum concentration of 20 mg kg�1, these being 0, 0.114, 0.182, 0.291,

0.466, 0.745, 1.19, 1.91, 3.05, 4.88, 7.81, 12.5 and 20.0 mg kg�1. The controls and the

concentrations of 0.291, 0.745, 1.91 and 4.88 were replicated five times, while all

other treatments only had single replicates.

EC50 values obtained from the range-finder experiment were used to design

the mixture experiment using the TU concept. Nine exposure levels ranging

between 0 and 6 TU were prepared at the five different ratios, using a log factor of

1.5 between concentrations (Fig. 1). Imidacloprid and thiacloprid stock solutions

prepared in ethanol were used for soil dosing. Make up volumes of ethanol were

used to ensure all treatments (including controls) received 16.6 ml of ethanol per

1.4 kg (dry wt.) soil. To ensure that the controls available would give a true

reflection of the variability in the system they were all used as solvent controls and

no pure controls were run. After addition of the stock solution, each replicate was

left uncovered in a dark fume cupboard for 72 h to ensure the solvent had

evaporated. After checking for removal of the solvent by checking for residual

wetting and smelling the soil, the soil was wetted to 60% of water holding capacity

and mixed to ensure a homogenous distribution of the pesticides. Horse manure

was dosed in the same way as the soil medium and then rewetted to 80% moisture

content, ready for weekly addition to each box as a food source (Spurgeon and

Hopkin, 1995).

Please cite this article as: Gomez-Eyles, J.L., et al., Measuring andCaenorhabditis elegans and Eisenia fetida. Ecotoxicol. Environ. Saf. (20

Fully mature earthworms were transferred from culture to the test boxes. Ten

worms per box were used for the single compound range-finder experiment and

eight for the mixture test. The total weight of the worms added to each replicate

was recorded. Six grams dry weight of the correspondingly dosed manure was

rewetted and then sprinkled onto the soil surface. The boxes were then covered to

prevent water loss and maintained at a constant 2071.5 1C under a 16:8 h

light:dark regime.

The mixture exposure was run for 21 days (Van Gestel et al., 1989). Manure

was removed and weighed weekly to provide a measure of feeding rate. Fresh

manure was then replaced. The number and weight of worms alive in each box was

also recorded, to determine survival rates and weight change relative to mean

initial weight. At the end of each test, soils were wet sieved and the number of

cocoons present counted to allow cocoon production rate (cocoon/worm/week) to

be calculated.

2.4. Data analysis

Single compound dose–response curves were fitted to the logistic model:

YðciÞ ¼Ymax

1þ ðci=EC50Þb , (2)

with the response Y(ci) being a function of the maximum response Ymax, the

exposure concentration (ci), the EC50 and the slope of the response curve (b), by

using the nonlinear fitting procedure in Genstat Release ver. 7 (Lawes Agricultural

Trust, Rothamsted Experimental Station), with extra code to calculate the

asymmetric EC50 95% confidence intervals. NOEC and LOEC values for the range-

finder experiment were calculated using one-way analysis of variance followed by

Tukey’s multiple comparison tests. An F-test was performed to test for significant

effects of PBO on the single compound concentration response relationships

calculated from the two imidacloprid–thiacloprid mixture experiments in

C. elegans.

Results from the mixture experiments were analysed using the MIXTOX model

of Jonker et al. (2005). This descriptive (rather than predictive) model offers a data

analysis framework that enables the detection and quantification of significant

deviations from either the CA or the IA models. To do this for CA, the fact that the

sum of TUs (Eq. (1)) of all chemicals should equal 1, where CA describes the data

can be rewritten as follows for binary mixtures:

c1

f�11 ðYÞ

þc2

f�12 ðYÞ

¼ expðGÞ, (3)

where c1 and c2 denote the concentrations of the individual chemicals in the

mixture, Y indicates the measured biological response, f�11 and f�1

2 indicate

the inverse dose–response functions (inverse of Eq. (2), so dependent on a joint

‘‘max’’ and individual EC50 and b values) for the individual compounds in the

mixture and G denotes an excess function to quantify deviations from the

CA model (Jonker et al., 2004). In cases where CA describes the data fully the

value of ‘‘exp(G)’’ should be 1, hence the value of the ‘‘deviation function’’ G

would be 0.

The procedure for modelling the data of an experiment of the design used here

following CA principles, relies on finding the values of the joint ‘‘max’’ and the

individual EC50 and b values that best allows the description of all data points in

the single compound response curves and the full mixture response surface where

G is 0 (for details see Jonker et al., 2005). This parameter fitting was done by

considering Eq. (3) for all data points simultaneously while minimising the sum of

the squared residuals (SS).

To assess if the CA model alone provided an adequate description of the

data, or there were systematic deviations where more or less severe effects

than should be anticipated from CA predictions, a stepwise approach was used.

In this, extra parameters that could describe biologically meaningful interactions

between the two chemicals were sequentially added to the deviation function

G. The first parameter added was a. This describes overall antagonistic (effect

less than predicted by CA) or synergistic (effect greater than predicted by CA)

effects. The second parameter added was either bi or iDL. These respectively

describe a ratio dependent (effect relative to CA dependent on the proportio-

nal contribution of each stressor) and effect level dependent (effect relative to CA

dependent on the magnitude of the response) deviations from CA.

Model with more parameters usually showed improved fits preventing direct

comparison between models. The sequential parameter addition used here,

however, creates a nested set of models which allows for testing the statistical

significance of the improvement in fit from the extra parameters. This significance

testing is completed by using the resulting SS for pairwise model comparison

through likelihood ratio testing at degrees of freedom equal to the difference in the

number of parameters in the two models through Chi-squared (w2) tests as

described by Jonker et al. (2005). The only two models that are not nested are the

ones for ratio- or effect level-dependence, hence these cannot be pairwise

compared using this approach. Numerical values calculated for the deviation

parameters can be interpreted using Table 1.

modelling mixture toxicity of imidacloprid and thiacloprid on08), doi:10.1016/j.ecoenv.2008.07.006

ARTICLE IN PRESS

Table 1Interpretation of additional parameters substituted into the concentration

addition (CA) reference model that define the functional form of the deviation

pattern

Parameter Value Interpretational meaning

Synergism/antagonism

a o0 Synergism

40 Antagonism

Ratio dependence

a o0 Synergism, except for those mixture ratios where significant

positive bi indicate antagonism

40 Antagonism, except for those mixture ratios where

significant negative bi indicate synergism

bi 40 Antagonism where the toxicity of the mixture is caused

mainly by toxicant i

o0 Synergism where the toxicity of the mixture is caused

mainly by toxicant i

Dose level dependence

a o0 Synergism at low concentrations and antagonism at high

40 Antagonism low concentrations and synergism at high

bDL 41 Change at concentrations lower than the EC50

¼ 1 Change at the EC50 concentration level

0obDLo1 Change at concentration levels higher than the EC50

Adapted from Jonker et al. (2005).

Table 2Dose–response curve parameters for Eq. (2)a (7S.E.) and for EC50 values (95%

confidence intervals) for the effect of imidacloprid and thiacloprid on the

reproduction of Caenorhabditis elegans in agar measured during the mixture

exposure in the absence of the PBO synergist and in a background of 6mM PBO

Exposure Parameter R2

Max b EC50 (mg l�1)

Imidacloprid No PBO 227(12.5) 2.43(1.38) 519(113) [329–820] 0.958

PBO 211(10.5) 0.95(0.44) 806(360) [315–2060] 0.978

Thiacloprid No PBO 223(11) 1.99(1.04) 289(68.1) [176–474] 0.968

PBO 204(7.0) 0.77(0.19) 325(91) [180–586] 0.991

a Eq. (2): Y(ci) ¼ Ymax/(1+(ci/EC50)b).

0

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100010010

No

of e

ggs

prod

uced

no PBOwith PBONo PBO - Reg.with PBO - Reg.

//0

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100010010

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of e

ggs

prod

uced

no PBOwith PBOno PBO - Reg.with PBO - Reg.

//0

Imidachloprid (mg l-1)

Thiachloprid (mg l-1)

Fig. 2. Effects of single compound exposures of (a) imidacloprid and (b)

thiacloprid both in the presence (filled circle, solid line) and absence (unfilled

circle, dashed line) of PBO. Lines for each concentration series represented a fitted

logistic model.

J.L. Gomez-Eyles et al. / Ecotoxicology and Environmental Safety ] (]]]]) ]]]–]]]4

3. Results

3.1. C. elegans exposures

Control worms produced an average of 49.8 eggs per individualduring the 24 h test. At the test temperature of 18 1C, this value isconsistent with expectation from previous life-history studies(Byerly et al., 1976) and previous ‘‘in house’’ studies in controlsprepared without solvent addition, indicating an acceptableperformance of control worms in the tests.

To initially validate that the logistic equation (Eq. (1)) wasapplicable for modelling the effects of imidacloprid and thiaclo-prid regressions were performed on the single compound dataobtained as part of the mixture experiment (i.e. disregarding themixture data for each compound) (Table 2). The logistic equationprovided a good fit to the data well in all cases (all R240.95,Table 2), with the only problem being that the upper asymptotefails to describe the apparently hormetic effects present atlow concentration of both pesticides in the exposure withoutPBO (Fig. 2).

Thiacloprid was more toxic than imidacloprid both with andwithout PBO present (Table 2). The sensitivity of C. elegans to bothpesticides was increased in the presence of PBO. This effect wasmost obvious at lower concentrations, with maximum effectsbeing largely unaffected (Fig. 2a, b). This did, however, result inEC50 and b (the slope) values decreasing for both pesticides in thepresence of PBO. Variation in the data was higher in the data fromthe exposures without PBO compared to with PBO. Due to thevariability in the data, an F-test comparison of the dose–responsecurves with and without PBO showed that these differences werenot statistically significant (p40.05) for either pesticide.

In the without PBO mixture test, fitting of the CA referencemodel to reproduction (combined number of eggs and hatchedjuveniles) data explained 51% of the variation in the data, but didnot provide the best description (Fig. 3). Adding the additionalparameter, a, to the deviation function G in Eq. (3) to describesynergism/antagonism reduced the sum of squared residuals andexplained 55% of the variation. Likelihood testing showed that theimprovement of this fit was significant (po0.05). The value of thisparameter a was �3.93, indicating that the effect observed was

Please cite this article as: Gomez-Eyles, J.L., et al., Measuring andCaenorhabditis elegans and Eisenia fetida. Ecotoxicol. Environ. Saf. (2

synergistic (see Table 1). Adding a second deviation parameter, bi

to describe ratio-dependent deviation did not improve thedescription of the data significantly (p ¼ 0.357). Including para-meter bDL, which describes a dose level-dependent deviation, didsignificantly improve the description of the data beyond thatprovide by the CA model with synergism/antagonism only(po0.005). In this case, the first additional parameter, a, had avalue of �5.96 and the second deviation parameter, bDL, had avalue of 1.05, with the model describing 60% of the variation in thedataset. The value of the first parameter suggests there issynergism at low concentration levels and antagonism at highconcentration levels, and the second parameter indicates the

modelling mixture toxicity of imidacloprid and thiacloprid on008), doi:10.1016/j.ecoenv.2008.07.006

ARTICLE IN PRESS

0

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30025020015010050 50 100 150 200 250 300

50 100 150 200 250 300 50 100 150 200 250 300

Fig. 3. Description of the effect of imidacloprid–thiacloprid on reproduction of Caenorhabditis elegans on agar. Observed data (number of eggs) related to modelled values

calculated using the MIXTOX model for (a) the CA reference model, (b) the synergistic/antagonistic model, (c) the dose ratio-dependent model and (d) the dose level-

dependent model. Note how the points distribute more equally around and draw closer to the 1:1 line representing agreement between data and model in (b) relative to (a)

and in (d) relative to (b), showing an improved description of the data by accounting for dose level-dependent deviation from the CA model.

J.L. Gomez-Eyles et al. / Ecotoxicology and Environmental Safety ] (]]]]) ]]]–]]] 5

switch from synergism to antagonism occurs at a concentrationlevel equal to 1/bDL ( ¼ 1/1.05 ¼ 0.95) of the EC50 (Table 1). TheMIXTOX model therefore suggests that a dose level-dependentdeviation from the CA model describes the data better than the CAmodel, with the major effect being synergism at low exposureconcentrations. This means that once the strength of the mixturedrops below 37% of the EC50 level the reduction in reproductionwould be more than two-fold that predicted by CA. In effect termsthis means that at a mixture of 37% of the EC50’s that waspredicted to cause a 19% reduction in reproduction would actuallycause a 38% reduction, with the under prediction of effects by CAincreasing at lower concentrations.

Analysis of the response surface for the binary mixturesexperiment with PBO included using the MIXTOX model indicatedthat the CA reference model described the effect of theimidacloprid–thiacloprid–PBO mixture better than in the absenceof PBO (Fig. 4), with 62% of the total variation described.Again, however, adding the additional deviation parameter, a, tothe G-element of Eq. (2) to describe synergism/antagonismimproved the description of the data explaining 68% of totalvariation (Fig. 4). Parameter a had a value of �5.20, again indi-cating synergism. Adding the bi parameter for ratio-dependentdeviation did not improve on this description (p ¼ 0.322).Including the parameter, bDL, to describe a dose level-dependentdeviation significantly (po0.005) improved the description of thedata (Fig. 4). In this case parameter a had a value of �7.33 andparameter, bDL a value of 0.908, with the model describing 72% oftotal variation. As in the test without PBO, the value of the firsttwo parameters suggests there is synergism at low dose levels andantagonism at high dose levels, but this time the secondparameter indicates the switch from synergism to antagonismoccurs at a dose level equal to 1/bDL ( ¼ 1/0.908 ¼ 1.1) times the

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EC50 (see Table 1). This means that once the strength of themixture drops below 45% of the EC50 level the reduction inreproduction would be more than 1.5-fold that predicted by CA. Ineffect terms this means that at a mixture of 45% of the EC50’s thatwas predicted to cause a 32% reduction in reproduction wouldactually cause a 48% reduction, with the level of under predictionof effects by CA increasing at lower concentrations.

3.2. E. fetida exposures

Single compound toxicities from the initial range-finderexperiments are given in Table 3 summarising all EC50, NOECand LOEC values. Only imidacloprid had a significant effect onE. fetida survival (po0.001) and after 2 weeks exposure, therewas significant (po0.001) mortality of worms at 1.91 mg kg�1 andabove (LC50 ¼ 2.36 mg kg�1, S.E. ¼ 0.18 mg kg�1), while there wasno significant effect (p40.05) of thiacloprid on E. fetida survivaleven at the highest concentration tested (20 mg kg�1). Bothimidacloprid and thiacloprid had a significant effect on percentageweight change (po0.001). Weight change data for concentrationsof imidacloprid above 1.91 mg kg�1 was not included in theanalysis as too few worms survived to permit reliable estima-tion. Calculated EC50 values for weight change were 2.77 and19 mg kg�1 for imidacloprid and thiacloprid, respectively.

Increasing concentrations of either imidacloprid or thiaclopridalso caused a significant decrease in the amount of manure eatenby the earthworms during the exposure (po0.001), with the effectof imidacloprid being significant at lower concentrations (1.91 mgkg�1, po0.05) than for thiacloprid (4.88 mg kg�1, po0.001). Bothcompounds had a highly significant effect on cocoon production(po0.001). Cocoon production in E. fetida was more sensitive to

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50 100 150 200 250 300 50 100 150 200 250 300

Fig. 4. Description of the effect of imidacloprid–thiacloprid–PBO on the egg production of Caenorhabditis elegans on agar. Observed data (number of eggs) related to

modelled values calculated using the MIXTOX model for (a) the CA reference model, (b) the synergistic/antagonistic model, (c) the dose ratio-dependent model and (d) the

dose level-dependent model. Note how the points distribute more equally around and draw closer to the 1:1 line representing agreement between data and model in (b)

relative to (a) and in (d) relative to (b), showing an improved description of the data by accounting for dose level-dependent deviation from the CA model.

Table 3EC50 (with 95% confidence intervals where calculable), NOEC and LOEC values for the effects of imidacloprid and thiacloprid on the earthworm Eisenia fetida in a Kettering

loam soil for three different endpoints: cocoon production, weight change and manure eaten

Endpoint Imidacloprid EC50

(mg kg�1)

Imidacloprid LOEC

(mg kg�1)

Imidacloprid NOEC

(mg kg�1)

Thiacloprid EC50

(mg kg�1)

Thiacloprid LOEC

(mg kg�1)

Thiacloprid NOEC

(mg kg�1)

Cocoon

production

1.41 (1.08–1.85) 1.91 0.745 0.968 (0.625–1.50) 0.291 o0.291

Weight change 2.77 (2.01–3.82) 1.91 0.745 19.0 (13.8–26.3) 1.91 0.745

Manure eaten 1.88 1.91 0.745 1.64 (1.08–2.50) 4.88 1.91

J.L. Gomez-Eyles et al. / Ecotoxicology and Environmental Safety ] (]]]]) ]]]–]]]6

thiacloprid than imidacloprid. Thiacloprid caused a significantdecrease in cocoon production at 0.291 mg kg�1 (po0.05) andimidacloprid at 1.91 mg kg�1 (po0.05). In all cases the logisticequation (Eq. (2)) was well suited for describing the dose–response relationship in all the data sets, with the worst fit beingfor the thiacloprid manure removal data, which still had an R2-value of 0.897.

Data for the effects of imidacloprid and thiacloprid and theircombination on all endpoints from the mixture test were fittedagainst CA using the MIXTOX model. Comparison of the referenceand deviation models for cocoon production indicated that CAdescribes 61% of total variation and that inclusion of theindividual deviation parameters from the CA reference modeldid not provide a significantly better description of the data(Table 4, Fig. 5). The same was the case when considering theamount of manure eaten as an endpoint, with CA accountingfor 76%, 70% and 50% of variation in weeks 1, 2 and 3, respectively.For both of these endpoints, the joint effects of imidaclopridand thiacloprid are, therefore, consistent with CA withoutinteraction.

For weight change, a dose ratio-dependent deviation provideda better description of the data than the CA reference model alone

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(Table 4). The description of the data was significantly (po0.001)improved from describing 43% to describing 61% of total variationwhen adding parameters a and bi to describe a dose ratio-dependent deviation using weight change after 3 weeks as theendpoint. The a and bi parameters took values of 9.45 and �24.34,respectively. These indicate that the mixture interaction isantagonistic when thiacloprid is the main component of themixture and synergistic when imidacloprid is the main compo-nent (in terms of TU) (Table 1). The switch from synergism toantagonism occurs when imidacloprid accounts for more than12% of the toxicity of the mixture. This same dose ratio-dependentdeviation on weight change was also significant when weightchange data that was collected at 1 and 2 weeks exposurewas modelled (po0.01). The descriptions improved from explain-ing 53% to 64% and from 50% to 60% of the total variation forthe weeks 1 and 2 datasets, respectively. The a and bi parameterstook values of 4.19 and �9.88 for week 1 and 10.7 and �20.7 forweek 2. This indicates that the switch from antagonism tosynergism occurs when imidacloprid caused more than 19% and18% of the mixtures toxicity in weeks 1 and 2, respectively.Further, the maximum difference between the CA reference modeland the ratio-dependent pattern observed would occur where

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Table 4EC50 values (with 95% confidence intervals where calculable) for the effect of imidacloprid and thiacloprid on Eisenia fetida measured during the mixture exposure in a

Kettering loam soil, and significance of additional parameters used to describe deviations from the CA model (S/A: synergistic/antagonistic deviation; DR: dose ratio-

dependent deviation; and DL: dose level-dependent deviation)

Endpoint Imidacloprid EC50 Thiacloprid EC50 S/A p(w2) DR p(w2) DL p(w2)

Cocoon production (3 weeks) 3.23 (1.80–5.81) 4.21 (0.702–25.2) 0.24 0.347 0.37

Manure eaten (1 week) 2.14 (1.28–3.59) 1.82 (0.901–3.68) 0.792 0.904 0.2

Manure eaten (2 weeks) 3.72 (2.35–5.88) 11.2 (1.42–87.9) 0.156 0.26 0.15

Manure eaten (3 weeks) 6.64 (1.87–23.6) 6.77 (1.36–34.3) 0.732 0.628 0.124

Weight change (1 week) 14.1 (6.79–29.1) 45.65 0.415 7�10�4 0.532

Weight change (2 weeks) 13.0 (4.45–38.2) 45.65 0.465 0.003 0.51

Weight change (3 weeks) 54.0 (0.847–3400) 45.65 0.818 9�10�5 0.268

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0.5 1 1.5 2 0.5 1 1.5 2

Fig. 5. Description of the effect of imidacloprid–thiacloprid on the cocoon production of Eisenia fetida in Kettering loam soil. Observed data (number of cocoons/worm/

week) related to modelled values calculated using the MIXTOX model for (a) the CA model, (b) the synergistic/antagonistic model, (c) the dose ratio-dependent model and

(d) the dose level-dependent model. Note that the points in (b), (c) or (d) do not distribute more equally around or draw closer to the 1:1 line representing agreement

between data and model relative to (a), indicating no improvement in the description of the data by addition of further deviation parameters to the CA model.

J.L. Gomez-Eyles et al. / Ecotoxicology and Environmental Safety ] (]]]]) ]]]–]]] 7

imidacloprid caused 85%, 87% and 91% of the total mixturetoxicity, and that the real weight losses here would be 1.39, 1.51and 1.72 times higher than CA predicted for weeks 1, 2 and 3.

4. Discussion

The effect of imidacloprid and thiacloprid on C. elegans

reproduction showed near identical dose level-dependent devia-tion from the CA reference model both with and without PBOpresent. A greater than predicted toxicity of the mixture wasfound at low concentration combinations, whilst a lower thanpredicted toxicity was found at higher concentrations. The underpredictions of effects at lower concentrations were 1.5–2-fold atexposure levels expected to cause 20–30% reductions in reproduc-tion. The fact that the same deviation pattern was found in both

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studies supports a conclusion that the response seen was notaccounted for by chance deviation in the data, but may represent areal difference in biological response to different combinedexposure levels imidacloprid and thiacloprid (with or withoutPBO and in an overall background of low ethanol as a result ofsolvent addition).

Regarding first the effects of PBO, here little influence onC. elegans sensitivity to the two pesticides was found. This is incontrast to the clear synergistic effect of PBO that have beenshown in co-exposures with neonicotinoids in insects. In thehoney bee (Apis mellifera) for example, PBO was found to increasethiacloprid (but not imidacloprid) toxicity by at least 150-fold(Iwasa et al., 2004), while PBO has also been found to suppress thedevelopment of neonicotinoid resistance associated with en-hancement of oxidative detoxification in the Colorado beetle

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(Leptinotarsa decemlineata) (Mota-Sanchez et al., 2006). That nosynergism was observed for either pesticide for C. elegans suggeststhat there may be difference between the nature of the enzymesinvolved in neonicotinoid metabolism between insects andnematodes. While it is possible that C. elegans may use otherdetoxification mechanisms rather than relying solely on cyto-chrome P450s when exposed to neonicotinoids, induction of theP450 detoxification mechanism has been shown for other organicxenobiotics in C. elegans (Menzel et al., 2001). This, therefore,suggest that PBO acts as only a poor inhibitor of the cytochromeP450 isoforms responsible for the detoxification of both neonico-tinoids in C. elegans. Confirmation of this, however, requiresextensive further work.

Regarding the joint effect of the two pesticides, to date manystudies with similar acting chemicals have shown a goodcapability of the CA model for predicting the joint effects ofsimilarly acting chemicals. The majority of these studies havebeen focused on mixtures of compounds which work through thenon-specific mode of action known as ‘‘narcotic’’ or ‘‘baseline’’toxicity (Altenburger et al., 2003). These have included mixtureswith just two and up to more than 50 compounds (Hermens et al.,1984a, b). Data for mixtures of similar specifically acting chemi-cals is more limited. In the majority of such studies, the mixtureeffective most often approximates to additivity (Lydy et al., 2004),although most only checked if CA provides a reasonable descrip-tion of observed data and do not include a significance testingaspect as is done here.

There are cases where interactions are found to occur that aremanifest by deviation in joint effects from model predictions forhigher organisation level endpoints (e.g. reproduction), this ismost often when the mixture is either simple, is dominated by afew compounds or is comprised of compounds with very specificmodes of action (Jonker et al., 2004; McCarty and Borgert, 2006).The work conducted here for imidacloprid and thiacloprid adds,therefore, to this list of studies that have shown interactions insimilar specifically acting simple mixtures.

Because deviations beyond overall synergism and antagonismare rarely studied, theories regarding their occurrence are not wellestablished except in a few cases, such as that between pyrethroidand carbamate insecticides where synergism has been attributedto a cascade of molecular events creating a negative feedback ofacetylcholinesterase release (Corbel et al., 2006). In a recent studywith a mixture of 14 nitrobenzenes, it was established that actualtoxicity exceeded the effect predicted by CA. This led the authorsto propose that not all nitrobenzenes follow the same mode ofaction and, therefore, that joint effects may actually be attribu-table to non-interactive IA (Altenburger et al., 2005). The mixtureinteractions between imidacloprid and thiacloprid in C. elegans

found here (both in the presence and absence of PBO) couldsuggest that the combined effect of these pesticides may occurthrough different modes of action. Studies on the binding sites ofneonicotinoids in insects have show that they appear to bind atthe [3H]imidacloprid binding site in the same manner for eachcompound and species examined (Zhang et al., 2000). Thismechanistic evidence seems to confirm selection of CA as themost appropriate model, with interactions occurring that influ-ence pesticide toxicokinetics or toxicodynamics at differentexposure levels.

For the earthworms, the LC50 of 2.36 mg kg�1 dry soil foundhere for imidacloprid corresponded well with a 14-day study onE. fetida where an LC50 of 2.3 mg kg�1 dry soil was found (Luoet al., 1999). Sublethal effects (sperm deformities) on E. fetida havebeen documented at imidacloprid concentrations above 0.5 mgkg�1 dry soil (Capowiez et al., 2006; Luo et al., 1999) and othersignificant sublethal effects of imidacloprid (weight loss andreduced burrowing activity) have been found at 0.5–1.0 mg kg�1

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dry soil in Allolobophora icterica and Aporrectodea nocturna

(Capowiez et al., 2005, 2006). These results are consistent withthe LOEC and NOEC values of 1.91 and 0.745 mg kg�1 obtainedhere for cocoon production, weight change and manure con-sumption.

There are few studies to date that have measured the effects ofthiacloprid of earthworms. A 14-day laboratory acute toxicity teston E. fetida indicated an LC50 above 50 mg kg�1 (Schmuck, 2001).As there was no significant mortality at the highest concentrationof 20 mg kg�1, the results of the current study are not incompa-tible. The EC50, LOEC and NOEC found here of usually less than2 mg kg�1 also supports the NOEC values of below 1.0 mg kg�1

found by Schmuck (2001).Although the only variation between the range-finder and the

mixture experiment was the duration of the exposure and thenumber of worms used per box, the EC50 values varied by a factorof 2–3 between the two experiments (and more so for adultweight change). Such variability is not uncommon and can be afeature even of standardised tests (Weyers et al., 2002). While notunexpected, it does highlight the need to always have the singlecompound exposures within a mixture experiment and to usethese EC50 values when describing the data, as values obtainedfrom previous experiments may skew the fit.

The E. fetida mixture experiment was designed using the TUconcept based on range-finder cocoon production EC50 values. Formeasured cocoon production, the CA model provides a good fit tothe data from the mixture experiment. Further the fit of the modelto the data was not significantly improved by addition of furtherparameters. As well as effects on cocoon production, thecombination of single chemical and mixture exposures usedcould be used to assess if the CA model could describe mixtureeffects on other endpoints. For the feeding rate endpoint at weeks1, 2 and 3, the CA model described the mixture effects and thiswas not improved by the addition of deviation parameters. Forweight change from initial at week 3, on the other hand, a doseratio-dependent deviation from the CA model was observed, witha lower combined toxicity associated with mixtures with a highproportion of thiacloprid in the mixture and an increased toxicityeffect to a high proportion of imidacloprid. As discussed abovefor the deviation seen for C. elegans, it is hard to relate theseobservations for imidacloprid and thiacloprid to any underlyingphysiological mechanism. It should be noted that the dose–response of this endpoint for thiacloprid was relatively flat andthis may have had an effect on the fitting of the data. Thus, theinteraction highlighted by the model can be considered less‘‘reliable’’ than the deviations seen in the two C. elegans

experiment, as the full response range for this parameter is notcovered.

As previously emphasised, use of the CA model to predict themixture toxicity of similarly acting chemicals is well establishedin (eco)toxicology (Altenburger et al., 2000). In this study, CAprovided a sound prediction of the joint effects of imidaclopridand thiacloprid for reproduction and feeding rate, but not adultweight change, in E. fetida, or in reproduction in C. elegans—wherean interaction was found. This suggests there may be bothendpoint and species specific responses to simple mixtures ofsimilarly acting chemicals.

When any study reveals apparent synergism between com-pounds (or in this case partial synergism), a question arises as tothe mechanisms that cause such effects. This is all the more truewhen the same chemical mixture is found to show non-interactive additivity in a different species. Present understandingof the toxicokinetics and toxicodynamics of imidacloprid andthiacloprid in nematodes and earthworms is insufficient to makeany firm conclusions as to why consistent deviations are found inthe experiment with C. elegans but not E. fetida. A number of

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factors, however, could be important for determining the betweenspecies differences observed.

Firstly, the two tests are conducted in vastly different media.C. elegans was tested in dosed agar, while earthworms were testedin soil. Since soil is a more complex media, it might be expectedfor interactions to be more likely here. That this was not the casesuggests that media characteristics alone may not account for theobserved difference. A second issue that could explain the speciesdifference is the fact that residual solvent ethanol remained in theC. elegans test (since solvent venting was not possible), but not inthe earthworm assay. Ethanol upregulates many response systems(e.g. heat shock, phases 1 and 2 metabolism) that are associatedwith responses to pesticide exposure (Kwon et al., 2004). Thismeans that we cannot exclude the possibility that ethanolmodulates the response between the chemicals leading to anapparent interaction in the C. elegans study. Finally and ratherobviously, the two species, while both soil dwelling, come fromseparate phyla located within different branches of the animalkingdom (Philippe et al., 2005). This means that the toxicokineticsand toxicodynamics of the two pesticides that govern internalexposure levels and target site interactions may ultimatelyprovide a true biological basis for the observed differencebetween the species.

While the majority of the literature indicates a high predictivepower of CA, the are also many other studies where CA (or IA)have over- or under-predicted actual mixture toxicity (Forgetet al., 1999; Kortenkamp and Altenburger, 1999; Silva et al., 2002).This indicates that the prediction of the CA model should not beconsidered definitive, but should be considered within a prob-abilistic context. Thus, by drawing together investigations like thisone conducted with different compounds and species, thedistribution of actual mixture effects against CA predictions canbe estimated. Incorporating these probabilities into ecological riskassessment using probabilistic theory may allow confidenceintervals to be provided to CA predictions that can support amore informed and robust risk assessment.

Acknowledgments

We would like to thank Peter Rothery at CEH Monks Wood forvaluable help with writing the Genstat code. The research wasfinancially supported by the European Union project ‘‘NoMiracle’’(European Commission, FP6 Contract no. 003956) and theUniversity of Reading.

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