370

Environmental Toxicology and Chemistry, Vol. 26, No. 2, pp. 370–379, 2007� 2007 SETAC

Printed in the USA0730-7268/07 $12.00 � .00

COMBINED USE OF BIOMARKERS AND IN SITU BIOASSAYS IN DAPHNIA MAGNATO MONITOR ENVIRONMENTAL HAZARDS OF PESTICIDES IN THE FIELD

CARLOS BARATA,*† JOANA DAMASIO,† MIGUEL ANGEL LOPEZ,‡ MARINA KUSTER,§ MARIA LOPEZ DE ALDA,§DAMIA BARCELO,§ MARI CARMEN RIVA,† and DEMETRIO RALDUA†

†Laboratory of Environmental Toxicology, Universitat Politecnica de Catalunya, Carretera nacional 150 Km 14.5, Terrassa 08220, Spain‡Aquatic Ecosystems Unit, Institut de Recerca & Tecnologia Agroalimentaries, San Carles de la Rapita Tarragona, Spain

§Department of Environmental Chemistry, Institud d’ Investigacions Quımiques i Ambientals de Barcelona,Consejo Superior de Investigaciones Cientıficas, Jordi Girona 18, 08034 Barcelona, Spain

(Received 24 April 2006; Accepted 25 September 2006)

Abstract—The aim of this investigation was to evaluate toxicity effects of pesticides in aquatic invertebrates using in situ bioassayswith the local species, Daphnia magna. Investigations were carried out in the Delta del Ebro (northeast Spain) during the maingrowing season of rice (May–August). Measures of energy consumption (i.e., algal grazing) and of specific biochemical responses(biomarkers) were conducted in individuals transplanted at four stations, including a clean site upstream of the affected area andthe three main channels that collect and drain the water from the rice fields into the sea. Seventeen pesticides were analyzed inwater by on-line solid-phase extraction–liquid chromatography–tandem mass spectrometry. The results obtained indicated highlevels of pesticides in water, with peak values of 487 �g/L for bentazone, 8 �g/L for methyl-4-chlorophenoxyacetic acid, 5 �g/Lfor propanil, 0.8 �g/L for molinate, and 0.7 �g/L for fenitrothion. Measured biological responses denoted severe effects on grazingrates; a strong inhibition of cholinesterases and carboxylesterases, which are specific biomarkers of organophosphorous and carbamatepesticides; and altered patterns of the antioxidant enzyme catalase and the phase II metabolizing enzyme glutathione S-transferase.Correlation analysis with pesticide residue levels converted to toxic units relative to its acute 48-h median lethal concentration ofD. magna indicated significant and negative coefficients between the dominant pesticide residues and the observed biologicalresponse, thus denoting a clear cause-and-effect relationship. The results emphasize the importance of considering specific (bio-markers) as well as more generalized and ecologically related (grazing) in situ responses to identify and evaluate biological effectsof environmental contaminants in the field.

Keywords—Pesticide Daphnia Feeding Biomarker In situ

INTRODUCTION

The need for ecologically relevant methods in environ-mental risk assessment has led to the development of newbioassays with caged, single species to determine pollutanteffects in situ. Key advantages of in situ bioassays comparedwith whole-effluent toxicity tests and biological surveys ofbenthic macroinvertebrate communities include a greater rel-evance to the natural situation, especially with respect to thecontamination scenario, and their ability to detect effects incaged individuals more rapidly (hours to days) than resultingchanges in community structure (months to years) measuredduring macroinvertebrate sampling [1]. Recently, the use ofin situ and cost-effective bioassays based on feeding responseshave allowed the detection of lethal and sublethal responsesthat are biologically linked with important life functions andkey ecological processes, such as detritus processing and algalgrazing rates [1–3]. Unfortunately, in most studies, in situbioassays have been focused on relating general toxicologicalresponses of caged organisms with environmental factors (i.e.,major contaminant sources) and/or responses of the same spe-cies (or whole community) in the field [1,3–5]. Thus, most insitu bioassays are still unable to identify toxic componentswithin complex mixtures [2].

Although aquatic animals invariably are exposed to con-taminant mixtures, only a small number of contaminants may

* To whom correspondence may be addressed(barata@intexter.upc.edu).

be responsible for the observed toxicity. The integrated use ofa set of biochemical biomarker responses to pollutants is con-sidered to be one of the best procedures for detecting impactsof specific contaminants in situ [6]. Therefore, by combiningseveral biochemical biomarkers in caged organisms to indicateexposure to or effects associated with specific stressors andtoxicological responses to indicate ecological impacts, itshould be possible to identify toxic components within com-plex chemical mixtures. Despite the potential benefits of com-bining biomarker and toxicological responses into a single insitu bioassay, their use in biomonitoring water-quality requiresvalidation in laboratory and field studies.

Recent studies conducted by McWilliam and Baird [3,7]demonstrated that in situ postexposure feeding depression inDaphnia magna was a sensitive, robust, and ecologically rel-evant end point to diagnose sublethal effects of toxic effluentsthat are rich in metals and agrochemicals. Laboratory studiesalso have shown that it is possible to identify specific pollutantscausing anticholinergic, oxidative stress, and phase II enzymeactivity responses in D. magna by using specific biomarkers[8,9]. Among them, B-esterases, such as acetylcholinesterase(AChE; EC 3.1.1.7) and carboxylesterase (CbE) activities,were found to be inhibited by organophosphorous and car-bamate pesticides in a concentration-related manner [9].Acethylcholinesterases are considered to be the target site oforganophosphorous and carbamates pesticides, and in D. mag-na, their persistent inhibition causes neurotoxic effects [9].Carboxylesterases are known to hydrolyze a wide range of

Daphnia magna in situ responses to pesticides Environ. Toxicol. Chem. 26, 2007 371

Fig. 1. Studied Spanish sites showing the location of the Ebro Riverbasin (A), that of Delta del Ebro with site 1 (B), and that of stations2 to 4 in the south hemidelta (C).

exogenous and endogenous esters, and they are assumed toprotect against poisoning with organophosphorous and car-bamate pesticides [10]. In D. magna, CbE was found to bemore sensitive then AChE to anticholinergic pesticides [9].Studying the mechanisms of toxicity of several redox-cyclingcompounds in D. magna, Barata et al. [8] found that the an-tioxidant enzyme catalase (CAT; EC 1.11.1.6), which reducesH2O2 to water, and the phase II enzyme glutathione S-trans-ferase (GST; EC 2.5.1.18), which catalyzes the conjugation ofglutathione with various electrophilic substances [11], werethe most responsive enzymes analyzed.

Here, we developed further the D. magna postexposure insitu assay combining toxicological (survival and feeding in-hibition) and biochemical responses (biomarkers) to evaluateenvironmental hazards and to identify toxic pesticides asso-ciated mainly with rice farming in the Delta del Ebro (northeastSpain). This is an area of 320 km2 in the northwest of theMediterranean Sea that holds the ‘‘Reserva del Delta del Ebro’’(areas protected for their plants and animal species by Direc-tive 79/409 of the European Commission) and the naturalaquatic resources of which receive large amounts of pesticidesbecause of the culture of rice [12]. Moreover, populations ofD. magna and other cladocerans occur naturally in the ricefields and in other, more permanent freshwater and brackishlagoons along the Delta del Ebro [13]; thus, measured re-sponses of caged organisms may reflect those responses oc-curring in natural populations. The biomarkers chosen includedthe B-esterases AChE and CbE, the antioxidant enzyme CAT,and the metabolizing enzyme GST, which are known to bealtered by pesticides and redox-cycling chemical contami-nants. In situ bioassays were complemented with a detailedchemical analysis of pesticide residues and other general water-quality variables to allow the identification of potential causalabiotic factors.

MATERIALS AND METHODS

Study site and pesticide use

The present study was conducted in the Delta del Ebro fromMay 2005 until August 2005, which includes the period duringwhich more pesticides are used. The Delta del Ebro is locatedat the end of the largest river in Spain, and it holds 21,600 haof rice fields, producing 113,500 tons of rice per year. It hasa triangular shape, going some 20 km into the sea, with onecoastal split in the north and one in the south that enclose twoshallow bays of capital importance for fish and mussel pro-duction [12]. The geological and hydrological dynamics of theDelta del Ebro have been altered for rice farming by meansof upstream barrages and canalization through the delta. Twomain channels, one on each side of the river, bring the waterfrom Xerta, some 30 km upstream, to the rice cultivation sys-tem. From these two channels, water is carried to and collectedfrom the rice fields by a network of irrigation and drainageditches, respectively. Three big drainage channels in either siteof the hemidelta collect all water and drive it straight to thesea. Rice farming involves the use of more than 25 agrochem-icals to treat plagues, with triazines, phenylureas, anilines, andorganophosphates being the compounds of broader use [12].The selected sites were located at the ends of three of the maindrainage channels (sites 2–4) (Fig. 1) that collect and transportthe water from the southern hemidelta rice fields into one ofthe sea lagoons. An additional site located upstream of theDelta del Ebro just before the collection of water into the mainchannel that irrigates the rice fields was selected as a surrogate

control (site 1) (Fig. 1). By selecting the main drainage chan-nels instead of individual rice fields, it was possible not onlyto detect but also to evaluate the effects of most pesticideresidues that are applied in the rice fields.

Physicochemical parameters

The following physicochemical parameters were measuredat each site at the start and the end of the exposure period:Flow rate, temperature, pH, conductivity, oxygen levels, andsuspended solids. Briefly, water flow rates were determinedwith a Mini Air prove (Technika-Schiltknecht, Phoenix, AZ,USA), and temperature, pH, conductivity, and dissolved ox-ygen concentration were measured in situ using a WTW Multi340i handheld meter (Wissenschaftlich-Technische Werkstat-ten [GmbH], Weilheim, Germany). Total suspended solidswere measured in the laboratory according to standard methods[14].

At the beginning of the exposures, water samples (3 L)were collected and stored at 4�C to determine pesticide residuelevels in water and to conduct whole-effluent tests. Chemicalanalyses were conducted within 24 h of water collection andrestricted to the most abundant 17 pesticides. These pesticideincluded organophosphates (fenitrothion, malathion, and dia-zinon), triazines (simazine, atrazine, and cyanazine), phe-

372 Environ. Toxicol. Chem. 26, 2007 C. Barata et al.

nylureas (chlortoluron, isoproturon, and diuron), an anilide(propanil), chloroacetanilides (alachlor and metalochlor), acid-ic herbicides (bentazone, 2-methyl-4-chlorophenoxyaceticacid [MCPA], 2,4-dichlorophenoxyacetic acid [2,4-D], andmecoprop), and thiocarbamates (molinate).

Analyses were performed using a previously describedmethod [15], which was based on on-line solid-phase extrac-tion–liquid chromatography–tandem mass spectrometry andadapted to include within the list of target compounds theinsecticides fenitrothion and malathion, which are known tobe used in the study area. In this method, fully automated, on-line trace enrichment was performed by passing 10 ml (flowrate, 8 ml/min) of the water samples (previously passedthrough 0.45-�m filters to remove suspended matter), the aque-ous standards, and the blanks through HySphere Resin GPpolymeric cartridges (length, 10 mm; inner diameter, 2 mm;particle size, 10–12 �m; Spark Holland, Emmen, The Neth-erlands) previously conditioned with 4 ml of acetonitrile and4 ml of liquid chromatography–grade water (flow rate, 4 ml/min). After sample loading, the cartridges were washed with1 ml of liquid chromatography–grade water at a flow rate of5 ml/min to complete transfer of the sample and to removematrix interferences. Elution of the trapped compounds di-rectly to the liquid chromatographic column was performedwith the mobile phase.

Chromatographic analysis was carried out in a reversed-phase Purospher STAR-RP-18e analytical column (length, 125mm; inner diameter, 2 mm; particle diameter, 5 �m; Merck,Darmstadt, Germany) using as mobile-phase gradient aceto-nitrile/water at a flow rate of 0.2 ml/min. Detection was per-formed with an electrospray ionization source under time-scheduled conditions, recording two different selected reactionmonitoring transitions (one for quantification and another forconfirmation) per compound [15]. Identification was accom-plished by comparing the retention time and the liquid chro-matography–tandem mass spectrometry signals of the targetcompounds in the samples with those of standards analyzedunder the same conditions, and quantification was performedby the external standard method.

The limits of detection and limits of quantitation of themethod were experimentally estimated from the on-line anal-ysis of spiked liquid chromatography–grade water samples asthe minimum concentration of analyte giving a signal to noiseratio of three and eight, respectively. The limits of detectionachieved were less than 5 ng/L except for metalochlor (12ng/L), alachlor (18 ng/L), malathion (39 ng/L), and fenitro-thion (50 ng/L). The limits of quantification were within therange of 0.8 to 98 ng/L except for fenitrothion (130 ng/L).

Test organisms: Culture conditions and acclimation

A single clone of D. magna was maintained in bulk culturesas described by Barata et al. [9]. Briefly, 10 adults were cul-tured in 2 L of American Society for Testing and Materials(ASTM) hard water [14] with the addition of a standard organicextract. Animals were fed daily with Chlorella vulgaris Bei-jerink (5 � 105 cells/ml, corresponding to 1.8 mg C/L). Pho-toperiod was set to 14:10-h light:dark and temperature to 20� 1�C.

Animals used for bioassay deployments were cultured for4 to 5 d until they reached their fourth instar.

In situ bioassays

Exposure regime. In situ D. magna deployments were con-ducted using the same test chambers and procedures described

by McWilliam and Baird [3] with only minor modifications,which included the use of 9 to 10 test chambers to allowcollection of animals for biomarker determination and to in-crease the number of replicates for measurements of postex-posure feeding rate. Chambers were constructed from clearpolyvinyl chloride cylindrical piping (length, 13 cm; externaldiameter, 5 cm). Each chamber had two rectangular windows(7 � 3.5 cm) cut into either side of the cage and covered with150-�m nylon mesh. Pipe ends were sealed with polypropyl-ene caps. Groups of four to five chambers were placed insidea 13-mm2, wire-mesh cylinder that was positioned in the streamperpendicular to flow. In each deployment, a control treatmentwith animals maintained in the laboratory and never exposedto the field also was included as a surrogate control. Deploy-ments were conducted simultaneously in four locations on fourdates (May 16th, June 20th, July 18th, and August 26th, 2005,hereafter referred to May, June, July, and August, respec-tively).

Briefly, the procedure for the in situ bioassays was as fol-lows. Juveniles were transported to field sites in groups of 10in 175 glass jars filled with ASTM hard water [3,14]. All fieldsites were less than a 3-h drive from the laboratory. At eachsite, five to seven chambers containing 10 individuals eachand four chambers containing 20 individuals each were placedinside a 13-mm2, wire-mesh cylinder that was positioned inthe stream perpendicular to flow.

Postexposure responses. After 24 h, animals were retrievedfrom the chambers. Surviving animals from those chambersholding groups of 20 individuals were pooled in an Eppendorftube, immediately frozen in liquid N2, and kept at �80�C untilfurther enzyme analysis. Shortly after exposure (within 1 h),five surviving juveniles from those chambers holding groupsof 10 animals were placed into 60-ml, screw-capped, glass jarscontaining 50 ml of ASTM hard water with C. vulgaris Bei-jerink (strain CCAP C211/12) at a concentration of 5 � 105

cells/ml and allowed to feed for 4 h [3]. Three jars containingno animals were used to establish initial algal densities. Bio-marker and postexposure feeding rates also were measured inanimals maintained in the laboratory for the duration of thedeployments and then transported to the field sites to includea surrogate laboratory control. Postfeeding experiments wereconducted in darkness to avoid algal growth and under constanttemperature conditions (20 � 2�C) provided by a thermosta-tized chamber. Individual feeding rates (cells/animal/h) weredetermined as the change in cell density over 4 h accordingto the method described by McWilliam and Baird [7]. Celldensity was estimated from absorbance measurements at � 650 nm in a dual-beam spectrophotometer (Uvikon 941; Kon-tron Instruments, Bletchley, UK) using standard calibrationcurves based on at least 20 data points (r2 0.98).

Whole-effluent bioassays

In situ bioassay responses generally have difficulties in dis-criminating contaminant from other environmental factorscausing biological effects (i.e., salinity). In an attempt to solvethis problem, laboratory toxicity tests with filtered water (re-tention size, 1.2 �m; GF/C glass fiber; Whatman, Maidstone,UK) obtained from the samples taken for pesticide analysiswere conducted within 48 h, and effects on survival and feed-ing responses were assessed. Bioassays followed previouslyreported procedures [16] and consisted of exposing groups offive fourth-instar juveniles in 150 ml of filtered water with analgal concentration of 5 � 105 cells/ml. Control treatments in

Daphnia magna in situ responses to pesticides Environ. Toxicol. Chem. 26, 2007 373

ASTM hard water and blank controls without animals alsowere included. Treatments were replicated five times. Follow-ing 24-h exposures, survival and feeding responses were de-termined as explained above in In situ bioassays.

Enzyme assays

Juveniles were homogenized at 4�C in a 1:4 wet weight tobuffer volume ratio in 100 mM phosphate buffer (pH 7.4)containing 100 mM KCl and 1 mM ethylenediaminetetra-ace-tic acid. Homogenates were centrifuged at 10,000 g for 10min, and the supernatants were immediately used as enzymesources. Biochemical measurements were carried out on a Uvi-kon 941 Plus dual-beam and Spectra-max Plus (Bio-Tek In-struments, Winooski, VT, USA) microplate reader spectro-photometer. Assays were run at least in duplicate. The activityof AChE was determined by a modification of the Ellmanmethod adapted for a microplate [9]. Acetylcholinesterase ac-tivity was measured in the presence of 1 mM acetylthiocholineand 0.1 mM 5,5�-dithiobis-2-dinitrobenzoic acid, and the in-crease of absorbance was measured at 405 nm. Carboxyles-terase activity was measured by the ultraviolet light methodas described by Mastropaolo and Yourno [17] in the presenceof 0.25 mM �-naphthyl acetate, and the formation of naphtholwas monitored by the increase in absorbance at 235 nm. Cat-alase activity was measured by the decrease in absorbance at240 nm because of H2O2 consumption (extinction coefficient,40 M/cm) according to the method described by Aebi [18].The reaction volume was 1 ml and contained 50 mM phosphatebuffer (pH 6.5) and 50 mM H2O2 [19]. Glutathione S-trans-ferase activity toward 1-chloro-2,4-dinitrobenzene (CDNB)was measured as described by Habig et al. [20]. The reactionmixture contained 100 mM phosphate buffer (pH 7.5), 1 mMCDNB, and 1 mM reduced glutathione. The formation of S-2,4 dinitrophenyl glutathione conjugate was evaluated by mon-itoring the increase in absorbance at 340 nm. Proteins weremeasured according to the method described by Lowry et al.[21] using serum albumin as standard.

Data analysis

Within each deployment date, postexposure feeding ratesand enzymatic activities at contaminated sites, reference sites,and surrogate laboratory controls were compared by one-wayanalysis of variance (ANOVA) followed by post-hoc Tukey’smultiple-comparison test [22]. Before analysis, data werechecked to meet ANOVA assumptions of normality and var-iance homoscedasticity and, if required, were log-transformed.

Pearson correlation analyses were performed on log-trans-formed environmental parameters and measured biological ef-fects to identify potential causal agents of toxicity to aquaticbiota. Before analyses, biological responses of field-exposedD. magna were converted to responses relative to surrogatelaboratory controls. From the 17 pesticides analyzed, onlythose detected in upstream and Delta del Ebro stations wereconsidered (13 pesticides). Measured pesticide concentrationsin water (Ci) were changed to toxic units (TU) according tothe procedure described by Liess and Von der Ohe [23] (Eqn.1):

CiTU (1)LC50i(D. magna)

where LC50i(D. magna) is the 48-h acute median lethal concen-tration reported for the measured compounds ([24]; http://www.pesticideinfo.org). The LC50s were set to 100,000

�g/L for mecropop and 2,4-D, 95,000 �g/L for alachlor,64,000 �g/L for bentazone, 11,000 �g/L for MCPA, 7,000�g/L for atrazine, 1,200 �g/L for propanil, 1,000 �g/L forsimazine and isoproturon, 600 �g/L for molinate, 400 �g/Lfor diuron, and 1 �g/L for fenitrothion and diazinon (basedon experimental data obtained in the laboratory [25]). The TUvalues were based on acute rather than sublethal end points,because for most of the compounds analyzed, this was the onlytoxicity information available. Because of the great disparityof reported data, the lowest value for those chemicals withmore than one reported LC50 was considered unless specifiedotherwise. Other potential environmental parameters, includ-ing conductivity, suspended solids, and water flow, also wereconsidered, because previous studies have shown that theseparameters may affect the studied responses [3].

RESULTS

Water quality

For the sake of clarity, water-quality parameters were av-eraged across deployments (Table 1). General water-qualityparameters of the studied drainage channels indicated increas-ing conductivity from site 1 toward site 4, reaching valuesgreater than 3 mS/cm in the latter station. Suspended solidsfollowed a similar trend, with drainage channels (sites 2–4)having approximately sixfold more solids suspended in waterthan the upstream reference station (site 1). Water flow in thedrainage channels was very low and variable compared toreference station 1, which may explain the lower oxygen levels(6 mg/L) found in stations 2 to 4 in comparison with station1. Other physicochemical variables were quite similar acrossstations.

Contaminants

Results depicted in Table 1 indicate that residues of 15 ofthe 17 pesticides analyzed were found in water collected atthe studied sites. Pesticide levels of compounds includingMCPA, bentazone, propanil, molinate, and fenitrothionshowed a clear increase on the drainage channels, whereaschlortoluron, atrazine, and metalochlor reached greater levelsat station 1. The other remaining compounds showed similarlevels in the four studied stations. Bentazone, MCPA, propanil,and atrazine had peak values higher than 1 �g/L, with ben-tazone reaching levels as high as 487 �g/L. On most occasions,other compounds including molinate and fenitrothion werepresent at levels higher than 0.1 �g/L. Cyanazine and mala-thion were the only compounds not detected.

Survival and feeding responses

The mean percentage of animals recovered (dead and alive)from the chambers after field exposures was always higherthan 90%. Mean survival rates of D. magna recovered fromin situ chambers were high in most field sites (90%) (Fig.2A) except for site 4 in May and June ( 70%). In all de-ployments, postexposure feeding rates of organisms deployedin the reference site (station 1) did not differ significantly ( p 0.05) from that of surrogate laboratory controls. Conversely,feeding responses of individuals deployed in the drainagechannels were significantly inhibited when compared withthose of the reference station (site 1), with individuals fromstation 4 showing the greatest effects, followed by those fromstations 3 and 2 (Fig. 2B). Effects on feeding also varied acrossdeploying dates, with July having the lowest effects and Mayand June having the greatest effects. In all deployments, feed-

374 Environ. Toxicol. Chem. 26, 2007 C. Barata et al.

Table 1. Mean � standard deviation (SD), minimum or limits of quantification (Min), and maximum (Max) of environmental parameters at thefour studied Spanish sites from May 2005 to August 2005a

Water quality

Station 1

Mean SD Min Max

Station 2

Mean SD Min Max

Flow (m/s) 0.4 0.1 0.3 0.5 0.01 0.01 0.01 0.1T (�C) 21.98 1.67 19.9 24 22.55 2.75 19.6 25.65Oxygen (mg/L) 8.75 1.02 7.83 10.15 6.44 0.39 6.01 6.95Conductivity (�S/cm) 938.63 169.12 769 1,091 1,523.6 245.06 1,374.5 1,886.5pH 8.48 0.07 8.39 8.55 7.9 0.15 7.77 8.12Suspended solids (mg/L) 1.7 0.17 1.6 1.9 8.42 2.23 6.2 10.65

PesticidesMCPA (�g/L) 0.076 0.038 0.038 0.128 2.242 3.731 0.267 7.835Mecoprop (�g/L) 0.009 0.002 0.007 0.012 0.008 0.008 0.002 0.022,4-D (�g/L) 0.024 0.009 0.015 0.036 0.024 0.023 0.008 0.057Bentazone (�g/L) 0.369 0.236 0.169 0.7 80.006 131.35 7.324 276.73Simazine (�g/L) 0.024 0.006 0.02 0.034 0.138 0.162 0.017 0.377Isoproturon (�g/L) 0.005 0.003 0.002 0.008 0.002 0.001 0.001 0.003Chlortoluron (�g/L) 0.001 0.001 0.005 0.002Atrazine (�g/L) 1.001 0.402 0.503 1.487 0.412 0.226 0.234 0.743Diuron (�g/L) 0.011 0.004 0.005 0.014 0.015 0.009 0.006 0.023Propanil (�g/L) 0.001 0.001 0.005 0.002 0.398 0.641 0.002 1.345Molinate (�g/L) 0.318 0.24 0.109 0.613Alachlor (�g/L) 0.095 0.135 0.018 0.286 0.218 0.435 0.018 0.871Metalochlor (�g/L) 0.018 0.007 0.012 0.027Fenitrothion (�g/L) 0.165 0.167 0.050 0.398Diazinon (�g/L) 0.001 0.000 0.001 0.001 0.011 0.022 0.001 0.044

a Empty cells indicate values not detected. MCPA 2-methyl-4-chlorophenoxyacetic acid; 2,4-D 2,4-dichlorophenoxyacetic acid.

ing rates were relatively constant in the surrogate laboratoryand reference station, with values ranging between 5 � 105

and 7 � 105 cells/individual/h.Whole-effluent survival and feeding responses of filtered

water collected at the deployment sites differed from thosemeasured in situ (Fig. 2C and D). Water samples from station3 had toxicity effects on survival and/or feeding in three ofour four occasions, and those from station 4 were only mar-ginally toxic in May.

Biomarker responses

The B-esterase activity enzymatic responses of field-ex-posed D. magna varied among sites and across deployments,showing the greatest inhibition effects in the drainage stationsduring May and June (Fig. 3A and B). Enzymatic activityresponse patterns of CbE and AChE were similar in May andJune but differed in July and August, with CbE showing greaterlevels of inhibition. In all deployments, AChE and CbE ac-tivities of individuals transplanted in the reference station werenot significantly ( p 0.05) different from those maintainedin the laboratory. Activities of CAT and GST also varied acrossstations and deployments (Fig. 3C and D). Individuals trans-planted in drainage channels showed CAT activities lower thanthose from the reference site and laboratory controls in June,but in August, measured CAT activities from site 1 were higherthan those of the laboratory control. Activities of GST werehigher in individuals transplanted in station 4 in May and Julyand in all field stations in August when compared to those oflaboratory controls.

Causal relationships between environmental parametersand biological effects

In all 16 deployments (four sampling dates at each of foursites), the estimated combined joint toxicity of the pesticideresidues measured in water collected in the drainage channelsin May, June, July, and August were, on average, 446-,

580-, 87-, and 1.5-fold more toxic than that of reference station1 (Table 2). From the 185 pesticide residues detected in water(Table 1), only 73 residues of nine pesticides had a toxicityof 0.0001 TU or greater (Table 2). For the sake of clarity, itwas assumed that pesticide residues with toxicity thresholdslower than 0.0001 TU could not be considered as potentiallytoxic to D. magna. The pesticide contributing the most to thecombined joint toxicity was diazinon in reference station 1(Table 2). In the drainage channels, fenitrothion and diazinoncontributed the most in 10 of 12 deployments and bentazonein 2 of 12 deployments (Table 2).

Bivariate correlation analyses denoted a clear, negative as-sociation between the predicted combined joint toxicity andmost measured biological effects, with CAT and GST activitiesbeing marginally inhibited and not affected, respectively (Ta-ble 3). Bentazone, MCPA, propanil, molinate, and fenitrothionwere inversely related to postexposure feeding responses, withpropanil also being negatively related to feeding responsesobtained in whole-effluent tests performed with filtered water.The B-esterase (AChE and CbE) and CAT activities were neg-atively related to propanil, molinate, and fenitrothion. Notice,however, that residual toxicity levels of propanil and molinatewere highly correlated (Pearson coefficients 0.84, p 0.05)to fenitrothion, which was the compound contributing the mostto the predicted toxicity, and those of MCPA with bentazone(Pearson coefficients 0.91, p 0.05). Activities of GSTshowed a negative and a positive relationship with fenitrothionand diazinon levels, respectively. Physicochemical water pa-rameters dramatically affected feeding rates and B-esterases(AChE and CbE), and conductivity marginally affected CATactivities. Nevertheless, it is important to consider that theseparameters were positively correlated (Pearson coefficients 0.6, p 0.05) to residual toxicity levels of propanil, molinate,and fenitrothion, which also affected B-esterases and/or feed-ing and CAT responses.

Daphnia magna in situ responses to pesticides Environ. Toxicol. Chem. 26, 2007 375

Table 1. Extended

Station 3

Mean SD Min Max

Station 4

Mean SD Min Max

0.02 0.01 0.01 0.1 0.03 0.02 0.01 0.123.66 2.13 21 25.5 24.13 2.49 22 26.75

5.92 0.74 5.35 6.9 6.42 0.77 5.6 7.352,044.6 669.28 1,412.5 2,980 3,363.8 672.15 2,675 4,250

7.79 0.2 7.59 8.06 7.84 0.13 7.66 7.9410.9 1.48 9.95 12.6 11.02 4.38 6.65 15.4

0.897 0.609 0.168 1.598 1.929 1.423 0.478 3.8540.007 0.005 0.003 0.014 0.005 0.003 0.003 0.010.012 0.006 0.005 0.019 0.011 0.005 0.004 0.017

95.466 132.42 14.573 293.18 153.6 223.04 23.149 487.540.053 0.024 0.028 0.084 0.023 0.022 0.0510.002 0.001 0.001 0.003 0.001 0.001 0.002

0.445 0.158 0.267 0.588 0.369 0.205 0.161 0.6260.017 0.008 0.008 0.027 0.007 0.008 0 0.0191.757 2.199 0.013 4.68 0.375 0.564 0.002 1.1950.331 0.304 0.146 0.784 0.256 0.156 0.134 0.4850.032 0.064 0.0018 0.127

0.317 0.398 0.050 0.681 0.065 0.098 0.050 0.2070.003 0.006 0.001 0.012 0.001 0.002 0.001 0.004

DISCUSSION

Environmental water parameters

Physicochemical water parameters measured in the up-stream reference site and at the drainage channel sites weredirectly related with the geological and hydrological dynamicsof the delta and Ebro River. The reported relatively low levelsof water flow in all the studied sites and the high conductivityvalues in the drainage channels indicated lenthic river con-ditions and an increasing salinization toward the east end ofthe delta (from site 2 to site 4). The previous water parametersare explained by the fact that most freshwater habitats of thedelta are at intertidal level. Thus, the water of drainage chan-nels is pumped into the sea when required, and water can reachmoderate to relatively high levels of conductivity, dependingon the income of seawater and the outflux of freshwater fromthe Ebro River [12]. Reference station 1 was located near asubmerged damp that regulates and, hence, slows the flow ofwater coming from the Ebro River, thus allowing collectionof water into the main distribution channels. The relativelyhigh amount of suspended solids found in stations 2 to 4 com-pared with that found in the upstream station 1 were directlyrelated to the constant flow of water and sediment comingfrom the rice fields into the drainage channels.

The levels of residue pesticides that we found support pre-vious studies indicating that herbicides, such as bentazone,MCPA, propanil, and molinate, were used in great amountsfor weed control in the Delta del Ebro [12], whereas othercompounds, including atrazine, simazine, alachlor, and me-talochlor, accounted for only 1 to 2% of the total herbicidesin the Delta del Ebro [26]. Nevertheless, it is important tomention that atrazine, chlortoluron, and metalochlor pesticideresidues reached greater levels in the Ebro River than in deltachannels, thus indicating that these herbicides also were usedalong the lower part of the Ebro River [27]. Organophospho-rous pesticides, despite having an extremely low environmen-tal persistence in the natural conditions occurring in the delta

[28], were present in drainage channels, with fenitrothionreaching levels near 0.7 �g/L in water, which were similar tothe amounts found in the water by Barcelo et al. [29] 1 d afterbeing detected in rice fields at 10 �g/L. This means that fen-itrothion is still massively applied to control rice-stem bores(Chilo supresalis) and the American crayfish (Procambusclarkii) [29,30]. Regarding the other compounds found in wa-ter samples, the organophosphate diazinon and the herbicidesmecoprop, 2,4-D, isoproturon, and diuron were found at lowlevels in most samplings across all studied sites, thus indicatingthat these agrochemicals are used along the lower part of EbroRiver [12].

In general, herbicides are known to pose negligible directenvironmental hazards because of their low toxicity when usedat recommended doses and to have short half-lives (1–2 d)[31]. Indeed, direct damage of herbicides to wildlife has neverbeen reported in the Delta del Ebro, but temporary abundancealterations of phytoplankton and zooplankton communities in-duced by herbicide use may indirectly affect wildlife by re-ducing food supply [12]. On the contrary, despite having shorthalf-lives, organophosphorous pesticides are known to be toxicfor vertebrate and invertebrate species, and adverse effects ofthese compounds in wildlife of the Delta del Ebro have beenreported [30]. In the present study, measured herbicide andorganophosphorous residues in water from May to Augustwere quite high compared to those reported elsewhere [12],thus suggesting that their massive and periodic use may com-pensate for their low environmental persistence and, hence,cause direct toxic effects in aquatic wildlife.

Evaluation of in situ effects

In situ–exposed D. magna from the upstream site (site 1)and those never deployed in the field but subject to the samehandling conditions (laboratory control) showed postexposurefeeding rates similar to those reported by McWilliam and Baird[3], thus indicating that this in situ assay was robust and con-

376 Environ. Toxicol. Chem. 26, 2007 C. Barata et al.

Fig. 2. Toxicological responses of Daphnia magna juveniles after in situ (A and B) and laboratory (C and D) exposure at the four studiedSpanish sites in May, June, July, and August. Responses of a laboratory control (Lab) also are depicted. Postexposure (PE) and whole-effluent(WE) feeding rates are depicted as 105 cells/individual/h. Distinct letters indicate significant differences ( p 0.05) among sites following analysisof variance and post-hoc Tukey’s multiple-comparison tests. Error bars indicate the standard errors (n 5–6).

sistent across different environmental conditions and studies.Daphnia magna were deployed at stations 2 to 4 and showedstrong, significant reductions of postexposure feeding rates in9 of 12 deployments, whereas survival was only affected in 2of 12 deployments during May and June. Bivariate correlationsindicated that the combined joint pesticide toxicity, propanil,molinate, fenitrothion, suspended solids, and salinity (mea-sured as conductivity) were negatively related with in situfeeding responses. Studies conducted by McWilliam [32] onthe effect of various environmental parameters, which includedtemperature, water hardness, pH, and suspended solids, foundthat only high flow rates in the presence of suspended solidshad negative effects on postexposure feeding rates. In the pres-ent study, the combination of low flow rates (0.01–0.03 m/s)and moderate amounts of suspended solids (6–15 mg/L) (Table1) should have prevented feeding rates from being affected inthe drainage channels. Furthermore, when water collected atthe studied sites was filtered and its toxicity assessed in thelaboratory, effects on feeding were detected in only 3 of 12

samples, thus suggesting that contaminants bound to suspend-ed solids rather than suspended solids alone or salinity wereaffecting the observed postexposure feeding responses in D.magna.

Many studies support the point of view that for organo-chlorine, organophosphorous, and pyrethroid insecticides, theparticulate-associated phase is an important route of avail-ability to river biota, whereas herbicides are transported indissolved form [33]. Schulz and Liess [4] found that mortalityin the amphipod Paramelita nidroculus exposed in situ toinsecticide runoff events was caused primarily by chlorpyrifosand endosulfan levels associated with suspended particles. InD. magna, the importance of contaminants bound to particleshas been analyzed directly by measuring the bioavailabilityand uptake of contaminants bound to particles and indirectlyby studying their effects on feeding [7,34]. Many contami-nants, including metals, polycyclic hydrocarbons, organochlo-rine (lindane) and pyrethroid pesticides, sorb to particles andimpair feeding in D. magna at lower concentrations than those

Daphnia magna in situ responses to pesticides Environ. Toxicol. Chem. 26, 2007 377

Fig. 3. Acetylcholinesterase (AChE; A) carboxylesterase (CbE; B), catalase (CAT; C), and glutathione S-transferase (GST; D) activities of Daphniamagna juveniles after in situ exposures at the four studied Spanish sites in May, June, July, and August. Responses of a laboratory control (Lab)also are depicted. Activities of AChE, CbE, and GST are depicted in nmol/min/mg protein and those of CAT in �mol/min/mg protein. Distinctletters indicate significant differences ( p 0.05) among sites following analysis of variance and post-hoc Tukey’s multiple-comparison tests.Error bars indicate the standard errors (n 3); in some sites, standard errors were too small to be depicted.

Table 2. Expected mean toxicities depicted as toxic units (TU) of the measured pesticide residues to Daphnia magna at the fourSpanish sites during May 2005 to August 2005a

Date Site MCPA Bentazone Simazine Atrazine Diuron Propanil Molinate Fenitrothion Diazinon �TU

May 1 — — — 0.0002 — — — — 0.0006 0.00092 — 0.0002 0.0001 0.0001 0.0001 0.0011 0.001 0.3981 0.0002 0.40083 0.0001 0.0002 0.0001 0.0001 0.0001 0.0039 0.0013 0.5886 0.0004 0.59474 — 0.0004 0.0001 0.0001 — 0.001 0.0008 0.2075 0.0001 0.2100

June 1 — — — 0.0001 — — — — 0.0003 0.00052 0.0002 0.0004 0.0004 0.0001 0.0001 0.0002 0.0007 0.1305 0.0009 0.13323 0.0001 0.0007 0.0001 0.0001 — 0.0019 0.0004 0.6806 0.0004 0.68434 0.0001 0.0008 — 0.0001 — 0.0003 0.0004 0.052 — 0.0536

July 1 — — — 0.0001 — — — — 0.0006 0.00082 0.0007 0.0043 — — — — 0.0002 0.1308 0.0437 0.17993 0.0001 0.0046 — 0.0001 — — 0.0003 — 0.0116 0.01684 0.0002 0.0076 — — — — 0.0003 — 0.0042 0.0124

Aug. 1 — — — 0.0001 — — — — 0.0008 0.0012 — 0.0001 0.0001 — — — 0.0002 — 0.0007 0.00113 — 0.0004 — — — 0.0001 0.0002 — 0.0006 0.00154 0.0004 0.0008 — — — — 0.0002 — 0.0006 0.002

a — cells with TUD. magna � 0.0001; �TU combined toxicity; MCPA 2-methyl-4-chlorophenoxyacetic acid.

affecting survival [7,16]. For organophosphorous pesticides,the disparity of results across compounds suggests that effectson feeding or related traits (growth and reproduction [16]) arecompound specific. For example, McWilliam and Baird [7] did

not find sublethal effects of pirimiphos-methyl and chlorpyr-ifos on feeding or related responses, whereas the oppositebehavior was reported for fenitrothion and diazinon [35]. Forherbicides, information is restricted to propanil, which ac-

378 Environ. Toxicol. Chem. 26, 2007 C. Barata et al.

Table 3. Bivariate correlation coefficients between expected toxicities (TUD. magna) of pesticide residues depicted in Table 2, suspended solids(SS), conductivities, and biological responses of Daphnia magna juvenilesa

AChE CbE CAT GST FED PE FED WE

�TU �0.70** �0.67** �0.45* �0.66** �0.44*MCPA �0.55**Bentazone �0.56**Propanil �0.71** �0.73** �0.48* �0.70** �0.50**Molinate �0.72** �0.68** �0.48* �0.63**Fenitrothion �0.80** �0.79** �0.43* �0.49* �0.48*Diazinon 0.41*Conductivity �0.52** �0.44* �0.88**SS �0.51** �0.55** �0.80** NA

a For clarity, only marginally significant (*p 0.1) and significant (**p 0.05) correlations are depicted. AChE acetylcholinesterase; CbE carboxylesterase; CAT catalase; GST glutathione S-transferase; FED PE postexposure feeding; FED WE whole-effluent feedingresponses; NA not applicable; MCPA 2-Methyl-4-chlorophenoxyacetic acid.

cording to Villarroel et al. [36] was able to impair feeding andsurvival of D. magna juveniles at 0.1 and 5 mg/L, respectively.In the present study, levels of pesticide residue in water con-verted to TU of MCPA, bentazone, propanil, molinate, andfenitrothion were negatively related with feeding responses.Notice, however, that postexposure feeding responses at sta-tions 3 and 4 were still dramatically inhibited in August, whenthe above-mentioned pesticide residues were too low to causeany effect, thus suggesting that regardless of these compounds,other contaminants should have accounted for the observedpatterns of feeding inhibition. Although not considered in thepresent study, pyrethroid insecticides and metals (i.e., chro-mium and copper) are used in the Delta del Ebro to controlalgae and invertebrate pest species in the water [12]. Thus,compounds other than those analyzed in the present study maybe present in the water of drainage channels and, hence, impairfeeding rates of D. magna.

Enzymatic activities can be used as fast and specific bio-logical responses to identify compounds causing toxic effectsin individuals [6]. For organophosphorous and carbamate pes-ticides, B-esterases constitute one of the best prognostic bio-markers to monitor biological effects in exposed biota [9],whereas many different contaminants (i.e., hydrocarbon qui-nones, nitroaromatics, organochlorine compounds, polycyclicaromatic hydrocarbons, and metals) and environmental factors(i.e., ultraviolet radiation and oxygen) may affect free radicalprocesses and, hence, may alter free radical scavenging sys-tems, such as antioxidant and GST enzyme activities [37]. InD. magna, B-esterases and antioxidant enzyme responses havebeen used successfully in the laboratory as prognostic andspecific indices of toxicity effects for many compounds [8,9].In this study, inhibition patterns of B-esterases of D. magnaindividuals deployed in drainage channels during May andJune likely were related to the quite high levels of fenitrothionresidue measured in water, which according to previous resultsshould be toxic to D. magna [35] and should inhibit AChEand CbE activities [9,30]. Residue levels in water of molinateand propanil, despite also of being negatively related to B-esterases activities, probably were too low (TU 0.01) tocause any effect.

For the other two studied enzymes, observed inhibition andinduction activity patterns precluded the identification of asingle causal relationship. For example, in August, when mea-sured pesticide residue levels were the lowest, both CAT andGST activities increased in D. magna deployed in the field,which indicates that other compounds not measured in thepresent study may be affecting those activities. In June, how-

ever, when pesticide levels were high, CAT activities wereinversely related with predicted combined joint toxicity. Al-though not analyzed in the present study, the fungicides pro-piconazole and tebufenozide, which are known to alter anti-oxidant and GST enzymes in mammals and insects [38,39],are massively used in August to preserve rice ears [12]. Cy-permethrin, which is used through the whole rice-growing sea-son, also is known to induce oxidative stress and to enhanceGST in insects and crustaceans [39,40]. Thus, antioxidant andGST responses observed in August may be associated withthe presence of pyrethroid and fungicide residues in water. InJune, high levels of herbicide residues in water may havecontributed to the observed inhibition patterns of CAT. Studiesconducted in fish and mammals have shown that molinate,propanil, and propanil’s major metabolite, 3,4-dichloroaniline,were able to cause oxidative stress and/or to disrupt antioxidantand GST enzyme activities [41,42]. Hence, the presence ofhigh residue levels of parental herbicides and/or their metab-olites may have contributed to the observed CAT and GSTactivity patterns in June.

In summary, our results lend positive support to the use ofin situ feeding bioassays in combination with biochemical re-sponses and chemical analyses to assess effects and to identifytoxic components within complex agrochemical mixtures inthe field, thus contributing to a more realistic assessment ofecological risks. Furthermore, the results reported here are inagreement with those obtained in other sentinel species, in-cluding crabs, mussels, and birds [12,30], thus indicating thatseveral species of the natural reserve and the nearby Medi-terranean Sea bays are at risk of being affected by pesticidesassociated with rice culture. Nevertheless, as a precautionaryremark, it is important to consider that whereas the approachused in the present study moves us closer toward ‘‘in situtoxicity identification evaluation,’’ issues arising from otherconfounding factors that can influence biomarker responsesstill sound a note of caution for interpreting the findings asconclusive diagnostic proofs of cause-and-effect relationshipsin individual substances.

Acknowledgement—This work is supported by the European Unionin ‘‘Integrated modeling of the river-sediments-soil-groundwater sys-tem; advanced tools for the management of catchment areas and riverbasins in the context of global change’’ [AQUATERRA, GOCE505428] and the Spanish Ministry of Education and Science[CGL2004-03514/HD, CTM2005-25168-E] and reflects the authors’view. The European Union is not liable for any use that may be madeof the information contained herein. Joana Damasio was supportedby the Portuguese Fundacion de Ciencia Tecnologia PhD grant (SFRH/

Daphnia magna in situ responses to pesticides Environ. Toxicol. Chem. 26, 2007 379

BD/23269/2005) and Carlos Barata by a Spanish-Ramon y Cajal con-tract.

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