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Passive Dosing in Chronic Toxicity Tests with the Nematode Caenorhabditis elegans Fabian Fischer,*,†,‡,∥ Leonard Böhm,‡ Sebastian Höss,§ Christel Möhlenkamp,† Evelyn Claus,† Rolf-Alexander Düring,‡ and Sabine Schaf̈ er† †

German Federal Institute of Hydrology (BfG), Am Mainzer Tor 1, 56068 Koblenz, Germany Institute of Soil Science and Soil Conservation, Research Center for BioSystems, Land Use, and Nutrition (iFZ), Justus Liebig University Giessen, Heinrich-Buff-Ring 26-32, 35392 Giessen, Germany § Ecossa, Giselastraße 6, 82319 Starnberg, Germany ‡

S Supporting Information *

ABSTRACT: In chronic toxicity tests with Caenorhabditis elegans, it is necessary to feed the nematode with bacteria, which reduces the freely dissolved concentration (Cfree) of hydrophobic organic chemicals (HOCs), leading to poorly defined exposure with conventional dosing procedures. We examined the efficacy of passive dosing of polycyclic aromatic hydrocarbons (PAHs) using silicone O-rings to control exposure during C. elegans toxicity testing and compared the results to those obtained with solvent spiking. Solid-phase microextraction and liquid−liquid extraction were used to measure Cfree and the chemicals taken up via ingestion. During toxicity testing, Cfree decreased by up to 89% after solvent spiking but remained constant with passive dosing. This led to a higher apparent toxicity on C. elegans exposed by passive dosing than by solvent spiking. With increasing bacterial cell densities, Cfree of solvent-spiked PAHs decreased while being maintained constant with passive dosing. This resulted in lower apparent toxicity under solvent spiking but an increased apparent toxicity with passive dosing, probably as a result of the higher chemical uptake rate via food (CUfood). Our results demonstrate the utility of passive dosing to control Cfree in routine chronic toxicity testing of HOCs. Moreover, both chemical uptake from water or via food ingestion can be controlled, thus enabling the discrimination of different uptake routes in chronic toxicity studies.

1. INTRODUCTION Miniaturized high-throughput toxicity tests using invertebrates are being increasingly employed in the ecotoxicity testing of anthropogenic chemicals, both to avoid the ethically questionable testing of vertebrates (fish and mammals) and to reduce costs.1−3 Free-living, nonparasitic nematodes are ubiquitous, diverse, and ecologically important invertebrates that inhabit soils and sediments.4,5 The nematode Caenorhabditis elegans has long been widely used to assess the toxicity of environmental pollutants.6−8 Among its advantages are its simple, cost-efficient cultivation and short generation time, which allow high-throughput toxicity testing,9 the development of a standardized test protocol,10,11 and test systems for assessing the toxicity of chemicals on molecular, organismal, and population scales.9,12,13 Defining and controlling the exposure concentrations of hydrophobic organic chemicals (HOCs) in aquatic toxicity testing is particularly challenging.14,15 The most common approach for introducing HOCs into toxicity tests is by preparing a concentrated stock solution of the test compound(s) in a water-miscible organic solvent and then adding a small volume of this solution to the test medium (solvent © 2016 American Chemical Society

spiking). The observed toxicity is then usually linked to the nominal concentration (Cnom), i.e., the spiked amount of chemical per volume of test medium. However, the toxicologically effective concentration is defined as the concentration of freely dissolved chemical (Cfree) and not Cnom.16−18 In toxicity tests with conventional solvent spiking, Cfree may be reduced significantly by sorption to organic matter (e.g., from food)19 and the test vessel surfaces,20 evaporation, degradation, and even biotransformation and uptake by the test organisms,21 all of which can result in a low test sensitivity.22 The sum of Cfree and the sorbed concentration is defined as the total concentration (Ctotal) in an aqueous test medium.23,24 In chronic toxicity tests, organic food particles, such as the Escherichia coli cells used in C. elegans toxicity tests, can substantially sorb HOCs and thereby affect the test sensitivity.25,26 Appropriate standardization of food type and quantity is thus a prerequisite of chronic toxicity testing, Received: Revised: Accepted: Published: 9708

June 14, 2016 August 5, 2016 August 5, 2016 August 5, 2016 DOI: 10.1021/acs.est.6b02956 Environ. Sci. Technol. 2016, 50, 9708−9716

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Environmental Science & Technology Table 1. Experiments Performed in This Study with Respective Methods and Test Substances exposure quantification method experiment

chemicals

(1) temporal changes in Cfree during C. elegans toxicity testing (2) toxicity testing at maximal Cfree

mixture of PHE, ANT, FLA, and PYR

(3) concentration−response testing (4) nfree and nfood at different food densities (5) toxicity at different food densities

NAP, ACE, FLO, PHE, ANT, FLA, PYR, BaA, CHR, and BaP PHE, PYR PHE, PYR PHE, PYR

passive dosing

hypothesis tested

Cfree measured by headspace SPME Cfree modeled by KE. coli,w

Cfree measured by headspace SPME

(i)

Cfree measured by in situ SPME

(ii)

Cfree modeled by KE. coli,w −

Cfree measured by in situ SPME nfree measured by in situ SPME nfood measured by LLE Cfree measured by in situ SPME CUfood measured by LLE

(ii) (iii)

solvent spiking

Cfree modeled by KE. coli,w CUfood modeled by KE. coli,w

especially in toxicological assessments of ingested chemicals.27 However, the role of food bacteria in altering apparent HOC toxicity on C. elegans is not fully understood. In the absence of bacteria, freely dissolved HOCs are the major contributor to chemical uptake in nematode tissue. 28,29 Toxicokinetic modeling of phenanthrene (PHE) uptake in C. elegans showed that only 9% of the total uptake flux was derived from bacterially associated PHE,30 suggesting that, in the presence of bacteria, the Cfree of HOCs is the main determinant of overall toxicity on C. elegans. However, minor contributions of foodassociated HOCs to the overall uptake might increase the observed toxic effect. As rates of bacterial food ingestion by C. elegans increase with increasing bacterial density in the medium,31 the chemical uptake rate via food (CUfood) may be higher at higher food densities. In laboratory test systems, Cfree of HOCs can be controlled and maintained by passive dosing. This method makes use of a biocompatible reservoir with a high absorption capacity for HOCs that is placed directly into the test medium. During the test, continuous partitioning of chemical(s) from the reservoir into the test medium compensates for chemical losses and thus maintains constant Cfree.14,32−34 Commercially available silicone O-rings (SRs) are often used as reservoirs in passive dosing studies due to their high practicality and versatility.15,35,36 To date, passive dosing has been used successfully to control Cfree during the toxicity testing of single HOCs,15,37,38 including acute toxicity testing with C. elegans,39,40 of chemical mixtures with defined composition,41,42 and of environmental samples.43−45 In passive dosing studies, Cfree can be measured after the test either by analyzing the chemical concentration in the passive dosing polymer (Cpolymer) and then dividing this value by the polymer-to-water partition ratio (Kpolymer,w)33 or by equilibrating the dosing polymer in a small volume of pure water, which is then analyzed.46 An alternative approach to measure the Cfree of HOCs is passive sampling followed by chemical analysis. In small test systems, solid-phase microextraction (SPME) can be applied either in situ (in situ SPME) or via the headspace (hsSPME). When SPME fibers are used in situ, Cfree can be determined by solvent extraction of the fiber polymer followed by analysis of the extract, provided that equilibrium partitioning between the polymer and the test medium was obtained and sampling was nondepletive.47 Nonetheless, in toxicity tests in which the test medium volumes are very small, avoiding chemical depletion may be challenging for HOCs. However, using hs-SPME, Cfree can also be measured during the test under nonequilibrium conditions. Even in low-volume toxicity tests, sampling with negligible depletion can be achieved by limiting the analyte uptake of the SPME fiber to the kinetic

(iv)

phase of the extraction, which prevents substantial effects on exposure concentrations during toxicity testing.48,49 Although passive dosing has been used in many toxicity studies to control the Cfree of HOCs, its utility in chronic toxicity tests that include complex test media containing food has yet to be demonstrated. Here, we describe a simple method for the toxicity testing of HOCs applied by passive dosing from SRs. It allows chronic toxicity testing with C. elegans as well as measurements of Cfree during and after toxicity testing. CUfood is determined based on measured E. coli-to-water partition ratios and C. elegans food-ingestion rates, assuming equilibrium between E. coli and the medium. In this study, Cfree, CUfood, and the apparent toxicity of passively dosed polycyclic aromatic hydrocarbons (PAHs) were compared to a conventional solvent spiking procedure. We hypothesized that (i) with conventional solvent spiking, Cfree would decrease during toxicity testing, especially due to sorption to food in the test medium, while passive dosing would allow the maintenance of a constant Cfree during the test. (ii) Thus, the PAH exposure of C. elegans would be higher with passive dosing than with solvent spiking, resulting in an increased apparent toxicity. (iii) The food density in the test medium would not affect the Cfree of PAHs with passive dosing, whereas CUfood would differ but could be adjusted by varying the food density. (iv) A higher CUfood with increasing food density could lead to a higher apparent toxicity when Cfree is simultaneously maintained by passive dosing. To test these hypotheses, different experiments comparing solvent spiking and passive dosing were performed (see Table 1 for an overview). Temporal changes in Cfree of PAHs were measured during the entire course of C. elegans toxicity testing (experiment 1). The toxicity of PAHs on C. elegans maximal Cfree was measured (experiment 2), and concentration−response testing was performed (experiment 3). The effects of different food densities on the amount of passively dosed PAHs in the water phase (nfree) and sorbed to food (nfood) were examined (experiment 4). In the test media containing different food densities, toxicity was related to Cfree and CUfood (experiment 5).

2. MATERIALS AND METHODS 2.1. Chemicals. Naphthalene (NAP), acenaphthene (ACE), fluorene (FLO), phenanthrene (PHE), anthracene (ANT), fluoranthene (FLA), pyrene (PYR), benzo[a]anthracene (BaA), chrysene (CHR), benzo[a]pyrene (BaP) (≥99.5% purity), and tetracycline (99% purity) were purchased from Sigma-Aldrich (Munich, Germany). Methanol (MeOH; 99.9% purity, Promochem, LGC Standards GmbH, Wesel, Germany), n-heptane (97% purity, Promochem), and acetone (99.8% purity, Carl Roth GmbH + Co. KG, Karlsruhe, 9709

DOI: 10.1021/acs.est.6b02956 Environ. Sci. Technol. 2016, 50, 9708−9716

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Environmental Science & Technology

temperature of 23 ± 2 °C. PAH mixtures below their maximal aqueous solubility were tested by incubating the SRs in the methanolic PAH solution followed by the addition of water to force partitioning of the PAHs into the silicone (Table S1); this method resulted in loading efficiencies of up to 82% (Figure S3). Single PAHs at their maximal water solubility were tested by loading the SRs with saturated methanolic PAH solutions (experiment 2); the presence of PAH crystals before and after the loading process was checked to ensure saturation.14 For concentration−response testing (experiment 3), PAH dilution series in MeOH with defined concentrations were prepared from saturated methanolic PAH solutions that were then used to load the SRs. The loaded SRs were rinsed three times for 10 min with bidistilled water and then dried with lint-free tissues to remove residual water. Most PAHs reached equilibrium in the test medium within 24 h, whereas time to equilibrium increased with increasing molecular weight (Table S2). In all passive dosing experiments, the SRs were added to the test medium and preincubated in the dark on an orbital shaker at 60 rpm for 24 h before the C. elegans juveniles were added. However, for the most hydrophobic chemicals tested (BaA, CHR, and BaP), equilibrium may not have been achieved within the 24 h preequilibration and 96 h test duration, wherefore 1000 rpm agitation was alternatively tested during pre-equilibration for enhanced equilibration kinetics (see page S4). 2.5. Cfree. We applied different approaches to determine Cfree in the experiments. The Cfree of four PAHs over time in the solvent-spiked or passively dosed samples were measured using hs-SPME (experiment 1). Sampling was carried out directly from the headspace above the test medium. The vials were equilibrated at 30 °C for 5 min prior to extraction of the samples for 30 min using 100 μm polydimethylsiloxane fibers (PDMS, fused silica 23Ga Red, Supelco, Sigma-Aldrich). Cfree was measured during the preincubation at 0.5, 4, 8, and 24 h and during C. elegans toxicity testing at 0.5, 4, 8, 24, 48, 72, and 96 h (n = 3 samples per time point). PAHs were quantified by means of an external calibration series with defined PAH concentrations and analyzed as described in section 2.7. In toxicity tests with passive dosing (experiments 2−5), Cfree was measured by SPME fibers [PDMS-coated glass core (coating: 30−31 μm, core diameter: 114−108 μm); Polymicro Technologies Inc., Phoenix, AZ] incubated in situ during the test and then removed from the test medium at the end of the experiments. Thereby, nondepletive sampling was achieved because analyte losses during sampling were compensated by the SR (see Table S2). The SPME fibers were cut to a length of 1 cm, yielding a PDMS volume of 0.136 μL cm−1.52 The fibers were precleaned three times for 10 min with MeOH in an ultrasonic bath and rinsed three times for 10 min with bidistilled water before their transfer to the test medium. For sampling, fibers incubated in the test medium were removed, dried using lint-free tissues, and extracted in 200 μL n-heptane for at least 24 h. Cfree (μg L−1) was calculated by dividing the concentration in the in situ SPME fiber (Cfiber, μg L−1) by the analyte-specific polymer-to-water partition ratios of the fibers (Kfiber,w, L L−1), as described by eq 2.

Germany) were used as organic solvents. PAH Mix 33, fluoranthene-d10, and benzo[a]pyrene-d12 were purchased from Dr. Ehrenstorfer GmbH (Augsburg, Germany) and served as internal standards. 2.2. C. elegans Toxicity Testing. C. elegans var. Bristol (strain: N2) were cultivated following standard procedures.11 The C. elegans toxicity test was performed according to ISO 10872, with slight modifications. The test medium consisted of E. coli (strain: OP50) suspended in M9-medium (6 g of Na2HPO4 L−1; 3 g of KH2PO4 L−1; 0.25 g of MgSO4 × 7H2O L−1; and 5 g of NaCl L−1) adjusted to a defined E. coli cell density of 500 formazine absorbance units (FAU) unless stated otherwise. The test medium was amended with 2 mg of tetracycline per L to inhibit bacterial growth during the experiments. This concentration was far below the noobserved-effect-concentration for C. elegans reproduction (NOEC = 10 mg L−1).50 Furthermore, synergistic effects with PAHs were expected to be negligible. Ten C. elegans juveniles (J1: first juvenile stage) were transferred to the PAHspiked test medium (experiment 1: 0.76 mL; experiments 2-5: 1 mL) in headspace glass vials (hs-vials, 46 × 22.5 mm with silicone−polytetrafluoroethylene screw caps, A-Z AnalytikZubehör GmbH, Langen, Germany) or 12 well multidishes (Nalgene Nunc, Rochester, NY). In experiment 1, the volume of the test medium had to be reduced to 0.76 mL so that the test design was comparable to that of the passive dosing experiments. The test vials were incubated at 20 °C in the dark for 96 h, after which the nematodes were heat-killed (15 min at 80 °C) and then stained with rose bengal. Juvenile offspring of the tested nematodes were counted and the number divided by the number of introduced test organisms (reproduction = offspring per test organism). Inhibition of reproduction compared to the control was calculated according to eq 1: %IR = 100 −

Ri × 100 mean R C

(1)

where % IR is the percentage of reproductive inhibition, Ri is the reproduction in replicate i, and RC is the reproduction in the control treatment. 2.3. Solvent Spiking. Acetone-based stock solutions with defined concentrations were prepared to yield the desired nominal concentration by adding 5 μL to the respective volume of test medium. For toxicity experiments in glass vials and well plates with 1 mL of test medium (experiments 2−5), this resulted in 0.5% acetone in the test medium, which is below the NOEC of 0.6% (data not shown). Expectedly, no significant difference was observed between water controls and solvent controls in any of the experiments (p > 0.05, Mann−Whitney U-test). The test vessels containing the spiked test medium were preincubated on an orbital shaker at 60 rpm in the dark for 24 h, after which C. elegans juveniles were added. 2.4. Passive Dosing. Silicone O-rings (14 mm inner diameter, ID: ORS-BS015, density 1.20 g cm−3, Altec Products Ltd., Cornwall, UK) were used as passive dosing reservoirs after verifying that their presence had no effect on the reproduction of C. elegans (Figure S1). The mean mass of the SRs was 142 ± 2 mg (n = 20), and their volume was 118 ± 1.7 μL. The SRs were precleaned three times for 10 min with MeOH in an ultrasonic bath. The SRs were loaded with single PAHs (experiments 2−5) or mixtures thereof (experiment 1) by incubating them in methanolic PAH solutions with defined PAH concentrations for at least 48 h at 60 rpm and a

measured Cfree =

Cfiber K fiber,w

(2)

Kfiber,w were calculated by multiplying the polymer to AlteSil silicone (Altec Products LTD) partition ratios (Kfiber,AlteSil, L 9710


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Environmental Science & Technology L−1)53 by the AlteSil silicone to water partition ratios (KAlteSil,w, L L−1),54 as shown in eq 3. K fiber,w = KAlteSil,w × K fiber,AlteSil

a concentration range of 0.1−25 μg L−1 in M9 test medium to ensure the comparability of the PAH distribution in the aqueous and gaseous phases.

(3)

Because nondepletive in situ SPME was not possible when applying solvent spiking, Cfree in the solvent-spiked samples was modeled by multiplying Cnom by the free fraction of PAHs in the test medium (experiments 2−5), which was calculated based on E. coli-to-water partition ratios (KE. coli,w, L kg−1), the mass of E. coli bacteria (mE. coli, kg dry weight), and the volume of test medium (Vmedium, L), as described in eq 4. modeled Cfree

⎛ =⎜ ⎜1 + ⎝

⎞ 1 ⎟×C mE. coli nom ⎟ × K E. coli ,w ⎠ V medium

3. RESULTS AND DISCUSSION 3.1. Temporal Changes in Cfree. The results of experiment 1 confirmed that Cfree decreased substantially over time with solvent spiking but remained constant during passive dosing after equilibrium partitioning between the SR and test medium was achieved [hypothesis (i), Figure 1]. Cfree of each of the four PAHs was reduced by 24−53% as early as 30 min after their solvent spiking. Because HOCs sorb to organic matter,19,55−57 including bacteria cells,58,59 this reduction probably resulted from sorption to E. coli cells. Cfree then remained constant until C. elegans juveniles were added, suggesting that equilibrium partitioning had already been attained within 30 min. This finding agrees with those of Lunsman and Lick,59 who reported that the sorption of HOCs to bacteria was in equilibrium within a few minutes. As expected, the reduction in Cfree differed between PAHs and generally increased with their increasing sorption tendency to E. coli, with analyte losses of 24% for PHE

(4)

The log KE. coli,w values were calculated as shown in eq 5 according to Baughman and Paris:51 log KE. coli,w = 0.907 × log Kow − 0.361

(5)

2.6. nfood and CUfood. In the solvent-spiked samples of experiments 2−5, nfood (μg) was modeled using the modeled Cfree (μg L−1) from eq 4. modeled n food = modeled Cfree × KE. coli ,w × mE. coli

(6)

We further determined KE. coli,w for all PAHs on the basis of the measured Cfree (in situ SPME) and Ctotal [liquid-liquid extraction (LLE), eq S3] in passively dosed samples (eqs S4 and S5) and compared these with the values obtained from Baughman and Paris.50 Determined KE. coli,w values were constant for the tested E. coli cell densities (Figure S4) and in good agreement with those calculated based on Baughman and Paris,50 confirming that KE. coli,w can be used to calculate nfood at varying food densities in chronic C. elegans toxicity tests. In passively dosed samples, nfood (μg) was thereof quantified by multiplying determined KE. coli,w (L kg−1) values with measured Cfree (μg L−1) and mE. coli (kg) values. measured n food = measured Cfree × KE. coli ,w × mE. coli

(7)

To compare exposure at different E. coli cell densities in both dosing approaches (experiment 5), CUfood (μg kgC. elegans−1 h−1) was calculated by means of the food ingestion rate of C. elegans at a specific cell density (IRa, kgE. coli kgC. elegans−1 h−1)30 using either modeled nfood in solvent spiking or measured nfood in passive dosing samples. n modeled or measured CUfood = food × IRa mE. coli (8) 2.7. Chemical Analysis. The PAH concentrations in extracts derived from in situ SPME and LLE were determined by gas chromatography−tandem mass spectrometry (GC−MS/ MS; 7000 Triple Quadrupole−GC−MS/MS, Agilent) with a HP-5MS column (30 m × 250 μm × 0.25 μm, 5% phenyl methyl siloxane, Agilent). For hs-SPME measurements, extraction and analysis were performed using an autosampler equipped with SPME (CTC-Analytics, Combi Pal PALSystem) and by GC−MS (Trace GC Ultra and ITQ 900 MS, Thermo Scientific, Waltham, MA) equipped with a TraceGOLD TG-XLBMS column (60 m × 250 μm × 0.25 μm, Thermo Scientific). PAHs were quantified by external standard calibration. Calibration samples for in situ SPME and LLE extracts were prepared at a concentration range of 1−2000 μg L−1 in n-heptane. Those for hs-SPME analysis were prepared at

Figure 1. Cfree during E. coli preincubation (0−24 h) and subsequent toxicity testing with C. elegans (24−120 h). PAHs were applied by (A) solvent spiking or (B) passive dosing. The blue line indicates the Cnom of the four PAHs (25 μg L−1) dosed by solvent spiking at time 0. For passive dosing, SRs were loaded with a 4.5 mg L−1 concentrated MeOH solution, yielding nominal concentrations between 3.9 (PYR) and 16.3 μg L−1 (PHE), and added to the test medium at time 0. 9711


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Environmental Science & Technology (log KE. coli,w: 3.87), 26% for ANT (log KE. coli,w: 3.92), 42% for FLA (log KE. coli,w: 4.40) and 53% for PYR (log KE. coli,w: 4.40). After the addition of C. elegans J1 juveniles to the test medium (24 h after PAH addition), the Cfree of all four PAHs decreased progressively, resulting in analyte losses of up to 89% for PYR by the end of the toxicity test (120 h, Figure 1A). A mass balance calculation showed that the sorption of PAHs to growing nematodes did not sufficiently explain the decrease in Cfree because the nematodes accounted for a maximum of 13% of the total lipid biomass in the test medium (see Table S3). Metabolic degradation by C. elegans might have contributed to the reduction in the Cfree, whereas metabolic degradation by E. coli was unlikely because after the immediate decrease due to sorption, the Cfree was not further reduced during the preincubation period. However, a fungus-like contamination can occasionally occur when chronic C. elegans toxicity testing is performed. Such an additional biomass might have contributed to the sorption of PAHs and, thus, could explain part of the continual decrease in the Cfree. Analyte losses due to processes such as evaporation, sorption to the test vial plastic walls, and photolysis can be considered as negligible because experiments were performed in closed glass vials in the dark. Figure 1B shows that, with passive dosing, equilibrium partitioning of all four PAHs was achieved during preincubation, such that the Cfree remained constant until the end of the toxicity test (120 h). This finding is in agreement with earlier studies in which HOCs were passively dosed in miniaturized aqueous toxicity tests with Daphnia magna,15 zebrafish embryos,35 and C. elegans.39,40 SRs have served as suitable dosing reservoirs in previous studies,42,61 including toxicity testing of single PAHs and mixtures thereof.62,63 However, those were acute toxicity tests performed without a food source in the test medium. Our results demonstrate that passive dosing using PAH-loaded SRs is also a suitable method of establishing constant Cfree of PAHs in standardized chronic toxicity tests with C. elegans, in which the nematodes are incubated in a complex food medium that acts as a strong sorptive sink for HOCs. This observation is promising regarding the application of passive dosing in other chronic toxicity tests. 3.2. Influence of the Dosing Method on the Apparent Toxicity of PAHs. Table 2 shows the half-maximal effect concentrations (EC50) after concentration−response testing of PHE and PYR (experiment 3). The reproductive toxicity on C. elegans was lower for passively dosed PAHs, whereas the solvent spiking of PHE and PYR in the same test vials resulted in EC50 values that were 2.6- and 3.8-fold higher, respectively (for the concentration−response curves, see Figure S8). These differences probably resulted from the uncertainty associated with using the modeled Cfree as exposure metric according to eq 4 because, as shown here, Cfree actually decreases progressively after C. elegans juveniles are added to the test medium (Figure

1A). In the plastic well plates, the EC50 values were 33- and 12fold higher than those determined in the passive dosing toxicity tests, probably as a result of evaporation (see page S9 for results and further discussion). Although earlier studies showed that passive dosing of HOCs can be performed in well plates,33,40 it is not yet clear whether passive dosing via SRs compensates for the large and rapid evaporative losses in toxicity tests of volatile chemicals in plastic wells containing C. elegans. PAHs mainly act as baseline toxicants, i.e., they impair the integrity and functioning of biological membranes.62 Because the membrane accumulation of PAHs is driven by passive diffusion from the test medium, a higher Cfree of PAHs results in a higher apparent toxicity,46 as observed in earlier studies of acute toxicity using invertebrate species.15,41 In our study, this correlation was illustrated by the activity of the chemicals in the test medium at maximal Cfree (experiment 2, Figure S7), in accordance with earlier investigations in which the baseline acute toxicity of neutral HOCs was in the 0.01−0.1 range.64 Similar to studies with daphnids, algae, and C. elegans,65,66 our results showed that, when applied by passive dosing, BaA and ANT exhibit chronic toxicity to C. elegans at chemical activities 0.05), whereas with increasing E. coli cell density, nfood increased significantly (Figure 2B, one-way ANOVA: p < 0.05, post-hoc Tukey: p < 0.05). The results demonstrate that even for chemicals with high sorption tendency to E. coli, such as PYR (log KE. coli,w: 4.40), passive dosing was able to maintain a constant nfree over a large range of E. coli cell densities [hypothesis (iii)]. Figure 3 relates the apparent toxicity of PHE to both Cfree and CUfood (experiment 5). When solvent spiking was applied, the apparent toxicity decreased significantly with increasing food density (Figure 3A, one-way ANOVA, p < 0.01, post-hoc Tukey: p < 0.01). This likely resulted from the decreased chemical uptake via the water phase because Cfree decreased considerably from 118.1 (FAU 125) to 28.5 μg L−1 (FAU 2000). Cfood likewise decreased from 691 (FAU 125) to 167 mg kgE. coli−1 (FAU 2000) as a result of the higher E. coli biomass in the test medium. However, CUfood remained relatively constant due to the increased food ingestion rate of C. elegans at higher food densities (Figure S10). Similar results were observed for PYR (Figure S9). However, these modeled exposure concentrations assume that the chemicals distributed solely between the water phase and the E. coli. The progressive losses that in fact occurred during solvent spiking might have altered Cfree and CUfood even further, thereby impeding the differentiation between uptake routes in chronic toxicity studies. In contrast to solvent spiking, the apparent toxicity of passively dosed PHE increased with increasing bacterial density, with significant differences between FAU 2000 and 500 as well as 125 and 500 (Figures 3B and S9, one-way ANOVA: p < 0.01, post-hoc Tukey: p < 0.05). These findings could be attributed to the increased CUfood because the Cfree of the PAHs remained constant over all E. coli cell densities. C. elegans reproduction thereby did not differ significantly in control groups (one-way ANOVA, p = 0.63, Figure S2), and, therefore, it seems unlikely that % IR was affected by the tested food densities, which is a

Table 2. Half-Maximal Effect Concentrations (EC50, μg L−1) of PHE and PYR for the Inhibition of C. elegans Reproduction after 96 h of Exposure to PAHs Provided by Solvent Spiking (n = 3) or Passive Dosing (n = 6) at 20 °C dosing procedure

solvent spiking

solvent spiking

passive dosing

test vessel exposure metric EC50 PHE (μg L−1) EC50 PYR (μg L−1)

plastic well plate modeled Cfree 1475 100

glass vial modeled Cfree 115 31

glass vial measured Cfree 44.8 8.2 9712


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prerequisite when testing toxicity at different food densities. Our results imply that the bacterially sorbed fraction is an important contributor to the overall toxic effect on C. elegans and are in contrast to the conclusion reached by Spann et al.,30 namely that only 9% of the total uptake flux of C. elegans is induced by bacterially associated PHE. However, in the experimental setup of Spann et al.,30 both Cfree and CUfood were consistently altered by varying the bacterial density as solvent spiking was used; thus, toxicity could not have been unequivocally assigned to a specific chemical fraction (dissolved or dietary). In this study, however, constant Cfree values of the PAHs were maintained while CUfood simultaneously increased with increasing bacterial densities. It may be that, with increasing ingestion, nematodes form more lipid reserves, which would result in higher internal chemical concentrations in the organism (due to the high sorption capacity of lipids for HOCs) and therefore a greater apparent toxicity of PAHs at higher food densities.67 Further studies are needed to investigate the contribution of food-bound chemicals to the overall toxic effect of HOCs on chronic end points, both in C. elegans and other invertebrate species. 3.4. Implementation for Toxicity Testing of Chemicals. Our study showed that passive dosing via SRs can eliminate the uncertainties of reduced Cfree caused by sorption of HOCs to food in chronic toxicity tests. Different SPME formats can thereby be used to measure and confirm exposure concentrations. The developed methods can thus improve the comparability and repeatability of the toxicity test, resulting in greater test sensitivity, more reliable toxicity testing, and better comparability of the toxicity data obtained in routine risk assessments of HOCs. A 24 h preincubation of the PAH-loaded SR in the test medium can be easily implemented in standardized nematode toxicity testing. However, higher agitation intensities during pre-equilibration might be needed prior toxicity testing to accelerate equilibrium partitioning when either very hydrophobic chemicals or high food densities are used. Passive dosing can furthermore be used to control both Cfree and CUfood by loading the passive dosing reservoir with a defined concentration while adjusting the food density in the test medium. Using this approach, we showed that CUfood contributes significantly to the overall reproductive toxicity of PAHs on C. elegans. In future studies, passive dosing can be employed to investigate the role of dietary chemical uptake in the toxicodynamics and toxicokinetics of HOCs. With this approach, the differences in measured toxicity can be assigned to a specific uptake route.

Figure 2. Masses of four PAHs in (A) the water phase (nfree, ng, mean ± standard deviation, n = 3) and (B) sorbed to food (nfood, ng, mean ± standard deviation, n = 3) at different E. coli cell densities (FAU 0, 125, 500, and 2000). The PAH-loaded SR was incubated in the test medium for 72 h. Different letters indicate significant differences between treatment groups with different bacterial densities (one-way ANOVA: p < 0.05, post-hoc Tukey: p < 0.05).

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ASSOCIATED CONTENT

S Supporting Information *

The Supporting Information is available free of charge on the ACS Publications website at DOI: 10.1021/acs.est.6b02956. Figures showing the reproduction of C. elegans under varying conditions, mean loading efficiencies of PAHs, nonlinear least-square regression of concentrations of five PAHs in in situ SPME fibers, E. coli to water partition ratios for PAHs, inhibition of reproduction of C. elegans after exposure of single PAHs at their maximal water solubility, and inhibition of reproduction of C. elegans after 96 hours of single-PAH exposure at their maximal chemical activity. Tables showing a schedule of water addition steps, time needed until 95% equilibrium

Figure 3. Inhibition of C. elegans reproduction (% IR mean ± standard deviation, n = 3) after 96 h exposures to PHE at different E. coli cell densities (FAU 125, 500, and 2000). PHE was provided by (A) solvent spiking or (B) passive dosing. Note, that the apparent toxicity was driven by Cfree with solvent spiking, whereas altered by CUfood while maintaining constant Cfree values with passive dosing. The data points were fitted using a logistic model for Cfree (solvent spiking; R2 = 0.85) and CUfood (passive dosing; R2 = 0.99).

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(9) Boyd, W. A.; Cole, R. D.; Anderson, G. L.; Williams, P. L. The effects of metals and food availability on the behavior of Caenorhabditis elegans. Environ. Toxicol. Chem. 2003, 22 (12), 3049−3055. (10) Traunspurger, W.; Haitzer, M.; Höss, S.; Beier, S.; Ahlf, W.; Steinberg, C. Ecotoxicological assessment of aquatic sediments with Caenorhabditis elegans (nematoda) - A method for testing liquid medium and whole-sediment samples. Environ. Toxicol. Chem. 1997, 16, 245−250. (11) ISO. Water quality − Determination of the toxic effect of sediment and soil samples on growth, fertility and reproduction of Caenorhabditis elegans (Nematoda); ISO 10872; International Organization for Standardization: Geneva, Switzerland, 2010. (12) Jager, T.; Á lvarez, O. A.; Kammenga, J. E.; Kooijman, S. Modelling nematode life cycles using dynamic energy budgets. Functional Ecology 2005, 19 (1), 136−144. (13) Van Assche, R.; Broeckx, V.; Boonen, K.; Maes, E.; De Haes, W.; Schoofs, L.; Temmerman, L. Integrating -omics: Systems biology as explored through C. elegans research. J. Mol. Biol. 2015, 427 (21), 3441−3451. (14) Mayer, P.; Wernsing, J.; Tolls, J.; de Maagd, P. G.; Sijm, D. T. Establishing and controlling dissolved concentrations of hydrophobic organics by partitioning from a solid phase. Environ. Sci. Technol. 1999, 33 (13), 2284−2290. (15) Smith, K. E.; Dom, N.; Blust, R.; Mayer, P. Controlling and maintaining exposure of hydrophobic organic compounds in aquatic toxicity tests by passive dosing. Aquat. Toxicol. 2010, 98 (1), 15−24. (16) Escher, B. I.; Hermens, J. L. Internal exposure: Linking bioavailability to effects. Environ. Sci. Technol. 2004, 38 (23), 455A− 462A. (17) Kiparissis, Y.; Akhtar, P.; Hodson, P. V.; Brown, R. S. Partitioncontrolled delivery of toxicants: A novel in vivo approach for embryo toxicity testing. Environ. Sci. Technol. 2003, 37 (10), 2262−2266. (18) Jahnke, A.; Mayer, P.; Schäfer, S.; Witt, G.; Haase, N.; Escher, B. I. Strategies for transferring mixtures of organic contaminants from aquatic environments into bioassays. Environ. Sci. Technol. 2016, 50 (11), 5424−5431. (19) Bö hm, L.; Schlechtriem, C.; Düring, R.-A. Sorption of hydrophobic organic chemicals to organic matter relevant for fish bioconcentration studies. Environ. Sci. Technol. 2016, 50 (15), 8316− 8323. (20) Schreiber, R.; Altenburger, R.; Paschke, A.; Küster, E. How to deal with lipophilic and volatile organic substances in microtiter plate assays. Environ. Toxicol. Chem. 2008, 27 (8), 1676−1682. (21) Gülden, M.; Mörchel, S.; Seibert, H. Factors influencing nominal effective concentrations of chemical compounds in vitro: Cell concentration. Toxicol. In Vitro 2001, 15 (3), 233−243. (22) Smith, K. E.; Jeong, Y.; Kim, J. Passive dosing versus solvent spiking for controlling and maintaining hydrophobic organic compound exposure in the Microtox (R) assay. Chemosphere 2015, 139, 174−180. (23) Bondarenko, S.; Gan, J. Simultaneous measurement of free and total concentrations of hydrophobic compounds. Environ. Sci. Technol. 2009, 43 (10), 3772−3777. (24) Düring, R.-A.; Bö hm, L.; Schlechtriem, C. Solid-phase microextraction for bioconcentration studies according to OECD TG 305. Environ. Sci. Eur. 2012, 24, 4. (25) Rose, R. M.; Warne, M. S.; Lim, R. P. Food concentration affects the life history response of Ceriodaphnia cf. dubia to chemicals with different mechanisms of action. Ecotoxicol. Environ. Saf. 2002, 51 (2), 106−114. (26) Pavlaki, M. D.; Ferreira, A. L.; Soares, A. M.; Loureiro, S. Changes of chemical chronic toxicity to Daphnia magna under different food regimes. Ecotoxicol. Environ. Saf. 2014, 109, 48−55. (27) Höss, S.; Schlottmann, K.; Traunspurger, W. Toxicity of ingested cadmium to the nematode Caenorhabditis elegans. Environ. Sci. Technol. 2011, 45 (23), 10219−10225. (28) Haitzer, M.; Burnison, B. K.; Hö ss, S.; Steinberg, C.; Traunspurger, W. Effects of quantity, quality, and contact time of dissolved organic matter on bioconcentration of benzo[a]pyrene in the

partitioning, masses of E. coli cells and C. elegans organisms in the test medium at the end of the toxicity test, measured maximal Cfree by in situ SPME, inhibition of reproduction of C. elegans, and ingestion rate of C. elegans as a function of the bacterial density in the test medium. Additional details on equilibrium partitioning between SR, test medium, and in situ SPME fiber; liquid−liquid extraction; experimental determination of log KE. coli,w and comparison to literature data; and discussion of the results shown in Table S4 and Figure S6. (PDF)

AUTHOR INFORMATION

Corresponding Author

*Phone: +49 341 235−1512; fax: +49 341 235−1787; e-mail: fabian.fi[email protected]. Present Address ∥

Helmholtz Centre for Environmental Research (UFZ), Department Cell Toxicology, Permoserstraße 15, 04318 Leipzig, Germany

Notes

The authors declare no competing financial interest.

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ACKNOWLEDGMENTS This study was financially supported by the German Federal Ministry for the Environment, Nature Conservation, Building and Nuclear Safety. We thank Philipp Mayer for kindly providing the in situ SPME fibers and Patrick Zurek for preliminary experiments. We gratefully acknowledge Julia Bachtin and Marina Ohlig for technical assistance and Benjamin Becker for helpful discussions. We are grateful to Beate Escher for a critical review of and helpful discussion on the manuscript. We thank the three anonymous reviewers for helpful comments on the manuscript.

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REFERENCES

(1) Breitholtz, M.; Rudén, C.; Hansson, S. O.; Bengtsson, B. E. Ten challenges for improved ecotoxicological testing in environmental risk assessment. Ecotoxicol. Environ. Saf. 2006, 63 (2), 324−335. (2) Baun, A.; Hartmann, N. B.; Grieger, K.; Kusk, K. O. Ecotoxicity of engineered nanoparticles to aquatic invertebrates: A brief review and recommendations for future toxicity testing. Ecotoxicology 2008, 17 (5), 387−395. (3) Diepens, N. J.; Arts, G. H. P.; Brock, T. C. M.; Smidt, H.; Van den Brink, P. J.; Van den Heuvel-Greve, M. J.; Koelmans, A. A. Sediment toxicity testing of organic chemicals in the context of prospective risk assessment: A review. Crit. Rev. Environ. Sci. Technol. 2014, 44 (3), 255−302. (4) Yeates, G. W. Nematode populations in relation to soil environmental factors - A review. Pedobiologia 1981, 22, 312−338. (5) Traunspurger, W. The biology and ecology of lotic nematodes. Freshwater Biol. 2000, 44 (1), 29−45. (6) Höss, S.; Williams, P. L. Ecotoxicity testing with nematodes. In Nematodes as Environmental Indicators; Wilson, M. J.; Khakouli-Duarte, T., Eds.; CABI Publishing: Wallingford, England, 2009; pp 208− 225.10.1079/9781845933852.0208 (7) Leung, M. C.; Williams, P. L.; Benedetto, A.; Au, C.; Helmcke, K. J.; Aschner, M.; Meyer, J. N. Caenorhabditis elegans: An emerging model in biomedical and environmental toxicology. Toxicol. Sci. 2008, 106, 5−28. (8) Meyer, D.; Williams, P. L. Toxicity Testing of Neurotoxic Pesticides in Caenorhabditis elegans. J. Toxicol. Environ. Health, Part B 2014, 17, 284−306. 9714


Article

Environmental Science & Technology nematode Caenorhabditis elegans. Environ. Toxicol. Chem. 1999, 18 (3), 459−465. (29) Haitzer, M.; Hö ss, S.; Traunspurger, W.; Steinberg, C. Relationship between concentration of dissolved organic matter (DOM) and the effect of DOM on the bioconcentration of benzo[a]pyrene. Aquat. Toxicol. 1999, 45 (2−3), 147−158. (30) Spann, N.; Goedkoop, W.; Traunspurger, W. Phenanthrene bioaccumulation in the nematode Caenorhabditis elegans. Environ. Sci. Technol. 2015, 49 (3), 1842−1850. (31) Nicholas, W. L.; Grassia, A.; Viswanathan, S. The efficiency with which Caenorhabditis briggsae (Rhabditidae) feeds on the bacterium, Escherichia coli. Nematologica 1973, 19 (4), 411−420. (32) Brown, S. L.; Pennello, G.; Berg, W. A.; Soo, M. S.; Middleton, M. S. Silicone gel breast implant rupture, extracapsular silicone, and health status in a population of women. Journal of Rheumatology 2001, 28 (5), 996−1003. (33) Smith, K. E.; Oostingh, G. J.; Mayer, P. Passive dosing for producing defined and constant exposure of hydrophobic organic compounds during in vitro toxicity tests. Chem. Res. Toxicol. 2010, 23 (1), 55−65. (34) Adolfsson-Erici, M.; Akerman, G.; Jahnke, A.; Mayer, P.; McLachlan, M. S. A flow-through passive dosing system for continuously supplying aqueous solutions of hydrophobic chemicals to bioconcentration and aquatic toxicity tests. Chemosphere 2012, 86, 593−599. (35) Butler, J. D.; Parkerton, T. F.; Letinski, D. J.; Bragin, G. E.; Lampi, M. A.; Cooper, K. R. A novel passive dosing system for determining the toxicity of phenanthrene to early life stages of zebrafish. Sci. Total Environ. 2013, 463−464, 952−958. (36) Vergauwen, L.; Schmidt, S. N.; Stinckens, E.; Maho, W.; Blust, R.; Mayer, P.; Covaci, A.; Knapen, D. A high throughput passive dosing format for the Fish Embryo acute toxicity test. Chemosphere 2015, 139, 9−17. (37) Kramer, N. I.; Busser, F. J.; Oosterwijk, M. T.; Schirmer, K.; Escher, B. I.; Hermens, J. L. Development of a partition-controlled dosing system for cell assays. Chem. Res. Toxicol. 2010, 23 (11), 1806− 1814. (38) Schmidt, S. N.; Holmstrup, M.; Damgaard, C.; Mayer, P. Simultaneous control of phenanthrene and drought by dual exposure system: The degree of synergistic interactions in springtails was exposure dependent. Environ. Sci. Technol. 2014, 48 (16), 9737−9744. (39) Kwon, H. C.; Roh, J. Y.; Lim, D.; Choi, J.; Kwon, J. H. Maintaining the constant exposure condition for an acute Caenorhabditis elegans mortality test using passive dosing. Environmental Health and Toxicology 2011, 26, e2011015−e2011015. (40) Roh, J. Y.; Lee, H.; Kwon, J. H. Changes in the expression of cyp35a family genes in the soil nematode Caenorhabditis elegans under controlled exposure to chlorpyrifos using passive dosing. Environ. Sci. Technol. 2014, 48 (17), 10475−10481. (41) Engraff, M.; Solere, C.; Smith, K. E.; Mayer, P.; Dahllöf, I. Aquatic toxicity of PAHs and PAH mixtures at saturation to benthic amphipods: Linking toxic effects to chemical activity. Aquat. Toxicol. 2011, 102 (3−4), 142−149. (42) Schmidt, S. N.; Holmstrup, M.; Smith, K. E.; Mayer, P. Passive dosing of polycyclic aromatic hydrocarbon (PAH) mixtures to terrestrial springtails: Linking mixture toxicity to chemical activities, equilibrium lipid concentrations, and toxic units. Environ. Sci. Technol. 2013, 47 (13), 7020−7027. (43) Bandow, N.; Altenburger, R.; Lübcke-von Varel, U.; Paschke, A.; Streck, G.; Brack, W. Partitioning-based dosing: An approach to include bioavailability in the effect-directed analysis of contaminated sediment samples. Environ. Sci. Technol. 2009, 43 (10), 3891−3896. (44) Booij, P.; Lamoree, M. H.; Leonards, P. E.; Cenijn, P. H.; Klamer, H. J.; van Vliet, L. A.; Akerman, J.; Legler, J. Development of a polydimethylsiloxane film-based passive dosing method in the in vitro DR-CALUX (R) assay. Environ. Toxicol. Chem. 2011, 30 (4), 898− 904. (45) Rojo-Nieto, E.; Smith, K. E.; Perales, J. A.; Mayer, P. Recreating the seawater mixture composition of HOCs in toxicity tests with

Artemia f ranciscana by passive dosing. Aquat. Toxicol. 2012, 120−121, 27−34. (46) Mayer, P.; Holmstrup, M. Passive dosing of soil invertebrates with polycyclic aromatic hydrocarbons: Limited chemical activity explains toxicity cutoff. Environ. Sci. Technol. 2008, 42 (19), 7516− 7521. (47) Mayer, P.; Tolls, J.; Hermens, J. L.; Mackay, D. Equilibrium sampling devices. Environ. Sci. Technol. 2003, 37 (9), 184A−191A. (48) Heringa, M. B.; Hermens, J. L. Measurement of free concentrations using negligible depletion-solid phase microextraction (nd-SPME). TrAC, Trends Anal. Chem. 2003, 22 (9), 575−587. (49) Bandow, N.; Altenburger, R.; Brack, W. Application of ndSPME to determine freely dissolved concentrations in the presence of green algae and algae-water partition coefficients. Chemosphere 2010, 79, 1070−1076. (50) Vangheel, M.; Traunspurger, W.; Spann, N. Effects of the antibiotic tetracycline on the reproduction, growth and population growth rate of the nematode Caenorhabditis elegans. Nematology 2014, 16 (1), 19−29. (51) Baughman, G. L.; Paris, D. F.; Hodson, R. E. Microbial bioconcentration of organic pollutants from aquatic systems - A critical-review. Critical Reviews in Microbiology 1981, 8 (3), 205−228. (52) Witt, G.; Lang, S. C.; Ullmann, D.; Schaffrath, G.; Schulz-Bull, D.; Mayer, P. Passive equilibrium sampler for in situ measurements of freely dissolved concentrations of hydrophobic organic chemicals in sediments. Environ. Sci. Technol. 2013, 47 (14), 7830−7839. (53) Gilbert, D.; Witt, G.; Smedes, F.; Mayer, P. Polymers as reference partition phase: Polymer calibration for an analytically operational approach to quantify multimedia phase partitioning. Anal. Chem. 2016, 88 (11), 5818−5826. (54) Smedes, F.; Geertsma, R. W.; van der Zande, T.; Booij, K. Polymer−water partition coefficients of hydrophobic compounds for passive sampling: Application of cosolvent models for validation. Environ. Sci. Technol. 2009, 43 (18), 7047−7054. (55) Gouliarmou, V.; Mayer, P. Sorptive bioaccessibility extraction (SBE) of soils: Combining a mobilization medium with an absorption sink. Environ. Sci. Technol. 2012, 46 (19), 10682−10689. (56) Hwang, S. C.; Cutright, T. J. Evidence of underestimation in PAH sorption/desorption due to system nonequilibrium and interaction with soil constituents. J. Environ. Sci. Health, Part A: Toxic/Hazard. Subst. Environ. Eng. 2004, 39 (5), 1147−1162. (57) Marschner, B. Sorption of polycyclic aromatic hydrocarbons (PAH) and polychlorinated biphenyls (PCB) in soil. J. Plant Nutr. Soil Sci. 1999, 162 (1), 1−14. (58) Xiao, L.; Qu, X. L.; Zhu, D. Q. Biosorption of nonpolar hydrophobic organic compounds to Escherichia coli facilitated by metal and proton surface binding. Environ. Sci. Technol. 2007, 41 (8), 2750− 2755. (59) Lunsman, T. D.; Lick, W. Sorption of hydrophobic organic chemicals to bacteria. Environ. Toxicol. Chem. 2005, 24, 2128−2137. (60) Moermond, C. T.; Traas, T. P.; Roessink, I.; Veltman, K.; Hendriks, A. J.; Koelmans, A. A. Modeling decreased food chain accumulation of PAHs due to strong sorption to carhonaceous materials and metabolic transformation. Environ. Sci. Technol. 2007, 41 (17), 6185−6191. (61) Oostingh, G. J.; Smith, K. E.; Tischler, U.; Radauer-Preiml, I.; Mayer, P. Differential immunomodulatory responses to nine polycyclic aromatic hydrocarbons applied by passive dosing. Toxicol. In Vitro 2015, 29 (2), 345−351. (62) Smith, K. E.; Schmidt, S. N.; Dom, N.; Blust, R.; Holmstrup, M.; Mayer, P. Baseline toxic mixtures of non-toxic chemicals: ″Solubility Addition″ increases exposure for solid hydrophobic chemicals. Environ. Sci. Technol. 2013, 47 (4), 2026−2033. (63) Bougeard, C.; Gallampois, C.; Brack, W. Passive dosing: an approach to control mutagen exposure in the Ames fluctuation test. Chemosphere 2011, 83 (4), 409−414. (64) Mayer, P.; Reichenberg, F. Can highly hydrophobic organic substances cause aquatic baseline toxicity and can they contribute to mixture toxicity? Environ. Toxicol. Chem. 2006, 25 (10), 2639−2644. 9715


Article

Environmental Science & Technology (65) Ahlers, J.; Riedhammer, C.; Vogliano, M.; Ebert, R. U.; Kühne, R.; Schüürmann, G. Acute to chronic ratios in aquatic toxicity Variation across trophic levels and relationship with chemical structure. Environ. Toxicol. Chem. 2006, 25, 2937−2945. (66) Ristau, K.; Akgul, Y.; Bartel, A. S.; Fremming, J.; Müller, M. T.; Reiher, L.; Stapela, F.; Splett, J. P.; Spann, N. Toxicity in relation to mode of action for the nematode Caenorhabditis elegans: Acute-tochronic ratios and quantitative structure-activity relationships. Environ. Toxicol. Chem. 2015, 34 (10), 2347−2353. (67) Endo, S.; Escher, B. I.; Goss, K. U. Capacities of membrane lipids to accumulate neutral organic chemicals. Environ. Sci. Technol. 2011, 45 (14), 5912−5921.

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