Notice: Wiley Online Library will be unavailable on Saturday 30th July 2016 from 08:00-11:00 BST / 03:00-06:00 EST / 15:00-18:00 SGT for essential maintenance. Apologies for the inconvenience.
If you can't find a tool you're looking for, please click the link at the top of the page to "Go to old article view". Alternatively, view our Knowledge Base articles for additional help. Your feedback is important to us, so please let us know if you have comments or ideas for improvement.
Aquatic contaminants—including pesticides, veterinary medicines, heavy metals, polycyclic aromatic hydrocarbons, and nutrients—can occur intermittently in surface waters. For example, pesticides can be repeatedly applied to crops and may enter the environment by a number of potential routes, including discharge of industrial effluents, spillage, spray drift, runoff events, overland flow, and drainage inputs 1–3. Timing and quantity of pesticide release to surface waters vary according to the nature of use, compound-specific physiochemical properties, and characteristics of the agricultural system. Consequently, pesticides are likely to be released to aquatic systems in pulses, which can vary considerably in both duration and magnitude, rather than as continuous discharge. For example, more than150 site-years of monitoring program data for atrazine in watersheds of the midwestern United States representing worst-case potential runoff concentrations (upper 20th percentile) indicate that, for small watersheds (typically 9–40 square miles), the median duration of peaks greater than 15 mg/L is 2 d, with a 90th percentile duration of 7 d (see Brain et al. 4 and citations therein). Standard laboratory toxicity tests employ continuous exposure conditions with set exposure durations. Therefore, these approaches may not provide a truly reflective estimate of the effects of a substance on aquatic organisms in the real world and could result in either over- or underestimation of impacts. Several currently used approaches are available that assess the effects of exposure under more realistic conditions. These include multispecies microcosm studies or the use of single-species studies that attempt to mimic exposure scenarios seen in the environment 5.
The relationship between toxicity and duration of contaminant exposure has been evaluated for a wide range of substances, including chlorpyrifos, esfenvalerate, fenvalerate, lambda-cyhalothrin, diazinon, thiocyanate, Zn, and Cd 6–11. However, more limited numbers of studies have examined the influence of multiple pulse fluxes and the effects of pulse interval and frequency 12, 13. Although these studies have typically been performed over relatively short durations, focusing primarily on mortality, the data demonstrate that exposure to multiple pulses may result in effects realized at concentrations similar to or higher than comparable continuous exposures. Reasons for this include the following: (1) organisms are able to detoxify or depurate accumulated test material during the between-pulse interval 14, 15, where the potential for recovery is defined by the duration interval prior to subsequent exposure and the toxic mode of action of the substance of interest; (2) biochemical adaptation can occur, in which the first pulse may strengthen survivors through alterations in biochemical pathways that adapt organisms to the presence of a toxicant 16, 17; and (3) tolerant individuals from a population can be selected, in which weaker individuals are removed by an initial pulse, resulting in the selection of more robust individuals.
However, examples also exist in which the resultant toxicity of a pulsed exposure is greater than an equivalent continuous exposure. Studies with permethrin, for example, have demonstrated enhanced toxicity when animals were exposed to two 1-h pulses with a 6-h interval, compared with a continuous 2-h pulse 13. Moreover, exposure pulse duration and frequency also have demonstrated importance. For example, Karen et al. 18 exposed mummichog (Fundulus heteroclitus) to four pulses of chlorpyrifos over either 4 d or four weeks and found that organisms exposed to four pulses over four weeks demonstrated lower inhibition of brain acetyl cholinesterase activity than animals exposed to four pulses over 96 h. Recovery of the affected biochemical pathway might have been facilitated by the longer recovery period between pulses in the weekly test.
Therefore, to assess more accurately and realistically the potential environmental impacts of chemicals likely to demonstrate pulsed exposure characteristics, it may be appropriate to evaluate repeated exposures explicitly during the risk assessment process. Several specifically designed experimental approaches could be used for this purpose, including the use of a time-weighted averages, in which an average concentration is calculated for the exposure period and compared with the results of standard toxicity data, or the use of modeling approaches.
Consequently, the present study investigated the ecotoxicological effects of sequential pulsed exposures (2-d pulsed, and 4-d pulses) of pesticides (isoproturon, metsulfuron methyl, and pentachlorophenol [PCP]) with a variety of toxic modes of action on Lemna minor relative to continuous exposures. As described by Cedergreen et al. 19, 20 effects from pulsed exposures depend largely on compound-specific uptake and degradation or dissipation rates and potential for plants to recover. Accumulation is defined largely by the lipophilicity and molecular charge of the compound 19, 20. Potential for recovery depends on the nature of the interaction of the compound with the target site, the biological mode of action, and the specific growth rate of the species 4. Compounds with higher KOW values, such as PCP (log KOW = 5.05), tend to enter plants rapidly compared with more hydrophilic compounds, such as metsulfuron-methyl (log KOW = 0.018), causing a more pronounced effect when applied in pulses 19, 20; isoproturon has intermediate lipophylicity (log KOW = 2.5). Plant recovery from exposure to photosystem II (PSII)-inhibiting herbicides such as isoproturon is typically rapid 21, 22, whereas for sulfonylurea herbicides, which inhibit branched-chain amino acid synthesis, a lag period is often experienced 23. Recovery characteristics for aquatic plants from exposure to PCP, which uncouples mitochondrial oxidative phosphorylation, are largely unknown. Therefore, isoproturon is expected to exhibit lower toxicity under pulsed relative to continuous exposure, metsulfuron-methyl is expected to demonstrate equivalent toxicity under pulsed and continuous exposures, and PCP is expected to show greater toxicity under concentrated pulsed exposure conditions relative to continuous exposure.
Unlike previous studies of effects of pulsed exposure that have used invertebrates and fish, the present study focuses on a sublethal endpoint (growth) and investigates effects over a prolonged exposure period of 42 d. The primary objectives were to: (1) evaluate whether the use of standard short-term plant ecotoxicity tests, using fixed concentrations of pesticides, are predictive of longer term effects; (2) explore the effects of pulsed exposures for herbicides with differing toxic modes of action; and (3) explore the utility of the time-weighted average approach for estimating impacts on plant growth.
MATERIALS AND METHODS
The effects of three herbicides—isoproturon, metsulfuron-methyl, and PCP—were evaluated under pulsed exposure conditions using a modification of the OECD guideline for testing with the aquatic macrophyte L. minor24. Specifically, modifications were made to the study length and the introduction of a subculturing stage. Two pulsed exposure scenarios were explored, and effects were compared with a continuous exposure scenario over a period of 42 d, using individually assessed compounds and a semistatic test procedure.
Organism and chemicals
Duckweed (L. minor) was purchased from Blades Biological. Analytical-grade isoproturon, metsulfuron-methyl, and PCP were acquired from British Greyhound Chromatography and Allied Chemicals.
The effects of each study compound on L. minor growth were assessed separately. For each of the three compounds, four concentrations were evaluated for continuous exposure (Table 1); equivalent concentrations were selected for the 2- and 4-d pulse scenarios (Table 1). Concentrations were selected such that each pulse exposure scenario delivered a dose per week equivalent to that of the continuous exposure. Triplicate beakers (borosilicate glass, 600-ml capacity) were employed for each treatment level (including controls) and exposure scenario. Concentrated stock solutions (in methanol) of each herbicide were diluted in standard Lemna growth medium 24 to prepare all test solutions. The final methanol concentration in each test solution was kept at ≤0.05% v/v to minimize any phytotoxic effects of the organic solvent. Each test vessel (beaker) received 100 ml of the appropriate test solution.
Table 1. Nominal exposure concentrations of pesticides used in continuous and pulsed exposure tests with Lemna minor
To initiate each test, three colonies (each comprising three fronds) of healthy, young L. minor plants were placed into each test vessel. Test solutions were renewed three times per week over a period of six weeks. Lemna minor was exposed exclusively to treatment media (herbicide) throughout the study duration for the continuous exposure scenario. For the 2-d pulse scenario, plants were exposed to the herbicide for two consecutive days per week and growth medium (media + appropriate volume of methanol) only for the remainder of the week. For the 4-d scenario, plants were exposed to the herbicide for four consecutive days per week and growth medium only for the remainder of the week. Control solutions were renewed on the same schedule and frequency as continuous exposure test solutions. For the pulsed and continuous exposures, samples of test media (100 ml) were taken once per week throughout the study for analysis of herbicide concentrations.
Plants were grown at 20 ± 1°C for 42 d under a cool white fluorescent light with an intensity of approximately 10,500 lux. Weekly subsampling and reinoculation of L. minor were essential, because the doubling time for growth of untreated plants is typically less than 2.5 d. Therefore, plants were subsampled at the end of each week. Subsamples similarly consisted of three colonies, with each colony comprising three fronds. Subsamples were then reinoculated into fresh exposure medium, and the experiment continued. To ensure that plants were in the exponential growth phase, the most recently developed fronds were selected whenever possible.
Digital images were taken of individual test colonies at the start of each study in replicate beakers to allow total frond area calculation by image analysis. Images were also taken prior to and after subsampling at the end of each week. The pH and dissolved oxygen were measured once per week. A Canon Powershot S45 digital camera was used to take all digital images, which were then analyzed in ImageJ 1.29x. Color images were initially converted to eight-bit gray scale images, which in turn were converted to binary images and analyzed to calculate the total frond area.
Calculation of frond area over time
The theoretical total frond area for each treatment over time was derived from the image analysis data using the following equation
where TSAt is the theoretical surface area (that is, the surface area expected at time t if no subsampling of L. minor occurred and growth was not limited), TSAp is the theoretical surface area at the end of the preceding week, SAt is the measured surface area of L. minor at time t, and SAi is the surface area of L. minor inoculated from the colony from the previous week. Average specific growth rate was also calculated for each treatment group and corresponding control.
Samples (100 ml) were acidified with 0.5 ml of concentrated H3PO4. Acidified samples were then extracted using C18 SPE cartridges (Kinesis Solutions) that had been preconditioned with 2 ml methanol and 4 ml conditioning buffer (250 ml Lemna growth media and 1.25 ml concentrated H3PO4). Samples were then passed through the cartridges at 10 ml/min, and the cartridges were subsequently washed with 4 ml washing buffer and air dried for 10 min. Herbicides were then eluted with 2 × 1 ml methanol, which was then analyzed by high-performance liquid chromatography (HPLC).
Analytical verification of exposure concentrations was performed with a Dionex Summit HPLC system consisting of a GINA 50 autosampler and a P580 quaternary gradient pump paired with a UVD 170S UV/visible detector. A 150 cm × 4.6 mm, 4 µm Genesis C18 column from Kinesis Solutions was used to perform separations. For isoproturon analysis, the mobile phase consisted of 50% acetonitrile:50% of a 0.05% TFA solution in water, and the detection wavelength was 240 nm. For PCP, the mobile phase was 60% acetonitrile:40% 0.05% TFA solution in water, and the detection wavelength was 238 nm. Because of the high limit of detection for sulfonylureas using HPLC analysis, it was not possible to analyze solutions containing metsulfuron-methyl. However, this compound is known to be relatively stable at environmental pH values 25, so degradation during the experiment was considered unlikely. This compound is also not particularly sorptive, so interaction with test vessels was not expected.
For each pesticide, a linear regression analysis was performed on the log-transformed area of L. minor. Initially, exponential modeling was explored, but it was found that variation increased with area; therefore, prior to all analyses, the area of Lemna was log transformed (base e). The covariate day and the factors concentration and exposure scenario, including their associated potential interaction, were fitted as effects in the regression model in order to estimate their effect on L. minor growth rate when compared to the control reference level. The following model was fitted
The group factor represents the dose by application method combination, which has 13 levels, including the control. The coefficients α and β1 represent the control effect across time. The coefficients β2j and β3j are the differences for each group, j = 1…12, over time compared with the control. The coefficient β3j represents the difference in growth rate between each group and the control.
Pair-wise differences between application methods, within concentration levels, were calculated using a Bonferroni adjustment for multiplicity. The adjusted 5% statistical significance level was 0.0021, and results were compared against this adjusted level to evaluate significance. The 42-d median effective concentration (EC50) values were modeled using nonlinear regression. All statistical analyses were performed in GenStat 8 statistical software.
Analytical recoveries for PCP and isoproturon were 82% (relative standard deviation [RSD] 5.4%) and 84% (RSD 8.6%), respectively. Comparison of mean measured concentrations in test solutions with nominal concentrations (Table 2) indicated that actual exposure concentrations were typically within 20% of expected concentrations. Therefore, for ease of interpretation, the data are reported subsequently as nominal concentrations, and concentrations of pulses are expressed in terms of their equivalent continuous exposure concentration. The results are summarized below.
Table 2. Analytical results for pulsed exposure studiesa
Continuous exposure concentration (µg/L)
Concentrations are expressed as percentage of nominal concentration; standard errors are given in parentheses.
For continuous and pulsed exposures at 100 µg/L isoproturon, time–response relationships for L. minor growth are illustrated in Figure 1. Nominal 42-d growth rate EC50 values for the continuous and 4-d isoproturon pulse scenarios both exceeded 100 µg/L; for the 2-d pulse scenario the EC50 was lower, at 97 µg/L (Table 3). Inhibition of L. minor growth rate under pulsed and continuous exposure scenarios is summarized in Figure 2. There were no significant differences in L. minor growth rate between the control and the different exposure scenarios (p > 0.0021) at treatment concentrations of 5 and 25 µg/L. At 50 µg/L, L. minor growth rate under the 2-d pulse exposure scenario was not significantly different from the control (p = 0.069); however, continuous and 4-d pulse application scenarios resulted in significantly reduced growth (p < 0.001). On comparison of the three application scenarios at this treatment level, growth rate of L. minor was significantly higher under the 2-d pulsed application compared with plants exposed continuously (p < 0.001), which in turn demonstrated a significantly higher growth rate than plants in the 4-d pulse exposure scenario (p < 0.001). Growth in all treatment scenarios was significantly lower than the control (p < 0.001) at an exposure concentration of 100 µg/L. At this treatment level, L. minor plants exposed under the 2-d pulsed application scenario had a significantly higher growth rate compared with the 4-d pulsed application (p < 0.001), which in turn had a significantly higher growth rate than plants under the continuous application scenario (p < 0.001).
Table 3. Nominal 42-d EC50 values (µg/L) for continuous and pulsed exposure studies with L. minor exposed to isoproturon, metsulfuron-methyl, and pentachlorophenola
Upper and lower 95% confidence intervals are given in parentheses.
NC = confidence limit not calculable; EC50 = median effective concentration.
Relationships of time–response for L. minor growth under continuous and pulsed exposures at 1 µg/L metsulfuron-methyl are illustrated in Figure 1. Nominal L. minor 42-d growth rate EC50 values for metsulfuron-methyl ranged from 0.90 to 0.99 µg/L (Table 3). Calculated EC50 values demonstrated that continuous exposure was less toxic than the pulsed exposures. Inhibition of L. minor growth rate under pulsed and continuous exposures at each test concentration are summarized in Figure 2. There were no significant differences in growth rate across all treatment scenarios compared with the control (p > 0.05) at a treatment concentration of 0.1 µg/L. At 0.25 µg/L, the continuous application scenario demonstrated a significantly higher rate of growth than the 2-d pulse application scenario (p < 0.001). At 0.5 µg/L, growth rates in all exposure scenarios were significantly lower than control (p < 0.001), and the 2-d pulse application demonstrated significantly lower growth than the continuous exposure scenario (p < 0.001). At 1 µg/L, growth rates of L. minor under the pulsed and continuous exposure scenarios were significantly lower than control (p < 0.001). The growth rate of continuously exposed plants was significantly lower than that of plants exposed to the 2-d pulse.
Relationships of time–response for L. minor growth under continuous and pulsed exposures at 5 mg/L PCP are illustrated in Figure 1. Nominal 42-d growth rate EC50 values for PCP ranged from 1.26 to 4.00 mg/L (Table 3). Calculated EC50 values demonstrated that the 2-d pulsed exposure scenario was more toxic than the 4-d pulsed treatment, which was in turn more toxic than the continuous exposure scenario. Inhibition of L. minor growth rate under pulsed and continuous exposures at each test concentration is summarized in Figure 2. At a treatment concentration of 0.10 mg/L, none of the exposure scenarios differed significantly from the control or from each other over time (p > 0.05). At a treatment concentration of 0.5 mg/L, the 2-d pulsed and continuous exposure scenarios induced a significant reduction in growth rate compared with the control (p < 0.001). At 1.0 mg/L, all three exposure scenarios demonstrated significantly reduced growth relative to the control (p < 0.001). Plants exposed to 2-d pulse grew significantly less than those exposed continuously (p < 0.001). At 5 mg/L, all three exposure scenarios demonstrated significantly reduced growth relative to the control (p < 0.001). Under the 4-d and 2-d pulse exposure scenarios, plants grew significantly slower than those in the continuous exposure scenario (p < 0.001).
Test concentrations for isoproturon and metsulfuron-methyl were not dissimilar from reported or expected concentrations in natural systems. For example, concentrations of isoproturon in drain flow frequently exceeded 20 µg/L in samples of drainage water from an agricultural clay hill slope in the United Kingdom 26. Peterson et al., 27 reported maximum expected environmental concentrations for metsulfuron-methyl of 3 µg/L. In contrast, tested concentrations for PCP were significantly higher than levels in the natural environment. The highest concentration seen in a review of European monitoring studies for PCP was 1.5 µg/L 28.
Experimental results indicated that metsulfuron-methyl was more toxic than isoproturon, which in turn was more toxic than PCP over 42 d. Based on continuous exposure, nominal 42-d growth rate EC50s for metsulfuron-methyl (0.99 µg/L), isoproturon (>100 µg/L), and PCP (4007 µg/L) are in agreement with values from short-term standard studies of 60 µg/L 29, 0.1 to 0.8 µg/L 20, 30, and 2,370 µg/L 31. Therefore, it appears that short-term assays with L. minor may provide adequate prediction of longer term impacts from continuous exposures.
Data from the present study indicate little difference in the effects of pulsed versus continuous exposure regimes for metsulfuron-methyl on L. minor at the two higher concentrations tested. For PCP, however, pulsed exposure at high concentrations (much higher than expected environmental concentrations) resulted in a greater relative impact compared with corresponding continuous concentrations. Pulses of high concentrations appeared to inhibit growth, whereas continuous lower exposure concentrations did not, suggesting that a concentration effects threshold might have to be exceeded before impacts are realized. Moreover, plants exposed to high pulsed concentrations did not appear to recover between pulses. Conversely, the highest pulsed scenario concentration for isoproturon had less effect than the continuous regime. These observations might be explained by differences in compound-specific mechanisms of action, uptake, translocation, and metabolism and differences in the nature of effect (e.g., reversibility). Potential explanations for the observations for the study compounds are suggested below. More work would be required to determine whether these mechanisms do indeed explain the results.
Isoproturon is a urea herbicide that inhibits PSII by blocking the flow of electrons between the primary acceptor (QB binding site of the D1 protein) and secondary acceptor (plastoquinone) in the PSII complex 32, 33. In plants, recovery from exposure to PSII-inhibiting herbicides is typically rapid 4. For example, in Elodea nuttallii exposed to 50 µg/L linuron, PSII electron flow reached steady-state inhibition within about 4 h, and recovery was complete approximately 6 h after treatment removal 22. After a 24-h pulse exposure to isoproturon, photosynthesis, as measured by pulse amplitude modulation (PAM) fluorometry, recovered within 2 h from 47 to 55% inhibition at 200 to 320 µg/L in Scenedesmus vacuolatus34. In chloroplasts of Spinacia oleracea, it was shown that up to 85% of PSII could be blocked by diuron, another urea herbicide, without significant inhibition of photosynthesis under high light 21, 22. Moreover, even under conditions of high light intensity resulting in 40 to 80% inhibition of photosynthesis in Spinacia oleracea, no net degradation of the D1 protein of PSII was observed 35. Under continual high exposures, the potential for sustained secondary reactive oxygen species (ROS) formation and corresponding damage increases once cellular defense mechanisms (ROS scavenging enzymes) become overwhelmed 36. Thus, inhibition from pulsed exposures would be expected to be more transient in nature and induce less toxic effects compared with continuous exposures.
Sulfonylurea herbicides such as metsulfuron-methyl inhibit the production of branched-chain amino acids (leucine, isoleucine, and valine) via specific inhibition of the enzyme acetolactate synthase, the first enzyme specific to the branched-chain amino acid biosynthetic pathway 37. In Lemna sp. exposed to a variety of sulfonylurea herbicides, higher exposure concentrations were found to cause longer lag periods for recovery (assessed over 7–10 d), although longer exposure periods caused slower growth rates without a lag period 23. Because the pulsed exposure delivered the same exposure as the continuous exposure over a shorter period, it is not surprising that the EC50s for both exposure regimes were similar given the lag period of recovery elicited by more concentrated short exposures. Cedergreen et al. 20 also found that 3-h pulse exposures of metsulfuron-methyl induced an effect similar to that of longer term exposures (4–7 d) at an approximately tenfold higher concentration. Recovery time was also found to be comparatively long at up to 4 d after a single 3-h exposure 20. Mechanistically, sulfonylureas appear to inhibit cell division before depletion of the branched-chain amino acid pools, suggesting a block in the cell cycle at G1 and G2, preventing cells from entering mitosis 38–40. Although the effects of sulfonylurea herbicides are reversible, there is a lag-time associated with the resumption of amino acid synthesis and consequently the cascading resumption of cell division 40. This may explain the parity in sensitivity of pulsed versus continuous exposures for metsulfuron-methyl in contrast to isoproturon. Moreover, plants typically have surplus energy reserves (photosynthate) to buffer a temporary loss in photosynthesis 41 and can resort to photorespiration, whereas metabolic reactions such as amino acid synthesis are typically highly regulated by negative feedback, negating surplus reserves.
Pentachlorophenol is a potent uncoupler of oxidative phosphorylation in mitochondria 42 and inhibitor of photosynthetic phosphorylation in chloroplasts 43. In Lemna gibba, the toxicity of phenols tends to increase as the number of chlorine substituents on the phenol ring increases; for example, the reported EC50 values for L. gibba exposed to PCP, 2,4,5,6-tetrachlorophenol, 2,4,5-trichlorophenol, 2,4-dichlorophenol, 4-chlorophenol, and phenol are 2, 1.2, 2.1, 9.2, 183, and 540 µM, respectively 44. Furthermore, chlorinated phenols, from phenol to PCP, are metabolized similarly by conjugation to glucose-forming phenylglucoside metabolites that are more polar than their parent compounds, which is considered a stereotypic response of Lemna to these contaminants 44. Therefore, it is likely that, at lower continuous exposures, Lemna are capable of metabolizing PCP to less toxic and more water-soluble metabolites; however, with higher short, pulsed exposures the metabolic enzymes may become overwhelmed, allowing PCP to elicit a more pronounced effect on the mitochondria and chloroplast. This effect might also be sustained after removal from the pulsed exposure, because the chlorinated phenols are typically conjugated to glucose via a glycosidic bond that is susceptible to cleavage at lower pHs 44. Finally, there is a negative correlation between log KOW and the difference between long-term and pulse exposure, in which compounds with a high KOW enter plants rapidly and therefore have a larger effect when applied in pulses 19, 20. Consequently, the greater impact of PCP under pulse exposure scenarios compared with continuous exposure may be a consequence of rapid uptake and associated exceedance of metabolic capacity resulting in a flux of electron transport uncoupling (adenosine triphosphate [ATP] starvation) and corresponding cellular dysfunction. In accordance with the concept of internal exposure outlined by Escher and Hermens 45, higher concentrations of PCP, which is a weak acid (pKa 4.71), would be expected to lower the pH of exposure medium, resulting in a greater proportion of the neutral species, which readily crosses membranes with a very high log KOW of 5.05, resulting in potentially greater exposure relative to lower continuous exposure scenarios.
In summary, the lower toxicity of isoproturon at higher concentrations under pulsed relative to continuous exposure may be due to rapid recovery and high protein turnover of PSII. In contrast, recovery from metsulfuron-methyl exposure is associated with a lag phase resulting from arrested cellular division, which likely negates appreciable recovery between exposure intervals relative to the continuous exposure scenario, resulting in relative parity in sensitivity. Finally, the greater toxicity exhibited by PCP under concentrated pulsed exposure relative to continuous may be due to rapid uptake across cellular membranes, resulting in greater internal exposure, particularly at lower pH.
These results demonstrate that the relative effects of pulsed exposure compared with equivalent continuous exposures vary depending on the mechanism of action of the toxicant and the concentration tested. Although the use of a time-weighted average approach, in which mean exposure concentrations are compared with standard effects data, may be suitable for addressing effects on L. minor from time-varying exposure for some pesticide types (e.g., metsulfuron-methyl), it may overestimate the impact of other substances that are rapidly depurated or when quick recovery of growth is possible (e.g., isoproturon); in other instances (e.g., highly lipophilic compounds), its use could underestimate impacts (e.g., PCP).
For the purposes of chemical risk assessment, the potential usefulness of pulsed exposure studies has also been recognized by others 2, 37. Several laboratory approaches have been proposed, including the following: (1) static exposures accounting for dissipation processes (e.g., volatilization, hydrolysis, photolysis, and biodegradation); (2) static renewal exposures that can simulate any exposure pattern, provided the renewals are frequent enough (particularly useful for simulating square pulses); and (3) flow-through exposures in which continuous changes in exposure concentration can be achieved 46.
However, experimental methods provide information only on the exposure profile investigated in the study. By using modeling approaches, it might in the future be possible to predict the impacts of the wide range of pulsed exposure scenarios that could occur for a particular substance. The use of more complex models, such as the threshold hazard model or the damage assessment model, has been recommended by Ashauer et al. 47 for assessment of potential impacts of pulsed pesticide exposures. Fluctuating exposure concentrations as well as recovery processes are integrally considered in these models. Model predictions may provide greater utility than traditional approaches such as standard laboratory test median lethal concentrations (LC50s) or no observed effect concentrations (NOECs), given that models allow for quantitative prediction of effects for various exposure scenarios, including simulation of actual profiles. Toxicokinetic–toxicodynamic models have been successfully used to simulate effects from repeated pulsed exposures to aquatic organisms 48–51. Modeling of the findings in the current study requires mechanistic effects models that accommodate the different modes of action of each compound. To our knowledge, such models have not yet been developed for the growth of Lemna. Toxicokinetic–toxicodynamic models for aquatic macroinvertebrates and survival may serve as inspiration for the development of such mechanistic effects models for Lemna in future work, taking into consideration the different physiology and endpoints. Alternatively, the modeling of energy budgets might provide a good description of growth in Lemna, also under chemical stress 52.
The current study demonstrates differential effects in L. minor from pulsed versus continuous exposure regimes for selected pesticides. Lipohphilicity, mechanism of action, potential for recovery between pulses, and pulse concentration appear to define the relative impact of a compound under pulsed exposure scenarios. Consequently, the time-weighted average approach may not be appropriate for predicting effects of pulsed exposures. Several authors 2, 5, 47 have identified the requirement for research to underpin the linking of regulatory exposure and effects assessment as a priority. Consequently, a framework for assessing effects of pulsed exposures could potentially include (1) development of generalized pesticide exposure scenarios; (2) development of guidelines describing the design, performance, and interpretation of experimental pulsed exposure studies; and (3) development and evaluation of mechanistic-based models for predicting effects of pulsed exposures on aquatic organisms. Such research, in addition to other studies addressing uncertainties in extrapolation from standard tests to field impacts, will be invaluable in more accurately assessing the environmental risk of chemicals.
The authors acknowledge the U.K. Pesticides Safety Directorate, who funded this study, and M. Clook and P. Ashby of PSD for useful discussions.