Arch. Environ. Contam. Toxicol. 39, 445– 451 (2000) DOI: 10.1007/s002440010126

A R C H I V E S O F

Environmental Contamination a n d Toxicology © 2000 Springer-Verlag New York Inc.

Toxicity and Bioavailability of Copper Herbicides (Clearigate, Cutrine-Plus, and Copper Sulfate) to Freshwater Animals B. J. Mastin, J. H. Rodgers, Jr. Department of Environmental Toxicology, P.O. Box 709, 509 Westinghouse Road, Clemson University, Pendleton, South Carolina 29670, USA

Received: 19 October 1999/Accepted: 13 June 2000

Abstract. In designing aquatic herbicides containing copper, an important goal is to maximize efficacy for target species while minimizing risks for nontarget species. To have a margin of safety for nontarget species, the concentration, duration of exposure (i.e., uptake), and form (i.e., species) of copper used for herbicidal properties should not elicit adverse effects on populations of nontarget species. To determine the potential for risk or adverse effects (conversely the margin of safety), data regarding the comparative toxicity of copper-containing herbicides are crucial. A series of comparative toxicity experiments was conducted, including baseline estimates of toxicity (LC50s, LOECs), sensitive species relationships (thresholds and exposure-response slopes), and bioavailability of toxic concentrations and forms of copper 7 days after initial herbicide application. Aqueous 48-h toxicity experiments were performed to contrast responses of Daphnia magna Strauss, Hyalella azteca Saussure, Chironomus tentans Fabricius, and Pimephales promelas Rafinesque to copper herbicides: Clearigate威, Cutrine威-Plus, and copper sulfate. D. magna was the most sensitive aquatic animal tested for all three herbicides; 48-h LC50s for organisms exposed to Clearigate, Cutrine-Plus, and copper sulfate were 29.4, 11.3, and 18.9 ␮g Cu/L, respectively. In terms of potency (calculated from the linearized portion of the exposure-response curves, which included 50% mortality), D. magna was the most sensitive animal tested. Organisms exposed to Clearigate, Cutrine-Plus, and copper sulfate had exposure-response slopes of 2.55, 8.61, and 5.07% mortality/␮g Cu/L, respectively. Bioavailability of Clearigate and Cutrine-Plus was determined by comparing survival data (LC50s) of test organisms exposed to herbicide concentrations during the first and last 48-h of a 7-day exposure period. Even in these relatively simplified water-only exposures, a transformation of copper to less bioavailable species over time was observed with a 100 –200% decrease in toxicity (i.e., an increase in 48-h LC50s) for all four test animals. This series of laboratory experiments provides a worst-case scenario for determining the risk associated with the manufacturer’s recommended application rates of Clearigate (100 –1,000 ␮g Cu/L), Cutrine-Plus (200 –1,000 ␮g Cu/L), and copper sulfate (100 –

Correspondence to: J. H. Rodgers Jr.

500 ␮g Cu/L) in natural waters for four nontarget freshwater animals.

Copper is both an essential element for organism survival and considered a priority pollutant by the U.S. EPA (US EPA 1980). Copper may exist in natural waters as free hydrated ions, complexed with inorganic and organic ligands, or sorbed onto particle surfaces (Deaver and Rodgers 1996). Copper toxicity to aquatic flora and fauna is principally due to soluble forms, such as the free ion Cu2⫹ (the most toxic form of copper to algae, protozoans, crustaceans, and fish) and some hydroxy and carbonate complexes (US EPA 1992; Nor 1987; Flemming and Trevors 1989). Copper has a lithic biogeochemical cycle, and precipitated and organically bound forms are generally less bioavailable to aquatic biota. In other words, only a fraction of the total metal mass introduced to an aquatic system is bioavailable for uptake and is subsequently able to elicit a response from aquatic biota (Suedel et al. 1996). To determine short-term margins of safety for nontarget species exposed to aquatic herbicides containing copper, knowledge of comparative toxicity, concentration, duration of exposure (i.e., uptake), and the bioavailable forms (i.e., species) of copper in the aquatic system is essential. As an aquatic herbicide, copper is commercially associated with a chelator to prevent rapid loss (i.e., precipitation and/or complexation) of copper from the water and maintain the toxicity of copper (in solution) over time (Elder and Horne 1978). For Clearigate威 and Cutrine威-Plus, a heterocyclic compound is produced when a central metallic ion (Cu2⫹) is covalently bonded to two or more organic compounds (e.g., emulsified surfactants); Cu2⫹ and its associated compounds are slowly released (due to photodecomposition) from their alliance with ethanolamine and ethylenediamine, respectively (Straus and Tucker 1993). Copper ion (Cu2⫹) penetrates the cell membranes of vascular and nonvascular aquatic vegetation, acting on receptors in photosynthetic cells, and interrupting the photosynthetic chain of electron transfer (Jursinic and Stemler 1983). In designing aquatic herbicides containing copper, an important goal is to maximize efficacy to target plant species while

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minimizing risk to nontarget species. Regulated in part by uptake, herbicide efficacy can be quantified by the magnitude of response of target species exposed to toxic concentrations and forms of copper. Residence time, aqueous and sediment characteristics (i.e., hardness, alkalinity, pH, and redox potential), photoperiod, compounds used in combination with copper, and the method and strategy of application to the aquatic system can also influence herbicide effectiveness. To have a margin of safety for nontarget species, the concentration, duration of exposure, and form of copper used for herbicidal properties should not elicit adverse effects on populations of nontarget species. To determine the potential for risk or adverse effects (conversely the margin of safety), data regarding the comparative toxicity of copper containing herbicides are crucial. A series of comparative toxicity and bioavailability experiments were conducted to evaluate the margins of safety of Clearigate, Cutrine-Plus, and copper sulfate for four nontarget freshwater animals as compared to the manufacturers’ recommended application rates (Table 1). It has been hypothesized that a significant portion of applied and residual copper in most water bodies exists in unavailable forms (i.e., complexed, precipitated, and/or sorbed to particulate matter or dissolved organics), whereas in most laboratory waters used for toxicity experiments copper is more readily bioavailable as free Cu2⫹ (US EPA 1992; Dobbs et al. 1994). Consequently, experiments employing “clean” laboratory waters overestimate actual risks to nontarget species that live in aquatic systems where additional binding sites (i.e., ligands) may decrease acute toxicity of copper (US EPA 1980; Gauss et al. 1985; Benson et al. 1994). The bioavailable fraction of copper observed by Deaver and Rodgers (1996) in laboratory waters was 67% of the total measured copper, considerably less than the presumed 100%, illustrating that even “clean” laboratory waters contain ligands that alter speciation of such metals as copper. When conducting laboratory experiments to estimate the margin of safety for nontarget organisms, we recognize that ligands present in laboratory water can influence the outcome of exposure experiments and these results must be further adjusted to account for specific field conditions when translating or scaling from the laboratory to the field. The margins of safety for copper-containing herbicides in aquatic systems generally increase as the concentrations and forms of copper from initial herbicide applications change over time. Consequently, it is assumed that bioavailability of toxic forms of copper to nontarget species will decrease over time (Kim et al. 1999). After initial application, manufacturers suggest 7–10 days to observe the effects of treatment on target plant species. Over long exposure periods (⬎ 96 h), however, the margin of safety for nontarget species is either difficult to measure (i.e., below detection limits), does not occur, or manufacturers have failed to identify it. Bioavailability of toxic forms of copper to nontarget species may not be accurately assessed when an organism remains in a test chamber for long periods of exposure (⬎ 96 h). Effects may be a result of exposure in the first or last 24 h and not a result of cumulative exposure. In this study, the margins of safety of nontarget species exposed to Clearigate, Cutrine-Plus, and copper sulfate over a 7-day exposure period were quantified by comparing survival of test organisms exposed to herbicide concentrations during the first and last 48 h of a 7-day experiment period. This

B. J. Mastin and J. H. Rodgers, Jr.

experimental design allowed for the identification of changes in concentrations of bioavailable forms of copper. To establish baseline margins of safety of aqueous copper herbicides for nontarget species, a series of comparative toxicity experiments were conducted. In short-term (48-h), aqueous laboratory exposures, nontarget species Hyalella azteca Saussure, Chironomus tentans Fabricius, Daphnia magna Straus, and Pimephales promelas Rafinesque were used (1) to determine the relative toxicities (i.e., LC50s and LOECs) and potency (i.e., thresholds and exposure-response slopes) of Clearigate, Cutrine-Plus, and copper sulfate; and (2) to quantify the response of nontarget species to bioavailable concentrations and forms of copper 7 days after initial herbicide application.

Materials and Methods Test Organism Culture Procedures All test organisms were cultured at the University of Mississippi Aquatic Ecotoxicology laboratory (University, MS). H. azteca culturing procedures followed the methods of de March (1981). Amphipods used for testing were removed from cultures and gently washed through a 1.0-mm mesh sieve. Organisms that passed through the 1.0-mm sieve but were retained by a 0.5-mm sieve (approximately 2–3 weeks old) were collected and used for testing. C. tentans culture methods followed those of Townsend et al. (1981). Midges used for testing were second instar larvae (10 –11 days old). D. magna and P. promelas culturing procedures followed the methods of Peltier and Weber (1985). D. magna were cultured in University of Mississippi Field Station (UMFS) spring water; adjusted for hardness and alkalinity with (0.1 g/L) NaHCO3 and CaCl2 to a total hardness of 80 mg/L as CaCO3 and alkalinity of 60 mg/L as CaCO3, respectively.

Experimental Design All experiments were conducted in light- and temperature-controlled incubators at 20 ⫾ 2°C with a 16 h light/8 h dark photoperiod. Experiments were started by adding 10 H. azteca (2–3 weeks), 6 C. tentans (10 –11 days), 10 D. magna (⬍ 24 h), or 10 P. promelas (⬍ 24 h) to each of three replicate 250-ml borosilicate glass beakers per six test treatments. In other words, sample size (n) for experiments conducted with H. azteca, C. tentans, D. magna, and P. promelas were 30, 18, 30, and 30 organisms per test concentration, respectively. Adjusted UMFS spring water was used as a control and to dilute the test compound to experimental test concentrations. Glass beads (150 – 212 ␮m; Sigma Chemical Co., St. Louis, MO) were used as a substrate in C. tentans experiments to allow for tube building and stress reduction (Suedel and Rodgers 1993). In all experiments, feeding regimes for each organism were as follows: D. magna— 0.5 ml of Selenastrum capricornutum algae daily; H. azteca— 0.5 g wet weight of leached and ground maple leaves at experiment initiation; C. tentans— 0.1 ml cerophyll suspension daily; P. promelas— 0 –20 newly hatched Artemia nauplii per fish daily. Water quality parameters were measured according to Standard Methods (APHA 1992) (Table 2).

Analytical Procedures Stock solutions used for water-only experiments were prepared by dissolving Clearigate (Applied Biochemists Inc., Germantown, WI), Cutrine-Plus (Applied Biochemists Inc.), and reagent-grade copper

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Table 1. Physical properties and fate characteristics of Clearigate, Cutrine-Plus, and copper sulfate Clearigate

Cutrine-Plus

Copper Sulfate

9.0 0.2–1.0b Copper-ethanolamine complex

25.4 0.05–0.5c CuSO4 䡠 5H2O

Chemical classa,b

3.825 0.1–1.0b Copper-ethanolamine in an emulsified complex Chelated elemental copper (Cu2CO3)

Copper salt

Mode of actiona,b Appearancea,b Odorb Water solubility (mg/L)a,b Boiling point (°C)b Melting point (°C)a,b Specific gravity (g/cm3)b pHb Vapor pressure (mm Hg)a,b

Cell toxicanta Blue viscous liquid Orange Miscible 100 N/A 1.1–1.3 9.5–10.0 Nonvolatile

Chelated elemental copper (Cu2CO3) Cell toxicanta Blue viscous liquid Slight amine Complete 100 N/A 1.21 10.0–11.0 Nonvolatile

b

% of Cu as elemental Application Rate (mg Cu/L) Formulationa,b

Cell toxicanta Blue crystalline N/A 316,000 N/A 110 N/A N/A Nonvolatile

a

Kamrin (1997). Applied Biochemists (1997). c Hohman and Martin (1995). b

Table 2. Range (n ⫽ 30) of water quality parameters measured at test termination in experiments conducted with freshwater organisms. UMFS water, adjusted for alkalinity and hardness, was used in all experiments

Temperature (°C) pH Dissolved oxygen (mg/L) Alkalinity (mg/L as CaCO3) Hardness (mg/L as CaCO3) Conductivity (␮mhos/cm)

Control

Clearigate

Cutrine-Plus

CuSO4

19.8–23.5 7.5–8.0 6.5–8.4 66–90 72–80 298–393

19.8–23.5 7.3–8.0 6.5–9.3 62–90 48–96 279–365

19.8–23.5 6.4–8.0 6.4–8.4 55–96 52–96 270–450

19.8–23.5 6.5–8.2 7.5–8.9 50–95 60–100 380–449

sulfate (Fisher Scientific Co., Fair Lawn, NJ) in Milli-Q™ water. Water samples collected for copper analysis were obtained from sacrificial beakers (one replicate per concentration) at the start of each experiment. Total copper concentrations are conserved over time (Kim et al. 1999) and did not change over the course of the exposure period. Water samples were acidified with redistilled nitric acid (Aldrich Chemical Co., Milwaukee, WI) to a pH of 1–2 prior to analysis. Copper concentrations were determined using a Buck model 200-A flame atomic absorption spectrophotometer (0.002 mg Cu/L detection limit). Forty-eight-hour exposures were used to evaluate bioavailability of copper herbicides because Cu2⫹ concentrations in aquatic systems typically fall below detection 48 h after initial application. Test chambers were checked daily, and dead organisms and debris were removed. After the initial 48-h exposure, survivors were tallied, all test organisms were removed, and the test chambers, including controls, were maintained in an incubator for another 72 h. For H. azteca and C. tentans, original maple leaf disks and glass bead substrates were not removed from their original test chambers. Five days after initial herbicide application, a new set of naı¨ve test organisms was introduced to the test chambers for 48 h, as described above.

Statistical Analysis Mean 48-h lethal concentration (LC50) and associated 95% confidence intervals for each replicate toxicity experiment was calculated by trimmed Spearman-Karber analysis (Hamilton et al. 1977). LC50s for test organisms introduced during the last 48 h of the 7-day experiment

period were calculated by probit (Stephan 1977) or trimmed Spearman-Karber analysis. Normality and homogeneity of variance tests were performed using Shapiro-Wilk’s and Bartlett’s tests, respectively. Analysis of variance (ANOVA) and Dunnett’s multiple range tests were used to compare control to treatment survival means and a multiple-comparison method determined differences among mean 48-h LC50 values. All statistical analyses were performed using SigmaStat (Jandel Corporation 1992/94).

Results and Discussion Test Organism Responses to Aqueous Copper Herbicide Exposures Survival of control organisms ranged from 78.0 to 100.0%. Resulting LC50 values are based on total recoverable copper concentrations measured (AA) from sacrificial test chambers (three replicates) at the initiation of each experiment. As illustrated in Table 2, ranges in water characteristics among testing compounds may have resulted in variations in copper concentrations available to testing organisms (e.g., D. magna and P. promelas); increasing water hardness and pH decrease the toxicity of copper (Flemming and Trevors 1989; Deaver and Rodgers 1996). Daphnia magna. Uptake of copper by daphnids occurs through two routes: directly from the water column (i.e.,

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through the gills and feeding apparatus) and indirectly from ingested phytoplankton with sorbed copper (Schuytema et al. 1984). For all three copper herbicides, D. magna was the most sensitive aquatic animal tested. The mean 48-h LC50s for D. magna exposed to Clearigate, Cutrine-Plus, and copper sulfate were 29.4, 11.3, and 18.9 ␮g Cu/L, respectively (Table 3). In similar toxicity experiments that exposed a heterogeneous assemblage of test organisms to copper, cladocerans (e.g., Ceriodaphnia and Daphnia) were the most sensitive species (Schubauer-Berigan et al. 1993; Dobbs et al. 1994). In aqueous studies by Dobbs et al. (1994) and Suedel et al. (1996), comparable mean 48-h LC50 values for D. magna exposed to copper as CuSO4 were 37 and 11.3 ␮g/L, respectively. Hyalella azteca. H. azteca is an epibenthic detritivore, exposed to dissolved metals at external binding sites (i.e., gill surfaces), through surface adsorption (Luoma 1983), and during feeding on metal-bound organic material on the sediment surface. Provision of maple leaf disks in test chambers may have provided additional sites for uptake of copper by sorption and feeding. The amphipod H. azteca was an order of magnitude less sensitive to Clearigate, Cutrine-Plus, and copper sulfate exposure than D. magna with mean 48-h LC50s of 433.4, 247.8, and 157.8 ␮g Cu/L (Table 3). A comparable 48-h LC50 value of 72.2 ␮g Cu/L was reported by Suedel et al. (1996) for H. azteca exposed to CuSO4. Schubauer-Berigan et al. (1993) reported pH-dependent (7.0 – 8.5) 48-h LC50s of 24 – 87 ␮g Cu/L for H. azteca exposed to Cu as Cu(NO3)2. Deaver and Rodgers (1996) reported 10-day LC50 values of 41.8 –142.7 ␮g Cu/L for H. azteca exposed to aqueous CuSO4 concentrations as alkalinity and hardness increased from ⬍ 10 to 64 mg/L as CaCO3. Deaver and Rodgers (1996) concluded that it was difficult to compare H. azteca data from varying sources due to differences in experimental designs, copper compounds tested, experimental waters, and animal care procedures. Chironomus tentans. The midge larvae C. tentans is a benthic infaunal organism that constructs a case from organic material during its larval stage. Copper uptake primarily occurs indirectly via feeding on organic material bound with copper (Giesy et al. 1990). Glass beads in C. tentans test chambers may provide additional sites for copper transfer via sorption. Mean 48-h LC50s of C. tentans exposed to Clearigate and Cutrine-Plus were 373.5 and 460.9 ␮g Cu/L, respectively (Table 3). Copper sulfate was an order of magnitude less toxic to C. tentans than Clearigate and Cutrine-Plus with a mean 48-h LC50 of 1,136.5 ␮g Cu/L. Dobbs et al. (1994) reported a C. riparius 48-h LC50 value of 1,170 ␮g/L when exposed to total dissolved copper from Blaine Creek, KY. Similarly, Gauss et al. (1985) and Suedel et al. (1996) exposed C. tentans to CuSO4 and reported 96-h EC50 and 48-h LC50 values of 997.5 ␮g Cu/L and 529 ␮g Cu/L, respectively. Pimephales promelas. Initially, exposure of copper to P. promelas occurs at external binding sites (i.e., gill surfaces) and is influenced by the rate of water transport across the gills (Phillips and Russo 1978; Pagenkopf 1983; Lauren and McDonald 1986). P. promelas was the least sensitive species tested for Clearigate with a mean 48-h LC50 of 480 ␮g Cu/L (Table 3). The mean 48-h LC50 of P. promelas exposed to Cutrine-Plus was 255.4 ␮g Cu/L. In this study, P. promelas was

B. J. Mastin and J. H. Rodgers, Jr.

Table 3. 48h LC50 values (⫾SD) for D. magna, H. azteca, C. tentans, and P. promelas Organism

Clearigate (␮g Cu/L)

Cutrine-Plus (␮g Cu/L)

Copper Sulfate (␮g Cu/L)

D. magna C. tentans H. azteca P. promelas

29.4 (⫾ 3.8) 373.5 (⫾48.7) 433.4 (⫾45.1) 480.1 (⫾40.4)

11.3 (⫾ 1.2) 460.9 (⫾35.0) 247.8 (⫾76.8) 255.4 (⫾21.5)

18.9 (⫾ 2.3) 1,136.5 (⫾138.6) 157.8 (⫾28.7) 19.2 (⫾3.1)

Clearigate ⫽ six independent experiments used to calculate mean 48-h LC50s; Cutrine-Plus and copper sulfate ⫽ three independent experiments used to calculate mean 48-h LC50s

as sensitive to copper sulfate as the microcrustacean D. magna, with a mean 48-h LC50 of 19.2 ␮g Cu/L (Table 3). Suedel et al. (1996) reported a comparable 48-h LC50 value of 20.2 ␮g/L for P. promelas exposed to CuSO4. At a pH of 7.0, Cusimano et al. (1986) measured the 48-h LC50 for Oncorhynchus mykiss (rainbow trout) exposed to CuCl2 as 7 ␮g/L. Straus and Tucker (1993) reported mean 96-h LC50 values of 54 –983 ␮g/L for Ictalarus punctatus (channel catfish) exposed to copper (8.5% Cu from mixed copper-triethanolamine-diethanolamine complexes) as water hardness increased from 16 to 287 mg/L CaCO3.

Relative Potency of Clearigate, Cutrine-Plus, and Copper Sulfate Based on Response-Slopes of Test Organisms Organism response slopes were calculated using the linearized portion of the exposure-response curves (approximately between 20% and 80% mortality) (Moore et al. 1998). If more than one inflection point existed, response slopes were calculated from the linearized portion of the exposure-response curve that included 50% mortality. Lower threshold responses occurred at the lowest test concentration of herbicide, where mortality was statistically different from the control. In terms of potency, D. magna was the most sensitive of the nontarget animals tested for all three copper herbicides. Due to the miscibility of Clearigate and Cutrine-Plus, one might expect that a planktonic or nektonic organism would be the most sensitive of the animals tested. D. magna had the steepest exposure-response slope and narrowest effects range; D. magna responded greatest to an incremental change in exposure concentration. The 48-h exposure-response slopes of D. magna exposed to Clearigate, Cutrine-Plus, and copper sulfate were 2.55, 8.61, and 5.07% mortality/␮g Cu/L, respectively; 20 –140 times more sensitive than C. tentans (0.13, 0.062, and 0.057% mortality/␮g Cu/L, respectively) (Table 4). Since potencies of these herbicides were compared and contrasted based on slopes of exposure-response relationships after initial thresholds (where slope ⬎ 0), it is important to also consider lower thresholds or intercepts as well as upper thresholds or saturation of the response (mortality in this case) (Figures 1–3). Lower threshold responses (LOECs) of D. magna, C. tentans, P. promelas, and H. azteca to Clearigate were 19.6, 208, 307, and 406 ␮g Cu/L, respectively (Figure 1). Upper thresholds of response to Clearigate exposures with 100% mortality were observed for D. magna, P. promelas, H. azteca, and C. tentans at 78, 1,133, 1,171, and 2,500 ␮g Cu/L, respectively.

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Table 4. Potency of Clearigate, Cutrine-Plus, and copper sulfate with their respective exposure-response slopes (% mortality/␮g Cu/L) Organism Sensitivity More sensitive

1 P P P 2

Less sensitive

Clearigate

Cutrine-Plus

Copper Sulfate

D. magna (2.55) P. promelas (0.21) H. azteca (0.15) C. tentans (0.13)

D. magna (8.61) H. azteca (0.35) P. promelas (0.275) C. tentans (0.062)

D. magna (5.07) P. promelas (2.85) H. azteca (1.0) C. tentans (0.057)

In the case of Cutrine-Plus (Figure 2), lower threshold responses (LOECs) of D. magna, H. azteca, P. promelas, and C. tentans were 7.6, 62.5, 125, and 250 ␮g Cu/L, respectively. An upper threshold response with 100% mortality was only observed for D. magna at 31.2 ␮g Cu/L. Upper thresholds of response (⬎ 90% mortality) were observed for H. azteca and P. promelas exposed to Cutrine-Plus at 1,000 ␮g Cu/L, and 2,000 ␮g Cu/L for C. tentans. Lower threshold responses of D. magna, P. promelas, H. azteca, and C. tentans to copper sulfate were 12.5, 20, 100, and 750 ␮g Cu/L, respectively (Figure 3). Upper thresholds of response (⬎ 92% mortality) were observed for P. promelas and D. magna to copper sulfate at 40 and 50 ␮g Cu/L, respectively. Upper thresholds (100% mortality) were observed for H. azteca and C. tentans at 300 and 6,000 ␮g Cu/L, respectively.

Clearigate and Cutrine-Plus Bioavailability Toxicity experiments are routinely performed to determine the risk of materials for test organisms through time. However, the bioavailability of toxic concentrations and forms of a material may not be accurately assessed when an organism remains in the testing chamber for long periods of exposure (⬎ 96 h). In other words, effects may be a result of exposure in the first or last 24 h and not a result of cumulative exposure. Based on responses of all four test animals to Clearigate and CutrinePlus, a transformation of copper to less bioavailable species over time was observed with a 100 –200% decrease in toxicity (i.e., an increase in 48-h LC50s) even in these relatively simplified water-only exposures. Forty-eight-hour LC50 values for D. magna, H. azteca, C. tentans, and P. promelas and Clearigate increased to 53.9, 712.8, 1,026.6, and ⬎ 1,800 ␮g Cu/L, respectively (Table 5). Survival of D. magna and P. promelas exposed to Cutrine-Plus the last 48 h of a 7-day experiment period increased 100 and 200%, respectively, compared to populations exposed after initial herbicide application. Fortyeight-hour LC50 values for D. magna and P. promelas increased to 22.1 and 634.2 ␮g Cu/L, respectively (Table 5). Suedel et al. (1996) concluded that for D. magna, H. azteca, and P. promelas exposed to copper as CuSO4, sensitivity increased with increased exposure time through a 10-day test period. However, Suedel et al. (1996) did not determine the fraction of the parent compound (i.e., concentration) that was bioavailable at test termination.

Fig. 1. Clearigate 48-h exposure-response curves for D. magna, C. tentans, H. azteca, and P. promelas in order of sensitivity

Summary The experimental design used in this study provides information for designing aquatic herbicides with maximum efficacy for target plant species while not overestimating risks for nontarget species. In natural waters, the propensity of copper to be transferred to ligands (e.g., sediments, suspended solids, flora, fauna, etc.) is generally greater than in simple aqueous testing systems. In closed laboratory systems, total copper concentrations are generally conserved. Although the forms of copper within the testing chambers may transform over time, total copper concentration from initial

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B. J. Mastin and J. H. Rodgers, Jr.

Fig. 2. Cutrine-Plus 48-h exposure-response curves for D. magna, H. azteca, P. promelas, and C. tentans in order of sensitivity Fig. 3. Copper sulfate 48-h exposure-response curves for D. magna, P. promelas, H. azteca, and C. tentans in order of sensitivity

herbicide application will remain constant. Exposing nontarget species to aquatic herbicides containing copper in simple aqueous laboratory experiments provides a worst-case scenario for determining the risk associated with herbicide application. In this study, the 48-h LC50s of D. magna, C. tentans, H. azteca, and P. promelas exposed to Clearigate were 29.4, 373.5, 433.4, and 480.1 ␮g Cu/L, respectively. The 48-h LC50s of D. magna, C. tentans, H. azteca, and P. promelas exposed to Cutrine-Plus were 11.3, 460.9, 247.8, and 255.4 ␮g Cu/L, respectively. Compared with the manufacturers’ recommended application rates of 100 – 1,000 ␮g Cu/L for Clearigate and 200 –1,000 ␮g Cu/L CutrinePlus, the margins of safety are minimal for all four nontarget species during the 48-h after initial herbicide application in simple aqueous exposures. The risk to nontarget species decreases as

knowledge of comparative toxicity, concentration, duration of exposure, form of copper, and water characteristics are factored into translation of margins of safety from the laboratory to the field. Knowledge of bioavailability of toxic concentrations and forms of copper over time (⬎ 96 h) is fundamental in translating these laboratory results to margins of safety for nontarget species in the field. In this study, the bioavailable concentration and forms of Clearigate and Cutrine-Plus were quantified by comparing the responses (i.e., survival) of nontarget animals exposed during the first and last 48 h of the 7-day experimental period. The margin of safety increased 100 –200% for all four

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Table 5. Test organisms introduced 5 days after initial herbicide application. Bioavailability (7 days) as 48-h LC50 values (⫾SD) for D. magna, H. azteca, C. tentans, and P. promelas (three independent experiments used to calculate mean 48-h LC50s) Organism

Clearigate (␮g Cu/L)

Cutrine-Plus (␮g Cu/L)

D. magna C. tentans H. azteca P. promelas

53.9 (⫾ 12.5) 1,026.6 (⫾ 78.2) 712.8 (⫾ 37.1) ⬎ 1,800

22.1 (⫾ 0.4) — — 634.2 (⫾ 103.2)

species exposed to Clearigate. Similarly, the margin of safety for D. magna and P. promelas exposed to Cutrine-Plus increased 100 and 200%, respectively. For long exposure periods (⬎ 7 days), new organisms can be introduced to test chambers for brief (48-h) exposures to evaluate compound bioavailability. More accurate estimates of actual risks from episodic events like herbicide applications can be provided by contrasting appropriate laboratory exposures with the concordant concentrations and forms encountered in field situations.

Acknowledgments. The authors thank Applied Biochemists and H. Knight for herbicide donation. The authors also thank B. Smith for technical assistance.

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