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a Dep. of Zoology, Private Bag X1, Univ. of Stellenbosch, Matieland 7602, South Africa
b Forensic Chemistry Lab., Dep. of Health, Cape Town 8000, South Africa
Corresponding author (R.Schulz{at}tu-bs.de)
Received for publication May 26, 2000.
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ABSTRACT |
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 TOPNOTES ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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Abbreviations: AZI, azinphos-methyl • BBA, Federal Biological Research Centre for Agriculture and Forestry, Germany • END, endosulfan • SDTF, Spray Drift Task Force
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INTRODUCTION |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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Spraydrift is regarded as one of the major routes of nonpoint-source pesticide input into surface waters (Groenendijk et al., 1994). Since spraydrift may result in high concentrations of water-diluted chemicals in surface waters, it usually serves as an exposure scenario for the aquatic risk assessment of pesticides (Gilbert and Bell, 1988). Specifically, orchard mist blowers result in a large amount of drift due to small droplet size and trajectory of release (Groenendijk et al., 1994). According to Payne et al. (1988), the combination of stable boundary layer, light wind, low relative humidity, and high air temperature results in large deposits on downwind water bodies from spray applications employing small drops. Droplet size distribution has been shown to be an important factor influencing the extent of spraydrift; the prevailing view has been that small droplets increase biological efficacy, whereas large droplets reduce downwind drift (Matthews, 1994). Recent results based on a stochastic model for pesticide application through hydraulic nozzles demonstrated that application of small droplets does not necessarily increase field efficacy (Ebert et al., 1999). Furthermore, very small droplets (<50 µm) are most exposed under field conditions to any convective upward air movement and are most liable to travel considerable distances from their source (Matthews, 1994). These atmospheric transport processes may result in contamination of nontarget ecosystems by pesticides applied in agricultural areas situated far away (Le Noir et al., 1999).
In addition to the physical properties of the spray itself, crop characteristics such as height, the amount of open area between trees, diffuse noninterceptance, and leaf area index influence the production and extent of pesticide that may be subject to drift (USEPA, 1999). Spray deposits downwind from orchard sprays reflect the atomizing system, orchard geometry, and seasonal and meteorological conditions, as well as the nontarget surface characteristics (Hall et al., 1996).
There are several generic scenarios for spray drift and spray deposition on surface waters. A large number of standardized drift studies conducted in Germany have been summarized by Ganzelmeier et al. (1995). The results for orchards were differentiated according to early and late growth stage; these were used to derive basic drift values widely used in EU countries for regulatory risk assessment and 95th-percentile values for deposited drift material for distances between 5 and 50 m. Predicted environmental concentrations of a chemical for regulatory exposure assessment purposes are then calculated by relating drift deposit rates per square meter to a water volume of 300 L (30-cm water depth) assuming immediate perfect mixing. Similar depths are used in the Netherlands (25 cm) and Canada (15 cm). The Environmental Fate and Effects Division (EFED) of the United States Office of Pesticide Programs uses a standard value of 5% of the application rate on 10 ha, which deposits on a 1-ha pond (2-m water depth) immediately adjacent to the orchard as an aquatic exposure scenario for airblasts (Aquatic Effect Dialogue Group, 1992). Recently, the Spray Drift Task Force's (SDTF) data set was analyzed and used to develop generic deposition curves with 95% confidence limits for distances between 0 and 549 m (USEPA, 1999), which are proposed for use in risk assessment. Deposition data were grouped into high drift potential orchards and low drift potential orchards.
Fruit orchards form an important agricultural crop in the Western Cape, South Africa, comprising some 440 km2 of growing area, which equals 82% of the orchards in South Africa. In contrast to other extensive fruit-growing areas (e.g., the Central Valley in California, which is protected from sea breeze by the coastal range of mountains), the orchards in the Western Cape are exposed to constantly high southwesterly winds during the pesticide application period (Table 1). This may cause spraydrift to be an important route of pesticide entry into nontarget ecosystems. However, spray deposition has not previously been investigated in this area.
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MATERIALS AND METHODS |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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In the 400-ha orchard area, mainly pears (Pyrus communis L.), plums (Prunus domestica L.), and apples (Malus domestica Borkh.) are grown. The pesticide application period in the study area's orchards proceeds from early August to the end of March. Organophosphorous insecticides, such as AZI (O,O-dimethyl S-[(4-oxo-1,2,3-benzotriazin-3(4H)-yl)methyl] phosphorodithioate) and chlorpyrifos [O,O-diethyl O-(3,5,6-trichloro-2-pyridinyl) phosphorothioate], are applied between October and February quite frequently to pears and plums. Endosulfan (6,7,8,9,10,10-hexachloro-1,5,5a,6,9,9a-hexahydro-6,9-methano-2,4,3-benzodioxathiepin 3-oxide) is applied mainly in apple orchards (Table 2).
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The orchard plots are separated from the Lourens River itself by a strip of vegetation (eucalyptus trees, shrubs, and grasses) between 20 and 100 m in width. In contrast, most of the tributaries are, at least in some stretches, directly bordering on orchard plots (distance: approx. 10–15 m). The Lourens River flows in a southwesterly direction, opposite to the main wind direction coming from the sea southwest of the farming area (Fig. 1). The orientation of the main river parallel to the main wind direction and the vegetated strips along the river make direct spray input into the river highly unlikely. Small side streams forming tributaries of the Lourens River flow more or less at a right angle to the river and to the main wind direction, with no vegetated strips, and are therefore at considerable risk of spraydrift-related pesticide input. Approximately 30% of the tributary surface area is covered by macrophytes, with cattail (Typha capensis Rohrb.) and rush (Juncus kraussii Hochst) predominating.
Application Characteristics
On 27 and 28 Jan. 1999, four trials with application of AZI to bearing pear orchards (average tree height: 6 m) were investigated. Normal spraying events with Jacto Arbus (Sao Paulo, Brazil) air-assisted mist blowers (nozzle type: J5-3; nozzle height: 0.7 to 1.6 m), which delivered AZI at 0.15 kg a.i. ha-1 in 1000 L of water at a pressure of approximately 1200 kPa, were monitored. The manufacturer stated a mean droplet diameter of 125 to 150 µm at the given spray volume and pressure. The registered formulation of AZI was Guthion (350 g kg-1 a.i.), a wettable powder with a recommended application rate of 0.5 kg ha-1. On 2 Feb. 1999, three trials with application of 1.425 kg a.i. ha-1 END to bearing apple orchards (average tree height: 4.5 m) were investigated. Thioflo, an emulsifiable concentrate (475 g L-1 a.i.), was used to supply the END, recommended at 3 kg ha-1 of product. Canopy and spacing characteristics in both orchard plots were of moderate density, with no space between trees.
Wind speed was measured during each trial at a height of 1.5 m above the ground in the tree row and 5 m downwind of the sprayed tree row using a portable anemometer (Ferropilot, Rellingen, Germany). Wind speeds were averaged for 1-min intervals from five measurements. During the AZI trials the average wind speed was 1.7 ± 0.1 m s-1, relative humidity 81.8 ± 0.8%, and air temperature 18.9 ± 0.4°C. The respective values for the END trials were 4.5 ± 0.2 m s-1, 77.6 ± 0.8%, and 20.9 ± 0.3°C.
Sampling Setup
Spray deposition during applications was studied at orchard plots adjacent to two different tributaries of the Lourens River situated downwind of the plots (Fig. 1). Distances of the tributaries from the edge of the treated area were 10 and 15 m for the AZI and END trials, respectively. Height of the vegetation layer in the area between orchard plots and tributary including stream banks was 
25 cm. Average width and depth of the tributaries was 0.89 x 0.30 m for the AZI trials and 1.13 x 0.39 m for the END trials. The current velocities were approx. 0.1 m s-1 in both tributaries.
Drift deposit was sampled at distances of 0, 5, 10 (tributary site), and 15 m downwind from the edge of the treated area during the AZI spraying (Fig. 1) and only at 15 m (tributary site) during the END spray application. Two replicate collectors were employed at each distance per trial, giving a total of eight replicates at each distance for AZI. A total of six replicates were performed only at the 15-m distance for the END spray application trials.
The drift deposit collectors consisted of acetone- and distilled water–rinsed flat straight-sided glass bowls containing 300 mL of distilled water and providing a surface area of 75 cm2 at a water depth of 4 cm. The sampling bowls were set horizontally on the ground. Vegetation in the vicinity of deposit samplers was removed to eliminate the possibility of spray interception. The bowls exposed at 10 m during the AZI trials, like those exposed at 15 m during the END trials, were supported 5 cm above the stream water surface by wooden supports. These bowls represented the drift deposition on the stream surface. Their collection surface was lower, due to the fact that the stream water surface was approx. 1 m below the average ground level. Following the spray event, the contents of the bowls were thoroughly stirred with a clean glass rod, poured into acetone- and distilled water–rinsed glass jars, and kept at 4°C in the dark until solid phase extraction was carried out.
Orchard tree rows were orientated perpendicular to the tributary, so that a more or less well-defined cloud of spraydrift moved from the orchard in the direction of the tributary. Each time the spraying machine turned around at the end of a row, it was observed in the field that between 10 and 15 m of stream length were exposed to spraydrift. Spraying was stopped when the spraying machine reached the end of one row and commenced again at the beginning of the next row (Fig. 1). In addition to the drift deposit collection, two different types of water samples were taken in 3-L glass jars from the tributaries in stretches without macrophyte coverage. First, 1-h composite samples (by combining 150-mL samples taken every 10 min) were collected at a site approximately 50 m downstream of the spray-receiving stretch of the tributary while the spraying took place on the adjacent orchard plots. Second, discrete water samples were taken approx. 30 s after the chemical had reached the water surface (visual determination) from the downstream section of the tributary stretch, which was covered with spray deposition (Fig. 1). All samples were taken by dipping closed sampling jars into the water column and opening the jars approx. 10 cm below the water surface to avoid contamination with surface film. The samples thus represent subsurface concentrations. Both types of samples were replicated six times during the AZI trials and three times during the END trials. The composite samples represent the average stream contamination for a time period of 1 h, whereas the discrete samples are intended to contain the potential peak pesticide concentrations, once the chemical has reached the water body. However, both types of samples may contain a certain amount of additional contamination due to airborne pesticide transport from orchard areas further away from the stream. To determine the potential contamination of the Lourens River, which receives the discharge of the investigated tributaries, 5-h composite samples (100 mL every 10 min) were taken at a site downstream of the farming area. This site was approx. 2.5 km downstream of the inlets of the two tributaries used for spraydrift monitoring. The river at this site has an average discharge in January of 0.28 m3 s-1. Sampling in the Lourens River was carried out on three days with spray application in the catchment and on three days without any spray application. On spraying days, pesticide application took place in parallel on approx. three plots adjacent to tributaries of the Lourens River.
Distilled water (300 mL), tributary, and river water samples (700 to 900 mL) were solid-phase extracted (SPE) within 10 h after sampling using Chromabond C18 columns (Macherey–Nagel, Dueren, Germany) that had been previously conditioned with 6 mL methanol and then 6 mL water. Samples were not filtered prior to solid phase extraction since all samples represent clear water samples with total suspended solid contents of less than 10 mg L-1. The columns were air-dried for 30 min and kept at -18°C until analysis.
Pesticide Analysis
Analysis was performed at the Forensic Chemistry Laboratory of the Department of National Health, Cape Town. Water samples were eluted from SPE columns with 2 mL hexane and then 2 mL dichloromethane. These extracts were dried in a stream of nitrogen and then taken up in 1 mL hexane. The hexane solutions of water samples were analyzed using gas chromatograph–electron capture–nitrogen/phosphorous detector (GC–ECD–NPD; 63Ni ECD temperature: 300°C with nitrogen as make up gas, NPD temperature: 300°C). The Hewlett–Packard (Palo Alto, CA) HP 5890 Series II gas chromatograph was equipped with an HP 7673 autosampler and a split–splitless injector and HP 5 capillary column (15-m length, 0.32-mm i.d., 0.25-µm film thickness) and with nitrogen as carrier gas (1.1 mL min-1). The temperature programs were: 170°C (1 min) 
(20°C min-1 300°C (1 min); 5 µL was injected with the splitter closed for 0.75 min. Measurements were confirmed using a gas chromatograph–flame photometric detector (GC–FPD; FPD temperature: 250°C). The HP 5890 gas chromatograph was equipped with an HP 7673 autosampler and a split–splitless injector and DB 210 capillary column (30-m length, 0.32-mm i.d., 0.25-µm film thickness; J&W) and with nitrogen as carrier gas (1 mL min-1). The temperature programs were: 150°C (0.5 min) 30°C min-1 210°C (1 min) 30°C min-1 240°C (1 min); 5 µL was injected with the splitter closed for 1 min.
Method validation was conducted on water matrices that were determined to have no detectable levels of AZI or END. The validation consisted of spiking water at eight spiking levels over the range of concentrations found in the actual samples. Overall mean recoveries were between 79 and 106%. For quality control, a matrix blank was analyzed with each extraction set. No AZI and END was detected in matrix blanks. The detection limits were 0.01 µg L-1.
Toxicity Tests
One additional bowl, containing water taken from the tributary before spraying, was placed next to the sampling bowls at the above-mentioned distances during each of the four AZI trials. Additionally, samples of the water taken from the tributary during spraying and from the Lourens River on spray application days were used for toxicity tests. Tributary water taken before spraying and more than 48 h after the last spraying anywhere in the catchment had taken place served as a control. The quantified characteristics of the waters are given in Table 3. The metal content of those waters (aluminium, copper, zinc, mercury, and lead) was lower than detection limits (0.005–0.25 mg L-1). Test water was taken to the laboratory and temperature-equilibrated in a water bath. Tests were commenced within 4 h after sampling. During the tests the jars were not aerated, and temperature, pH, and dissolved oxygen were monitored. Static 24-h acute toxicity tests were performed in 1-L glass jars following the general procedures of acute aquatic toxicology (Sprague, 1969, 1970). Four replicates each containing 500 mL of test substance were performed for each sample. Short-term exposures were employed because they most closely represent the "pulse" exposure typical of contamination from operational sprays.
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RESULTS AND DISCUSSION |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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Spray deposition measured for AZI was compared with median basic drift values used in exposure assessment for orchard spraying in Germany and those suggested for use in the USA (Fig. 3). Measured deposition rates at distances of 10 and 15 m were approximately 25 and 59% lower than these basic drift values, which are in fact very similar in the BBA and USEPA approach for distances up to 20 m. One possible reason might be that the average wind speed during AZI application was only 1.7 ± 0.1 m s-1, so it cannot be considered a worst-case scenario. Average wind speeds in the BBA trials with late growth stage orchards (downwind of plot) and in the SDTF trials with high-drift orchards (inside orchard) were 3.1 ± 0.4 and 2.8 ± 0.5 m s-1. Our wind speed values were all obtained at approx. mid-crop height downwind of the orchard plots and should therefore be comparable with inside-orchard readings. At distances of 0 to 5 m, measured deposition of AZI was not appreciably different from BBA and SDTF estimates (Fig. 3).
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Resulting In-Stream Concentrations
Measured tributary water peak concentrations from discrete sampling were compared with the calculated concentrations based on deposit in the water bowls (Fig. 4 and 5). To calculate the concentrations, deposition rates (mg m-2) measured at 10 m in the AZI trials and at 15 m in the END trials were related to the water volume covered by 1 m2 tributary surface area (for the AZI tributary 300 L and for the END tributary 390 L).
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The maximum END concentration in the tributary directly after settling of spraydrift was calculated, on the basis of an average stream depth of 0.39 m, to be 13.0 ± 0.54 µg L-1 (Column B1 in Fig. 5), which is approx. 3 µg L-1 higher than the measured peak concentration in the tributary (10.1 ± 1.2 µg L-1; Column B2 in Fig. 5). The estimated in-stream concentrations according to the basic drift values given by Ganzelmeier et al. (1995) and SDTF (USEPA, 1999) are 9.1 µg L-1 (Column A in Fig. 5) and 12.1 µg L-1, respectivley, which are slightly different from the measured peak concentrations. The measured 1-h average concentration of END in the tributary was 0.9 ± 0.16 µg L-1 (column C in Fig. 5).
The fact that for both pesticides the measured and calculated in-stream concentrations were very similar is evidence of the efficiency of the collection methods and their suitability for spray deposit measurements. It has been shown that results from water-filled bowls are also well in accordance with those from glass fibre filter collectors (Ernst et al., 1991).
Pesticide spray deposit produced detectable contamination in the drift-receiving tributaries. The subsurface peak concentrations of AZI (0.3 m stream depth; 1.68 µg L-1) and END (0.39 m stream depth; 10.1 µg L-1) measured in tributary water (B2 in Fig. 4 and 5) were slightly lower than estimates based on Ganzelmeier et al. (1995) for AZI (2.2 µg L-1; A in Fig. 4), and slightly higher than estimates for END (9.1 µg L-1; A in Fig. 5). Apart from physical spray parameters, the differences in wind conditions, as discussed above, may be one reason for this. An assessment of exposure based on EFED (Aquatic Effect Dialogue Group, 1992) would, in contrast, predict much lower contamination if the usual depth of 2 m is used (0.4 and 3.6 µg L-1). On the other hand, an EFED assessment using 0.3 m water depth would overestimate the AZI trials (2.5 µg L-1) and strongly overestimate the END contamination (23.7 µg L-1). A prediction based on SDTF's data set and 0.3 m water depth gives 2.4 µg L-1 for AZI, and 12.1 µg L-1 for END.
The derivation of worst-case scenarios is usually based on the assumption of immediate perfect mixing of the deposited chemical into the water column, which is probably not the case under field conditions. The water depth used for these calculations is of particular importance. The subsurface peak tributary concentrations measured during this study compare quite well with calculations based on a 0.3-m-deep water body assuming perfect mixing, while a depth of 2 m would lead to a considerable underestimation of the concentrations. Both current and macrophyte vegetation might contribute to a relatively fast dissipation of the chemical into the water column. A different situation may prevail when a body of standing water receives spray drift.
Pesticide Concentrations in the Lourens River
Both pesticides occurred in the Lourens River at increased 5-h average levels during days with pesticide spray application to fruit orchards in the river catchment compared with those without any spraying. On days when field plots were sprayed, the AZI concentration was increased by more than a factor of 4, from nondetectable levels (<0.01 µg L-1) to 0.041 µg L-1, while the increase for END amounted to a factor of 11, from 0.006 to 0.067 µg L-1 (Table 4). The fact that END was detectable even during days without spray application in the catchment may be related to the tendency of END to accumulate in the environment, as it is not readily detoxified by soil microorganisms (Goebel et al., 1982).
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Ecotoxicity Testing
Mortality of midges exposed for 24 h during the AZI field trials decreased with decreasing pesticide concentration in the samples (Table 5). Mortalities of 56.3 and 45% occurred at AZI concentrations of 17.2 and 5.1 µg L-1, respectively.
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For both endpoints, the effects decreased with increasing distances from the sprayed orchard, which is in accordance with the measured concentration levels, as well as with the results of other studies employing aquatic or terrestrial test species (Davis et al., 1994; Helson et al., 1993).
It follows from the results with field samples that a mortality of 50% occurred at an estimated concentration of aproximately 10 µg L-1, which equals a distance of aproximately 13 m downwind from the edge of the sprayed area. This estimated field LC50 compares fairly well with the 24-h LC50 (95% C.I.) obtained from spiked water tests with AZI in the laboratory: 7.3 (5.7–9.9) µg L-1. This laboratory-measured 24-h LC50 implies that the unidentified Chironomus species we used is less sensitive than Chironomus tentans Fabricius, which has a 96-h LC50 of 0.37 µg L-1 (Ankley and Collyard, 1995). The estimated EC50 for tube formation in the field samples is approximately 2.4 µg L-1. It also compares fairly well with the 24-h EC50 (95% C.I.) of 2.0 (1.7–2.4) µg L-1 obtained from spiked water tests.
The fact that the toxicity data obtained from field samples indicate LC50 and EC50 concentrations similar to those found from laboratory data gives evidence for the comparability of the two test designs. A potential reason might be that water samples from surface waters contaminated by drift are quite comparable in their physicochemical conditions (e.g., turbidity and bioavailability) with those used in the laboratory.
Mortality in the tributary samples averaged 11%, while no mortality was discernible in the Lourens River samples. Failure of tube formation in the tributary water samples ranged between 22 and 40%, and was thus significantly increased compared with the control group with a level of 1.2%. Failure of tube formation was also significantly increased in the Lourens River samples in 12% of the test organisms.
The results of this study indicate that the transient (1 h) pesticide concentrations measured in tributary water were in the range of concentrations that require longer exposure periods in order to be acutely toxic for a number of other test species (Table 2). For example, the 48-h EC50 of 1.6 µg L-1 AZI for water flea (Daphnia magna) (Dortland, 1980) as well as the 96-h LC50 of 0.3 µg L-1 END for rainbow trout (Oncorhynchus mykiss) (Lemke, 1981) and the 24-h LC50 of 7.75 µg L-1 END for threespined stickleback (Gasterosteus aculeatus L.) (Ernst et al., 1991) were exceeded. The detected END levels in tributary water and in the Lourens River on spraying days exceed the target water quality range (TWQR) of <0.01 µg L-1 established by the South African Department of Water Affairs and Forestry (Department of Water Affairs and Forestry, 1996).
Based on the measured short-term peak concentrations, toxic effects in the tributaries or the Lourens River as a result of spray deposit in its tributaries are unlikely. However, much concern has been voiced about potential chronic effects following short-term exposure of aquatic organisms (Hosmer et al., 1998; Liess and Schulz, 1996; Schulz and Liess, 2000). Ecological effects of pollution in Western Cape rivers have to be considered carefully since many of the aquatic invertebrate and fish species present in the rivers are endemic to a relatively small area (Davies and Day, 1998), and their extinction cannot be compensated by recolonization from other regions.
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ACKNOWLEDGMENTS |
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NOTES |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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REFERENCES |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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