Ecological Impact in Ditch Mesocosms of Simulated Spray
Drift from a Crop Protection Program for Potatoes
Gertie HP Arts,*Laura L Buijse-Bogdan,J Dick M Belgers,Caroline H van Rhenen-Kersten,Rene PA van
Wijngaarden,Ivo Roessink,Steve J Maund,`Paul J van den Brink,and Theo CM Brock
Alterra, Wageningen University and Research Centre, 6700 AA Wageningen, The Netherlands
`Syngenta Crop Protection AG, 4002 Basel, Switzerland
(Received 30 December 2004; Accepted 12 July 2005)
Outdoor aquatic ditch mesocosms were treated with a range of pesticides to simulate various spray drift rates resulting
from a typical crop protection program used in the cultivation of potatoes in The Netherlands. The main experimental aims
of the present study were to provide information on the fate and ecological effects of drift of the pesticides into surface
water and to evaluate the effectiveness of drift-reduction measures in mitigating risks. The pesticides selected and the
dosage, frequency, and timing of application were based on normal agricultural practices in the potato crop. Applications of
prosulfocarb, metribuzin (both herbicides), lambda-cyhalothrin (insecticide), chlorothalonil, and fluazinam (both fungicides)
were made in the sequence typical of the spray calendar for potatoes. A total of 15 treatments with the various compounds
were made by spray application to the water surface at 0.2%,1%, and 5%of the recommended label rates. Chemical fate
and effects on ecosystem function and structure (phytoplankton, zooplankton, chlorophyll-a, macroinvertebrates,
macrophytes, breakdown of plant litter) were investigated. To interpret the observed effects, treatment concentrations
were also expressed in toxic units (TU), which describe the relative toxicity of the compounds with standard toxicity test
organisms (Daphnia and algae). After treatment, each compound disappeared from the water phase within 2 d, with the
exception of prosulfocarb, for which 50%dissipation time (DT50) values ranged between 6 and 7 d. At the 5%treatment
level, an exposure peak of 0.9 TU
was observed, which resulted in short-term responses of pH, oxygen, and
phytoplankton. At the 5%treatment level, exposure concentrations also exceeded 0.1 TU
, and this resulted in long-
term effects on zooplankton and macroinvertebrates, some of which did not fully recover by the end of the present study. At
the 1%treatment level, only slight transient effects were observed on a limited number of zooplankton and macro-
invertebrate species and on pH. At the 0.2%level, no consistent treatment-related effects were observed. Most of the
observed effects were consistent with the results from higher-tier and mesocosm studies with the individual compounds.
Multi and repeated stress played a small role within the applied pesticide package, because of rapid dissipation of most
substances and the absence of many simultaneous applications. This suggests that risk assessments based on the individual
compounds would in this case have been sufficiently protective for their uses in a crop protection program.
Keywords: Mesocosms Crop scenario Pesticide package Invertebrates Risk assessment
The use of pesticides in agriculture may lead to the entry
of small amounts of these compounds into aquatic ecosys-
tems through a variety of potential routes including spray
drift, drainage, surface runoff, or accidental spills (FOCUS
2001). In the flat, polder landscape of The Netherlands,
spray drift into edge-of-field ditches is considered to be the
main entry route of pesticides to surface water (Beltman and
Adriaanse 1999). Here, these pesticides can exhibit undesir-
able side effects on the survival and recruitment of nontarget
aquatic organisms. The entry rate of spray drift depends on
the crop, application technique, implementation of risk
mitigation measures, and environmental conditions on the
day of application. Subsequent aquatic exposure patterns
then depend on the number and frequency of applications
and the dissipation rate of the compound (Boxall et al.
Regulatory schemes for pesticides are currently based on
assessments of individual compounds (see EU 1997; USEPA
1998). It is common practice, however, for several different
pesticides to be applied to protect crops and for their uses to
be simultaneous and/or repeated. Only relatively few studies
have investigated the effects of combinations of pesticides in
aquatic mesocosm experiments (see Fairchild et al. 1994;
Van den Brink et al. 2002; DeLorenzo and Serrano 2003),
and such combinations of compounds are often not related to
treatment programs used in specific crops. Only 1 previous
study describes a crop-based approach to assess the combined
and repeated effects of pesticides: The effects of a crop-
protection program for flower bulb crops (Van Wijngaarden
et al. 2004).
In this article, we describe an experiment to evaluate the
fate and effects in outdoor ditch mesocosms of a typical
(with respect to compounds, treatment rate, frequency, and
timing) crop-protection program for potatoes in The Nether-
lands. Potato crops cover a large area of The Netherlands
(Deneer 2003), and because of the disease pressure they face
(e.g., blights such as Phytophtora infestans), are characterized
by an intensive use of pesticides (see Caux, Bastien, et al.
1996). The spray drift treatment rates selected followed
Dutch regulatory scenarios. The mesocosms are representa-
tive of the macrophyte-dominated drainage ditches that are
typical in the Dutch agricultural landscape and harbor
populations of aquatic organisms representative for several
*To whom correspondence may be addressed: firstname.lastname@example.org
Integrated Environmental Assessment and Management — Volume 2, Number 2—pp. 105–125
Ó2006 SETAC 105
MATERIAL AND METHODS
Experimental systems and study design
The experiment was performed in 12 ditch mesocosms,
each with a length of 40 m, a width of 3.3 m at the water
surface and 1.6 m at the sediment surface, a water depth of
0.5 m, a sediment depth of 0.25 m, and a total volume of
approximately 55 m
(Drent and Kersting 1993). Ditch
sediment consisted of sandy loam with a moderate nutrient
content. Construction and technical details of the mesocosms
are described by Drent and Kersting (1993) (Figure 1). The
systems and their aquatic community structure resemble
shallow, macrophyte-dominated drainage ditches on sandy
loam and clay in the agricultural landscape. The mesocosms
are not dominated by free-floating Lemna species, as is
predominantly the case in the hypertrophic ditches in the
agricultural Dutch landscape. Therefore, they may suffer from
a higher aquatic exposure to pesticides entering the systems
by spray drift compared with that of Lemna-dominated
To increase similarity between ditch mesocosms (and hence
increase experimental power), 2 measures were taken at the
start of the experiment. First, ditches were reset to a pioneer
stage by removal of the organic detritus layer in the autumn of
2000. After removal, recovery of macrophytes and macro-
fauna was stimulated by introduction of Elodea nuttallii shoots
and macroinvertebrates originating from ditches that were
situated in an area that was not in agricultural use
(‘‘Veenkampen’’). Gammarus pulex was derived from the
groundwater basin at the experimental station. In 2001, fine
detritus and macroinvertebrates from the same area were
seeded into the ditches for a 2nd time. In 2002, the year in
which the experiment was performed, extra individuals of
Asellus aquaticus and Proasellus juv. from ditches in the
Veenkampen area were added to the ditches in the pretreat-
ment period. Second, water was circulated between the
systems for about 1 month in the pretreatment period.
During the experiment, however, each ditch was isolated
from the others, meaning that both duration of exposure
(owing to lack of dilution) and external recovery of organisms
by water (owing to lack of supply of diaspores and organisms)
were more or less the worst case.
The experimental setup followed a regression design.
Because spray drift is considered to be the main emission
route of pesticides to surface water in the flat polder
landscape of The Netherlands, we applied spray drift rates
that varied from 0.2% to 5% of label-recommended rates of
pesticides. Four spray drift scenarios were applied: 0%, 0.2%,
1%, and 5%. Four ditches served as control ditches. Two
ditches received a low spray drift emission of 0.2%, which is
the percentage that can be achieved if several mitigation
measures are applied simultaneously (Ganzelmeier et al.
1995; Huijsmans et al. 1997). In 3 ditches, a spray drift
emission of 1% was applied, which is the drift emission rate
used for regulatory purposes for crops such as potatoes. In 3
ditches, a spray drift emission of 5% was applied, which is the
spray drift emission without mitigation measures (Ganzel-
meier et al. 1995; Huijsmans et al. 1997). The last scenario is
meant as a realistic worst-case. The 4 treatments were
Figure 1. Photo of ditch mesocosms.
106 Integr Environ Assess Manag 2, 2006—GHP Arts et al.
randomly assigned to 1 of 3 rows of ditches at the field station
(randomized block design).
The typical pesticide treatment regimes for ware and starch
potatoes (Deneer 2003) were combined to develop a typical
application scenario for potatoes based on active ingredients
currently on the Dutch market. In potatoes, herbicides are
applied at the beginning of the cultivation period, followed by
various fungicide and insecticide treatments. Before harvest,
aboveground crop plants are removed either mechanically or
herbicidally; in this case, we assumed mechanical removal and
did not apply an herbicide at the end of the season.
The pesticides used and the sequence and frequency of
their application in the spray program are presented in Table
1. The use rate of the compounds was based on the label
recommendations for 2003. The treatment program included
prosulfocarb, metribuzin, lambda-cyhalothrin, chlorothalonil,
and fluazinam (Table 1). Prosulfocarb is a thiocarbamate
herbicide that inhibits weed shoot growth (Tomlin 2000).
Metribuzin is a triazinone herbicide that inhibits photo-
synthetic electron transport and is widely used for the control
of grasses and broadleaf weeds in numerous crops (Fairchild
and Sappington 2002). Lambda-cyhalothrin is a synthetic
pyrethroid insecticide that disrupts ion channels in nerves
(Maund et al. 1998). Fluazinam is a dinitroaniline fungicide
that uncouples mitochondrial oxidative phosphorilation in
fungi (Tomlin 2000). Chlorothalonil is a widely used,
nonsystemic chlorophenyl fungicide used to control various
plant diseases in wetter regions in the United States (Caux,
Kent, et al. 1996) and in The Netherlands. At least part of
chlorothalonil toxicity is due to respiratory disruption (Davies
Endpoints in the ditch mesocosms (Table 2) were inves-
tigated for a period of 30 weeks (18 March–14 October
2002), including a pretreatment period of 5 weeks, a
treatment period of 18 weeks, and a posttreatment period
of 7 weeks. In the pretreatment period, all endpoints were
sampled once or twice to establish starting conditions and to
check sampling methods. In the posttreatment period, all
endpoints were sampled at least twice. Measurements and
sampling activities (fate sampling, artificial substrates and
litter bags, macrophyte bioassays, phytoplankton and zoo-
plankton sampling and measurement of water quality
endpoints) were nonrandomly assigned to specific ditch
segments in order to avoid mutual disturbance. On each
sampling date, measurements and sampling activities were
located in their assigned, specific ditch segments. Endpoints
and sampling frequencies are presented in Table 2.
Pesticide application and sampling
To mimic spray drift deposition of pesticides on the water
surface of the ditch mesocosms, pesticides were applied by
means of a shielded spray boom (Design 1990 Technical
Department of Wageningen University and Research Centre,
formerly Institute of Agricultural and Environmental Engi-
Table 1. Application scheme, use rates, emission levels, calculated intended target concentrations, and nominal
concentrations of pesticides (lg/L active ingredients) in the ditches
Use rate (g/ha) Concentration
Emission level (lg/L)
PEC 5%drift PEC 1%drift PEC 0.2%drift
0 3200 Target 76 15 3.0
Nominal 76.4 15.9 3.2
2 350 Target 8.3 1.7 0.33
Nominal 8.2 1.5 0.27
5 and 9 5 Target 0.120 0.024 0.0048
6–9 1010 Target 24 4.8 0.96
10–17 200 Target 4.8 0.95 0.19
Number of weeks after first pesticide (prosulfocarb) application.
Target concentrations were calculated according to the method outlined below. Nominal concentrations were based on measurements in
the spray solution, the amounts of solution sprayed, and the estimated amount of water (55 m
) in the ditches. The range in nominal
concentrations is given in brackets.
The following calculation method for predicted environmental concentrations (PEC) values is based on the PEC calculation for a standard
ditch as described in Beltman and Adriaanse (1999). The given example is a 5%emission of prosulfocarb.
Assume a ditch width of 1.0 m and a depth of 0.3 m with a ditch width at the sediment surface of 0.4 m and a side-slope angle of 1:1
(horizontal:vertical). By use of these assumptions, the ditch capacity per metre is calculated by
0.4 0.3 þ20.3 0.5 0.3 ¼0.21 m
/m ¼210 L/m
where the use rate of prosulfocarb ¼3,200 g active ingredient (A.I.)/ha ¼0.32 g A.I./m
¼320 mg A.I./m
sediment surface. Assuming a 5%
emission rate, 0.05 320 mg A.I./m
sediment surface ¼16 mg/m
sediment surface. Therefore, 16 mg/m
sediment surface divided by
¼0.076 mg/L ¼76 lg/L.
Impact in Ditch Mesocosm of Simulated Spray Drift—Integr Environ Assess Manag 2, 2006 107
neering) (Ronday et al. 1998). By this method, the desired
spray volume is dosed accurately and evenly over the water
surface of each ditch and the risks of cross-contamination are
minimized. The spray boom is shielded from the wind to
enable its use irrespective of wind conditions. To prevent
prolonged stratification of pesticide concentrations in the top
layer after application, we mixed the upper water layer by
moving a metal plate, which was fixed to the platforms of the
movable bridges and extended to a water depth of approx-
imately 10 to 20 cm, from one end to the other end of the
Prosulfocarb was applied as the formulated product Boxert
(806 g/L prosulfocarb; density, 1.020 g/mL; Syngenta, UK).
Metribuzin was applied as 70% metribuzin (powder; DIC
1468 70 WG, Bayer, Germany). Lambda-cyhalothrin was
applied as Karate Zeont(98 g/L lambda-cyhalothrin; density,
1.064 g/mL; Zeneca Crop Protection, UK). Chlorothalonil
was applied as the formulated product Bravot(40.4% w/w
chlorothalonil; density, 1.254 g/mL; formulation YF 10934,
Syngenta). Fluazinam was applied as the formulated product
Shirlant(39.5% w/w fluazinam; density, 1.273 g/mL;
Syngenta). In the laboratory, a known amount of the
compound was weighed into a brown glass bottle, and
tapwater was added. The glass bottles were transported to
the experimental site, and just before application, the entire
contents of a bottle were transferred to a spraying vessel that
was filled with groundwater to an end volume of 12 L,
weighed, and thoroughly mixed. The ditches were sprayed for
about 60 to 70 s at a pressure of 3.5 bar, resulting in a spray
volume of 7 L. After the application, the spraying vessel with
remaining spray solution was weighed to calculate the amount
that was sprayed. Two subsamples were transferred from the
remaining spray solution directly into high-pressure liquid
chromatography (HPLC) vials for measuring the concen-
tration of the application solution.
The intended treatment concentrations were calculated
from product use rates, emission percentages and predicted
environmental concentration (PEC) calculations for the
standard Dutch ditch (Table 1) (Beltman and Adriaanse
1999). Initial concentrations in the ditches were calculated
from residue measurements in the spray solution, the
amounts of solution sprayed, and the estimated amount of
water (55 m
) (Drent and Kersting 1993) in the ditches
(Table 1). To measure exposure concentrations, water samples
were collected at intervals after application (2, 8, 24, 72, 168,
336, and 672 h) in the water compartment of each ditch
mesocosm or over shorter periods if dissipation was antici-
pated to be rapid. Depth-integrated water samples were
collected with a Perspex tube (length, 50 cm; internal
diameter, 3.9 cm) at 3 fixed sampling locations in each ditch.
The tube was inserted vertically into the water column and
collected from all ditches in this way. The water samples
were transferred into a 1-L glass beaker and were thoroughly
mixed. A subsample was then transferred into a 500-mL flask,
which was sealed and transferred to the laboratory for
extraction within 2 h.
Prosulfocarb, metribuzin, and fluazinam were extracted
from the water samples by solid-phase extraction with OASIS
extraction columns. The bottle containing the water sample
was weighed before extraction. The OASIS columns (Waterst
preconditioned with 1 mL acetonitrile (LAB-SCAN HPLC
quality) and 5 mL distilled water for prosulfocarb and
metribuzin and with 2 mL of LAB-SCAN methanol (HPLC
quality) and 5 mL distilled water for fluazinam. The bottle
with the water sample was connected to the extraction
column with stainless steel tubing, and water was filtered
through the column. After extraction, the bottle with water
sample was weighed again to measure the amount of water
extracted. Prosulfocarb, metribuzin, and fluazinam were
eluted from the extraction columns with acetonitrile (2 3
0.5 mL) into volumetric tubes. For each compound, both
eludates were collected in a 10-mL graduated tube, and the
samples were diluted with distilled water to a known final
volume. A subsample was transferred into an HPLC vial, and
the samples were analyzed on an HPLC system equipped
with a ultraviolet detector.
Table 2. Ecological endpoints and sampling frequencies
Endpoint Frequency once in:
Sampling (combined with phytoplankton and chlorophyll-a) 3 weeks
Oxygen, pH, temperature
Alkalinity and conductivity
Concentration in water 3 weeks
108 Integr Environ Assess Manag 2, 2006—GHP Arts et al.
Table 3 summarizes the parameters used for the HPLC
analysis. Concentration calculations were based on external
standard samples. Recovery efficiencies from spiked water
samples (control ditches) were 92.2% (n¼15; SD 7.9%),
96.6% (n¼11; SD 3.1%), and 83.9% (n¼59; SD 3.3%) for
prosulfocarb, metribuzin, and fluazinam, respectively. Con-
centrations were corrected for recovery.
Lambda-cyhalothrin and chlorothalonil were extracted
from water samples by liquid–liquid extraction with either
hexane (own distillate; lambda-cyhalothrin) or toluene
(Across Organics, Belgium, toluene pro analysis; chlorotha-
lonil). A known volume (;300 mL) of water was transferred
to a flask, and 35 mL of petroleum ether was added. The
flasks were closed and shaken thoroughly for at least 15 min.
The organic layer was concentrated by evaporation of the
petroleum ether (own distillate) on a rotary evaporator, and
the residue was dissolved in 1.5 mL hexane. Chlorothalonil
samples (;20 mL water) were extracted with 2 mL toluene
by shaking thoroughly for 15 min in 35-mL tubes. Samples
were analyzed without cleanup by gas chromatography with
an electron capture detector. Table 3 summarizes the
parameters used for the gas chromatography and electron
capture detector analysis. Concentration calculations were
based on external standard samples. Recovery efficiencies
from spiked water samples (control ditches) were 97.5% (n¼
3; SD 3.7%) for lambda-cyhalothrin and 104.9% (n¼12; SD
10.6%) for chlorothalonil. Concentrations were corrected for
The time required to dissipate 50% of the mass originally
present (DT50) values were calculated according to the
Organization for Economic Cooperation and Development
(OECD) (2000) guideline 308 and were based on linear least-
Table 3. Parameters for high-pressure liquid chromatography and gas chromatography/electron capture detector analysis of
the pesticides prosulfocarb, metribuzin, fluazinam, lambda-cyhalothrin, and chlorothalonil
High-pressure liquid chromatography analysis
Prosulfocarb Metribuzin Fluazinam
Model Waters M590 þPerkin Elmer ISS 100 autosampler þWaters LC 90 ultraviolet detector
Injection volume 100 lL
Column Waters XTerra
MSC 18, 3.5 lm, 4.6 3150 mm Waters NovapaktC-18, 4 lm,
Guard column Waters XTerra MSC 18, 3.8 lm, 3.9 320 mm Waters Novapak C-18, 4 lm,
3.9 320 mm
Mobile phase Acetonitrile, 65%
Acetic acid, 0.1%
Flow 1 mL/min
Oven temperature 408C
Wavelength (nm) 220 296 260
Retention time (min) 8.4 8.2 4.1
Detection limit (lg/L) ,0.5 ,0.05 ,0.05
Gas chromatography analysis/electron capture detector analysis
Model HP-5890 gas chromatograph þHP 6890 autosampler
Injection volume 3 lL splitless 3 lL split
Column WCOT 25 m 30.32 mm CP Sil 8 (film thickness, 0.4 lm)
Mobile phase Helium 15 psi
Injection temperature 2408C 1508C
Temperature program Initial temperature, 808C
Initial time, 1 min
Final temperature, 2808C
Final time, 4 min
Initial temperature, 1208C
Initial time, 1 min
Final temperature, 2808C
Final time, 4 min
Detection temperature 3258C
Retention time (min) 10.3 5.8
Detection limit (lg/L) ,0.005 ,0.1
Impact in Ditch Mesocosm of Simulated Spray Drift—Integr Environ Assess Manag 2, 2006 109
squares regression of the natural logarithm of the concen-
tration versus time. They were calculated for mean concen-
trations at the 0.2%, 1%, and 5% treatment levels and were
based on measurements above detection limit.
Water quality endpoints
Temperature, pH, and dissolved oxygen (DO) were
measured 3 times a week between 9:00 AM and 11:00 AM.
The DO was measured with a Wissenschaftlich-Technische
¨tte (WTW) Oxi 320 Set oxygen meter at 10 and 40
cm below water surface, and pH was measured with a WTW
pH196 pH meter at a water depth of 25 cm. Alkalinity and
conductivity were measured every 4 weeks. Conductivity
was measured by means of a WTW LF 96 conductivity
meter at a depth of 25 cm. Alkalinity was analyzed in 100
mL samples, taken from a depth of 25 cm, by titration with
0.02 M hydrochloride down to pH 4.2.
Depth-integrated water samples filtered through a 40-lm
mesh net (100 mL) were stored in labeled polyethylene vials
in a deep freezer at a temperature below 208C for nutrient
analysis. After termination of the experiment, a Skalar 5100
Autoanalyser (Breda, The Netherlands) was used to colori-
metrically analyze for nitrate/nitrite, ammonium, orthophos-
phate, and chloride.
Macroinvertebrates were sampled by means of artificial
substrates and litter bags. The use of artificial substrates is
described in detail by Brock et al. (1982). The litter bag
technique is described in the section ‘‘decomposition’’ . Four
weeks before the first sampling, 4 pebble baskets and 4 litter
bags were placed in the ditches in order to allow
colonization by microfauna and macrofauna. Pebble baskets
and litter bags were assigned to 4 specific ditch segments,
evenly distributed over each ditch. The pebble baskets were
sampled every 4 weeks and, in total, 7 times during the
experiment. Macroinvertebrates that were present on the
substrates were identified and counted alive and then
released again into the corresponding ditch. The species
composition of macroinvertebrates was determined to the
lowest practical taxonomic level. For each ditch the
abundance of macroinvertebrates on pebble baskets and in
litter bags were summed before analysis of the data.
Phytoplankton and zooplankton
Depth-integrated plankton samples were taken every 3
weeks. Phytoplankton and zooplankton were collected in 1
sample. Plankton samples were taken with Perspex tubes
(0.4 m long, 0.8 L in volume) from 3 sampling locations
assigned to specific ditch segments (;5 samples taken at
random per sampling location). Samples were mixed to
create 1 sample (;13 L volume) per experimental ditch.
Five liters of each sample were filtered through a 55-lm
mesh net to concentrate zooplankton and were preserved
with formalin (4%). Eight liters were filtered through a 40-
lm mesh net to concentrate phytoplankton and were
preserved with formalin (4%). Phytoplankton samples were
split up on a weight basis. One set of samples was used for
identification. Of the remaining unprocessed water sample, 1
L was collected for chlorophyll-aand nutrient analysis. For
the chlorophyll-adeterminations, 1 L of the water sample
was concentrated over a Schleicher and Schuell glassfiber
filter (GF52; diameter, 4.7 cm; mesh size, 1.2 lm), by use of
a vacuum pump. The filter was stored in a labelled Petri dish,
wrapped in aluminium foil, at a temperature below 208C
for a maximum period of 3 months. Extraction of
chlorophyll-awas performed by use of the method described
by Moed and Hallegraeff (1978). Chlorophyll-acontent was
analyzed by spectrophotometric measurement (dichromatic
following protocol NEN 6520).
Subsamples of approximately 2 to 3 mL were taken from
1 preserved sample. Subsamples were recalculated to the
overall sample and the numbers in 1 L by weight. Cladocera
were identified to species level. The remaining zooplankton
taxa (e.g., Copepoda, Ostracoda, and Rotifera) were identi-
fied to the lowest practical level. For a few samples, it was
checked whether the counting of 1 subsample was sufficient
by taking an extra subsample.
Phytoplankton species composition was studied by count-
ing the number of cells of a known volume. Taxa and
number of cells were based on 40 counting fields of an object
glass under a microscope (magnification, 3400). In the case
of colony-forming and filamentous algae, the number of
colonies/filaments was counted. Identification took place to
the lowest practical taxonomic level.
Seasonal development of macrophyte species composition
was investigated by monitoring macrophyte cover, abun-
dance, and structure every 6 to 7 weeks (6 times during the
experiment). Cover/abundance (percentage) of each macro-
phyte species was estimated per 5-m ditch length by direct
measurement. To describe the macrophyte structure, the
mean, minimum, and maximum height of the vegetation was
measured per 5-m ditch length by direct measurement.
Macrophyte bioassays were performed to determine the
effects of the pesticide application regime on growth of
submerged water plants. The macrophyte we used in the
bioassays was Myriophyllum spicatum. Plant shoots (length,
10 cm) were derived from plants growing in the ditches. Six
to 8 shoots (total weight of 5–7 g) were planted in a
flowerpot. The flowerpots were filled with ditch sediment.
In each ditch, 6 flowerpots were placed on the sediment
surface at each of 2 locations in specific ditch segments.
After an incubation period of 6 weeks, above- and below-
sediment plant material was harvested, rinsed to remove
sediment particles, dried in aluminium foil (105 8C, 24 h),
and weighed. The material from 2 locations in each ditch
was pooled. Sampling of macrophyte biomass was repeated 3
times at intervals of 6 weeks.
Leaf litter decomposition
Decomposition of coarse particulate organic material was
studied by means of the litter bag technique described by
Brock et al. (1982). The organic matter used in these litter
bags were leaves of Populus 3canadensis (2 g dry weight
dried at 60 8C). These leaves had been soaked 3 times for 2 d
to remove soluble humic compounds. After soaking, they
were dried for 72 h at 60 8C and stored until use. We studied
decomposition of these leaves as influenced by shredding
macroinvertebrates. To achieve this, 2 holes of 0.5 mm were
placed in the stainless steel wire (mesh size, 0.7 30.7 mm)
around the glass Petri dishes (diameter, 11.6 cm) in order to
allow macroinvertebrates to enter. Two litter bags were
110 Integr Environ Assess Manag 2, 2006—GHP Arts et al.
introduced at the sediment surface of a ditch at each of 2
locations. After an incubation period of 2 weeks, each litter
bag was emptied in a white tray to separate coarse
particulate organic material and macroinvertebrates. Macro-
invertebrates on the outer surface of the litter bags (snails,
leeches) were removed and not included in the sample.
Macroinvertebrates were identified alive, counted, and
released again into the ditch. Remaining organic plant
material was dried in aluminium foil at a temperature of
105 8C. After 24 h, dry weight was determined. The
decomposition over a 2-week period was expressed as
percentage of remaining organic material. New litter bags
were incubated 2 weeks before the next sampling date.
Decomposition measurements were repeated 7 times.
To assist with the interpretation of the observed effects,
we calculated Toxic Units (TU) and applied the concept of
concentration addition (Deneer 2000). The TU were based
on acute toxicity data of the most sensitive standard test
species and measured exposure concentrations in the water
compartment of the ditches. The toxicity for sensitive
invertebrates (zooplankton and macroinvertebrates) was
scaled to the toxicity of the 5 pesticides for Daphnia magna:
is the actual concentration of the compound i,
is the geometric mean 48-h EC50 of compound ifor
D. magna (Table 4), and yis the resulting fraction of TU. The
toxicity for phytoplankton was scaled to the toxicity of the 5
pesticides for algae by using Equation 1. EC50
geometric mean EC50 for Selenastrum capricornutum.
The toxicity of the pesticide package for macrophytes was
calculated on the basis of the toxicity of the herbicides
prosulfocarb, metribuzin, and chlorothalonil to Lemna minor
or L. gibba. No data were available for either of the other
compounds. Previous studies with pesticides in experimental
ecosystems have demonstrated that effects on primary
producers are likely to occur at TU
.0.1, and effects
on invertebrates are likely to occur at TU
(Brock, Lahr, et al. 2000; Brock, Van Wijngaarden, et al.
In the data analysis, we used a combination of multivariate
(principal response curves [PRC]) and univariate (Williams
test) techniques. The PRC technique serves to identify the
treatment levels that affect the community and to indicate
the taxa for which the responses affiliate most to the
treatment regime. The Williams test is used to analyze the
dynamics at the taxon level and helps to further identify the
number of taxa affected. A graphical interpretation of the
univariate analysis is often needed to decide whether the
statistically significant deviations can be considered consis-
tent effects and whether these deviations make any sense in
relation to variations in abundance. At the taxon level, no
Table 4. Geometric means of toxicity data of species representative of primary producers and invertebrates for the
pesticides used in the experiment
Pesticide Taxon EC50 (lg/L) Nr Exposure time
Prosulfocarb Selenastrum capricornutum 86.7 1 96 h
Lemna gibba 690 1 14 d
Daphnia magna 1531 2 48 h
S. capricornutum 39.7 72–96 h
L. minor 36.5 96 h
D. magna 14,983 48 h
S. capricornutum .1000 96 h
Lemna sp. — —
D. magna 0.35 48 h
Oncorhynchus mykiss 0.32 96 h
Chlorothalonil S. capricornutum 190.4 3 72–120 h
L. gibba 510 1 14 d
D. magna 105.2 7 48 h
Fluazinam S. capricornutum 160 1 96 h
L. gibba ——
D. magna 201.9 4 48 h
Geometric mean effect concentrations (EC50s) (lg/L) are based on toxicity data for standard laboratory test species (Nr ¼number of
values) commonly used in the 1st-tier risk assessment procedurefor the administration of pesticides. Data were from the ECOTOX database
(www.epa.gov./ecotox/), from the RIVM database (Rijksinstituut voor Volksgezondheid en Milieu, Bilthoven, The Netherlands), and from
data kindly provided by industry (Syngenta, Jealotts Hill, UK).
Data originate from Brock et al. (2004).
Data originate from Van Wijngaarden et al. (2004).
Impact in Ditch Mesocosm of Simulated Spray Drift—Integr Environ Assess Manag 2, 2006 111
observed effect concentrations (NOEC) resulted from the
Williams test (ANOVA, p0.05) (Williams 1972). Analyses
were made with the Community Analysis computer program
(Hommen et al. 1994). Data of macrophyte bioassays and
leaf litter composition were also analysed with the Williams
test (ANOVA, p0.05) (Williams 1972).
Before univariate and multivariate analysis of abundance
values, the macroinvertebrate data were ln(2xþ1) trans-
formed, the zooplankton data were ln(10xþ1) transformed,
and the phytoplankton data were ln(0.2xþ1) transformed,
where xequals the abundance value. This transformation was
performed to down-weight high abundance values and
approximate a normal distribution of the data (for rationale,
see Van den Brink et al. 2000).
The effects of the treatment on the community level of
macroinvertebrates and zooplankton were analyzed by the
PRC method, which is based on the redundancy analysis
ordination technique, the constrained form of principal
component analysis (Van den Brink and Ter Braak 1998,
1999). The PRC method is a multivariate technique specially
designed for the analysis of data from microcosm and
mesocosm experiments. PRC produces a diagram that shows
the sampling weeks on the x-axis and the 1st principal
component of the treatment effects on the y-axis. This yields
a diagram that shows the deviations in time of the treatments
compared with the control. The species weights (b
on the right side of the diagram, can be interpreted as a
correlation of each species with the response given in the
diagram. Thus, the species that has the highest weight is
indicated to have decreased most at the highest treatment
level. The species with the lowest negative weight has
increased most at the highest treatment level. A complete
description and discussion of the PRC method is given by Van
den Brink and Ter Braak (1998, 1999). PRC analysis was
performed by means of Canoco for Windows software
package, version 4 (Ter Braak and Smilauer 1998).
In the Canoco computer program, redundancy analysis is
accompanied by Monte Carlo permutation tests to assess the
statistical significance of the effects of the explanatory
variables on species composition of the samples (Van den
Brink et al. 1996). The significance of the PRC diagram in
terms of displayed treatment variance was tested by Monte
Carlo permutation of entire time series in the redundancy
analysis from which PRC is obtained, by using an F-type test
statistic based on the eigenvalue of the component (Van den
Brink and Ter Braak 1999).
Although the 1st principal component extracts the
maximum amount of information from the multivariate
treatment effects, it does not necessarily describe the effects
of each treatment on all taxa in sufficient detail. Further
components can be extracted from the residual variation. The
significance of the 2nd and higher PRC diagrams was tested
by means of Monte Carlo permutation tests. It showed that
the 2nd and higher components were not significant and were
therefore not considered. The results of the PRC analysis can
be evaluated in terms of the fractions of the variance
explained by the factors time and treatment. Which fraction
of the variance is explained by treatment, is shown in the PRC
Monte Carlo permutation tests were also performed for
each sampling date, by use of the ln-transformed treatment
levels as the explanatory variable (for rationale, see Van den
Brink et al. 1996). This procedure served to test the
significance of the treatment for each sampling date. In
addition to the overall significance of the treatment regime,
we tested which treatments differed significantly from the
controls, so as to infer the NOEC at the community level
). The NOEC
done by applying the Williams test to the sample scores of
the 1st principal component of the PCA of each sampling
date in turn (for the rationale of this, see van den Brink et al.
Overall, initial treatment concentrations (based on meas-
urements in the spray solution and the amounts of solution
sprayed) in the ditch mesocosms were close to the intended
peak concentrations (Table 1). The only exception was that
the 2nd application of lambda-cyhalothrin was approximately
50% lower than intended, which could be ascribed to an
incorrect 50% lower concentration in the application solution.
Average treatment concentrations in the treatment regimes as
a function of time are shown in Figure 2.
Prosulfocarb disappeared from the water layer most slowly.
Mean calculated dissipation DT50 value from the water layer
was about 7 d at the 5% treatment level (Table 5). Disspation
DT50 values for metribuzin and fluazinam in the water
compartment were between 1 and 2 d (Table 5). For lambda-
cyhalothrin, a DT50 value in the water layer of about 1 d was
calculated. With water DT50 values of 0.3 to 0.5 d,
dissipation of chlorothalonil was the most rapid (Table 5).
Peak exposure concentrations reached 0.88 TU for algae at
the highest treatment level (5%) (Figure 3). The largest
component of the toxicity was caused by prosulfocarb in the
1st weeks of the experiment. Based on toxic units, metribuzin
was expected to contribute to the toxicity for algae as well
(;0.1 TU; Figure 3). However, data from mesocosm studies
indicate that effects of metribuzin at the ecosystem level are
unlikely to occur at the applied concentrations. The contri-
bution of chlorothalonil to the toxicity to algae was low (short
exposures to 0.12–0.13 TU), whereas the contribution of
fluazinam was negligible. At the 1% treatment level, only
prosulfocarb reached concentrations at which toxicity for
algae could be expected (Figure 3).
For invertebrates, peak exposure concentrations reached
0.4 TU at the highest treatment level (5%) (Figure 3). For
these groups, lambda-cyhalothrin and chlorothalonil were
expected to contribute most to the observed toxicity.
Contributions of the other compounds at the 5% treatment
level and of lambda-cyhalothrin and chlorothalonil at the 1%
treatment level lie in the range 0.01 to 0.1 TU. In this range,
low levels of effects cannot be ruled out. At the 0.2% level,
TU values were less than 0.014 for invertebrates and less than
0.06 for algae, so no effects were anticipated.
Water quality endpoints
Nutrient concentrations in the water layer were low (Table
6). Nutrients and chloride were not different between the
Table 7 presents the results of NOEC calculations for pH
and DO. There was a significant decrease in pH of
approximately 0.5 to 1 pH unit from day 1 to day 16 at the
5% treatment level, and from day 1 to day 9 at the 1%
treatment level. In addition, there was a significant decrease
112 Integr Environ Assess Manag 2, 2006—GHP Arts et al.
of oxygen of approximately 0.75 mg/L on day 3 during this
period at the 5% treatment level. The responses observed can
most probably be attributed to prosulfocarb. A significant
decrease in pH also occurred from day 49 to day 53 at the 5%
treatment level, overlapping slightly with the application
period of chlorothalonil. In the last week of the application
period (days 112 through 123), DO in the water column
significantly increased at the 5% treatment level (Table 7).
Increased oxygen levels coincided with an increase in
turbidity of the water at the 1% and 5% treatment levels
(data not shown), and an increase in chlorophyll-a (NOEC
week 14 ¼0.2%), indicating an increase in algal densities.
Over the experimental period, a total of 81 different
macroinvertebrate taxa were identified as belonging to 8
Figure 2. Trends in concentrations of pesticides applied in the potato spray program. Concentrations have been measured in the water compartment of the
ditch mesocosms in the 0.2%,1%,and5%treatment regimes. Note that the different pesticides were applied on different dates (see Table 1).
Impact in Ditch Mesocosm of Simulated Spray Drift—Integr Environ Assess Manag 2, 2006 113
classes, with almost half of the taxa (34) being Insecta. The
mesocosms were characterized by Crustacea (dominant
species: G. pulex, Proasellus meridianus/coxalis), Insecta
(Ephemeroptera; dominant species: Cloeon dipterum, Caenis
horaria; Odonata; Diptera: Chaoborus sp., Tanypodinae,
Ceratopogonidae), Hirudinea (dominant species: Erpobdella
octoculata), Gastropoda (dominant species: Armiger cristata)
and Turbellaria (dominant species: Polycelis nigra/tenuis).
Of the total variance, 24% could be attributed to the
treatment regime by the PRC analysis (Figure 4). The PRC
diagram displayed a significant amount of the treatment
variance (p¼0.003). Clear treatment effects were observed
at the 5% and 1% treatment levels compared with the controls
from week 7 onward (Figure 4 and Table 8). At the 5% level,
the PRC diagram shows long-lasting effects on the macro-
invertebrate community without full recovery (Figure 4).
These effects first became evident after the 1st application of
lambda-cyhalothrin. Effects at the community level also
persisted during the application periods of chlorothalonil
and fluazinam, presumably because there was no recovery
from the lambda-cyhalothrin effects. At the 1% level, effects
at the community level were 1st observed in week 7 and
seemed to be associated with the application of lambda-
cyhalothrin in week 5. A similar effect was not observed after
the 2nd application of lambda-cyhalothrin, possibly because
dosages were lower than the intended peak concentration and
just below or above threshold levels (Figure 2 and Table 11).
Statistical calculations resulted in an isolated community
at the 0.2% treatment level and an
of 1% (Table 8).
Taxa with high positive weights (i.e., those taxa that
decreased most strongly in association with changes in
community structure) in the PRC analysis (Figure 4) were
G. pulex (Figure 5A), Chaoborus sp. (Figure 5B), C. horaria
(Figure 5C), and C. dipterum (Figure 5D). The populations of
these taxa declined significantly at the highest treatment
Figure 3. Toxicity of the pesticide package for invertebrates and algae at 1%and 5%drift emission of label-recommended rates of pesticides expressed as Toxic
Units (TU) and the contribution of each pesticide to the toxicity (0.2%not shown). See text for rationale for these values. The lines of 0.01 and 0.1 TU are
Table 5. Calculated dissipation DT50 (dissipation time, time required to dissipate 50%of the mass originally present) values
(d) for the compounds in the pesticide package in the water column of the ditches
Prosulfocarb Metribuzin Chlorothalonil Lambda-cyhalothrin Fluazinam
Med Low High Med Low High Med Low High Med Low High Med Low High
0.2 6.1 5.6 6.7 1.3 1.1 1.6 0.3 0.1 1.0 NC NC NC 2.0 1.1 10.0
1 6.0 5.4 6.7 1.7 1.3 2.6 0.3 0.1 2.7 1.2 0.5 2.7 1.7 1.2 2.9
5 7.0 6.6 7.5 1.7 1.4 2.2 0.5 0.3 1.2 0.9 0.7 1.2 1.4 1.3 1.6
The table presents median values over time for the 3 different treatments, as well as 95%lowest and highest values. NC ¼could not be
calculated because of very fast dissipation of the compound; med ¼median; low ¼lowest value; high ¼highest value.
114 Integr Environ Assess Manag 2, 2006—GHP Arts et al.
level. Recovery was observed for the affected insect taxa but
was not observed for G. pulex. Taxa with high negative
weights shown in the PRC diagram (Figure 4) comprise Dero
sp. (Figure 5F), Stylaria lacustris (Figure 5I), and Orthocla-
diinae/Chironomini (Figure 5H). The population abundance
of these taxa all increased, at least temporarily.
Statistical analysis of treatment-related responses for
individual macroinvertebrate populations resulted in NOECs
for 8 taxa (Table 9). Adverse effects of treatments were
apparent from week 7 onward. After week 15, recovery of
most insect populations has occurred. Consistent responses,
defined as statistically significant deviations pointing in the
same direction on at least 2 consecutive sampling dates, were
observed for all the presented taxa in Table 9. Longer-lasting
reductions were observed at the 5% treatment levels and
occurred with G. pul e x (Figure 5A). Recovery of the
populations of this species was hindered by isolation of the
ditches. At the 1% treatment level, long-lasting reductions
were observed within Chaoborus sp. (Figure 5B) and C.
horaria (Figure 5C). The lowest consistent NOECs were
calculated for Chaoborus sp. at the 0.2% treatment level. For
C. horaria and Orthocladiinae/Chironomini, an isolated
NOEC was calculated of 0.2%. NOECs for other species
and sampling dates were at the 1% treatment level.
Phytoplankton and zooplankton
Over the experimental period, a total of 59 different
zooplankton taxa were identified. The majority of the taxa
(40) belonged to the Rotifera, followed by Cladocera (13),
Copepoda (3), Ostracoda (1), and Diptera (3). Copepoda and
Ostracoda were not identified to species level. Rotifera were
the most abundant.
Of the total variance, 18% could be attributed to the
treatment regime by the PRC analysis (Figure 6). This
percentage is in line with results from other experimental
ponds (see Sanderson 2002). The PRC diagram displayed a
significant amount of the treatment variance using a threshold
value of 0.10 for p(p¼0.068). It shows a clear trend,
particularly for the 5% treatment level (Figure 6). Monte
Carlo permutation tests showed significant reductions at this
level from week 8 through week 14 after application (Table
8). The 1st effects on zooplankton followed the application of
the insecticide lambda-cyhalothrin. The period in which
effects on zooplankton were apparent also overlapped with
the application period of the fungicide chlorothalonil.
Recovery had already started within the application period
of the fungicide fluazinam. Statistical calculations resulted in a
for the zooplankton at the 1% treatment level
A number of Rotifera, including Polyarthra remata and
Hexarthra sp., showed high negative weights in the PRC (Fig
6; see species weights b
), indicating a treatment-related
decline. Cladocera had positive weights in the PRC, indicating
treatment-related increases in abundance. Treatment-related
responses for individual taxa were statistically analyzed and
resulted in a range of NOECs (Williams test, p0.05) for 3
zooplankton taxa (Table 9). For Hexarthra sp. and Daphnia
group galeata, a NOEC of 0.2% was calculated on an isolated
sampling date as well as a NOEC of 1% on several other
sampling dates. For P. remata a NOEC of 1% was calculated.
The dynamics of zooplankton populations showing treat-
ment-related responses after application of the pesticide
package are presented in Figure 7. Significant long-lasting
reductions in numbers occurred in 2 populations of Rotifera
species. In the population of Hexarthra sp., reductions were
observed at the 1% and 5% levels (Figure 7A). In the
population of P. remata, reductions were observed at the 5%
level (Figure 7B). Within the group of Calanoida, reductions
were observed at the 5% level (Figure 7D). Short-term
increases were observed in the population of Daphnia group
galeata at the 1% and 5% levels (Figure 7C).Both positive and
negative effects were apparent from week 8 onward.
A total of 92 phytoplankton taxa were identified. Dominant
groups were green algae, mainly Chlorococcales, Desmidia-
ceae, and Diatoms. The PRC diagram for the phytoplankton
did not display a significant amount of the treatment variance
and, thus, is not shown. Statistical calculations resulted in a
for the phytoplankton at the 1% treatment
level (Table 8), which was calculated for 1 sampling date
(week 5). The data also suggest a trend of an effect in week 2,
although not statistically significant (p¼0.08; Table 8). Of the
phytoplankton taxa, only Flagellatae showed a consistent
response (increase in numbers) at the 5% treatment level in
weeks 2 and 5 (Table 9). For Flagellatae, a NOEC of 1% was
calculated. Figure 8 presents the dynamics of the population
Chlorophyll-aconcentrations of phytoplankton did not
show a consistent treatment-related response. Only 1 isolated
NOEC could be calculated for week 14 (0.2%).
The aquatic vegetation in the ditches was dominated by
Chara globularis ssp. virgata. Other abundant macrophytes
were Elodea nuttallii and Sagittaria sagittifolia. Submerged and
floating algal beds were locally abundant.
Of the percentage cover/abundance data of macrophyte
taxa, only floating filamentous algae showed a consistent
response (Table 9) after the application period of the pesticide
package. Floating filamentous algae developed in controls and
0.2% treatments and resulted in significantly higher percent-
age cover compared with that of (Table 10) the 1% and 5%
treatments. Dynamics of submerged filamentous algae
showed a consistent treatment-related response as well (Table
10). There were no significant treatment-related responses in
Table 6. Concentrations of nutrients and chloride (mg/L) in surface water of ditch mesocosms
Sum nitrate/nitrite Ammonium Orthophosphate Chloride
Controls 0.079 60.089 0.032 60.012 0.006 60.002 6.656 60.740
0.2%0.081 60.074 0.036 60.011 0.006 60.002 7.026 60.890
1%0.094 60.114 0.029 60.010 0.007 60.003 7.124 60.819
5%0.082 60.104 0.029 60.014 0.006 60.002 6.977 60.737
Impact in Ditch Mesocosm of Simulated Spray Drift—Integr Environ Assess Manag 2, 2006 115
4 in situ bioassays performed with M. spicatum (NOECs .
5%). Total macrophyte biomass in bioassays varied from 0.68
60.19 to 0.84 60.25 g in week 5 and from 1.42 60.35 to
1.98 60.02 g in week 22.
Leaf litter decomposition
Significant treatment-related effects on leaf litter decom-
position were not observed in the 7 decomposition experi-
ments we performed (NOECs .5%). Only at week 8 was a
Tab l e 7 . Results of no observed effect concentration (NOEC) calculations of the
physicochemical parameters (drift percentage; Williams test, p,0.05)
–4 — NM
88 — —
1 0.2%—91— —
2 0.2%—92— —
3 0.2%—94— —
4— 1%98 — —
9 0.2%—99— —
14 1%1%101 — —
15 0.2%— 105 — —
16 1%1%106 — —
21 — — 108 — —
23 — — 112 — 1%
29 — — 113 — 1%
36 — — 115 — —
42 — — 119 — 1%
43 — — 120 — —
49 1%— 123 — 1%
50 1%— 126 — —
53 1%— 127 — —
56 — — 130 — —
57 — — 133 — —
60 — — 134 — —
63 — — 137 — —
64 — — 140 — —
67 — — 141 — —
70 — 1%144 — —
71 — — 147 — —
74 — — 148 — —
77 — — 151 — —
78 — — 154 — —
81 — — 161 — —
84 — — 169 — —
85 — — 176 — —
Only endpoints showing a consistent response (NOECs calculated for 2 consecutive sampling dates)
are displayed. Em dashes indicate NOEC 5%.
NM ¼not measured.
116 Integr Environ Assess Manag 2, 2006—GHP Arts et al.
single NOEC of 1% calculated. The decay rate varied from
19.29 62.24% in 5% treatments to 21.78 61.75% in 0.2%
treatments over a 2-week period. In the control ditches, the
decay rate was 20.63 62.46% over a 2-week period.
The treatment regime of the ditch mesocosms was
considered a worst case, despite the absence of herbicide
application at the end of the season. Instead of this,
mechanical removal was simulated because of limitations set
by the ditch mesocosms (no persistent compounds could be
applied) and limitations set by the compounds (compounds
should be permitted for use in The Netherlands at that
moment and in near future). Because ditch mesocosms were
not connected to each other during the experiment, the
absence of recovery via water contributed to the worst-case
situation. Recovery from neighboring ditches via air was, of
course, possible and reflected more or less the field situation.
Except for prosulfocarb, the dissipation of other pesticides
from the water of the ditch mesocosms was very rapid, and
concentrations generally reached detection limits before the
next application was made (Figure 2 and Table 5). The field
dissipation measured covers the combined aqueous losses
owing to photolysis, hydrolysis, sorption, volatilization, and
biodegradation. Dissipation DT50 values from the literature
(where available, field values) were compared to those
measured in this experiment. For prosulfocarb, no other field
studies were available and in the standard laboratory tests
DT50 is 1 d in water and 150 to 380 d in sediment in a
sediment–water study (data from Syngenta). For metribuzin,
a dissipation half-life of 5 d was found in experimental pond
mesocosms by Fairchild and Sappington (2002), whereas a
range of 6 to 9.4 d (mean, 7.1 d) was reported by Brock et al.
(2004) in a ditch enclosure study. In the present study, the
dissipation of metribuzin from the water layer was faster than
reported elsewhere. One possible explanation for this could
be differences in photolysis, because this process is reported
to be important for the dissipation of metribuzin (Muszkat et
al. 1998, 2002).
For lambda-cyhalothrin, DT50 values ranged from 0.7 to
1.1 d in laboratory microcosms (Van Wijngaarden et al. 2004)
and from 0.24 to 0.27 d in ditch enclosures (Leistra et al.
2003; Roessink et al. 2005). Wendt-Rasch et al. (2004)
calculated a mean DT50 in water of 0.86 d in outdoor Elodea-
dominated microcosms and a mean water DT50 of 2.4 d in
outdoor Lemna-dominated microcosms. Hand et al. (2001)
found a DT50 of less than 0.125 d for dissipation from the
water column in an indoor macrophyte-dominated micro-
cosm study and a DT50 of less than 0.125 d for the whole
system. The dissipation of lambda-cyhalothrin in the present
study was in the same range as reported by Van Wijngaarden
et al. (2004) and in the Elodea-dominated microcosms
reported by Wendt-Rasch et al. (2004) but was slower than
that reported by Leistra et al. (2003), Roessink et al. (2005),
and Hand et al. (2001). Alkaline hydrolysis in the water near
the surface of macrophytes is considered to be the main
dissipation process for lambda-cyhalothrin (Leistra et al.
Available water DT50 values for fluazinam range from 1.5
to 3.3 d in indoor microcosms (Van Wijngaarden et al. 2004)
and from 0.9 to 1.1 d in outdoor microcosms (Wendt-Rasch
et al. 2004). The dissipation of fluazinam in our study fell
within these ranges. For chlorothalonil, field DT50 values in
water range from 0.17 (Davies 1988) to 1.25 d (Ernst et al.
Figure 4. Principal response curves with species weights (b
) for the macroinvertebrate data set, indicating the effects of applications of compounds of the
pesticide package. Of the variance, 38%could be attributed to sampling date and is displayed on the horizontal axis. Differences between replicates accounted
for 38%of the variance. Twenty-four percent of the variance could be attributed to the treatment scenario. Of this percentage, 34%is displayed on the vertical
axis. The species weight can be interpreted as the affinity of the taxon to the principal response curve. Taxa that have a species weight between 0.5 and 0.5
have a low correlation with the response curve and are therefore not displayed. The PRC diagram does display a significant amount of the treatment variance
(p¼0.003). Application of herbicides, medium shading; application of insecticide, light shading; application of fungicides, solid.
Impact in Ditch Mesocosm of Simulated Spray Drift—Integr Environ Assess Manag 2, 2006 117
1991). In the present study, the dissipation of chlorothalonil
was somewhat slower than that reported by Davies (1988).
In conclusion, when compared with the field DT50 values
reported in other studies, the fate of the compounds was
overall very comparable. The relatively fast dissipation of
metribuzin, lambda-cyhalothrin, chlorothalonil, and fluazi-
nam from the water of the ditch mesocosms suggests that in
our test systems these pesticides cause short-term stress only,
despite the repeated application of some of these compounds.
However, the longer exposures to the herbicide prosulfocarb
may have also led to chronic stress on the phytoplankton
The 2nd application of lambda-cyhalothrin, which was
approximately 50% lower than intended, had no consequen-
ces for the 5% treatment. Concentrations still were above
threshold levels (Tables 1 and 11). It might have had
consequences for the 1% treatment, because realized nominal
concentrations were around threshold concentrations. If the
target concentration (just above threshold concentration)
would have been realized instead, the effect at the 1%
treatment level might have been consistent between weeks 5
and 9 (community-level NOEC
at the 0.2%
treatment level [Table 8] on both dates and the effects would
have been short lived). However, this would not have changed
the results and conclusions of this experiment: a NOEC for
the community of 0.2% and clear but transient effects at the
1% treatment level. Effects of lambda-cyhalothrin are in line
with effects found in other mesocosm experiments (Van
Wijngaarden et al. 2004; Roessink et al. 2005).
Secondary effects and interaction
In addition to direct effects on sensitive species, pollutants
may exert effects on tolerant species by a number of
ecological mechanisms. Such effects are called indirect (or
secondary) contaminant effects (Fleeger et al. 2003). Com-
petitive release and trophic cascades are indirect effects that
are very common (Fleeger et al. 2003). Only in studies at the
population, community, or ecosystem level can indirect
contaminant effects be detected. Because indirect effects
may be an important part of contaminant effects at these
levels, we will consider some important indirect effects that
were detected in the experiment more profoundly.
The trend of an effect on phytoplankton after the
application of prosulfocarb in week 2 (Table 8), the observed
effect on phytoplankton in week 5 (Table 8), and effects on
Flagellatae (increase) in weeks 2 and 5 suggest that the effects
observed can be attributed to herbicide application and not to
the application of lambda-cyhalothrin. It seems most likely
that these effects were a result of exposure to prosulfocarb
(Table 12). The ecological reason for the increase in
Flagellatae is difficult to explain from the data derived in
the experiment. An increase in numbers usually indicates the
occurrence of an indirect effect, such as a result of a decrease
in other taxa. This primary effect, however, could not be
demonstrated by the data. Chlorophyll-aconcentrations of
phytoplankton did not show a consistent treatment-related
response, despite the temporal increase in Flagellatae. The
decreases in pH and dissolved oxygen described above suggest
that prosulfocarb decreases the net photosynthesis of the
community of primary producers at the 5% treatment level in
particular. However, consistent declines in populations of
primary producers could not be observed during the first 5
weeks after prosulfocarb application. In addition, significant
declines in zooplankton and macroinvertebrate taxa were also
not observed in this period. A decrease in pH was also
observed in the 8th week of the experiment (Table 7). A
Table 8. Results of the Monte Carlo permutation test (pvalue) and no observed effect concentrations (NOECs) on the
zooplankton, phytoplankton, and macroinvertebrate community level (drift percentage; Williams test, p,0.05 ) for the
different treatment levels of a pesticide package containing prosulfocarb, metribuzin, lambda-cyhalothrin, chlorothalonil,
Zooplankton Phytoplankton Macroinvertebrates
pNOEC pNOEC pNOEC
–1 .0.10 .5%.0.10 .5%.0.10 .5%
2.0.10 .5%0.08 .5%——
5 0.080 .5%0.04 1%——
8 0.009 1%.0.10 .5%——
11 0.020 1%.0.10 .5%0.005 1%
14 0.015 1%.0.10 .5%——
15 — — — — 0.017 1%
17 .0.10 .5%.0.10 .5%——
19 — — — — 0.008 1%
20 .0.10 .5%.0.10 .5%——
23 — — — — 0.001 1%
24 .0.10 .5%.0.10 .5%——
118 Integr Environ Assess Manag 2, 2006—GHP Arts et al.
decrease in pH suggests a decrease in the net photosynthesis
of the primary producers. Increases in oxygen levels, turbidity,
and chlorophyll-ain water in weeks 14 to 16 (Table 7) suggest
an increased algal production. However, effects on algae
community structure were not observed in the 3 examples
mentioned above. One possibility might be that effects on
functional endpoints last longer and can be detected, whereas
effects on structural endpoints might be too transient to
detect with the chosen sampling frequency. A 2nd possibility
might be the filtration of the samples through a net with 40-
lm mesh size. As a consequence of this, the smallest species
(among others, nanoplankton) are not considered in this
experiment. Therefore, effects on phytoplankton in this
experiment are limited to the larger species. The decrease in
filamentous algae from week 13 onward and the persistence
of this effect at the 5% treatment level may be another
secondary effect, of which the cause is not clear.
Some Rotifera showed clear direct effects (Figure 6). The
decrease in the population of P. remata may be owing to direct
effects of chlorothalonil, but indirect effects may also play a
role. Because of the observed indirect effect of increasing
daphnid abundance owing to a decrease in its predator
Chaoborus, competitive exclusion may have led to a decline in
The significant reductions in G. pulex at the 5% treatment
level did not influence the extent to which leaf litter was
decomposed in the ditch mesocosms consistently. Only in
week 8 was a single NOEC of 1% calculated. A reduced
decomposition can probably be attributed to the effect of
lambda-cyhalothrin, but the effect was not consistent.
Obviously, other shredders such as Proasellus meridianus/
coxalis, which were unaffected by the pesticide package, were
able to maintain this function.
The adverse effect on Chaoborus obscuripes (Figures 4 and
5B) contributed to several indirect effects. In the 0.2% and
Figure 5. Dynamics of macroinvertebrate populations showing consistent treatment-related responses (Table 9) after applications of compounds of the
pesticide package. Numbers per artificial substrata are geometric mean abundance numbers of Gammarus pulex (A), Chaoborus sp. (B), Caenis horaria (C),
Cloeon dipterum (D), Haliplidae (E), Dero sp. (F), Lymnaea stagnalis (G), Orthocladiinae/Chironomini (H), and Stylaria lacustris (I).
Impact in Ditch Mesocosm of Simulated Spray Drift—Integr Environ Assess Manag 2, 2006 119
1% treatments, C. obscuripes was eliminated. Hence, preda-
tion of C. obscuripes did not reduce the numbers of Calanoida
at these treatment levels (Figure 7D). At the 5% level,
Calanoida show a strong decrease owing to direct effects after
treatment with lambda-cyhalothrin and chlorothalonil. In the
controls, numbers of Calanoida were reduced by predation of
C. obscuripes. It may be owing to such nonlinearity in the
observed responses that apparent effects were absent.
Significant reductions in the abundance of the mayfly C.
dipterum by lambda-cyhalothrin also resulted in several
indirect effects. As a result of competitive release, other
water column inhabitants like S. lacustris,Dero sp., and
Orthocladiinae/Chironomini may have increased.
Other indirect effects observed in the ditch mesocosms
were an increase of the flatworms Dugesia lugubris and
Mesostoma and an increase of the snail Lymnaea stagnalis. The
latter phenomenon has been observed in other mesocosm
experiments (Van Wijngaarden et al. 2004; Roessink et al.
2005) and is probably owing to competitive release and
subsequent higher food availability. Farmer et al. (1995)
found higher algal biomass and higher abundances of copepod
nauplii and Rotifera as indirect effects of lambda-cyhalothrin
at levels above the highest concentration in our study. Effects
were probably owing to effects on Asellidae and Gammaridae.
In the present study, Asellidae did not decrease significantly.
Summary of effects and ecological threshold concentrations
of individual compounds
In the present study, the total pesticide package resulted in
short-term responses of pH, oxygen, and phytoplankton and
in long-term effects on zooplankton and macroinvertebrates
at the 5% treatment level. At the 1% treatment level, only
slight, transient effects were observed on a limited number of
zooplankton and macroinvertebrate species and on pH. At the
0.2% treatment level, no consistent treatment-related effects
The effects at the community level were compared with
known ecological threshold concentrations (NOEC and Class
2 LOEC values) for the individual compounds based on
microcosm and mesocosm studies with these individual
compounds (Table 11 with data from Brock et al. 2004;
Van Wijngaarden et al. 2004; Roessink et al. 2005; data not
Table 9. Results of no observed effect concentration (NOEC) calculations of zooplankton, phytoplankton, macroinvertebrate
taxa, and percentage cover of floating filamentous algae (drift percentage; Williams test, p,0.05 ) for the different
treatment levels of a pesticide package containing prosulfocarb, metribuzin, lambda-cyhalothrin, chlorothalonil,
Zooplankton Phytoplankton Macrophytes
Hexarthra sp. Polyarthra remata Daphnia group galeata
(;5–6 micron, 2 flag.) Filamentous algae
–1 — — — — —
2— — — 1%"—
5— — — 1%"—
11 — 1%#1%"——
— 17 0.2%#———
19 — — — — 0.2%#
20 1%#—— ——
24 — — — — 1%#
pulex Chaoborus sp.
depterum Haliplidae Dero sp.
–1 — — — — — — 0.2%"——
3— — —— ——— — —
11 1%#0.2%#1%#1%#——— 1%"1%"
15 — 1%#0.2%#1%#—1%"1%"—1%"
19 1%#— — — 0.2%#1%"1%"1%"—
Only taxa showing a consistent response (NOECs calculated for 2 consecutive sampling dates) are displayed. Blank cells indicate NOEC
5%.#¼populations were significantly reduced at concentrations above NOEC; "¼populations were significantly increased at
concentrations above NOEC
120 Integr Environ Assess Manag 2, 2006—GHP Arts et al.
shown). This table shows that the effects of metribuzin and
fluazinam at the community level are unlikely at the
concentrations applied in the experiment. The treatment
concentrations in the experiment lie below or in the range of
the threshold concentrations for these compounds (for
metribuzin, see Fairchild and Sappington 2002; Brock et al.
2004). Effects of chlorothalonil might be expected at the 5%
treatment level but are unlikely at the 1% treatment level.
The treatment concentrations of lambda-cyhalothrin lie above
the thresholds, indicating that clear effects would be expected
at both the 1% and 5% treatment levels. For prosulfocarb only
the First Tier Uniform Principles Standard is available (8.6 lg/
L), indicating that probably effects would be expected at both
the 1% and 5% treatment levels.
Table 12 provides an overview of the effects of the
treatments on the various endpoints studied. In this table,
observed effects were summarized into effect classes de-
scribed by Brock, Lahr, et al. (2000) and Brock, Van
Wijngaarden, et al. (2000). At the 5% level, a strong decline
was observed for G. pulex,Chaoborus, and mayfly larvae, most
likely caused by direct toxic effects of lambda-cyhalothrin
(Tables 1 and 11). Also Haliplidae decreased at the end of the
experiment. Increase in numbers was observed in populations
of Dero sp., L. stagnalis, Orthocladiinae/Chironomini, and S.
lacustris. This may be, at least partly, a result of competitive
exclusion due to the strong negative effect on the mayfly C.
dipterum by lambda-cyhalothrin. During the period of
chlorothalonil application, decreases in population densities
of Rotifera and Calanoida were observed. At the 5%
treatment level, the ecological threshold concentrations for
chlorothalonil were exceeded (Tables 1 and 11). The effects
on Rotifera and Calanoida can therefore most probably be
attributed to exposure to chlorothalonil. Although based on
laboratory data, effects of fluazinam on the zooplankton
might have been expected. However, they were not observed
in this experiment. At the community level, effects of
fluazinam are unlikely (Tables 1 and 11). Effects on the
higher levels might be absent because of functional redun-
dancy between species. At the 5% level, a short-term decrease
in pH and DO and an increase of Flagellata were observed in
the 1st weeks of the experiment. Because effects of
metribuzin at the community level are unlikely at the
concentrations applied in the experiment (Tables 1 and 11),
these effects can be attributed to prosulfocarb. Filamentous
algae were lower than were the controls from week 13
onward. In weeks 14 to 15, algal productivity was increased.
So, at the 5% treatment level, pronounced effects were
observed and could be attributed to 3 of the 5 compounds
within the pesticide package. Most of these effects persisted
until the end of the experiment (effect class 4 in Table 12).
Recovery was observed within the group of insects (Figure 5B
to D), which may be attributed to the multivoltine character
of their propagation and their recolonization abilities by air
from neighboring control ditches. Recolonization by other
routes was not possible in this experiment because of the
hydrological isolation of the ditches. Hence, recovery of the
Figure 6. Principal response curve (PRC) with species weights (b
) for the zooplankton data set, indicating the effects of applications of compounds of the
pesticide package. Of the variance, 44%could be attributed to sampling date and is displayed on the horizontal axis. Differences between replicates accounted
for 38%of all variance. Eighteen percent of the variance could be attributed to treatment scenario. Of this percentage, 24%is displayed on the vertical axis. The
species weight can be interpreted as the affinity of the taxon to the PRC. The PRC diagram does display a moderate significant amount of the treatment
variance (p¼0.068). Application of herbicides, medium shading; application of insecticide, light shading; application of fungicides, solid.
Impact in Ditch Mesocosm of Simulated Spray Drift—Integr Environ Assess Manag 2, 2006 121
population of G. pulex could not occur because of a lack of
external supply. Other macroinvertebrates without a terres-
trial stage in their life cycle were also hindered to recover.
Rotifera and Calanoida recovered rapidly as a result of
survival in less-affected microhabitats or production of
resistant eggs. The importance of life cycle characteristics to
the recovery potential of affected habitats and ditches has also
been stressed by other authors (Maund et al. 1997; Sherratt et
At the 1% treatment level, transient decreases were
observed in populations of Chaoborus and C. horaria, most
likely caused by lambda-cyhalothrin (Tables 1 and 11), and on
Hexarthra sp., likely caused by lambda-cyhalothrin and
chlorothalonil (Table 9). A clear negative effect was observed
on Haliplidae at the end of the experiment. Clear increases of
population numbers of L. stagnalis and Orthocladiinae/
Chironomini were observed, although abundance values of
these invertebrates were small (Figure 5G and H). Another
clear indirect effect is the temporary increase of Cladocera,
most likely caused by the direct toxic effects of lambda-
cyhalothrin on the predator Chaoborus. Prosulfocarb most
probably caused a short-term decrease in pH in the 1st weeks
of the experiment. Floating filamentous algae decreased from
week 13 onward but were not significantly different any more
from controls and 1% treatments by the end of the experi-
ment. In weeks 14 to 15, algal production is increased. We can
conclude that at the 1% level only slight, short-term effects
For macroinvertebrates a NOEC was calculated at the
treatment level of 0.2%. The NOECs for other invertebrates
(zooplankton) and phytoplankton were equal to a treatment
level of 1%. For the pesticide package, the most sensitive
organisms within the community were the macroinverte-
brates, with Chaoborus and C. horaria as the most sensitive
species and lambda-cyhalothrin as the most toxic compound.
Figure 7. Dynamics of zooplankton populations showing consistent treatment-related responses (except for Calanoida, Table 9) after applications of a
pesticide package containing prosulfocarb, metribuzin, lambda-cyhalothrin, chlorothalonil, and fluazinam. Numbers are geometric mean abundance numbers
of Hexarthra sp. (A), Polyarthra remata (B), Daphnia group galeata (C), and Calanoida (D).
Table 10. Dynamics of submerged and floating filamentous algae expressed as mean cover percentage per treatment level
per sampling date plus associated standard deviation
Submerged filamentous algae Control 0.2%1%5%
–1 29.47 625.54 6.94 67.51 20.83 636.08 6.67 63.57
7 21.91 625.42 4.19 65.92 1.06 61.06 2.31 62.31
13 21.78 619.50 21.38 619.27 9.58 616.06 31.00 624.00
19 14.50 67.12 19.81 64.68 4.38 65.56 8.25 67.72
23 26.75 613.74 20.31 61.33 4.88 64.28* 3.13 61.44*
Floating filamentous algae
–1 8.44 67.67 2.38 63.01 10.42 616.87 0.83 61.44
7 5.97 64.37 6.04 64.71 9.56 69.56 2.19 62.19
13 15.06 611.01 14.25 64.95 8.21 614.00 4.94 61.81
19 19.88 66.84 21.06 64.33 6.79 66.37* 5.54 60.71*
23 14.41 63.65 23.06 66.28 8.46 69.90 2.54 61.50*
Asterisks indicate significant differences (Williams test, p,0.05).
122 Integr Environ Assess Manag 2, 2006—GHP Arts et al.
Hence, overall, we consider the 0.2% treatment level as the
NOEC for the community.
Toxicity of the pesticide package
Effects on primary producers are likely at TU
(Brock, Lahr, et al. 2000). At the 5% level, a temporal exposure
peak up to 0.9 TU
was observed (Figure 3), which resulted
in short-term responses of pH, oxygen, and phytoplankton
(Table 7). For algae, effects of treatment predominantly
resulted from prosulfocarb treatments and resulted in short-
term responses of pH, oxygen, and phytoplankton.
After repeated application, effects on invertebrates are likely
to occur at TU
.0.1 (Brock, Van Wijngaarden, et al.
2000). At the 5% level, exposure concentrations amounted to
and resulted in long-term responses (see Figures
4 and 6). For invertebrates, effects were probably mainly
caused by lambda-cyhalothrin and chlorothalonil (Table 11).
Effects on zooplankton were most apparent from week 8
onward, after the application of lambda-cyhalothrin in week 5
Table 11. Threshold levels (no observed effect concentration [NOEC]) and lowest observed effect concentration causing
slight effects (Class 2 LOEC) for several endpoints in microcosm and mesocosm studies with individual compounds
Ecological threshold concentrations (NOEC – Class 2 LOEC)
Microcrustaceans 18–56 #10–25 #*
3–10 #* 2–10 #
Rotifers 18–56 #" 10 #"*10#* 2–10 #
Phytoplankton 5.6–18 #.250 10 #"* 10–50 "
Macrophytes .180 .250 100 .250
Macrocrustaceans — 10–25 #*.300 10–50 #
Insects — 10 #* 100–300 —
Other macroinvertebrates — 100–250 #" 100 50–250
pH, DO 5.6–18 #.250 100 .50
NOEC ¼no observed effect concentration; Class 2 effects ¼slight effects; LOEC ¼lowest observed effects concentrations causing Class 2
effects; #¼populations were significantly reduced at concentrations above the thresholds; "¼populations were significantly increased at
concentrations above the thresholds; #" ¼ populations were both significantly reduced and significantly increased at concentrations above
Metribuzin: Brock et al. (2004); lambda-cyhalothrin: Van Wijngaarden et al. (2004); Roessink et al. (2005) (macrophyte-dominated
mesocosms); chlorothalonil: Syngenta, United Kingdom, unpublished data; fluazinam: Alterra, Wageningen, The Netherlands,
Asterisks indicate values below nominal concentrations used in this study.
Table 12. Summary of effects observed in ditch mesocosms treated with pesticides used in potato cultivation
Microcrustaceans 1 2"2"
PRC Zooplankton 1 2#3#
Phytoplankton (Flagellata) 1 2"3"
Macrophytes (Filamentous algae) 1 3#4#
Macrocrustaceans 1 1 4#
Other macroinvertebrates 1 2#" 4#
PRC macroinvertebrates 1 2#4#
pH/dissolved oxygen 1 2#3#
Treatment levels are equal to 0.2%,1%, and 5%spray drift emission of label-recommended rates. The numbers in the table refer to effect
classes described by Brock, Lahr et al. (2000) and Brock, Van Wijngaarden et al. (2000) and included in the European Union Guidance
Document on Ecotoxicology (SANCO 2002). PRC¼principal response curve; 1 ¼no effect, 2 ¼slight effect, 3 ¼clear short-term effect, full
recovery observed, 4 ¼clear effect, no full recovery observed at the end of the experiment, 5 ¼clear long-term effects; #¼decrease of
endpoint; "¼increase of endpoint; #" ¼ both decrease and increase of endpoint is observed.
Impact in Ditch Mesocosm of Simulated Spray Drift—Integr Environ Assess Manag 2, 2006 123
and applications of chlorothalonil in weeks 6 and 7 (Figures 3
and 6). The effects on zooplankton continued into the period
in which fluazinam was sprayed. However, some zooplankton
populations were already showing recovery within this period
(Figure 7B and C). The calculations of TU show that treatment
concentrations for fluazinam lie in the range in which effects
are border line. This is in accordance with Tables 1 and 11,
which show that effects of fluazinam are unlikely.
Effects based on multi- and repeated stress
The exposure concentrations in toxic units (Figure 3) may
indicate potential risks owing to multi or repeated stress.
During the experiment 2 periods of multi-stress can be
recognized. The first is the application period of the 2
herbicides, in which multistress of the 2 herbicides seems to
exhibit a potential risk. However, it is not an actual risk,
because effects of metribuzin at the community level are
unlikely at the concentrations applied in the experiment
(Tables 1 and 11) and effects can be attributed to prosulfo-
carb. The 2nd is the simultaneous application of lambda-
cyhalothrin and chlorothalonil as indicated by Figure 3.
Experimental data seem to confirm an effect by the 2
stressors. The maximum TU was reached after week 9 (Figure
3) after the application of both lambda-cyhalothrin and
chlorothalonil. This was followed by a clear, but not
significant, effect on zooplankton at the 1% level (Figure 6).
This effect was not observed after the application of 1 of the 2
compounds separately in the weeks before. Similar effects at
the 1% level are also clear at the population level within the
population of Hexarthra sp. (Figure 7 and Table 9). At the 5%
level, the effects proceed until recovery starts, which was not
until week 15 (Figure 6). It can be concluded that at the 1%
level minor effects on zooplankton were observed, which
might result from multistress by lambda-cyhalothrin and
chlorothalonil. These effects on zooplankton and macro-
invertebrates seem to be more severe at the 5% level.
During the experiment repeated stress was potentially
caused by the multiple applications of lambda-cyhalothrin
and chlorothalonil. Repeated stress might have played a role
in the effects of chlorothalonil on zooplankton and in the
effects of lambda-cyhalothrin on macroinvertebrates,
although in the experiment these effects cannot be separated
from the effects of the combined application of chlorothalonil
with lambda-cyhalothrin on the 4th chlorothalonil applica-
tion. We can conclude that multi- and repeated stress played a
small role within the applied pesticide package. This was
probably mainly because most of the substances rapidly
dissipated and the absence of many simultaneous applications.
Comparing the findings from previous experiments with
those observed in the present study, it can be concluded that
the patterns of effects observed from the application of the
pesticide program were in line with the effects we could
deduce from the thresholds and concentrations for the
individual compounds, although repeated applications were
used. This may be because of the short dissipation time of
If no effects on aquatic ecosystems are accepted and if spray
drift is considered the only emission route, emission reduction
measures to values below 1% spray drift are necessary. The
current aquatic risk assessment procedure, based on individual
compounds, the Uniform Principles, and a drift emission of
1%, may sufficiently protect aquatic ecosystems if slight and
transient effects are accepted. The crop approach is a
promising approach for risk assessment and evaluation of
effects of realistic pesticide stress. Moreover, a crop approach
is also promising from a risk management point of view,
because it will improve the practicality and acceptance.
Acknowledgment—The present study was financially sup-
ported by the Ministry of Agriculture, Nature Conservation,
and Food Safety and by Syngenta Crop Protection. We thank
Steven Crum, Arie
¨nne Matser, Alex Schroer, Jos Sinkeldam,
Marie-Claire Boerwinkel, and Frans van Tilburg.
Beltman WHJ, Adriaanse PI. 1999. Proposed standard scenarios for a surface water
model in the Dutch authorization procedure of pesticides. Method to define
standard scenarios determining exposure concentrations simulated by the
TOXSWA model. Wageningen (NL): DLO Winand Staring Centre. Report 161.
Figure 8. Dynamics of Flagellatae sp. population belonging to the phytoplankton. This taxon is the only taxon that showed a consistent treatment-related
response (Table 9).
124 Integr Environ Assess Manag 2, 2006—GHP Arts et al.
Boxall ABA, Brown CD, Barrett KL. 2002. Review. Higher-tier laboratory methods
for assessing the aquatic toxicity of pesticides. Pest Manag Sci 58:637–648.
Brock TCM, Crum SJH, Deneer JW, Heimbach F, Roijackers RMM, Sinkeldam JA.
2004. Comparing aquatic risk assessment methods for the phototsynthesis-
inhibiting herbicides metribuzin and metamitron. Environ Pollut 130:403–426.
Brock TCM, Huijbrechts CAM, Van de Steeg-Huberts MJHA, Vlassak MA. 1982. In
situ studies on the breakdown of Nymphoides peltata (Gmel.) O. Kuntze
(Menyanthaceae): Some methodological aspects of the litter bag technique.
Hydrol Bull 16:35–49.
Brock TCM, Lahr J, Van den Brink PJ. 2000. Ecological risks of freshwater
ecosystems, part I: Herbicides. Wageningen (NL): Alterra Wageningen
University and Research Centre. Report 088.
Brock TCM, Van Wijngaarden RPA, Van Geest GJ. 2000. Ecological risks of
insecticides in freshwater ecosystems, part 2: Insecticides. Wageningen (The
Netherlands): Alterra Wageningen University and Research Centre. Report
Caux P-Y, Bastien C, Crowe A. 1996. Fate and impact of pesticides applied to
potato cultures: The Nicolet River Basin. Ecotoxicol Environ Saf 33:175–185.
Caux P-Y, Kent RA, Fan GT, Stephenson GL. 1996. Environmental fate and effects
of chlorothalonil: A Canadian perspective. Crit Rev Environ Sci Tech 26:45–93.
Davies PE. 1988. Disappearance rates of chlorothalonil (TCIN) in the aquatic
environment. Bull Environ Contam Toxicol 40:405–409.
DeLorenzo ME, Serrano L. 2003. Individual and mixture toxicity of three pesticides:
Atrazine, chlorpyrifos, and chlorothalonil to the marine phytoplankton species
Dunaliella tertiolecta.J Environ Sci Health B 38:529–538.
Deneer JW. 2000. Toxicity of mixtures of pesticides in aquatic systems. Pest Manag
Deneer JW. 2003. Bestrijdingsmiddelen en teelten. Een verkenning van de risico’s
van het gebruik van bestrijdingsmiddelen voor waterorganismen binnen
enkele landbouwkundige teelten. Wageningen (NL): Alterra. Report 494.
Drent J, Kersting K. 1993. Experimental ditches for research under natural
conditions. Research Note. Water Res. 27:1497–1500.
Ernst W, Doe K, Jonah P, Young J, Julien G, Hennigar P. 1991. The toxicity of
chlorothalonil to aquatic fauna and the impact of its operational use on a
pond ecosystem. Arch Environ Contam Toxicol 21:1–9.
[EU] European Union. 1997. Commission proposal for a council directive
establishing Annex VI to Directive 91/414/EEC concerning the placing of
plant protection products on the market. Brussels (BE): EU. Official
Fairchild JF, Lapoint TW, Schwartz TR. 1994. Effects of an herbicide and insecticide
mixture in aquatic mesocosms. Arch Environ Contam Toxicol 27:527–533.
Fairchild JF, Sappington LC. 2002. Fate and effects of the Triazinone herbicide
Metribuzin in experimental pond mesocosms. Arch Environ Contam Toxicol
Farmer D, Hill IR, Maund SJ. 1995. A comparison of the fate and effects of 2
pyrethroid insecticides (lambda-cyhalothrin and cypermethrin) in pond
mesocosms. Ecotoxicology 4:219–244.
Fleeger JW, Carman KR, Nisbet RM. 2003. Indirect effects in aquatic ecosystems.
Sci Total Environ 317:207–233.
[FOCUS] FOrum for Co-ordination of pesticide fate models and the USe. 2001.
FOCUS Surface Water Scenarios in the EU Evaluation Process under 91/414/
EEC. Report of the FOCUS Working Group on Surface Water Scenarios.
Brussels (BE): EC SANCO. Report SANCO/4802/2001-rev.2
Ganzelmeier H, Ra utmann D, Spangenberg R, Streloke M, Herrmann M,
Wenzelburger H, Walter H-F. 1995. Studies on the spray drift of plant
protection products. Mitteilungen aus der Biologischen Bundesanstalt fu¨r
Land-, und Forstwirtschaft, Heft 304. Berlin (DE): Blackwell Wissenschafts-
Verlag. 111 p.
Hand LH, Kuet SF, Lane MCG, Maund SJ, Warinton JS, Hill IR. 2001. Influences of
aquatic plants on the fate of the pyrethroid insecticide lambda-cyhalothrin in
aquatic environments. Environ Toxicol Chem 20:1740–1745.
Hommen U, Veith D, Ratte HT. 1994. A computer program to evaluate plankton
data from freshwater field tests. In: Hill IA, Heimbach F, Leeuwangh P,
Matthiesen P, editors. Freshwater field tests for hazard assessment of
chemicals. Boca Raton (FL): Lewis. p 503–513.
Huijsmans JMF, Porskamp HAJ, Van de Zande JC. 1997. Spray drift (reduction) in
crop protection application technology. Evaluation of spray drift in orchards,
field crops and nursery tree crops spraying (state-of-the-art December 1996).
Wageningen (NL): DLO Institute of Agricultural and Environmental Engineer-
ing (IMAG-DLO). IMAG-DLO Report 97–04.
Leistra M, Zweers AJ, Warinton JS, Crum SJH, Hand LH, Beltman WHJ, Maund SJ.
2003. Fate of the insecticide lambda-cyhalothrin in ditch enclosures differing
in vegetation density. Pest Manag Sci 60:75–84.
Maund SJ, Hamer MJ, Warinton JS, Kedwards TJ. 1998. Aquatic ecotoxicology of
the pyrethroid insecticide lambda-cyhalothrin: Considerations for higher-tier
aquatic risk assessment. Pestic Sci 54:408–417.
Maund SJ, Sherratt TN, Stickland T, Biggs J, Williams P, Shillabeer N, Jepson P.
1997. Ecological considerations in pesticide risk assessment for aquatic
ecosystems. Pest Sci 49:185–190.
Moed JR, Hallgraeff GM. 1978. Some problems in the estimation of chlorophyll-a
and phaeopigments from pre- and post-acidification spectrophotometric
measurements. Int Rev Gesamten Hydrobiol 63:787–800.
Muszkat L, Feigelson L, Bir L, Muszka KA. 1998. Reaction patterns in
photooxidative degradation of two herbicides. Chemosphere 36:1485–1492.
Muszkat L, Feigelson L, Bir L, Muszka KA. 2002. Photocatalytic degradation of
pesticides and bio-molecules in water. Pest Manag Sci 58:1143–1148.
[OECD] Organization for Economic Cooperation and Development. 2000. OECD
guidelines for the testing of chemicals. Draft proposal for a new guideline
308. Aerobic and anaerobic transformation in aquatic sediment systems.
Roessink I, Arts GHP, Belgers JDM, Bransen F, Maund SJ, Brock TCM. 2005. Effects
of lambda-cyhalothrin in two ditch microcosm systems of different trophic
status. Environ Toxicol Chem 24:1684–1696.
Ronday R, Aalderink GH, Crum SJH. 1998. Application methods of pesticides to an
aquatic mesocosm in order to simulate effects of spray drift. Water Res
[SANCO] Sante´ des Consommateurs. 2002. Guidance document on aquatic
ecotoxicology in the context of the directive 91/414/EEC. Brussels, Belgium:
European Commission Health and Consumer Protection Directorate General.
SANCO/3268/2001 rev. 4 (final).
Sanderson H. 2002. Pesticide studies—Replicability of micro/mesocosms. Environ
Sci Pollut Res 9:429–435.
Sherratt TN, Roberts G., Williams P, Whitfield M, Biggs J, Shillabeer N, Maund SJ.
1999. A life-history approach to predicting the recovery of aquatic
invertebrate populations after exposure to xenobiotic chemicals. Environ
Toxicol Chem 18:2512–2518.
Ter Braak CJF, Smilauer P. 1998. CANOCO 4. Ithaca (NY): Microcomputer Power.
Tomlin CDS, editor. 2000. The pesticide manual. 12th ed. Furnham (UK): British
Crop Protection Council. 1250 p.
[USEPA] US Environmental Protection Agency. 1998. Guidelines for ecological risk
assessment. Cincinnati (OH): EPA 630/R-95/02F.
Van den Brink PJ, Hartgers EM, Gylstra R, Bransen F, Brock TCM. 2002. The effects
of a mixture of two insecticides on freshwater microcosms, II: Water quality,
responses of zooplankton, phytoplankton and periphyton and ecological risk
assessment. Ecotoxicology 11:181–197.
Van den Brink PJ, Hattink J, Bransen F, Van Donk E, Brock TCM. 2000. Impact of
the fungicide carbendazim in freshwater microcosms, II: Zooplankton, primary
producers and final conclusions. Aquat Toxicol 48:251–264.
Van den Brink PJ, Ter Braak CJF. 1998. Multivariate analysis of stress in
experimental ecosystems by principal response curves and similarity analysis.
Aquat Ecol 32:163–178.
Van den Brink PJ, Ter Braak CJF. 1999. Principal response curves: Analysis of time-
dependent multivariate responses of biological community to stress. Environ
Toxicol Chem 18:138–148.
Van den Brink PJ, Van Wijngaarden RPA, Lucassen WGH, Brock TCM, Leeuwangh
P. 1996. Effects of the insecticide Dursban 4E (active ingredient chlorpyrifos) in
outdoor experimental ditches, II: Invertebrate community responses and
recovery. Environ Toxicol Chem 15:1143–1153.
Van Wijngaarden RPA, Cuppen JGM, Arts GHP, Crum SJH, Van den Hoorn MW,
Van den Brink PJ, Brock TCM. 2004. Aquatic risk assessment of a realistic
exposure to pesticides used in bulb crops, a microcosm study. Environ Toxicol
Wendt-Rasch L, Van den Brink PJ, Crum SJH, Woin P. 2004. The effects of a
pesticide mixture on aquatic ecosystems differing in trophic status: Response
of the macrophyte Myriophyllum spicatum and the periphytic algal
community. Ecotoxicol Environ Saf 57:383–398.
Williams DA. 1972. The comparison of several dose levels with a zero dose
control. Biometrics 28:519–531.
Impact in Ditch Mesocosm of Simulated Spray Drift—Integr Environ Assess Manag 2, 2006 125