Degradation and Transport of the Chiral Herbicide S-Metolachlor at

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Degradation and transport of the chiral herbicide S-metolachlor at the catchment scale: combining observation scales and analytical approaches Marie Lefrancq, Sylvain Payraudeau, Benoît Guyot, Maurice Millet, and Gwenaël Imfeld Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b02297 • Publication Date (Web): 22 Oct 2017 Downloaded from http://pubs.acs.org on October 26, 2017

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Environmental Science & Technology

Infiltration Runoff

Soil Degradation Volatilization SM MESA Dissolved

PLOT

SM Particulate

MESA MOXA

SM Dissolved SM Plot export Particulate 8.1% ACS Paragon Plus Environment

MOXA Catchment export 11.2%


1

Degradation and transport of the chiral herbicide S-

2

metolachlor at the catchment scale: combining

3

observation scales and analytical approaches

4

Lefrancq Marie.a,b, Payraudeau Sylvain.a*, Guyot Benoit.a, Millet Maurice.c, Imfeld

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Gwenaël.a

6 7

a

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University of Strasbourg, CNRS, ENGEES, 1 Rue Blessig, 67084 Strasbourg cedex, France

9

b

Laboratory of Hydrology and Geochemistry of Strasbourg (LHyGeS UMR 7517),

LETG-Angers (UMR CNRS 6554), University of Angers, 2 bd Lavoisier, 49045 Angers,

10

France

11

c

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Strasbourg, CNRS, 1 rue Blessig, 67084 Strasbourg cedex, France

Atmospheric Physical Chemistry Department (ICPEES UMR 7515), University of

13 14

*Corresponding author: [email protected]

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Laboratory of Hydrology and Geochemistry of Strasbourg (LHyGeS UMR 7517), Université

16

de Strasbourg, CNRS, ENGEES, 1 Rue Blessig, 67084 Strasbourg cedex, France. phone: + 33

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(0)3 6885 0437 and fax: +33 (0)3 68 85 04 02

18 19

Manuscript for Environmental Science and Technology

ACS Paragon Plus Environment

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Abstract

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Evaluating pesticide degradation and transport in the soil-surface water continuum remains

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challenging at the catchment scale. Here we investigated the dissipation of the chiral herbicide

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S-metolachlor (SM) in soil in relation to its transport in runoff. Analyses of SM,

24

transformation products (TPs, i.e. MESA and MOXA), and enantiomers were combined to

25

determine SM degradation at plot and catchment scales. Assisted by modelling, we found that

26

the main dissipation pathways of SM at the plot scale were degradation (71%), volatilization

27

(5%), leaching (8%) and runoff (3%), while 13% of SM persisted in top-soil. This highlights

28

the relevance of degradation processes. TPs could trace the different discharge contributions:

29

MOXA prevailed in runoff water, whereas MESA was associated with slower flowpaths. At

30

the catchment outlet, 11% of SM applied was exported in dissolved or particulate phases or as

31

TPs (in SM mass equivalent). A single event one week after application exported 96% of SM,

32

which underlined the potential importance of severe rainfall on seasonal SM export.

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Enantioselective degradation enriched SM in the R-enantiomer over longer periods and may

34

be associated with slower flowpaths. Altogether, combining observation scales and analytical

35

approaches enabled to quantify SM degradation and to identify how degradation controls SM

36

export at the catchment scale.



37

Introduction

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Understanding pesticide degradation in agricultural systems is essential as degradation

39

is the only process that can reduce mass transport of pesticides.1 Pesticide degradation is

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controlled by a complex interplay between physico-chemical characteristics of molecules and

41

environmental conditions, including crop management practices, climatic forcing and water

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flowpaths.2–4 Degradation processes are often estimated at agricultural plot scales,5,6 whereas

43

water management measures are defined on a larger scale, such as that of the hydrological

44

catchment.7 Degradation is rarely evaluated at the catchment scale8 because degradative

45

processes can hardly be distinguished from non-degradative ones, such as dilution or

46

sorption.1 As a result, relevant quantitative information is mostly lacking to understand factors

47

driving pesticide degradation in relation to transport at the catchment scale.

48 49

Chloroacetanilides belong to the top ten pesticide classes used worldwide.9 They are

50

used for pre-emergent control of annual weeds, mainly on corn, sugar beet and sunflower.

51

Chloroacetanilides are frequently detected in surface and ground water together with their

52

transformation products (TPs), notably ethane sulfonic acid (ESA) and oxanilic acids

53

(OXA).10–14 S-metolachlor (SM) is one of the most used chloroacetanilides and equates to

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4.2% of the global use of pesticides.1 Similar to approximatively 25% of currently used

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chemical pesticides, SM is a chiral compound, i.e. with two enantiomers which are non-

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superimposable mirror images of each other.15 Enantiomers of chiral pesticides have identical

57

physico-chemical properties.16 However, enantiomers may interact selectively with molecular

58

receptors and thus differ in their biological efficacity, toxicity and environmental

59

behaviour.17–20 Enrichment of specific enantiomers in environmental compartments can

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indicate biodegradation of chiral pesticides.21,22 The combination of different analytical

61

approaches, such as parent compound and TP quantification, and enantiomeric signatures can

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help to identify and disentangle the main dissipation pathways of SM (degradation vs.


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export).21,23 To the best of our knowledge, simultaneous evaluation of SM, its enantiomers

64

and major TPs in surface runoff at the catchment scale has not been reported yet.

65 66

Therefore, the purpose of this study was to evaluate the degradation of SM in soil in

67

relation to its export in runoff concomitantly with its major TPs, i.e., metolachlor-ESA

68

(MESA) and metolachlor-OXA (MOXA). SM was used as a proxy for moderately soluble

69

herbicides. Our specific objectives were to (i) identify the dissipation pathways of SM in

70

relation to hydrological conditions, and (ii) to quantify transport of SM and its main TPs at

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two spatial scales: an experimental plot (77 m²) nested within a catchment (47.8 ha).

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Measurements on both scales were strengthened with pesticide mass-balance modelling to

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quantify the contribution of degradation and transport of SM in the soil-surface water

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continuum. Beyond the quantification of SM, enantiomer fractionation and a mass-balance

75

analysis of MESA and MOXA production in the soil and export in runoff enabled to estimate

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the contribution of degradation to SM dissipation at the catchment scale.

77 78

Material and methods

79

Study catchment. The study site is a 47.8-ha catchment that is part of the Zorn catchment of

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750 km², situated in the lower Alsatian hills (Bas-Rhin, France) (Figure 1).24 Briefly, it is

81

characterized by silty-clay soils and a mean annual precipitation of 605 mm. Arable land

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comprises 88% of the catchment area (of which, in 2012, 68% was corn, 16% winter wheat

83

and 4% sugar beet). Corn and sugar beet are sown between mid-March and end of May.

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Natural drainage and a 5-cm diameter subsurface pipe draining an unknown part of the study

85

area discharged into a ditch located 102 m upstream of the catchment outlet (Figure 1). This

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drainage continuously sustained the baseflow observed at the catchment outlet during the

87

study period (0.43 ± 0.15 L-1s-1).

88



89

To isolate the contribution of subsurface and runoff flow, a 77 m² portion of a sugar

90

beet plot was studied in the upstream area of the catchment (plot is in dark orange in

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Figure 1). This plot is considered representative of the corn and sugar beet area in the

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catchment because soil types, slope and crop management techniques are spatially

93

homogeneous within the catchment.24 The plot was hydrologically isolated from runoff and

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subsurface flow in the plowing layer from neighboring plots with a 60-cm high HDPE shield

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to a depth of 30 cm below the ground surface, as saturation excess overland flow is unlikely

96

to occur24. The study was carried out from March 12 to July 10 (Day 0 to 120) at the plot

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scale, and up to August 14, 2012 (Day 155) at the catchment scale.

98 99

SM characteristics and applications. SM is nonionic and moderately soluble in water,

100

whereas MESA and MOXA have a higher polarity and solubility and lower Kd values as

101

reported in the Supporting Information (Table S1).25,26 Racemic metolachlor has been applied

102

since the early 1990s at the study site, and, since 1998-2000, it has been replaced by SM,

103

which is enriched in the herbicidally active S-enantiomer.27,28 A total of 10.95 kg of SM was

104

applied in 9 plots in the catchment during the study period (Figure 1), with the percentage of

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total SM by formulation of Mercantor Gold® (Day 31, 4.9% and Day 50, 4.9%), Camix® (Day

106

63, 78.9%) and Dual Gold® (Day 62, 11.3%). Only Mercantor Gold® was applied twice (Day

107

31 and Day 50) in the experimental plot. On Day 31, 16 petri dishes were used to quantify SM

108

deposition on the experimental plot (Annex S1), which ranged from 192 to 336 g ha-1 for 288

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g ha-1 reported by the farmer. This highlights the spatial variability of SM deposition at the

110

plot scale. Drift and rainfall depositions of metolachlor were previously measured during 64

111

days for similar rainfall and higher application rate.29 This may represent 0.7 µm) to mass equivalent

164

load of SM (MELSM) was evaluated as a percentage of total loads of SM (illustrated here for

165

MESA [%]).

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 MWSM  load (MESA) ×    MWMESA  × 100 %MESA = MELSM

(2)

167

Enantiomer fraction. The enantiomer fraction (EF [-]) was used to describe enantiomer

168

composition of SM:31 =

+

(3)

169

The EF of S-metolachlor in the commercial formulations is approximately 88% S isomer and

170

12% R isomer32, and is obtained through a selective and likely identical synthetic

171

manufacturing process by Syngenta.32–34 The EF of the commercial product Mercantor Gold®

172

applied in the studied plot was 0.866 ± 0.008 (n = 8), which is consistent with previous

173

studies for Dual II®, Dual magnum®32,33 or for Mercantor Gold®23, indicating an EF value

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ranging between 0.872-0.882. Shift in the enantiomer fractions was calculated as follow: ∆ = −

175 176

(4)

A positive ∆EF thus indicates S-enantiomer enrichment, and, inversely, a negative one indicates R-enantiomer enrichment.

177 178

Pesticide mass balance model. A parsimonious model35,36 was developed to estimate the

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different terms of the SM mass balance (i.e. degradation, volatilization, infiltration, runoff

180

loads and remaining mass in top-soil). Rainfall partitioning between infiltration and runoff

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was determined using the Green-Ampt method37, which is adapted to the studied catchment.24

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Retention and release of SM in the top-soil was simulated using a mixing layer approach36,38



183

and assuming linear equilibrium sorption39 (Table S3). Despite a moderate risk of

184

volatilization for SM according to its Henry’s law constant (Table S1), significant

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volatilization rates have been observed previously40 and were thus implemented in the model

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(Annex S3). Long range atmospheric transport of metolachlor has never been shown to be

187

significant,41 and was therefore neglected. Between rainfall events, SM degradation and

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volatilization were estimated by a first order dissipation rate (kb), fitted on the observed SM

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concentrations in the top-soil. Predictions were compared with the observed dissolved

190

concentrations in runoff water (Figure S1). Degradation led to the formation of TPs and it was

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assumed that MESA and MOXA were the only TPs for SM (in testing the full range of

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MESA/MOXA production ratios) and that they have the degradation half-lives and sorption

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coefficients reported in Table S1. Abiotic degradation (e.g. photolysis and hydrolysis) of SM

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was considered negligible compared to biodegradation in soil.42 Plant uptake was not

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explicitely integrated in the model as no correlation could be found between the remaing mass

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of SM in the top-soil and the sugar beet growth (from 0 to 40 cm during the investigation

197

period) or the surface cover (from 0 to 100%). The confidence intervals (95% CI) of model

198

predictions, including parameter uncertainties for SM sorption and degradation, were obtained

199

from best runs (Nash-Sutcliffe criterion > 0.55, n = 138) using a Monte-Carlo method (1000

200

permutations) (Table S3).

201 202

Results and discussion

203 204

SM dissipation in the top-soil. Before applications, SM concentrations in top-soil (0 to 1 cm)

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reached 0.006 µg gdw-1 emphasizing the low risk of SM runoff after a year. Following SM

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applications, SM concentrations in the top-soil gradually decreased from 2.1 to 0.5 µg gdw-1

207

after 89 days (Day 120) (Figure 2B and Table S4). Changes in SM concentrations in the top-


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soil was fitted with the model by a dissipation rate (kb) including both degradation and

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volatilization. Accordingly, kb was evaluated to 0.05 day-1 (DT50 = 14 days), which is in the

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range of literature values.43,44 Among the tested range of cumulative volatilization losses, only

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the low volatilization hypothesis, corresponding to 4.5% of SM, was consistent with SM

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concentrations measured in top-soil (Figure S1C). However, the dissipation rate

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underestimated the mass of SM remaining in the soil during the second part of the study

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period (model ‘kb constant' in Figure 2B). Therefore, a gradual decrease of the rate, which

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reached 0.005 day-1 (i.e. DT50 = 140 days) 30 days after application, was fitted to estimate the

216

SM mass in soil until the end of the study (model ‘2 successive kb’ in Figure 2B). As most of

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the volatilization occurred during the 24h following the application40, the use of a lower

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dissipation rate during the second study period suggests that SM sorption onto the soil

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reduced its availability for degradation over time, thereby increasing its persistence in the

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soil.45,46 SM reached its maximum concentration in the depth 2-5 cm (Figure 3A). Overall, the

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SM mass remaining in the plot corresponded to 13% (± 2%) of the applied mass in the top-

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soil (0-1 cm) after 89 days (Day 120) (or 25% (± 4%) in the first soil meter, Day 122) (Table

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1).

224 225

The concentrations of MESA and MOXA in top-soil were systematically below the

226

quantification limits, i.e. 1 µg gdw−1. This was confirmed by MESA and MOXA predicted

227

concentrations (Figure S2). SM was enriched in the R-enantiomer before SM application

228

(∆EF = -0.057 ± 0.013) (Figure 2B). This highlights preferential degradation of the S-

229

enantiomer in the top-soil during the winter preceeding the study period. Following

230

applications, the ∆EF values in the top-soil averaged 0.001 in the soil, ranging from -0.005 to

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0.01 (Figure 2B). These values reflect the signature of freshly applied, non-degraded SM. No

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significant shift of enantiomer fractions could be observed between the first SM application



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(Day 31, April 12) and Day 99 (June 19) (Figure 2B). This suggests either no significant

234

biodegradation of SM during the first 68 days following applications, or no significant

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enantioselective processes. Then, the R-enantiomer was preferentially degraded in the top-soil

236

(0-5 cm) (∆EF = 0.032 ± 0.004, provided as mean ± σ over the samples). In contrast, the S-

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enantiomer was preferentially degraded in deeper soil layers (∆EF = -0.028 ± 0.013) (5-

238

100 cm) (Figure 3B). Different shifts of enantiomer fractions with increasing soil depth

239

suggest changes of bacterial community47, and/or to decreasing oxygen concentrations with

240

depth. Shifts of enantiomer fractions were also observed in lab-scale wetlands, where ΔEF of

241

SM was lower in the anoxic zone than in oxic one.21

242 243

Athough the SM pool in soil was depleted, in situ SM dissipation pathways could not

244

be precisely identified because TPs in the soil could not be detected and shifts of enantiomer

245

fractions present opposite trends in surface and deeper layers. A detailed mass balance is thus

246

needed to infer the contribution of SM dissipation pathways in the soil (i.e., degradation vs

247

export).

248 249

SM mass balance: degradation versus dissolved and particulate export. At the

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experimental plot scale, dissolved concentrations of SM in runoff correlated with the SM

251

mass in top-soil (ps < 0.001) (Figure S1). While successive runoff gradually decreased the SM

252

pool in top-soil, SM remained largely available for off-site transport across the entire study

253

period. Based on measured concentrations and predicted discharges, SM export was estimated

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to 3.7% of the applied SM mass, with a dissolved and particulate contribution of 97.9% and

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2.1%, respectively (Table 1). The main predicted SM dissipation pathways in the top-soil

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(± 95% confidence interval) were degradation (71 ± 4% of applied SM), followed by SM


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leaching via infiltration (8 ± 3%), volatilization (5 ± 4%), runoff (3 ± 1%) and SM remaining

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in the top-soil (13 ± 2%) at the end of the observation period (Day 120) (Figure S1C).

259 260

At the catchment scale, 3.4% of the applied SM was exported during the study period

261

in both the dissolved and particulate phases of runoff (Table 1). This is an order of magnitude

262

higher than export coefficients previously reported at the event scale (0.16%)4 or for several

263

years (0.072%)48 and similar to yearly exports coefficients (1.0 ± 1.6%) reported for 54 US

264

streams.14,49 Export of SM mainly occurred during nine significant runoff events (> 10 m3)

265

that represented 69% of the total outflow water volume (i.e. 17,419 m3) (Table S5), with 53%

266

for one single event (May 21). The catchment lag time, defined as the time between the

267

rainfall gravity center and the resulting peak discharge, ranged between 10 and 184 min.24

268

The contribution of drainage was 8.6 ± 6.7% during the nine major runoff events (> 10 m3)

269

and 33% during the entire study period (Table S5)24, as evaluated with the constant-slope

270

method for hydrograph separation.50 Discharge was then predominantly controlled by fast

271

flowpaths driven by Hortonian runoff (i.e unsaturated overland flow), with a lower volumetric

272

contribution of slow flowpaths driven by drainage.24 The hydrochemical signature of the drain

273

water (e.g. TSS, TIC, DIC, TOC, DOC, PO43−, Ptotal, NO2− and NH4+) also differed from that

274

of runoff from the catchment (pW < 0.05), suggesting different flowpaths (Table S6 and S7).

275

The TSS export during the study period represented 10.1 t ha-1, i.e. a top-soil loss of

276

0.4 mm over the catchment, which is in the range of regional annual erosion for a corn

277

catchment.51 Thus, 46% of the SM was exported in association with the particulate phase

278

(>0.7 µm), in agreement with previous estimates at the plot scale (20 - 46%)52,53 but higher

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than export reported at the catchment scale (12%)54. The largest runoff event occurred on May

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21, one week after applications, with 54 mm in 4 h and a runoff coefficient of 40.8%24, and

281

accounted for, respectively, 92%, 42%, 87% and 96% of TSS, DOC, POC and SM exported



282

during the study period. TSS (60 g L-1), NO3- (92 mg L-1) and SM (62 µg L-1 and 34 µg g dw -1)

283

concentrations were particularly high during this intense event, whereas other physico-

284

chemical characteristics, especially DOC or POC, remained at background concentrations

285

(Figure S3). This suggests that the extreme event did not present particular characteristics,

286

except for TSS, which can be explained by the high intensity and duration of the rainfall

287

event. The study catchment is frequently prone to significant mudflows,55 and following this

288

event, the soil surface of the arable land was crusted, which enhanced runoff in subsequent

289

rainfall events. Such an erosive event may mobilize not only freshly applied SM, but also a

290

persistent fraction of SM in the soil.56 Without this event, SM export from the catchment

291

would represent only 0.1%, of which 85% of the total loads should have been found in the

292

dissolved phase (< 0.7 µm).

293 294

Rainfall intensity positively correlated with TSS and SM concentrations in both the

295

dissolved and particulate phases (pS < 0.001). The Kd values of SM negatively correlated with

296

TSS concentrations at the catchment scale (pS < 0.001) (Table S8 and Annex S5), which is

297

consistent with lower POC concentrations with increasing TSS concentrations (ps < 0.05).

298

This suggests that organic carbon of eroded materials progressively get depleted, which in

299

turn decreased SM sorption onto TSS.57,58 However, Koc values largely varied over time with

300

a large range of values (7 – 16180 ((L kg-1)/(gC g-1)), Table S8), whereas stable Koc over time

301

was expected if POC concentrations were the only reason for SM sorption. Koc values were

302

about two orders of magnitude lower than those estimated previously.57 This underlines that

303

Koc or Kd values obtained from pesticide concentrations in surface water57,59 can hardly be

304

compared as equilibration sorption time, and to a lesser extent sorbent characteristics (size of

305

the particles, aromaticity and polarity), largely vary between runoff events.

306


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Overall, the complete mass balance of SM enabled to quantify the contribution of

308

degradation versus export at the plot scale. Similar export coefficients of SM at the plot and

309

catchment scales (3.7% and 3.4%) (Table 1) support the idea that SM dissipation mainly

310

occurred at the plot scale, and to a lesser extent during transport from the plot to the

311

catchment outlet. The analysis of MESA and MOXA may thus provide another line of

312

evidence to estimate SM degradation at the catchment scale.

313 314

Production and export of MESA and MOXA from soil. Concentrations of MESA and

315

MOXA were systematically below the quantification limits in soils and TSS. At the

316

experimental plot scale, MESA and MOXA in runoff generally contributed equally or more to

317

the MELSM compared to SM (%MOXA + %MESA = 52.1 ± 44.4%) (Figure 4C). The

318

seasonal MELSM export coefficient reached 8.1% of the applied SM mass, including 3.7% of

319

MOXA and 0.7 % of MESA (in SM mass equivalent). At the catchment scale, the seasonal

320

MELSM export coefficient reached 11.2% of the applied SM, including 5.5% of MOXA and

321

2.4% of MESA. MOXA was frequently detected (88% and 53% detection frequency at the

322

plot and catchment scales, respectively), and more so than MESA (25% and 29%,

323

respectively) (Table S9).

324

At the catchment scale, SM and MOXA concentrations were significantly higher

325

during runoff events compared to low-flow periods mainly controlled by drainage

326

(pW < 0.01), whereas MESA concentrations did not significantly differ (pW > 0.05). The ratios

327

of SM/MOXA/MESA at both scales highly varied over time (Fig 4C and 4D). Before major

328

applications (Day 29 for plot and Day 58 for catchment, as only 10% of SM was applied in

329

upstream areas before), TPs predominantly contributed to MELSM with mainly MOXA at the

330

plot and MESA at the catchment scale. At the plot scale, the two runoff events following

331

applications on Day 36 and 50 exported mainly freshly applied SM. Then from Day 78



332

onwards, the contribution of MOXA to MELSM gradually increased until the end of the study

333

period (Day 120), which indicates SM degradation. SM degradation was also supported by a

334

shift of the S-enantiomer fraction in top-soil samples from Day 91 (Figure 2B). At the

335

catchment scale, the temporal variability of the ratios SM/MOXA/MESA was less clear (Fig

336

4D), likely due to overlapping and complex hydrological pathways. At Day 8 and at the end

337

of the study period (Day 120-155), most of the exported MELSM seemed to be SM instead of

338

TPs. However, this finding may be an artifact because LOQ was two orders of magnitude

339

lower for SM compared to TPs.

340 341

The high contribution of TPs to the MELSM in runoff water at both scales, although

342

TPs were not detected in the soils or TSS, suggests that MESA and MOXA were mostly

343

present in the soil solution due to their high water solubility and low Koc.12,25,42,60 The higher

344

mass export rate of MOXA compared to MESA at both scales may be due either to a smaller

345

production of MESA in the top-soil or due to more significant off-site transfer of MOXA

346

from top-soil to runoff. Both processes likely coexisted in the top-soil. Fluctuations of soil

347

moisture in top-soil (Figure 4A) with an alternation of oxic and anoxic conditions may favour

348

MOXA production in soils, as previously observed in an artificial wetland during wet/dry

349

cycles.23 In addition, MESA is less sorptive than MOXA60 (Table S1). Therefore, it may

350

preferentially leach from top-soil at the beginning of rainfall and thus be less available, when

351

ponding and runoff occur.25,35,42,61

352

An estimated TP formation ratio of 70/30 (MOXA/MESA) for unsaturated zones61

353

supports the idea that MOXA may be preferentially produced. Similarly, about two moles of

354

MOXA were exported from the catchment for each mole of MESA, which corresponds

355

approximately to the ratio of runoff (67%) to drainage (33%) volumes in the study catchment.

356

However, without the May 21 event, MELSM reached 2.8%, with 0.43% for MOXA and


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2.27% for MESA. Hence, without the impact of the May 21 event, about four moles of MESA

358

would have been exported for each mole of MOXA giving a ratio of runoff (22%) to drainage

359

(78%). This suggests that both SM and MOXA were associated with faster surface flows,

360

such as Hortonian runoff, as a result of shorter half-life (Table S1), whereas MESA prevailed

361

in slower and longer flowpaths that sustain baseflow. This is in agreement with previous

362

predictions, which showed that MESA was associated with groundwater recharge, whereas,

363

SM stemmed from event-driven runoff

364

the unsaturated zone compared to MOXA.14,61

62

and that MESA was more deeply transported into

365 366

The TP analyses in runoff directly confirmed that at least 7.9% (5.5% for MOXA and

367

2.4% for MESA) of SM was degraded at the catchment scale. In addition, the model

368

predictions underscored that a significant fraction of MESA and MOXA leached (57% of the

369

total mass of applied SM) or was degraded (4%) at the plot scale during the study period.

370

However, the plot-based pesticide model disregards the travel time and the different pathways

371

of SM from the different plots to the catchment outlet. In addition, SM degradation may

372

generate other TPs that are mostly unknown. In this case, enantiomer analyses may strengthen

373

the evaluation of SM degradation at the catchment scale.

374 375

Enantioselective behaviour of SM in runoff. Prior to SM applications, negative ∆EF were

376

observed in runoff water from the catchment (∆EF = -0.37 ± 0.01) (Day 8) and in runoff

377

water from the plot (-0.067 ± 0.009) (Day 29) (Figure 4E and 4F). Hence, SM in discharge

378

water from the catchment outlet contained only 49.1% of the S-enantiomer (i.e. EF=0.49),

379

whereas the applied commercial product contained 86.6% of the S-enantiomer. This is

380

consistent with the finding that the SM pool available for runoff in the top-soil was also

381

depleted in the S-enantiomer (∆EF = -0.057 ± 0.013) (Day 8) (Figure 2B). The pool of SM



382

prone to runoff may thus partly correspond to the SM fraction degraded in the top-soil where

383

enantioselective degradation occured. However, the ∆EF value of catchment discharge was

384

one order of magnitude lower than that of the plow layer at Day 8 (-0.37 and -0.057

385

respectively). Enantioselective processes may have occurred deeper in the soil as suggested

386

by higher ∆EF values in the top-soil compared to those found in deeper soil layers (Figure 3)

387

and/or during slower flowpaths driven by natural drainage in fall and winter. This also

388

supports previous findings showing that ∆EF of SM negatively correlated with catchment

389

area, suggesting that catchment retention leads to a greater loss of the S-enantiomer.63

390 391

Following SM application, no significant shift of ∆EF could be observed in runoff

392

water from the plot or the catchment (Figure 4E and 4F). Hortonian runoff prevailed during

393

the season and mainly mobilized SM, which was not enantiomerically enriched in the top-soil.

394

Enantiomeric analysis may thus not indicate degradation or export flowpaths of SM in the

395

first month following SM application. Enantioselective microbial degradation of metolachlor

396

has been shown in laboratory conditions27,64 but remains controversial in experimental

397

wetlands.21,23,65,66 Under field conditions, the fate of SM enantiomers remains poorly

398

described. Enantioselective shift was first accounted for an environmental response to the

399

introduction of non-racemic SM in the 2000s31, and enantioselective degradation was then

400

assumed in 2006-200767. In this study, SM concentrations in the different environmental

401

matrices (dissolved and particulate phases in runoff water, sediment and soil samples)

402

positively correlated with ΔEF values (pS < 0.05). This indicates that samples with low SM

403

concentrations were associated with low ΔEF, i.e. depleted in the S-enantiomer compared to

404

the commercial product. In addition, SM in a grab water sample (< 0.7 µm) collected at Day

405

253 was depleted in the S-enantionmer with ∆EF = -0.050 ± 0.009 (data not shown). This

406

confirmed that enantioselective processes, enriching SM in the R-enantiomer, occured over


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Page 19 of 32


407

longer time scales and may be associated with slower flowpaths at the catchment scale.

408

Altogether, the results emphasized a slight increase of ΔEF in the top-soil during the growing

409

season, and negative ΔEF several months after the applications in fall or winter, at the

410

beginning of the new growing season. These opposite trends may be related to a seasonal

411

effect (temperature and nutrient availability) over the year, impacting the diversity of

412

microoganisms associated with SM degradation in the agricultural soil.68 A year-long field

413

survey would be required to confirm the enantioselective trends observed during SM

414

degradation.

415 416

Environmental implications for herbicide transport in agricultural catchments

417

By means of two investigative scales assisted with a modelling approach, we could

418

evaluate SM dissipation pathways at the catchment scale. A combination of analytical

419

approaches enabled to disentangle SM degradation from other non-degradative processes.

420

One significant implication of our results is that export of TSS-bound pesticides should be

421

accounted for to evaluate risks of moderately soluble pesticides transport in loess covered

422

catchments prone to extreme rainfall-runoff events. Export of SM sorbed onto TSS did not

423

prevail at the plot scale (2.1%) but represented 46% of the total load of SM exported by

424

runoff at the catchment scale, mainly due to a single and extreme rainfall-runoff event. The

425

impact of this event also underlines how severe rainfalls may alter seasonal export of SM.

426

Pesticide application during periods prone to muddy floods that occur in the beginning of

427

spring when bare soil prevails should therefore be subject to caution. Approximately 450 km²

428

of crops on sensitive loess surface soil drain into the Zorn River (Figure 1B). This underlines

429

the potential contribution of both dissolved and particulate SM flows from more than 900

430

headwater catchments with features similar to those of the study catchment. Sorption to and

431

transport with TSS may reduce, bioavailability45,69, biodegradation and toxicity70 of SM. In



432

parallel, the gradual degradation of the S-enantiomer leads to exported SM residues, which are

433

also herbicidally less active over time. Such temporal changes of SM fate may in turn control

434

the exposure level of biota to SM and its enantiomers.

435 436

From a methodological point of view, the combined scales approach (in isolating the

437

runoff component at the plot) and the combined analysis (especially MOXA and MESA

438

ratios) may help to distinghuish flowpaths (fast and slow) contributing to catchment

439

discharge, by analogy to groundwater dating with MESA.12 This potential approach to trace

440

the contribution of different flowpaths can be reused in other agricultural catchments because

441

metolachlor is often detected concomitantly with MESA and MOXA.14

442 443

Multiple lines of evidence for SM degradation were observed. TP export represented

444

more than 50% of the total molar loads of SM exported at both plot and catchment scales, and

445

pronounced enantiomer fractionation of SM could be observed in the field. MESA and

446

MOXA concentration and loading patterns provided direct evidence of SM degradation and

447

preferential water flowpath in the short term. Complementarily, R/S enantiomers analysis may

448

emphasize enantio-selective degradation on medium and longer terms. However, only few

449

quantitative information on persistence of enantiomers and TPs have been reported so far.

450

From a regulatory point of view, there is a growing need not only to improve the

451

interpretation of physico-chemical characteristics of TPs but also to consider pesticide active

452

ingredients manufactured as isomeric mixtures. Chiral analysis of pesticides to evaluate in-

453

situ degradation processes must be done with caution because mechanistic aspects of

454

enantioselectivity of biotic and abiotic reactions have not been fully explored yet. Reference

455

experiments for enantiomer fractionation are currently missing to evaluate the extent of

456

degradation and relate it to TP formation. From a modelling point of view, predicting


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Page 21 of 32


457

pesticide transport in runoff is burdened by numerous uncertainties as sorption and

458

degradation processes may counterbalance each other.71 Models may incorporate enantiomer

459

fractionation under field conditions to improve the estimation of degradation, and thus assist

460

interpretation of complex data sets, while reducing uncertainties in risk assessments.

461 462

Acknowledgements

463

This work was funded by the European INTERREG IV program Upper Rhine

464

(PhytoRET project C.21). Support from REALISE, the Network of Laboratories in

465

Engineering and Science for the Environment in the Alsace Region (http://realise.unistra.fr) is

466

also gratefully acknowledged. Marie Lefrancq was supported by a fellowship from Région

467

Alsace. The authors thank the farmers, the soil laboratory EOST UMS 830 CNRS and

468

Martine Trautmann, Anne Véronique Auzet, Tristan Meyer, Diogo Reis, Jeanne Dollinger,

469

Thomas Dreidemy, René Boutin, Sophie Gangloff, Marie-Pierre Ottermatte, Eric Pernin and

470

Brian Sweeney for their support in data collection, sampling and analysis. We would like to

471

thank Richard Coupe and Pablo Alvarez-Zaldivar for proof-reading the manuscript and three

472

anonymous reviewers for relevant comments.

473 474

Supporting information

475

Additional information contains the summary of the physico-chemical properties of

476

SM, MESA and MOXA, detailed description of analysis standards and norms, values and

477

sources of the dissipation model parameters, SM concentrations and enantiomeric signature in

478

soil and sediment samples, hydrological characteristics of the runoff events, hydrochemical

479

data of the plot, drain and catchment outlets, detailed calculations for Kd and Koc values, data

480

on SM, MESA and MOXA detection frequencies, predicted concentrations of SM, MESA and



481

MOXA, detailed procedure for extraction of SM, MESA and MOXA from solid samples and

482

for the SM deposition measurement during the application, detailed calculations for

483

volatilization or rainfall deposition estimation. This material is available free of charge via the

484

Internet at http://pubs.acs.org/.

485

References

486 487 488 489 490 491 492 493 494 495 496 497 498 499 500 501 502 503 504 505 506 507 508 509 510 511 512 513 514 515 516 517 518 519 520 521 522 523 524

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Figures

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Figure 1. Location of the study catchment in Bas-Rhin, in France (A) within the Zorn

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catchment (B), and the experimental setup and landuse in 2012 (C).

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Figure 2. Daily rainfall and temperature, weekly top soil moisture and SM application at the

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plot scale (A), compared with the measured and predicted remaining mass of SM in the top

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soil (0-1 cm), with constant or successive dissipation rates (kb) model in relation with its

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enantiomer enrichment (∆EF) (B). Error bars correspond to the analytical uncertainty

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calculated via error propagation and incorporating accuracy and reproducibility of n = 3

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measurements. Confidence intervals (95%) were derived from 123 runs (Nash-Sutcliffe

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criterion > 0.55) using a Monte-Carlo approach (1000 runs).

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Figure 3. SM mass in the soil profile on July 12th (Day 122) normalized by the SM mass

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applied in relation to the organic matter content (A), and compared with the vertical

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enantiomer enrichment in the soil profile (B). Error bars correspond to the analytical

722

uncertainty calculated via error propagation and incorporating accuracy and reproducibility of

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n = 3 measurements.

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Figure 4. Temporal changes of SM use, hydrological conditions and TSS loads at the plot (A)

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and catchment scales (B), compared with weekly MELSM values (white diamonds: low flow,

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black triangles: high flow, defined by a runoff event > 10 m3) with relative fractions of MESA

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and MOXA and of dissolved and particulate SM in runoff at the plot (C), and at the catchment

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outlets (D) in relation to enantiomer enrichments of SM in soil, runoff ( 0.7

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µm) and sediment samples at the plot (E) and the catchment scales (F).



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Figure 1.

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Page 29 of 32

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Figure 2

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Page 30 of 32

Figure 3.

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Figure 4.

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Table 1. SM mass balance from March 12 to July 10, 2012 (120 days) at the plot scale and up to August 14 (155 days) at the catchment scale. The unit haapplied refers to the SM load normalized by the surface where SM was applied. Unit [ha] [ha]

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Page 32 of 32

Plot 7.7 10-3 7.7 10-3

Catchment 47.8 12.5

Area Application area SM application and stock in the soil Stock in the soil (top 1cm) before applications [g] 6.9 10-3 ± 1.1 10-3 nc Application in 2012 [g] 4.4 10.95 103 Application in 2012 [g haapplied-1] 571.4 874.6 Stock in the soil (top 1cm) at day 36 nc [g] (% SM applied) 2.4 ± 0.4 (55 ± 9) Stock in the soil (top 1cm) at day 120 nc [g] (% SM applied) 0.6 ± 0.1 (13 ± 2) Stock in the soil (top 1m) at day 122 nc [g] (% SM applied) 1.1 ± 0.2 (25 ± 4) Hydrology Outflow discharge (0-120 days)a [m3 day-1 ha-1] 5.2 ± 28.3 (0 - 270.2) 3.0 ± 18.1 (0 -195.2) a,b Runoff coefficient (0-120 days) [%] 30.9 ± 27.2 (0.7 - 83.1) 7.8 ± 12.6 (0.2 - 40.8) Erosion TSS export (0-120 Days) [t ha-1] 1.3 10.0 SM and TPs export in runoff/dischargeb 20.8 ± 1.7 15.8 ± 1.3 Dissolved export (0-120 days) [g haapplied-1] c -1 0.4 ± 0.1 13.5 ± 2.2 Particulate export (0-120 days) [g haapplied ] 3.7 ± 0.3 3.4 ± 0.4 Total export (dissolved, particulate) (0-120 days) [%] -1 3.9 ± 1.1 20.7 ± 5.7 MESA-dissolved export (0-120 days) [g haapplied ] 21.3 ± 8.2 47.7 ± 18.4 [g haapplied-1] MOXA-dissolved export (0-120 days) -1 46.4 ± 11 97.7 ± 27.5 Total export (dissolved, particulate, TPs) (0-120 days) [g haapplied ] 8.1 ± 1.9 11.2 ± 3.1 Total export (dissolved, particulate, TPs) (0-120 days) [%] -1 nc 98.4 ± 27.6 [g haapplied ] Total export (dissolved, particulate, TPs) (0-155 days) nc 11.2 ± 3.2 Total export (dissolved, particulate, TPs) (0-155 days) [%] a Mean ± σ (min – max), bsimulated plot runoff or measured catchment discharge, c >0.7 µm, nc: non-calculated 31


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