Long-term relationships among pesticide applications, mobility, and soil erosion in a vineyard watershed

Lake Sediment.

The lake-bottom sediment cores were characterized in terms of their color, grain size, loss on ignition (LOI), and sedimentary structure. The upper 41 cm consists of olive-gray silty clay with constant fractions of carbonate (30%) and organic content (7.5%) (Fig. 2B). The grain size distribution of this upper sequence is homogeneous [median (D50) = 12.7 ± 1.8 µm] and exhibits two main populations centered at 0.3 µm (carbonate fraction) and 14 µm (Fig. 2A).

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Fig. 2.

Data from core SANP2. From left to right: (A) photograph and grain size contour plot with two primary populations centered at 0.3 and 14 µm, (B) LOI at 550 °C (organic matter) and 950 °C (carbonates), (C) ²¹⁰Pb_ex activity, (D) ¹³⁷Cs activity, and (E) the age model.

The levels of major and trace elements were measured using an X-ray fluorescence (XRF) core scanner (Avaatech) and were subjected to principal component analysis (PCA; Fig. S1) to constrain sediment end-members (21). This PCA of the bulk sediment resulted in the identification of four geochemical end-members: (i) Al, Si, K, Fe, Ti, Rb, Ba, and Zr, which are related to terrigenous input from the watershed (aluminosilicates and heavy minerals present in marls); (ii) Ca and Sr, which are linked to the carbonate productivity in the lake; (iii) S and Mn, which are related to the lake’s oxidation state; and (iv) a Cu source that may be correlated with periods of significant vineyard-related activities in the watershed, during which a blend of copper sulfate and calcium hydroxide (Bordeaux mixture) was sprayed as a fungicide.

A chronological framework was established via measurements of short-lived radionuclides (22, 23). A logarithmic plot of (²¹⁰Pb_ex) activity (Fig. 2C) shows a general decrease with three distinct linear trends. According to the “constant flux, constant sedimentation rate” (CFCS) model (22, 24), as applied to each part of the profile, the levels of ²¹⁰Pb indicate mean accumulation rates of 2.9 ± 0.2 mm⋅y⁻¹ between depths of 41 and 26.5 cm, 5.2 ± 0.6 mm⋅y⁻¹ between 26.5 and 17 cm, and 8.7 ± 1.3 mm⋅y⁻¹ in the upper 17 cm of the core (Fig. 2C). The plot of ¹³⁷Cs data (Fig. 2D) displays a peak at a depth of 29.5 ± 1 cm, which apparently correlates with the maximum atmospheric production of ¹³⁷Cs in 1963 (25). This temporal correlation is supported by the ²⁴¹Am peak at the same depth, which was a result of the decay of ²⁴¹Pu in fallout from atmospheric nuclear weapons tests (26). In the upper part of this core, at a depth of 20.5 ± 0.5 cm, a second ¹³⁷Cs peak corresponds to the time of the Chernobyl accident in 1986 (Fig. 2D). The good agreement between the ages derived from the ²¹⁰Pb_ex-CFCS model, and the artificial radionuclide peaks provide a well-constrained, continuous age–depth relationship (Fig. 2E) within the sediment sequence, with two primary sedimentation rate changes in ∼1973 ± 5 y and 1994 ± 2.5 y.

Sediment Chronology of Pesticides Use.

No significant variations in the grain size distribution or the organic content were observed during the last century. Thus, these two parameters could not have affected the absorption/degradation of pesticides in this sediment sequence (27). On the basis of the age–depth relationship, the down core variations in pesticides were presented in relation to the age and expressed in flux (ng⋅cm⁻²⋅y⁻¹) (Fig. 3). Historical vineyard practices were obtained through a survey of winemakers and agricultural technicians working in the study area. Three herbicides (or their metabolites) were identified in the Lake Saint André sediment (Fig. 3A): AMPA [a metabolite of glyphosate, a main ingredient in Roundup (Monsanto)], which is used as a nonselective postemergence herbicide; deisopropyl atrazine (a metabolite of triazine herbicides, such as atrazine and simazine); and diflufenican [a main ingredient in Buffalo (Bayer)], which is used as a preemergence herbicide. Monsanto introduced glyphosate into the market in the 1970s, but it has been used widely and increasingly more often only since the early 1990s. High levels of AMPA were found in the core representing deposition during the previous 20 y, with a primary increase since 1990. AMPA is also present in low but significant concentrations before this period, most likely because of contamination of the deeper part of the core by downward smearing of the very high concentrations found in the upper layers. The metabolite of atrazine, which was used at the end of the 1950s and was banned in 2003, was observed in a sample that dates to the period between 1960 and 1970. Diflufenican, which was introduced at the end of 1990s and is still allowed, was identified in the sediments deposited beginning in 2005.

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Fig. 3.

Chronological variations in pesticide fluxes. (A) Herbicides: AMPA, deisopropyl atrazine, and diflufenican; (B) fungicides: Bordeaux mixture (Cu/Rb), captan, dimethomorph, mancozeb (Zn/Rb), and pyrimethanil; (C) insecticides: dicofol, bromopropylate, and bifenthrin; (D) DDT and metabolites: DDT, DDE/DDT, DDD/DDE, and ƩDDT. The gray area in the AMPA profile denotes the lower detection limit for this compound.

Three fungicides were identified in the core and at various concentrations with depth (Fig. 3B). Higher fluxes of captan [e.g., Sigma (Arysta LifeScience)], which was introduced in the early 1950s, and dimethomorph [e.g., Acrobat and Forum (BASF Agro)], which was introduced in the early 1990s, were observed in the samples representing the periods between 1960 and 2005 and after 1990, respectively. Both of these fungicides were used to combat mildew in vineyards, and both are still allowed. Pyrimethanil (e.g., Scala and Toucan), which was introduced during the 1990s and is used to treat the Botrytis fungi (“pourriture grise”) in grapes, was identified in the sediments dating to the period between 2000 and 2008. Metals are also currently used against fungi in vineyards: Cu in the Bordeaux mixture is used as an antimildew and powdery mildew treatment, and Zn in mancozeb [(C₄H₆MnN₂S₄)x(Zn)y] is primarily used against black rot (caused by Guignardia bidwellii fungi). Because Cu and Zn are metals, they may accumulate in vineyard soil, and long-term applications have led to elevated concentrations (1). Thus, we can use certain geochemical ratios (i.e., metals/terrigenous elements) to quantify the amounts of these fungicides. Rubidium was selected as a background metal tracer because it can be attributed to the terrigenous input from the watershed (Fig. S1), it is associated with the thinner sediment particles (clay minerals) with greater specific surface areas for metal adsorption, and it is readily measured using the XRF core scanner. We therefore used the Cu/Rb ratio as a proxy for the Bordeaux mixture and the Zn/Rb ratio as a proxy for mancozeb. The Cu/Rb ratio (Fig. S2) displays a constant increase dating from 1900 and a marked increase at the end of WWII. This trend is consistent with the historic use of the Bordeaux mixture beginning at the end of the 19th century and its widespread use after WWII. The Zn/Rb ratio increased after 1965, with 2 primary peaks in ∼1980 and 2010 separated by low values in ∼1995. Mancozeb was first marketed in the early 1960s.

Four insecticides were identified in the sediment: dicofol, bromopropylate (e.g., Néoron), bifenthrin (Fig. 3C), and DDT (dichlorodiphenyltrichloroethane) and its metabolites [DDD (dichlorodiphenyldichloroethane) and DDE (dichlorodiphenyldichloroethene)] (Fig. 3D). Insecticides used in vineyard treatments against tetranychid and eriophyid mites were present in the post-1940 sediments. The level of dicofol increased in 1970 and exhibited peaks in ∼1970, 1992, and 2000. The level of bromopropylate increased in 1985 and exhibited a primary peak in ∼1998, and the level of bifenthrin increased in 1994. These three insecticides were not detected or were present at very low levels in the youngest part of this core, and their earliest appearances in the sediment are consistent with their introductions into the environment. The ΣDDT profile displays a complex trend, with a first appearance between 1910 and 1920, a small peak in 1970, and a more prominent peak during the 1990s. DDT was first used in agriculture after WWII, and it was banned in 1972. Thus, the ΣDDT record does not correlate well with the known history of this insecticide. The apparent pre-WWII appearance of this compound may be explained by its possible diffusive behavior in the lake sediment: Others (28) have found, by computer simulation, that this compound diffuses at a rate of 10⁻⁸ cm²⋅s⁻¹, which could fully explain this earlier appearance. The 1970 peak may be related to the timing of maximum use of this insecticide (29, 30). The decreasing trend of ΣDDT after this date reflects the banning of DDT. A later peak during the 1990s cannot be attributed to the agricultural application of DDT but, rather, is most likely a result of the reworking of old sources of contamination, which is supported by high concentrations of DDT metabolites (high DDE/DDT ratio; Fig. 3D) in this upper part of the sediment core. The ratio of DDD/DDE is an indicator of the degradation pathways of DDT because DDE and DDD are, respectively, the aerobic and anaerobic degradation products of DDT (31). Thus, the first peak in ΣDDT levels corresponds to a high DDD/DDE ratio, indicating that degradation under anaerobic conditions most likely occurred in situ in the lake sediment after its maximum use around 1970, whereas the low DDD/DDE ratio corresponding to the second peak indicates that degradation of the precursor DDT was prevalent under the aerobic conditions operating on this secondary source of DDT (e.g., in subaerial soil), and these compounds were then remobilized before their deposition on the lake bottom (32).

This chronology of pesticide (herbicides, fungicides, insecticides) use, which was reconstructed from high-resolution lake sediment analysis and validated by the historical introduction and application of these compounds in vineyards, makes it possible to apply a wide range of methods to the reconstruction and evaluation of the long-term effects of agricultural practices on the transport and fate of chemical contaminants in the environment.

Herbicides and Soil Erosion.

We observed three changes in the sedimentation rate (Fig. 2) in this sediment sequence, which display a general increase in the terrigenous flux into the lake, going from ∼0.9 g⋅cm⁻²⋅y⁻¹ (1900–1972) to 1.7 g⋅cm⁻²⋅y⁻¹ (1973–1993), and then to as much as 3.4 g⋅cm⁻²⋅y⁻¹ (1994–2005). A drastic decrease was then observed during the following years (Fig. 4). These variations in the terrigenous sediment supply from the watershed may be directly attributed to soil erosion via vineyard practices. In the early 1970s, the local use of heavy farm machinery (12), which is known to contribute to soil erosion (11) and is associated with the first application of preemergence herbicides (Atrazine metabolite) to combat grass between the rows of vines, may have induced the first increase of terrigenous flux into the lake (Fig. 4). In 1990, we observed synchronous increases in AMPA, ΣDDT (with a low DDD/DDE ratio), and terrigenous soil fluxes into the lake (Fig. 4). In the early 1990s, applications of postemergence herbicides increased widely, including the use of Roundup, as indicated by the high flux of AMPA dating from this period (Fig. 4). It has been demonstrated that application of this chemical has a strong effect on soil erosion, as it acts on grass development and leads to permanently bare soil (12). Moreover, this high flux of sediment supply to the lake is synchronous with the reemergence of banned pesticides, such as DDT and its aerobic metabolites (DDE), which were most likely stored in the vineyard and other agricultural soils in the watershed (33, 34) and subsequently remobilized by the herbicide-triggered rise in soil erosion. Several studies of similar environmental situations (proglacial lakes) also resulted in the identification of a second peak in the levels of DDT and its metabolites in lake sediments in the 1990s, which is more than 20 y after the chemical’s prohibition. This delayed contamination was correlated with old pesticides stored in the ice cap and released by the recent retreat of glaciers (35, 36). In the context of agricultural watersheds, pollutants stored in soil have rarely been taken into consideration (37). In this study, we demonstrate for the first time to our knowledge that the recent widespread use of herbicides (glyphosate) induced an important release and reemergence of contaminants into the environment 20 y after their use was banned. The soils underwent a change in storage conditions, converting from sinks to sources of pesticides. The decrease in pesticide concentrations during the most recent years (Figs. 3 and 4) may be attributed to French and European regulations controlling the use of micropollutants in agriculture.

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Fig. 4.

Chronological variation in levels of AMPA, deisopropyl atrazine, subaerial flux, and sum of DDT and DDE/DDT. The horizontal dotted lines denote the two primary changes in the sedimentation rate.

In summary, our study demonstrates the possibility of reconstructing the use of various pesticides (herbicides, fungicides, insecticides) in an agricultural watershed over the last century, using sedimentary archives. The dates of first use and prohibition of products used to control pests in vineyards and the changes in the soil erosion flux are recorded in the lake sediments. This work demonstrates that this high-resolution analysis of lake sediment allowed us both to reconstruct past agricultural practices in this watershed and to precisely determine the 100-y-long dynamics of chemicals (organic and inorganic) used in vineyards. In particular, this study highlights the effects of postemergence herbicides (glyphosate) on soil erosion and the remobilization of banned remnant pesticides (DDT) stored in vineyard soil. Thus, any study of the environmental fates of pesticides should take into account these potential pesticide–environment interactions.

In light of this study, it appears crucial that ecotoxicological risk assessment and the management of new agricultural practices take into account such mechanisms of pesticide mobility in the environment (38, 39).