Sorption behavior of herbicides
In terms of Kf parameters, our findings are in agreement with those shown by Wauchope et al. (1992) and Rytwo et al. (2005), which addressed the strong affinity of pendimethalin to solid phases due to its hydrophobic nature. Kočárek et al. (2018) previously studied sorption isotherms of pendimethalin under laboratory conditions and obtained a higher Kf value (270.1 cm3/n µg1 − 1/n g− 1) in a silt loam soil. Aclonifen showed significantly lower Kf values than that shown in PPDB (Kf= 138.1 cm3/n µg1 − 1/n g− 1). On the other hand, Trevisan et al. (1999) tested the sorption of aclonifen in 9 different soil types corresponding to our soil characteristics and determined significantly higher Kd values between 8.54-and 602.60 mL g− 1. As mentioned above, the Freundlich isotherm for flurochloridone adsorption is defined as an S-type curve, since n values were < 1 in the three soils and ranged from 0.76 to 0.85, respectively. This indication shows that flurochloridone has a relatively higher affinity for all three soils at a low concentration and that adsorption can decrease at a high concentration (Pinna et al. 2014). Sorption coefficient (Kf) values (Table 3) for s-metolachlor, pethoxamid, and dimethenamid-p show that they have a considerably lower sorption behavior compared to the other three herbicides. This was probably due to their physicochemical properties such as a smaller polar surface area (Å2) and high-water solubility (Table 2). However, the Kd values for s -metolachlor were within the range of previously reported findings for soils with similar properties, e.g. 0.51–3.40 cm3 g− 1 (Si et al. 2009), 0.76–16.67 cm3 g− 1 ( Westra et al. 2015), and 0.6–5.7 cm3 g− 1. (Weber et al. 2003). The Kf values of pethoxamid for three soils are within the range of 1.88–5.61 cm3/n µg1 − 1/n g− 1, while its highest sorption was indicated in Haplic Fluvisol. Pose-Juan et al. (2018) recently investigated the sorption of pethoxamid in amended and unamended sandy loam soil, resulting in significantly low Kf values 0.20 cm3/n µg1 − 1/n g− 1 for the unamended soil. Furthermore, a high level of linearity was observed in the unamended soil by the pethoxamid adsorption curve (nf = 1,68), whereas its sorption by amended soil was near to the linearity of (C-type) (Pose-Juan et al., 2018). Dimethenamid-p showed the lowest Kf values in the studied soils and was relatively smaller than the value in PPDB (2019), (Kf=3.69 cm3/n µg1 − 1/n g− 1). Kočárek et al. (2018) also reported a small sorption value (Kf = 2.12 cm3/n µg1 − 1/n g− 1) for dimethenamid-p in silt loam soil. Westra et al. (2015) examined dimethenamid-p sorption in 25 different soils and the mean Kd values calculated for soils were 2.3 cm3 g− 1.
The effect of soil properties on herbicide sorption.
Table 4 shows the statistical correlation between herbicide’s Kf coefficients and soil properties. Consistent with the herbicide’s properties and pKa values (Table 2), their molecules are in neutral form and behave as a non-ionic organic compound. Recently published authors have stated that hydrogen bonding is a key mechanism for the sorption of non-ionic molecules on soil OM and clay minerals. The effect of OM on the sorption of non-ionic herbicides was evaluated under different soil conditions. For the compounds flurochloridone, s-metolachlor, dimethenamid-p, and pethoxamid, our findings support the predicted interactions described in the literature (Weber et al. 2000; Weber et al. 2004; Pinna et al. 2014; Pose-Juan et al. 2018). The Kf coefficient of herbicides was strongly correlated with OM and CEC (at p < 0. 001), (Table 4). This indicates that they are the major predictors of sorption and increase affinity to soils with a high OM content. Besides, a positive correlation was observed between OM and CEC (r ≥ 0. 989, p < 0. 001), and no relationship was found between clay content and CEC (Table 5).
Table 4
Statistical correlations between the Kf coefficients and soil properties: pHKCl, pHH2O, organic matter content (OM), CaCO3 content, cation exchange capacity (CEC), sand, silt and clay content.
|
pHKCl
|
pHH2O
|
OM
|
CaCO3
|
CEC
|
Sand
|
Silt
|
Clay
|
Pendimethalin
|
-0.147
|
-0.829***
|
0.981***
|
-0.513*
|
0.974***
|
-0.095
|
0.108
|
-0.013
|
Aclonifen
|
-0.286
|
-0.899***
|
0.996***
|
-0.633*
|
0.995***
|
-0.240
|
0.252
|
0.133
|
Flurochloridone
|
-0.210
|
-0.863***
|
0.990***
|
-0.569*
|
0.986***
|
-0.161
|
0.173
|
0.053
|
S-metolachlor
|
-0.332
|
-0.919***
|
0.996***
|
-0.669*
|
0.996***
|
-0.286
|
0.297
|
0.180
|
Pethoxamid
|
-0.320
|
-0.914***
|
0.995***
|
-0.661*
|
0.996***
|
-0.275
|
0.287
|
0.169
|
Dimethenamid-p
|
-0.378
|
-0.933***
|
0.992***
|
-0.705**
|
0.997***
|
-0.333
|
0.344
|
0.228
|
Correlation significant at: * p < 0.05; ** p < 0.01; *** p < 0.001. |
Table 5
The correlation coefficients describing relationship between soil properties: pHKCl, pHH2O, organic matter content (OM %), CaCO3 content (%), cation exchange capacity (CEC mmol+ kg− 1), sand, silt and clay contents (%).
|
pHKCl
|
pHH2O
|
OM
|
CaCO3
|
CEC
|
Sand
|
Silt
|
Clay
|
pHKCl
|
1
|
|
|
|
|
|
|
|
pHH2O
|
0.660*
|
1
|
|
|
|
|
|
|
OM
|
-0.305
|
-0.905***
|
1
|
|
|
|
|
|
CaCO3
|
0.907***
|
0.895**
|
-0.647*
|
1
|
|
|
|
|
CEC
|
-0.346
|
-0.919***
|
0.989***
|
-0.680*
|
1
|
|
|
|
Sand
|
0.976***
|
0.622*
|
-0.262
|
0.903***
|
-0.301
|
1
|
|
|
Silt
|
-0.976***
|
-0.631*
|
0.274
|
-0.908***
|
0.313
|
-1.000***
|
1
|
|
Clay
|
-0.964***
|
-0.535*
|
0.155
|
-0.851**
|
0.196
|
-0.994***
|
0.992***
|
1
|
Correlation significant at: * p < 0.05; ** p < 0.01; *** p < 0.001. |
In general, organic matter has a high sorption capacity, in particular compounds with lower water solubility (Carringer et al. 1975), even though the structure of the pendimethalin and aclonifen molecules have considerably low water solubility (0.33 and 1.4 mg L− 1) and high Kfoc values (13,792 and 7,126 cm3/n µg1 − 1/n g− 1) (Table 2). However the findings suggested that the adsorption of all six herbicides was positively related to OM (r ≥ 0.981, p < 0.001) regardless of their water solubility and Kfoc values. It should be noted that some of the hydrophobic pesticide mobility can be increased in soils with a high OM content due to increased solubility of compounds in the presence of dissolved organic carbon (DOC) as shown by Kodešová et al. (2012) for chlorotoluron and in the case of DDT. Soil minerals and organic components are used to determine the soil sorption potential of pesticides, particularly non-ionic compounds such as atrazine (Spark and Swift 2002; Ben Hur et al. 2003). Ben Hur et al. (2003) the contribution of clay minerals to pesticides can be significant, when the ratio between clay minerals and a fraction of organic carbon is greater than 30. A small ratio was observed between soils, except for Haplic Fluvisol (the ratio clays/OC was 35). However, the effect of clay content on herbicide sorption was not observed for all tested soils, as Kf values and clay content was not significantly correlated (r ≤ 0.228) (Table 4). On the other hand, it is difficult to assess the influence of a single soil property on soil sorption, since they are always correlated. Table 5 shows the relationship between soil properties where soil pHH2O was negatively correlated with OM (r ≥ − 0. 905, p < 0. 001), and CEC (r ≥ − 0. 919, p < 0. 001).
The Kf was negatively correlated to soil pHH2O (r ≥ − 0. 829, p < 0. 001). A negative but not significant correlation was also observed between Kf and pHKCL for all herbicides (Table 4). Weber et al. (1989) indicated that pesticide adsorption by hydrophobic bonding is pH-independent. Ahmad et al. (2014) reported that the sorption behavior of the molecule is influenced by the pKa of the herbicide and the pH of the soil solution. As the sorption of non-ionic molecules is less affected by soil pH (Kan and Tomson 1990). In our study, the pKa values of pendimethalin and aclonifen appeared to be insignificant for soil sorption. Based on the information provided in PPDB (2019), both pendimethalin and aclonifen are acidic herbicides and the dissociation of pendimethalin is somewhat inconsistent (Kočárek et al. 2018), but stable in the pH range 4–9 (Sakaliene et al. 2007). Aclonifen, which has a pKa of -3.15 (PPDB, 2019), was neutral in the current experiment, as the pH of soil solutions ranged from 7.8 to 8.2 (Table 1). This means that those two herbicides cannot be completely dissolved. In this case, sorption processes could not have occurred through anionic-cation pathways, but through hydrogen bonds, Van der Waals force, or hydrophobic partitioning (Trevisan et al. 1999; Shetti et al. 2019). There was a negative correlation between the Kf coefficient and the CaCO3 content (r ≥-0.513, p < 0.05) for all herbicides. Sorption of 2,4-D was studied by Rodriguez-Rubio et al. (2006), and found that soil with a very large amount of CO32− content was necessary for sorption. In our case, however, the opposite effect has occurred, and it is predicted that the adsorption of herbicides would not be favored by CaCO3. Also, Kodešová et al. (2011) suggest that the CaCO3 is not sufficient for the evaluation of herbicide sorption, as the resulting correlation may be based on extremely high or very low CaCO3 values. As stated above, the clay and silt content were not correlated with the sorption coefficient and a slightly negative but not significant correlation was observed with sand content (Table 4). This indicates that they were not the main predictors of the sorption of the studied herbicides, and support the findings of (Weber et al. 2004; Alletto et al. 2013; Westra et al. 2015), while the findings of Peter and Weber (1985) suggest that the clay content has a direct influence on the sorption of alachlor and metolachlor. The correlations between Kf values in all six herbicides are summarized in Table 6. A strong positive correlation was observed between the sorption coefficient of the herbicide. As already discussed, this may be due to the sorption mechanism of these herbicides, which prevails as non-ionizable compounds and are mainly present in their neutral form. In addition, similar behavior of herbicides can be predicted under the same soil conditions.
Table 6
Correlation matrix between the Kf coefficients of different herbicides.
|
Pendimethalin
|
Aclonifen
|
Flurochloridone
|
S-metolachlor
|
Pethoxamid
|
Dimethenamid-p
|
Pendimethalin
|
1
|
|
|
|
|
|
Aclonifen
|
0.998***
|
1
|
|
|
|
|
Flurochloridone
|
0.999***
|
1.000***
|
1
|
|
|
|
S-metolachlor
|
0.985***
|
0.993***
|
0.992***
|
1
|
|
|
Pethoxamid
|
0.996***
|
0.999***
|
0.999***
|
0.997***
|
1
|
|
Dimethenamid-p
|
0.972***
|
0.983***
|
0.981***
|
0.998***
|
0.989***
|
1
|
Correlation significant at: * p < 0.05; ** p < 0.01; *** p < 0.001. |
The effects of soil properties on herbicide degradation
Several authors have proposed that sorbed chemicals are less available for transport and degradation (Selim et al. 1999; Koskinen et al. 2001). It can be assumed that the longest DT50 values were possibly due to the higher OM content in the soil (Table 1). The high DT50 values could also be explained by the sorption coefficient (Kf) of herbicide. The DT50 values may also be increased if microbiological degradation is the key dissipation mechanism (Felsot and Dzantor 1995; Munoz-Leoz et al. 2013).
The degradation rate of two herbicides significantly correlated only with soil pH; a positive correlation for pethoxamid, and a negative but not significant correlation for flurochloridone. For other herbicides, weak positive correlations were observed between soil pH and degradation rate. Aclonifen and pethoxamid showed faster degradation and shorter dissipation half-lives in the studied soils. The average values were 18 days for aclonifen and 20 days for pethoxamid. Shortest values were found in Haplic Chernozem and Arenic Regozem (Table 7). The obtained half-life for aclonifen was much shorter than the value presented (62.3 days) in PPDB (2019), and for pethoxamid, it is significantly higher than those stated in the literature (Pose-Juan et al. 2018; Rodriguez-Cruz et al. 2019). Vischetti et al. (2002) also observed a high aclonifen half-life (40.3 to 49.1 days) in the laboratory under different temperature and soil moisture conditions in sandy clay loam soil. As shown in Table 8, the rate of degradation of aclonifen was positively correlated with clay content (r ≥ 0. 880, p < 0.001), positive but not significantly correlated with soil pH, and a negative but not significant correlation was observed with OM, CaCO3, and CEC for all three soils. Based on data from the Pesticide Properties DataBase (PPDB), aclonifen is stable between pH 5 to 9 at 22–70 ° C and photolysis is not a major route of degradation (PPDB 2019). It is known that microbiological degradation takes more time and the results suggest that degradation of aclonifen was mainly caused by chemical hydrolysis and/or oxidation due to its shortest half-life among the herbicides.
The faster degradation rate and shorter DT50 values obtained for pethoxamid could be explained by its low adsorption capacity (Table 3), which were more available in dissolved organic carbon (DOC). According to Marín-Benito et al. (2012) dissolved organic carbon can adsorb herbicide and may increase the bioavailability for degradation. Pethoxamid dissipation in soil was studied by Dhareesank et al. (2005), and stated that rapid dissipation could be explained by its irreversible adsorption and hydrophobic character. Rodriguez-Cruz et al. (2019) investigated the dissipation of pethoxamid at different initial concentrations, and the results indicated that organic residues would decrease the dissipation rate at the lowest dosage rate (2 mg kg− 1).
The dissipation half-lives for dimethenamid-p in the three soils ranged from 19 to 34 days, which were almost within the range of DT50 (7.7–31.5 days) reported in EFSA (2018). A longer dissipation half-life value of 34 days was found in Haplic Fluvisol, and significantly lower values were observed in Haplic Chernozem and Arenic Regozem (Table 7). The degradation rate was negatively correlated with OM (r ≥ − 0. 807, p < 0.001), and CEC (r ≥ − 0. 803, p < 0.001), positive but not significantly correlated with clay content, as well as soil pH and CaCO3 for all three soils (Table 8). Based on the EFSA (2018) data, dimethenamid-p is characterized as stable in hydrolysis, while photolysis in water and soil takes place relatively quick (3–5 weeks). In this study, an increase of dissipation half-life of herbicides was observed mainly in soils with higher OM content, and this has also been observed for dimethenamid-p in similar conditions, even though dimethenamid-p has higher water solubility and lower Kfoc values (Table 2) in comparison with other herbicides. It should also be noted that our findings are compatible with the previous study, which indicates that dimethenamid-p appears to be bound immediately to soil particles or dissolved in soil solution and then desorbed into the soil solution, where readily degraded by soil microorganisms (Kočárek et al. 2018). The pendimethalin’s half-life reported here was shorter than the value of 182.3 days as presented in PPDB (2019). The calculated half-life values were extremely short, especially in Haplic Chernozem (more than 10 times) and Arenic Regozem (around 8 times) (Table 7). Despite this, the average pendimethalin half-life value was within the same range (24.4–34.4 days) as those calculated by Kočárek et al. (2016) in a Haplic Chernozem soil under greenhouse conditions. These findings indicate that pendimethalin was rapidly degraded in the studied soils; thus, the rate of dissipation was not related to the hydrophobicity and the degree of adsorption, even though the pendimethalin was strongly adsorbed by all three soils (Table 3). There was a positive correlation between degradation rate and clay content (r ≥ 0. 536, p < 0.05) and, a negative correlation with OM (r ≥- 0. 741, p < 0. 01) and CEC (r ≥-0. 737, p < 0. 01) and, a negative but not significant correlation with CaCO3 (Table 8). The dissipation half-lives (from 38 to 51 days) for flurochloridone in three soils were consistent with those previously reported, ranging from 9 to 66 days (EFSA, 2010). A longer dissipation half-life (51 days) was found in Arenic Regozem (Table 7). Available data on flurochloridone soil persistence under laboratory conditions is scarce. Walker. (1987) obtained a similar DT50 value (40 and 90 days) for flurochloridone incubated in the laboratory at 20 and 10°C in sandy loam soil. According to information presented in PPDB (2019), flurochloridone is moderately persistent and its main degradation pathway is by chemical hydrolysis. The degradation rate of flurochloridone positively correlated with clay content (r ≥ 0. 994, p < 0. 001), and negatively with CaCO3 (r ≥ − 0. 783, p < 0. 05), and a positive but not significant correlation was observed with OM and CEC. A slightly negative correlation was found with soil pH for all three soils (Table 8). The clay content seems to be the predominant factor in the determination of the dissipation of flurochloridone due to its strong correlation with the rate of degradation. Indeed, there was not a significant relationship between the OM content and the rate of degradation. In addition, it could also be hypothesized that the degradation of flurochloridone may decrease (longer half-life) in soils with higher pH. The observed half-life values for s-metolachlor ranged from 37 to 57 days, and the highest value was found in Haplic Fluvisol (Table 7). The degradation of s-metolachlor has previously been evaluated in several studies. Our results were comparable to those reported by Rice et al. (2002) for a half-life of 9. 6–81 days and slightly higher than those reported by Wu et al. (2011) (37.9–49.5) days in five different soils. Assuming that s- metolachlor is more persistent in high OM soil than in low OM soil. Therefore, a negative correlation between the degradation rate and OM content (r ≥ − 0. 583, p < 0.05) and CEC (r ≥ − 0. 579, p < 0.05) were observed, while a significant positive correlation was found with clay content (r ≥ 0. 702, p < 0.01) for all soils (Table 8). Consequently, a decrease in soil OM content and an increase in clay content result in a reduction in the half-life of s-metolachlor in soils.
In summary, it seems likely that the degradation of herbicides in soils was affected by several parameters, including OM content, clay content, CEC, and soil pH. However, in most cases, OM content showed a statistically significant relationship between the herbicides adsorption coefficient and the rate of degradation. This suggests that they may be inversely related. Correlation analysis shows that soil pH was negatively correlated with herbicide sorption (Table 4), and according to Villaverde et al. (2008) it is often estimated that degradation is supported by a higher amount of chemical present in the solution. Finally, as predicted, there was a negative association between the adsorption coefficient (Kf) and the rate of degradation (Table 8). The correlation has shown that the Kf sorption coefficient is useful in predicting the degradation of the studied herbicides.