Introduction

Intense and massive agricultural applications led to the accumulation of high concentrations of pesticides in the soil that may ultimately become hazardous to the various forms of life. In the soil, the accumulation and dissipation of pesticides depend on the nature of pesticides and the soil characteristics (Liu et al. 2007; Morillo et al. 2000).

Chlorpyrifos (O,O-diethyl O-(3,5,6-trichloro-2-pyridyl)phosphorothioate) is one of the most frequently used organophosphorus insecticides that is effectively used against various pests and insects of important cash crops. It is also frequently used to control the termites, flies, mosquitoes, and various household and veterinary pests (Zhang et al., 2011). In soils, its t0.5 usually ranges between 60 and 120 days, but different soil conditions such as pH, soil type, and temperature may extend it over 1 year. The frequent use of chlorpyrifos has been reported to contaminate the soils and water as well as destroy the non-target organisms (EC 2005; Mugni et al. 2012). Moreover, its abundant use in agriculture has been linked to various health problems that may emanate from dietary exposure to the residues of chlorpyrifos present in the food (Mugni et al. 2012). Its high acute toxicity may affect the cardiovascular system, respiratory system, and the central nervous system (Dam et al. 2000). The environmental fate of chlorpyrifos is governed by both the biotic and abiotic processes, such as chemical hydrolysis, microbial degradation, and photolysis (Sreekumaran and Pradeep 1999; Zalat et al. 2014).

A study evidenced that if chlorpyrifos is applied continuously to the soil for 360 days, the soil contained 22 % 3,5,6-trichloropyridinol (TCP), less than 8 % of 3,5,6-trichloro-2-methoxypyridine, and 27–88 % of CO2 [Walia et al. 1988].

Chlorpyrifos degrades in the soil by photoinduced reactions. Three different photochemical reactions are reported to take place under UV irradiation involving dechlorination, hydrolysis, and oxidation (Walia et al. 1988). During photoirradiation of soil, different products are formed by dehalogenation and oxidation of chlorpyrifos that are further photolysed to yield O,O-diethyl phosphorothioic acid and chloropyridinols. The rate of hydrolysis of oxon is greater than that of chlorpyrifos so it does not accumulate in the soil. UV irradiation is also known to cause a decrease in the levels of chlorinated pyridinols thereby confirming their mineralization [Jabeen et al. 2014].

Sewerage sludge and organic wastes are frequently used as a source of organic material in agricultural soils. However, some studies have evidenced that the addition of nitrogenous and phosphate fertilizers and organic manure can alter the rate of degradation of pesticides in soils (Caracciolo et al 2005; Han et al. 2003; Topp et al. 1996; Zhou et al. 2008). The strong binding of pesticides with clay minerals and soil organic matter may cause them to persist in the soil as well as to disperse slowly. Thus, the potential of the pesticides to cause long-term effects on beneficial soil microorganisms as well as aquatic species is increased (Van Zwieten et al. 2003; Liu et al. 2007; Gaw et al. 2003). Moreover, these materials contain large quantities of hazardous heavy metals that may accumulate in the soil upon successive application of sewage sludge and thus cause the problems arising from high contents of heavy metals.

Heavy metals play dual roles in the degradation of pesticides in soil. Enhanced levels of some pesticides in the soil such as Cu2+ and Zn2+ may reduce the dissipation of deltamethrin and cypermethrin while some other metals like Fe may enhance their rate of dissipation and thus reduce their t0.5. This was explained in terms of reduction in the activity of bacterial biomass (Liu et al. 2007; Zheng and Ye 2001; Rafique and Tariq 2014). The persistence/dissipation of pesticides in the presence of metals has also been explained on the basis of their influence on the growth of bacterial populations involved in pesticide degradation (Vágvölgyi et al. 2010; Škrbić et al. 2010). Cu2+ ions have the potential to retard the dissipation of ethylenethiourea produced on degradation of ethylene bis-dithiocarbamate fungicides. On the other hand 2,4D degradation was highly accelerated by Variovorax in the presence of Cu2+ ions. Manganese, zinc, and copper (20–50 mg L−1) were also found to accelerate the carbendazim and diuron degradation (Vágvölgyi et al. 2010; Škrbić et al. 2010). Guzsvany et al. studied the potential of AlFe-pillared clay and Fe-ZSM-5 catalysts for the removal of aqueous imidacloprid under UV irradiation or visible light. The comparison of the obtained results was also provided with the data obtained by using the photo-Fenton reaction and TiO2 Degussa P25 (Guzsvany et al. 2010)

Thus, the present study was formulated to determine the influence of Cu2+ and Fe2+ ions on the persistence/dissipation of soil-incorporated chlorpyrifos. The importance of the study stems from the fact that enhanced levels of certain metals present in the soil may trigger the photodegradation process of pesticides. Thus, the problems arising from repeated and excessive applications of pesticides to various crops and their subsequent accumulation in the soils may be overcome to some extent.

Materials and methods

Soil collection, pretreatment, and characterization

For the present study, bulk soil was obtained from the botanical garden of National Agricultural Research Center (NARC) Islamabad at a depth of 0–15 cm and placed in zip-mouthed, high-density, polythene bags. No pesticides or fertilizers had previously been used on the sampling site for 10 years. Prior to use, the soil samples were homogenized periodically and sieved to obtain particles of less than 2 mm mesh sizes. A water holding capacity (WHC) of 75 % was maintained for the soil samples at 0.33 bars by using the method of Frank et al. (2002)). Then, these soil samples were stored in the dark, at 20 °C until analysis. The texture of soil was determined by a hydrometer (Bao 2000). Soil physico-chemical properties (Table 1) were determined by using the standard method of Ryan et al. (2001). For bioactive control experiments, soil samples were autoclaved in capped flasks for 2 h at 121 °C for sterilization (100 mL capacity). An atomic absorption spectrophotometer was used for the estimation of metals in soil samples before use.

Table 1 Physico-chemical properties of the studied soil
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Materials

Chlorpyrifos (99 % purity) standard and HPLC-grade acetonitrile were obtained from Sigma Aldrich (Germany). The solutions of chlorpyrifos at a concentration of 5 mg L−1 were prepared in acetonitrile and stored in a freezer at 4 °C. Analytical grade FeSO4·7H2O and CuSO4·5H2O were purchased from Merck, Germany. Analytical grade reagents and chemicals were used without any further purification. Highly pure distilled water as prepared by a Milli-Q system of Millipore Waters Co. (Bedford, MA) was used during these experiments.

Spiking procedure

The spiking solution of chlorpyrifos at a concentration of 0.5 mg/mL was prepared by diluting the stock solution appropriately with acetonitrile. A 10.0-g portion of pre-dried soil was mixed thoroughly with water (7.5 mL) to prepare the soil slurries. This slurry was uniformly spread on the petri plate in the form of thin films of 2 mm depth and then added with pre-requisite concentrations of pesticide standards. Subsequently, these soil slurries were added with appropriate concentrations of two metals, i.e., Cu2+ and Fe2+, separately. The final concentrations of these metals spiked with the soil are provided in Table 1. Microsyringe was used to evenly dispense the spiking solution across the soil surface. The soil samples were then thoroughly mixed and uniformly spread on the plate. In the case of control experiments, no addition of metal and pesticides was made. All the samples were prepared in triplicate. After irradiation, the controls and samples were removed from the photoreactor.

Metal speciation studies in soil

Metal speciation studies were performed by using the modified method of Tessier et al. (1979) that distributed the metals into six operationally defined fractions. According to the method, a 1.0-g portion of air-dried, homogenized soil sample was sequentially treated with deionized distilled water, MgCl2 (1.0 M), sodium acetate solution (1.0 M), hydroxylamine hydrochloride (0.04 M), nitric acid (0.02 M) + 30 % H2O2, and HNO3 + HClO4, to obtain the respective fractions such as water-soluble (S1), exchangeable (S2), carbonate-bound (S3), Fe–Mn-oxide-bound (S4), organic-bound (S5), and residual fractions (S6). (Salbu et al. 1998) For each extraction, the mixture was centrifuged for 30 min at 3000 rpm; the supernatants were removed with pipette, filtered with Whatman filter paper, and analyzed for metals by using atomic absorption spectroscopy (AAS). Before starting the next extraction step, the sample was shaken for 30 min with 8 cm3 of double distilled water (DDW) water and centrifuged, and the wash solution was discarded. All extractions were conducted in triplicate.

The calibration curve method was adopted for the quantification of results while using the standard solutions in concentration range of 2–8 mg L−1 and recording their corresponding absorptions. At least four standard solutions were run on the instrument covering the absorption range of samples. The precision of quantitative results was ensured by running the triplicate samples. In order to check the accuracy of data obtained by AAS, standard reference material SRM 2711“Montana Soil” was run with the samples. The results of the analysis were considered to be reliable if analysis error for repeat samples was less than 5 %, and an analytical precision of ±10 % was obtained for replicate samples.

Photochemical experiments

The soils were irradiated with a UV tube of 8 W (Atlas, Linsengericht, Germany) equipped in a self-designed photoreactor. The vessels used for irradiating the samples were provided with quartz lids at the top. In order to control the temperature, water was kept circulating beneath the samples, throughout the floor of the photolysis chamber. A 3.0-V electric fan was installed inside the photoreactor to allow the sample headspace to be purged continuously. The vertical distance between spiked soil samples and UV tube was maintained at 23 cm. The reference comprised an unspiked soil sample that was irradiated in a UV photoreactor for a time interval equal to the sample. In order to maintain the WHC at 75 %, the moisture content of soil samples was regularly monitored, initially after every 60 min and later on after every 12 h. DDW was used to maintain the initial moisture content and weight of each soil sample after every sampling, where necessary. Subsequently, these soil samples were irradiated in the UV photoreactor for different time intervals, i.e., 0, 4, 24, 48, 96, 192, 384, and 762 h.

Control test experiments

The control test samples of two different types were prepared including unsterilized and sterilized ones. The unsterile dark control samples were prepared by adding a 10-g portion of soil sample with prerequisite concentration of chlorpyrifos and two metals. These samples were incubated in the dark at 25 °C for 0, 1, 2, 4, 8, 16, 32, 64, and 128 days. Before spiking, a WHC of 75 % was maintained for each of the soil samples. For sterile control experiments, 10 g soil samples were autoclaved thrice at 121 °C (at intervals of 24 h) for 30 min in air-tight flasks. A water content of 75 % was achieved by adding deionized water to the autoclaved soil samples. These samples were then added with solutions of chlorpyrifos and metals and incubated for 0, 1, 2, 4, 8, 16, 32 64, and 128 days respectively in the dark, at 25 °C.

Pesticide extraction and analysis

The chlorpyrifos-spiked soil samples were extracted after irradiation according to the procedure of Graebing and Chib (2004). In brief, UV-irradiated samples were extracted three times with (90:10) mixture of acetonitrile and water (10 mL) containing 1 N H3PO3. During extraction, the contents were vortexed thoroughly for 1 min and sonicated for 10 min in an ultrasonic bath (35 kHz, 50/60 Hz, D-7700 Singen/HtW T-700, Elma, Germany). Subsequently, these samples were centrifuged for 10 min. The extraction of residue was carried out twice by using the same mixture. The combined soil extracts were concentrated to 5 mL on a rotary evaporator at 40 °C, cleaned thrice with 10 mL CH2Cl2, and then evaporated again to 0.5 mL. These extracts were then dissolved in a 2-mL volume of acetonitrile, cleaned by passing through a 0.45-μm polyethersulfone-based filter membrane present in a syringe, and reduced to 1.0 mL at ambient temperatures, under N2 atmosphere. The extracts obtained were analyzed by HPLC. Chlorpyrifos recoveries obtained by this method were in the range of 79.9 % ± 2.05 for 0.5 mg kg−1 to 94 % ± 1.19 for10 mg kg−1.

The method of Zalat et al. (2014)) was used for the determination of chlorpyrifos. The analyses by HPLC were carried out by using a reversed-phase C18 column with an internal diameter of 250 × 4.6 mm and a 5-μm particle size. The instrument was equipped with a binary LC Pump 250 PE Nelson 900 Series Interface UV/Vis Detector for separation. The mobile phase consisted of a mixture of glacial acetic/water/acetonitrile (0.1 v/10 v/90 v) that was used at a 1 mL/min flow rate and an injection volume of 20.0 μL. A wavelength of 290 nm was used for the detection of chlorpyrifos at the retention time of 6.7 min. The adopted HPLC method was observed to be linear in the concentration range of 0.05 to 50 μg mL−1 (with R 2 = 0.999 ± 0.04). The LOD and LOQ obtained for this method were 0.05 and 0.15 μg mL−1, respectively. Calibration of the instrument was performed by using external matrix-matched standards, every time before the analysis of samples and the linear regression analyses were used for quantification.

Photocatalytic degradation kinetics

The data for the photodegradation of pesticides in the soil was studied by fitting the Langmuir–Hinshelwood (L–H) model according to the equation:

$$ \mathrm{r}=\mathrm{d}\mathrm{C}/\mathrm{d}\mathrm{t}=\mathrm{k}\mathrm{K}\mathrm{C}/2+\mathrm{K}\mathrm{C} $$

where C represents the pesticide concentration, k = rate constant, r = rate of pesticide mineralization, and K = adsorption coefficient of the pesticide.

If the initial concentration of pesticide is in the range of parts per billion or parts per million, then L–H kinetics may be approximated to first-order kinetics according to the equation:

$$ \mathrm{r}=\mathrm{d}\mathrm{C}/\mathrm{d}\mathrm{t}=\mathrm{k}'\mathrm{C}=\mathrm{k}\mathrm{K}\mathrm{C} $$

The integrated forms of equation are:

$$ \mathrm{C}\mathrm{t}={\mathrm{C}}_{\mathrm{o}}{\mathrm{e}}^{-\mathrm{k}'\mathrm{t}}\mathrm{orln}\left({\mathrm{C}}_{\mathrm{o}}/{\mathrm{C}}_{\mathrm{t}}\right)=-\mathrm{k}\mathrm{K}\mathrm{t}=-\mathrm{k}'\mathrm{t} $$

where k = first-order rate constant (time1) [Konstantinou and Albanis 2003). The t0.5 of pesticide is determined by the equation = (1/k) ln 2.

Results and discussion

Chlorpyrifos is degraded in the soil on exposure to chemical reactions, sunlight, and microorganisms (Reddy et al. 2013; Luebke and Hum 2002). Its half-life has been recorded to range from 48–190 days in the dark, but it may exceed 1 year depending on its formulation, rate of application, soil type, climate, etc. pH of soil was found to have a profound influence on the degradation of pesticide. On increasing the pH from 7 to 8, an increase in rate of degradation was observed with a corresponding decrease in t0.5 from 100 to 1.5 days.

The microorganism-assisted degradation of chlorpyrifos is a good technique for its dissipation, but incomplete and slow degradation are the major constraints that limit this process [Walia et al. 1988; Roberts and Hutson 1999]. In water, the photodegradation t0.5 of chlorpyrifos was found to be 3–4 weeks. So far, no studies have been reported focusing the metal-assisted photodegradation of chlorpyrifos in the soil.

Photodegradation studies of soil-incorporated chlorpyrifos

On exposure to sunlight, the degradation rate of chlorpyrifos in the soil was observed to be 3.5 × 10−2 day−1 with t0.5 of 19.8 ± 2.3 days (r = 0.988). The natural logarithmic decline corresponding to the photodegradation of soil-incorporated chlorpyrifos on UV irradiation verses irradiation time is provided in Fig. 1. The photodegradation curves for dark unsterile and sterile controls are also cast in this figure for comparison. In moist soil, chlorpyrifos photodegradation followed the first-order kinetics. It was found that the photodegradation of chlorpyrifos in soil was quite rapid under UV irradiation with half-life being reduced to 19.8 days from an initial half life of 40.8 days (R 2 of 0.988, Table 2). The reaction rate was observed to be varied from 1.7 × 10−2 to 3.5 × 102 per day, at a p value of <0.05.

Fig. 1
figure 1

Photodegradation profile of chlorpyrifos in soil

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Table 2 Dissipation statistics for chlorpyrifos in soil amended with trace metals
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The microbial degradation and abiotic hydrolysis are responsible for transforming the chlorpyrifos. The significance of biotic degradation of soil-incorporated chlorpyrifos was indicated by a large difference in degradation rates of pesticide in unsterilized and sterilized soils, i.e., 6 × 10−2 day−1 and t0.5 of 25 days (Fig. 1 and Table 2), thus evidencing the significant role of microbes in dissipating and detoxifying the residues of chlorpyrifos in the soil (Jabeen et al. 2014; Rokade and Mali 2013). Chlorpyrifos also inhibits the growth of bacterial populations present in the soil and interrupts the degradation of soil organic matter. As is mostly applied aerially, it accumulates in the top soil where it affects the soil biochemical properties. It also retards the growth of nitrogen-fixing symbiotic bacteria and adversely affects the enzyme activity of soil that is the degradation index for organic matter in the soil [Laksmikantha 2000].

The data furnished in Table 2 shows that in comparison with photodegradation, the t0.5 of chlorpyrifos in unsterilized treatments was enhanced from 19.84 to 30.13 days. These results agreed well with those of Singh et al. (2002)) who reported the t0.5 of chlorpyrifos to be in the range of 34–46 days (Singh et al. 2002). TCP has been reported as the antimicrobial degradation product of chlorpyrifos that is accumulative in the soil and slows the rate of microbial degradation (Robertson et al. 1998).

Under sterilized conditions, the soil microbial population was eliminated and thus the pesticide became persistent. The degradation of chlorpyrifos was very slow in the dark sterile soil conditions with a rate constant being 1.7 × 10−2 and a t0.5 double than that under UV irradiation (Fig. 1). Hydrolysis was the main process that caused the dissipation of chlorpyrifos in sterile soil kept in the dark because the possibilities of microbial dissipation and photodegradation were eliminated by incubating the samples in the dark. Racke et al. (1996) also reported a small degradation of chlorpyrifos in soils with high pH, under sterile conditions (Racke et al. 1996).

Effect of initial pesticide concentration on photodegradation

An increase in initial concentration of chlorpyrifos in soil was observed to decrease its rate of photodegradation. The data in Table 3 evidenced that on increasing the initial concentration of soil-incorporated chlorpyrifos from 50 to 100 mg/kg, a decrease in its rate of degradation to 2.4 × 10−2 took place. Simultaneously, the t0.5 of pesticide was enhanced from the initial half-life of 19.74 to 28.40 days at p < 0.05. Further confirmation was obtained by reducing the chlorpyrifos concentration to 25 mg kg−1, where t0.5 was significantly reduced to 12.81 days. It was also depicted by an increase in degradation rate constants from 3.51 × 10−2 to 5.41 × 10−2. Thus, the degradation rate of chlorpyrifos was dependent on its initial concentration. In fact, UV sources produce photons. On increasing the initial concentration of chlorpyrifos, the competition between TCP and chlorpyrifos for UV photons was increased which led to a reduction in dissipation of chlorpyrifos.

Table 3 Dissipation statistics for different concentrations of chlorpyrifos
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The rate of degradation of chlorpyrifos in the dark was also found to be dependent on the initial pesticide concentration in the soil (Table 3). A four-time increase in chlorpyrifos concentration (i.e., from 25 to 100 mg kg−1) led to a twofold increase in t0.5. Singh et al. (2003)) have reported that even after repeated applications of chlorpyrifos, no chlorpyrifos-degrading microbial population was developed in the soil (Singh et al. 2003).

Metal speciation studies in soil

The form of the metal in soil decides its behavior in the soil. Thus, the metal speciation analysis in the soil was performed after spiking the soil with metals according to the modified Tessier’s scheme. The results of the study pertaining to the distribution of the metals in various operationally defined fractions in surface soil evidenced that Cu and Fe mostly belonged to Fe–Mn-oxide-bound fraction and organic fraction. The appreciable concentration of Cu was also present in residual fraction. The least amount of the two metals was present in water-soluble and exchangeable fractions, respectively.

Effect of Cu2+ on degradation of soil-incorporated chlorpyrifos

When the soil-incorporated chlorpyrifos was spiked with Cu2+ at a concentration of 20 mg kg−1, the rate of its photodegradation was increased from 3.5 × 10−2 to 1.6 × 10−1 day−1 (p < 0.05, Fig. 2). A sudden reduction in chlorpyrifos t0.5 was also observed from 19.84 to 4.4 days (Table 2). In fact, Cu compounds are known to catalyze the photodegradation of various pesticides under UV irradiation (Tariq et al. 2014). An increase in Cu2+ concentration from C0 + 20 mg kg−1 to C0 + 40 mg kg−1 caused an increase in the rate of photodegradation from 1.58 × 10−1 to 1.65 × 10−1 day−1. Thus, a clear reduction in the dissipation half-life of the pesticide was recorded from 105 to 98 h (Table 4). Figure 2 depicts that during 32 days of incubation, the rate of chlorpyrifos dissipation was increased significantly in unsterilized soil at p > 0.05 in comparison with the control dark sterile treatments.

Fig. 2
figure 2

Degradation profile of chlorpyrifos in Cu2+-amended soil

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Table 4 Dissipation statistics for chlorpyrifos in soil amended with different Cu2+ and Fe2+ levels under UV irradiation
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In unsterile control treatments, chlorpyrifos t0.5 was reduced from 22.8 to 13.1 days that evidenced that Cu2+ influenced the activity of the microbes in the soil that degrade the pesticide. Copper also influenced the abiotic dissipation of chlorpyrifos. Thus, the t0.5 of chlorpyrifos dissipation was reduced from 40.8 to 18.73 days when it was spiked with catalytic amounts of Cu2+ but on UV exposure, a further decrease in half-life was observed from 18.73 to 4.4 days. This was explained on the basis of the photon capturing potential of Cu2+ ions and their subsequent efficient role in catalyzing the photolysis of chlorpyrifos.

Effects of Fe2+ on chlorpyrifos degradation

Fe2+ is known to boost the photodissipation of numerous halogenated herbicides and pesticides due to its photosensitizing ability (Rafique and Tariq 2014; Salah et al. 2006; Tajeddine et al. 2010). When the soil-incorporated chlorpyrifos was spiked with Fe 2+, an increase in the rate of photodegradation of chlorpyrifos was observed from 3.5 × 10−2 to 5.5 × 10−2 (p < 0.05) and its t0.5 was decreased from 19.74 to 12.6 days (Table 2). The role of Fe2+ in catalyzing the photodegradation of pesticides has also been reported by other authors (Guzsvany et al. 2010; Kochany 1992; Tariq et al. 2014; Zheng and Ye 2001).

The increased soil Fe2+ levels have been reported to enhance the degradation of chlorpyrifos in soil (Satapanajaru et al. 2003; Sayles et al. 1997). The data provided in Table 4 and Fig. 3 evidenced that in comparison with controls, a 94 to 96 % of initial chlorpyrifos concentration was dissipated when it was continuously irradiated with UV light for 4 days in the presence of increased Fe2+ concentration (C0 + 10 and C0 + 30 mg kg−1). This increased degradation was due to the direct photodegradation catalyzed by Fe2+ or by an indirect photolysis resulting from the reaction of iron ions with hydroxyl radicals present in the moist soil (Sayles et al. 1997).

Fig. 3
figure 3

Degradation profile of chlorpyrifos in Fe2+-amended soil

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Under dark unsterile conditions, the rate of chlorpyrifos dissipation was only negligibly affected (0.02 at p < 0.05) due to increased Fe2+ levels that eliminated the possibility of microbial dissipation (Fig. 3). The t0.5 of chlorpyrifos in Fe2+-amended soil was decreased from 43.36 to 26.6 days (p < 0.05) indicating that Fe2+ also affected the abiotic degradation of chlorpyrifos (Table 2). Under anoxic conditions, Fe0 has been successfully used to remediate the chlorpyrifos-contaminated soils at acidic pH. Thus, hydrolysis followed by reductive dechlorination was the major process for bringing about abiotic degradation of chlorpyrifos (Sayles et al. 1997).

Conclusions

The presence of selected trace metals, i.e., Cu2+ and Fe2+, caused an increase in the rate of chlorpyrifos photodegradation with a resultant decrease in t0.5 from 19.8 to 4.39 and 19.25 days, respectively. Cu2+ caused a six-time increase in photodegradation rate and a twofold increase in microbial degradation of chlorpyrifos. Fe2+, on the other hand, negligibly affected the biotic chlorpyrifos degradation. The selected trace metals also influenced the abiotic degradation in the order Cu2+ > Fe2+.