Introduction

1,1,1-trichloroethane (1,1,1-TCA) and trichloroethylene (TCE) are typical chlorinated organic solvents (COSs) that are extensively found in contaminated soil and groundwater, posing environmental problems globally (Gu et al. 2012; Zhang et al. 2015). The COSs are widely used in various industrial applications such as in detergents, pesticides, paint thinners, resins, and glues, but their non-eco-friendly disposal results in considerable contamination of soil and groundwater (McCarty and Perry 2010). The COS contaminants in soil and groundwater resources cause severe threats to human life and environment due to a number of harmful characteristics, e.g., carcinogenicity and cytotoxicity harmful for the central nervous system, the immune system, and the endocrine system (Bellona and Drewes 2007; Clewell et al. 1995; Wigle et al. 2007; Zhao et al. 2015). In recent years, scientists and researchers have focused to develop efficient and state-of-the-art technologies such as enhanced solubilization flushing (Watts and Teel 2005), permeable reactive barriers (Vogan et al. 1999), in situ air stripping (Adams et al. 2011), thermal treatment, and in situ bioremediation (Miao et al. 2015a, b, c) to remediate the chlorinated solvent-contaminated groundwater. Most of the biological technologies are cost-effective for remediation, but their prolonged treatment period makes them inefficient for practical and commercial applications (Oller et al. 2011). In contrast, in situ chemical oxidation (ISCO) is considered as an efficient method for effective degradation of COSs (Akyol and Yolcubal 2013; Wu et al. 1999).

Several studies utilized potassium permanganate, persulphate, ozone, and fenton reagent in chemical oxidation processes as chemical oxidants for remediation of contaminated groundwater and soil (Huang et al. 2001; Liang et al. 2004; Neyens and Baeyens 2003; Sunder and Hempel 1997). Fenton reagent represents an attractive option due to its high reactivity-degrading COSs with low environmental impacts (Burbano et al. 2005; Che and Lee 2011; Li et al. 2015). In Fenton process, hydrogen peroxide is decomposed by Fe2+ to generate very active species, mainly hydroxyl radicals (OH●), as shown in Eqs. 1 and 2 (Che and Lee 2011).

$$ {\mathrm{H}}_2{\mathrm{O}}_2+{\mathrm{Fe}}^{2+}\to {\mathrm{O}\mathrm{H}}^{\bullet }+{\mathrm{O}\mathrm{H}}^{\hbox{--} }+{\mathrm{Fe}}^{3+} $$
(1)
$$ {\mathrm{Fe}}^{3+}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{Fe}}^{2+}+{\mathrm{H}\mathrm{O}}_2+{\mathrm{H}}^{+} $$
(2)

Hydroxyl radicals (OH● = 2.76 V) generated during this process show higher oxidation potential than permanganate (1.68 V), persulfate (2.01 V), and ozone (2.01 V) (Liang et al. 2004) and can react with COSs at higher degradation rates. OH● reactions with organic compounds are diffusion-controlled with second-order rate constant K >109 M−1s−1 (Miao et al. 2015a, b, c). However, in Fenton process, H2O2 is unstable with a short life-time when injected into the groundwater. Moreover, Fenton process is strongly pH-dependent at an optimum value of 3–4 (Watts and Teel 2005). Precipitation of Fe2+ to (Fe(OH)3) near neutral pH causes loss of catalyst and injection-well plugging problems (Klamerth et al. 2010; Watts and Teel 2006). Nano zero valent iron (nZVI) due to its high surface area and reactivity has been used as a reagent in many environmental remedies since ferric ions are converted into ferrous ions at an increased rate as shown in Eq. 3 (Crane and Scott 2012).

$$ 2{\mathrm{Fe}}^{3+}+{\mathrm{Fe}}^0\to 3{\mathrm{Fe}}^{2+} $$
(3)

Fe0 is considered to be an efficient catalyst by generating more Fe2+ and ultimately increasing the production of OH● (Han et al. 2015; Ramirez et al. 2007). However, the main drawback is the significant leaching of cation of iron in the acidic pH range using heterogeneous Fenton-like systems (Liou et al. 2004).

In order to address such issues, iron precursors supported on cation exchange resins (Fukushima et al. 2000; Shin et al. 2008) and minerals (Sun et al. 2012; Xu and Wang 2012) have been synthesized in heterogeneous Fenton-like systems. In this study, a novel catalyst is synthesized using the economically viable and abundantly available natural zeolite having an Al/Si ratio of 1.56 that provided a support to attach nano iron particles. So far, no data is available for the degradation of organic contaminants (1,1,1-TCA and TCE) using natural zeolite-supported iron nano (Z-nZVI) composite (Liou et al. 2005) in the presence of hydrogen peroxide. Natural zeolite has very short specific surface area (0.6018 m2/g) due to the presence of impurities like feldspar and quartz (Miura et al. 2009). The zeolites are cheap and extensively utilized for soil fertility improvement and as adsorbents (Nishimoto et al. 2015) but not utilized as catalyst for degradation of COSs (Devi et al. 2009; Fukuchi et al. 2010, 2013).

A number of studies have been conducted to find alternative oxidants having identical oxidation potential to Fenton reagent such as sodium persulfate (SPS), calcium peroxide (CaO2), and sodium percarbonate salt (SPC, 2Na2CO3•3H2O2) (Anipsitakis and Dionysiou 2004; Bogan et al. 2003; Christina and Sedlak 2008; Liang et al. 2004; Miao et al. 2015a, b, c; Northup and Cassidy 2008). Recently, SPC has attracted much attention as an alternative to liquid H2O2 (Devi et al. 2009) due to its stability in handling, storage, and transportation together with low cost. Solid-powdered SPC ensures the safety in applications in order to reduce the chances of splashes and spilling during transportation as compared to the liquid H2O2. In recent ISCO practices, SPC is being used as an alternative oxidant (de la Calle et al. 2012) due to similar functional benefits as liquid H2O2. Equation 4 shows the reaction of SPC in Fenton-like process.

$$ 2{\mathrm{Na}}_2{\mathrm{CO}}_3\bullet 3{\mathrm{H}}_2{\mathrm{O}}_2\to 2N{a}_2C{O}_3+3{H}_2{O}_2 $$
(4)

As compared to the Fenton reagent, the SPC is more advantageous; e.g., (i) the SPC is considered to be applicable in a wider pH range, (ii) the SPC decomposition products are non-toxic for microbes that make it compatible for further bioremediation, and (iii) the oxidation products of SPC are water, sodium carbonate, and carbon dioxide that are naturally occurring substances (Neamtu et al. 2004). Therefore, in this study we synthesized natural zeolite-supported Z-nZVI composite to investigate its activity as a catalyst to degrade 1,1,1-TCA and TCE with SPC as an oxidant. To the best of our knowledge, natural zeolite-supported Z-nZVI composite-catalyzed SPC system has not been well studied yet. Moreover, the addition of reducing agent (RA) such as hyroxlaminechroide (NH2OH•HCl) seems to be very effective and potentially applicable in the degradation of organic solvents by boosting the reduction of ferric into ferrous ions and further promoting the generation of free radicals (Miao et al. 2015a, b, c; Savant et al. 2009; Wei et al. 2015). To elucidate the 1,1,1-TCA and TCE degradation performance, free radical probe compounds were used to investigate the presence of dominant free radicals in Z-nZVI-catalyzed SPC system with and without the addition of the RA. The spectrophotometer was used to investigate the presence of the iron ions in the Z-nZVI-catalyzed SPC system. A novel Z-nZVI composite was synthesized in this research study, and the availability of the oxidative degradation of 1,1,1-TCA and TCE with SPC and Z-nZVI composite with and without addition of RA was evaluated.

Materials and methods

Materials and chemicals

1,1,1-TCA (99.0 %) and TCE (99.0 %) as organic solvents were obtained from Aladdin (Shanghai, China). Sodium borohydride (99.5 %), nitrobenzene (NB, 99.0 %), carbon tetrachloride (CT, 99.5 %), ethanol (99.8 %), ferrous sulfate heptahydrate (99.0 %), sodium nitrate (99.0 %), sodium chloride (99.5 %), sodium sulfate (99.0 %), sodium bicarbonate (99.5 %), SPC (98 %), and aluminosilicates (99.0 %) were acquired from Acros Organics (Shanghai, China). Hydroxylamine hydrochloride (99.0 %), 5,5-potassium biphthalate (99.0 %), and 1,10-phenanthroline monohydrate (98.0 %) were obtained from Shanghai Jingchun Reagent Ltd., Co. (Shanghai, China). The stock solution of 0.15 mM (1,1,1-TCA or TCE) was prepared separately by dissolving in ultra-pure Milli-Q water. To adjust the solution pH, 0.1 M NaOH and 0.1 M H2SO4 were used. All reagents were of analytical grade that were utilized without further purification. The deionized water prepared by Milli-Q ultra-pure water process (Classic DI, ELGA, 102 Marlow, UK) using distilled water was used in all experiments.

Preparation of the Z-nZVI composite

The Z-nZVI composite was synthesized using wet ion exchange method (Chen et al. 2011; Wang et al. 2010). To reproduce the results, a detailed description of the utilized method is presented. In this method, natural zeolite (0.5 g) and ferrous sulfate heptahydrate (1.0 g) were mixed in ultra-pure water (250 mL) in a three-neck flask. The initial solution pH was maintained at 4 to inhibit the oxidation of Fe2+ using nitric acid (1.0 M), and the mixture was treated with ultrasound for 10 min. The mixture was then stirred briskly at room temperature for another 30 min. A strong RA, i.e., NaBH4 (1.0 M), was added for effective reduction of Fe2+. Black solid particles were separated from solution using vacuum filtration flask with a 47-mm diameter membrane filter, and sulfate ions from black solid particles were removed through deionized water. Black particles were finally washed with ethanol eliminating traces of sulfate ions and water and further vacuum-dried. Nitrogen atmosphere was provided throughout the process of reduction of Fe2+.

Characterization of Z-nZVI composite

Surface characteristics and morphology of natural zeolite and Z-nZVI composite were tested using scanning electron microscopy (SEM, JSM-630LV, Tokyo, Japan) and transmission electron microscopy (TEM, JEM-1400 electron microscope operated at an accelerating voltage at 80 kV), respectively, before and after contaminant degradation. Energy-dispersive X-ray spectra (EDS) was acquired using SEM to estimate the iron content and other elements in the Z-nZVI composite. Total pore volume, specific surface area, and pore size of zeolite and Z-nZVI composite were measured using the Brunauer-Emmett-Teller (BET) method (Quantasorb Jr. Instrument Quantachrome Co., UK). In order to determine the nature and crystal structure of the Z-nZVI composite (before and after usage), X-ray diffractometer (XRD) along with Rigaku D/max diffractometer (2550VB/PC; CuKα (λ = 0.154 nm radiation) were used respectively.

Experimental procedures

Stock solutions of 1,1,1-TCA (0.15 mM) and TCE (0.15 mM) were prepared separately by allowing the pure non-aqueous phase liquid 1,1,1-TCA or TCE to equilibrate with Milli-Q ultra-pure water under gentle stirring for 60 min in the absence of light. A cylindrical glass reactor (inner diameter of 6.0 cm and height of 9.0 cm) was utilized with two openings at the top of the reactor; one is for the dosing purpose and the other for sampling. A temperature bath (DC Ningbo, China) maintained the temperature of the reactor at 20 ± 0.5 °C by circulating the water through the jacket of the reactor. A known quantity of SPC and Z-nZVI was added in the reactor containing 1,1,1-TCA or TCE solution, and the effective homogeneity of the solution was maintained through a magnetic stirrer. The initial and final pH of solution was recorded. All reagents were added into the reactor before SPC and Z-nZVI addition. Samples were taken at a specific interval of time from the reactor and immediately quenched with n-hexane before COS analyses. All experimental tests were performed twice, mean values were reported, and standard deviations were in the range of 0.015–0.041. Probe tests were performed to find the presence of the reactive oxygen species in the Z-nZVI/SPC system during the reaction period. Same procedure was adopted for probe tests as that of 1,1,1-TCA and TCE degradation except the replacement of 1,1,1-TCA and TCE with NB (OH● probe) or carbon CT (O2 ●– probe). Control tests were conducted simultaneously for each experiment without addition of Z-nZVI composite.

Analytical methods

Samples (1.0 mL) of 1,1,1-TCA, TCE, CT, and NB were withdrawn and extracted with n-hexane (1.0 mL) using a vortex stirrer and were analyzed using a gas chromatograph (Agilent GC 7890A, Palo Alto, CA, USA). For 1,1,1-TCA and TCE, oven and injector temperatures were constant at 75 and 240 °C, respectively, and the detector (ECD) was at the temperature of 260 °C. However, in the case of CT, the oven temperature was changed to 100 °C and the rest of the procedure was identical as that of 1,1,1-TCA and TCE. NB measurements were obtained from a flame ionization detector (FID) where the oven, injector, and the detector temperatures were at 175, 250, and 300 °C, respectively. N-hexane is commonly used as a non-polar solvent in chromatography technique for the extraction of COSs from the aqueous solution due to its exceptional advantage of having strong resistance against de-protonation. The average extraction recovery rates (a cleaner GC trace) of N-hexane to extract 1,1,1-TC was in the range of 85–95 %, whereas TCE, CT, and NB have 89.7, 87.5, and 93.1 %, respectively. Spectrophotometer (T-60 UV–VIS Spectrophotometer, Purkinje General Instrument Co., Ltd., Beijing, China) at an absorbance of 512 nm was used to determine Fe2+ and Fe3+ ions (Tamura et al. 1974). pH meter (Mettler Toledo Delta 320, Greifensee, Switzerland) was used to determine pH of the solution.

Results and discussions

Characterization of natural zeolite and Z-nZVI composite

The summarized characterization results based on inorganic elemental contents, specific surface areas (SSAs), specific total pore volume, and average pore diameter of the natural zeolite and Z-nZVI composite are presented in Table 1. To demonstrate the adsorption-desorption, pore size, and pore volume distribution, N2 isotherm of natural zeolite and Z-nZVI composite is shown in Figs. S1 and S2 (Supplementary Material). Morphologies of natural zeolite and Z-nZVI are compared, and the TEM and SEM images before and after 1,1,1-TCA and TCE degradation are shown in Fig. 1. Natural zeolite has a thin layer structure, whereas nZVI particles are distributed on the natural zeolite sheet. It has been previously reported that nZVI particles (spherical shape with an average diameter of 80 nm) tended to aggregate (Wang et al. 2010) and were arranged in net chain due to their intrinsic van der Waals, magnetic interactions, and higher surface energies (He et al. 2007). This aggregation reduced the mobility and reactivity of nZVI particles (Liu et al. 2007). Our results (see Fig. 1a) indicate that the zeolite has a smooth surface as compared to Z-nZVI (Fig. 1b) that has an abrasive surface. Fe0 particles are present in the form of light spots over the entire surface of the natural zeolite (Fig. 1c), and the average size of these particles is in the range of 50–100 nm by measuring the diameter of over 100 particles in the grid of different areas. These particles are quasi spherical in shape and distributed on the entire surface of natural zeolite without any noticeable aggregation. Literature also ensures that the support materials for iron particles limit the aggregation that is consistent with our results (Ahmad et al. 2015). Our observations are consistent with the explanation that exchanged Fe2+ ions act as seeds to grow Fe0 (Qu et al. 2013), and significant shares of Fe2+ ions are located in the sodalite cage and a portion in the hexagonal prism (Tong et al. 1996) on the external surface of the zeolite framework which ensured the production of dispersed nZVI particles on the natural zeolite surface (Wang et al. 2010). BET N2 measurement analysis indicated that natural zeolite has a particle size of 9.97 μm and surface area of 0.60 m2/g, whereas in the case of Z-nZVI, particle size reached 81.0 nm (decreased) and surface area of 72.3 m2/g (increased). It has been reported in literature that increased surface area of the Z-nZVI composite is most likely due to the non-aggregation of the nZVI particles (Kim et al. 2013). After TCE degradation, the average particle size of Z-nZVI was increased from 81.0 to 278.9 nm and the surface area was decreased to 21.5 m2/g. Similar results were also observed in the case of 1,1,1-TCA degradation (see Table 1). The decrease in surface area of the spent Z-nZVI could be due to (i) sintering of the iron particles (Beydoun and Amal 2002), (ii) reactions occurring on nano particle surfaces (Wang et al. 2010), and (iii) the by-products, i.e., Fe (oxy) hydroxide formed during contaminant degradation process (Seol et al. 2013). Existence of iron in the Z-nZVI composite was verified through the EDS analysis (see Fig. S3) and the elemental composition (see Table 1). Total dissolved iron content was present in lower quantity, and Fe2+ ions were predominant in solution initially that made Z-nZVI more advantageous than other Fenton-like catalysts (Fig. S4). The XRD patterns of Z-nZVI before and after 1,1,1-TCA and TCE degradation have been recorded and shown in Fig. S5. The structure of the natural zeolite has noticeable peaks that indicate strong crystalline behavior, whereas the Z-nZVI composite structure shows resemblance with amorphous structure. The XRD results of Z-nZVI after 1,1,1-TCA degradation (Fig. S5b) show peaks 1, 2, 3, and 4 at 2 theta 23.86, 25.75, 26.21, and 27.80 that are likely due to the oxides of iron such as lepidocrocite and maghemite (γ-Fe2O3). The peaks 6 and 7 could be due to SiO2 associated with the zeolite. In addition, peak 8 may be associated with FeO on the Fe0. The literature suggested that iron particles in XRD are composed of disordered aggregates (Lim et al. 2006; Zied et al. 2008). The XRD pattern of Z-nZVI after TCE (Fig. S5c) had similar structure to that after 1,1,1-TCA degradation.

Table 1 Summary of the elemental composition by weight % and characterization of natural zeolite and Z-nZVI
Full size table
Fig. 1
figure 1

Morphologies of natural zeolite and Z-nZVI. a SEM analysis of natural zeolite; b TEM analysis of Z-nZVI; c SEM analysis of Z-nZVI after 1,1,1-TCA degradation; d SEM analysis of Z-nZVI after TCE degradation

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Comparative degradation of 1,1,1-TCA and TCE in Z-nZVI-catalyzed SPC system

All experiments were performed with an initial concentration of 0.15 mM of both 1,1,1-TCA or TCE, 3.1 mM of SPC and 50 mg/L of Z-nZVI in aqueous solution. Control tests were performed without addition of SPC and Z-nZVI to find the mass loss of 1,1,1-TCA and TCE. The results showed a minor loss (<1.5 %) of 1,1,1-TCA and TCE under all experimental conditions. The performance of 1,1,1-TCA and TCE degradation in SPC system catalyzed by Z-nZVI is shown in Fig. 2. About 49.6 % of 1,1,1-TCA and 39.7 % of TCE were degraded in a duration of 3 h. It is important to note that both 1,1,1-TCA and TCE removals followed the pseudo-first-order kinetic model.

Fig. 2
figure 2

1,1,1-TCA and TCE degradation performances in the Z-nZVI/SPC system. [T = 20 ± 0.5 °C]

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It was observed that pH of the solution started to increase from an initial pH (4.0) as the reactions proceeded, and the final pH of the solution after termination of reactions was at 9.87–11.89. The results indicated that Z-nZVI has substantial potential to degrade both types of contaminants, i.e., 49.6 % of 1,1,1-TCA and 39.7 % of TCE, but could not degrade the contaminants completely. It is noteworthy that a larger amount of 1,1,1-TCA was degraded than TCE during the same interval of time under identical experimental conditions, indicating that Z-nZVI is more efficient to degrade 1,1,1-TCA as compared to TCE.

Role of reactive oxygen species by probe compound tests

Activation of different types of oxidants with transition metals to generate free radicals, such as OH● and superoxide anion radical (O2 ●–), is a complex process (Anipsitakis and Dionysiou 2004). In Fenton and Fenton-like systems, many oxidation species like OH● and reduction species like O2 ●– have been observed, but OH● has been proved to be the most dominant and momentous to degrade chlorinated organic contaminants (Kim et al. 2001; Yang et al. 2010). Therefore, a sequential set of experiments was performed in this study to investigate the presence of reactive oxygen species in solution by detecting free radicals using probe chemical compound tests. The data regarding material selection of probe compounds can be found in the Supplementary Material. The initial NB, CT, and SPC concentrations were set at 2.0, 0.05, and 3.1 mM, respectively. Approximately 59.33 % of NB was degraded indicating the presence of OH● in the system (see Fig. 3). Our findings are in line with the results published in the literature (Oturan 2000). Previous studies showed that 95.0 % of organic contaminants can react with OH● immediately at diffusion-control rates (Hoigne and Bader 1976). Moreover, about 43.19 % of CT was degraded in the Z-nZVI/SPC system which revealed the existence of O2 ●−, and the results were consistent with the literatures (Liang and Su 2009; Navarro et al. 2009). Fenton process has been considered not to react with strong oxidizing contaminants like hexachloroethane (Watts et al. 1999; Xu et al. 2015) due to low reactivity with OH●. However, the results obtained from this study demonstrated that the Z-nZVI-catalyzed SPC system has the ability to react with highly oxidizing and reducing contaminants for groundwater remediation.

Fig. 3
figure 3

Degradation of probe compounds in the Z-nZVI/SPC system. [NB]0 = 1.0 mM, [CT]0 = 0.05 mM, [Z-nZVI] = 50 mg/L, [T = 20 ± 0.5 °C]

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Influence of reducing agent on degradation of 1,1,1-TCA and TCE

From the experiments, we observed that 1,1,1-TCA and TCE degradation was not very appreciable in the presence of Z-nZVI and SPC. An RA was employed in this study to examine its effect on transformation of iron ions and increase in 1,1,1-TCA and TCE degradation. Hydroxylamine (NH2OH•HCl) was selected as an RA due to its strong capability to reduce Fe3+ to Fe2+. Degradation performances of 1,1,1-TCA and TCE with RA addition showed that 1,1,1-TCA was degraded more than 83.0 % and TCE more than 99.0 % (see Fig. 4). However, the degradation rates of both contaminants were lower in the absence of the RA, i.e., 48.0 % (1,1,1-TCA) and 39.0 % (TCE). The calculated rate constants (Kobs) in the presence of RA of 1,1,1-TCA and TCE (based on the pseudo-first-order kinetic model) were 0.0096 and 0.0249 min−1, respectively, which were 2.5- and 8.9-fold greater than the Kobs values obtained in the absence of RA. Similarly, the half-lives of 1,1,1-TCA and TCE degradation were 72 and 27 min−1, respectively, showing a significant improvement with addition of RA as compared to the half-lives of 247 and 182 min−1, respectively, in the absence of RA. The degradation of TCE followed the predominantly oxidative pathway with OH● transforming TCE into formic, dichloroacetic, oxalic, and glyoxylic acids and finally to carbon dioxide (Zhang et al. 2015). It has also been reported that addition of RA significantly enhanced the generation of OH●, leading toward the degradation of organic substrates (Fukuchi et al. 2013). On the contrary, O2 ●– contributes in the degradation of 1,1,1-TCA and TCE (Gu et al. 2012; Zhang et al. 2015). The significant increase in 1,1,1-TCA and TCE degradation with the addition of RA was therefore attributed to the generation of highly reactive species such as OH● and O2 ●–. About 98.0 % of NB was degraded very efficiently with an addition of RA into the Z-nZVI/SPC system, and reduction of CT reached at a maximum value of 76.0 % (Fig. 5) in comparison with nearly 59.33 % of NB and 43.19 % of CT removals without RA addition. Furthermore, the calculated rate constants for NB and CT were increased from 0.0027 to 0.0052 and 0.0022 to 0.0045 min−1, respectively, with the addition of RA. The results indicated that addition of RA nearly doubled the rate of reaction of NB and CT, thus confirming more OH● and O2 ●– species generated in the system.

Fig. 4
figure 4

Effect of reducing agent (NH2OH•HCl) on the degradation of 1,1,1-TCA and TCE, [T = 20 ± 0.5 °C]

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Fig. 5
figure 5

Degradation of probe compounds in the Z-nZVI/SPC system with NH2OH•HCl addition. [NB]0 = 1.0 mM, [CT]0 = 0.05 mM, [Z-nZVI] = 50 mg/L, [T = 20 ± 0.5 °C]

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The significant improvements in the generation of OH● and O2 ●– with the addition of RA in the Z-nZVI/SPC system was accredited due to the high concentration of Fe2+ and fast transformation of Fe3+ to Fe2+. Therefore, the changes of Fe2+ and Fe3+ ions in the Z-nZVI/SPC system were determined (Fig. S4). A considerable increase in ferrous ions was observed with the addition of RA due to enhanced ferric reduction to ferrous and further promoting the OH● and O2 ●– radicals in the Z-nZVI/SPC system. Similar results were reported by Du when Fe2+ was present in large amount in the homogeneous Fenton-like system (Du et al. 2006). The reactions accompanying catalytic decomposition of RA and H2O2 and reduction of Fe3+ to Fe2+ are shown in Eqs. 5, 6, 7, 8, 9, 10, 11, 12, and 13.

$$ {\mathrm{Fe}}^{3+} + {\mathrm{NH}}_2\mathrm{O}\mathrm{H}\to {\mathrm{NH}}_2{\mathrm{O}}^{\bullet }+{\mathrm{Fe}}^{2+}+{\mathrm{H}}^{+} $$
(5)
$$ 2{\mathrm{N}\mathrm{H}}_2{\mathrm{O}}^{\bullet}\to {\mathrm{N}}_2+2{\mathrm{H}}_2\mathrm{O} $$
(6)
$$ {\mathrm{Fe}}^{3+}+{\mathrm{NH}}_2{\mathrm{O}}^{\bullet}\to \mathrm{N}\mathrm{H}\mathrm{O}+{\mathrm{Fe}}^{2+}+{\mathrm{H}}^{+} $$
(7)
$$ 2\mathrm{N}\mathrm{H}\mathrm{O}\to {\mathrm{N}}_2\mathrm{O}+{\mathrm{H}}_2\mathrm{O} $$
(8)
$$ {\mathrm{Fe}}^{3+}+2{\mathrm{H}}_2\mathrm{O}+{\mathrm{NH}}_2{\mathrm{O}}^{\bullet}\to {{\mathrm{NO}}_3}^{\hbox{--} }+{\mathrm{Fe}}^{2+}+6{\mathrm{H}}^{+} $$
(9)
$$ {{\mathrm{NO}}_3}^{\hbox{--} } + {\mathrm{NH}}_2\mathrm{O}\mathrm{H}+{\mathrm{OH}}^{\hbox{--}}\to {\mathrm{NO}}^{\hbox{--} }{{+\mathrm{NO}}_2}^{\hbox{--} }+2{\mathrm{H}}_2\mathrm{O} $$
(10)
$$ {\mathrm{Fe}}^{3+}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{Fe}}^{2+}+2{\mathrm{O}\mathrm{H}}^{\bullet } $$
(11)
$$ {\mathrm{Fe}}^{3+}+2{\mathrm{O}\mathrm{H}}^{\bullet}\to {\mathrm{Fe}}^{2+}+{{\mathrm{O}}_2}^{\bullet \hbox{--} }+2{\mathrm{H}}^{+} $$
(12)
$$ {\mathrm{Fe}}^{2+}+{\mathrm{H}}_2{\mathrm{O}}_2\to {\mathrm{Fe}}^{3+}+{\mathrm{O}\mathrm{H}}^{\bullet }+{\mathrm{O}\mathrm{H}}^{\hbox{--} } $$
(13)

Addition of RA strongly encouraged the reduction of Fe3+ to Fe2+ and more generation of OH● and O2 ●– radicals in the Z-nZVI/SPC system. Moreover, by-products like NO3 −, NO2 −, NO−, and N2O were also produced with RA addition in the Z-nZVI/SPC system. It could be seen that the addition of RA produces part of N2O as an end product that is highly soluble in water (Chen et al. 2011) and all other by-products are considered to have no toxic effects on the environment. The overall experimental results indicate that the Z-nZVI-catalyzed SPC system with the addition of RA is a promising technique for remediation of the groundwater contaminated with 1,1,1-TCA and TCE. Further expansion of the Z-nZVI-catalyzed SPC system applications in different directions will be emphasized in our future research study.

Conclusion

This study investigated the performance of our novel synthesized composite, i.e., Z-nZVI-catalyzed SPC, in a heterogeneous Fenton-like system for the degradation of 1,1,1-TCA and TCE as target COSs. Several important conclusions can be withdrawn from this study.

  • Reduced aggregation and increased surface area validated that the natural zeolite is an effective stabilizer and dispersant for zero valent iron. Z-nZVI-catalyzed SPC degraded 1,1,1-TCA and TCE to a certain extent; however, the addition of RA (NH2OH•HCl) substantially increased the degradation of both contaminants by promoting the generation of reactive free radicals.

  • The Z-nZVI composite increased the reduction of Fe3+ to Fe2+ on its surface that elevated generation of OH●. The presence of reactive oxygen species demonstrated that OH● and O2 ●– were exclusively responsible for 1,1,1-TCA and TCE degradation.

  • EDS and BET analysis of the spent Z-nZVI indicated that it contained large amounts of iron that could persistently be utilized for contaminant degradation.

  • The Z-nZVI/SPC has substantial potential to degrade both types of contaminants, i.e., 49.6 % of 1,1,1-TCA and 39.7 % of TCE, but could not degrade the contaminants completely. On the contrary, larger amount of 1,1,1-TCA was degraded than TCE during the same interval of time under identical experimental conditions, indicating that Z-nZVI is more efficient to degrade 1,1,1-TCA as compared to TCE.

  • The Z-nZVI/SPC oxidation along with RA (NH2OH•HCl) can be used as a promising technique for remediation of groundwater contaminated with 1,1,1-TCA and TCE. However, further improvements in Z-nZVI/SPC degradation performance with a novel approach will be focused in our future research work.