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Stimulating Nitrate Removal with Significant Conversion to Nitrogen Gas Using Biochar-Based Nanoscale Zerovalent Iron Composites

College of Urban and Environmental Sciences, Northwest University, Xi’an 710127, China
Author to whom correspondence should be addressed.
Water 2022, 14(18), 2877;
Submission received: 24 August 2022 / Revised: 9 September 2022 / Accepted: 11 September 2022 / Published: 15 September 2022
(This article belongs to the Special Issue Groundwater Quality and Public Health)


For efficient and environmentally friendly removal of nitrate from groundwater, biochar-based nanoscale zerovalent iron composites were prepared, where biochar was derived from pine sawdust at 4 different pyrolysis temperatures. The results show that biochar with different pyrolysis temperatures played a great role in both nitrate removal efficiency and nitrate conversion rate to nitrogen gas for the prepared composites. Specifically, the composite with biochar pyrolyzed at 500 °C, ZB12-500, showed the best performance in both nitrate removal and conversion to nitrogen gas. With an initial solution pH from 5 to 10, ZB12-500 maintained high removal efficiencies varying from 97.29% to 89.04%. Moreover, the conversion of nitrate to nitrogen gas increased with the initial nitrate concentration, and it reached 31.66% with an initial nitrate concentration of 100 mg/L. Kinetics analysis showed that the nitrate removal process fit well with a two-compartment first-order kinetic model. Meanwhile, the test of nitrate removal by ZB12-500 in synthetic groundwater showed that HCO3 and SO42− limited nitrate removal but improved nitrate conversion to nitrogen gas. Furthermore, the nitrate removal mechanism suggested that biochar could facilitate electron transfer from zero valent iron to nitrate, which led to high nitrate removal efficiency. In addition, the interaction of ferrous ions and the quinone group of biochar could increase the nitrate conversion to nitrogen gas. Therefore, this study suggests that ZB12-500 is a promising alternative for the remediation of nitrate-contaminated groundwater.

1. Introduction

Groundwater plays an important role in human freshwater resources. About 20% of the world’s fresh water supply is provided by groundwater. Therefore, ensuring the safety of groundwater is of great importance to human beings. However, as a result of rapid expansion in modern agriculture and industry, inorganic nitrogen pollution has become a major problem in the world. The main types of nitrogen present in water are nitrate, nitrite, and ammonia, but the most common pollutant is nitrate, and nitrate nitrogen pollution has been reported as a worldwide pollution problem [1,2,3]. Nitrate nitrogen entering water may lead to eutrophication or other negative effects on water quality [4,5,6,7]. On the other hand, nitrate can also end up in the human body through groundwater and cause methemoglobinosis, also known as “blue baby syndrome,” and can be converted into carcinogenic nitrite amine preforms [8]. The WHO limited the concentration of NO3-N to 10 mg/L [9]. Therefore, advanced treatment technologies are needed to remove nitrate from groundwater in an economical and environmentally friendly manner.
At present, mainstream nitrate removal technologies in groundwater include chemical catalysis of nitrate reduction, anion exchange, low-pressure reverse osmosis, and microbial methods. Among them, the anion-exchange method and reverse osmosis method both have the disadvantage of requiring frequent regeneration medium and producing secondary pollutants [10,11], while biological denitrification requires a long repair time and generates sludge, which requires a large amount of maintenance cost [12]. Compared with other methods, the chemical reduction method provides the benefits of quick effect, lower cost, and no secondary pollution, and is more suitable for nitrate remediation in groundwater [13].
Nanoscale zero-valent iron (nZVI) has been reported to effectively remedy groundwater contamination in recent years because of its finer particle size, higher reactivity, and larger specific surface area to remove pollutants. Although nZVI can effectively reduce nitrate, it also has some disadvantages, such as easy agglomeration, easy corrosion, poor electronic selectivity, low reactivity, and poor product selectivity [14]. For these reasons, a variety of modification methods for nZVI have been studied. The bimetallic method is to dope some high-potential metals into the nZVI so that nZVI and the doped metals can make up micro-macroscopic coupled electrode systems to get rid of more nitrate and product selectivity of the material. Sparis et al. used ZVI-5%Cu particles doped with Cu ions into ZVI to reduce more than 80% of nitrate in 20 min and completely remove it in 1 h, whereas ZVI alone removed only 74.5% of nitrate in 1 h [15]. Lubphoo et al. found that a trimetal (Pd-Cu)-ZVI material using Cu and Pd as catalysts was more efficient in reducing nitrate than pure nZVI, as well as an increased ratio of N2 production [16]. Zhang et al. found that the addition of palladium increased the gas production of nitrate, nitrite, and ammonia recovered from aqueous and solid-phase supports [17]. Although the dopped metal materials could be beneficial for nitrate removal and its selective conversion to nitrogen gas, they also mean increased material cost and environmental risk. Pre-magnetization could also boost nZVI’s reactivity due to its superiority in magnetic memory. The nitrate removal rate of the pre-magnetized Fe0 system was 1.99 times better than that of the non-pre-magnetized one [18]. During nitrate reduction, the magnetic field gradient force drove nitrate gathering at the surface of the pre-magnetized Fe0 system, and meanwhile led to more nitrate conversion to nitrogen gas [19]. However, it is difficult to apply the pre-magnetization method on a large scale in the field of groundwater treatment engineering. In fact, nZVI prepared by the loading method can prevent aggregation of ZVI in the reaction, provide more active sites, and have a wider range of pH application conditions [20]. The application scenario of the material is more suitable for practical groundwater remediation.
The carrier materials commonly used in the loading method include organic materials such as alginate matrix [21], polymeric styrene anion exchanger [22], and inorganic materials such as activated carbon [23], biochar [24], zeolite [20], etc. Biochar often presents with better porosity and specific surface area, which makes it a promising carrier for nanoscale materials [25]. Namasivayam et al. used a waste coconut shell to prepare ZnCl2 activated carbon to recover nitrate from water. The results showed that pH had a great impact on the recovery of nitrate, and the desorption rate could reach 58% and 92% at pH 2 and 11, while the desorption rate was negligible at pH 3–10 [26]. Kamyar et al. prepared TBC by impregnating magnetic nanoparticles on tea biochar to remove heavy metals and nutrients in water, up to 147.84 mg/g of Ni2+, 160.00 mg/g of Co2+, 49.43 mg/g of NH4+ and 112.61 mg/g of PO43− could be adsorbed onto tested biochar [27]. For nZVI, nitrate was mainly conversed to NH4+ (93.5%) instead of N2 (5.7%), while the N2 conversion ratio of ZVI/BC composite can reach 60.1% [28]. Oh et al. used straw as raw material to prepare biochar-loaded nZVI material at 900 °C and reaction results showed that NO3-N was almost completely removed and the selectivity of the N2 product was also very high [29]. Gao et al. produced biochar-loaded ZVI at 400 °C to remove Cr6+ from aqueous solutions and reached a maximum removal capacity of 126 mg/g at pH 2.5, whereas ZVI was highly agglomerated at the same pH [30]. Wei et al. prepared BC/nZVI composites with different mass ratios from straw to remove nitrate nitrogen from water. The prepared composite presented superiority to nZVI, and its removal capacity was 229 mg NO3-N/g [28]. Some studies have concluded that the process of nitrate removal by ZVI composite materials can be explained by Equations (1)–(10) [31,32,33,34]. Obviously, biochar-supported nZVI composites have great potential for remediation of nitrate contamination. As a low-cost and environmentally friendly natural material, biochar could also mediate environmentally related abiotic redox processes [35,36], so it is desirable to use composite materials for nitrate treatment in groundwater. When biochar is used as the carrier of composite materials, pyrolysis temperature affects the properties of composite materials by changing the physicochemical properties of biochar (including specific surface area, functional group, hydrophobicity, and graphitization) [37,38,39]. However, little is known about the effect of the pyrolysis temperature of biochar support on the product selectivity for nitrate removal by nZVI/BC composites.
Fe 0 + 2 H + Fe 2 + + H 2
2 Fe 0 + O 2 + 2 H 2 O 2 Fe 2 + + 4 OH
Fe 0 + NO 3 + 2 H + Fe 2 + + NO 2 + H 2 O
4 Fe 0 + NO 3 + 10 H + 4 Fe 2 + + NH 4 + + 3 H 2 O
3 Fe 0 + NO 2 + 8 H + 3 Fe 2 + + NH 4 + + 3 H 2 O
5 Fe 0 + 2 NO 3 + 12 H + 5 Fe 2 + + N 2 g + 6 H 2 O
8 Fe 0 + 3 NO 3 + 9 H 2 O BC 4 Fe 2 O 3 + 3 NH 4 + + 6 OH
10 Fe 0 + 6 NO 3 + 3 H 2 O BC 5 Fe 2 O 3 + 3 N 2 g + 6 OH
3 Fe 0 + NO 3 + 3 H 2 O BC Fe 3 O 4 + NH 4 + + 2 OH
18 Fe 0 + 3 Fe 2 + + 10 NO 3 + 2 H 2 O BC 7 Fe 3 O 4 + 5 N 2 g + 4 OH
Our previous study have proved the feasibility of the nZVI/BC for nitrate removal from groundwater [28]. In order to explore the effect of biochar prepared at different pyrolysis temperatures on the removal of nitrate from groundwater by nano zero-valent iron/biochar composite here, nZVI/BC composites with a mass ratio of 1:2 (ZB12) were prepared by using biochar at different pyrolysis temperatures. The prepared composites were characterized using surface analysis techniques, including a scanning electron microscope, X-ray diffraction pattern, Fourier transform infrared spectroscopy, X-ray photoelectron spectroscopy, and Brunauer–Emmett–Teller (BET) specific surface area measurement. Moreover, nitrate removal efficiencies and their product selectivity were evaluated under different conditions of groundwater, including dosage, initial pH, initial nitrate concentration, and co-existing ions. In addition, the nitrate removal kinetics were investigated, and a mechanism of nitrate removal was proposed.

2. Materials and Methods

2.1. Materials

Anhydrous ethanol (CH3CH2OH, 99.7%) was obtained from Fuyu Chemical Co., Ltd. (Tianjin, China). Sodium borohydride (NaBH4, 97.0%), potassium nitrate (KNO3, 99.0%), potassium persulfate (K2S2O8, 99.0%), ammonium chloride (NH4Cl, 99.8%), and sodium nitrite (NaNO2, 99.0%) were acquired from Kermel Chemical Reagent (Tianjin, China). Ferrous sulfate heptahydrate (FeSO4·7 H2O, 99.0%) was purchased from Sheng Ao Chemical Reagent (Tianjin, China). The raw pine sawdust was collected from Weinan, Shaanxi Province, China. The sawdust was rinsed three times with deionized (DI) water and then dried overnight in an oven at 80 °C. nZVI used for comparison were purchased from Xiangtian Nanomaterials Co., LTD., Shanghai, China.

2.2. Synthesis of ZB12

Biochar was prepared by pyrolysis. Briefly, the crucible was filled with pre-treated pine sawdust, which was then put into a muffle furnace for carbonization. The carbonization was under a nitrogen atmosphere with a nitrogen flow rate of 80 mL/min and a heating rate of 10 °C/min. The carbonization process lasted 4 h at target temperatures of 350, 500, 650 and 800 °C, respectively. Then the heating stopped, and the muffle furnace was cooled down to room temperature. Next, the biochar was removed and pulverized through a 150-mesh sieve for later use.
The synthesis of ZB12 was conducted based on a sodium borohydride reduction method, as in Equation (11) [40]. According to our previous study [28], 1:2 was the optimal mass ratio of iron content to biochar, so this ratio was used in the sample preparation in this study. Briefly, 1.39 g FeSO4·7H2O was dissolved in 50 mL deionized water, then 0.56 g of the prepared biochar sample was added, and next, the mixture was placed in an anaerobic flask for 1 h under ultrasound to make full contact with the iron solution. Then, nitrogen was purged for 30 min to create an anoxic environment. According to the reaction formula, NaBH4 with a slightly excess amount was dissolved in deionized water of 20 mL, and then added dropwise. The Fe2+ was completely reduced to Fe0 after reaction at 120 rpm in a shaker for half an hour. The prepared composite was rinsed using deionized water and then anhydrous ethanol, and such rinsing was repeated three times and then dried in a constant temperature water bath under nitrogen protection. The composites prepared from biochar pyrolysis at 350, 500, 650 and 800 °C were labeled ZB12-350, ZB12-500, ZB12-650, and ZB12-800, respectively. The preparation process for ZB12 is shown in Figure 1.
Fe 2 + + 2 BH 4 + 6 H 2 O Fe 0 + 2 B OH 3 + 7 H 2

2.3. Characterization

According to previous studies, the nitrate removal reaction can reach about half of the reaction process after 1 h, and the reaction is almost complete after 24 h [41]. Therefore, the sample that reacted for 1 h is selected as the sample during the reaction, and the sample that is reacted for 24 h is the sample after the reaction.
The morphology of the samples before and after the reaction was investigated using a scanning electron microscope (SEM). The specific surface area (SSA) was measured using a surface area BET analyzer. X-ray diffraction (XRD) patterns were determined using an X-ray diffractometer. The chemical composition of the samples before, during, and after the reaction was analyzed according to the diffraction peaks. The scanning range 2θ was 10°~70°. In addition, the surface function groups were analyzed using Fourier transform infrared spectroscopy (FTIR). Meanwhile, the surface composition of the samples was investigated using X-ray photoelectron spectroscopy (XPS).

2.4. Experiments for Chemical Reduction of Nitrate by ZB12

Initial concentrations of nitrate were investigated in the range of 30–100 NO3-N mg/L, which was set according to a survey of nitrate-contaminated groundwater in China [42]. In addition, nitrate removal experiments were conducted in 40 mL nitrate solution (30 NO3-N mg/L) with 0.2 g of the prepared samples (nZVI or ZB12 samples). Moreover, the role of dosage in a 40 mL solution (30 NO3-N mg/L) was investigated with 0.08, 0.12, 0.16, 0.2, 0.24, and 0.28 g of ZB12 samples. The impact of the solution pH was studied between 5 and 10, which was adjusted with HCl or NaOH solutions. On the other hand, the effects of co-existing ions were also investigated according to Table 1, which was obtained from a survey of groundwater samples in Weinan City, Shaanxi Province, China. In the above experiments, all the reactions happened in anaerobic bottles, which were sealed with butyl rubber stopper.
Kinetic and mechanistic experiments were carried out as follows: 0.2 g of ZB12 samples was reacted with nitrate with an initial concentration of 30 NO3-N mg/L in 40 mL solution in anaerobic bottles. Samples in each bottle were measured in turn at presetting time points. The measurement included pH, DO, oxidation-reduction potential (ORP), and the concentrations of nitrogen species (NO3-N, NO2-N, NH4+-N and total nitrogen (TN)). The values of DO, pH, and ORP were monitored by a portable DO meter and a pH meter, respectively. The concentrations of NO3-N, NO2-N, NH4+-N and TN were determined by UV-spectrophotometric method, spectrophotometric method, Nessler’s reagent spectrophotometry, and alkaline potassium persulfate digestion spectrophotometry, and the instrument for determination was spectrophotometer. To explore the selectivity of nitrate reduction, gas samples were gathered using a microsyringe. N2 and N2O in the gas samples were analyzed by GC-TCD with a Molecular Sieve 5A column and a Porapak Q column, respectively [43]. NOx (NO2 and NO) were analyzed using the chemiluminescence detection method [44]. The models of all the instruments used in the experiment are shown in Table 2. All the experiments were conducted at room temperature (25 ± 2 °C) with 120 rpm shaking.

2.5. Analysis Method

2.5.1. Calculation Method

Many previous studies have shown that nitrite, ammonia, and nitrogen are the main products in the process of nitrate reduction, while NOx and N2O are basically not found [45,46], the same result was also found in the preliminary experiment of this study. Therefore, in this experiment, only nitrite nitrogen, ammonia nitrogen, and nitrogen were considered in the reduction products. The removal efficiency of NO3-N (η) and the product conversion ratio (Sproduct) are calculated as follows:
η = C 0 C NO 3 C 0 × 100 %
S NH 4 + = C NH 4 + C 0 C NO 3 × 100 %
S NO 2 = C NO 2 C 0 C NO 3 × 100 %
S N 2 = C 0 C NO 3 C NO 2 C NH 4 + C 0 C NO 3 × 100 %
where, C0 is the initial concentration of NO3-N, mg/L; CNO3, CNH4+ and CNO2 are the concentration of NO3-N, NH4+-N and NO2 after reaction, mg/L.

2.5.2. Kinetic Studies

Many studies have shown that the removal process of nitrate from solution by ZVI composites conforms to first-order or second-order kinetics [47]. The removal process of NO3-N from the solution by the ZB12-500 composite can be divided into two stages: the fast removal stage and the slow removal stage. Therefore, the first-order kinetic model, second-order kinetic model, and two-compartment first-order kinetic model were used for analysis and fitting in this experiment, and the fitting equations were as follows:
C t C 0 = e k 1 t
1 C t 1 C 0 = k 2 t
C t C 0 = f f     e k f t + f s     e k s t
where Ct (mg/L) is the residual concentration of nitrate, C0 (mg/L) is the initial concentrations of nitrate; k1 and k2 (1/h) are the reaction rate constants of first and second order reaction kinetics, respectively; ff and fs are the proportion of fast and slow compartment removal in the total removal, respectively, and ff + fs = 1; kf and ks (1/h) are the fast and slow compartment reaction rate constants, respectively.

3. Results

3.1. Characterization of ZB12-500

The SEM analysis results of ZB12-500 before, during, and after the reaction are shown in Figure 2a–c. Figure 2a shows the ZB12-500 material before the reaction. The particle size of the composite material is about 30 μm, roughly spherical, smooth surface, and rich honeycomb channels. nZVI particles are evenly distributed on the surface and pores of biochar particles, such as a spider web, and nZVI particles in different pores are separated by a carbon skeleton. This is because of the rich porous structure and large specific surface area of biochar [48]. This image shows that the ZB12-500 composite was well prepared by the sodium borohydride method. Figure 2b shows ZB12-500 during the reaction. Compared with before the reaction, the nZVI particles on the surface of the ZB12-500 were obviously corroded, and the corrosion of nZVI particles was mostly surface corrosion. Some of the porous structures on the surface of the ZB12-500 were blocked because the nZVI lost electrons, and the iron ions diffused from the inner core of the nZVI to the outer core, forming iron oxides on the surface of the particles. In this process, there may be mechanisms such as iron dissolution, migration, and reagglomeration of nZVI particles, and migration of iron ions between different particles [49]. Figure 2c shows ZB12-500 after the reaction. The surface morphology of the material was highly crystallized, and the porous structure on the surface of the composite material completely disappeared and was covered by iron oxide.
The SSA analysis results of ZB12-500 before, during, and after the reaction are shown in Table 3. After the reaction, the SSA of the material increased. The possible reason is that the iron oxides generated by the nZVI reaction were redistributed on the biochar support, which may cover the original nZVI particles or other active sites, resulting in a larger gap between the iron oxides and an increase in SSA.
The XRD analysis results of ZB12-500 before, during, and after the reaction are shown in Figure 2d. ZB12-500B, ZB12-500D, and ZB12-500A represent ZB12-500 samples before, during, and after the reaction, respectively. The characteristic peaks at 2θ = 44.8°, 65.32°, and 82.60° represent planes (110), (200), and (211) of the body-centered cubic crystal structure in nZVI particles, respectively [50]. Before the reaction, the ZB12-500B sample showed a diffraction peak at 2θ = 44.8°, which corresponds to the body-centered cubic α-Fe0 (110) crystal plane, indicating that the prepared ZB12-500 material contains α-Fe0, and nZVI particles are successfully loaded on the surface of biochar support [51,52]. The peak intensity of Fe0 in the ZB12-500D sample decreases significantly, while the diffraction peak of Fe0 in the ZB12-500A sample completely disappears, indicating that in the process of reducing NO3-N by ZB12-500, nZVI was all consumed or covered by reaction products. The characteristic peaks of Fe2O3 and Fe3O4 located at 2θ = 35.5° appear in sample ZB12-500D, which is widened to a certain extent in sample ZB12-500A, while these two peaks are not found in sample ZB12-500B, indicating that the main iron oxides generated after the reduction reaction are Fe2O3 and Fe3O4. Some scholars have found that Fe3O4 is formed at the interface between nZVI and iron oxide, while Fe2O3 is formed at the interface between iron oxide and water [53,54,55].
The results of FTIR analysis of ZB12-500 before, during, and after the reaction are shown in Figure 3. As can be seen from the figure, these three samples contain abundant functional groups, among which a characteristic peak with high strength appears near 3430 cm−1, which corresponds to the tensile vibration of the –OH bond [56,57,58,59]. The characteristic peak intensity at 1620 cm−1 is slightly smaller and is related to the C=O and –OH bonds of carbonyl or carboxyl groups. The intensity of characteristic peaks at 1320 cm−1 and 1100 cm−1 is small, which represents the tensile vibration of the C–O bond. The characteristic peak intensity at 670 cm−1 is weak, which is the characteristic peak of the Fe–OH bond [60,61,62]. The number and types of functional groups will greatly affect the physical and chemical properties of materials, especially the –OH functional group, which will affect the absorption of other substances by biochar [63]. Compared with ZB12-500B, the intensity corresponding to the –OH characteristic peak of ZB12-500A and ZB12-500D at 3430 cm−1 increased, while the intensity of the characteristic peaks at 1620 cm−1, 1320 cm−1 and 670 cm−1 decreased slightly, which may be due to the formation of C–O–Fe, Fe–O–Fe, and C–N–Fe.
Figure 4a–c shows the full spectrum scan of ZB12-500 material before, during, and after the reaction. The peaks of binding energies at 710 eV, 530 eV, and 284 eV correspond to the absorption peaks of Fe 2p, O 1s, and C 1s, respectively [60]. The main elements in the three samples are Fe, O, and C, but the content of the elements is different. Among them, the Fe element is mainly from nZVI and iron oxide, the O element is mainly from oxygen-containing functional groups in biochar, and C element is mainly from biochar. By comparing the scanning spectra, it can be found that the content of the Fe element increased continuously after the beginning of the reaction. This is because although nZVI was consumed during the reaction, the increased content of Fe oxide exceeded the consumed nZVI content. One hour after the reaction between ZB12-500 and NO3-N solution, the content of the Fe element increased greatly, and in the following 23 h, the content of the Fe element increased slightly, indicating that the reaction was nearly complete after one hour of reaction, which is consistent with the change of the iron element in XRD. The content of the O element in the ZB12-500 increased first and then decreased in the reaction process, which may be related to the consumption of nZVI and the formation of iron oxides in the reaction process. The change in the –OH group content in the FTIR spectra also showed the same trend. The content of element C decreased by 12.98% in 1 h after the reaction, and then decreased by 0.74% in the following 23 h, indicating that the functional group containing C participated in the reaction. This is consistent with the variation of the C element in the XRD and FTIR spectra.
Figure 4d–f shows the high-resolution XPS scan spectra over Fe 2p of ZB12-500 before, during, and after the reaction. The peaks at the binding energy of 707 eV, 710 eV, and 712 eV represent the absorption peaks of Fe0, Fe2+ and Fe3+ in Fe 2p3/2, respectively. The peaks at the binding energy of 721 eV, 723 eV, and 725 eV represent the absorption peaks of Fe0, Fe2+ and Fe3+ in Fe 2p1/2, respectively [64,65]. The relative contents of the three different valence iron elements changed obviously before, during, and after the reaction of ZB12-500, especially the peak of Fe0 at 707 eV binding energy, and the intensity of this peak almost disappeared after the reaction. Table 4 shows the relative contents of Fe in different valence states in ZB12-500 before, during, and after the reaction. Compared with the ZB12-500 before the reaction, the content of Fe0 in the composite during and after reaction decreased by 3.84% and 5.00%, respectively, indicating that nZVI was continuously consumed during the whole reaction. The increased of Fe3+ content mainly occurred in the first hour of the reaction process, and then decreased in the next 23 h. Meanwhile, the increased of Fe2+ content mainly occurred in the last 23 h of the reaction process, indicating that Fe3+ was converted to Fe2+ in the later reaction.

3.2. Effect of the Pyrolysis Temperature of ZB12-500 on Biochar

Figure 5 shows ZB12-500 was the most effective reactant, followed, in decreasing order, by ZB12-350, ZB12-650, ZB12-800, and nZVI with removal efficiencies of 93.49%, 86.92%, 86.65%, 84.90%, and 40.14%, respectively. In the five different composites, the N2 conversion ratios from large to small were ZB12-500 > ZB12-350 > ZB12-800 > ZB12-650 > nZVI, corresponding to 27.34%, 24.12%, 24.11%, 16.83%, and 5.71%, respectively.

3.3. Effect of the Dosage of ZB12-500

Figure 6 shows how the dosage of ZB12-500 affected the removal efficiency of nitrate. When the dosages of ZB120-500 were 2, 3, 4, 5, 6, and 7 g/L, the removal efficiencies were 54.62%, 72.62%, 89.49%, 93.98%, 94.78%, and 96.04%, respectively. A positive correlation was observed between the dosage of ZB12-500 and the efficiency of nitrate removal. In terms of the selectivity of nitrogen products, when the dosage was less than 5 g/L, the proportion of various reduction products basically did not change, and the N2 conversion ratio was about 27%, while when the dosage was more than 5 g/L, the N2 conversion ratio decreased to about 24%, and the conversion ratio of NH4+-N and NO2 increased slightly with the increase in dosage.

3.4. Effect of pH

Figure 7 shows the effect of the initial pH on nitrate removal by ZB12-500. Obviously, the removal efficiency was negatively correlated with the initial pH of the solution. When the initial pH increased from 5 to 10, the removal efficiency decreased from 97.29% to 89.04%. The N2 conversion ratio was the highest when pH = 5, followed by pH = 6, and the lowest when pH = 7, which were 27.13%, 26.38%, and 21.92%, respectively. In particular, ZB12-500 exhibited similar nitrate reduction results at initial pH values of 5 and 6.

3.5. Effect of Initial Nitrate Concentration

Figure 8 shows the effect of the initial concentration of nitrate (NO3-N) on the removal of nitrate by ZB12-500. When the initial concentration of NO3-N were 30 mg/L, 50 mg/L, 70 mg/L and 100 mg/L, the removal efficiencies were 93.94%, 82.60%, 75.3%, and 51.42%, respectively, and the N2 conversion ratios were 26.82%, 28.45%, 29.74%, and 31.66%, respectively.

3.6. Effect of Co-Existing Ions

Figure 9 shows the effect of the co-existing ions on nitrate removal by ZB12-500. Common ions (Na+, Mg2+, Ca2+, Cl, SO42−, and HCO3) in groundwater were investigated. According to the experimental data in the previous sections, without adding co-existing ions, the removal efficiency of NO3-N exceeded 93%, and the conversion ratio of N2 exceeded 25% under the same reaction conditions. However, after adding co-existing ions Na+, Mg2+, Ca2+, Cl, SO42−, and HCO3 into the solution, the removal efficiencies of NO3-N were 91.14%, 95.20%, 94.33%, 92.88%, 80.20%, and 57.00%, and the conversion ratios of N2 were 26.30%, 26.01%, 25.02%, 25.82%, 29.49%, and 37.01%, respectively.

3.7. Kinetics

The kinetics data were fitted using first-order kinetic, second-order kinetic, and two-compartment first-order kinetic equations, as shown in Figure 10. The whole removal process was obviously divided into two stages: 0–2 h for the rapid removal stage and 2–24 h for the slow removal stage. The kinetics parameters fitted by the kinetics model are shown in Table 5. The results show that the R2 values of the three kinetics models were all above 0.99, indicating that both adsorption and reduction reactions existed in the removal process. The R2 of the two-compartment first-order kinetic model was the highest, which was 0.997, indicating that it is more reasonable to use the two-compartment first-order kinetic model to explain the process of NO3-N removal by the ZB12-500 composite. From the two-compartment first-order kinetic parameters, the main stage in the whole removal process was the fast compartment reaction stage, accounting for 92.5%, while the slow compartment reaction stage only accounted for 7.5%. The fast compartment reaction rate constant was 3.093 h−1, and the slow compartment reaction rate constant was 0.038 h−1.

4. Discussion

4.1. Effect of the Pyrolysis Temperature of ZB12-500 on Biochar

For nZVI, the removal efficiency of nitrate is greatly affected by agglomeration. The low N2 selectivity of pure nZVI is also affected by the natural defects of the materials. In the reduction process, after the surface active site of the aggregate is inactivated by the reaction, the internal active site is also inactivated by being covered. For ZB12, the removal efficiency of NO3-N firstly increased and then decreased with the increase of biochar pyrolysis temperature. However, the selectivity of the nitrogen products had no obvious rule with the pyrolysis temperature of the biochar carrier. Although increasing the temperature of pyrolysis increases the electronic conductivity and the degree of graphitization of biochar, it also leads to the loss of functional groups [37,66]. Some scholars have pointed out that with an increase in pyrolysis temperature, the number of functional groups of wood-based biochar (mainly –OH and aliphatic C–H functional groups) and grass-based biochar (mainly C–O functional groups) decreases [67]. The trend of nitrate removal efficiency and nitrogen product selectivity with pyrolysis temperature proves to some extent that when the pyrolysis temperature of biochar is 500 °C, the types and number of functional groups related to reduced nitrate in ZB12 reach an ideal equilibrium state with their electrical conductivity and graphitization degree. The NO2-N produced by the reaction of the five kinds of samples with nitrite nitrogen was very small, indicating that NO2-N was the intermediate product of the reaction. Briefly, ZB12-500 was the best reactant for nitrate removal and was selected for the following studies.

4.2. Effect of the Dosage of ZB12-500

When the dosage is low, the removal efficiency increased rapidly with increasing dosage because the increased dosage provided more Fe0 active sites, which is the reason for the positive correlation. At the same time, there were sufficient reaction raw materials for ZB12-500 in the whole reaction process, and the active site of ZB12-500 could be fully utilized in the reduction process. Therefore, when the dosage was less than 5 g/L, ZB12-500 experienced the same reaction conditions and environment, resulting in the same nitrogen product selectivity. When the dosage was greater than 5 g/L, nitrate dilution significantly inhibited the reactivity of the residual nitrate, leading to incomplete reduction of more nitrate, i.e., an increased proportion of intermediates accumulated, resulting in a slight change in nitrate removal efficiency and a slight reduction in N2 selectivity. Some previous studies have found similar results [68,69]. Accordingly, a dosage of 5 g/L is optimum for this study.

4.3. Effect of pH

From the changing trend of NO3-N removal efficiency with the initial pH, it can be seen that the reaction between the ZB12-500 material and NO3-N is an acidophilic reaction. Hao et al. also reached this conclusion in their research results [70]. The reasons are as follows: Under acidic conditions, H+ in the solution is increased, the adsorption of NO3-N on the biochar surface is enhanced, and protons also participate in the reduction process of nitrate nitrogen [71]. Under alkaline conditions, the increase of OH leads to more metal hydroxide precipitation (Fe(OH)2 and Fe(OH)3) and metal carbonate (FeCO3). This led to increased corrosion of the nZVI. These iron oxides limited the diffusion of nitrate ions and coated the zero-valent iron, reducing the active sites on the surface of ZB12-500 composite and reducing the removal efficiency of NO3-N [72]. Many studies have found that pH can affect the reduction of nitrate by ZVI particles [19,73,74]. However, in this section, the composite material can achieve good selectivity on N2 (21.92–27.13%) within the initial pH range of 5–10. This may be related to the rich functional groups, huge specific surface area, and abundant active sites of biochar carriers with pyrolysis temperature of 500 °C [75]. The reason N2 conversion ratio of ZB12-500 was the lowest under neutral conditions was that the content of active sites or functional groups on biochar surface decreases under neutral conditions.

4.4. Effect of Initial Nitrate Concentration

The initial nitrate concentration is negatively correlated with its removal efficiency. The study of Sparis et al. also showed that the higher the initial nitrate nitrogen concentration, the lower the reduction rate constant [15]. In fact, when nitrate concentration in solution is low, the active sites on the ZB12-500 composite are relatively abundant, which makes the utilization rate of ZVI very high. However, for higher concentrations of nitrate in solution, nitrate ions compete for limited active sites on ZB12-500 composites, which makes it easier for nitrate anions to squeeze onto the ZB12-500 surface and then rapidly oxidize nZVI particles on the ZB12-500 surface, eventually seriously blocking the porous structure of ZB12-500. As a result, the material loses more active sites more quickly, which inevitably hinders nitrate reduction. In the experiment, N2 conversion ratio was positively correlated with the initial nitrate concentration, which might be the result of the influence of solution ion concentration on the surface charge density of the material. Mikami et al.’s study also showed that increased nitrate concentration was conducive to the transformation of reduced products into N2 rather than NH4+-N [76]. As the density of N-species on the surface of the composite became high with the increased in NO3-N concentration, the selectivity to N2 increased.

4.5. Effect of Co-Existing Ions

In this study, Na+, Mg2+, Ca2+ and Cl had insignificant effects on the removal efficiency of NO3-N, while SO42− and HCO3 significantly reduced the removal efficiency of NO3-N. General cations and low concentrations of Cl have a negligible effect on the removal of anionic NO3. However, the outer-spherically sorbing anions, especially SO42−, have a significant interference effect on nitrate adsorption, and adsorption competition will occur on the limited adsorption sites on the surface of ZB12-500 [77,78]. HCO3 can be ionized to form CO32− and H+ or hydrolyzed to form OH and H2CO3. However, the degree of hydrolysis of HCO3 is greater than the degree of ionization, so more OH is generated by hydrolysis, which can form Fe(OH)2, Fe(OH)3 and FeCO3 with iron ions in the system. Thus, the corrosion of nZVI was aggravated, and the formation of iron oxides will cover the outer surface of nZVI particles, preventing it from continuing to react with NO3-N, and ultimately reducing the reduction efficiency of NO3-N. The reason HCO3 obviously improved the product selectivity of N2 may be related to the atomic structure of HCO3. The atomic arrangement of HCO3 is planar, and the carbon in the center is bonded with three oxygen atoms (one C=O, one C–OH, and one C–O–). All of these functional groups are involved in the reduction process of NO3-N. If the ZB12-500 composite is applied to actual groundwater, the influence of HCO3 and SO42− should be considered.

4.6. Kinetics

The kinetics of nitrate removal complied with the two-compartment first-order kinetic equations. In the rapid removal stage, that is, within 2 h of reaction, the concentration of NO3-N decreased almost in a straight line, indicating that the reaction was strong and rapid at this time. In the slow removal stage, that is, within 2 h to 24 h after the reaction, the concentration of nitrate nitrogen decreased extraordinarily little, indicating that the reaction in the system was extremely slow and that the reaction was almost complete. At the beginning of the removal reaction, ZB12-500 reacted rapidly with NO3-N in solution when it just entered the system. This is because the initial concentration of nitrate is relatively high and the mass transfer driving force is large, so NO3-N is easily sent to the active site on the surface of the composite materials. This is manifested by the film diffusion of nitrate ions from the bulk liquid phase to the external surface of ZB12. As the reaction produced a large amount of iron oxide, the active site of nZVI was blocked so that nZVI could not be exposed to the system to react with the remaining NO3-N. Moreover, as the concentration of nitrate in the solution decreased, the mass transfer driving force also decreased, resulting in a slow reaction in the later stage.

4.7. Nitrate Reduction Mechanism

During the nitrate removal process, ZVI nanoparticles in ZB12 also corroded. In the XRD pattern, the disappearance of Fe0 peak and the appearance of Fe3O4 and Fe2O3 peaks in the composite also indicate the corrosion of nZVI and the appearance of iron oxides. The infrared spectrum shows that the peak strength and peak width of the –OH functional group increased significantly, indicating that –OH was involved in the reaction between ZB12 and NO3-N. XPS survey spectra shows that Fe2+ was consumed and generated during the reaction.
In order to deeply explore the reaction process between ZB12 and NO3-N, this section analyzes the nitrogen species (NO3-N, NO2-N, NH4+-N and TN), pH, ORP, and DO during the reaction process. The experimental results are shown in Figure 11 and Figure 12. According to the change in the concentration of each substance in the reaction process, the reaction can be basically divided into three stages.
The first stage occurred in 0–0.5 h. Here, the concentration of NO3-N decreased to 8.85 mg/L in a straight line because the active sites of ZVI on ZB12 were abundant at this time, which could fully react with nitrate in the solution. The concentrations of NO2-N, NH4+-N and TN at the time of 0.5 h were 0.5 mg/L, 10.18 mg/L and 19.53 mg/L, respectively, indicating that nitrite and ammonia were generated during this stage (Equations (3) and (4)). The decrease in TN concentration indicated that N2 was generated during the reaction (Equation (6)). Additionally, the pH increased from 7.5 to 9.67, indicating that H+ in the system participated in the reaction (Equation (1)); DO decreased from 7 mg/L to 0.9 mg/L, indicating that dissolved oxygen in the system competed with NO3-N and participated in the reaction (Equation (2)); and ORP decreased, indicating that the system was transformed into a reducing environment and Fe2+ was generated.
The second stage occurred at 0.5–8 h. Here, the concentration of NO3-N decreased from 8.85 mg/L to 1.65 mg/L, which slowed the reaction speed. The concentration of NO2-N decreased from 0.50 mg/L to 0.05 mg/L, indicating that nitrite was an intermediate product of the reaction (Equation (5)). The concentrations of NH4+-N increased from 10.18 mg/L to 18.88 mg/L, becoming the main component of TN. The pH dropped from 9.67 to 8.9, probably buffered by oxygen-containing functional groups in the biochar. The rise in ORP indicated the consumption of Fe2+ in the system (Equation (10)). The rise in DO was due to the release of adsorbed oxygen or the buffering effect of functional groups from the biochar. Combined with the characterization analysis, the slow reaction rate was affected by the formation of iron oxide in the reaction process. In addition, functional groups in biochar were also involved in the reaction (Equations (7)–(9)).
In the last stage (8–24 h). The concentration of NO3-N, TN and NO2-N were still decreasing, and the concentration of NH4+-N was slightly increasing, indicating that the reduction reactions still occur. pH, ORP, and DO leveled off, indicating that the reaction proceeded at a very low rate.
The reaction mechanism of the ZB12 composite with NO3-N is shown in Figure 13. The nZVI particles loaded on the surface and pores of biochar can react with NO3-N or DO in solution. During this period, electrons escape from the core and shell of nZVI to form Fe2+ and generate NO2, NH4+ and N2, with NH4+ being the main product and NO2 being partially converted to NH4+ as an intermediate. nZVI is converted into iron oxide to cover the surface of biochar or exist in solution, and the Fe2+ generated by the reaction contributes to the reduction of nitrate [79]. Carboxyl groups on the surface of biochar will exist in the form of esters [80], while ester groups mainly undergo hydrolysis reaction, so ester groups in biochar will react with H2O, nZVI, and Fe2+ to generate quinone groups (Equation (10)), which is consistent with the results of the FTIR spectrum. The generated quinone group can also provide H+ for the reaction, which converts NO3-N to N2 and itself into an ester group. In addition, ester groups on the surface of biochar can dissociate in a wide pH range, thus buffering the pH of the reaction and slowing down the reaction inhibition at a higher pH.
In addition to the role of biochar in regulating the pH of the system and increasing the reactivity, the conductivity of the carbon matrix and the electron-mediated ability of functional groups may also contribute to the reduction of nitrate nitrogen. The electrical conductivity of biochar contributes to the transfer of electrons from nZVI to nitrate, and the higher the pyrolysis temperature of biochar, the stronger the electrical conductivity of the carbon matrix [81]. The electron-mediating capacity of functional groups can be divided into electron donation capacity and electron acceptor capacity, among which the electron donation capacity is attributed to phenolic functional groups, and the electron acceptor capacity is attributed to quinone groups and concentrated aromatic hydrocarbons [82]. These REDOX functional groups can mediate the electron transfer process among nZVI, biochar, and nitrate.
In modified ZVI materials, carbon with high potential is used as the cathode and iron as the anode. When the two contact, many microscopic galvanic cells will be generated. The potential difference causes electrons to transfer from the ZVI core shell to the carbon. Some studies have shown that nitrate nitrogen reduction occurs primarily on the carbon surface with a higher electric potential, which reduces the hindering effect of iron oxides on nitrate reduction [15,83]. Therefore, for the ZB12 composite material in this paper, a huge number of miniature galvanic couples can be formed between biochar and nZVI. nZVI acts as the anode, losing electrons and being oxidized to ferrous ions, and biochar acts as the cathode for nitrate reduction.

5. Conclusions

On the basis of the 1:2 mass ratio of nZVI to biochar studied in our previous work [28], ZB12 was successfully prepared. The effects of different biochar pyrolysis temperatures on nitrate removal by the ZB12 composite were explored. It was found that the best biochar pyrolysis temperature was 500 °C. In addition, ZB12 has a higher N2 conversion ratio (21.9~27.13%) in a wide pH range (5–10) under the premise of high nitrate removal efficiency (89.04–97.59%), which is more environmentally friendly in practical application. Increasing the initial concentration of nitrate would lead to a decrease in the removal efficiency, but a higher density of N-species on the surface of the composite might lead to an increase in the conversion ratio of N2. The co-existence of HCO3 or SO42− in the solution can reduce the removal efficiency of NO3-N to 57.00% or 80.20% and increase the conversion ratio of N2 to 37.01% or 29.49%, respectively. The removal of nitrate by ZB12 was in accordance with the two-compartment first-order kinetics. Biochar plays a mediating role in the reduction of nitrate by nZVI. The pyrolysis temperature of biochar will affect its electrical conductivity and the electron-mediated ability of its surface functional groups, thus affecting the removal of nitrate by the ZB12 composite and the formation of products. Therefore, the results of this study provide further references for the eco-friendly removal of nitrate from groundwater.

Author Contributions

Conceptualization, A.W.; methodology, S.L. (Siyuan Liu) and A.W.; software, S.L. (Siyuan Liu); validation, S.L. (Shaopeng Li) and W.X.; formal analysis, S.L. (Siyuan Liu); investigation, W.X.; resources, X.H. and S.L. (Shaopeng Li); data curation, S.L. (Siyuan Liu); writing—original draft preparation, S.L. (Siyuan Liu) and X.H.; writing—review and editing, A.W.; visualization, X.H.; supervision, A.W. and X.H.; project administration, A.W.; funding acquisition, A.W. All authors have read and agreed to the published version of the manuscript.


This research work was funded by the National Natural Science Foundation of China (NSFC) (Grant No. 51208424).

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

Not applicable.

Conflicts of Interest

All authors certify that they have no affiliations with or involvement in any organization or entity with any financial interest or non-financial interest in the subject matter or materials discussed in this manuscript. The authors declare no conflict of interest.


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Figure 1. Preparation process of ZB12.
Figure 1. Preparation process of ZB12.
Water 14 02877 g001
Figure 2. SEM images of ZB12-500 (a) before, (b) during, and (c) after the reaction. (d) XRD patterns of ZB12-500 before, during, and after the reaction.
Figure 2. SEM images of ZB12-500 (a) before, (b) during, and (c) after the reaction. (d) XRD patterns of ZB12-500 before, during, and after the reaction.
Water 14 02877 g002aWater 14 02877 g002b
Figure 3. FTIR spectra of ZB12-500 before, during, and after the reaction.
Figure 3. FTIR spectra of ZB12-500 before, during, and after the reaction.
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Figure 4. XPS survey spectra of ZB12-500 (a) before, (b) during, and (c) after the reaction. High-resolution XPS scan spectra over Fe 2p of ZB12-500 (d) before, (e) during, and (f) after the reaction.
Figure 4. XPS survey spectra of ZB12-500 (a) before, (b) during, and (c) after the reaction. High-resolution XPS scan spectra over Fe 2p of ZB12-500 (d) before, (e) during, and (f) after the reaction.
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Figure 5. Nitrate removal by nZVI and ZB12 prepared with different pyrolysis temperatures (initial NO3-N concentration: 30 mg/L, pH 6, dosage: 5 g/L).
Figure 5. Nitrate removal by nZVI and ZB12 prepared with different pyrolysis temperatures (initial NO3-N concentration: 30 mg/L, pH 6, dosage: 5 g/L).
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Figure 6. Effect of ZB12-500 dosage (initial NO3-N concentration: 30 mg/L, pH 6, ZB12 sample: ZB12-500).
Figure 6. Effect of ZB12-500 dosage (initial NO3-N concentration: 30 mg/L, pH 6, ZB12 sample: ZB12-500).
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Figure 7. Effect of initial pH (initial NO3-N concentration: 30 mg/L, Dosage: 5 g/L, ZB12 sample: ZB12-500).
Figure 7. Effect of initial pH (initial NO3-N concentration: 30 mg/L, Dosage: 5 g/L, ZB12 sample: ZB12-500).
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Figure 8. Effect of initial nitrate (NO3-N) concentration (pH 6, dosage: 5 g/L, ZB12 sample: ZB12-500).
Figure 8. Effect of initial nitrate (NO3-N) concentration (pH 6, dosage: 5 g/L, ZB12 sample: ZB12-500).
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Figure 9. Effect of co-existing ions (initial NO3-N concentration: 30 mg/L, pH 6, Dosage: 5 g/L, ZB12 sample: ZB12-500).
Figure 9. Effect of co-existing ions (initial NO3-N concentration: 30 mg/L, pH 6, Dosage: 5 g/L, ZB12 sample: ZB12-500).
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Figure 10. Kinetics of ZB12-500 for NO3-N. (initial NO3-N concentration: 30 mg/L, pH 6, Dosage: 5 g/L, ZB12 sample: ZB12-500).
Figure 10. Kinetics of ZB12-500 for NO3-N. (initial NO3-N concentration: 30 mg/L, pH 6, Dosage: 5 g/L, ZB12 sample: ZB12-500).
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Figure 11. Evolution of the concentration of nitrogen species during the reaction (initial NO3-N concentration: 30 mg/L, pH 6, Dosage: 5 g/L, ZB12 sample: ZB12-500).
Figure 11. Evolution of the concentration of nitrogen species during the reaction (initial NO3-N concentration: 30 mg/L, pH 6, Dosage: 5 g/L, ZB12 sample: ZB12-500).
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Figure 12. Evolution of DO, pH, and ORP during reaction (initial NO3-N concentration: 30 mg/L, pH 6, Dosage: 5 g/L, ZB12 sample: ZB12-500).
Figure 12. Evolution of DO, pH, and ORP during reaction (initial NO3-N concentration: 30 mg/L, pH 6, Dosage: 5 g/L, ZB12 sample: ZB12-500).
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Figure 13. Reaction mechanism of ZB12 composite with NO3-N.
Figure 13. Reaction mechanism of ZB12 composite with NO3-N.
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Table 1. Water quality data of groundwater and experimental configuration water.
Table 1. Water quality data of groundwater and experimental configuration water.
Water TypeMonitoring IndexMaximum ValueMinimum ValueAverage ValueExperimental Value
Note: K+, Na+, Mg2+, Ca2+, Cl, SO42−, HCO3, NO3 are in mg/L.
Table 2. Models of experimental instruments.
Table 2. Models of experimental instruments.
Scanning electron microscopeVega-3XMU, Tescan, Czech
Surface area BET analyzerMicromeritics ASAP 2020, Norcross, USA
X-ray diffractometerTTR-III, Rigaku, Japan
Fourier transform infrared
Equinox 55, Bruker Banner Lane, Germany
X-ray photoelectron spectroscopyEscalab 250, Thermo Fisher Scientifific, USA
DO meterHQ30d, Hach, USA
pH meterMP220, Mettler Toledo, Switzerland
spectrophotometerUV-1802, BeifenRuili, China
GC-TCDGC-8A, Shimadzu, Japan
Table 3. SSA of ZB12-500 before, during, and after the reaction.
Table 3. SSA of ZB12-500 before, during, and after the reaction.
ZB12-500 SamplesSSA/(m2/g)
before reaction10.69
during the reaction63.47
after reaction83.32
Table 4. The relative content of Fe valence states measured with XPS.
Table 4. The relative content of Fe valence states measured with XPS.
Fe Valence StatesZB12-500BZB12-500DZB12-500A
Table 5. Kinetic parameters of NO3-N removal by ZB12-500.
Table 5. Kinetic parameters of NO3-N removal by ZB12-500.
SampleMaterialFirst-OrderSecond-OrderTwo-Compartment First-Order
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Liu, S.; Han, X.; Li, S.; Xuan, W.; Wei, A. Stimulating Nitrate Removal with Significant Conversion to Nitrogen Gas Using Biochar-Based Nanoscale Zerovalent Iron Composites. Water 2022, 14, 2877.

AMA Style

Liu S, Han X, Li S, Xuan W, Wei A. Stimulating Nitrate Removal with Significant Conversion to Nitrogen Gas Using Biochar-Based Nanoscale Zerovalent Iron Composites. Water. 2022; 14(18):2877.

Chicago/Turabian Style

Liu, Siyuan, Xiao Han, Shaopeng Li, Wendi Xuan, and Anlei Wei. 2022. "Stimulating Nitrate Removal with Significant Conversion to Nitrogen Gas Using Biochar-Based Nanoscale Zerovalent Iron Composites" Water 14, no. 18: 2877.

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