Study on the Adsorption Behavior and Mechanism of Nitrate Nitrogen in Sewage by Aminated Reed Straw

Abstract

Nitrate pollution in water bodies has become a global environmental problem, and its excessive presence not only leads to eutrophication of water bodies but also threatens human health through the drinking water pathway. Therefore, it is urgent to develop new adsorbents with high adsorption capacity, good selectivity and excellent regeneration performance to solve the problem of nitrate pollution. In this study, reed straw (RS), trimethylamine-modified reed straw (MRS) and triethylamine-modified reed straw (ERS) were prepared by quaternary amination modification for nitrate removal. The adsorption performance, desorption performance, adsorption characteristics under disturbed environment and dynamic adsorption performance were investigated experimentally, and the adsorption mechanism was analyzed by various characterization means. The adsorption performance followed the order ERS (12.25 mg·g⁻¹) > MRS > RS, demonstrating that quaternary amination modification, particularly with triethylamine, significantly enhanced the NO₃⁻-N adsorption capacity. ERS exhibited excellent regeneration stability (over 80% after nine cycles) and high selectivity towards NO₃⁻-N in the presence of competing anions (Cl⁻, SO₄²⁻, humic acid). In the dynamic adsorption experiment, ERS had a breakthrough time of 290 min at a packing height of 3.3 cm, with an adsorption capacity of 10.74 mg·g⁻¹ and good adaptability to flow rate. In the actual wastewater application, the initial NO₃⁻-N removal rate was over 95%, the dynamic desorption rate reached 99.2% and the peak nitrate concentration of the desorbed solution reached 27 times of the initial value, confirming its high efficiency regeneration and enrichment ability. The study shows that the amine-modified reed straw adsorbent has a good potential for application and provides a new way for wastewater treatment plants to solve the problem of nitrate removal 12.25 mg·g⁻¹.

1. Introduction

Accelerated urbanization, excessive application of nitrogen fertilizers in agricultural production leading to nitrate nitrogen entering the water body through surface runoff as well as incomplete treatment of domestic sewage discharge and industrial nitrogenous waste discharge, make nitrate nitrogen pollution prominent. The traditional biological denitrification process is poorly adapted to low-temperature, low carbon and nitrogen ratio sewage. This results in nitrate nitrogen concentration in effluent water often exceeding 15 mg/L. Under anaerobic conditions, nitrate nitrogen can be converted to nitrite, and the resulting nitrosamine compounds formed through combination with amines are strongly carcinogenic. High concentrations of nitrate nitrogen can cause algal blooms, leading to eutrophication of the water body. The self-purification ability synchronously declines, and other organisms in the food chain are therefore affected, thus adversely affecting human health [1]. According to the World Health Organization’s (WHO) regulations, the concentration of nitrate nitrogen in drinking water should not exceed 50 mg/L [2].

Nitrate exists in the water body in ionic state. The traditional physical treatment is difficult to separate, so the current method of removing nitrate nitrogen from the water body are mainly biological and chemical methods, electrodialysis and reverse osmosis. However, the biological method has a long treatment cycle, a long time to achieve the desired removal effect and the temperature has a greater impact, as at lower temperatures, the removal efficiency is significantly reduced [3]. The chemical reduction method has a higher removal rate, but it can easily cause secondary pollution and is easy to be affected by pH [4]. As the electrodialysis method needs to be applied to the operation of the electric field, the cost is higher [5]. Although reverse osmosis has a high nitrate nitrogen removal rate, it requires high-pressure equipment, which makes the operation cost high, the removal of nitrate nitrogen in water is not selective and it is not the most ideal technology to remove nitrate nitrogen [6].

The abundant biomass resources and low-cost advantages of straw make it widely used in the preparation of adsorbent materials, and its porous structure and surface-active sites provide a natural basis for adsorption. Reed in Baiyangdian accounts for 60% of the area covered by aquatic vegetation in the precipitation area, with an estimated annual biomass of about 135,000 t, based on regional productivity. Exploring the high-value utilization of RS plays an important role in the whole Baiyangdian ecosystem. The branching amine group-modified straw adsorbent showed the advantages of fast adsorption rate, good selectivity and excellent regeneration performance, making it a new type of adsorbent.

Reed straw (RS) is a promising adsorbent precursor due to its abundant hydroxyl groups (from cellulose/hemicellulose) for chemical modification, natural porous structure for mass transfer and high mechanical strength. Selecting the stem ensures better structural integrity than other plant parts. Trimethylamine (TMA) and triethylamine (TEA) were chosen to introduce quaternary ammonium groups for anion exchange. The differing alkyl chain lengths (methyl for TMA, ethyl for TEA) may affect hydrophobicity, steric hindrance and site accessibility, influencing adsorption performance and selectivity. While reed-based materials have been studied for pollutant removal (e.g., heavy metals, dyes), research on quaternary ammonium-modified RS for nitrate nitrogen (NO₃⁻-N) removal, especially comparing TMA and TEA modifications, is limited. This study aims to address this gap.

The aim of this paper is to develop and evaluate two RS-based amination-modified adsorbents, MRS and ERS, and to use them for efficient removal of nitrate (NO₃⁻-N) from water. In the study, the materials were firstly prepared by a three-step method of washing–alkalization–amidation, and their structures and modification mechanisms were characterized using Fourier transform infrared spectroscopy (FTIR) (PerkinElmer Co., Ltd., Shelton, CT, USA), scanning electron microscopy (SEM) (TESCAN (China) Co., Ltd., Shanghai, China), X-ray photoelectron spectroscopy (XPS) (Thermo Fisher Scientific (China) Co., Ltd., Shanghai, China) and Brunauer–Emmett–Teller (BET) (Kubo X1000, Beaud Electronic Technology Co., Ltd., Beijing, China)analysis, among others [7]. The batch adsorption performance (effect of dosage, pH, concentration, temperature, adsorption time), interference resistance (SO₄²⁻, Cl⁻, HA) and cyclic regeneration performance of the two materials were compared in detail. The dynamic adsorption behavior was further evaluated by fixed-bed experiments and analyzed in conjunction with Thomas and Yoon–Nelson models. Finally, the application potential of ERS and the feasibility of resourcefulness of dynamic desorption were verified in simulated and actual urban wastewater scenarios, elucidating that the hydrophobicity, spatial site resistance and selectivity advantages of ERS, due to its long-chain alkyl groups are the key to its performance beyond MRS.

2. Materials and Methods

2.1. Source of Material

The reed straw (RS) used in this study was collected from the Baiyangdian Lake area in Xiong’an New Area, China. It was cleaned, dried and ground to pass through a 60-mesh sieve (particle size < 250 μm) for further modification.

All chemicals used were of analytical grade and were used as received without further purification. The reagents used include: potassium nitrate (KNO₃) (Kemiou Chemical Reagent Co., Ltd., Tianjin, China), potassium chloride (KCl)) (Kemiou Chemical Reagent Co., Ltd., Tianjin, China), zinc sulfate heptahydrate (ZnSO₄·7H₂O)) (Kemiou Chemical Reagent Co., Ltd., Tianjin, China), potassium bicarbonate (KHCO₃)) (Kemiou Chemical Reagent Co., Ltd., Tianjin, China), potassium carbonate (K₂CO₃)) (Kemiou Chemical Reagent Co., Ltd., Tianjin, China), anhydrous ethanol (C₂H₆O)) (Kemiou Chemical Reagent Co., Ltd., Tianjin, China), potassium sulfate (K₂SO₄)) (Kemiou Chemical Reagent Co., Ltd., Tianjin, China) Sodium hydroxide (NaOH) (Huihang Chemical Technology Co., Ltd., Tianjin, China), N,N-Dimethylformamide (HCON(CH₃)₂) (Damao Chemical Reagent Factory, Tianjin, China), epichlorohydrin (C₃H₅ClO) (Damao Chemical Reagent Factory, Tianjin, China), sulfuric acid (H₂SO₄) (Damao Chemical Reagent Factory, Tianjin, China),hydrochloric acid (HCl) (Damao Chemical Reagent Factory, Tianjin, China), Sodium bicarbonate (NaHCO₃) (McLean Biochemical Technology Co., Ltd., Shanghai, China), L-tryptophan (C₁₁H₁₂N₂O₂) (McLean Biochemical Technology Co., Ltd., Shanghai, China), sodium humate (C₉H₈Na₂O₄) (McLean Biochemical Technology Co., Ltd., Shanghai, China), triethylamine (C₆H₁₅N) (McLean Biochemical Technology Co., Ltd., Shanghai, China),trimethylamine (C₃H₉N) (McLean Biochemical Technology Co., Ltd., Shanghai, China).

Real sewerage samples were collected from the inlet of a municipal wastewater treatment plant. The samples were filtered (0.45 μm) and spiked with KNO₃ to adjust nitrate nitrogen concentrations for adsorption experiments.

2.2. Material Preparation

Three-stage preparation of modified aminated RS: In pretreatment, the dried RS was washed and dried with distilled water several times after crushing and passing through a 60-mesh sieve; in alkalization, 8.0 g of the pretreated sample was added into a 20% sodium hydroxide solution with magnetic stirring at 25 °C for 2 h. This treatment can effectively remove the impurities on the surface of the material and expose the reactive hydroxyl groups in the cellulose. At the stage of amination modification, 4.0 g of alkaline samples were placed in a 1000 mL three-necked flask, and 40 mL of N,N-dimethylformamide (DMF) and 20 mL of epichlorohydrin (ECH) were added in a thermostatic water bath at 80 °C for 30 min to carry out a ring-opening cycloaddition with hydroxyls on the RS surface and amination with trimethylamine (TMA) and triethylamine (TEA), respectively. After the reaction, the product was washed with anhydrous ethanol and deionized water alternately until the filtrate was colorless and transparent, without irritating odor and without bubbles. Finally, the products were dried at 80 °C in an electrically heated constant temperature blast drying oven (DHG-9023A, Jinghong Experimental Equipment Co., Ltd., Shanghai, China) until constant weight.

Taking the ERS modification process as an example, the specific mechanism is shown in Figure 1:

Scanning electron microscopy (SEM, TESCAN MIRA, TESCAN Co., Ltd., Shanghai, China), Fourier transform infrared spectroscopy (FT-IR, Spectrum 3, PerkinElmer Co., Ltd., Shanghai, China), X-ray photoelectron spectroscopy (XPS, ESCALAB 250Xi), specific surface area and pore size distribution analyzer (BET, Kubo X1000, Piotek Electronic Technology Co., Ltd., Beijing, China) were used to characterize the surface micro-morphology, functional group and elemental composition valence and pore structure of the adsorbent before and after modification and adsorption.

The technology roadmap for this paper is shown in Figure 2.

2.3. Performance Measurement

The adsorption reaction system was placed in a digital thermostatic water bath oscillator (SHY-2A, Changzhou Putian Instrument Manufacturing Co., Ltd., Changzhou, China), and the supernatant was filtered through a 0.45 μm nylon membrane at 25 °C at a constant speed of 180 rpm for 30 min, and the concentration of nitrate nitrogen (NO₃⁻-N) was determined according to the “Determination of Nitrate in Water Quality by Ultraviolet Spectrophotometry” [8] (HJ/T 346-2007). All experiments were repeated in groups of three to reduce the error. The dissolved organic matter (DOM) in municipal wastewater was determined by a three-dimensional fluorescence spectrophotometer (F-7100, Hitachi Scientific Instruments Ltd., Beijing, China), and three-dimensional fluorescence spectroscopy (EEM) combined with fluorescence region integration (FRI) and parallel factor analysis (PARAFAC) was used to efficiently analyze the DOM fractions and to record the excitation-emission matrix (EEM) data [9,10].

Batch adsorption experiments were conducted to evaluate the nitrate removal performance of the adsorbent. In a typical procedure, a predetermined amount of adsorbent was added to 20–50 mL of nitrate solution in glass flasks. The mixtures were shaken in a temperature-controlled shaker at 180 rpm and 25 °C for a specified contact time. After adsorption, the suspensions were centrifuged (if necessary), and the supernatants were filtered through 0.45 μm membrane filters to remove residual adsorbent particles. The filtrates were analyzed for residual nitrate concentration using a UV-Vis spectrophotometer at 220,275 nm, with corrections for nitrite interference when applicable. All experiments were performed in duplicate or triplicate; the average values are reported.

2.3.1. Batch Adsorption and Desorption Performance

A multi-gradient experimental design was used by adding 2–18 g/L of aminated RS and 50 mL of nitrate nitrogen standard solution (initial concentration range of 10–90 mg/L with gradient interval of 10 mg/L) into conical flasks, respectively. The initial pH of the reaction system was adjusted to 4.0–9.0, and the conical flasks were sequentially removed at 0.5, 1, 2, 3, 4, 5, 10, 15, 30, 45, 60, 90, 120, and 180 min for measurement, and the NO₃⁻-N removal and adsorption capacity of the materials were calculated (Equations (1) and (2)) with the formulae, respectively:

$$\mathrm{R} = {\displaystyle \frac{{\mathrm{C}}_{0 }{- \mathrm{C}}_{\mathrm{t}}}{{\mathrm{C}}_0}} \times 100\%,$$

(1)

where R is the removal rate of NO₃⁻-N by adsorbent at adsorption equilibrium, %; C₀ is the initial concentration of NO₃⁻-N, mg/L; and Ct is remaining concentration of NO₃⁻-N at adsorption equilibrium, mg/L.

$$\begin{array}{c}{\mathrm{q}}_{\mathrm{e}}={\displaystyle \frac{\left({\mathrm{C}}_0{ - \mathrm{C}}_{\mathrm{t}}\right)}{\mathrm{m}}} \times \mathrm{V},\end{array}$$

(2)

where q_e is the adsorption capacity of the adsorbent for NO₃⁻-N at adsorption equilibrium, mg·g⁻¹; V is the volume of NO₃⁻-N solution, L; and m is the mass of the adsorbent, g.

Determination of static desorption performance: In total, 0.4 g of NO₃⁻-N saturated and adsorbed MRS (MRS-N) and ERS (ERS-N) were placed in a 100 mL conical flask, and 50 mL of desorbents (KCl, NaOH, KHCO₃, K₂CO₃, K₂SO₄) of 0.1 mol/L were added. Then, the desorbents were oscillated for 30 min under the conditions of 25 °C and 180 rpm. The residual concentration of NO₃⁻-N in the desorbed solution was determined, and the desorption rate was calculated according to the Formula (3) to screen the optimal desorbent. Desorption rate calculation formula:

$$\mathsf{\zeta }\left(\%\right) = {\displaystyle \frac{{\mathrm{C}}_{\mathrm{t}} \times \mathrm{V}}{{\mathrm{q}}_{\mathrm{e}} \times \mathrm{m}}} \times 100\%,$$

(3)

where $\mathsf{\zeta }$ is the desorption efficiency of the adsorbent, %; and C_t is the NO₃⁻-N concentration in the solution after desorption and regeneration, mg/L.

Determination of batch adsorption performance under disturbed environment: In total, 0.4 g of MRS and ERS were added to 50 mL of municipal wastewater treatment plant primary sedimentation tank influent samples, respectively, with 180 rpm oscillation for 60 min. Competitive adsorption was used to investigate the effects of SO₄²⁻, Cl⁻ and humic acid (HA) on the adsorption performance of the adsorbents. A simulated wastewater (pH = 7.0) with an initial concentration of NO₃⁻-N of 20 mg/L was prepared, and a mixed solution of 50 mL was prepared by adding gradient concentrations of SO₄²⁻/Cl⁻ (0–350 mg/L) and HA (0–40 mg/L), respectively, and then 0.4 g of MRS and ERS were added after transferring them to a 100 mL conical flask.

2.3.2. Dynamic Adsorption Performance

An adsorption column apparatus made of glass (20 mm in diameter and 150 mm in height) was used to carry out the experiments. In order to remove the impurities from the material, the quartz sand was soaked in 4 mol/L hydrochloric acid for 24 h and then rinsed and dried repeatedly with deionised water. The adsorption column was made of 1 cm thick quartz cotton as the bottom support layer, 5 g of pre-treated quartz sand was chosen to be evenly spread, and then a certain amount of adsorption material was evenly covered. Finally, a layer of 15 g of quartz sand was spread to fill the column. This layered design not only prevents the adsorbent material from being washed away by the water flow but also allows the solution to flow evenly through the entire adsorbent medium. At the same time, it ensures the stability and uniformity of the adsorbent bed. The nitrate solution enters from the bottom of the column and is controlled by a peristaltic pump to flow the solution through the adsorbent bed at a specific flow rate.

Through the fixed bed dynamic adsorption experiment, the filling amount (1 g, 2 g and 3 g, corresponding to the packing height of 1.1 cm, 2.2 cm and 3.3 cm) was explored. Under the condition of NO₃⁻ concentration 20 mg/L and flow rate 5 mL/min, the water samples were taken regularly to measure the concentration, and the influence of packing height on the adsorption effect was evaluated. The concentration of NO₃⁻ was fixed at 50 mg/L, the amount of adsorbent was 2 g, the flow rate was adjusted (3 mL/min, 5 mL/min, 7 mL/min) and the samples were sampled at an interval of 10 min to investigate the effect of flow rate on the adsorption effect. Thomas and Yoon–Nelson models were used to fit the dynamic adsorption behavior (Equations (4) and (5)).

Thomas model:

$$\begin{array}{c}{\displaystyle \frac{{\mathrm{C}}_{\mathrm{t}}}{{\mathrm{C}}_0}} = {\displaystyle \frac{1}{1 + \exp \left(\frac{{\mathrm{K}}_{\mathrm{TH}}{\mathrm{q}}_{\mathrm{e}}\mathrm{m}}{\mathrm{Q}}{- \mathrm{K}}_{\mathrm{TH}}{\mathrm{C}}_0\mathrm{t}\right)}},\end{array}$$

(4)

where t is the running time of the adsorption column, min; and K_TH is the Thomas model parameter, min⁻¹.

Yoon–Nelson model:

$$\begin{array}{c}{\displaystyle \frac{{\mathrm{C}}_{\mathrm{t}}}{{\mathrm{C}}_0}} = {\displaystyle \frac{{\exp (\mathrm{K}}_{\mathrm{YN}}{\mathrm{t} - \mathsf{\tau }\mathrm{K}}_{\mathrm{YN}})}{{1 + \exp (\mathrm{K}}_{\mathrm{YN}}{\mathrm{t} - \mathsf{\tau }\mathrm{K}}_{\mathrm{YN}})}},\end{array}$$

(5)

where $\mathsf{\tau }$ is the time required for the ratio of the concentration of the effluent of the adsorption column to the concentration of the influent to reach 0.5, min; and K_YN is the Yoon–Nelson model constant, min⁻¹.

3. Results and Discussion

3.1. Material Characterization

SEM analysis shows that the unmodified RS surface has a smooth, dense, regular rod-like structure (Figure 3a,d). After modification by TMA and TEA, the surface of the material became rough, loose and swollen (Figure 3b,e), which was attributed to the breakage of cellulose microcrystals and formation of nanopores due to ring opening of epoxy groups and grafting of quaternary ammonium groups in the amination reaction [11].The fold depth of ERS (ca. 200 nm) was significantly larger than that of MRS (ca. 80 nm), which might originate from the stronger spatial potential resistance due to longer molecular chains of TEA. After nitrate adsorption, uniformly distributed submicron particles appeared on the surfaces of MRS-N and ERS-N (Figure 3c,f, shown by arrows), and these attachments were not seen to be present before adsorption, which provided a preliminary basis for nitrate adsorption on the material surfaces.

BET characterization is shown in Figure 4: the specific surface area of MRS decreased from 0.6481 m²/g to 0.2292 m²/g, a decrease of 65%, which was attributed to the short-chain quaternary ammonium groups occupying the surface of the pore channel and blocking part of the mesopores [12]. The pore size distribution showed a decrease in the total pore volume of MRS, and the hysteresis ring was changed from the H3 slit pore to the H2-type ink-bottle pore with the change in pore geometry. ERS showed a 30-fold increase in specific surface area to 19.3177 m²/g, which was attributed to the three-dimensional effect of long-chain alkyl groups to increase the cellulose layer spacing (from about 0.8 nm to 1.6 nm), as well as the hydrolysis of lignin to generate new mesopores during alkali washing. The H2-type lag ring was dominated by ink-bottle pores with a centralized pore size distribution averaging 2.26 nm, which is conducive to the diffusion of ions. Comparing the changes in specific surface area before and after adsorption, MRS-N decreased from 0.2292 m²/g to 0.18 m²/g, with the disappearance of micropores and the blockage of mesopore entrance; ERS-N decreased from 19.3177 m²/g to 1.0865 m²/g, with a decrease of 94.4%; and the hysteresis ring H2 was reversed to H3, with the loss of pore connectivity. The severe collapse of ERS proved that its high adsorption capacity (12.25 mg·g⁻¹) depends on pore space.

The FT-IR results are shown in Figure 5, with typical cellulose absorption peaks at O-H stretching vibration at 3340 cm⁻¹, C-H stretching vibration at 2920 cm⁻¹ and C-O-C glycosidic bond at 1029 cm⁻¹ for both materials, which confirms the preservation of the main structure of the biomass. The intensity of the O-H peaks of MRS/ERS is weakened and red-shifted (3340 shifted to 3325 cm⁻¹). This was attributed to the disruption of the hydrogen bonding network by alkali treatment and the depletion of -OH to form ether bonds (C-O-C) by the epoxy group ring-opening reaction. The disappearance of the lignin ester bond C=O at 1734 cm⁻¹ and the aryl ether C-O-C peaks at 1230 cm⁻¹ indicated that alkali treatment effectively hydrolyzed the lignin–carbohydrate complexes (LCCs) [13], exposing the cellulosic substrate. The new peaks at 1470 cm⁻¹ were ascribed. The new peak at 1470 cm⁻¹ was attributed to the C-H bending vibration of the quaternary ammonium group, whose intensity was higher in ERS. The area of ERS peak increased by 32%, confirming that the quaternary ammonium density was enhanced by the long-chain alkyl group [14]. The C-N⁺ stretching vibration of ERS peak at 1635 cm⁻¹ confirmed the successful grafting of triethylamine, and the NO₃⁻ was specifically captured by the site through electrostatic gravitational force, where the binding energy of 45 kJ/mol. 1635 cm⁻¹ and 1470 cm⁻¹ peaks synergistically formed a “dual” peak. The 1635 cm⁻¹ and 1470 cm⁻¹ peaks synergistically formed a “double fingerprint signal”, which proved that the quaternary ammonium group was the active center of NO₃⁻ adsorption. The higher intensity of the 1470 cm⁻¹ peak of ERS was positively correlated with its adsorption capacity (12.25 mg·g⁻¹) (R² = 0.94), and the long alkyl chain also enhanced the hydrophobicity, which inhibited the competition for adsorption by water molecules [15].

The evolution of chemical valence states of surface elements of RS, MRS and ERS before and after adsorption of nitrate was revealed by XPS analysis (Figure 6). The characteristic peak of the base C element located at 284.8 eV was used as a charge correction benchmark to ensure the accuracy of the subsequent analysis. The characteristic peaks of O and N were observed at 530.08 and 400.08 eV, respectively (the N peaks of MRS and ERS were shifted). The 200 eV Cl peaks only appeared in the MRS and ERS samples, which indicated that Cl was successfully introduced during the modification process, and the disappearance of the Cl peaks after adsorption confirmed that the adsorption of nitrate was dominated by the exchange of Cl⁻-NO₃⁻ ions, which verified the dominant mechanism of chemical adsorption. The dominant mechanism of chemisorption was verified.

The adsorption mechanism was further elucidated by fine spectral analysis of N elements (Figure 7). The N peak of pristine RS (400.08 eV) originated from the nitrogen-containing fraction of straw itself (Figure 7a,d), whereas the N peaks of MRS and ERS were shifted to 402.3 and 403 eV, respectively (Figure 7b,e), which were attributed to the altered electronic environments of branched amine groups, where the larger steric hindrance effect of TEA and the difference in nitrogen lone electron pairs led to their higher binding energies. After the adsorption of nitrate, new peaks appeared in the interval of 406.5–406.7 eV between MRS-N and ERS-N, which were attributed to the N signal of NO₃⁻ (Figure 7c,f). Furthermore, combined with semi-quantitative analysis, they showed that the peak intensity of ERS-N was significantly higher than that of MRS-N, which indicated that ERS had a superior nitrate adsorption performance.

3.2. Static Performance Analysis

3.2.1. Batch Adsorption Performance

Increasing the dosage can provide more adsorption sites and promote the coordination or electrostatic interaction between nitrate and functional groups on the adsorbent surface [16]. As shown in Figure 8a, the removal rate of RS was only 5%, and the adsorption capacity was less than 0.1 mg·g⁻¹. The removal rates of NO₃⁻-N by MRS and ERS increased and then slowed down with the increase in the dosage: when the dosage was in the range of 8–10 g/L, the removal rate of MRS was about 80%, and the removal rate of ERS reached 80%~85%, and the growth rate was significantly reduced thereafter, and the adsorption capacity decreased with the increase in the dosage; after the increase in the dosage >8 g/L, the excess adsorbent led to the aggregation of particles, and the specific surface area of adsorbents was about 1.2%. After >8 g/L, the excess adsorbent led to particle aggregation and decreased specific surface area utilization. While the NO₃⁻-N concentration in the solution decreased, the mass transfer driving force was weakened, which was manifested as a decrease in adsorption capacity. Combining the removal efficiency and adsorption capacity, both adsorbents reached the performance equilibrium at the dosage of 8 g/L. When the dosage exceeded 10 g/L, the removal rate of the two adsorbents increased by less than 5%, indicating that the active sites were almost completely occupied and the economic benefits of increasing the dosage were reduced.

As shown in Figure 8b, the adsorption performance of the two types of adsorbents showed a single peak curve with the change in pH value. Under acidic conditions, NO₃⁻ in the solution was surrounded by a large amount of free H⁺, forming a ‘hydrated hydrogen ion shell’, which hindered the direct contact between NO₃⁻ and the active site of the adsorbent. The high concentration of H⁺ increased the ionic strength of the solution, compressed the diffusion double layer on the surface of the adsorbent, and weakened the electrostatic adsorption force. In addition, the increase in H⁺ concentration would destroy the structure of the material to a certain extent, so that the removal rate decreased rapidly [17]. Under neutral conditions, the protonation and deprotonation reactions reached equilibrium, and the utilization rate of adsorption sites and mass transfer efficiency were optimal [18]. The competitive adsorption caused by the increase in OH⁻ concentration under alkaline conditions affects the adsorption effect of the adsorbent. Because the pH of urban sewage is between 6 and 9, it has little effect on the performance of the adsorbent, and there is no need to adjust the pH in the subsequent experiments.

As shown in Figure 8c, the removal efficiency of MRS and ERS decreased with the increase in initial nitrate concentration, while the equilibrium adsorption capacity increased. At low initial concentrations (<18 mg/L), the equilibrium adsorption capacities of MRS and ERS are comparable, indicating limited enhancement from amination under dilute conditions. When the initial concentration exceeded 60 mg/L, the equilibrium adsorption capacity of ERS began to be significantly higher than that of MRS, and the gap between the two gradually expanded with the increase in concentration. In the low concentration range, the two adsorbents showed similar adsorption capacity. Under high concentration conditions, ERS exhibits better adsorption performance. This indicates that, in addition to the mass transfer driving force enhanced by the increase in initial concentration, the characteristics of the ERS material itself play a more critical role in improving the adsorption performance at high concentrations.

The adsorption kinetics study showed that the adsorption process of MRS and ERS showed significant two-stage characteristics (Figure 9a,b) [19,20]. In the rapid adsorption stage (0–30 min), both adsorbents showed high adsorption rates at three concentrations (the removal rates reached 91.8% and 89.8% within 30 min), which was attributed to the dense active sites on the surface of the material and the strong mass transfer driving force provided by the high ion concentration. At the equilibrium stage (>30 min), the diffusion resistance in the particles increased, the concentration of the solution decreased, the driving force of mass transfer decreased, and the adsorption rate decreased significantly. At this stage, the two materials showed different adsorption behaviors. MRS tended to be stable after 30 min, and ERS reached the adsorption equilibrium state at about 45 min.

It can be seen from

Table 1. Kinetic model fitting parameters for nitrate adsorption.

Materials	Concentration (mg·L−1)	Proposed Primary Adsorption Kinetic Model			Proposed Secondary Adsorption Kinetic Model
		qe (mg·g−1)	K1 (min−1)	R2	qe (mg·g−1)	K2 (g.(mg·min)−1)	R2
MRS	20	2.24	2.21	0.997	2.45	2.34	0.995
	50	4.79	2.66	0.988	4.85	3.95	0.998
	100	10.36	3.35	0.954	9.93	4.23	0.921
ERS	20	2.31	2.43	0.994	2.34	2.55	0.998
	50	5.67	3.27	0.996	5.77	3.65	0.999
	100	10.97	3.98	0.953	11.01	4.08	0.977

that the higher fitting degree (R² = 0.921–0.999) and the rate constant k₂ of the pseudo-second-order model increased significantly with the concentration, confirming that chemical adsorption was dominant. Combined with the adsorption equilibrium time of the two materials and the calculated deviation of q_e (pseudo-second-order model <5%), 30–45 min is recommended as the optimal adsorption contact time.

From Figure 10 and

Table 2. Adsorption isotherm model fitting parameters.

Material	Temperature (k)	Langmuir			Freundlich
		qmax (mg·g−1)	KL	R2	n	KF	R2
MRS	288	8.54	0.139	0.972	2.29	1.69	0.997
	298	8.43	0.143	0.975	2.43	1.78	0.995
	308	8.76	0.122	0.976	2.31	1.72	0.997
ERS	288	12.25	0.071	0.966	2.02	1.56	0.977
	298	10.75	0.102	0.938	2.11	1.59	0.984
	308	11.56	0.082	0.984	2.35	1.76	0.995

, it can be seen that the goodness of fit of the Freundlich model for MRS and ERS (R² = 0.977–0.997) is significantly higher than that of the Langmuir model (R² = 0.938–0.984) [21,22], indicating that the adsorption of NO₃⁻ by the two materials is a multilayered non-homogeneous process [23]. The Freundlich model parameters “n” were all between 2 and 10, indicating strong adsorption affinity (MRS: 2.29–2.43; ERS: 2.02–2.35) [24].

As shown in

Table 3. Adsorption thermodynamic model fitting parameters.

Material	Temperature (K)	∆G (kJ/mol)	∆H (kJ/mol)	∆S (J/mol)
MRS	288.15	−0.34	−4.07	−12.96
	298.15	−0.21
	308.15	−0.08
ERS	288.15	−1.47	−11.12	−33.53
	298.15	−1.12
	308.15	−0.80

, the adsorption process of NO₃⁻-N by both MRS and ERS is a spontaneous (ΔG < 0), exothermic (ΔH < 0) and entropy-decreasing (ΔS < 0) chemical adsorption [25]. ΔG increases (negatively decreases) with temperature, indicating that low temperatures are more favorable for adsorption (e.g., ΔG = −1.47 kJ/mol for ERS at 288.15 K, the adsorption drive is the strongest). The entropy reduction effect arises from the ordered transition of NO₃⁻ from the solution free state to immobilization on the adsorbent surface, resulting in a decrease in the degree of freedom of the system. ERS outperforms MRS due to stronger enthalpy drive (ΔH = −11.12 kJ/mol) and significant entropy reduction (ΔS = −33.53 J/mol-K), and both achieve the optimal adsorption efficiency at low temperature (<298 K).

3.2.2. Desorption Properties of Materials

Experiments were carried out to desorb MRS-N and ERS-N using various reagents such as KCl (Figure 11a). The results showed that KCl, NaOH and KHCO₃ exhibited higher desorption efficiencies, and this difference originated from the selective adsorption properties of the quaternary aminated materials for anions. Considering that NaOH may damage the structure of the material, reagents containing CO₃²⁻ and SO₄²⁻ will interfere with the active sites due to competition for adsorption and nitrate exists in the form of potassium salt in the other original adsorption experiments, 0.1 mol/L KCl was finally identified as the optimal desorption reagent.

Through nine adsorption-regeneration cycles (Figure 11b), the removal rate and adsorption capacity of MRS and ERS decreased slowly with the increase in cycles, indicating that both of them had good regeneration performance, but the regeneration stability of ERS was significantly better than that of MRS. This is because the short methyl chain structure of MRS results in a small steric hindrance, and competitive anions (such as Cl⁻) are more likely to be retained at the active site during the regeneration process, reducing the adsorption efficiency of subsequent cycles.

3.3. Interference Studyperformance Analysis

3.3.1. Inorganic Component

The concentration of SO₄²⁻ in urban wastewater (usually 200–500 mg/L) was significantly higher than that of NO₃⁻ (about 30–50 mg/L), and in order to investigate the interference mechanism, the construction of a SO₄²⁻ and NO₃⁻ binary system (Figure 12a). The results showed that the adsorption of NO₃⁻ by both adsorbents was inhibited with increasing SO₄²⁻ concentration, but the interference resistance of ERS was significantly better than that of MRS. When SO₄²⁻ = 350 mg/L, the MRS removal rate decreased by 52.1%, adsorption capacity decreased by 59.4%, the residual removal rate 35.6% and the adsorption capacity was 0.89 mg·g⁻¹. The ERS removal rate decreased by 17.7%, adsorption capacity decreased by 19.5%, residual removal rate by 72.3% (2.03 times of MRS) and the adsorption capacity was 1.82 mg·g⁻¹ (2.04 times of MRS). This was due to the difference in ion selectivity, as the high hydration energy of SO₄²⁻ (−1100 kJ/mol) made it easier to be adsorbed in hydrophilic environments, whereas the hydrophobic surface of ERS preferentially bound ions with low hydration energies (NO₃⁻: −314 kJ/mol) [26]; and the spatial site-blocking effect of the long alkyl chains of ERS. The long alkyl chains of ERS produce a spatial site-blocking effect, which selectively prevents SO₄²⁻ with a large hydration radius from approaching its active adsorption site and thus adsorbs NO₃⁻ more efficiently in complex aqueous environments [27]. The hydrophobic chains of ERS weaken the electrostatic attraction, which reduces the affinity of ERS for multivalent anions (SO₄²⁻) [28]. Therefore, under high sulfate interference (≥350 mg/L), ERS exhibited more stable nitrate adsorption performance due to the ion-sieving ability conferred by the hydrophobic modification.

Cl⁻ is commonly found in municipal wastewater (concentration of 100–300 mg/L), and its physical and chemical properties are similar to NO₃⁻ (monovalent anion), which leads to the difficulty of selective adsorption, for which a binary system of Cl⁻-NO₃⁻ was constructed (Figure 12b). The adsorption performance of MRS and ERS on NO₃⁻ tended to decrease with increasing Cl⁻ concentration, but the interference intensity was only 1/2 of that of SO₄²⁻, because the monovalent property of Cl⁻ reduced the electrostatic attraction strength with amine groups (weaker than that of SO₄²⁻), confirming that anion valence is one of the reasons affecting the competitive adsorption. In addition, low hydration energy ions are more susceptible to dehydration adsorption (the hydration energy of NO₃⁻ is −314 kJ/mol, which is smaller in absolute value than that of Cl⁻ (−381 kJ/mol)). ERS hydrophobic long chains further amplify the effect of the difference in hydration energies, and their spatial site resistance excludes Cl⁻ (hydration radius of 0.332 nm > NO₃⁻ of 0.26 nm). ERS maintained 73% removal rate (9% higher than MRS) under Cl⁻ interference, and the results of this experiment are consistent with the literature [29].

3.3.2. Organic Component

To assess the impact of dissolved organic matter (DOM) on nitrate adsorption by aminated reed straw (ERS) in real wastewater, three-dimensional fluorescence (EEM) coupled with parallel factor analysis (PARAFAC) was applied (Figure 12c) [30,31]. PARAFAC resolved three key fluorescent components: humic-like (C1, Ex/Em ≈ 250(350)/450 nm), fulvic-like (C2, Ex/Em≈250/420 nm) and protein-like (C3, Ex/Em ≈ 275/340 nm) substances (Figure 12d,e).

After ERS adsorption, the fluorescence intensity of C1 decreased by ~18% and C2 by ~12%, indicating effective removal of humic and fulvic acids [32]. This is likely due to hydrophobic interactions enhanced by the long alkyl chains on ERS and electrostatic attraction between the positively charged quaternary ammonium groups and the negatively charged humic substances. In contrast, the intensity of C3 decreased only slightly (~5%), suggesting minimal adsorption of protein-like substances, possibly due to their smaller size, higher hydrophilicity and weaker affinity with the adsorbent.

These results demonstrate that ERS can simultaneously adsorb nitrate nitrogen and partially remove humic substances from wastewater. Although DOM components, particularly humic acids, may compete with nitrate for adsorption sites, ERS maintained a high nitrate removal efficiency (initial rate >95%). This highlights the strong selectivity of the quaternary ammonium-functionalized surface for nitrate ions over most DOM fractions. The findings confirm ERS’s robust anti-interference capability and significant potential for practical application in treating complex urban wastewater [33,34].

3.4. Dynamic Adsorption Performance

3.4.1. Effect of Operational Parameters and Model Fitting

The adsorption penetration point was set at 5% of the effluent concentration up to the influent concentration, and the adsorption saturation point was set at 95% of the effluent concentration up to the influent concentration. Under this criterion, as shown in Figure 13a, the adsorption performance of ERS was significantly better than that of MRS: the penetration time (70 min) and saturation time (150 min) of ERS were 7 times and 3.3 times of that of MRS, respectively, at 1 g filler volume. The adsorption capacity reached 11.06 mg·g⁻¹, which was 3.5 times of that of MRS (3.18 mg·g⁻¹). As shown in

Table 4. Effect of filling volume and initial flow rate on dynamic adsorption performance.

Material	Filler Volume (mL)	Initial Flow Rate (mL/min)	Penetration Point (min)	Saturation Point (min)	Total Adsorption (g)	Adsorption Capacity (mg·g−1)	Removal Rate (%)
MRS	1	5	10	45	3.18	3.18	70.67
	2	5	45	75	6.05	3.03	80.67
	3	5	95	125	11.15	3.72	89.20
	2	3	25	90	3.91	3.91	43.44
	2	5	20	55	3.99	3.99	72.55
	2	7	10	45	4.03	4.03	89.56
ERS	1	5	70	150	11.06	11.06	73.73
	2	5	130	260	21.92	10.96	84.31
	3	5	290	370	32.23	10.74	87.11
	2	3	60	280	11.57	11.57	41.32
	2	5	55	180	12.2	12.2	67.78
	2	7	40	150	13.49	13.49	89.93

, the increase in filling amount can prolong the penetration and saturation time, but the unit adsorption efficiency decreases: when the filling amount increases from 1 g to 3 g, the adsorption capacity of MRS decreases by 38%, which is attributed to the clogging of mesopores leading to the surge of mass transfer resistance. ERS decreases by only 27%, which can be attributed to the high specific surface area and hierarchical pore structure to alleviate the interference of competing ions and maintain the high efficiency of mass transfer, which indicates that ERS has a stability advantage at high filling amounts. The decrease in ERS was only 27%, which was attributed to the high specific surface area and hierarchical pore structure to alleviate the interference of competing ions and maintain the efficient mass transfer.

The initial flow rate is the core regulating parameter of the dynamic adsorption process. The differences in the dynamic adsorption performance between MRS and ERS filled with 2 g on simulated water samples at different flow rates (3 mL/min, 5 mL/min and 7 mL/min) were investigated. As shown in Figure 13b, the flow rate was increased from 3 mL/min to 7 mL/min: the penetration point of MRS was shortened from 25 min to 10 min, with a decrease of 60%; the penetration point of ERS was shortened from 60 min to 40 min, with a decrease of 33%, which was attributed to the fact that the high flow rate weakened the resistance of the liquid film and accelerated the diffusion of pollutants into the surface of the material, but it also shortened the effective contact time between adsorbate and the active site, resulting in an earlier penetration. As shown in Table 4, the removal rate of MRS increased from 43.44% to 89.56% with the increase in flow rate, which confirms that its mass transfer is limited, and the forced convection at high flow rate can alleviate the pore clogging of MRS due to ionic competition. The adsorption capacity of ERS fluctuated narrowly from 11.57 to 13.49 mg·g⁻¹ (RSD < 8%), which was attributed to the high selectivity of ERS and the hierarchical pore structure.

The Thomas model assumes that the adsorption process is controlled by surface reaction rather than limited by internal diffusion [35], and as shown in

Table 5. Fit parameter.

Material	Filler Volume (g)	Inlet Flow Rate (mL·min−1)	Thomas Model			Yoon–Nelson Models
			KTh (mL·min−1·mg−1)	qe (mg·g−1)	R2	KYN (min−1)	τ (min)	R2	τ/KYN
MRS	1	3	0.421	4.48	0.967	0.084	67.73	0.967	806.309
	1	5	0.796	4.04	0.983	0.159	40.35	0.983	253.774
	1	7	0.831	4.02	0.971	0.166	28.71	0.971	172.952
	2	7	0.588	4.21	0.81	0.179	58.46	0.916	327.598
	3	7	0.317	6.14	0.863	0.086	123.32	0.954	1433.953
ERS	1	3	0.143	11.62	0.968	0.025	188.72	0.952	7248.8
	1	5	0.241	12.09	0.964	0.048	120.85	0.966	2517.708
	1	7	0.381	15.79	0.976	0.054	107.8	0.939	1996.296
	2	7	0.243	12.6	0.907	0.046	178.7	0.942	3884.783
	3	7	0.365	15.17	0.984	0.065	323.51	0.951	4977.077

, the equilibrium adsorption capacity (q_e) of ERS is 3.9 times higher than that of MRS under the same conditions (1 g, 7 mL/min). Although the initial adsorption rate of MRS was faster than that of ERS due to its predominantly surface adsorption, the deep mass transfer was hindered, which ultimately led to a decrease in adsorption capacity.

The Yoon–Nelson model—a semi-empirical model based on the adsorption rate being proportional to the product of the probability of the adsorbate not being captured and the probability of the adsorption site remaining—and the correlation (R²) for both adsorbents exceeded 0.9 confirm that the model can describe the dynamic adsorption behavior of MRS and ERS [36,37]. As shown in Table 5, the K_YN of MRS was significantly higher than that of ERS, indicating the rapidity of the initial adsorption of MRS, which was usually accompanied by the ease of saturation of the sites due to low selectivity; the τ value of ERS was 3.8 times higher than that of MRS, which was attributed to the high selectivity of the triethylamine moiety and the hierarchical pore structure. When the flow rate was changed from 3 mL/min to 7 mL/min, the τ value of ERS decreased by 42.9%, whereas that of MRS plummeted by 57.6%, highlighting the resistance of ERS to flow rate fluctuations; the τ value of ERS linearly increased by 200% (K_YN fluctuation < 20%) when the filling volume was increased from 1 g to 3 g, whereas the K_YN decayed by 48% in the case of MRS, exposing the defect of internal diffusion limitation. Therefore, ERS is suitable for long-lasting adsorption processes, and MRS is only recommended for short-duration high-load scenarios, and ERS at 3 mL/min and τ/K_YN = 7548 is the optimal energy-efficiency balance.

3.4.2. Practical Applicability

The actual water sample was pretreated by 0.45 μm microporous membrane, and the average concentration of nitrate nitrogen was 12.18 mg/L by double-beam ultraviolet spectrophotometer (TU-1810, Puxi General Instrument Co., Ltd., Beijing, China). The nitrogen removal efficiency of ERS in actual sewage was further investigated. As shown in Figure 14a, when ERS ran to 40 min, the system approached the dynamic adsorption breakthrough point, with an average removal rate of more than 97.5%. The adsorption capacity began to show a gradient attenuation at 120 min and reached the adsorption saturation state at about 178 min. After 190 min of continuous operation, the total adsorption capacity of the material reached 12.6 mg, the unit adsorption capacity was 6.3 mg/g, and the average removal rate before adsorption saturation was always higher than 65%. The adsorption efficiency of ERS before 120 min is basically consistent with the simulation data, indicating that ERS has application potential in actual complex water bodies, but its pretreatment process needs to be optimized to reduce the impact of environmental interference and achieve its long-term stability.

As shown in Figure 14b, the continuous flow of the desorption reagent (1 mol/L KCl) disrupted the solid–liquid boundary layer, forced convection to break through the mass transfer limit, and accelerated the diffusion of reagent into the interior of the material, resulting in a significantly higher dynamic desorption rate (99.2%) than the static desorption rate (85.6% ± 2.1%). At the same time, the nitrate concentration of the desorbed solution reached a peak of 553.64 mg/L (27 times the initial concentration of 20.5 mg/L), which can be directly used for nitrate resource recovery and reduce the subsequent treatment cost. In addition, dynamic desorption can remove the desorbed ions in time to avoid the reverse reaction, a mechanism that is limited in static systems. Therefore, dynamic desorption provides an efficient pathway for adsorbent regeneration and pollutant resourcing.

4. Conclusions

In this study, trimethylamine-based adsorbent (MRS) and triethylamine-based adsorbent (ERS) were successfully prepared from quaternary amine-modified reed straw (RS), and their removal efficacy and mechanism of nitrate removal from water were systematically evaluated and the potential for practical application was verified, the findings are consistent with previous studies on amine-functionalized biosorbent [38,39,40], strengthening the scientific basis of the adsorption process. The batch adsorption capacity of ERS reached 12.25 mg·g⁻¹, which was significantly higher than that of MRS. From the simulated two-stage kinetic model with R² > 0.92, it is clear that both of them were dominated by chemisorption. In complex water bodies, ERS has excellent selectivity and interference study properties due to the preferential adsorption of low hydration energy ions on hydrophobic surfaces and steric shielding effect. In the dynamic column experiments, ERS had a breakthrough time of 290 min at a packing volume of 3 g and a flow rate of 5 mL/min, with an adsorption capacity of 10.74 mg·g⁻¹, which was three times higher than that of MRS, and the τ value of ERS was 2.6 times higher than that of MRS in the Yoon–Nelson model fitting, which indicated that ERS had a strong resistance to flow rate fluctuations, and it was suitable for long-lasting operation. After nine cycles of regeneration, the adsorption capacity retention rate of ERS was more than 80%, which was better than that of MRS, and its stability is superior to that of the remaining biosorbents, which usually degrade after a few cycles [41]. The peak nitrate concentration of desorbed liquid was 27 times of the initial value, which was feasible for resource recovery. Leaching tests were conducted on fresh ERS under neutral pH conditions, and no detectable release of organic carbon or nitrogenous compounds was observed, indicating that the modified reed straw does not contribute to secondary pollution during use. After exhaustion and multiple regeneration cycles, the spent adsorbent can be safely disposed of via incineration or landfill, as it is primarily composed of carbonized biomass. Alternatively, it may be repurposed as a soil amendment or biofuel feedstock, aligning with circular economy principles.

In summary, ERS offers an efficient, renewable, and low-cost solution for the deep denitrification of wastewater with a low carbon-to-nitrogen ratio, owing to the hydrophobicity of its long alkyl chains and the spatial site-barrier effect, which collectively enhance ionic selectivity, interference resistance, and mass-transfer efficiency. This makes ERS particularly suitable for treatment scenarios characterized by high sulfate or organic backgrounds, providing a technologically innovative and engineering-adaptable approach for the green transformation of wastewater treatment plants. It should be noted, however, that regeneration of spent ERS generates a concentrated nitrate solution with a high chloride content, which must be further treated—via processes such as biological denitrification, membrane concentration, or catalytic reduction—before safe discharge. While this study primarily focuses on the development and performance of the adsorbent, future work will aim to integrate downstream treatment strategies to enable complete and sustainable management of the regeneration effluent.