Unveiling the Potential of Novel Struvite–Humic Acid Composite Extracted from Anaerobic Digestate for Adsorption and Reduction of Chromium

Abstract

A novel struvite–humic acid composite (S–HA) was derived from an anaerobic digestate and evaluated for the adsorption and reduction of chromium [Cr (VI)] in this study. The results indicated that the struvite–humic acid composite (S–HA) contains higher contents of oxygen-containing and aromatic functional groups (47.05% and 34.13%, respectively) and a higher specific surface area (19.3 m² g⁻¹). These special characteristics of S–HA contributed to its higher adsorption capacity (207.69 mg g⁻¹ and 254.47 mg g⁻¹ for pseudo-first and second-order kinetic models, respectively) for chromium. Furthermore, XPS analysis showed that a portion of the bonded Cr (VI) was reduced to Cr (III) by carboxyl and hydroxyl functional groups, which oxidized and changed into ketone and phenol functional groups. Based on the findings, it was concluded that the phosphate–humic acid composite has an outstanding chromium adsorptive and reduction capacity. However, more research is needed to fully understand the potential of the struvite–humic acid composite for chromium adsorption and reduction.

1. Introduction

Chromium (Cr) is a highly hazardous heavy metal that occurs naturally and is widely used in industrial processes [1]. It has been reported that about 11% and 14% of the contaminated sites in the U.S. and Japan are polluted by Cr, respectively [2]. Likewise, Cr-contaminated land accounts for more than 5% of China’s total potentially heavy metal-contaminated lands [3]. Due to the high toxicity, accumulation, and refractory degradation of Cr, the existence of Cr threatens the soil ecosystem and adversely affects human health via the food chain [4]. Therefore, it is necessary to identify and implement effective strategies to remediate Cr-contaminated soil.

Over the years, various approaches have been developed to remediate Cr-contaminated soils, such as capping, stabilization, and extraction technologies [5,6]. Among them, stabilization has been considered the most effective and preferable method to remediate Cr-contaminated soil due to its high efficiency and low cost. Generally, the stabilizing agents reduce the mobility and bioavailability of heavy metals in the soil by inducing various physicochemical reactions [5,7]. To date, multiple materials have been evaluated for the in-situ fixation of heavy metals, including phosphate-based materials (e.g., ammonium phosphate, apatite, and hydroxyapatite) and organic matters (e.g., compost, manure, and humic substances) [8,9,10]. Generally, these materials can reduce heavy metals’ bioavailability and leaching potential by mediating adsorption, complexation, precipitation, and redox processes.

Phosphate-containing materials, as commonly-used remediation agents, can form stable compounds with a wide range of heavy metals to stabilize metal amendments in contaminated soil [11]. Struvite (MgNH₄PO₄.6H₂O), as a critical fraction of these phosphate-containing materials, has been reported to interact with various heavy metals, such as copper (Cu), cadmium (Cd), and zinc (Zn) in contaminated media [12]. Tang et al. (2019) reported that the abundance of amino groups, hydroxyl groups, and phosphate groups in struvite could coordinate or adhere to heavy metal hydroxides through hydrophilic interactions [8]. Moreover, previous studies have confirmed that hydrogen bonding and electrostatic forces are the main driving forces for inducing interactions between heavy metal particles and struvite crystals [13,14].

Furthermore, humic acid (HA), a macromolecular organic compound widely existing in nature, contains complex structures and functional groups and shows an excellent potential in removing heavy metal pollutants [15,16,17]. Compared with other adsorbents, HA contains abundant active functional groups, including carboxyl, hydroxyl, and aromatic, which contribute to the adsorption of various heavy metals [18]. Moreover, HA removed Cr by forming stable complexes through electrostatic adsorption and complexation reactions [19,20]. Likewise, the functional groups (e.g., quinone groups) in the HA served as an electron-transferring shuttle, providing electron donors continuously to mediate biotic or abiotic redox reactions, thus driving the conversion of Cr (VI) to Cr (III) [10]. Moreover, as an essential component of organic matter in the soil, HA does not cause secondary pollution to metal-contaminated soil and can be readily taken up by plants to promote plant growth, thereby promoting the efficiency of the soil remediation process [21].

Although the ability of commercial phosphate-containing materials and HA for heavy metal remediation is unquestionable, there is no research on the combined use of phosphate-containing materials and HA for Cr-containing soil remediation. Moreover, the commercial phosphate-containing materials and HA are mainly found in non-renewable resources, i.e., phosphate rock and lignite, respectively [22,23]. Therefore, finding a renewable, low-cost alternative resource for obtaining phosphate-containing materials and HA is necessary. Anaerobic digestate, a by-product of the anaerobic digestion process, is rich in various macronutrients (e.g., P, N, K, Mg) and active molecular components (e.g., humic acid) [24,25]. Our previous studies have successfully obtained phosphate-containing materials (struvite) and HA from anaerobic digestate via the alkaline–acid alternate chemical extraction method and conducted preliminary investigations on their structures and functions [21,26]. However, the effect of the simultaneous extraction of humic acid and struvite from the anaerobic digestate and the potential of this composite (phosphate–humic acid) for the remediation of Cr-contaminated soil is still unclear.

To address this knowledge gap, a struvite and humic acid composite was first extracted from anaerobic digestate via an alternate alkaline–acid chemical extraction method. Then, the structural and functional properties of struvite, humic acid, and their composite were analysed. Finally, the potential of the novel struvite and humic acid composite for the adsorption and reduction of Cr (VI) was explored.

2. Results and Discussion

2.1. Characteristics of S–HA

The SEM micrographs showed that the surface structure of raw HA is relatively smooth (Figure 1a). In comparison, some cylindrical and interlaced rhombohedral-shaped struvite crystals were observed in the S–HA (Figure 1c). Likewise, the EDS results perceived that N, O, and C were the major elements of raw HA, with lower P contents (0.91%), while the higher content of P (5.27%) and Mg (2.81%) in the S–HA material indicated the possible formation of phosphorus-rich minerals on the surface of S–HA (Figure 1d–f and

Table 1. The elemental composition of struvite, humic acid, and struvite–humic acid composite.

Element	Struvite (%)	Humic Acid (%)	Struvite–Humic Acid (%)
C	17.27	58.67	48.72
N	3.36	8.90	8.96
O	47.60	20.19	21.60
Na	1.52	1.37	3.48
Mg	12.44	/	2.81
P	15.70	0.91	5.27
K	0.59	0.59	0.37
Si	0.50	0.41	0.29
S	0.16	1.61	0.70
Ca	0.34	0.14	/
Cl	0.52	7.20	7.80

) [11]. XPS analysis was used to further confirm the chemical bonds of the functional groups in the S–HA (Figure S1). The energy peaks of C 1s, N 1s, O 1s, P 2p, and Mg 1s appeared in the S–HA, indicating the existence of C, N, O, P, and Mg in the materials [27]. The high-resolution XPS spectra of the Mg 1s and P 2p regions were further analysed by peak fitting (Figure S1b,c). The results showed that the primary forms of the bonds in the S–HA were P-O (134.2 eV), P-Mg-O (133 eV), P-C-H (135 eV), Mg-OH (1302.5 eV), and Mg-Cl (1305 eV), indicating the incorporation of struvite into the HA structure via Mg/P-O chemical bonds.

The HA and S–HA functional groups were determined by the FTIR spectra (Figure 2). According to previous studies, the oxygen-containing functional groups (900–1400 cm⁻¹, including C-O, -COOH, and -OH) [21,28], amide groups (1400–1600 cm⁻¹, -NH) [29], and aromatic functional groups (1600–1850 cm⁻¹, including C=O, C-C, and ketones) [30] were correlated with the adsorption efficiency of heavy metals. Thus, peak fitting at these wavelengths was performed for semi-quantitative analysis. The results showed that the relative contents of oxygen-containing functional groups (900–1400 cm⁻¹), amide groups (1400–1600 cm⁻¹), and aromatic functional groups (1600–1850 cm⁻¹) in the HA structure were 33.54%, 27.63%, and 38.82%, respectively. Contrasted with HA, S–HA contained a higher proportion of oxygen-containing groups (47.05%) and aromatic functional groups (34.13%). However, the content of amide groups (18.82%) in S–HA was lower than in HA. Generally, the abundance of oxygen-containing and aromatic functional groups could be attributed to the complexing between heavy metals and HA [31]. The higher content of aromatic functional groups is significantly correlated with the stability of HA [32]. These findings indicate that S–HA has a higher heavy metal complexation capacity and stability value.

Moreover, the specific surface area and average pore diameter of HA and S–HA were estimated based on the N₂ adsorption/desorption isotherms combined with Brunauer–Emmett–Teller (BET) multilayer adsorption theory (Figure S2). The results showed a significant difference between HA and S–HA (p < 0.05). The specific surface area and average diameter of the pores of HA were 5.1 m² g⁻¹ and 10.25 nm, respectively, which was lower than for S–HA (19.3 m² g⁻¹ and 4.3 nm, respectively). These findings indicated that the struvite coated on the surface of HA affected the physical properties and pore structure of the S–HA, thus increasing the complexing capacity of S–HA [33].

2.2. Cr (VI) Sorption by S–HA

HA is a macromolecular organic compound with fluorescent characteristics. When HA forms stable complexes with heavy metal ions, the fluorescence intensity will decrease—called fluorescence quenching [11,34]. Generally, the degree of fluorescence quenching has a significantly positive correlation with the adsorption efficiency of heavy metals [35]. Thus, to explore the Cr (VI) complexing ability of S–HA, the EEM spectra were firstly used to confirm the evolution of the fluorescence of S–HA during the complexing process (Figure 3a,b). An individual fluorophore-centred humic-like compound (Ex/Em = 325 nm/416 nm) was observed in the EEM spectra of S–HA without Cr (VI) [11,36]. With the addition of Cr (VI), the fluorescence intensity of the fluorophore was decreased, indicating the complexing of S–HA and Cr (VI) by forming the S–HA–Cr complex [35,37].

To further investigate the efficient complexation of S–HA and Cr (VI), Cr (VI) adsorption experiments were performed (Figure 3c,d). As shown in Figure 3c, S–HA had a high adsorption ability for Cr (VI) in the initial stage. The adsorption capacity quickly reached around 174 mg g⁻¹ within 12 h, much faster and higher than that of struvite and HA. Previous studies have also reported that the large surface area and functional groups contributed to the rapid adsorption kinetics [38], consistent with the results of BET and FTIR in the current study. Cr (VI) adsorption kinetic data from HA, struvite, and S–HA were fitted using the pseudo-first order and pseudo-second order kinetic models, respectively (Figure 3d and

Table 2. The Cr (VI) kinetic data and best-fit parameters of pseudo first- and second-order kinetic models.

Kinetic Models	Parameters	Adsorbents
		Struvite	Humic Acid	Struvite-Coated Humic Acid
Pseudo first order kinetic model	Qe (mg g−1)	58.0224	172.2368	205.6920
	k1 (h−1)	0.1374	0.0766	0.1147
	R2	0.9465	0.9501	0.9726
	ΔQ (%)	12.3	9.41	6.36
Pseudo second order kinetic model	Qe (mg g−1)	69.7625	229.6012	254.4748
	k2 (mg g−1 h−1)	0.0002	0.0003	0.0004
	R2	0.9263	0.9642	0.9409
	ΔQ (%)	14.6	8.32	11.5

). The results showed that the correlation coefficients (R²) of the pseudo-first-order kinetic models (0.94, 0.95, and 0.97 for struvite, HA, and S–HA, respectively) were higher than those of the pseudo-second-order models (0.92, 0.94, and 0.96, respectively), indicating that the pseudo-first-order kinetic model could provide a better interpretation of the Cr (VI) adsorption behaviour on struvite, HA, and S–HA caused by the strong chemical forces at the solid/solution interface [39,40]. Likewise, the values of the kinetic adsorption constants (K₁) were in the order of struvite (0.1347) > S–HA (0.1147) > HA (0.0766), and the kinetic adsorption constants (K₂) values were in the order of S–HA (0.0004) > HA (0.0003) > struvite (0.0002). These results indicated that only physical forces are involved in the adsorption of Cr (VI) to struvite. In contrast, physicochemical forces contributed to the adsorption of Cr (VI) onto HA and S–HA [9]. Moreover, the calculated maximum adsorption capacity (Qe) of S–HA was 254.47 mg g⁻¹, which is higher than that of struvite (69.76 mg g⁻¹) and HA (229.60 mg g⁻¹). Contrary to the findings of the current study, Jang et al. (2020) employed chemically functionalized amorphous and mesoporous silica nanoparticles and observed the maximum adsorption of Cr (VI) to be 34.0 and 42.2 mg g⁻¹ [41]. Similarly, Bayuo et al. (2019) observed that the maximum chromium adsorption on ground nut shell was 5.56 mg g⁻¹ [42]. Recently, Bachmann et al. (2022) reported an adsorption of 0.32 to 403.23 mg Cr (VI) per gram of lignocellulosic waste adsorbent [43]. Therefore, it is concluded that the struvite–humic acid composite derived herein is an effective adsorbent for chromium. The fitness of the pseudo-first-order model can also be verified by calculating the percentage deviation between the experimental amount of adsorbed Cr (VI) (Qe_exp) and the calculated amount (Qe_cal) [44,45]:

$$\mathsf{\Delta }\mathrm{Q}(\%)=\frac{{\mathrm{Qe}}_{\exp } - {\mathrm{Qe}}_{\mathrm{cal}}}{{\mathrm{Qe}}_{\mathrm{cal}}}\times 100$$

(1)

2.3. Cr (VI) Reduction by S–HA

To explore the Cr (VI) reduction by struvite, HA, and S–HA, the Cr speciation distribution in these materials was determined by XPS analysis (Figure 4). Three peaks were observed in the XPS spectra of struvite at 583.5 eV, 581 eV, and 580 eV, which contributed to Cr (VI) (Figure 4a) [19,38]. Compared with the struvite, additional peaks appeared in the XPS spectra of HA and S–HA at 585 eV corresponding to Cr (III), indicating that adsorbed Cr (VI) on the HA and S–HA surfaces was reduced into Cr (III) [46]. Moreover, the Cr 2p peaks areas of HA and S–HA were further determined by semi-quantitative analysis. It was observed that the absolute intensity of Cr (III) in HA was higher than that in S–HA. This means that HA exhibits a stronger Cr (VI) reduction than S–HA. Cr (III) concentration on HA and S–HA was also determined (Figure S3). The results showed that Cr (III) concentration on HA and S–HA increased with reaction time. At the end of the reaction, the content of Cr (III) on HA (59.96 mg g⁻¹) was higher than that on S–HA (57.46 mg g⁻¹). These findings are consistent with our XPS analysis results.

2.4. Mechanisms of Cr (VI) Sorption and Reduction

The interaction between Cr (VI) and HA extracted from soil followed an adsorption-reduction mechanism reported previously [20,47]. Briefly, Cr (VI) is first bound to the surface of humic acid through a complexation/ion-exchange pathway. The bound Cr (VI) is then reduced into Cr (III) via the functional groups of HA. Zhang et al. (2018) confirmed through the 2D COS analysis method that during the adsorption-reduction process, the phenol, hydroxyl, and carboxyl functional groups might be the main groups involved in the adsorption and reduction of Cr (VI) by HA. However, other studies have reported the carboxyl and aromatic functional groups of HA as the groups contributing to the adsorption and reduction of Cr (VI), respectively. This contradiction in the reported results might be due to the different sources of HA under analysis.

XPS analysis was used to analyse the species of functional groups involved in the adsorption and reduction of Cr (VI) (Figure 5). The primary forms of the bonds in the HA and S–HA were C-O (534.1 eV), C=O (535.2 eV), -COOH (533.2 eV), and -OH (531.5 eV), confirmed through peak-fitting analysis [27]. When comparing changes in the functional groups before and after adsorption of Cr (VI), it was observed that the intensity of -COOH and -OH, representing carboxyl and hydroxyl, decreased significantly during the reaction of HA/S–HA with Cr (VI). Meanwhile, the intensity of C-O and C=O, representing phenol and quinone, increased throughout the reaction process. These findings indicate that the carboxyl and hydroxyl groups might be the main functional groups mediating the reaction of HA/S–HA with Cr (VI).

Moreover, pseudo-first-order and second-order kinetic models were utilized for the time-dynamic evolution analysis of Cr (VI) adsorption by S–HA (Figure 4). The correlation coefficients (R²) were all less than 0.96, which indicated that the reaction between Cr (VI) and S–HA might involve processes such as physical adsorption, chemical adsorption, and redox processes [46]. Therefore, based on the above information, a possible reaction process of Cr (VI) with S–HA is proposed in this study (Figure 6). Firstly, the Cr (VI) in solution reaches the surface of S–HA via liquid film diffusion and is then adsorbed onto the S–HA surface by electrostatic forces. Then, under the action of active functional groups such as carboxyl and hydroxyl groups, Cr (VI) establishes a stable complex with S–HA. As a result, the carboxyl and hydroxyl functional groups reduce a portion of the bonded Cr (VI) to Cr (III). Meanwhile, the carboxyl and hydroxyl functional groups are oxidized and converted into ketone and phenol functional groups. Finally, the Cr (III) may stabilize the S–HA surface via the complexation and adsorption process.

3. Materials and Methods

3.1. Materials

Chicken manure anaerobic digestate was collected from the Deqingyuan Biogas plant (Beijing, China). The plant was operating under mesophilic conditions. The digestate was taken and separated to lessen the impact of suspended solids and then kept at 4 °C until usage. The characteristics of the anaerobic digestate are presented in

Table 3. Characteristic of anaerobic digestate in this study.

Parameter	Unit	Value
pH	/	7.90
Phosphate	mg L−1	230
Total solids	g L−1	3.20
Potassium	mg L−1	1574
Calcium	mg L−1	26.8
Magnesium	mg L−1	12.6
Ammonia nitrogen	mg L−1	4700

. Analytical grade potassium dichromate with a purity ≥99.8%, magnesium chloride hexahydrate with a purity ≥99.8%, sodium dihydrogen phosphate with a purity ≥99.8%, and ammonium chloride with a purity ≥99.8% were purchased from Beijing Chemical Works, Beijing, China.

3.2. Extraction of Phosphate and Humic Acid Composite

This study used the alkaline–acid alternate chemical extraction method to obtain the struvite and humic acid composite (S–HA). Our previous published studies’ methods were followed for HA extraction and struvite formation from anaerobic digestate [21,26]. Briefly, the digestate was first mixed with 0.1M NaOH at a 1:10 ratio and shaken at 200 rpm for 24 h. Then, the mixed solution was acidified to pH 2 with 6 M HCl and left for 12 h. Subsequently, MgCl₂·6H₂O, NaH₂PO₄, and NH₄Cl were added to the mixed solution to maintain an equimolar ratio of Mg²⁺, NH₄⁺, and PO₄⁻³ (1:1:1) and shaken under 100 rpm at room temperature for 24 h. Finally, the solution was filtered through a 0.45 μm membrane and centrifuged at 5000 rpm for 10 min to obtain the S–HA composite.

3.3. Characterization Methods

The surface structure and elemental compositions of struvite, HA, and S–HA were analysed using a scanning electron microscope (SEM, SU8020, Tokyo, Japan) and energy-dispersive X-ray spectroscopy (EDS). The chemical bonds and functional groups were analysed by X-ray photoelectron spectroscopy (XPS, Thermoescalab 250Xi, Waltham, MA, USA) and Fourier-transform infrared spectra (FTIR, Nicolet IS10, Akron, OH, USA), respectively [9,48]. Moreover, the N₂ adsorption/desorption isotherms of struvite, HA, and S–HA were measured on Quanta-chrome NOVA2000e instruments. N adsorption-desorption tests at 77 K were used to determine the specific surface area of each sample employing a Micromeritics ASAP2020, Norcross, GA, USA. The pore characteristics (including total pore volume, mesopore volume, and micropore volume) were calculated following the method described in Zhang et al. (2020) [27].

3.4. Adsorption and Reduction Experiments

The Cr (VI) adsorption and reduction experiments were conducted following the method described in Zou et al. (2021) [38]. Briefly, 100 mL of Cr (VI) solution (30 mg/L) was mixed with 0.5 g sorbent (struvite, HA, and S–HA, respectively) in a 200 mL conical flask at 25 °C. The pH was adjusted to 5 using HCl or NaOH solution. Then, the mixtures were shaken at 200 rpm and equilibrated for 48 h. After the reaction, an excitation–emission matrix (EEM) spectra analysis was conducted using a fluorescence spectrophotometer [21]. The emission and excitation wavelengths were set from 250 to 600 nm with 5 nm and 3 nm increments.

To analyse the Cr adsorption and reduction efficiency, samples were taken at different intervals (0, 2, 4, 6, 8, 10, 12, 24, 36, 48 h) to measure the concentrations of Cr (VI) and Cr (III), respectively. Briefly, the samples were filtered and analysed by inductively coupled plasma-atomic emission spectrometry (ICP-AES, Perkin-Elmer Plasma 3200, OH, USA) and the 1, 5-diphenylcarbazide colorimetric method, respectively [38]. Then, pseudo-first-order kinetic (Equation (2)) and pseudo-second-order kinetic (Equation (3)) models were used to analyse the adsorption behaviour [49]:

$$\ln \left({\mathrm{Q}}_{\mathrm{e}}-{\mathrm{Q}}_{\mathrm{t}}\right)=\ln {\mathrm{Q}}_{\mathrm{e}}-k_1·t$$

(2)

$$\frac{t}{{\mathrm{Q}}_{\mathrm{t}}}=\frac{1}{k_2{\mathrm{Q}}_{\mathrm{e}}^2}+\frac{t}{{\mathrm{Q}}_{\mathrm{e}}}$$

(3)

where Q_t (mg g⁻¹) represents the adsorption capacity of the adsorbent at time t; Q_e (mg g⁻¹) represents the adsorption equilibrium of the adsorbent at time t; and k₁ (h⁻¹), and k₂ (g mg⁻¹ h⁻¹) are the kinetic adsorption constants for the pseudo-first-order and pseudo-second-order kinetic models, respectively.

4. Conclusions

In this study, a novel struvite–humic acid (S–HA) complex material was obtained from anaerobic digestate via an alkaline–acid alternate chemical extraction method and applied as an adsorbent for the adsorption and reduction of Cr (VI). The characterization of S–HA indicated that it contains higher contents of oxygen-containing and aromatic functional groups, and a higher specific surface area. Moreover, the Cr (VI) adsorption and reduction experiments showed that the maximum pseudo-first-order and second-order kinetic adsorption capacities were 207.69 mg g⁻¹ and 254.47 mg g⁻¹, respectively, which is higher than those of humic acid and struvite. The XPS results revealed that carboxyl and hydroxyl functional groups play the dominant role in the adsorption and reduction of Cr (VI). Our results will advance understanding of the interaction mechanisms between Cr (VI) and humic acid-based materials in the soil and support further research toward strategies that combine Cr (VI) remediation and nutrient recovery from wastewater.