Ternary surface complexation promotes simultaneous removal of arsenic and uranium by TiO₂

Significance

Groundwater containing naturally occurring arsenic and uranium is widely used for drinking purpose in numerous countries, presenting a significant public health concern. However, there is a lack of economically efficient technology for the simultaneous removal of arsenic and uranium. Our research uncovered a mechanism involving the formation of a ternary surface complex [Ti–U(VI)–As(V)] on the TiO₂ surface confirmed through density functional theory calculations. This provides valuable insights into the adsorption processes of arsenic and uranium, contributing to the development of strategies for their simultaneous removal. Our study offers a promising solution for addressing the simultaneous removal of arsenic and uranium from groundwater, thereby reducing the human health risks associated with using this groundwater as a drinking water source.

Abstract

The co-occurrence of arsenic and uranium in groundwater has been found in many countries, posing a significant challenge to human health. Here, we have demonstrated the efficient simultaneous removal of arsenic and uranyl-carbonate complexes from groundwater using {001}-TiO₂. Surprisingly, the presence of U(VI) greatly enhanced the adsorption of As(V) on {001}-TiO₂, while As(V) had a negligible impact on U(VI) adsorption. Through in situ ATR-FTIR spectroscopy, we uncovered a mechanism involving the formation of a ternary surface complex [Ti–U(VI)–As(V)] on the surface of {001}-TiO₂. This ternary surface complex formed through the substitution of CO₃^2- from uranyl coordination sites. Furthermore, the adsorbed As(V) and U(VI) can be easily recovered using a sodium hydroxide solution, and {001}-TiO₂ can be used repeatedly. Our findings offer a promising solution for the simultaneous removal of As(V) and U(VI) from groundwater and provide valuable insights into the mechanisms involved in their removal.

Arsenic in groundwater is a serious public health concern due to its high toxicity and widespread distribution. It is estimated that over 140 million people in at least 70 countries have been exposed to arsenic-contaminated groundwater (1), leading to severe skin alterations and organ cancer. (2) Recently, uranium, a radioactive element, has been found in arsenic-containing groundwater in over 10 countries (SI Appendix, Table S1), where rhyolitic rocks are considered to be the primary source of both arsenic and uranium. (3–5) In these countries, groundwater containing arsenic and uranium is widely used for drinking purposes (SI Appendix, Table S1). In addition to the chemical toxicity, uranium also possesses radioactivity. Chronic exposure to arsenic and uranium in drinking water may further aggravate the threat to human health. The World Health Organization (WHO) and the US Environmental Protection Agency (EPA) have established maximum concentrations for arsenic and uranium in drinking water at 10 and 30 μg/L, respectively. (6, 7) However, the coexistence of arsenic and uranium presents a challenge for simultaneous and efficient removal of them from groundwater.

In groundwater, arsenic and uranium commonly coexist in the oxidized forms, As(V) and U(VI), respectively, under oxidizing conditions. (8–10) U(VI) tends to form more mobile uranyl-carbonate complexes with CO₃^2- in groundwater. (7) However, most studies have primarily focused on removing either As(V) or uranyl-carbonate complexes alone from groundwater, with little attention given to their simultaneous removal. Adsorption is a simple and effective method for removing multiple contaminants due to its extensive adaptability, convenience, and low cost. Traditional arsenic adsorbents such as goethite, ferrihydrite, bentonites, kaolinite, and clinoptilolite can also be used to remove uranyl ions. However, the adsorption of U(VI) is significantly decreased in neutral and weakly alkaline solutions due to the formation of uranyl-carbonate complexes. (11–13) Recently, TiO₂ has been proven capable of capturing uranyl-carbonate complexes because of its high affinity for U(VI). (14) Additionally, TiO₂ is considered one of the optimal materials for efficient As(V) adsorption. (15, 16) TiO₂ possessed multiple exposed facets that have been shown to highly influence its adsorption performance for heavy metals, organic matter, and gases. (17–19) The various exposed facets of TiO₂ exhibit specific morphologies and unique surface properties, providing more active sites for adsorption. Therefore, we believe that TiO₂ is a favored material for the simultaneous removal of As(V) and uranyl-carbonate complexes from groundwater.

The coexistence of multiple contaminants has been demonstrated to have a significant impact on adsorption performance of adsorbents. (20–23) The coadsorption behaviors of As(V) and U(VI) on TiO₂ can be influenced by alterations in the surface properties of TiO₂, such as changes in surface functional groups and surface charges. On the one hand, the strong adsorption affinity of hydroxyl groups on TiO₂ for both As(V) and U(VI) creates the potential for competition between these contaminants. On the other hand, As(V) and U(VI) are primarily present as negatively charged or neutral species [H₂AsO₄^-, HAsO₄^2-, Ca₂UO₂(CO₃)₃ and CaUO₂(CO₃)₃^2-], (24, 25) resulting in the accumulation of negative charges on the surface of TiO₂, which enhances electrostatic repulsion. These factors likely contribute to a decrease in the coadsorption performance of As(V) and U(VI) on TiO₂.

However, our results demonstrated that {001}-TiO₂ can simultaneously and efficiently remove As(V) and U(VI). Surprisingly, the presence of U(VI) greatly enhanced the adsorption of As(V), while As(V) had a negligible impact on U(VI) adsorption. We employed in situ ATR-FTIR spectroscopy to explore the coadsorption mechanism of As(V) and U(VI). Our findings provide the evidence of a ternary surface complex [Ti-U(VI)-As(V)] formation on {001}-TiO₂, which is the primary mechanism responsible for the significant enhancement of As(V) adsorption on {001}-TiO₂ facilitated by U(VI). Importantly, the adsorbed As(V) and U(VI) can be recovered, and {001}-TiO₂ can be used repeatedly. These advantages make {001}-TiO₂ a promising material for the simultaneous removal of As(V) and U(VI) from groundwater.

Results and Discussion

Simultaneous Adsorption of As(V) and U(VI) on Three Types of TiO₂ Facets.

Three types of anatase TiO₂ with {001}, {201}, and {101} as the dominant facets were synthesized by hydrothermal methods. SEM and TEM images revealed that each of the synthesized TiO₂ predominantly exhibited a unique facet (SI Appendix, Fig. S1). The XRD pattern further confirmed that all TiO₂ samples were the anatase crystal phase (SI Appendix, Fig. S2). Detailed characterization of the TiO₂ samples was described in SI Appendix, Text S1. The adsorption efficiencies of the three types of TiO₂ for the simultaneous removal of As(V) and U(VI) (10 μM) were evaluated at pH 7.0 in the presence of 2 mM NaHCO₃ and 1 mM Ca²⁺, where HAsO₄^2- was the dominant As(V) species (56%), and Ca₂UO₂(CO₃)₃ was the dominant U(VI) species (62%) (SI Appendix, Fig. S3). Notably, {001}-TiO₂ exhibited the highest adsorption efficiencies for As(V) (94.0%) and U(VI) (95.8%) within 24 hours, followed by {201}-TiO₂ (83.1% As(V) and 48.6% U(VI)) and {101}-TiO₂ (77.2% As(V) and 43.2% U(VI)) (SI Appendix, Fig. S4). The adsorption performance of the TiO₂ samples is generally affected by the specific surface areas (SBET) and the zeta potential values (ZP). {001}-TiO₂ exhibited a 2.2-fold higher SBET compared to {201}-TiO₂ and {101}-TiO₂ (SI Appendix, Fig. S5), suggesting that {001}-TiO₂ may have more available adsorption sites. Additionally, the surface charge of all three types of TiO₂ was negative, with the order of {001}-TiO₂ (-1.8 mV) > {201}-TiO₂ (-6.2 mV) > {101}-TiO₂ (-7.5 mV) at pH 7.0 (SI Appendix, Fig. S6). As(V) and U(VI) predominantly exist as negatively charged or neutral species at pH 7.0 [H₂AsO₄-, HAsO₄^2-, Ca₂UO₂(CO₃)₃, and CaUO₂(CO₃)₃^2-] (SI Appendix, Fig. S3). The stronger electrostatic repulsion between {201}-TiO₂, {101}-TiO₂, and the negatively charged As(V) and U(VI) species may contribute to the lower adsorption efficiencies of As(V) and U(VI). Our results demonstrated that the superior simultaneous removal performance of {001}-TiO₂ for As(V) and U(VI) was likely due to its larger specific surface area and more positive surface potential. Therefore, {001}-TiO₂ was selected for adsorption of As(V) and U(VI) in the following experiments. The results of individual adsorption experiment demonstrated that {001}-TiO₂ exhibited a remarkable maximum adsorption capacity for As(V) and U(VI) at 222.22 μmol/g and 79.36 μmol/g (SI Appendix, Fig. S7), respectively. These results indicated that {001}-TiO₂ is an optimal material for efficient adsorption of As(V) onto U(VI).

The simultaneous adsorption efficiencies of As(V) and U(VI) (10 μM) were further evaluated at different {001}-TiO₂ concentrations. The results revealed a significant increase in the adsorption rates and efficiencies of both As(V) and U(VI) as the concentration of {001}-TiO₂ increased (SI Appendix, Fig. S8). At low concentrations of {001}-TiO₂ (0.1 g/L), only 26.4 and 39.4% of As(V) and U(VI) were removed, respectively. This can be attributed to the limited availability of adsorption sites on the surface of {001}-TiO₂. However, at a higher concentration of {001}-TiO₂ (0.4 g/L), the adsorption efficiencies of both As(V) and U(VI) exceeded 99.0% within 24 h, and the residual concentrations of As(V) and U(VI) were found to be below the safe drinking water levels set by the WHO and the US EPA [10 μg/L for As(V) and 30 μg/L for U(VI)].

We further investigated whether conventional adsorbents, such as iron oxides, manganese oxides, and activated carbon, exhibited simultaneous adsorption and removal of arsenic and uranium. As shown in SI Appendix, Fig. S9A, goethite, MnO₂, and activated carbon exhibit certain adsorption efficiencies for As(V), with removal rates of 23.54%, 10.86%, and 3.89%, respectively. However, these conventional adsorbents exhibited markedly lower affinity toward U(VI), achieving removal efficiencies of only 6.51%, 3.02%, and 1.75% for goethite, MnO₂, and activated carbon, respectively (SI Appendix, Fig. S9B). Among all tested materials, only TiO₂ exhibits significant potential for the simultaneous removal of both As(V) and U(VI). These results clearly demonstrate the promising potential of {001}-TiO₂ adsorption as an innovative solution for the simultaneous and efficient removal of As(V) and U(VI) from groundwater.

U(VI) Promotes As(V) Adsorption on {001}-TiO₂.

Generally, the coadsorption behavior of multiple contaminants on an adsorbent is competitive, such as the adsorption of As(III), As(V), and F^- on {201}-TiO₂, (21) as well as the adsorption of Cu²⁺, Zn²⁺, and Mn²⁺ on Fe modified clinoptilolite. (20) In this study, coadsorption experiments were conducted at different initial concentration ratios of As(V) and U(VI) to explore the potential competitive adsorption of As(V) and U(VI) on {001}-TiO₂. Surprisingly, the addition of U(VI) did not inhibit the adsorption of As(V) on {001}-TiO₂. In contrast, the adsorption efficiency of As(V) on {001}-TiO₂ increased from 27.2 to 93.5% as the initial concentration of U(VI) increased from 0 μM to 100 μM, resulting in a significant 3.4-fold enhancement (Fig. 1A). This finding suggests that the coexistence of U(VI) can substantially enhance the adsorption of As(V) on {001}-TiO₂. However, the adsorption efficiency of U(VI) remained relatively stable with increasing initial concentration of As(V) compared with U(VI) alone (Fig. 1B), indicating that the coexistence of As(V) had a negligible effect on U(VI) adsorption. Our results highlight the involvement of more complex sorption mechanisms in the coadsorption of As(V) and U(VI) than simple competitive adsorption.

Fig. 1.

Simultaneous adsorption of As(V) and U(VI) on {001}-TiO₂ (0.1 g/L). (A) The adsorption efficiencies of As(V) (10 μM) at different initial concentrations of U(VI). The ratios of As(V) and U(VI) were 1:1, 1:5, and 1:10. (B) The adsorption efficiencies of (10 μM) at different initial concentrations of As(V). The ratios of As(V) and U(VI) were 1:1, 5:1, and 10:1. The coadsorption experiments were carried out in 10 mM NaCl solution at pH 7.0 in the presence of 2 mM NaHCO₃ and 1 mM Ca²⁺.

The phenomenon of U(VI) promoting the removal of As(V) may be attributed to the formation of metaschoepite (UO₃·2H₂O), which subsequently adsorbs As(V). However, MINTEQ calculations reveal that no metaschoepite precipitation occurs under the experimental conditions (SI Appendix, Fig. S3B), and the stable U(VI) concentration in the absence of {001}-TiO₂ further confirms this observation (SI Appendix, Fig. S10). Consequently, the formation of metaschoepite can be ruled out, thereby eliminating the possibility of a metaschoepite-mediated As(V) adsorption mechanism. Although the results preclude the possibility of metaschoepite precipitation under our experimental conditions, we further investigated the adsorption behavior of metaschoepite precipitation for As(V), which were synthesized in the absence of carbonate species. When 10 μM U(VI) was introduced to a carbonate-free system, metaschoepite precipitates formed within 72 h (SI Appendix, Fig. S11). The subsequent introduction of 10 μM As(V) showed negligible As(V) concentration changes over 72 h, indicating minimal adsorption capacity of metaschoepite for As(V). These results suggest that metaschoepite adsorption is not the mechanism responsible for the simultaneous removal of As(V) and U(VI) by {001}-TiO₂. In addition, the concentrations of As(V) and U(VI) remained stable in the absence of {001}-TiO₂ (SI Appendix, Fig. S10), demonstrating that As(V) and U(VI) do not form complex precipitates in the solution. Previous research has shown that surface precipitates can form when the complexation of multiple contaminants on the surface exceeds surface saturation. For example, As(V) complexes with Zn(II) on goethite to gradually form surface precipitate [Zn₂(AsO₄)OH]. (26) Therefore, XRD patterns were used to determine the structure of As(V) and U(VI) adsorption samples at different initial concentration ratios (Fig. 2A). The results indicated that the diffraction peaks of all As(V) and U(VI) adsorption samples matched well with the standard PDF cards of Anatase TiO₂. No diffraction peaks corresponding to As(V)–U(VI) precipitates or As(V)–U(VI)–Ca precipitates were observed, confirming the absence of significant precipitates on the {001}-TiO₂ surface. The surface morphologies and the As(V)/U(VI) atomic ratio of the adsorption samples were characterized using SEM-EDS (Table 1 and Fig. 3). SEM-EDS images revealed the coexistence of As(V) and U(VI) on the {001}-TiO₂ surface after adsorption (Fig. 3). The As(V)/U(VI) atomic ratio of As(V) and U(VI) adsorption samples showed a positive correlation with the variation of the initial concentration ratios of As(V) and U(VI) (Table 1). In the high concentration groups (As(V):U(VI) = 1:10 and 10:1), the As(V)/U(VI) atomic ratios were 0.31 and 2.90, respectively, which were significantly different from the As(V)/U(VI) atomic ratios of As(V)–U(VI) precipitates and As(V)–U(VI)–Ca precipitates (close to 1). In addition, surface precipitation processes typically involve the mutual facilitation of contaminant removal. However, in our experiments, only U(VI) was observed to enhance the removal of As(V), whereas As(V) had a negligible effect on U(VI) adsorption. These results demonstrated that surface precipitation is not the dominant mechanism for the simultaneous adsorption of As(V) and U(VI) on {001}-TiO₂.

Fig. 2.

(A) XRD patterns of {001}-TiO₂ before and after adsorption of U(VI) and As(V) at pH 7.0 in the presence of 10 mM NaCl, 2 mM NaHCO₃, and 1 mM Ca²⁺. (B) Zeta potential of {001}-TiO₂, As(V) preadsorbed on {001}-TiO₂, U(VI) adsorbed on {001}-TiO₂ after As(V) preadsorption. (C) Zeta potential of {001}-TiO₂, U(VI) preadsorbed on {001}-TiO₂, As(V) adsorbed on {001}-TiO₂ after U(VI) preadsorption. Adsorption experiments were carried out at pH 7.0 in the presence of 2 mM NaHCO₃, 1 mM Ca²⁺, and 10 mM NaCl on 0.1 g/L {001}-TiO₂.

Table 1.

The As(V)/U(VI) atomic ratios on {001}-TiO₂ after adsorption of As(V) and U(VI)

	Atomic (%)
Sample	As(V)	U(VI)	As(V)/U(VI)
10 μM As(V)+5 μM U(VI)+TiO2	56.25	43.75	1.29
10 μM As(V)+10 μM U(VI)+TiO2	54.17	45.83	1.18
10 μM As(V)+50 μM U(VI)+TiO2	36.36	63.63	0.57
10 μM As(V)+100 μM U(VI)+TiO2	23.44	76.56	0.31
5 μM As(V)+10 μM U(VI)+TiO2	42.86	57.14	0.75
10 μM As(V)+10 μM U(VI)+TiO2	54.17	45.83	1.18
50 μM As(V)+10 μM U(VI)+TiO2	62.07	37.93	1.64
100 μM As(V)+10 μM U(VI)+TiO2	74.36	25.64	2.90
As(V)-U(VI) precipitations	55.34	44.66	1.24
As(V)-U(VI)-Ca precipitations	52.68	47.32	1.11

The coadsorption experiments were carried out at pH 7.0 in the presence of 0.1 g/L {001}-TiO₂, 10 mM NaCl, 2 mM NaHCO3, and 1 mM Ca2+. As(V)-U(VI) precipitation was synthesized at pH 7.0 in the presence of 100 μM As(V) and 100 μM U(VI). As(V)-U(VI)-Ca precipitation was synthesized at pH 7.0 in the presence of 100 μM As(V), 100 μM U(VI), and 1 mM Ca2+.

Fig. 3.

SEM images and EDS mapping images of {001}-TiO₂ after coadsorption of U(VI) and As(V). (A) 10 μM U(VI) and 10 μM As(V). (B) 100 μM U(VI), and 10 μM As(V). (C) 10 μM U(VI) and 100 μM As(V).

Another possible adsorption mechanism of As(V) and U(VI) is changing the surface charge of {001}-TiO₂. For example, Pb(II) has been shown to promote the adsorption of SO₄^2- on goethite through electrostatic effects. (27) The surface potential of {001}-TiO₂ after adsorption of As(V) alone or U(VI) alone was measured to determine the electrostatic interaction on the surface of {001}-TiO₂ (Fig. 2 B and C). The point of zero charge (PZC) of {001}-TiO₂ shifted from 6.7 to approximately 4.5 after adsorption of As(V) and from 6.7 to approximately 5.9 after adsorption of U(VI), suggesting the formation of negatively charged inner-sphere complexes on {001}-TiO₂ after adsorption of As(V) or U(VI). Meanwhile, As(V) and U(VI) predominantly exist as negatively charged or neutral species at pH 7.0 [H₂AsO_4-, HAsO₄^2-, Ca₂UO₂(CO₃)₃ and CaUO₂(CO₃)₃^2-] (SI Appendix, Fig. S3). Consequently, the negatively charged inner-sphere complexes cannot facilitate the removal of the remaining As(V) or U(VI) in the solution through electrostatic attraction.

Formation of Ternary Surface Complexes on {001}-TiO₂.

In situ ATR-FTIR was further employed to investigate the adsorption mechanism of As(V) and U(VI) on {001}-TiO₂. Initially, the ATR-FTIR spectra of U(VI) preadsorption (10 μM) on {001}-TiO₂ were measured as a function of time, where uranyl-carbonate complexes were the dominant U(VI) species (Fig. 4). Two peaks emerged at 1,440 to 1,240 cm⁻¹ and 1,600 to 1,440 cm⁻¹ upon the addition of U(VI) solution, and the intensity of the peaks increased significantly with the increasing adsorption time (Fig. 4A). Peak fitting analysis was performed on the in situ ATR-FTIR spectra at the range of 1,440 to 1,240 cm⁻¹ and 1,600 to 1,440 cm⁻¹. Four main peaks were deconvolved at 1,527, 1,500, 1,382, and 1,337 cm⁻¹ (Fig. 4 C and E). The pair of peaks at 1,527/1,337 cm⁻¹ were assigned to ν₃, as(C-O) and ν₃, s(C-O) of uranyl-carbonate complexes, while the pair of peaks at 1,500/1,382 cm⁻¹ correspond to CO₃²⁻ monodentate surface complexes. (28, 29) The pair of peaks at 1,527/1,337 cm⁻¹ revealed the formation of a titanium–uranyl–carbonate surface complex [Ti-U(VI)-CO₃²⁻] on {001}-TiO₂ active sites in the presence of uranyl-carbonate complexes (Fig. 5A).

Fig. 4.

(A) U(VI) preadsorption on {001}-TiO₂. (B) 10 μM As(V) adsorption on the U(VI) preadsorbed {001}-TiO₂. (C) Peak fitting analysis of in situ ATR-FTIR spectra of U(VI) preadsorbed on {001}-TiO₂ at 1,600 to 1,440 cm⁻¹. (D) Peak fitting analysis of in situ ATR-FTIR spectra of As(V) adsorption on the U(VI) preadsorbed {001}-TiO₂ after 3 h at 1,600 to 1,440 cm⁻¹. (E) Peak fitting analysis of in situ ATR-FTIR spectra of U(VI) preadsorbed on {001}-TiO₂ at 1,440 to 1,240 cm⁻¹. (F) Peak fitting analysis of in situ ATR-FTIR spectra of As(V) adsorption on the U(VI) preadsorbed {001}-TiO₂ after 3 h at 1,440 to 1,240 cm⁻¹.

Fig. 5.

Schematic illustration of the mechanism for the simultaneous adsorption of As(V) and U(VI) on {001}-TiO₂ through (A) adsorption of U(VI)-CO₃²⁻ complexes and (B) the formation of Ti–U(VI)–As(V) ternary surface complex.

Subsequently, when As(V) (10 μM) was flowed through over the U(VI) preadsorbed {001}-TiO₂, the intensity of the peaks within the range of 1,440 to 1,240 cm⁻¹ and 1,600 to 1,440 cm⁻¹ was obviously decreased with the increase of As(V) adsorption time (Fig. 4D). The peak fitting analysis further revealed a significant decrease in the intensity of the pair of peaks at 1,527/1,337 following As(V) adsorption (Fig. 4 D and F), demonstrating a decrease in the number of uranyl-carbonate bonds on the surface of {001}-TiO₂. (28, 29) There are two possible explanations for this phenomenon. First, it could be due to the adsorption of As(V), which leads to the desorption of adsorbed uranyl-carbonate complexes on the surface of {001}-TiO₂. Second, it is possible that As(V) has replaced CO₃²⁻ ligands in the uranyl-carbonate complexes on the surface of {001}-TiO₂. The coadsorption experiments revealed that As(V) had a negligible effect on U(VI) adsorption (Fig. 1B). Based on this result, the decrease in the number of uranyl-carbonate bonds was likely attributed to the substitution of As(V) for the CO₃²⁻ ligands from the adsorbed uranyl-carbonate complexes (Fig. 5B), rather than the desorption of U(VI) induced by the competitive adsorption of As(V). Overall, these results demonstrated the formation of Ti–U(VI)–As(V) ternary surface complex, where As(V) replaces the CO₃²⁻ ligand of the adsorbed uranyl-carbonate complexes (Fig. 5B). Similar results were made in the experiment of adsorption of 10 μM U(VI) followed by 100 μM As(V) (SI Appendix, Fig. S12). As the concentration of introduced As(V) increased, the intensity of ν3, as(C–O) and ν3, s(C–O) of uranyl-carbonate complexes at 1,527/1,337 cm⁻¹ were more obviously decreased (SI Appendix, Fig. S12 D and F). This result indicated that more As(V) replaced the CO₃²⁻ ligand of the adsorbed uranyl-carbonate complexes, which was more favorable to the formation of the Ti–U(VI)–As(V) ternary surface complex.

The previous studies revealed the mechanism of uranyl-carbonate adsorption on the surface of TiO₂ and established a comprehensive theoretical model for this process. (30, 31) Based on this model, we employed DFT calculations to investigate the thermodynamic feasibility of CO₃²⁻ substitution by HAsO₄²⁻ within the uranyl coordination sphere. Our computational analysis revealed that the substitution of a single CO₃²⁻ ligand by HAsO₄²⁻ in a uranyl complex initially coordinated with two carbonate groups yields a favorable energy change of −2.08 eV (Fig. 6). This negative substitution energy demonstrates the thermodynamic preference for HAsO₄²⁻ coordination over CO₃²⁻ in the uranyl complex. Notably, the subsequent replacement of the second CO₃²⁻ ligand by HAsO₄²⁻ produces an even more favorable energy change of −3.01 eV (Fig. 6), indicating the formation of a highly stable bidentate structure between uranyl and HAsO₄²⁻. These computational findings provide strong theoretical evidence that HAsO₄²⁻ exhibits superior complexation affinity for adsorbed uranyl compared to CO₃²⁻, thereby offering molecular-level insights into the enhanced As(V) removal efficiency in the presence of U(VI).

Fig. 6.

The substitution process of the CO₃²⁻ ligand in uranyl-carbonate complex adsorbed on TiO₂ by HAsO₄²⁻ obtained from DFT calculations.

Effects of Cations, Anions, Organic Matter, and pH.

Groundwater in arsenic-uranium-contaminated areas typically contains cations such as Mn²⁺ and Zn²⁺, as well as anions such as SO₄²⁻ and Cl⁻. (32–34) Additionally, groundwater also contains dissolved organic matter, such as humic acid and citric acid. (35) Therefore, we further investigated the effect of potential competing ions on the simultaneous removal of As(V) and U(VI) by TiO₂. Results demonstrated that varying concentrations of cations (Mn²⁺ and Zn²⁺), anions (SO₄²⁻ and Cl⁻), and dissolved organic matter (humic acid and citric acid) did not have a significant impact on the U(VI) promotion of As(V) removal (SI Appendix, Fig. S13). This further validates the potential application of TiO₂ as an excellent adsorbent for As(V) and U(VI) in complex water environments, demonstrating its promising prospects for practical engineering applications. Coexistence of As(V) and U(VI) in groundwater typically occurs in environments with neutral to slightly alkaline pH conditions (36, 37). We examined the effect of pH on the coadsorption behavior of U(VI) and As(V) at 7.0, 7.5, and 8.0. The results demonstrated that the adsorption capacity of {001}-TiO₂ for As(V) remained relatively stable as the pH increased from 7.0 to 8.0 (SI Appendix, Fig. S14). This phenomenon can be attributed to the predominant existence of As(V) as HAsO₄²⁻ species within the studied pH range, exhibiting high adsorption stability. Meanwhile, a slight decrease in U(VI) adsorption efficiency was observed as pH increased from 7.0 to 8.0, primarily attributed to the increased in the proportion of less adsorbable uranyl tricarbonate complexes (SI Appendix, Fig. S3B). This finding indicates the robust performance of {001}-TiO₂ for the simultaneous removal of As(V) and U(VI) in groundwater with pH conditions typical of As(V) and U(VI) co-occurrence environments.

Regeneration and Reusability of {001}-TiO₂.

The effective removal of adsorbed As(V) and U(VI) is crucial for achieving the regeneration of {001}-TiO₂. Various eluents including HNO₃, NaHCO₃, Na₂CO₃, and NaOH were tested for the removal of the adsorbed As(V) and U(VI). Results indicated that 10 and 20% HNO₃ were able to extract more than 96.4% adsorbed U(VI), while the recovery efficiencies for As(V) were less than 35.0% (Fig. 7 A and B). Na₂CO₃ exhibited a higher recovery ability for both As(V) and U(VI) than NaHCO₃, and the recovery efficiencies of As(V) and U(VI) by using 2 mM Na₂CO₃ were 95.9 and 90.1%, respectively (Fig. 7 C and D). NaOH could almost completely recover the adsorbed As(V) and U(VI) with the recovery efficiencies of 98.7 and 98.6%, respectively, when its concentration is greater than 5 M (Fig. 7 E and F). In addition, NaOH exhibited faster recovery kinetics of As(V) and U(VI) compared to Na₂CO₃. Therefore, 5 M NaOH was selected to regenerate {001}-TiO₂ and evaluate its reusability.

Fig. 7.

Regeneration and reusability of {001}-TiO₂. (A and B) The recovery efficiency of As(V) and U(VI), respectively, by using HNO₃. (C and D) The recovery efficiency of As(V) and U(VI), respectively, by using NaHCO₃ and Na₂CO₃. (E and F) The recovery efficiency of As(V) and U(VI), respectively, by using NaOH. (G and H) The successive removal and recovery efficiency of As(V) and U(VI) for 5 cycles, respectively, using {001}-TiO₂ (0.4 g/L) at pH 7.0 in the presence of 10 mM NaCl, 2 mM NaHCO₃, and 1 mM Ca²⁺.

Throughout the five cycles, {001}-TiO₂ exhibited consistent and high removal and recovery efficiencies for both As(V) and U(VI). The average removal and recovery efficiencies of As(V) were 97.6 and 93.5%, respectively. Similarly, the average removal and recovery efficiencies of U(VI) were 97.0 and 92.3%, respectively (Fig. 7 G and H). After the fifth cycle, the adsorption kinetics of both As(V) and U(VI) remained similar to those in the first cycle (SI Appendix, Fig. S15 A and B), and the morphology of {001}-TiO₂ did not show significant changes compared to that before adsorption (SI Appendix, Figs. S1A and S16). These results demonstrate that {001}-TiO₂ exhibits a stable structure and performance, highlighting its potential for effectively removing As(V) and U(VI) from groundwater in practical applications.

Conclusions

The coexistence of arsenic and uranium in groundwater presents significant challenges to human health. In this study, {001}-TiO₂ has demonstrated excellent performance for efficiently simultaneous removal of the toxic As(V) and U(VI) from groundwater. Interestingly, the presence of U(VI) significantly enhanced the adsorption of As(V), whereas As(V) had a negligible effect on U(VI) adsorption. The maximum adsorption capacities of {001}-TiO₂ for As(V) and U(VI) individually are as high as 222.22 μmol/g and 79.36 μmol/g, respectively. The presence of U(VI) can increase the adsorption capacity of As(V) to 3.4 times. We have first revealed that the formation of ternary surface complex involving Ti–U(VI)–As(V) is the primary mechanism of the enhanced adsorption. Specifically, dissolved As(V) replaces the CO₃²⁻ ligands of the adsorbed uranyl-carbonate complexes on {001}-TiO₂, a process confirmed thermodynamically favorable through DFT calculations. Importantly, the residual concentrations of As(V) and U(VI) after treatment can be both below the safe drinking water level set by WHO and the US EPA. In addition, the adsorbed As(V) and U(VI) can be effectively recovered from the surface of {001}-TiO₂, and the regenerated {001}-TiO₂ maintains its efficient adsorption capacity and rapid adsorption kinetics even after multiple uses, demonstrating the robustness and reusability of {001}-TiO₂. Overall, our study provides a highly efficient and promising method to tackle the co-occurrence of arsenic and uranium in groundwater and offers valuable insights into the mechanisms involved in their simultaneous removal.

Materials and Methods

Chemicals and Materials.

Titanium(IV) isopropoxide (TTIP, 97%) and P25 TiO₂ were purchased from Shanghai Macklin Biochemical Co., Ltd. Titanium butoxide (TB, 99%), isopropyl alcohol (IPA), diethylenetriamine (DETA), N, N-dimethylformamide (DMF), acetic acid (HAc), KOH, NaHCO₃, and CaCl₂ were purchased from Shanghai Aladdin Biochemical Technology Co., Ltd. Na₂HAsO₄·7H₂O was purchased from Alfa Aesar. U(VI)O₂(CH₃COO)₂·2H₂O was purchased from Sinopharm Chemical Reagent Co., Ltd. All chemicals used in this study were of analytical grade or higher quality.

Synthesis of Three Types of TiO₂.

Three types of TiO₂ with predominantly exposed {001}, {201}, and {101} facets, respectively, were synthesized with modified hydrothermal methods. (38–40) To prepare {001}-TiO₂, a typical experiment involved adding DETA (0.05 mL) dropwise to 60 mL of IPA, followed by the addition of 2.5 mL of TTIP dropwise to the solution after gently stirring for 10 min. The resulting solution was transferred to a 100 mL Teflon-lined stainless autoclave and heated in an electric oven at 200 °C for 24 h. After the autoclave cooled to room temperature, the precipitate was collected by centrifugation at 10,000 rpm for 10 min, thoroughly washed with ethanol several times, and then dried at 60 °C overnight. The obtained products were further calcined at 400 °C for 2 h to achieve a highly crystalline anatase phase.

For the synthesis of {201}-TiO₂, a typical procedure involved adding 1.5 mL of TB as a Ti precursor dropwise to a mixed organic solvent of 24 mL HAc and 36 mL DMF. The resulting solution was then transferred to a 100 mL Teflon-lined stainless autoclave and heated in an electric oven at 210 °C for 12 h. The subsequent steps were identical to those used for {001}-TiO₂.

To synthesize {101}-TiO₂, a two-step hydrothermal procedure was employed. In the first step, 39.2 g of KOH was dissolved in 70 mL of ultrapure water, and then 1 g P25 was added under magnetic stirring to form a white suspension. This suspension was transferred to a 100 mL Teflon-lined stainless autoclave and heated in an electric oven at 200 °C for 24 h. After the reaction, the resulting precipitate was separated by centrifugation, washed multiple times with ultrapure water, and then dried at 60 °C overnight to obtain potassium titanate nanowires (KTNWs). In the second step, 200 mg of KTNWs were stirred in 60 mL of ultrapure water and heated in a 100 mL Teflon-lined stainless autoclave at 170 °C for 24 h. The precipitate obtained was collected by centrifugation at 10,000 rpm for 10 min, washed several times with ultrapure water, and then dried at 60 °C overnight. Finally, the products were calcined at 400 °C for 2 h to achieve a highly crystalline anatase phase.

As(V) and U(VI) Adsorption.

Batch adsorption experiments of As(V) and U(VI) (10 μM) on three types of TiO₂ (0.25 g/L) were performed in serum bottles in the presence of 10 mM NaCl, 2 mM NaHCO₃, and 1 mM Ca²⁺ at pH 7.0. The experimental procedure involves sequentially adding CaCl₂, NaHCO₃, and precalculated volumes of U(VI) and As(V) stock solutions to a background electrolyte of 10 mM NaCl, followed by adjusting the pH to the target pH. Finally, the 55 mL mixed solution is transferred to a 55.5 mL serum bottle containing the adsorbent, which is sealed with a rubber stopper and aluminum foil to protect it from light and create a closed system. Serum bottles were sealed with rubber stoppers and aluminum foil to avoid light and create closed systems. The bottles were then placed on a constant-temperature shaker at 25 °C and 125 r/min (ZWY-2012C, Shanghai Zhicheng, China). Dosage experiments (0.1 to 0.4 g/L) were carried out to evaluate the effectiveness of {001}-TiO₂ for the simultaneous removal of As(V) and U(VI) (10 μM). A series of As(V) and U(VI) coadsorption experiments on {001}-TiO₂ (0.1 g/L) were performed with different initial concentration ratios of As(V) and U(VI) (1:1, 1:5, 1:10, 5:1 and 10:1). Blank experiments of As(V) and U(VI) without {001}-TiO₂ were conducted under the same experimental conditions to exclude the formation of As(V)-U(VI) complex precipitates. The adsorption isotherm of As(V) alone or U(VI) alone on 0.1 g/L {001}-TiO₂ was performed with initial As(V) or U(VI) concentration ranging from 1 to 300 μM. The adsorption kinetic of As(V) alone or U(VI) alone on 0.1 g/L {001}-TiO₂ was conducted with 10 μM As(V) or U(VI).

In Situ ATR-FTIR Spectra of As(V) and U(VI) Adsorption on TiO₂ Film.

In situ flow cell ATR-FTIR measurements were performed using FTIR spectrometer (Nicolet IS50R, Thermo Fisher Scientific, U.S.) equipped with a liquid-nitrogen-cooled mercury-cadmium-telluride (MCT) detector, a 45° ZnSe ATR crystal mounted horizontally in a flow cell (PIKE Technologies, U.S.) connected to a LC pump (LC-10AT, Shimadzu, Japan), and an adsorbate solution reservoir. Briefly, 1 mL of TiO₂ suspension (1 g/L) was spread on the ZnSe crystal surface and dried overnight at room temperature to form a uniform TiO₂ film. The flow cell with the coated film was washed with 10 mM NaCl solution using an LC pump at a flow rate of 0.5 mL/min for 2 h to flush out the uncoated particles and equilibrate the TiO₂ film at pH 7.0 until no further changes were observed in the ATR-FTIR spectrum. The background spectra for TiO₂ were acquired after the rinsing process was complete. Subsequently, the electrolyte solution containing 10 μM U(VI), 10 mM NaCl, 2 mM NaHCO₃, and 1 mM Ca²⁺ was pumped into the ATR cell at pH 7.0 to perform in situ preadsorption experiment. After reaching the adsorption equilibrium of U(VI), 10 μM or 100 μM As(V) solution with the same electrolyte composition flowed through over the U(VI) preadsorbed TiO₂ film. Adsorption spectra were averaged with 256 scans every 10 min at 4 cm⁻¹ resolution until the adsorption equilibrium of As(V) or U(VI) was reached (about 3 h). All samples were in the dark during the spectra collection. Data collection and analysis were carried out with Omnic 9.2 software (Thermo Fisher Scientific, U.S.) and Peakfit 4.12 software (Systat Software Inc., U.S.), respectively. The numbers and positions of the peaks were verified using the second derivatives. The curve-fitting analysis of the overlap peak was conducted using the Gaussian line shape.

Computational Details.

Our spin-polarized density functional theory (DFT) calculations (41–43) were carried out in the Vienna ab initio simulation package based on the plane-wave basis sets with the projector augmented-wave method. (44, 45) The exchange-correlation potential was treated by using a generalized gradient approximation with the Perdew–Burke–Ernzerhof parametrization. (46) The van der Waals correction of Grimme’s DFT-D3 model was also adopted. (47) A three layers 4 × 4 supercell of pristine TiO₂(001) surface was built with bottom two-TiO-layer fixed and top two layers relaxed. Meanwhile, a vacuum region of about 20 Å was applied to avoid the interaction between adjacent images. The energy cutoff was set to be 450 eV. The Brillouin-zone integration was sampled with a Γ-centered Monkhorst–Pack mesh (48) of 2 × 2 × 1. The bottom layer was fixed to bulk structure, and top two layer and adsorbates were fully relaxed until the maximum force on each atom was less than 0.01 eV/Å, and the energy convergent standard was 10⁻⁵ eV.

Regeneration and Reusability of TiO₂.

After the adsorption of As(V) and U(VI), the spent {001}-TiO₂ was collected and regenerated in different solutions (HNO₃, NaOH, NaHCO₃, and Na₂CO₃). The regenerated {001}-TiO₂ was then washed with ultrapure water several times until the pH value was close to 7.0. After regeneration, {001}-TiO₂ (0.4 g/L) was reused for simultaneous removal of As(V) and U(VI) (10 μM) in the presence of 10 mM NaCl, 2 mM NaHCO₃, and 1 mM Ca²⁺ at pH 7.0. Five cycles of As(V) and U(VI) simultaneous removal and recovery experiments were carried out to evaluate the reusability of {001}-TiO₂.

Synthesis of As(V)-U(VI) Precipitate and As(V)-U(VI)-Ca Precipitates.

The synthesis reactions were carried out in the absence of carbonate to facilitate the complexation of As(V) with U(VI) and obtain more precipitates.

As(V)-U(VI) precipitate was synthesized in 600 mL solution containing 100 μM As(V) and 100 μM U(VI) at pH 7.0. The solution was then reacted in a constant-temperature shaker for 24 h, and the resulting precipitate was collected by filtration, washed with ultrapure water several times, and dried at 60 °C to obtain the As(V)-U(VI) precipitate. To synthesize the As(V)-U(VI)-Ca precipitate, the same conditions and steps as the As(V)-U(VI) precipitate were followed, with the additional addition of 1 mM Ca²⁺.

Analytical Methods.

After adsorption, all the samples were filtered with 0.22 μm filters (polyethersulfone, ANPEL Scientific Instrument Co., Ltd., China). The concentrations of As(V) and U(VI) were determined by the inductively coupled plasma mass spectrometry (ICP-MS, NexION 300X, PerkinElmer, U.S.).

Characterization of TiO₂ and As(V)-U(VI) Complex Precipitates.

The morphologies of TiO₂ and As(V)–U(VI) complex precipitates were acquired using a field-emission scanning electron microscope (FE-SEM, SU8020, HITACHI, Japan). Morphology of TiO₂ was further obtained using high-resolution transmission electron microscopy (HR-TEM, F20, FEI, U.S.). The crystal structures of TiO₂ and As(V)-U(VI) complex precipitates were characterized by using X-ray diffraction equipment (XRD, X’Pert PRO MPD, PANalytical, Netherlands). The Brunauer–Emmet–Teller specific surface area and Barrett–Joyner–Halenda pore size distribution was determined from N₂ adsorption–desorption isotherms by using an automated gas sorption instrument (ASAP 2460, Micromeritics, U.S.). The surface potential of TiO₂ was measured using Zeta potential instrument (Zetasizer Nano ZS90, Malvern Instrument Ltd., UK).

Data Calculation.

The pseudo-second-order model was used to simulate the kinetics data of U(VI) adsorption according to the following equation (49):

$$q_t=\frac{q_e^{2}kt}{1+q_{e}kt},$$ [1]

where q_t is the adsorption capacity of each electrode at time t (μmol), t is the adsorption time (h), q_e is the equilibrium adsorption capacity of each electrode (μmol), and k is the rate constant of the pseudo-second-order model (h⁻¹).

The Langmuir model was used to simulate the isotherms experimental data of U(VI) adsorption according to the following equation (50):

$$q_e=\frac{k{C}_e{q}_m}{1+k{C}_e},$$ [2]

where q_e is the equilibrium adsorption capacity of each electrode (μmol), q_m is the maximum adsorption capacity of each electrode (μmol), k is the Langmuir constant (L/μmol), and C_e is the equilibrated U(VI) concentration (μM).

The adsorption rate within a certain time was calculated according to the following equation: (7)

$$v_a=\frac{C_0-C_t}{t},$$ [3]

where $v_a$ is the adsorption rate within time t (μM/h), C₀ is the initial U(VI) concentration (μM), C_t is the U(VI) concentration at time t (μM), and t is the adsorption time (h).

The speciation of As (V) and U (VI) in simulated groundwater was calculated using the Visual MINTEQ 3.1 software. (51) The database employed by the software is an extended version of MINTEQA2, (52) with the solubility constant (log K) of schoepite is 5.39.

Supplementary Material

Appendix 01 (PDF)

Data, Materials, and Software Availability

All study data are included in the article and/or SI Appendix.