Characteristics of strontium adsorption onto aquatic sediments in Southwest China

Abstract

The adsorption of strontium (Sr) onto aquatic sediments plays a crucial role in regulating its environmental mobility and level of ecological risk. This study investigated the adsorption behavior of Sr onto selected sediments through batch equilibrium experiments, revealing that adsorption follows the Langmuir and the Freundlich equations. The adsorption temperature was positively correlated with the maximum adsorption capacity (Q_m) of Sr onto sediments (R²: 0.78–0.88). As temperature increases, enhanced molecular kinetics promotes adsorbate diffusion and active site accessibility, resulting in a temperature-dependent augmentation of adsorption capacity (T: 277.15–308.15 K, Q_m: 2926–12628 mg·kg^− 1, an increase of 1.03–2.10 times). During the adsorption process of sediment, calcium ions (Ca²⁺) and strontium ions (Sr²⁺) exhibit significant competitive adsorption behavior. The results show that an increase ion strength significantly inhibits the adsorption capacity of Sr²⁺ (Ca²⁺ concentration: 0–0.05 mol·L^− 1, the maximum inhibition decreased of Q_m: 19.38 times). As the concentration of the equilibrium concentration (C_e) increased, the pH (pH_f) values gradually decreased within the range of 7.05–6.35. The adsorption removal efficiency (R%) of Sr increased with the initial pH (pH_i) (2.17%-49.65%). This study of the adsorption nature of sediments provides an empirical reference for the treatment of radionuclide contamination in sediments.

Supplementary Information

The online version contains supplementary material available at 10.1038/s41598-026-38190-7.

Introduction

Understanding the processes responsible for the long-term transportation and destination of metals in the environment is essential for the protection and remediation of the water-sediment system. In recent years, environmental contamination resulting from heavy metals in sediments has become a matter of concern^1–3. Strontium (Sr) is a widely distributed alkaline heavy metal, a calcium analogue in its physical and chemical properties, with high mobility. It disperses easily and is not directly affected by variation in redox conditions^4–6. The high mobility of Sr is primarily attributed to its highly hydrated nature (hydration enthalpy = − 1480 kJ·mol^{− 1})⁷. Once in the environment, Sr primarily exists as the divalent cation Sr²⁺⁸. Sr also constitutes a major health hazard due to its easy replacement of Ca²⁺ in the body, serving as a threat to both the ecosystem and human health⁹. Therefore, it is essential to understand the processes that influence the migration and transformation of Sr between groundwater, surface water, and saturated sediments. Even though it has high mobility, the migration and transportation of Sr are affected by its adsorption from sediments, a process that depends on the pH, ionic strength, and other factors^10,11. For example, Boyer et al.¹² demonstrated that the cation exchange capacity (CEC) content is positively correlated with Sr adsorption capacity, a relationship specifically manifested through two distinct pathways: in inorganic components, CEC increases with elevated clay content due to the dominance of clay minerals’ ion-exchange properties; in organic components, CEC is primarily derived from specific carbon functional groups (e.g., carboxyl, phenolic hydroxyl groups) that contribute to metal-binding sites. Together, they constitute the key regulatory factors of Sr adsorption capacity. Nguyen et al.¹ showed that as adsorption on river sediment is partly prevented because sediment organic matter (SOM) may inhibit the binding of As(III) and As(V) with functional groups of sediment surface.

Heavy metals are deposited onto sediments primarily through adsorption, complexation, and precipitation. In this context, metals are sustained in a dynamic balance with the aqueous phase¹³. In a water-sediment system, sediments function as significant sinks for dissolved Sr. This characteristic enables them to act as an efficient natural barrier, effectively impeding the spread of groundwater-bearing Sr pollution. Sediments can reduce the concentration of Sr in aquatic ecosystems. However, when the environmental conditions change, heavy metals can re-enter the aquatic environment through transformation in the water-sediment interface, thereby emerging as a secondary pollution source. Therefore, the migration and transformation of Sr in sediments play pivotal roles in influencing the quality of water. Some studies have shown that most of the Sr discharged into water bodies becomes fixed onto sediments^14,15. Zhang et al.¹⁵ found that the migration and distribution of heavy metals were largely influenced by the pH, redox potential, aging, and levels of nutrients in sediments. Šíma et al.¹⁶ found that the preferential uptake of metals in wetland sediments was closely associated with the chemical actions of sulfur and iron. Previous studies have found that the impact of Sr on the environment was altered by the variation in the biogeochemistry of sediments. A substantial volume of early research was devoted to investigating Sr adsorption onto soils^17–20. Relatively little is known about the possible factors influencing the adsorption of Sr in sediments. In this study, three sediments were selected to investigate the influence of temperature and ionic strength on the adsorption of Sr. This study aimed to explore the environmental fate and mobility of Sr and to provide empirical guidance for site restoration and risk assessment. This information will contribute to understanding the behavior of Sr in the environment.

Materials and methods

Preparation of sediment samples

The sediment samples were obtained from three distinct sites in the Mianyang urban area. The three sediment samples were collected from Sichuan Province, Southwest China, at the Southwest University of Science and Technology (Fujiang River (S1): N31°32’ 32.92”, E104°41’ 14.92”; Longshan Creek (S2): N31°32’ 29.29”, E104°42’ 2.20”; Jiuzhou Lake (S3): N31°32′ 14.16′′, E104°41′ 44.32′′). The altitudes of the three areas are 470 m for S3, 465 m for S1, and 503 m for S2. Three surface sediment samples (0–10 cm depth) were collected from 10 m² plots at the three sites and sent to laboratory for analysis. Site S3 remained persistently waterlogged, with a standing water depth of 100 ± 5 cm (measured from the sediment surface), resulting in sustained anaerobic conditions throughout the year. S2 and S1 exhibited variable saturation conditions; the two sediments were seasonally scoured by water, resulting in more oxygen-rich sediment environments. The three sediment samples were stored in plastic basins, while the residual portion was air-dried at room temperature. Decomposing leaves, roots, and gravel were removed. All sediment samples underwent milling and were subsequently sieved through a 100-mesh screen. The sediment samples were sealed and stored under standardized environmental conditions.

Characteristics of sediment samples

The pH was measured using a glass electrode under the condition that the ratio of water to sediments was 2.5 (PHS-2 C Precision Acidity pH meter, Shanghai Jingke Leici, Shanghai, China). The specific surface area (SSA) of the three sediment samples was determined by the nitrogen adsorption method according to the Brunauer-Emmett-Teller theory using an ST-08 specific surface area analyzer (Quantachrome Instruments, Boynton Beach, FL, USA).

The total organic carbon (TOC), total carbon (TC), and dissolved organic matter (DOM) contents of the sediment samples were determined using a Total Organic Carbon Analyzer (Elementar Analysensysteme GmbH, Langenselbold, Germany). The total inorganic carbon (TIC) was calculated from the difference between TC and TOC²¹. The elemental analysis of the sediment samples was performed by using an elemental analyzer (Elementar Analysensysteme GmbH, Langenselbold, Germany).

The cation exchange capacity (CEC) of the sediment samples was assessed by the neutral ammonium acetate method. Each sample was measured in triplicate, and the mean values were calculated. The physicochemical properties of sediments are shown in Table 1.

Table 1.

Physicochemical properties of the three tested sediments.

Sediment1)	SSA2) (m2·g− 1)	TOC3) wt%	TC4) wt%	Elemental composition and atomic ratio of sediment organic components5)					pH	DOM6) (mg·L− 1)			CEC10) (cmol·kg− 1)
				H wt%	N wt%	S wt%	H/C	C/N		DTC7)	DOC8)	DTN9)
S1	5.41	0.89	1.75	0.348	0.09	0.019	2.4	22.7	8.36	1.96	–	–	58.7
S2	5.86	4.37	9.78	1.53	0.62	0.144	1.9	18.3	7.99	2.57	0.65	–	386
S3	28.8	1.26	1.27	0.844	0.12	0.004	8.0	12.4	8.97	3.24	1.48	–	273

⁽¹⁾ S1: the sediments from the Fujiang River; S2: the sediments from Longshan Creek; S3: the sediments from Jiuzhou Lake; ⁽²⁾ SSA: Specific surface area; ⁽³⁾ TOC: Total organic carbon; ⁽⁴⁾ TC: Total carbon; ⁵⁾The elemental composition of the organic constituent was determined by subtracting the ash content and was finally calculated as a mass fraction. ⁶⁾ DOM: Dissolved organic matter; ⁷⁾DTC: Dissolved total carbon; ⁸⁾DOC: Dissolved organic carbon; ⁹⁾DTN: Dissolved total nitrogen; ¹⁰⁾CEC: Cation exchange capacity; “-” indicates that the corresponding factor was not detected, that is, the indicator concentration is below the analytical detection limit (0.01 mg·L^− 1); Each sample was measured in triplicate, and the mean values were calculated.

Chemical reagents

SrCl₂·6H₂O (analytical grade; reported purity > 99.5%) was obtained from Tianjin Kwangfu Fine Chemical Research Institute (Tianjin, China) and was used without further purification. Ultrapure water was used for preparing the solutions throughout the experiment (resistivity was 18.2 MΩ cm^− 1). CaCl₂ and HNO₃ were obtained from Chengdu Kelon Chemical Reagent Plant (Chengdu, Sichuan, China). HNO₃ was of guaranteed reagent grade. All other chemicals utilized in this investigation were of analytical reagent grade and were used as supplied.

Batch adsorption experiment

The isothermal adsorption curves of strontium (Sr) for individual samples were obtained by batch adsorption experiments. The isothermal adsorption experiments were performed by weighing (BSA224S Electronic Balance, Beijing Sartorius Scientific Instrument Company, Beijing, China) 40 mg samples and adding 50 mL of the Sr solution at different concentrations (0, 1, 2, 4, 6, 8, 10, 15, 20, 25, 30, 40, and 60 mg·L^− 1). The isothermal adsorption experiments were conducted at temperatures of 277.15 K, 288.15 K, 298.15 K, and 308.15 K. The adsorption lasted for 60 min without light at a constant temperature for each temperature. For each Sr concentration, all experiments were designed to include two parallel samples and two blank controls (to which no sample was added). After the adsorption was complete, the suspensions were then centrifuged at 4000 r·min^− 1 for 10 min (TGL-16 C Desktop High Speed Centrifuge, Shanghai Anting Scientific Instrument Factory, Shanghai, China). The supernatants were separated from the solid fraction by filtration through a 0.45-µm filter. An appropriate aliquot of the supernatant was diluted with 1% HNO₃ for the measurement of Sr concentration. The pH of the supernatant solutions was determined at the equilibrium stage.

For the experiments investigating the effect of calcium ions (Ca²⁺) concentration on Sr adsorption by sediments, 40 mg of the sample was weighed and added to 50 mL of the Sr solution with concentrations of 0, 1, 2, 4, 6, 8, 10, and 15 mg·L^− 1. The Ca²⁺ concentration was measured as 0.001, 0.01, and 0.05 mol·L^− 1. The suspensions were stored in an oscillator for 60 min under light-excluded conditions at the constant temperature of 25 °C ± 0.1 °C (ZWY-211 C Thermostatic Oscillator, Shanghai Zhicheng Analysis Instrument Co., Ltd., Shanghai, China). Following the completion of the adsorption experiments, the subsequent procedures were the same as those used in the isothermal adsorption experiments.

The concentration of Sr in the solution was analyzed using a flame atomic absorption spectrometer (PE900T Flame Atomic Absorption Spectrometer, PerkinElmer Instruments). The detection wavelength was at 460.7 nm, and the detection limit was 0.03 mg·L^− 1. The current of the lamp was 15 mA; the slit width was 0.7 nm; the air flow rate was 17 L·min^− 1, and the acetylene flow rate was 17 L·min^− 1. Finally, the adsorption capacity of Sr on sediments was calculated, and the final results were presented as mean values.

Data analysis

Statistical analysis and mapping of relevant data were conducted by using Excel 2013 and Origin 2024 software.

The equilibrium Sr adsorption capacity was calculated using Eqs. (1) and (2) derived from the concentration differential between experimental samples and corresponding blank controls.

1

2

where Q_t (mg·kg^− 1) and Q_e (mg·kg^− 1) represent the adsorption amount at time t and at equilibrium, respectively; C_e (mg·L^− 1) is the equilibrium concentration in solution; C₀ (mg·L^− 1) is the initial concentration in solution; V represents the volume of solution in the experiment (mL); and m denotes the mass of the experimental sample (g).

The adsorption data were fitted using the Freundlich Eq. (3) and the Langmuir Eq. (4) isothermal models.

3

4

5

where Q_m (mg·kg^− 1) represents the maximum adsorption amount; Q_e (mg·kg^− 1), C_e (mg·L^− 1), and C₀ (mg·L^− 1) have the same definitions as in Eq. (2); K_L (L·mg^− 1) is a constant related to the binding strength; K_f ((mg·kg^− 1)·(mg·L^− 1)^−N) represents the Freundlich regression parameter; N represents the adsorption constants of the Freundlich model; and R_L is a dimensionless factor derived from the Langmuir model’s regression parameters. In this study, the value of R_L was calculated for the highest initial concentration studied (0–60 mg·L^− 1).

The adsorption removal efficiency of Sr was calculated using Eq. (6), derived from the concentration differential between experimental samples and corresponding blank controls.

6

where R (%) is the adsorption removal efficiency of Sr; and C₀ (mg·L^− 1) and C_e (mg·L^− 1) are defined as in Eq. (2).

Results and discussion

Characterization of the tested sediments

The basic physicochemical characteristics of the three tested sediments are presented in Table 1. The pH values ranged from 7.99 to 8.97 for the three tested sediments. The pH values were slightly higher than those reported in the literature; the pH values of all wetland substrates were below the value of 5.1 reported by Boyer et al.¹². Previous studies have shown that strontium (Sr) itself is not a redox-sensitive species, remaining at its Sr²⁺ oxidation state below pH 13 regardless of the redox potential²². Thus, the oxidation state of Sr²⁺ was assumed to be unaffected by the natural pH in these experiments. The specific surface area (SSA) was in the order of 5.41 m²·g^{− 1} for S1, 5.86 m²·g^{− 1} for S2, and 28.79 m²·g^{− 1} for S3. The SSA range was relatively narrow (5.41–5.86 m²·g^{− 1}) for S1 and S2. The values of SSA were relatively equivalent to those observed by Jurina et al.²³, who showed that the value of SSA for surface sediments of the delta plain ranged from 2.5 to 34.7 m²·g^{− 1}.

The elemental analysis revealed that the percentage composition (%) of total carbon (TC) for the three sediments followed the sequence S2 > S1 > S3, with values significantly lower than those in sediments from the intertidal regions of the Yellow River Delta, China²⁴. The total organic carbon (TOC) was in the order: S2 > S3 > S1. The total inorganic carbon (TIC) was calculated as the difference between TC and TOC^5,21. The TIC of the three sediments was as follows: 0.01% for S3, 0.86% for S1, and 5.41% for S2. The H/C value reflects the aromaticity of a sediment. Aromaticity was enhanced with the decrease in the H/C value. The H/C value was 8 for S3, indicating that the carbon in S3 was largely composed of aliphatic carbon. The aromaticity values of S1 and S2 were greater than that of S3, which was composed of aromatic carbon, and the carbon was highly stable in S1 and S2. In comparison to aliphatic TOC, aromatic TOC exhibits higher reactivity and a greater ability to complex metals¹⁷. The molar C/N ratios were 12.4, 22.7, and 18.3 for S3, S1, and S2, respectively. The distinct ratios are representative of the corresponding sources of soil nitrogen. The nitrogen in the sediments was of microbial, emergent vegetation, and/or algal origin²⁵. The results are generally in line with those of Boyer et al.²⁶, who reported molar TOC/TN (total nitrogen) values of 22 ± 4 for swamp sediments and 12 ± 2 for marsh sediments. Simpson et al.²⁵ found that most of the organic carbon in the sediments was contributed by microbes, mostly in the form of peptide and protein structures. For older decaying organic matter, protein-rich organic compounds are present in sediments. As the age of the organic matter increases, more proteins occur in the sediments²⁶.

The variation in hydrologic regimes, temperature, vegetation, and sustained anaerobic conditions leads to distinct dissolved organic matter (DOM) concentration profiles for sediments. Due to its structural (containing hydroxyl, carboxyl, carbonyl, and other active functional groups) and physicochemical properties, DOM is considered one of the most active chemical components in sediments, as it significantly influences the fate and transport of contaminants²⁷. In this study, the surface water dissolved organic carbon (DOC) concentrations ranged from 0.65 to 1.48 mg·L^{− 1}, significantly lower than the values reported in literature, where the range is typically from 30.3 to 90.4 mg·L^{− 128}. According to the dissolved total carbon (DTC) and dissolved total organic carbon (DTOC) contents, we were surprised to find that the dissolved total inorganic carbon (DTIC) was fairly similar (1.76–1.92 mg·L^{− 1}) for the three sediments. Previous research findings indicated that the processing of DOC within lakes depends on the extent to which DOC is mineralized to the atmosphere or settles as particulate organic carbon to bottom sediments and thereby remains in the lake¹⁴. Dissolved total nitrogen (DTN) was not detected in any of the sediments.

The cation exchange capacity (CEC) differed notably among the three sediments, and the CEC of S2 (386.44 cmol·kg^{− 1}) was 6.59 times that of S1 (58.67 cmol·kg^{− 1}). Shafie et al.²⁹ found that the CEC of urban river sediments ranged from 151.95 to 155.05 cmol·kg^{− 1}, which is within the range of CEC reported in this study.

Sr adsorption isotherms of sediments

The isothermal adsorption profiles of strontium (Sr) in the three selected sediments with different temperatures are shown in Fig. 1, and the regression parameters derived from fitting the Langmuir and the Freundlich models are listed in Table 2. The adsorption data exhibited a good fit to both the Langmuir and the Freundlich equations (R² ≥ 0.95). The isothermal adsorption curves for Sr in the three selected sediments were similar. Based on the isothermal adsorption profiles, the adsorption capacity increased with the Sr concentration and tended to an equilibrium, and the adsorption capacity increased with the temperature. The 60-minute equilibrium time used in this study, based on preliminary kinetic experiments, is sufficient to ensure equilibrium for bulk external diffusion. However, for samples with high specific surface area, a longer duration might facilitate more complete intraparticle diffusion. Future studies could further optimize this parameter.

Fig. 1.

Adsorption isotherms of Sr onto three selected sediments at different temperatures.

Table 2.

Parameters obtained from the Langmuir and Freundlich models for Sr adsorption isotherms with the three selected sediments at different temperatures.

Sediment1)	T2) (K)	Langmuir			Freundlich
		Q m 3) (mg·kg− 1)	K L 4) (L·mg− 1)	R 2	Kf[(mg·kg− 1) ·(mg·L− 1)−N] 5)	N 6)	R 2
S1	277.15	2926 ± 127	0.035 ± 0.003	0.99	195 ± 23	0.58 ± 0.03	0.96
	288.15	3003 ± 103	0.075 ± 0.007	0.98	435 ± 20	0.45 ± 0.01	0.99
	298.15	3462 ± 129	0.072 ± 0.007	0.98	521 ± 28	0.43 ± 0.02	0.98
	308.15	4175 ± 54	0.047 ± 0.001	0.99	410 ± 38	0.51 ± 0.03	0.97
S2	277.15	5320 ± 214	0.039 ± 0.003	0.99	390 ± 48	0.57 ± 0.04	0.95
	288.15	5783 ± 88	0.32 ± 0.02	0.99	2128 ± 133	0.26 ± 0.02	0.95
	298.15	7950 ± 117	0.12 ± 0.005	0.99	1689 ± 99	0.38 ± 0.02	0.97
	308.15	11,169 ± 200	0.074 ± 0.003	0.99	1611 ± 133	0.44 ± 0.02	0.96
S3	277.15	6142 ± 146	0.04 ± 0.002	0.98	480 ± 45	0.56 ± 0.03	0.97
	288.15	6443 ± 164	0.45 ± 0.06	0.96	2882 ± 142	0.22 ± 0.02	0.96
	298.15	8059 ± 187	0.16 ± 0.01	0.98	2060 ± 87	0.34 ± 0.01	0.98
	308.15	12,628 ± 281	0.10 ± 0.006	0.99	2256 ± 103	0.41 ± 0.01	0.99

⁽¹⁾ S1: the sediments from the Fujiang River; S2: the sediments from Longshan Creek; S3: the sediments from Jiuzhou Lake; ⁽²⁾ T is the reaction temperature; ⁽³⁾ Q_m is the maximum adsorption amount; ⁽⁴⁾ K_L is a constant related to the binding strength; ⁽⁵⁾ K_f represents the Freundlich regression parameter; ⁽⁶⁾ N represents the adsorption constants of the Freundlich model.

The Langmuir maximum adsorption capacity (Q_m) of Sr was in the order S1 < S2 < S3. The Q_m value of S3 at the temperature of 277.15 K was 6142 mg·kg^− 1, which was 2.1 times that of S1. The Q_m value of Sr on the three tested sediments increased with the temperature, and the adsorption capacities of S3 and S2 were greater than that of S1. The Q_m values at temperatures of 288.15 K, 298.15 K, and 308.15 K were 1.03, 1.18, and 1.43 times that at 277.15 K for S1; 1.09, 1.49, and 2.10 times that at 277.15 K for S2; and 1.05, 1.31, and 2.06 times that at 277.15 K for S3 (Table 2).

To provide a clearer picture of the fate of Sr at the sediment-water interface, we considered the relationship between cation exchange capacity (CEC) and total organic carbon (TOC). The results are shown in Fig. 2. The CEC values of the three sediments were linearly correlated with the TOC content, with an R² value of 0.94. The TOC content for S2 was 3.47 times that for S3 and 4.91 times that of S1, and the CEC content for S2 was 1.41 times that of S3 and 6.58 times that of S1, consistent with the hydrological conditions at the sampling locations and the presence of dense vegetation. Hydrological fluctuations extend the diffusion path length of particulate matter, thereby decreasing the preservation of TOC in sediments. In this research, the CEC values were 273 cmol·kg^{− 1} for S3 and 386 cmol·kg^{− 1} for S2. These values were significantly higher than those of Boyer et al.¹², who found that decaying cattail litter present in wetland water demonstrated a CEC of 47.9 cmol·kg^{− 1}, and two wetland sediment samples collected from a different site of the Chalk River area had CEC values of 52.6 and 65.0 cmol·kg^{− 1}.

Fig. 2.

Relationship between cation exchange capacity (CEC) and total organic carbon (TOC) in the sediments.

The correlation between the maximum adsorption capacity of Sr on the three chosen sediments and temperature is shown in Fig. 3. The results show that temperature was positively correlated with the maximum adsorption capacity of Sr onto sediments, with R² values ranging from 0.78 to 0.88 for the three sediments. The correlation between the maximum adsorption capacity and temperature of S1 was stronger than that for the other two sediments. The Q_m of Sr on the three tested sediments increased with the temperature of the system. In the adsorption process, the degree of adsorption of Sr increased with the rise of temperature in the system, and the probability of collision between Sr and organic matter adsorption sites was greatly increased.

Fig. 3.

Relationship between the maximum adsorption capacity of Sr for the three selected sediments and temperature.

The factor influencing the adsorptive property was not the total number of adsorption sites, but rather the number of effective adsorption sites. The effective adsorption sites refer to those sites in which the adsorbent can overcome the diffusion barrier between particles and with particles, which was affected by the temperature and particle size (100-mesh sieve) in the adsorption system. The diffusion barrier of the adsorption ions will be significantly reduced as the temperature increases. The adsorption of Sr onto selected sediment surfaces exhibits a temperature-dependent enhancement, consistent with thermodynamic principles governing surface reactivity. As temperature increases, molecular thermal motion intensifies, leading to elevated collision frequency between Sr and adsorption sites. Therefore, the adsorption of Sr by sediments increases with increasing temperature.

Effect of ionic strength on Sr adsorption onto sediments

The effect of the calcium (Ca) ion concentration on the adsorption of strontium (Sr) by sediments at 277.15 K is shown in Fig. 4. The adsorption capacity increased with the Sr concentration at different Ca concentrations. The adsorption capacity of Sr by sediments significantly decreased with the increase in the Ca concentration in the solution from 0.001 to 0.5 mol·L^{− 1}. The inhibition sequence of Sr adsorption capacity onto sediments by Ca concentration was determined as 0.05 mol·L^{− 1} > 0.01 mol·L^{− 1} > 0.001 mol·L^{− 1}. Wallace et al.¹⁰ also found that the adsorption capacity of Sr decreased with the increase in ion strength. In many studies, Sr was electrostatically bound to the surface of sediments in the form of outer-sphere complexes, with ion exchange as the dominant sorption mechanism^30,31. Outer-sphere complexes are weakly associated with surface binding sites and hence are generally reversibly bound and mobilized with the solution’s increase in ionic strength³¹.

Fig. 4.

Influence of the calcium ion concentration on the adsorption of Sr by sediments.

Table 3 lists the regression parameters of the Langmuir and Freundlich models for Sr adsorption capacity onto the three selected sediments with different ion strengths of Ca at 277.15 K. The correlation coefficients ranged from 0.95 to 0.99. The maximum adsorption capacity (Q_m) of Sr (the concentration of Sr was between 0 and 15 mg·L^{− 1}) decreased markedly at concentrations of Ca from 0 to 0.05 mol·L^{− 1} in the order S3 (from 6643 to 326 mg·kg^{− 1}, a reduction by 19.38 times) > S2 (from 7532 to 753 mg·kg^{− 1}, a reduction by 9.00 times) > S1 (from 2271 to 678 mg·kg^{− 1}, a reduction by 3.35 times). Sr adsorption decreased significantly in the presence of higher ionic strength solutions, as also found by Kaplan et al.⁷.

Table 3.

Parameters obtained by fitting the Langmuir and Freundlich models for Sr adsorption isotherms with the three selected sediments under different concentrations of calcium ions.

Sediment1)	(mol·L− 1)	Langmuir			Freundlich
		Q m 2) (mg·kg− 1)	K L 3) (L·mg− 1)	R 2	Kf[(mg·kg− 1) ·(mg·L− 1)−N] 4)	N 5)	R 2
S1	0.000	2271 ± 50	0.18 ± 0.009	0.99	468 ± 17	0.48 ± 0.02	0.99
	0.001	2208 ± 45	0.13 ± 0.005	0.99	345 ± 21	0.55 ± 0.03	0.99
	0.010	764 ± 16	0.36 ± 0.02	0.99	243 ± 18	0.39 ± 0.04	0.96
	0.050	678 ± 82	0.09 ± 0.02	0.97	73 ± 5.8	0.63 ± 0.04	0.98
S2	0.000	7532 ± 357	0.14 ± 0.01	0.99	1179 ± 23	0.56 ± 0.009	0.99
	0.001	6755 ± 250	0.10 ± 0.006	0.99	773 ± 58	0.64 ± 0.04	0.98
	0.010	1431 ± 35	0.37 ± 0.03	0.99	457 ± 19	0.40 ± 0.02	0.99
	0.050	753 ± 46	0.20 ± 0.03	0.98	168 ± 11	0.47 ± 0.03	0.98
S3	0.000	6643 ± 196	0.27 ± 0.02	0.99	1733 ± 65	0.45 ± 0.02	0.99
	0.001	6070 ± 63	0.25 ± 0.007	0.99	1481 ± 90	0.48 ± 0.03	0.98
	0.010	414 ± 13	1.0 ± 0.16	0.97	231 ± 6	0.22 ± 0.01	0.99
	0.050	326 ± 42	0.11 ± 0.03	0.95	43 ± 6	0.58 ± 0.05	0.95

⁽¹⁾ S1: the sediments from the Fujiang River; S2: the sediments from Longshan Creek; S3: the sediments from Jiuzhou Lake. ⁽²⁾ Q_m is the maximum adsorption amount; ⁽³⁾ K_L is a constant related to the binding strength; ⁽⁴⁾ K_f is the Freundlich regression parameter; ⁽⁵⁾ N is the adsorption constant of the Freundlich model.

The relationship between the equilibrium concentration (C_e) and pH (pH_f) of Sr adsorption on the three selected sediments for different calcium ion concentrations is shown in Fig. 5. As the concentration of C_e increased, the pH_f values gradually decreased, with the range of 7.05–6.77 at 0.001 mol·L^− 1 Ca, 6.58–6.35 at 0.01 mol·L^− 1 Ca, and 6.60–6.38 at 0.05 mol·L^− 1 Ca. At Ca concentrations of 0.001, 0.01, and 0.05 mol·L^− 1, the pH_f values of the adsorption system for S1, S2, and S3 were generally higher than those of S0. This result is consistent with the inherent pH range (7.99–8.97) of S1, S2, and S3. For S3, the pH_f decreased gradually as the Ca concentration increased. As adsorption progresses, there may be cation exchange adsorption between Sr and hydrogen ions, resulting in an increase in hydrogen ions in the solution and a slight decrease in the pH of the solution.

Fig. 5.

Relationship between C_e and pH at different calcium ion concentrations and Sr adsorption on the three selected sediments. S0 is the pH of the blank Sr solution (without adding sediments). pH_f is the pH of the solution after the adsorption experiment.

The correlations between the pH of the solution before and after adsorption at different concentrations of Ca are shown in Fig. 6. The relationship between pH_i and pH_f was linear for the three sediments (R²≥ 0.86). The buffer capacity of a sediment represents its ability to maintain a constant pH when exposed to acidic or alkaline environments⁵. All sediments were weakly alkaline, in the order S3 > S1 > S2 (Table 1). This indicated that S3 had the strongest adsorption capacity for protons. Under Ca concentrations of 0.001, 0.01, and 0.05 mol·L^{− 1}, the slopes of the three sediments were generally greater than 1. This may have been due to the competitive adsorption between hydrogen and Sr during the experiment, resulting in an increase in pH values after adsorption.

Fig. 6.

Relationship between the pH of the solution before and after adsorption of Sr by sediments. pH_i and pH_f are the pH of the solution before and after the adsorption experiment, respectively.

The pH is a crucial factor influencing the adsorption of heavy metals by sediments¹². The influence of the solution pH (pH_i) on the adsorption removal efficiency of Sr is shown in Fig. 7. The adsorption removal efficiency (R%) of Sr increased with pH_i. At a Ca concentration of 0.001 mol·L^{− 1}, the R values for S1, S2, and S3 reached 16.8%, 29.8%, and 49.8%, respectively. These removal efficiencies are in line with those reported by Šíma et al.¹⁶, who showed that the average removal efficiency of Sr from municipal wastewater in the constructed wetland was 27.7%.

Fig. 7.

Effect of the solution pH_i on the adsorption removal efficiency of Sr. pH_i is the initial pH of the solution before the adsorption experiment.

The pH range of this study is limited to 6–8. However, all sediments are inherently weakly alkaline (pH 7.99–8.97). At highly alkaline pH (> 12.5), Fuller et al.³² found that Sr adsorption increased with increasing ionic strength and pH, due to the specific adsorption of Sr (as the form of SrOH⁺), and that Sr was not outcompeted by other ions. Therefore, Sr may form hydroxo complexes such as SrOH⁺ under alkaline conditions, potentially shifting the adsorption mechanism from ion exchange to surface complexation or even precipitation.

Conclusion

The adsorption data of strontium (Sr) were well described by both the Langmuir and Freundlich equations. The adsorption temperature (T) was positively correlated with the maximum adsorption capacity (Q_m) of Sr onto sediments (R²: 0.78–0.88). As temperature increases, thermally enhanced molecular kinetics promote adsorbate diffusion to the adsorbent surface and augment the accessibility of active sites, leading to a temperature-dependent augmentation of adsorption capacity (T: 277.15–308.15 K, Q_m: 2926–12628 mg·kg^− 1, an increase of 1.03–2.10 times). During the adsorption process of sediment, calcium ions (Ca²⁺) and strontium ions (Sr²⁺) exhibit significant competitive adsorption behavior at sediment surface sites due to their similar ionic radius and charge characteristics. The results show that an increase in Ca²⁺ concentration significantly inhibits the adsorption capacity of Sr²⁺ (Ca concentration: 0–0.05 mol·L^− 1, the maximum inhibition decreased of Q_m: 19.38 times). As the concentration of the equilibrium concentration (C_e) increased, the pH (pH_f) values gradually decreased within the range of 6.35–7.05. The adsorption removal efficiency (R%) of Sr increased with the increase of pH (pH_i) (2.17%-49.65%). During the macroscopic external diffusion process of sediment, strontium may undergo cation exchange adsorption with hydrogen ions, resulting in a slight decrease in the pH of the solution. This study focused on the competitive effect of Ca²⁺, highlighting the importance of competition from ions of the same valence. However, the co-existence of multiple cations in real environments may lead to more complex competitive or synergistic effects. Future research will validate these findings within a more complex ionic matrix.

Supplementary Information

Below is the link to the electronic supplementary material.

Author contributions

Xiang Luo and Yungui Li conceived and designed the experiments; Xiang Luo, Dan Zhang, and Mingluo Zhou performed the experiments and analyzed the data; Xiang Luo and Dan Zhang wrote the paper; and Mingluo Zhou, Xiang Luo, and Yungui Li revised the paper. All authors have read and approved the final manuscript.

Funding

This study was supported by Key Lab of Process Analysis and Control of Sichuan Universities (No.GCFX2025004); Yibin University’s University-Level Project (No.2025SZ403).

Data availability

The datasets generated during and/or analyzed during this study are available from the corresponding author upon reasonable request.

Declarations

Ethics approval and consent to participate

This study does not involve human participants, animal experiments, or data collected from social media platforms. Therefore, ethical approval and consent to participate are not applicable.

Consent for publication

All authors have read and agreed to the manuscript’s published version.