Effects of Oxygenation Resuspension on DOM Composition and Its Role in Reducing Dissolved Manganese in Drinking Water Reservoirs
* These authors contributed equally to this work.
1 Key Laboratory of Environmental Aquatic Chemistry, State Key Laboratory of Regional Environment and Sustainability, Research Center for Eco-Environmental Sciences, Chinese Academy of Sciences. No. 18 Shuangqing Road, Haidian, Beijing, 100085, China
2 School of Ecology and Environment, Inner Mongolia University. No.235 West College Road, Saihan, Hohhot, 010021, China
3 University of Chinese Academy of Sciences. No. 19A Yuquan Road, Shijingshan, Beijing, 100049, China
✉ Correspondence: Changwei Lü <lcw2008@imu.edu.cn>, Min Yang <yangmin@rcees.ac.cn>
KEYWORDS
Dissolved oxygen (DO); Manganese; Dissolved organic matter (DOM); Sediment resuspension; Drinking water reservoir
ABSTRACT
Anaerobic conditions in source water sediments are a key driver of manganese (Mn) release in drinking water systems. Enhancing sediment oxidation can inhibit Mn release, but the mechanisms of Mn speciation under varying oxidative conditions remain unclear. This study examined sediment exposure to oxygenated water layers at controlled dissolved oxygen levels (0, 2, 5, 7 mg L-1) through laboratory simulations. Results showed Mn release is negatively correlated with DO (\(R^2=0.93\), \(p\)=0.034), with oxygen driving reactions between dissolved organic matter (C2 and C3 components) and forming functional groups (-OH, -COOH) that remove Mn through adsorption or complexation (C2: \(R^2\)=0.57, \(p\)<0.001; C3: \(R^2\)=0.53, \(p\)<0.001). Field studies in six reservoirs identified operational thresholds for sediment resuspension to mitigate Mn risks (compensation threshold: 17.4 μg L-1; risk threshold: China: 95.5 μg L-1; WHO: 70.8 μg L-1). These findings clarify Mn-organic matter interactions and can provide practical guidance for Mn and algae removal in source water systems.
SYNOPSIS
This study proposes proactive sediment resuspension for manganese removal, offering a novel solution for source water reservoirs.
Introduction
Manganese (Mn) is a naturally occurring element found in the Earth’s crust and is essential for both environmental and biological systems1,2. However, elevated concentrations of Mn in drinking water systems can present significant challenges3,4. High levels of Mn in water can cause discoloration (yellow or black water)5–7, form brownish precipitates in pipes, and stain clothing when exposed to oxidants. Mn oxides can adsorb toxic metal ions (e.g., Cr, Cu, Fe, Pb)8–10, and their dissociation may release these ions, further compromising water safety. Additionally, studies suggest that elevated Mn exposure in drinking water may lead to neurobehavioural issues and reduced intelligence quotient (IQ), particularly in children11–13.
Mn removal has been extensively studied, with various technologies such as oxidation, biological filtration, membrane processes, aeration, electrocoagulation, and catalytic materials being widely investigated14–16. However, these methods often fail to handle sudden Mn concentration spikes in source water, resulting in treatment inefficiencies and exceedances. Therefore, preventing Mn exceedance at the source is a more fundamental approach. Mn speciation is closely linked to water redox conditions: Mn (IV), typically in the form of MnO2, is insoluble and settles in sediment, while Mn (II), the soluble form, is the primary cause of exceedances in drinking water17. In reservoir systems, deeper waters can become hypoxic or anoxic during summer stratification18, reducing sedimentary Mn (IV) to Mn (II), which then dissolves. When the thermocline breaks down in autumn, vertical mixing can transfer Mn (II) to upper water layers, potentially causing exceedances in the drinking water supply19–21.
Hypolimnetic oxygenation (HOx) systems22 increase redox potential (Eh) by introducing oxygen to the hypolimnion, keeping metals like Mn in oxidized forms and reducing their release from sediments23, while a recent study found that 94% of removed Mn in this way will be re-mobilized within one week24. In 2024, we proposed a novel algae control approach based on sediment resuspension (SR). Demonstrated in drinking water reservoirs, this method effectively inhibits algal growth by reducing underwater light availability and enhancing flocculation, while also removing dissolved phosphate through adsorption onto sediment particles—all without introducing external materials25. However, the potential for SR to either elevate water-column Mn levels or suppress its release via enhanced oxidative conditions remains unclear, which is critical for its broader application.
SR may influence Mn dynamics through two competing mechanisms: (1) physical disturbance could release sediment-bound Mn (II) into the water column21, potentially increasing dissolved Mn concentrations; and (2) enhanced oxygen transfer at the sediment-water interface during resuspension could promote the oxidation of Mn (II) to insoluble Mn (III/IV) oxides26, thereby reducing dissolved Mn levels. Additionally, dissolved organic matter (DOM) plays a crucial role in controlling the speciation and bioavailability of heavy metals. Components of DOM, such as humic acid (HA) and fulvic acid (FA), can undergo redox reactions or form complexes with metals like Mn, significantly influencing their speciation. For example, DOM can complex with metal ions through carboxyl and phenolic functional groups, altering metal bioavailability27. Specific interactions include Zn complexing with HA via carboxyl groups and Hg ions binding with carboxyl, phenolic, and alcoholic groups28,29. Furthermore, DOM can participate in abiotic oxidation processes involving oxidants like H2O230, Mn (III/IV) oxides, and Mn (III)-ligand complexes31–36, which can alter its chemical properties37.
To investigate the mechanisms governing Mn speciation transformation under varying oxygenation conditions and the interaction between DOM and Mn in resuspended sediment, we conducted a controlled laboratory experiment. Dissolved oxygen (DO) levels were systematically regulated from anaerobic to aerobic, with stirring applied to simulate SR. Additionally, field applications were carried out in six drinking water reservoirs to evaluate the efficiency of SR in Mn removal and assess potential risks of Mn release. Based on these findings, we proposed two threshold values to define the boundary conditions for the safe application of SR technology in mitigating Mn-related risks. This study provides novel insights into the regulatory role of oxygen in DOM-Mn interactions and speciation transformation, while establishing critical prerequisites for the practical implementation of SR in drinking water reservoirs.
Materials and methods
Laboratory Simulation Experiments
Sediment samples were collected from Reservoir F (Fig. S1) using a Petersen grab sampler. Subsurface sediments (10–20 cm depth) were placed in 5 L polyethylene containers, sealed, transported under dark, refrigerated conditions, and frozen upon arrival at the laboratory until use.
Prior to the experiments, all sediments were homogenized to remove stones and plant debris. The prepared sediments were then transferred to 10 simulators (black polyethylene buckets, 5 L each). Each simulator was filled with 4 L of oxygen-free water (oxygen removed via aeration with N2) and 742.8 g of sediment to achieve a sediment-to-water ratio of 40 g L-1. This ratio was selected based on preliminary results (Fig. S2), which indicated that metal release was highest at this sediment-to-water ratio. The sediment moisture content was adjusted to 78.46%38. To simulate the temperature increase typically observed during SR operations, the simulators were maintained at 20 °C and allowed to equilibrate for two weeks in the dark before initiating the experiments. During this period, DO, oxidation-reduction potential (ORP), and pH were monitored to ensure that all setups reached a consistent initial state (Fig. S3).
Eight simulators were stirred with N2 and/or O2, and a DO meter (JPB-607A) was used to continuously monitor and maintain DO concentrations at 0, 2, 5, and 7 mg L-1. These levels were maintained for 15 min, with each condition including a parallel group. After 15 min, water samples were collected at regular intervals, immediately refrigerated, and subsequently filtered. An additional simulator was prepared without the initial 15-minute stirring but was aerated with O2 to maintain a DO concentration of 7 mg L-1. Furthermore, a control simulator was established without stirring or aeration, resulting in a DO level of 0 mg L-1 (Fig. S4). Sediment samples were collected both before and after the experiment and promptly freeze-dried. The physical characterization of the sediments used in the experiment is detailed in Table S1 and Fig. S5.
Field applications in reservoirs
Field applications were conducted in six drinking water reservoirs (A, B, C, D, E, and F; Fig. S1), with experimental (labeled as FE) and control (FC) zones established in each. Water samples were stratified and collected from the overlying water column, filtered through 0.45 μm membranes, and refrigerated for later metal concentration analysis.
SR operations were applied with varying frequencies and areas across the reservoirs. In Reservoir A, SR was conducted at a frequency of 0.625 d-1 over a 25 ha area. In Reservoir B, SR was implemented across 12.08 ha. In Reservoir C, SR was performed at a frequency of 1.172 d-1 over 18 ha. In Reservoir D, SR was conducted at a frequency of 1.049 d-1 across 26.5 ha. In Reservoir E, SR was applied at a high frequency of 17.31 d-1 over a 3.8 ha area. In Reservoir F, SR was conducted at a frequency of 5.461 d-1 across 12 ha.
Measurements for water and sediment samples
Water samples were immediately pre-filtered using a 4 μm filter to separate particulate and dissolved fractions. The filtrate was further filtered through 0.22 μm and 1.0 μm membranes. Samples passing through the 0.22 μm filter were designated as truly dissolved metals, while those in the 0.22–1.0 μm range were categorized as colloidal metals39,40. Both sample types were stored at 4°C in the dark for subsequent analyses.
For metal concentrations in the overlying water, the filtrates were acidified with 1% nitric acid and analyzed using inductively coupled plasma mass spectrometry (Thermo Fisher iCAP Q). DOM in the water was characterized by three-dimensional fluorescence spectroscopy (F-7000, Hitachi). Fluorescence measurements were conducted with an excitation wavelength range of 200–500 nm, an emission wavelength range of 250–550 nm, and an emission slit width of 5 nm. Milli-Q water was used as the blank control, and the scanning speed was set to 12,000 nm/min.
Surface sediments were dried using a vacuum freeze dryer (FreeZone 4.5) and subsequently ground with an agate mortar before being sieved through a 100-mesh screen for sediment property analyses. Sediment particle size was determined using a laser particle size analyzer (Mastersizer 3000), and specific surface area was measured with an automated surface area and micropore analyzer (ASAP2020HD88).
The extraction of different phosphorus (P) fractions was conducted using a modified method based on ref.41, ref.42, and ref.43. Details of the extraction procedure are provided in Table S2.
Risk Model
Based on SR experiments conducted (Table S3) at six reservoirs, a correlation model (Eq. 1) was established to describe the relationship between the mean Mn concentration at FC (\(c_{\text{FC}}\)) and the ratio of the average Mn concentration (\(\rho\)) in the overlying water at FC and FE (\(\rho = \overline{c_{FE}} / \overline{c_{FC}}\)):
\[ \rho = \left(A_0^{-1} + A_1 \log_{10}\overline{c_{\text{FC}}}\right)^{-1} \tag{1}\]
Here, \(c_{\text{FC}}\) represents the control condition without SR, while \(c_{\text{FE}}\) represents the condition with SR. This correlation was used to determine the compensation threshold (\(\text{CT}\), Eq. 2), which indicates the point at which SR becomes ineffective at removing dissolved Mn from the water column under elevated Mn concentrations:
\[ \text{CT} = \left(1 - A_0^{-1}\right) A_1^{-1} \tag{2}\]
A log-log correlation model (Eq. 3) was developed to describe the relationship between the maximum Mn concentrations observed during routine reservoir monitoring (\(\text{max}(c_{\text{FC}})\)) and after SR (\(\text{max}(c_{\text{FE}})\)):
\[ \log_{10} \text{max}(c_{\text{FE}}) = B_0 + B_1 \log_{10} \text{max}(c_{\text{FC}}) \tag{3}\]
Here, \(B_0\) and \(B_1\) are the intercept and slope of the log-log correlation curve, respectively. This model was used to determine the risk threshold (\(\text{RT}\), Eq. 4), which indicates the point at which SR may introduce a potential risk (guidance concentration, provided by drinking water guidelines, \(\text{GC}\)) of Mn presence in source water, necessitating subsequent treatment at drinking water plants:
\[ \text{RT} = 10^{B_1^{-1} (\log_{10} \text{GC} - B_0)} \tag{4}\]
Two guidance concentrations were used for the \(\text{RT}\): 1) \(\text{GC}_1 = 100\,\text{µg L}^{-1}\), based on the Chinese National Standard for Drinking Water Quality (GB 5749-2022)44; and 2) \(\text{GC}_2 = 80\, \text{µg L}^{-1}\), based on the WHO guideline for Mn concentrations in drinking water (WHO/HEP/ECH/WSH/2021.545).
Data analysis
Definition of DOM Components
The method of DOM components measurement are described in Fig. S7, and the definition of each component is as follows:
C1 (Ex 225 nm, Em 325 nm): Protein-like matter, linked to degraded proteins, particularly tryptophan-like compounds46. C2 (Ex 230 nm, Em 415 nm): Terrestrial humic-like substances with aromatic structures, large molecular size47, and hydrophobicity48,49. C3 (Ex 263 nm, Em 464 nm): Likely FA-like terrestrial humic substances50–52.
The cut-off for sufficient spectral congruence was set as:
\[ \theta = \theta_{\text{ex}} \times \theta_{\text{em}} \tag{5}\]
Statistical Analysis
All experimental data were processed and visualized in R (v4.0) using ggplot2. Pearson’s correlation coefficients were calculated, with statistical significance assessed at \(p\) < 0.05. Boxplots display medians, and data points outside 1.5 × IQR were removed to exclude outliers, ensuring accurate representation of central tendency and dispersion. Mn, Fe, and other metal elements were quantified using standard calibration curves, with results considered valid when the R2 value exceeded 0.99. Spike recovery tests demonstrated recovery rates within the range of 80%–120%, which is within the acceptable range for environmental analysis.
Results
Migration and transformation of Mn under various DO concentration in simulators
Independent simulators were employed to investigate Mn concentration dynamics under varying anaerobic and aerobic conditions (Fig. S4). Maximum Mn concentrations were recorded approximately 30 min after a stirring operation (lasting 15 min) across all simulators (Fig. 1A). Subsequently, Mn concentrations declined significantly, following a second-order kinetic model (Table S4, Fig. 1A). The modeled maximum Mn concentrations (\(c_0\)) demonstrated a strong correlation with DO levels (\(R^2 = 0.93\), \(p\) = 0.034; Fig. 1B).The highest second-order rate constant (\(k\)) was observed at 5 mg L-1 DO (\(k = 5.75 \times 10^{-4} \text{min}^{-1}\)), while the lowest occurred at 0 mg L-1 DO (\(k = 1.31 \times 10^{-5} \text{min}^{-1}\)). Notably, at the highest DO concentration (7 mg L-1), Mn degradation was inhibited, with \(k = 2.57 \times 10^{-5} \text{min}^{-1}\). Similarly, the proportion of dissolved Mn relative to colloidal Mn exhibited a decreasing trend with increasing DO levels (0 mg L-1: 87.5%, 2 mg L-1: 65.5%, and 5 mg L-1: 45.7%). However, at DO = 7 mg L-1, it increased significantly to a very high level (86.3%, Fig. 1C).
To further clarify the effects of stirring, four simulators were used to represent anaerobic & no-stirring, aerobic & no-stirring, aerobic & stirring, and anaerobic & stirring conditions (Fig. 1D and Fig. S6). Rapid decreases in Mn (II) concentrations were observed under both aerobic & stirring (93.11%) and anaerobic & stirring (90.75%) conditions, compared to the aerobic & no-stirring and anaerobic & no-stirring conditions.
Correlation between DOM components and Mn concentrations
PARAFAC analysis identified three fluorescent components (Fig. S7): protein-like (C1), terrestrial humic-like substances (C2, C3). To evaluate the effects of aeration and stirring on DOM fluorescence intensity and components, experiments involving aeration alone (Fig. 2A) and stirring alone (Fig. 2B) were performed, compared to a control set (no aeration, no stirring). Aeration alone significantly increased the fluorescence intensity of C2 (2.32 fold change) and C3 (3.71 FC), but not C1 (1.08 FC), resulting in an overall increase in total DOM fluorescence intensity (2.03 FC). In contrast, stirring alone significantly increased the fluorescence intensity of C1 by 2.55 FC, but not C2 (1.33 FC) or C3 (1.08 FC), leading to an overall increase in total DOM fluorescence intensity (1.86 FC). Notably, aeration alone shifted the dominant DOM component from C1 to C3, whereas stirring alone did not alter the dominant DOM component.
Further, we analyzed the impact of DO concentration on the fluorescence intensity of DOM components. Overall, the fluorescence intensity of DOM components increased with rising DO concentrations (Fig. 2C and 2E, Fig. S8A). The highest fluorescence intensity of C2 and C3 were observed at DO = 5 mg L-1, while C1 (Fig. S8A) reached its highest fluorescence intensity at DO = 7 mg L-1. Moreover, the fluorescence intensity of C2 and C3 showed a significant correlation with Mn (II) concentrations during the decreasing phase (C2: \(R^2\) = 0.57, \(p\) < 0.001, Fig. 2D; C3: \(R^2\) = 0.53, \(p\) < 0.001, Fig. 2F). In contrast, the correlation between C1 (\(R^2\) = 0.14, \(p\) = 0.022, Fig. S8B) and Mn (II) was weaker.
The transformation of Mn fractions in the sediment under various operations
The transformation of Mn fractions (mild acid-soluble, reducible, and oxidizable fractions; Table S5, Table S6) in sediment was assessed by comparing its composition before and after the experiment. Under aerobic & no-stirring conditions compared to anaerobic & no-stirring conditions, all Mn fractions exhibited increases: mild acid-soluble by 13.0%, reducible by 29.1%, and oxidizable by 80.4% (Fig. 3A). Comparing anaerobic & stirring to anaerobic & no-stirring conditions, the mild acid-soluble and reducible fractions showed slight changes, decreasing from 100.2% to 96.9% and increasing slightly from 97.2% to 98.1%, respectively. However, the oxidizable fraction displayed a substantial rise from 39.7% to 109.4%, despite its relatively low absolute amount. Joint oxygenation and stirring operations caused significant decreases in the mild acid-soluble and reducible fractions by 27.2 and 14.8%, respectively, compared to anaerobic and no-stirring conditions, while the oxidizable fraction increased considerably by 64.9%, from 39.7% to 104.6%.
To investigate the role of DO, Mn removal rates were analyzed under varying DO conditions (Fig. 3B). Increased DO concentrations significantly enhanced Mn removal rates for the mild acid-soluble and reducible fractions, with strong statistical correlations (mild acid-soluble: \(R^2\) = 0.85, \(p\) = 0.076; reducible: \(R^2\) = 0.81, \(p\) = 0.097; Fig. 3B). In contrast, the oxidizable fraction exhibited a weaker but still positive correlation with DO concentrations (\(R^2\) = 0.69, \(p\) = 0.170).
Field oxygenation experiments in F Reservoir further clarified Mn dynamics. The highest removal rates for the mild acid-soluble and reducible fractions, 42.1 and 35.8%, respectively, were achieved on the 5th day after stirring (Fig. 3C). Subsequently, both fractions rebounded, with their removal rates decreasing to 25.7 and 28.9% by the 10th day. Conversely, the oxidizable fraction reached its peak removal rate of 20.5% on the 10th day.
Field applications in six reservoirs
To evaluate the effects of oxygenation and stirring on Mn transformation, in-situ SR experiments were conducted in six reservoirs across different seasons. The water depths of these reservoirs ranged from 4 m to 20 m. Detailed information, including water temperature, pH, and other environmental factors, is summarized in Table S3. Mn concentrations in the reservoirs ranged from 0.25 μg L-1 to 55.53 μg L-1 (Fig. 4A - Fig. 4F). Post-SR, Mn concentrations generally decreased in Reservoirs A–E but increased in Reservoir F.
In Reservoir A, Mn (II) decreased from 0.25 μg L-1 to 0.047 μg L-1 in 5 days. Reservoir B saw a decrease from 1.22 μg L-1 to 0.87 μg L-1 over 21 days, while in Reservoir C, levels declined from 4.58 μg L-1 to 0.63 μg L-1 over 39 days. Reservoir D’s Mn (II) dropped from 5.94 μg L-1 to 3.64 μg L-1 in 14 days, and Reservoir E had a smaller decrease from 8.62 μg L-1 to 7.86 μg L-1 over 29 days.
In contrast, Reservoir F showed increased Mn (II) levels (112.87 μg L-1) compared to the control site (FC, 55.53 μg L-1) during the SR operations (Fig. 4F). Field observations revealed high Mn (II) concentrations at both FC and FE initially, with subsequent declines. Reservoir F exhibited elevated Mn risk in the early stages. A risk model was developed with two thresholds: the \(\text{CT}\) (\(\text{CT}\) = 17.4 μg L-1) and the \(\text{RT}\) (\(\text{RT}\) = 95.5 μg L-1 for national standards, 70.8 μg L-1 for WHO standards). Dissolved iron showed minimal variation in Reservoirs A and D, with stable removal efficiency across other reservoirs except Reservoir D (Fig. S9).
Discussion
The role of oxygenation and stirring in transforming Mn speciation
This study investigated the effects of oxygenation and stirring on Mn speciation and its release from reductive sediments into the water column. We found that oxygen concentration significantly influenced the transformation of Mn into colloidal and particulate forms, promoting its removal through sedimentation. Mn (II) concentration exhibited a strong negative correlation with DO concentration (\(R^2\) = 0.93, \(p\) = 0.034), with lower Mn concentrations observed at higher DO levels. Compared to completely anoxic conditions (DO = 0 mg L-1), the Mn concentration in the aerobic group (DO = 7 mg L-1) was significantly reduced (806.0 μg L-1). This reduction is likely due to the rapid oxidation of certain substances in the water, particularly organic matter, which indirectly decreased Mn (II) concentrations. These findings align with ref.37, supporting the concept that in a ternary system containing mineral surfaces, Mn (II), and organic matter, a Mn redox cycle coupled with organic carbon degradation is established.
Furthermore, the rate of Mn concentration decline in the water column was slowest under anoxic conditions (1.31 × 10-5 min-1) and increased nearly 44-folds under moderately oxygenated conditions (DO = 5 mg L-1; 5.75 × 10-4 min-1), coinciding with a significant rise in the proportion of colloidal Mn. Specifically, during the decrease of total Mn, as the DO concentration increases from 0 mg L-1 to 5 mg L-1, both the rate of total Mn decline (Table S4) and the proportion of colloidal Mn (Fig. 1C) increase. This indicates that, under these conditions, the process of dissolving Mn into colloidal Mn is predominant, with the transformation of colloidal Mn into particulate Mn occurring as a secondary process. However, at higher DO levels (7 mg L-1), the proportion of colloidal Mn decreased. This can also be confirmed from Fig. S10, where throughout the entire experimental phase, as the DO concentration increases, the time for Mn in the water to begin to decrease becomes earlier and the peak Mn concentration also decreases. We hypothesize that elevated oxidative conditions further promote the aggregation of colloidal Mn into large particles, which is mechanistically plausible and consistent with findings in previous studies53. A recent study also demonstrated that higher oxygen levels significantly enhance the aggregation of colloidal P into particulate forms40.
A comparison of four experimental conditions—stirring vs. no-stirring, and anoxic vs. oxygenated—revealed that Mn concentrations in stirred groups were significantly lower than those in no-stirring groups. This suggests that mechanical stirring may influence Mn oxide-organic matter interactions through two mechanisms: 1) stirring increases the contact frequency between DOM and Mn oxide particles, promoting the adsorption of Mn (II) to the solid phase. This is consistent with the results of our previous investigation into the properties of sediment (Table S1), which showed that the sediment of F reservoir is mainly dominated by medium particles with rapid settling characteristics, while containing fine particles. It also has a high specific surface area (23.31 m 2/g) and surface activity, making it highly capable of adsorbing heavy metals. Ref.37 showed that adsorbed divalent Mn became the primary component at the beginning of DOM and Mn (II) reactions in the solid phase. 2) stirring also enhances the collision frequency between particles, facilitating the flocculation process of Mn oxides and organic matter. Research indicates that organic matter can enhance the stability of Mn oxide colloids through electrostatic repulsion and steric hindrance effects54. Moreover, the resuspension process increases the exposure of smaller mineral particles with larger specific surface areas, such as ferrihydrite, which has been shown to adsorb DOM55,56. Additionally, divalent cations (such as Mg2+, Ca2+) may promote the adsorption of organic matter-Mn oxide complexes31,37,54, without affecting the oxidation processes. This finding aligns with recent research on the complexation mechanisms between algal-derived extracellular organic matter and Mn57. Therefore, these observations suggest that DO plays a key role in promoting the adsorption and complexation of Mn (II) by DOM, while stirring enhances the aggregation of Mn with natural organic matter (Mn-NOM), facilitating its rapid sedimentation.
The role of organic matter components in the fractionation of Mn
The composition of organic matter influences its interactions with heavy metals53,58. In this study, we examined different organic matter components to analyze their complex relationship with DO and Mn, aiming to identify the mechanisms behind the decrease in Mn (II) concentration during SR due to oxygenation.
In the absence of oxygenation, Mn oxides (such as MnO2, MnOOH, and Mn3O4) in the water can adsorb, fractionate, and oxidize DOM, thereby being reduced to Mn (II)37. Then, DOM is degraded into smaller organic molecules (such as pyruvate, acetone, formaldehyde, acetaldehyde)33,59, leading to higher Mn (II) concentrations in the water. Some studies suggest that oxygen during resuspension reduces Mn (II) concentrations by accelerating the oxidation of Fe (II) and Mn (II) into oxides, which adsorb the dissolved metals60. During the entire experiment, the pH of the overlying water ranged from 6.6 to 6.9 (Fig. S3C), which is a mildly acidic environment unfavorable to redox reactions between Mn oxides and DOM60,61. However, under near-neutral to slightly acidic pH conditions in our experiments, the oxidation rate of Fe (II) and Mn (II) was limited62, and the decrease occurred within 40 hours, indicating that adsorption by Mn oxides does not significantly contribute to this process.
Upon the introduction of oxygen, oxygen competes with DOM for Mn oxides63, reducing the Mn (II) generated from redox reactions. Simultaneously, oxygen oxidizes DOM, altering its composition, with an increase in the proportions of C2 and C3 components (Fig. 2A), though their quantities remain relatively unchanged (Fig. S11). C2, characterized by a higher abundance of aromatic moieties, exhibits a strong affinity for cationic metal complexation59,64. Its pronounced hydrophobicity promotes adsorption onto sedimentary particles and metal oxide surfaces, significantly reducing its mobility in aquatic systems50. In contrast, C3 displays marked hydrophilicity, enabling extensive transport within fluvial and groundwater systems and exerting a substantial influence on metal ion complexation dynamics65. The molecular structure of C3 includes diverse reactive functional groups—such as carboxyl (-COOH), phenolic hydroxyl (-OH), carbonyl (-C=O), and ether (-O-) moieties—which provide a strong chelation capacity for cationic metals. Among these, carboxyl (-COOH) and phenolic hydroxyl (-OH) groups serve as the primary binding sites for interactions with cationic heavy metal species66–68. This explains the observed negative correlation between C2, C3, and Mn (II) concentrations (Fig. 2D and 2F). The pH reduction observed during the experiment further suggests that DOM enhances the adsorption of Mn oxides through surface complexation and ligand exchange31. This also explains the increase in the proportion of colloidal Mn in the water as oxygen concentration rises during the stirring and oxygenation process (Fig. 1C).
The differentiation between C2 and C3 components is mechanistically significant, as it highlights the distinct pathways through which different DOM fractions interact with metal species in aquatic environments. Specifically, the preferential adsorption of C2 onto particulate matter facilitates the removal of associated metals through sedimentation processes. In contrast, the persistent mobility of C3 promotes the long-range transport of complexed metals. These divergent behaviors directly influence metal speciation, bioavailability, and geochemical cycling. Moreover, understanding these mechanistic differences may provide critical insights for predicting the fate of metals and developing targeted remediation approaches50 in both natural and engineered aquatic systems.
We also monitored the influence of fluorescence intensities of C2 and C3 on the concentrations of other metals in the water (Fig. S12). The results showed that as the fluorescence intensities of C2 and C3 increased, the concentrations of most metals (Al, Ba, Mg, and Sr) decreased, while the concentration of Fe remained relatively stable. Previous studies have indicated that the presence of Al can significantly affect the structure and stability of Fe-DOM co-precipitates. Al primarily binds to DOM through functional groups such as carboxyl groups, which may competitively inhibit the binding of Fe to DOM, leading to the weak correlation between Fe and C2/C369. This observation aligns with the higher Al concentrations detected in our study, suggesting that the presence of Al is likely the main reason for the lack of a clear relationship between Fe and C2/C3.
In reservoirs, the removal of Mn primarily relies on heterogeneous oxidation, as the efficiency of homogeneous oxidation of Mn (e.g., reactions with free chlorine or DO) is 7–8 orders of magnitude lower than that of Fe (II)70. Consequently, homogeneous oxidation is unlikely to be the dominant mechanism for Mn removal in these systems. We propose that SR operations may enhance the adsorption of Mn (II) onto mineral surfaces, such as iron oxides70,71, thereby facilitating heterogeneous oxidation of Mn. Iron plays a pivotal role in Mn removal processes. Fe (III) oxides not only catalyze the oxidation of Mn (II) and promote the formation of Mn (II)-doped ferrihydrite70, but also influence the oxidation kinetics of Fe (II) and the formation of Fe colloids71. Moreover, in Fenton-like oxidation systems, Mn (II) can accelerate the reduction of Fe (III), enhance the generation of reactive oxygen species, and improve Fe cycling and oxidation efficiency72. Therefore, the observed decrease in Mn concentrations during SR operations is not solely attributable to DOM but is also closely linked to heterogeneous oxidation processes. These interactions between Mn and Fe, coupled with the catalytic role of Fe oxides, make the Mn removal process during SR more complex and intriguing.
Framework for the overall process
Higher oxygen concentrations during stirring significantly reduce mild acid-soluble and reducible Mn fractions while partially suppressing the increase of the oxidizable fraction (Fig. 3B). This indicates that oxygenation decreases the bioavailability of Mn, consistent with the findings of ref.73, who demonstrated that resuspension techniques mitigate heavy metal remobilization in aquatic ecosystems. However, oxygenation of overlying water alone is insufficient to effectively reduce bioavailable Mn in sediment. Anaerobic stirring moderately decreases active Mn levels, while stirring with oxygenation not only results in a pronounced reduction in mild acid-soluble and reducible fractions but also suppresses the increase of the oxidizable fraction (Fig. 3A). In summary, oxygenation without stirring increases bioavailable Mn as Mn transitions from particulate matter to sediment. In contrast, stirring reduces the amount of bioavailable Mn, with oxygenation further accelerating this process and improving sediment quality. The underlying mechanism of this process is illustrated through a comprehensive framework that integrates experimental data and literature-reported mechanisms. This framework highlights the complex interactions between Mn and DOM, emphasizing their potential mechanisms and ecological significance (Fig. 5).
Aeration significantly influences sediment metal speciation and bioavailability. In the mild acid-soluble fraction, aerobic conditions reduce NH\(_4^+\) levels in the water column, promoting sediment adsorption of metal ions and increasing mild acid-soluble metals, whereas anoxic conditions lead to their reduction74. In the reducible fraction, Fe (II) oxidizes to Fe (III) under oxic conditions, with iron and Mn hydroxides serving as key adsorbents due to their large surface area, notably affecting the partition of Fe, Mn, Cr, and Ni between dissolved and particle fractions75,76. For the oxidizable fraction, aeration accelerates the decomposition of organic matter, releasing bound metals; however, this effect is limited by the fraction’s relatively low proportion in sediments77. Overall, no-stirring aeration increases the bioavailability of metals in sediments, while stirring aeration decreases it. The positive correlation between sediment bioavailability and metal concentrations in the overlying water78 aligns with the observations in Fig. 4, where SR resulted in a decline in Mn concentrations in the water column.
Metal interactions exhibit both competitive and synergistic effects on Mn adsorption, precipitation, and redox transformations in aquatic systems. Fe and Mn can act as bridging agents between negatively charged DOM and other ions, such as As, enhancing adsorption processes through the formation of ternary complexes79. These bridging configurations likely play a significant role in Mn removal kinetics. Additionally, metals such as Cu2+ may catalyze redox reactions that influence Mn speciation, similar to how Cu2+ catalyzes the dissolution of Fe compounds from sediment surfaces80. When dissolved metals return to sediments, metal sulfide precipitation exerts strong control, which may explain the eventual equilibrium observed in later stages81. Research indicates that salinity gradients influence these metal interactions, with increased salinity promoting metal mobility in the order: Cd > Mn > Cu > Pb82. Collectively, these complex metal interactions contribute to the observed rapid decrease in dissolved Mn concentrations through multiple mechanisms.
Implications
The results indicate that stirring sediments under oxygenated conditions can remove dissolved Mn from water through complex mechanisms. This suggests that the Mn-related risks previously associated with SR for algal control may be less severe than anticipated in practical applications. This understanding motivated us to test SR in six drinking water reservoirs with varying background Mn concentrations.
In five reservoirs, SR reduced dissolved Mn concentrations to varying degrees. The differences in Mn removal efficiency may not only be related to initial Mn concentrations but also to sediment bioavailability, aeration conditions, and other environmental factors. These properties should be further investigated in relation to the season when SR is applied, reservoir depth, water temperature, and other variables. For instance, in Reservoir F, where dissolved Mn was exceptionally high (112 μg L-1), SR was ineffective at reducing Mn levels. This elevated concentration is likely linked to anaerobic sediment release, as indicated by the black coloration of the sediment, reflecting prolonged anoxic conditions. Based on field applications in six reservoirs, we propose a critical threshold (CT) of 17.4 μg L-1, meaning that SR is unlikely to reduce dissolved Mn when concentrations exceed this level. Additionally, a risk threshold (RT) of 95.5 μg L-1 is suggested, as SR at higher concentrations may elevate Mn levels above the drinking water standard of 100 μg L-1, necessitating further Mn removal measures at water treatment facilities. The World Health Organization (WHO) recommends avoiding long-term exposure to Mn levels above 80 μg L-1, implying that source water Mn concentrations should ideally remain below 70.8 μg L-1. Notably, SR operations conducted near this range did not cause sharp increases in Mn levels, suggesting minimal impact on water treatment processes. However, since the number of field reservoirs studied is still limited, the risk model and its proposed thresholds may not be universally applicable to other reservoirs. Verification of the model results should be performed before full-scale application.
Although these thresholds may vary under different reservoir conditions, they offer essential insights for guiding the implementation of SR-based algal control technologies. The compensation threshold (\(\text{CT}\)) helps identify when SR is ineffective, while the risk threshold (\(\text{RT}\)) serves as a warning that the operation may result in excessive Mn release, based on established restriction rules. This is particularly critical for drinking water sources, where the use of chemical algicides is prohibited. Clarifying the Mn release risks and removal capabilities associated with SR-based techniques holds significant practical relevance for ensuring safe and sustainable water management. However, when Mn concentrations in reservoirs are high, SR may fail to remove Mn effectively and can even lead to increased concentrations. This highlights the need to explore new, suitable approaches for non-invasive sediment treatment, such as leveraging Mn-oxidizing microorganisms83.
Global carbon storage, oxygen cycling, and the mobility of heavy metals in sedimentary systems have been prominent topics of discussion in recent years84. In terrestrial and marine environments, interfacial reactions between minerals and organic carbon facilitate interactions with clay minerals, leading to the formation of stable structures that effectively protect organic carbon from microbial degradation85. In marine sediments, Fe and Mn oxides play a crucial role in transforming labile organic carbon into more complex and recalcitrant forms, enabling the annual sequestration of approximately 4.1 gigatons of carbon86. While our study closely aligns with these topics, further modeling efforts are needed to better evaluate these processes in the context of the global carbon-oxygen balance and Mn mobility within the Earth system.
FIGURES
- Fig. 1: Dynamics of manganese (Mn) under varying conditions, including dissolved oxygen (DO) concentrations (0, 2, 5, and 7 mg L-1) and operational modes (stirring vs. no stirring) in the water. (A) Second-order kinetic model depicting total Mn dynamics over time; (B) Relationship between peak total Mn concentration (\(c_0\)) and DO levels; (C) Changes in the proportions of Mn species, including dissolved Mn and colloidal Mn, within the water column (Samples were collected 9 min after the stirring and/or aeration operations); (D) Comparison of the effects of stirring and aeration on dissolved Mn concentrations.
- Fig. 2: Components of DOM under varying conditions, including operational modes (stirring vs. no stirring) and DO concentrations (0, 2, 5, and 7 mg L-1): (A) Comparison of DOM components at 0 mg L-1 and 7 mg L-1 DO under no-stirring conditions; The percentage represents the proportion of fluorescence intensity in each respective category. The outermost ring shows the proportion of each component (C1, C2, C3) at three stages of the experiment: increasing (before peak), decreasing, and stabilization. (B) Comparison of DOM components under stirring and no-stirring conditions at 0 mg L-1 DO; (C, E) fluorescence intensity of the C2 (C) and C3 (E) components under different DO levels with stirring; (D, F) Correlation between the fluorescence intensity of C2 (D) and C3 (F) components and Mn (II) concentration.
- Fig. 3: Mn fractions (mild acid-soluble, reducible, and oxidizable) under various experimental conditions in sediment before and after the simulation experiment and in the field application: (A) The changes in Mn fractions in the sediment before and after the simulation experiment are represented by the ratio of the values after the experiment to those before the experiment; (B) Correlation models depicting Mn reduction under oxygenic conditions with varying DO levels (0, 2, 5, and 7 mg L-1); (C) Mn removal rate over operational time during SR in Reservoir F.
- Fig. 4: Field applications and risk thresholds. (A-F) Comparison of Mn concentrations between FE (sites with SR) and FC (control sites without SR) across six reservoirs (Reservoirs A-F). (G) Temporal dynamics of Mn concentrations in the water column at FC and FE in Reservoir F over a 29-day SR operation. (H) As defined in Eq. 2, determination of the CT for Mn removal via SR, based on the modeled relationship between the mean Mn concentration at FC and ρ. (I) As defined in Eq. 4, determination of the RT for Mn release during SR, derived from the modeled relationship between the maximum Mn concentrations at FC and FE.
- Fig. 5: Schematic diagram of the principle of the decrease in Mn (II) concentration during SR.
ASSOCIATED CONTENT
Supporting Information
The Supporting Information is available free of charge at https://pubs.acs.org/doi/10.1021/acs.est.5c00235.
Geographic locations of sampling sites (Fig. S1); Sediment Mn release experimental results (Fig. S2); Time-series monitoring data of pH, dissolved oxygen (DO), and oxidation-reduction potential (ORP) in simulated systems(Fig. S3); Schematic diagram of the laboratory simulation apparatus (Fig. S4); Sediment adsorption-desorption isotherms of Mn (Fig. S5); Comparative analysis of colloidal Mn across simulated systems (Fig. S6); Fluorescence excitation-emission matrix (EEM) spectra of dissolved organic matter (DOM) (Fig. S7); Correlation between DOM component C1 and DO concentrations (Fig. S8); Spatial-temporal variations of iron (Fe) in six reservoir sediments (Fig. S9); Dynamics of dissolved Mn and colloidal Mn in simulation experiments (Fig. S10); Relationship between DO and dissolved organic carbon (DOC) in simulated systems (Fig. S11); Pearson correlation analysis of DOM components (C2, C3) with metals (Al, Ba, Fe, Mg, Sr, Zn) (Fig. S12); Physicochemical properties of sediment samples (Table S1); Sequential extraction protocols for phosphorus (P) fractions (Table S2); Field experiment details (sampling dates, locations, and water depths) (Table S3); Fitted parameters of second-order kinetic models for Mn release (Table S4); Chemical extraction methods of Mn speciation (Table S5); Environmental significance of Mn speciation (table s6); References (Text) (PDF)
ACKNOWLEDGEMENTS
This work was financially supported by the National Key R&D Program of China (2022YFC3203603), Shanghai Municipal Investment (Group) Corporation (CTKY-ZDZX-2023-004), and the National Natural Science Foundation of China (52030002, 51878649, W2412156, 42167028).
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