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Article

Effects of Ammonium and COD on Fe and Mn Release from RBF Sediment Based on Column Experiment

by
Xuelian Xia
,
Yanguo Teng
* and
Yuanzheng Zhai
*
Engineering Research Center of Groundwater Pollution Control and Remediation of Ministry of Education of China, College of Water Sciences, Beijing Normal University, Beijing 100875, China
*
Authors to whom correspondence should be addressed.
Water 2023, 15(1), 120; https://doi.org/10.3390/w15010120
Submission received: 27 November 2022 / Revised: 22 December 2022 / Accepted: 26 December 2022 / Published: 29 December 2022
(This article belongs to the Special Issue River Ecological Restoration and Groundwater Artificial Recharge II)
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Abstract

:
Riverbank filtration (RBF) is an important part of the surface water–groundwater cycle, and it intercepts and retains many pollutants in rivers. However, RBF affects the biogeochemical process which enables aquifer sediments to release iron (Fe) and manganese (Mn). In this study, column experiments were performed to investigate the effects of ammonium ions and organic matter on Fe and Mn release from anaerobic RBF sediments. In addition, high-throughput sequencing technology was used to characterize the microbial community. The results showed that the ammonium ions (NH4+) and organic matter (COD) in groundwater promote the release of Fe and Mn from aquifer sediments. The trends of Fe and Mn release were similar during the leaching process. The maximum concentrations of Fe and Mn were 0.32 and 40 μg/L, respectively. The structural diversity and abundance of the microbial communities in the groundwater were closely related to the Fe/Mn content. Actinobacteriota, Proteobacteria, Acidobacteriota, Bacteroidota, and Chloroflexi were the dominant phyla, while Rhodococcus, Ochrobactrum, and Pseudarthrobacter were the dominant genera. These functional microbes are actively involved in the biogeochemical cycling of Fe, Mn, and N. In summary, contaminants and the microbial-community structure have dual effects on the release of Fe and Mn from RBF aquifers.

1. Introduction

As an important interface and natural filter of surface water and groundwater systems, riverbank filtration (RBF) is an active biogeochemical transition zone [1]. The presence of certain pollutants in natural RBF systems enhances the heavy-metal-release process in sediments [2]. Iron (Fe) and manganese (Mn) are widely distributed in aquifer sediments, and they can enter groundwater through leaching and become common water-chemistry components [3]. High Fe and Mn concentrations in groundwater affect the purification efficiency of RBF. The concentrations of Fe and Mn in groundwater are highly spatially variable, which is related to the geochemical processes of Fe and Mn in aquifer sediments [4,5]. The superposition of exogenous contaminants and chemical components in the aquifer may complicate hydrogeochemical processes [6]. Therefore, it is necessary to understand the flow and release of Fe and Mn from aquifers.
Fe and Mn in groundwater are mainly derived from aquifer sediments and anthropogenic activities [7]. This not only pollutes groundwater to varying degrees, but it also alters the environmental conditions of the groundwater, thus affecting the natural environmental processes of Fe and Mn in groundwater [8,9]. In recent years, it has been shown that the Fe and Mn concentrations are usually higher in groundwater with higher ammonia and organic-matter contents, for example, in the aquifers in the United States [10], China [11], and some other regions [12]. It is worth noting that the premise of the above study was the observation under homogeneous redox zones, and since redox zones are usually inhomogeneous in natural environments, a newer method of aquifer redox zoning was proposed to compare the redox zoning and morphology of Fe and Mn in aquifers by fusing the Honduras aquifer data; as a result, anomalous concentrations of Fe and Mn were also detected [13]. In addition, at the site scale, such as in landfills, the Fe and Mn concentrations in the groundwater significantly increase after a period of operation, while the background values in the groundwater are not high [11]. Various organic and nitrogen-containing substances are among the main pollutants in these sites [14]. Agriculture, urbanization and industrialization determine the nitrogen pollution of soils, surface water and groundwater, and hydrological and biological processes between rivers and aquifers can control the concentration and fate of nitrogen compounds. In groundwater multilayer systems, the interaction between groundwater and surface water is influenced by changes in river flow and morphology, which can have an impact on the global nitrogen cycle system [15]. In addition, the detection of high concentrations of Fe and Mn in groundwater due to the presence of organic matter is usually associated with the occurrence of redox processes. In the hydrogeological context, multi-layered alluvial aquifers are directly connected to surface waters heavily affected by eutrophication, and redox processes involving organic loads in groundwater control the distribution and reaction rates of Fe and Mn [16]. Thus, formation of Fe- and Mn-containing groundwater is the result of a combination of a series of water–rock interactions and the hydrogeochemistry.
The release of Fe and Mn from aquifers is influenced by the physicochemical properties and microbial response [17]. Microbes can change the morphology and potential parameters of Fe and Mn in aquifers, such as through redox reactions, which affect the circulation and distributions of Fe and Mn in groundwater [18]. Native iron-reducing bacteria (IOB) and manganese-reducing bacteria (MnOB) play important roles in the migration processes of Fe and Mn in aquifers [19]. Fe-reducing bacteria (IRB) and Mn-reducing bacteria (MRB) can reduce iron and manganese oxides through metabolic activities, making them susceptible to migration in groundwater [20]. In addition, sulfate-reducing bacteria can indirectly reduce Fe/Mn oxides, and sulfate-reducing bacteria can stimulate the reaction of SO42− in sediments with Fe/Mn minerals to form synthetic minerals such as Fe-S/Mn-S, which may also inhibit some Fe/Mn migration [21]. However, the effects of the microbial community structure and function on Fe and Mn release from RBF sediment and their potential mechanisms are not clear. Therefore, the effect of the coupled processes of exogenous pollutant species and microbes on the behavior of Fe and Mn in the actual RBF system requires further investigation.
In this study, an anaerobic column experiment simulating the RBF system was developed to investigate the effects of ammonium ions and organic matter (COD) on the mobility of Fe and Mn in aquifer sediments. The objectives of this study were as follows: (1) to determine the enhancement effects of ammonium and COD in groundwater on the release of Fe and Mn from aquifer sediments through time-series analysis, (2) to analyze the structure and dynamics of the response processes of the microbial community based on Illumina MiSeq high-throughput sequencing, and (3) to reveal the mechanisms behind these effects. This study will aid in better understanding the range of biogeochemical effects of contaminants on Fe in groundwater systems, and we hope that the results of this study will lead to more effective control of groundwater pollution.

2. Materials and Methods

2.1. Sediment Sampling and Geochemical Analysis

The aquifer sediments were sampled in Northeast China at a specific point (45°50′ N, 126°40′ E) located east of the Songnen Plain, which belongs to the Songhua River Basin. One hole was drilled in the study area in July 2021 to collect aquifer sediment samples from 0 to 2 m, 5 to 6 m, 10 to 11 m, 15 to 16 m, and 20 to 21 m below the surface, and the lithological variation (sand, clay, or silty clay) was recorded. The physicochemical properties of the sediments are given in Table 1, and the test methods and detection limits for the physical and chemical properties of the sediments are given in Table 2. The in situ sediment samples were stored in sterile tubes, kept frozen in a cold environment, and transported to the laboratory within 3 days for storage in a −70 °C refrigerator for microbial-community analysis.

2.2. Column Experiment

A one-dimensional column system was constructed to simulate the roles of ammonium and COD in the release of Fe and Mn from the sediments to groundwater. The sediments were air-dried and sieved to a particle size of less than 2.0 mm before the column experiments. Three polyvinyl chloride columns with an inner diameter of 8 cm and a height of 80 cm were designed (Figure 1), and called column A (blank control), column B (ammonium solution), and column C (COD solution). A total of five outlets were designed, with each column filled to a height of approximately 60 cm with a sediment weight of approximately 1.2 kg. The top and bottom of the column were lined with 1 cm of quartz sand, and a layer of glass wool was placed at the bottom of the quartz sand at the lower end of the column. The sediment was tapped several times with a pounding stick after each filling. After filling, the outer wall of the sediment column was wrapped with a shade cloth, which was used to simulate the original light-proof environment of the underground aquifer.
The column experiments were conducted at room temperature (20 ± 3 °C) and protected from light to analyze the migration patterns of sediment Fe and Mn during contaminant fixed-head drenching tests. Before the experiment was started, ultrapure water (column A), 50 mg/L ammonium solution (column B), and 50 mg/L COD solution (column C) were injected into the columns at a flow rate of 2.5 × 10−5 cm/s, and the columns were protected from light for treatment. Ammonium and COD for the leaching test were artificially prepared with NH4Cl and potassium hydrogen phthalate, respectively. The sampling times were 0 h, 3 h, 5 h, 9 h, 18 h, 20 h, 24 h, 30 h, …, 742 h (31 days). The samples from the sampling points were analyzed, and the experiments were stopped when the pollutant concentrations at the outlets stabilized. The detected parameters were the pH, oxidation-reduction potential (Eh), dissolved oxygen (DO) concentration, ammonium concentration, Fe concentration, and Mn concentration. Owing to the special characteristics of DO, the DO concentration was measured immediately after sample collection. The other parameters were measured after storing the sample in a low-temperature environment. The solution used to determine the Fe and Mn concentrations was stored with a protective agent.

2.3. Aqueous-Phase Analysis

Approximately 5 mL of the existing water solution was collected from sampling ports 1–5 on the column with a sediment sampler. The samples were filtered through a 0.45-μm membrane, and the solutions for measurement of the Fe and Mn concentrations were preserved with acid. The pH, DO concentration, and Eh were measured by a water quality parameter meter (TE-600Plus, USA). The Fe and Mn concentrations were determined by inductively coupled plasma atomic emission spectrometry (Optima 8000, PerkinElmer, Waltham, MA, USA) [22]. Determination of the Fe(II) concentration was performed by the 1,10-o-phenanthroline method with a detection limit of 5 μM [23]. The difference between the total Fe concentration and Fe2+ concentration was used as the concentration of Fe3+ in the aqueous phase of the effluent. The ammonium concentration was measured by ultraviolet spectrophotometry.
All of the reagents were analytical grade. All of the glassware was carefully treated before use to ensure that the data were of high quality. The glassware was immersed in 10% HCl for 4 h, rinsed three times with ultrapure water, and then dried at 180 °C for 4 h.

2.4. Microbial Community Analysis

Sediment samples at five depths (20, 40, 60, 80, and 100 cm) of columns A–C were collected, which are called A1–A5, B1–B5, and C1–C5, respectively. Total DNA extraction from the microbial community was performed according to the instructions of the E.Z.N.A. soil DNA kit (Omega Bio-tek, Norcross, GA, USA), the quality of DNA extraction was determined by 1% agarose gel electrophoresis, and the DNA concentration and purity were determined by a NanoDrop2000 spectrophotometer. The 16S rRNA gene V3–V4 variable region was amplified by polymerase chain reaction (PCR) using 338F (5′-ACTCCTACGGGAGGCAGCAG-3′) and 806R (5′-GGACTACHVGGGTWTCTAAT-3′). PCR amplification of the variable region was performed as follows: 95 °C pre-denaturation for 3 min, 27 cycles of 95 °C denaturation for 30 s, 55 °C annealing for 30 s, and 72 °C extension for 30 s, 72 °C stable extension for 10 min, and finally storage at 4 °C (ABI GeneAmp 9700 PCR system, ABI, USA). The PCR reaction system was 4 μL of 5× TransStart FastPfu buffer, 2 μL of 2.5 mM dNTPs, 0.8 μL of upstream primer (5 μM), 0.8 μL of downstream primer (5 μM), 0.4 μL of TransStart FastPfu DNA polymerase, and 10 ng of template DNA (made up to 20 μL). There were three replicates per sample.
The PCR products from the same sample were mixed and recovered using 2% agarose gel, purified using an AxyPrep DNA gel extraction kit (Axygen Biosciences, Union City, CA, USA), detected by 2% agarose gel electrophoresis, and quantified with a Quantus fluorometer (Promega, Madison, WI, USA). Library construction was performed using a NEXTFLEX Rapid DNA-Seq kit with the following steps: (1) splice linkage, (2) removal of the splice self-linked fragments by magnetic-bead screening, (3) enrichment of the library template by PCR amplification, and (4) recovery of the PCR products by magnetic beads to obtain the final library. Sequencing was performed using Illumina’s Miseq PE300/NovaSeq PE250 platform (Shanghai Meiji Biomedical Technology Co., Ltd., Shanghai, China).

3. Results and Discussion

3.1. Changes of pH, DO and Eh

In the control experiment (column A) (Figure 2, Table S1), the average leachate pH was 7.1, and the pH first decreased and then slowly increased with increasing infiltration time. The pH decreased to 5.92 at sampling port 3, and then the pH basically stabilized. With continuous introduction of ultrapure water, the pH of the column increased after 576 h. In the ammonium test (column B) (Figure 2), the average pH of the leachate was 6.2. With increasing infiltration time, the pH first decreased and then slowly increased, and the pH fluctuated before 432 h. The pH increased to 7.12 at 48 h of the experiment, and it then decreased to 5.8 at 144 h. During this period, ammonium underwent a hydrolysis reaction to produce H+, so the pH briefly decreased. After 432 h of the experiment, the reaction in the column tended to be stable, so there was no significant change in the pH in the later stage. In the COD test (column C) (Figure 2), the average pH was 6.2, and the pH fluctuated and then reached a stable state after 430 h. The pH increased at 300 h. The pH was the lowest at the beginning of the test, because the COD solution was acidic (pH = 5–6). Overall, compared with the ultrapure water group, the pH values in the ammonium and COD tests were lower, the pH values of the sampling points were similar, and they all reached a stable state after 430 h.
In column A (Figure 2), the average DO concentration in the leachate was 10.2 mg/L. The DO concentration first decreased, then increased, and then decreased, and the DO concentration fluctuated in the different water-extraction holes. Before 96 h, the DO concentration showed a decreasing trend, during which the column consumed O2. After 300 h, the DO concentration in the column began to stabilize. After 528 h, the DO concentration greatly fluctuated, which may be due to a change in the DO concentration caused by contact with air at the bottom of the column. In column B (Figure 2), the average DO concentration was 12.8 mg/L. The DO concentration fluctuated before 144 h, indicating a strong reaction in the column. In column C (Figure 2), the average DO concentration was 13.2 mg/L, and the DO concentration showed a downward trend before 48 h. During this period, aerobic microbes were relatively active in the column. After 48 h, the DO concentration in the column began to decrease and was then relatively constant at 300 h, indicating that the reaction was not intense during this period. The DO concentrations in the ammonium and COD tests were higher than that in the control test.
In column A (Figure 2), the average Eh of the leachate was 210 mV. Eh showed the trend of first decreasing, then increasing, and then decreasing. Eh is highly susceptible to the influence of redox substances. Eh greatly fluctuated before 324 h, and complex reactions occurred in the column during this period. After 324 h, Eh tended to a relatively stable state, indicating that the column had reached a relatively balanced state. In column B (Figure 2), the average Eh was 222 mV. Before 276 h, Eh greatly fluctuated, with obvious biochemical effects and strong microbial metabolic processes. After 276 h, the column tended to a relatively stable state. In column C (Figure 2), the average Eh was 212 mV. The lowest Eh was 178 mV at 96 h, the highest Eh occurred between 200 and 400 h, and Eh tended to be stable at 670 h.

3.2. Changes of Fe and Mn

With the increase in infiltration time, the concentration of Fe and Mn in leachate increased to the highest value, and then decreased gradually (Figure 3, Table S2). The concentration trends of Fe and Mn in different leaching depths were similar, but the concentration values were different. In column A, Fe was released faster before 48 h (Figure 3), and the maximum Fe concentration was 0.051 mg/L at 60 h for sampling point 1. For sampling points 2–5, the maximum Fe concentrations of 0.06, 0.07, 0.081, and 0.078 mg/L were reached at 140 h, respectively, and they then gradually decreased. The concentration of Fe in the leachate increased with increasing infiltration distance, and the mean Fe concentrations for sampling points 1–5 were 0.033, 0.039, 0.049, 0.05, and 0.053 mg/L, respectively. At 140 h, the Mn concentrations reached the maximum concentrations of 4, 4.5, 4.5, 4.1, and 5.9 μg/L, respectively (Figure 3). The Mn concentrations then gradually decreased. The average Mn concentrations for sampling points 1–5 were 2, 2, 3, 4, and 4 μg/L, respectively, and the Mn concentrations increased with increasing infiltration distance.
In column B, Fe was released at a fast rate before 50 h (Figure 3), and the concentration of Fe released at sampling point 1 reached a maximum of 0.12 mg/L at 144 h. At sampling points 2–5, the maximum Fe concentrations of 0.2, 0.28, 0.22, and 0.28 mg/L occurred at 240 h, respectively, and they then gradually decreased and reached a stable state at 576 h. The average Fe concentrations for sampling points 1–5 were 0.073, 0.11, 0.156, 0.131, and 0.148 mg/L, respectively, and the Fe concentrations increased with increasing infiltration distance. The Mn release rate was slow before 144 h (Figure 3), and the Mn concentration did not significantly change. The maximum Mn concentration was 0.03 mg/L at 525 h for sampling point 1. For sampling points 2 and 5, Mn showed the maximum concentrations of 0.068 and 0.08 mg/L at 550 h, respectively. The average Mn concentrations for sampling points 1–5 were 11, 16, 14, 13, and 10 μg/L, respectively, and the Mn concentrations decreased with increasing infiltration distance.
In column C, Fe was released at a fast rate before 48 h (Figure 3), and the maximum Fe concentrations for sampling points 1, 2, and 3 were 0.2, 0.28, and 0.34 mg/L at 150 h, respectively. The maximum Fe concentrations for sampling points 4 and 5 were 0.31 and 0.28 mg/L at 200 h, respectively, and they then gradually decreased and reached a stable state at 380 h. The average concentrations of Fe for sampling points 1–5 were 0.086, 0.131, 0.122, 0.114, and 0.084 mg/L, respectively. The concentration of Mn fluctuated before 144 h (Figure 3). For sampling point 1, the maximum Mn concentration of 25 μg/L occurred at 190 h. For sampling point 4, the maximum Mn concentration of 35 μg/L occurred at 240 h. The maximum Mn concentration for sampling point 5 was 30 μg/L at 576 h, and the Mn concentration then gradually decreased and reached a stable state at 380 h. The mean Mn concentrations for sampling points 1–5 were 11, 14, 14, 19, and 13 μg/L, respectively.
The leaching trends of Fe and Mn in the three columns were similar. The concentrations of Fe and Mn rapidly increased and reached peaks in the early stage of leaching, and they then gradually decreased until they reached a stable state in the middle to late stages of leaching. Ammonium and COD can enhance the release of Fe/Mn in aqueous media. Compared with the ultrapure water group, the average release times of Fe/Mn were 2.8/10 and 3/7 times higher for columns A and B, respectively. By comparing the five sampling points, the leaching concentration order of Fe and Mn was 5 > 4 > 3 > 2 > 1.
Areas with high Fe and Mn contents in groundwater are accompanied by ammonium pollution, which has attracted great attention. For example, the chemical data of 43,334 wells in the United States were tested [24], and the results showed that high Mn concentration (>300 μg/L) was present in the groundwater. The potential value of the redox reaction of Fe, Mn, and ammonium is related to the order of the reaction speed. Therefore, under limited conditions, Fe can first undergo an oxidation reaction [25]. In the reaction process, the ammonium reaction is incomplete, resulting in the accumulation of nitrite ions. Ammonium solutions are weakly acidic (pH = 5.5–6.5), and the sediment will dissolve during leaching. In addition, the migration processes of Fe and Mn in the column are controlled by the adsorption–desorption equilibrium, and this equilibrium moves towards desorption when the concentration increases.
The ammonium concentration of the whole sediment column first decreased and then increased (Figure 4). The decrease in the ammonium concentration can be attributed to heterotrophic denitrification and dissimilatory nitrate reduction [26]. Ammonium-coupled Fe2+ oxidation is the main pathway of nitrate reduction under oxidizing conditions [27,28]. The Fe3+ content in the column was higher than the Fe2+ content. In addition to rapid oxidation of Fe2+ to Fe3+ owing to the instability of Fe2+, Fe3+ can also form by the reaction between Fe2+ and ammonium in the column. Mn and Fe have similar oxidation characteristics. We speculate that nitrification driven by Mn2+ oxidation is the main reason for the decrease of the ammonium concentration, which is confirmed by the fact that dissolved reduced Mn can be used as a reducing agent for nitrogen oxide species [29]. The increase in the ammonium concentration may be due to the fact that the reaction in the sediment column tends to be stable in the later stage of the experiment, and the content of pollutants reacting with Fe and Mn oxides in the sediment becomes less. Compared with the change in the ammonium concentration, the COD concentration did not significantly change. Previous studies have shown that COD has certain effects on the dissolution of minerals [30], and COD can act as a reducing agent or complex formal ligand. Because COD is a good metal-chelating agent, it can provide electrons to promote reduction and dissolution of minerals [31]. COD creates a more suitable environment for the presence of reducing substances (such as Fe and Mn) by decreasing Eh in groundwater [32].

3.3. Response of Microbial Community

Microbial community sequence analysis revealed a series of groups associated with Fe-, Mn-, ammonium-, and organic-related processes. At the phylum level (Figure 5, Table S3), in column A, Actinobacteriota, Proteobacteria, Acidobacteriota, and Chloroflexi were the dominant phyla, with relative abundances of 23.1–40.5%, 11.4–45.5%, 1.4–15.2%, and 2.7–14.1%, respectively. Among the top ten bacterial groups, Actinobacteriota, Proteobacteria, Acidobacteriota, Firmicutes, and Bacteroidota were the common IOB and MOB. Their cumulative abundance was 64.9%. In column C, Proteobacteria, Actinobacteriota, Bacteroidota, and Acidobacteriota were the dominant phyla with relative abundances of 16.7–54.4%, 15.2–42.1%, 3.8–7.0% and 2.8–6.3%. The cumulative relative abundance of IOB and MOB increased by 74.7%, and organic-degrading bacteria (1.3–3.6%) appeared. Compared with column A, the cumulative relative abundance of IOB and MOB in columns B and C increased by 19.5% and 15.1%, respectively, indicating that ammonium and COD promote the release of iron. At the genus level, Rhodococcus, Ochrobactrum, and Pseudarthrobacter were common IOB and MOB in column A, with a cumulative abundance of 20.8%. The relative abundances of IOB and MOB in columns B and C were 23.3% and 30.3%, respectively, which were 12.0% and 45.6% higher than that in column A, respectively. This verified that ammonium and COD promote the release of Fe and Mn.

3.4. Fe and Mn Release Mechanism

The changes of the Fe and Mn concentrations in the sediments in ammonium and COD solution, and the structure of the microbial community indicate that microbe-mediated reductive dissolution of Fe/Mn oxides may still be the main mechanism of Fe/Mn release in the system (Figure 6). With the addition of a large number of pollutants, the DO was consumed in large quantities, and the microbial community in the system was mainly composed of anaerobic/supplementary anaerobic bacteria, in which IRB/MRB played an important role. The metabolic activity of Acidobacteriota affects the pH of the environment and the dissolution of Fe and Mn minerals through a change in the environmental conditions, and Fe and Mn minerals are then reduced under the action of IRB and MRB. In the presence of ammonium, Nitrospirae, Denitratisoma, and Nitrosomonas enhance the process of Fe release by nitrification and denitrification of NH4+ [33,34]. Through the reductive dissolution of Fe/Mn minerals, Fe and Mn attached to the surface of iron and manganese oxides are released and reduced to Fe2+ and Mn2+ under the action of IRB and MRB in the aqueous environment [35]. In the presence of COD, IRB and MRB enhance the Fe- and Mn-release processes by using organic matter, and organic-matter-degradation bacteria can enhance the Fe-release process [36]. In addition, Desulfobacterota (an anaerobic microorganism) was detected in the COD test, and S can be used as a strong reducing agent to participate in the reaction of iron oxide in the system, accelerating the reduction and release of iron/manganese oxides [37,38], while the Fe/Mn and S elements coexist. The S2− or HS− ions produced by reduction can combine with Fe2+ in water to form secondary minerals, which can inhibit the increase in the Fe/Mn concentration in groundwater.

4. Conclusions

The effect of anthropogenic pollutants on the release of Fe and Mn from sediments has been investigated by column experiments. The results showed that ammonium and COD significantly enhance the release of Fe and Mn from sediments to groundwater. The leached concentrations of Fe and Mn were similar, and the maximum leached concentrations were 0.32 and 40 μg/L, respectively. The diversity and abundance of the microbial community structure was closely related to the contents of Fe and Mn. Actinobacteriota, Proteobacteria, Acidobacteriota, Bacteroidota, and Chloroflexi were the dominant phyla, and Rhodococcus, Ochrobactrum, and Pseudarthrobacter were the dominant genera in the sediments. These microbes not only promote the release of Fe/Mn by reducing iron and manganese minerals, but they also actively participate in the biogeochemical cycles of Fe and Mn.
This study provides a good explanation for why the concentrations of Fe and Mn in groundwater spatially vary in the natural environment. The results show that the discharge of ammonium and COD can increase the concentration of Fe and Mn in groundwater. Therefore, more attention should be paid to the importance of ammonium ions and organics in the geochemical processes of Fe and Mn in sediments, which will facilitate the further development of groundwater pollution control measures.

Supplementary Materials

The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/w15010120/s1, Table S1: The average value of pH, DO concentration and Eh over time; Table S2: The average value of Fe and Mn concentration over time; Table S3: The proportion of microbial community abundance in sediment samples of different soil columns.

Author Contributions

Methodology, Y.T.; formal analysis, X.X.; investigation, Y.Z.; data curation, Y.T.; writing—original draft, X.X.; writing—review and editing, X.X.; supervision, Y.T.; funding acquisition, Y.Z. All authors have read and agreed to the published version of the manuscript.

Funding

This work was supported by the National Natural Science Foundation of China (No. 41877355, 42077170 and 41831283), Beijing Advanced Innovation Program for Land Surface Science, and the 111 Project of China (No. B16020).

Data Availability Statement

Not applicable.

Conflicts of Interest

The authors declare no conflict of interest.

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Water 15 00120 g001 550
Figure 1. Schematic diagram of the column system.
Figure 1. Schematic diagram of the column system.
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Figure 2. Changes of the pH, DO concentration, and Eh with time.
Figure 2. Changes of the pH, DO concentration, and Eh with time.
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Figure 3. Changes of Fe and Mn concentrations.
Figure 3. Changes of Fe and Mn concentrations.
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Figure 4. Changes of the concentrations of ammonium and COD in the column.
Figure 4. Changes of the concentrations of ammonium and COD in the column.
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Figure 5. Histogram of the species abundance in the microbial community.
Figure 5. Histogram of the species abundance in the microbial community.
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Figure 6. Effects of microbes, ammonium, and COD on the release of Fe and Mn from sediments.
Figure 6. Effects of microbes, ammonium, and COD on the release of Fe and Mn from sediments.
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Table
Table 1. Physical and chemical properties of the aquifer sediments used for the experiment.
Table 1. Physical and chemical properties of the aquifer sediments used for the experiment.
Sampling Site (m)Depth *
(m)		Sediment
Texture		pHFe
(g/Kg)		Mn
(g/Kg)		TOC
(%)		NH4+
(mg/kg)
0–21–2Clay5.257.360.42211.5
5–65–6Clay7.8715.400.38521.3
10–1110–11Clay and sand7.869.60.29120.9
15–1615–16Sand6.312.80.2081.3
20–2120–21Clay and sand6.615.400.1871.2
Note(s): * Refers to the buried depth of the sampling site underground.
Table
Table 2. Methods for determining the physicochemical properties of the sediment and the detection limits.
Table 2. Methods for determining the physicochemical properties of the sediment and the detection limits.
Item (Unit)Detection MethodDetectable Limit
pH (/)Glass electrode method–
NH4+ (mg/L)Gas phase molecular absorption spectrometry0.15 mg/kg
TOC (mg/L)Combustion method–
Fe (mg/L)ICP-AES (PerkinElmer Optima 8000)0.06 mg/kg
Mn (mg/L)ICP-AES (PerkinElmer Optima 8000)0.02 mg/kg
Note(s): “–” means that there is no detection limit.
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Xia, X.; Teng, Y.; Zhai, Y. Effects of Ammonium and COD on Fe and Mn Release from RBF Sediment Based on Column Experiment. Water 2023, 15, 120. https://doi.org/10.3390/w15010120

AMA Style

Xia X, Teng Y, Zhai Y. Effects of Ammonium and COD on Fe and Mn Release from RBF Sediment Based on Column Experiment. Water. 2023; 15(1):120. https://doi.org/10.3390/w15010120

Chicago/Turabian Style

Xia, Xuelian, Yanguo Teng, and Yuanzheng Zhai. 2023. "Effects of Ammonium and COD on Fe and Mn Release from RBF Sediment Based on Column Experiment" Water 15, no. 1: 120. https://doi.org/10.3390/w15010120

APA Style

Xia, X., Teng, Y., & Zhai, Y. (2023). Effects of Ammonium and COD on Fe and Mn Release from RBF Sediment Based on Column Experiment. Water, 15(1), 120. https://doi.org/10.3390/w15010120

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