1 Introduction

Phosphorus (P) is an important environmental factor driving the steady-state transformation of aquatic ecosystems and is a significant limiting factor for water eutrophication (Wu et al. 2022). The primary sources of P in water encompass external input, internal release of sediment, and a small amount of atmospheric deposition. Some management and engineering measures have effectively controlled the input of external P. However, water eutrophication remains unresolved, which can mainly be caused by the internal P release (Li et al. 2023c). Studies have suggested that persistent P release from sediments can delay lake recovery by approximately 10 years (Søndergaard et al. 2012). Therefore, it makes sense to find a way to reduce internal P releases.

In situ capping is regarded as an effective remediation strategy to control water eutrophication due to its relatively modest cost and considerable remediation effectiveness (Pan et al. 2012; Sun et al. 2023). Biochar is produced through the thermal decomposition of biomass in the oxygen (O2)-restricted condition, and is regarded as an outstanding in situ capping material (Abel and Akkanen 2018; Zhu et al. 2019a). With its high porosity, numerous functional groups, and adsorption sites, biochar can immobilize contaminants via processes such as physical adsorption and redox reactions (Silvani et al. 2017; Zhao et al. 2022). A mount of research has found that biochar capping can effectively reduce the release of nitrogen from sediments using physical barriers, chemical adsorption and microbial transformation (Zhu et al. 2019a, b). Additionally, biochar has successfully remediated contaminated sediments by reducing the mobility of various pollutants (Yang et al. 2021b), such as organic compounds (Abel and Akkanen 2018), heavy metals and metalloids (Qian et al. 2022). Although biochar is often regarded as ‘black gold’ in ecological remediation engineering (Tan and Yu 2024), research on the use of biochar capping for sediment P remediation has only just been conducted (Gao et al. 2021). The effect of biochar capping on the bioavailability and mobility of P in sediments is not yet known.

Recently, studies have reported that biochar application can improve P availability in soils (Yang et al. 2021a). The water-soluble P content of most soils increased by 11–253% when biochar was added at 1% (w/w) (Parvage et al. 2012). Biochar has a high P content of 15–11,600 mg kg−1 due to the loss of volume of the biomass by 60–90% via the high-temperature carbonization process (Parvage et al. 2012; Sun et al. 2018). Therefore, it can act as a source of available P. In addition, biochar-derived dissolved organic matter (BDOM), a highly reactive product, plays an important role in the transport and transformation of P (Qi et al. 2023). BDOM may affect P mobilization in two ways. First, BDOM has a lot of reactive groups, such as ROH and RCOOH, which can replace adsorbed P on minerals (Schneider and Haderlein 2016). Second, BDOM can drive the reduction of Fe oxide as an electron shuttle, subsequently releasing P (Yang et al. 2021a). Consequently, we hypothesize that biochar addition may accelerate the P desorption from the sediment solid phase, which ultimately leads to internal P release, thus exacerbating water eutrophication. It is therefore crucial to accurately assess the impact of biochar capping on sediment P release.

Traditional chemical extraction procedures cannot provide comprehensive information about the P adsorption and desorption in sediments (Li et al. 2021, 2023b). Moreover, conventional sediment porewater sampling methods, sediment core slicing, centrifugation and subsequent analysis are not only generally limited in their spatial resolution, but also may change the analyte species in the porewater (Li et al. 2022b; Pedersen et al. 2015). Recently, multiple high-resolution sampling techniques have emerged that can compensate for these shortcomings, including diffusive gradients in thin films (DGT) (Davison and Zhang 1994), and high-resolution dialysis sampler (HR-Peeper) (Chen et al. 2022). The resupply kinetics of solutes can be characterized by combining these high-resolution sampling techniques (Zhang et al. 2020a). Therefore, fine-scale measurement of P mobilization processes can be achieved using multiple high-resolution sampling techniques.

Accordingly, this study aimed to assess how biochar remediation affects the bioavailability and mobility of P in sediments using high-resolution sampling techniques and microbiome analysis methods. Our findings provide a more comprehensive insight into the influence of biochar capping on the bioavailability and mobility of P in sediments, thus providing a scientific basis for the optimization of sediment restoration techniques.

2 Materials and methods

2.1 Incubation experiment setup

The water and sediment were sampled from an area of eutrophication in Lake Taihu (Fig. S1, Tables S1 and S2) (Duan et al. 2023). The sediment was sifted through a 0.6 mm mesh to remove macrofauna and large particles, and subsequently mixed evenly. The water sample was filtered through a diaphragm vacuum pump (0.45 µm pore size). The homogenized sediments were transferred into 21 Plexiglas tubes (11 cm in diameter and 50 cm in height) and 3 transparent rhizoboxes (height × length × width: 30 cm × l10 cm × 50 cm) with detachable front windows (Fig. 1). The height of sediment was set at 25 cm in the Plexiglas tubes, with overlying water added to reach 20 cm above the interface between water and sediment (SWI). The sediment height in the rhizoboxes was 15 cm and overlying water was added at 10 cm SWI. Both the tubes and the rhizoboxes were precultured at 25 °C for 2 weeks before the addition of biochar (Fig. 1).

Fig. 1
figure 1

Microcosm experiment design

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Biochar (Runkelvyuan Biological Engineering Co., Ltd, Shandong, China) produced from fruit tree branch was applied in this experiment, which was obtained at 420 °C for a duration of 1 h in a muffle furnace with limited oxygen conditions. Biochar was processed to pass through 80-mesh sieves before use. The detailed parameters of the biochar in this study are provided in Table S3. The sediment cores were treated in three ways: the control group (no biochar), 1-cm biochar capping group and 2-cm biochar capping group, respectively. Using the field capping restoration approach, pre-treated biochar was sprayed into the water column and settled into the sediment, forming a capping layer of approximately 1 or 2 cm (n = 9, respectively). All sediment tubes were cultured for 80 days at 25 °C and overlying water was also added continuously to maintain the 20 cm height during incubation. The sediment samples used to determine the P fractions and for DNA extraction were collected on day 80.

2.2 Measurement of the overlying water samples

A total of 17 days of overlying water physicochemical measurements, including SRP, UV254, dissolved oxygen (DO), and pH were conducted in Plexiglas tubes during the 80-day experiment. For the SRP and UV254, the water samples (approximately 1 mL) were obtained by the injector and subsequently filtered (0.45 μm pore size). The SRP and UV254 were determined immediately. SRP concentration was determined using molybdenum blue method by microplate spectrophotometer (Ren et al. 2020). UV254, as an alternative indicator for dissolved organic material (DOM), was determined using a quartz microtiter plate and a microplate spectrophotometer to estimate the DOM concentration (Zhang et al. 2018). The microplate spectrophotometer has been widely used for the detection of trace water samples. In addition, DO and pH in the water column were measured using a portable parameter meter (DPL-LS12) inserted directly into the water.

2.3 Acquisition of high-resolution data at the sediment–water interface

2.3.1 2D-distribution of O2 and pH in sediments

The distribution of pH and O2 in the SWI can be mapped using two-dimensional (2D) imaging techniques using pH and O2 planar optode, respectively (Li et al. 2019). Briefly, pH planar optodes (dimension: 10 cm × 5 cm) and O2 planar optodes (dimension: 10 cm × 5 cm) were affixed to the transparent detachable windows of rhizoboxes. pH and O2 imaging were performed before biochar capping, on the 1st, 2nd, 3rd, 15th, and 30th day after biochar capping. The imaging system PO2100 was provided by Easysensor Ltd (Nanjing, China). Considering the pH optodes lifetime (~ 7 days), the pH optodes were replaced on the 14th and 29th day. Fabrication and calibration information for pH and O2 planar optodes is provided in the Supplementary Materials.

2.3.2 1D-distribution of SRP/soluble Fe and labile P/Fe

The HR-Peeper sampler was used to collect the porewater for determining the SRP and soluble Fe on the 0th, 30th, and 80th days. Additionally, the labile P and Fe in the sediments were determined using ZrO-Chelex DGT on those same days. Both they were deoxygenated with nitrogen (N2) for 16 h before use. They were deployed in the sediment for 48 h and 24 h, respectively. Treatment of the HR-Peeper and Zr-oxide DGT probes was carried out according to the methods of Ma et al. (2020). Detailed operational processes can be found in the Supplementary Materials. Labile P was detected using the same method as that used for SRP. Fe was detected using the phenanthroline colorimetric method (Ren et al. 2020).

2.3.3 2D-distribution of soluble Fe(II) and labile S(−II) imaging

On day 30 and day 80, the 2D-distribution of soluble Fe(II) and labile S(−II) in sediments was determined using the combined DET/DGT probe according to the method of Robertson et al. (2008). To minimize solute diffusion in the gel, the DET/DGT probe was prepared by immobilizing the DET gel and DGT gel on both sides of the polyethylene terephthalate (PET) plate in a back-to-back manner. The DET/DGT probe was deployed in the sediments for 24 h. The DET staining gel was soaked in Ferrozine reagent for at least 12 h. After retrieval, the staining gel was placed on the DET gel, forming a stack of sheet/DET gel/staining gel/sheet. After 15 min coloration, the stack was scanned at 600 dpi resolution. The AgI DGT gel was briefly rinsed and directly scanned using a flat-bed scanner (Canon 5600F). The soluble Fe and labile S(−II) could be quantified based on the color change in the gel using computer-imaging densitometry techniques (Wang et al. 2019). More information on the DGT gel and DET gel fabrication and calibration processes is provided in Supplementary Materials.

2.4 Measurement of the sediment sample

DNA extraction, PCR amplification and sequence analysis were performed at Shanghai Personal Biotechnology Co., Ltd (Shanghai, China). DNA was extracted from freeze-dried homogenized sediment samples using the DNeasy PowerSoil Kit (QIAGEN, Inc., the Netherlands). The primers 515F and 907R (5ʹ-GTGCCAGCMGCCGCGGTAA-3ʹ and 5ʹ-CCGTCAATTCMTTTRAGTTT-3ʹ, respectively) were used for the PCR amplification of 16 S rRNA in the V3–V4 region. The analysis of amplicon sequences was conducted using Qiime2 (https://qiime2.org/). Specific details regarding the processing of microbial data can be found in Ma et al. (2023). The raw DNA sequencing data has been submitted to the NCBI SRA (PRJNA1047949).

On day 80, the sediment cores were sectioned under nitrogen protection conditions, and freeze-dried for further analysis of the P fraction and alkaline phosphatase activities (ALP). The P fraction included exchangeable P (NH4Cl-P), Fe-adsorbed P (BD-P), Al-adsorbed P (NaOH-P), organic P (Org-P), Ca-bound P (HCl-P), and residual P (Res-P). Detailed extraction methods for inorganic and organic P can be found in Janssen et al. (2021). The ALP activities in the sediments were measured using an S-AKP kit (Boxbio, China).

2.5 Data analysis

The pH and DO optodes were calculated according to color ratiometric principles (Li et al. 2022a). The labile P and Fe concentrations were calculated according to (Chen et al. 2018). The net diffusion fluxes of soluble and labile P within 20 mm at SWI were calculated based on Fick's diffusion law (Wang et al. 2024). The grayscale of the DET gel images and DGT gel images were applied to determine the soluble Fe(II) concentration and labile S(−II) accumulation quantity (Li et al. 2022b). Details of the calculations are provided in the Supplementary Materials. The parameter R, which indicated the P resupply capacities from the solid phase to the porewater, was calculated from the ratio of the labile P and the SRP concentration (Wang et al. 2018).

The diffusion kinetics and resupply of P in sediments were revealed by the model of DGT induced fluxes in sediments and soils (DIFS) (Wu et al. 2018). The main output parameters of the model include K1 and K−1, representing the adsorption rate constant, and desorption rate constant, respectively (Menezes-Blackburn et al. 2016). The desorption rate was calculated according to Chen et al. (2022). Details of the calculations are provided in the Supplementary Materials (Table S4).

The alpha diversity indexes and principal coordinates analysis (PCoA) were determined using the “Vegan” package. In addition, the potential effects of environmental variables on the microbial community were determined using canonical correlation analysis (CCA). PICRUSt2 was used to predict the P-cycle function. According to KEGG orthologs, the P-cycle functional genes were classified into four types, namely inorganic P solubilization, organic P mineralization, P transportation and P regulation (Hu et al. 2023). All graphs were produced using Origin 2018, R v.4.1.2. and Surfer 13. Pearson correlation analysis and t-test were conducted in SPSS 25 to evaluate the association between parameters.

3 Results and discussion

3.1 Changes in pH, DO, UV254 and SRP in the incubation experiment

After biochar capping, the pH and redox conditions in the incubation system were significantly changed (Fig. 2 and Fig. S2). The average pH of the three different treatment groups was maintained at around 8.46, 8.53 and 8.65 for the control, 1-cm and 2-cm biochar capping groups, respectively. The water pH after biochar capping was significantly higher than the control group (p < 0.05), and it increased with the amount of biochar added (Fig. 2a). The increase in pH was more evident at the sediment-biochar interface, and the pH gradually decreased with depth (Fig. 2b). The 2-cm biochar capping led to a greater area of pH increase than 1-cm capping in the sediment. This phenomenon is consistent with previous findings at the soil-biochar interface (Buss et al. 2018). A large amount of inorganic ash and negatively charged functional groups in the biochar bound tightly to the H+ ions, leading to an increase in pH (Chintala et al. 2013). The DO in the overlying water decreased by approximately 70% after biochar capping (Fig. S2). It was initially decreased in the aqueous phase specifically at 0.5 cm above sediment and then increased to its original level on day 30. The concentration of organic matter was increased in the water column after biochar capping (Fig. 2a). The decomposition of organic matter consumed oxygen, which reduced the amount of DO in the water column (Cui et al. 2022). Additionally, the aerobic layer at the SWI decreased following the biochar capping.

Fig. 2
figure 2

a Changes in physicochemical properties in overlying water over time. b pH changes in the sediment–water profile over time

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Biochar capping caused a sharp increase in organic matter (UV254) and SRP concentration in the overlying water, which was associated with the amount of biochar capping (Fig. 2a). The UV254 increased from 0.0247 in control to 0.0494 and 0.0701 in the 1-cm and 2-cm capping groups, respectively. During the sediment incubation, the capping groups consistently maintained higher UV254 levels compared with the control group. As a carbon-rich solid material, biochar contains a significant amount of organic matter that can be dissolved in the water, resulting in higher dissolved organic carbon (DOC) concentration (Azeem et al. 2023). The SRP concentration changes in the water column were similar to those for UV254, with a sharp increase and a maximum value on the first day of biochar capping. The highest SRP concentrations were 1.11 mg L−1 and 2.10 mg L−1 in the 1-cm and 2-cm capping groups, respectively. The solubilization of dissolved P in biochar likely contributed to the elevated SRP concentrations in the water column (Padilla et al. 2023). Subsequently, the SRP concentration gradually decreased and returned to its original level after 20 days. This study revealed that biochar capping led to a sudden increase in organic matter and P concentration, which may adversely affect the water quality. Furthermore, biochar capping-induced changes in pH, redox conditions and UV254 at the sediment interface could affect the P transport and transformation processes in sediments.

3.2 Effects of biochar capping on P mobilization in SWI

The spatial distribution of soluble and labile P in the sediment profiles showed that the mobility and bioavailability of P can be altered by biochar capping (Fig. 3). On day 30, the profile-averaged SRP concentration of 1-cm capping group (0.87 mg L−1) was significantly higher than that of the control group (0.68 mg L−1) (p < 0.001), while there was no significant difference in the profile-averaged SRP concentration of 2-cm capping group (0.67 mg L−1) compared to the control group (p > 0.05) (Fig. 3a and Fig. S3). The changes in labile P were consistent with the SRP in the sediment profiles (Fig. 3b and Fig. S3). The 2-cm biochar capping can effectively control porewater SRP concentration at the surface, in particular in the surface sediment (0–50 mm), and the SRP concentration was significantly lower than that of the control and 1-cm capping groups (p < 0.05). After 30 days of the biochar capping, the K−1 value in the treatment groups increased significantly, while K1/K−1 values of the 1-cm capping group decreased and those of the 2-cm capping group increased (Table S5). The P desorption from solid was dominated in the 1-cm capping group, and the P adsorption from solution was dominated in the 2-cm capping group. The adsorption of phosphorus by biochar involves a rapid chemisorption phase, followed by a surface diffusion-controlled phase. Therefore, 2-cm biochar capping provided enough adsorption sites for SRP, leading to a decline in SRP concentration in the porewater in the short term. For 1-cm biochar capping, however, the surface adsorption sites of biochar were saturated and subsequent adsorption of P via intraparticle surface diffusion was also limited, which led to a higher desorption than adsorption of phosphorus by the biochar (Table S5). In addition, it was found that soluble and labile Fe decreased in the surface sediments (Fig. 3), which could be attributed to enhanced Fe adsorption by the biochar (Chen et al. 2022). Also, the significant positive correlation (p < 0.01) between Fe and UV254 indicated that organic matter released from the biochar may bind with Fe oxides, contributing to a decline in the adsorption of Fe oxides on P. Consequently, in the early stages of biochar capping, the adsorption–desorption processes of biochar on P are the key factors controlling the P mobilization.

Fig. 3
figure 3

a Distribution of soluble Fe, SRP and UV254 in the sediment–water profile on days 30 and 80, respectively. b Distribution of labile Fe, labile P and R in the sediment–water profile on days 30 and 80, respectively

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The effects of biochar aging on P mobilization were further investigated. On day 80, both the SRP and labile P concentration in the sediments significantly increased in the capping groups compared with the control group (p < 0.001) (Fig. 3a, b, and Fig. S3). The SRP concentration increased by 35.50% and 23.84% in the 1-cm and 2-cm biochar capping groups, respectively. Furthermore, the labile P concentration increased by 36.56% and 23.06%, respectively. The P desorption rate in the treatment groups increased significantly, especially in the 2-cm capping group (Table S5). The increases in P mobility and bioavailability could be attributed to the competitive organic matter adsorption and biochar aging. As biochar ages, more BDOM is released, which could facilitate P desorption through surface adsorption of Fe oxides (Xu et al. 2014). In addition, the surface of aged biochar is rich in oxygen-containing functional groups, which increases the negative charge density of the biochar (Soinne et al. 2014; Yang et al. 2021a). The resultant electrostatic repulsion can lead to the desorption of sorbed P on the biochar, thus increasing P mobility and bioavailability. Biochar capping also increased the anaerobic conditions in the deep sediments. The reduction of SO42− and the formation of FeS can also promote the release of sediment P (Chen et al. 2016). In this study, it was found that both soluble Fe(II) and labile S(−II) decreased as the capping time of the biochar increased (Fig. 4), resulting from the formation of FeS. Consequently, the upward flux of reduced Fe decreased, resulting in a reduced availability of Fe(II) for re-oxidation and P sequestration in the surface sediment, thereby reducing the P retention capacity.

Fig. 4
figure 4

Distribution of soluble Fe(II) and labile S(− II) in the sediment–water profile on a day 30 and b day 80

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Furthermore, it was found that 2-cm biochar capping decreased the release flux of SRP by 73.1%, while the 1-cm biochar capping significantly increased the release flux of SRP by 51.0%, on day 30 (Fig. 3). However, the resupply ability (R) of P in the 1-cm biochar capping group was significantly lower compared to the control and 2-cm biochar capping group (p < 0.001; Fig. 5 and Fig. S3). The rising changes in SRP and labile P concentrations caused by biochar capping were disproportionate over the short-term period of biochar capping, resulting in differences in P resupply ability. After aging of biochar (80 days), the resupply capacity of sediment P improved, resulting in increases of more than 121.5% and 33.7% in the release fluxes of SRP in 1-cm and 2-cm capping groups compared to the control group, respectively (Fig. 5 and Fig. S3). The P fraction analysis also further confirmed a rise in the sediment P resupply. The solid labile P and Ca-P contents were lower than those for the control group (Table 1). Labile P and Ca-P are recognized as one of the main P fractions adsorbed by biochar (Jung et al. 2015; Zhang et al. 2020b). In summary, in the short term, a large amount of biochar capping can reduce SRP concentrations and release flux by chemisorption of P, whereas with biochar aging, the remediation effects of biochar capping decrease, leading to a rise in the resupply capacity and release flux of sediment P.

Fig. 5
figure 5

Changes in net diffusion fluxes of SRP and labile P across the sediment–water interface after biochar capping

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Table 1 The P fractions in control and capping sediment within 40 mm on day 80
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3.3 Response of the microbial community composition to biochar capping

To further investigate the impact of biochar capping on P mobilization processes, the response of microbial community composition to biochar capping was analyzed. After DNA metabarcoding and bioinformatic analysis, a total of 21,224 bacterial amplicon sequence variants (ASVs) were identified in sediments. Proteobacteria accounted for 39.7–43.7% of the total bacterial sequences in sediments, which was the dominant phylum (Fig. 6a), followed by Acidobacteria, Chloroflexi, and Nitrospirae. The relative abundance of Proteobacteria and Bacteroidetes in sediments was higher after biochar capping, whereas some microorganisms, such as Acidobacteria and Chloroflexi, showed opposite changes (Fig. 6a). In addition, PCoA showed that microbial communities were significantly separated in sediments after biochar capping (Fig. 6b). Changes in bacterial community composition after the addition of biochar may be attributed primarily to the alkalinity of the biochar (Fig. S4). Acidobacteria is thought to be the phylum most sensitive to biochar and is dominant in low pH soils (Xu et al. 2016). In this study, it was found that Acidobacteria were significantly reduced in biochar-capped sediments, which is consistent with previous studies (Sheng and Zhu 2018). Benthic microorganisms are mostly responsible for the material cycling process in sediments (Ma et al. 2023). Biochar capping may alter the specific microbial activity and thus affect the process of material transport and transformation (Li et al. 2022c). On one hand, the nutrients adsorbed by biochar could provide some nutrients to benthic microorganisms and increase microbial richness (Fig. 6c) (Xu et al. 2019). On the other hand, biochar can provide a better environment for microorganisms to survive and reproduce due to its porosity and high specific surface area. These pore structures can allow microorganisms to hide to avoid predation, thus enhancing the abundance of microorganisms in the soil (Ameloot et al. 2013; Liang et al. 2010).

Fig. 6
figure 6

a The dominant bacterial communities at the phylum level in the control, 1-cm and 2-cm biochar capping sediments. b The composition and structure differences of microbial communities at the ASV level. c The richness, relative abundance of Pseudomonadaceae and Geobacteraceae and the ALP activity for the control, 1-cm and 2-cm biochar capping sediments

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Functional P-cycling genes were analyzed to understand how biochar affects microbial-controlled P-cycling processes in sediments. There was no significant effect on the overall abundance of P-cycling genes associated with inorganic P solubilization, organic P mineralization, P transportation, and P regulation. However, the relative abundance of individual genes, such as ppa and phnO was significantly affected by biochar (Fig. S5). The biochar capping increased the relative abundance of ppa-involved inorganic P solubilization and decreased the relative abundance of phnO-involved organic P mineralization. In addition, the abundance of phosphate-solubilizing bacteria, particularly Pseudomonadaceae (the dominant genus), was influenced by biochar (Fig. 6c). The relative abundance of Pseudomonadaceae in the treatment group was lower compared to the control group (Fig. 6c). ALP activity also showed a decreasing trend with the increase in biochar, decreasing from 1714.3 U g−1 in the control group to 1048.8 U g−1 and 526.7 U g−1 in the 1-cm and 2-cm capping groups, respectively (Fig. 6c). Previous studies found increases in phosphate-solubilizing bacteria and ALP activity in flooded rice rhizospheres following biochar addition (Chen et al. 2022). However, our findings suggest that the presence of excess labile P in sediments may contribute to the decrease in ALP activity. Long-term capping of biochar promoted P desorption. High concentrations of dissolved inorganic P can have a negative effect on organic P mineralization processes in sediments (Zhai et al. 2014).

Phosphorus cycling in sediments has also been reported to be strongly linked to Fe redox conditions (Liang et al. 2022). The biochar capping group had a higher abundance of Geobacteraceae, which involved in the reduction of Fe(III), compared to the control group (Fig. 6c). Fe(III) reduction bacteria-mediated reductive dissolution of Fe oxides enhances P release from sediments (Li et al. 2023a). The capping of biochar provided suitable conditions for the growth of Fe(III) reduction bacteria, which increased the biotic reduction of Fe oxides. Also, the capping of biochar also increased BDOM concentration and promoted the desorption of P (Fig. 3). In this study, the Fe(II) reduced by microorganisms was bound by biochar and S(−II), and the adsorption rate of Fe(II) was higher than the reduction rate of Fe oxides, resulting in lower concentrations of soluble and labile Fe(II) in the capping group compared to the control group (Fig. 3). The decrease in the amount of Fe(II) that can be oxidized on the sediment surface caused by biochar capping reduces the sediment's capacity to retain P and increases P release from the sediment. Overall, biochar capping can alter the sediment microbial community structure and function, and, in turn, lead to the P release from sediments.

3.4 Implications of biochar capping on sediment restoration

Biochar has been extensively used for the remediation of contaminated sediments; however, the accompanying adverse effects of biochar capping on the aquatic environment, especially on sediment P release, have not been fully understood. In this study, we examined the mobility, bioavailability and release risk of P under different biochar additions and treatment times using microcosm experiments (Fig. 7). The results showed that both the capping amount and the treatment time affected the mobility and P release in sediments. In the short term, biochar capping can effectively control SRP concentrations in porewater in surface sediments by rapid chemisorption. After a longer incubation period (80 days), the sediment P resupply capacity increased, resulting in higher P release from sediments in the biochar capping group compared to the control group (Fig. 5). The changes in biochar composition and surface chemistry, as well as sediment microbial community structure after biochar capping over time, reduce the effectiveness of biochar-based sediment remediation (Yang et al. 2021b). Furthermore, the sharply increased process of SRP concentration in the overlying water due to biochar capping during the initial phase should be of wide concern. Thus, modifying biochar in a way that sequesters phosphorus release and slows aging is critical for biochar remediation of water bodies and sediments.

Fig. 7
figure 7

Conceptual diagram of the effects of biochar remediation on phosphorus bioavailability and mobility at the SWI

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In summary, biochar capping has adverse effects on sediment P release, although it is effective for controlling P release in short-term applications by using a large amount of biochar. Therefore, the effectiveness of biochar in the long term is an essential foundation for exploring the potential environmental risks of biochar. Additionally, a systematic assessment of the effects of biochar capping on aquatic ecosystems by combining aquatic organisms, water environment and other elements needs to be further explored.

4 Conclusions

Using advanced fine-scale measurement methods, this study demonstrated the effects of biochar capping on the P mobilization processes in sediments. The application dosage and capping time of the biochar were two important factors that influenced sediment P release. Adequate biochar capping can act as a barrier to sediment P release in the short term, while the sharp increase in SRP and organic matter in the overlying water should be noted. After the aging of biochar (80 days), the bioavailability and P release flux were increased, leading to the P desorption from the aging biochar. Furthermore, biochar capping intensified the microbial-mediated Fe reduction, which was subsequently sequestrated by biochar and S(−II), reducing the sediments retention capacity for P and thereby enhancing P mobility. The novel P mobilization processes in sediment from biochar capping determined in this study will contribute to the assessment of the ecological risks of biochar remediation for ecosystem function and service.