1 Introduction

Water pollution by antibiotics has become a global environmental problem due to incomplete metabolism in humans or animals, unsatisfactory removal by conventional processes, and unintentional wastewater discharges from pharmaceutical plants and hospitals (Li et al. 2018, 2024; Liu et al. 2023a). Traditional wastewater treatment technologies, such as adsorption, flocculation, biotechnology, and membrane technology, fail to completely remove antibiotics from the environment (Rodriguez-Narvaez et al. 2017). Therefore, develo** cost-effective approaches to remediate antibiotic-contaminated water is an important goal. Advanced oxidation processes (AOPs) have received increasing attention in recent studies (Zhang et al. 2020; Sun et al. 2010). Synergies between biological invasions and other drivers of ecosystem degradation exacerbate current invasions and increase their scope and impacts (Pyšek et al. 2020). Once invasive plants become dominant, they can displace native vegetation and disrupt the original ecosystem (Zhou et al. 1c) further confirmed the homogeneous distribution of C, N, O, Co, and S species within the catalyst. Additionally, Additional file 1: Fig. S2 presents the N2 adsorption–desorption isotherms data, elucidating the pore size distribution. In accordance with the IUPAC classification, the CoS@MLBC isotherm conformed to a type IV isotherm exhibiting an H3 hysteresis loop, indicative of a mesoporous structure.

Fig. 1
figure 1

SEM images of CoS@MLBC (a), (b), EDX elemental map** of CoS@MLBC (c), TEM image of CoS@MLBC (d), and HRTEM imagines of CoS@MLBC (e), (f)

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The XRD pattern of CoS@MLBC is depicted in Fig. 2a, revealing distinct diffraction peaks at 2θ = 30.56°, 35.26°, 46.98°, and 54.28° match the (100), (101), (102), and (110) crystal planes, respectively, of CoS (PDF#75–0605). This analysis result was consistent with that of HRTEM. Furthermore, the diffraction peaks at 15.1° and 22.5° match the (001) and (100) crystal planes of g-C3N4(PDF#87-1526), respectively, indicating the synthesis of carbon nitride structure. The XRD pattern of BC indicated that the catalyst also formed the structure of CN (Fig. 2a). Compared with BC, the (001) and (100) peaks of CoS@MLBC were further weakened, indicating that the metal successfully combined with BC under the condition that did not change the structure of CN. As shown in Additional file 1: Fig. S3a, the distinct diffraction peaks of pure Co catalyst evidenced the synthesis of CoS crystal. In summary, the XRD patterns illustrated that the carbon nitride structure was successfully synthesized using a one-step HTC method based on ASM as biomass.

Fig. 2
figure 2

The XRD pattern of the prepared catalysts (a), FTIR image (b), and XPS spectra of CoS@MLBC: C 1s (c), N 1 s (d), Co 2p (e), S 2p (f)

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The FTIR spectra also revealed similar observations (Fig. 2b). The peaks observed at the range of 1000 ~ 1680 cm−1 in the two curves can be attributed to the stretching vibration of the C–N and C = N–C heterocyclic skeletons (Bharathi et al. 2023; Shaheen et al. 2022; Yu et al. 2023). And the peak at 890 cm−1 corresponds to the bending vibration of tri-s-triazine rings (Wang et al. 2019). Moreover, the peak observed at 600 cm−1 and 1100 cm−1 can be ascribed to the surface oxidation of sulfur species on the catalyst and the presence of Co-S bonds, respectively (Mahmoud et al. 2021; Yuan et al. 2020). The stretching vibrations of –CH3, –CH2– appeared at 2900 cm−1. A broad peak between 3100 and 3600 cm−1 is due to the N–H and O–H stretching modes, which can be attributed to the N–H groups on the uncondensed surfaces and the absorbed H2O (Qiu et al. 2024; Wen et al. 2023).

XPS was used to analyze the change of elemental composition and valence states on the catalyst surface. The full XPS survey scan spectrum indicated that CoS@MLBC was comprised of C, O, N, S, and Co elements (Additional file 1: Fig. S4a). Figure 2c–f shows the high-resolution spectra of C 1 s, N 1 s, Co 2p, and S 2p peaks in CoS@MLBC. The C 1 s spectra of CoS@MLBC (Fig. 2c) could be deconvoluted into three distinct peaks at 284.8 eV, 286.5 eV, and 288.5 eV, which correspond to adventitious carbon (C–C bond), C-NHx and sp2-bonded carbon (N–C = N), respectively (Qiu et al. 2024; Balakrishnan et al. 2023; Liu et al. 2021). Figure 2d displays the three peaks at 398.62 eV, 399.68 eV and 400.73 eV were attributed to C = N–C, C–N, and C–N–H, respectively (Peng et al. 2021). The four peaks at 782.00 eV, 784.47 eV, 797.77 eV, and 800.26 eV can be ascribed to Co3+2p3/2, Co2+2p3/2, Co3+2p1/2, Co2+2p1/2, respectively (Fig. 2e). The percentage of Co2+ and Co3+ were about 38% and 62%, respectively, and Co3+ was considered to be the surface oxidation of CoS (Fang et al. 2022). As shown in Fig. 2f, the two distinct peaks at 162.18 eV and 164.31 eV can be attributed to the S 2p3/2 and S 2p1/2, respectively.

3.2 Performance analysis

3.2.1 Catalyst comparison

The CIP was selected as the model antibiotic and the removal efficiencies of CIP by CoS@MLBC/PMS, Co/PMS, BC/PMS, and PMS systems were 100%, 98.13%, 17.01%, and 7.68%, respectively, within 10 min (Fig. 3a). The reaction rate constants (kobs) of CoS@MLBC/PMS, Co/PMS, BC/PMS and PMS systems were 0.6865, 0.3570, 0.0062, and 0.0053 min−1, respectively (Additional file 1: Fig. S5a). The control tests showed that CIP was relatively stable in the catalyst-free PMS system. The results indicate that the oxidant cannot decompose CIP alone. Compared with that of BC/PMS, the CoS@MLBC/PMS system showed an obvious superiority in CIP removal; the kobs of CoS@MLBC/PMS showed an almost 110-fold increase (kobs increased from 0.0062 min−1 to 0.6865 min−1). And according to the contact angle analysis (Additional file 1: Fig. S6), the Co-doped enhanced the hydrophilicity of catalysts, which indicated that metal ions played a critical role in PMS activation. Apparently, CoS@MLBC had better performance than CoS, indicating that the hydrophilic structure of CN loaded in biochar enhanced the adsorption of the catalyst and improved catalytic performance.

Fig. 3
figure 3

Removal efficiency of CIP by different systems (a), different antibiotics (b), and effect of catalyst dosage (c), PMS dosage (d). Conditions: [CIP]0 = 10 mg L−1, [Cata.] = 0.1 g L−1, [PMS]0 = 0.2 g L−1, pHini = 7.0, T = 25 ℃

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According to the XRD patterns analysis of Co-Corn and Co-WS (Additional file 1: Fig. S3b and c), it was observed that there were no characteristic peaks of g-C3N4. The results showed that the invasive plants ASM as biomass could be used to successfully synthesize the catalyst with CN structure using the one-step HTC method. Besides, the CIP adsorption efficiencies of Co-Corn and Co-WS were 21% and 22% in 30 min, respectively (Additional file 1: Fig. S7a) and the CIP degradation efficiencies were only 55% and 48% in 20 min, respectively (Additional file 1: Fig. S7b). It was further illustrated that the CN structure could effectively improve the catalytic performance of catalysts.

To test the generalizability of the CoS@MLBC/PMS system in removing antibiotics, NOR, CIP, Levo, and TC were selected as target pollutants. The removal efficiencies of NOR, CIP, Levo, and TC were 99%, 100%, 97%, and 100%, respectively (Fig. 3b), indicating that the CoS@MLBC/PMS system exhibited good degradation performance for the abovementioned antibiotics. TC was more susceptible to degradation probably because its molecular structure was susceptible to attack by radicals (Fan et al. 2023a). The molecular masses of NOR, CIP, and Levo were similar, resulting in comparable degradation outcomes. The CIP was chosen as the paradigmatic antibiotic for forthcoming experiments.

In this work, we carried out an application evaluation of the CoS@MLBC/PMS system for the removal of CIP in actual water. As shown in Additional file 1: Fig. S5c and d, the degradation efficiency of CIP was 92% within 10 min and it was completely degraded within 20 min. Compared to the control test, the kobs of CIP in real water (0.2720 min−1) was lower than that of the control test (0.6863 min−1). The possible reason for this is that the organic matter and inorganic ions in the water column affect the degradation efficiency of CIP, which agrees with the study of Fan et al. (Fan et al. 2023b). In addition, the comparison of various monometallic catalysts for the activation of PMS to pollutant degradation is presented in Additional file 1: Table S2. The proposed CoS@MLBC catalysts had the simplest preparation method while guaranteeing a high degradation efficiency. The comparative results showed that the CoS@MLBC catalyst efficiently activated the PMS degradation of CIP.

3.2.2 Effect of catalyst and PMS dosage

The impact of catalyst concentration on CIP degradation is illustrated in Fig. 3c. The results demonstrated a positive correlation between catalyst concentration and the removal efficiency of CIP, with an increasing trend observed within the range of 0.05 g L−1 to 0.1 g L−1 and kobs increased from 0.1316 min−1 to 0.6863 min−1. The degradation efficiency of CIP exhibited a slight decrease as the catalyst dosage was increased from 0.1 g L−1 to 0.4 g L−1. The enhancement of degradation efficiency could be attributed to increased active sites as the catalyst concentration increases. However, the excess catalyst will activate PMS to produce many radicals quickly, which will cause radical quenching and reduce the degradation efficiency (Hong et al. 2019). Therefore, the concentration of 0.1 g L−1 of the catalyst was chosen for forthcoming experiments.

The concentration of PMS also affected the CIP degradation efficiency (Fig. 3d). The increase in PMS dosage from 0.1 g L−1 to 0.2 g L−1 led to a significant enhancement of CIP degradation efficiency, as evidenced by the observed kobs increasing from 0.1664 min−1 to 0.6865 min−1. The possible reason is that an elevation in PMS concentration leads to increased production of radicals. The degradation efficiency of CIP remained relatively stable within the PMS concentration range of 0.2 g L−1 to 0.4 g L−1. On one hand, the ability of the reaction site to adsorb PMS was limited. On the other hand, the excess radicals would react with one another, which forms other active species with lower oxidation potential, thus reducing the degradation efficiency (Eqs. 1–2) (Fan et al. 2023a). Therefore, the concentration of 0.2 g L−1 of PMS was chosen for forthcoming experiments.

$$SO_{4}^{{\cdot - }{}} + SO_{4}^{\cdot - } \to S_{2} O_{8}^{{{\mathbf{2}} - }}$$
(1)
$$SO_{4}^{{\cdot - }{}} + \, \cdot OH \to \,HSO_{5}^{ - }$$
(2)

3.2.3 Effect of pH and anions

The impact of initial pH on the removal efficiency of CIP was detailed. The removal efficiency of CIP decreased from 100 to 94% (kobs decreased from 0.6865 min–1 to 0.1992 min–1) within 10 min when the initial pH was reduced from 7 to 3 (Fig. 4a). This might be due to the high H+ concentration preventing the adsorption of PMS on the active sites, limiting its activation process (Long et al. 2021) (Eqs. 3–4). In addition, excessive H+ inhibits the production of CoOH+, which will scavenge hydroxyl radicals and sulfate radicals; thus, the degradation efficiency decreases under strong acid conditions (Eqs. 5–6). Furthermore, HSO5− and SO52− are present at pH < 7.6, with HSO5− the dominant species and HSO5− more oxidizing and generating SO4− than SO52− capacity (Cao et al. 2020b). Thus, the kobs increased from 0.6865 min–1 to 0.8046 min–1 at pH = 5. At pH > 7.6, SO52− was the dominant species, which reacts more readily with ·OH to generate S2O8·2– and further quench SO4·− (Eqs. 7–8). Therefore, the degradation efficiency of CIP was slightly inhibited under alkaline conditions. In conclusion, it is obvious that the removal efficiencies of CIP at pH = 3, 5, 7, 9, 11 were 96%, 100%, 100%, 98%, and 90%, respectively, indicating that CoS@MLBC/PMS maintained high efficiency in degrading CIP over a wide pH range.

$$H^{ + } + SO_{4}^{\cdot - } + \, e^{ - } \to HSO_{4}^{ - }$$
(3)
$$H^{ + } + \cdot OH + \, e^{ - } \to H_{2} O$$
(4)
$$Co^{{\mathbf{2} {+} }} + H_{2} {O} \to CoOH^{ + } + H^{ + }$$
(5)
$$CoOH^{ + } + HSO_{5}^{ - } \to CoO^{ + } + H_{2} O + SO_{4}^{{ - }{}}$$
(6)
$$\cdot OH + {\text{SO}}_{{5}}^{{\mathbf{2}} - } \to S_{2} O_{8}^{{ \cdot {\mathbf{2}} - }}$$
(7)
$$SO_{4}^{\cdot - } + \, S_{2} O_{8}^{{\cdot{\mathbf{2}} - }} \to S_{2} O_{8}^{\cdot - } + \, SO_{4}^{{{\mathbf{2}} - }}$$
(8)
Fig. 4
figure 4

Effect of pH (a), Cl− (b), NO3− (c) and the kobs of different concentrations of Cl− and NO3− (d). Conditions: [CIP]0 = 10 mg L−1, [Cata.] = 0.1 g L−1, [PMS]0 = 0.2 g L−1, pHini = 7.0, T = 25 ℃

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The effect of common ions (Cl− and NO3−) in water on the CoS@MLBC/PMS system was investigated as shown in Fig. 4b, c. The inhibitory effect was significant when the Cl− concentrations ranged from 10 to 50 mM, as depicted in Fig. 4b. At lower Cl− concentrations, oxidation of Cl− by SO4·− and ·OH leads to the formation of chlorides or hypochlorite ions (Wang et al. 2021), expressing the reaction (Eqs. 9–11). As shown in Fig. 4c, it could be observed that removal efficiency of 98%, 96%, and 93% corresponded to NO3− concentrations of 10 mM, 50 mM, and 100 mM, respectively. The results indicated that the presence of NO3− did not significantly impact the removal efficiency, but the reaction rate decreased with the increase in ion concentration. The decrease in CIP removal efficiency may be attributed to the reaction between external anions and ROSs, which form other active species, as demonstrated by the reaction in Eqs. 12–13.

$$Cl^{ - } + \cdot OH \to HOCl^{ \cdot - }$$
(9)
$$Cl^{ - } + SO_{4}^{\cdot - } \to SO_{4}^{{{\mathbf{2}} - }} +Cl \cdot$$
(10)
$$HOCl^{ \cdot - } + H^{ + } \to Cl \cdot + H_{2} O$$
(11)
$$NO_{3}^{ - } + \, SO_{4}^{ \cdot - } \to SO_{4}^{{{\mathbf{2}} - }} + \, NO_{3}\cdot$$
(12)
$$NO_{3}^{ - } + \cdot OH \to OH^{ - } + NO_{3}\cdot$$
(13)

3.2.4 Reusability and stability of CoS@MLBC/PMS system

The reusability of the CoS@MLBC/PMS system is a critical factor for practical applications. As presented in Additional file 1: Fig. S8a, the CoS@MLBC/PMS system showed no obvious loss of degradation efficiency over the 4 cycles and maintained its performance. After 20 min reaction, the removal efficiencies of CIP in 4 cycles were 100%, 100%, 99%, and 98%, respectively, and the kobs is shown in Additional file 1: Fig. S8b. Furthermore, based on the XPS analysis of the catalyst before and after the reaction (Additional file 1: Fig. S4), no significant change was found in the Survey scan before and after use, and it is speculated that the decrease in CIP removal efficiency may be due to the change in the valence state of Co (Long et al. 2021b). In addition, carbonaceous catalysts are easily deactivated during PMS activation, which is a possible reason for the reduced CIP degradation efficiency (Yang et al. 2021). However, as shown in Additional file 1: Fig. S8c, the XRD patterns of the used catalysts did not exhibit any specific peaks, indicating no change in catalyst surface structure and composition. The findings suggest that the CoS@MLBC/PMS system exhibited a remarkable capability for efficient degradation of CIP and demonstrates excellent reusability. The FTIR spectra of fresh and used catalysts (Additional file 1: Fig. S8d) showed no significant change, further confirming the stability of CoS@MLBC. In addition, the CoS@MLBC/PMS system exhibited a lower leaching rate of Co ions than the Co/PMS system (Additional file 1: Fig. S5b). The results indicate that the biochar structure effectively immobilized Co ions and reduced ion leaching. The relevant calculations are given in Additional file 1: Text S5.

3.2.5 Toxicity test

In this work, we designed a Vigna radiata cultivation experiment to test the potential toxicity and ecological effects of CoS@MLBC. The control group (CG) was Vigna radiata cultivated with ultrapure water, and the treatment group (TG) was Vigna radiata cultivated with ultrapure water containing CoS@MLBC (the catalyst concentration of 0.1 g L−1). According to Additional file 1: Fig. S10, the average stem length of CG was 20.52 cm, and the average root length was 10.86 cm. The mean stem length of TG was 20.97 cm and root length was 9.24 cm. The results for CG and TG were not significantly different, indicating that CoS@MLBC had no impact on the growth of Vigna radiata. In summary, CoS@MLBC is a simple and environmentally friendly catalyst.

3.3 Degradation mechanism

3.3.1 Identification of ROSs

The radical scavengers and EPR analysis were employed to identify the active species produced in the CoS@MLBC/PMS system. Tert-Butanol (TBA), methanol (MeOH), 1,4-Benzoquinone (pBQ) and NaN3 were added in the CoS@MLBC/PMS system to quench ·OH, SO4·−, O2·− and 1O2, respectively. As shown in Fig. 5a, the addition of excess TBA significantly inhibited the degradation efficiency of CIP by 59%, providing evidence for the generation of ·OH in the CoS@MLBC/PMS system. Furthermore, adding MeOH inhibited 63% of CIP degradation, indicating that the contribution of SO4·− was also significant. Furthermore, NaN3 inhibited CIP degradation by 55%, indicating that 1O2 played a key role. In contrast, excess pBQ inhibited CIP degradation by 12%, indicating that O2·− may not be the primary active species. Overall, O2·− had little effect on the degradation of CIP, and ·OH, SO4·− and 1O2 promoted the degradation of CIP. Based on the above analysis, the pathway of O2·−, ·OH, SO4·− and 1O2 production was explored. Firstly, HSO5− could self-decomposition to generate SO52−, which react with ·OH to produce HO2·− (Additional file 1: Eqs. S3–5). And the O2·− was generated by the self-decomposition of HO2·−, SO4·− and ·OH could be generated by the oxidation of Co2+ (Additional file 1: Eqs. S6-7), and HSO5− could gain electrons to produce SO4·− (Additional file 1: Eqs. S8). Besides, SO4·− could react with H2O or OH− to produce ·OH (Additional file 1: Eqs. S9-10). In addition, the CIP degradation under N2 conditions was carried out, and the results showed that the effect of N2 on CIP degradation was negligible (Additional file 1: Fig. S11), suggesting that 1O2 was not formed by dissolved oxygen but by PMS activation. Furthermore, ·OH had a shorter lifespan than 1O2, which might lead to its conversion to 1O2 (Additional file 1: Eqs. S11). Therefore, 1O2 might be generated by the oxidation of PMS. In addition, Fig. 5b–d illustrates EPR analysis in the CoS@MLBC/PMS system, and the presence of ·OH, SO4·−, 1O2 and O2·− was confirmed. The results were consistent with the quenching experiments.

Fig. 5
figure 5

Effect of radical scavengers on CIP removal in CoS@MLBC/PMS system (a), EPR spectra of ·OH and SO4‧− (b), 1O2 (c), O2− (d)

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3.3.2 The role of CoS and BC

According to Additional file 1: Fig. S4b, it is observed that the proportions of Co2+ decreased from 38 to 29%, and Co3+ increased from 62 to 71%. A part of Co2+ on the catalyst surface was oxidized to Co3+ during the activation of PMS. Fig. S4c illustrates that the S2− proportion decreased from 65 to 58%, and the Sn2− increased from 35 to 42%. Due to the low redox potential, S2− was oxidized by PMS to expose Co2+, which participates in the activation of PMS (Additional file 1: Eqs. S6-7). Previous studies have shown that the reaction of S2− with PMS does not produce ROSs, but the low electronegativity of S2− can promote the reduction of Co3+ to Co2+ (Additional file 1: Eqs. S12-13). The CoS/PMS could remove CIP to 100% in 20 min, while in the Co2+/PMS system, the CIP removal efficiency was only 32% (Additional file 1: Fig. S12). This indicated that S2− promotes the generation of Co2+ to activate PMS more effectively. For the mechanically mixed CoS and BC (1:1 mass ratio), the CIP removal was only 59%, which evidenced that chemical synthesis (CoS@MLBC) enhanced the binding of CoS and BC to improve the activation performance of PMS. The result indicates that BC could effectively enhance the electron transfer capacity of CoS to promote PMS decomposition in the CoS@MLBC/PMS system. The electrochemical analysis was performed to verify this conclusion. The cyclic voltammetry curve illustrated that CoS@MLBC had the highest current density compared to CoS and BC, indicating that CoS@MLBC had the highest reduction capacity (Fang et al. 2022) (Fig. 6a). The electron transfer potential of different catalyst surfaces was measured via electrochemical impedance spectroscopy (EIS). As shown in Fig. 6b, it could be observed that the arc radius of CoS@MLBC was smaller than that of CoS and BC, indicating that CoS@MLBC had lower impedance and higher electron transfer efficiency (Ren et al. 2019). Therefore, the combination of CoS and BC accelerated the electron transfer rate, and the carbon nitride structure improved the electron transfer performance of the catalysts in PMS activation, which was consistent with the results of previous studies (Shang et al. 2019).

Fig. 6
figure 6

Cyclic voltammetry curves (a) and EIS in 0.1 M Na2SO4 (b) of CoS, BC and CoS@MLBC

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Based on the above-mentioned analysis, the possible mechanism of CIP degradation in the CoS@MLBC/PMS system was as follows (Scheme 2): The O–O bond in PMS was broken by the absorption electrons from CoS@MLBC to generate ·OH and SO4·−, and Co2+ could combine with HSO5− to produce ·OH, SO4·− and Co3+ (Additional file 1: Eqs. S6-7). The generated Co3+ could further react with PMS to produce SO5−· (Additional file 1: Eqs. S14), indicating that the Co2+/Co3+ cycle significantly contributed to the activation of PMS (Hu and Long 2016). Additionally, SO5−· could self-react or react with H2O to further generate 1O2 (Additional file 1: Eqs. S15-16). Overall, the CoS@MLBC could efficiently activate PMS to generate ROSs for degradation of CIP.

Scheme 2.
scheme 2

Possible mechanism of CoS@MLBC/PMS system on CIP degradation

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3.4 Intermediate identification and toxicity prediction

3.4.1 Degradation pathways and intermediate analysis

In the CoS@MLBC/PMS system, twelve significant intermediates generated from the degradation process of CIP were successfully identified using HPLC/MS analysis. Three possible degradation pathways of CIP were initially proposed, and the molecular weights and structures of the intermediates were analyzed in detail (Fig. 7). For Pathway I, the formation of P1 (m/z 362) could be attributed to the cleavage of the piperazine ring, resulting in a dialdehyde derivative structure. Then, P1 lost two formaldehyde groups one after another to form P2 (m/z 334) and P3 (m/z 306). The secondary amine nitrogen lost from P3 was rapidly oxidized, resulting in the formation of P4 (m/z 291), and subsequent loss of formaldehyde from P4 led to the generation of P5 (m/z 263), ultimately leading to complete cleavage of the piperazine ring. Pathway II began with the oxidation of the cyclopropyl group and cleavage to create P6 (m/z 292). Subsequently, the piperazine ring underwent fragmentation, resulting in the generation of P7 (m/z 223). Finally, the elimination of the carboxyl group from the quinolone ring yielded P8 (m/z 179). Pathway III was designated as the defluorination (−OH substitution -F) to form P9 (m/z 330). On this basis, the cleavage and elimination of both the cyclopropyl group and carboxyl group from the quinolone ring by the Kolbe decarboxylation reaction generated P10 (m/z 304), P11 (m/z 177). According to Additional file 1: Fig. S13, the intermediates of Pathway I and Pathway III had higher intensity in the ion chromatogram. They were the primary degradation pathways of the degradation process, while Pathway II was the minor pathway.

Fig. 7
figure 7

Possible degradation pathways of CIP in the CoS@MLBC/PMS system

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3.4.2 Toxicity prediction of CIP and its intermediate

In order to predict and comprehend the potential ecotoxicity of CIP and its degradation products on aquatic organisms during catalyst degradation, we utilized the EPI Suite software to calculate the acute and chronic toxicity levels of CIP and its degradation products on marine organisms across three distinct trophic levels. The toxicity assessment was based on the Chinese hazard evaluation guidelines for new chemical substances (HJ/T 154-2004) and European Union criteria (Additional file 1: Table S3). The acute toxicity of the substance was evaluated by determining LC50 values for fish and Daphnia, as well as EC50 values for green algae, based on the data presented in Fig. 8 and Additional file 1: Table S4. The results revealed that CIP exhibited no detrimental effects on aquatic organisms. Intermediates with LC/EC50 and ChVs higher than 100 and 10 mg L−1, respectively, indicate that the negotiators were non-toxic and harmless to the organism. The mediators of the three pathways exhibited minimal toxicity towards fish, Daphnia, and green algae, except for P5, P8, and P12. It was observed that P5 displayed toxicity towards Daphnia and had detrimental effects on the health of fish and green algae. Similarly, P8 was found to be harmful to fish and green algae while exhibiting toxic effects on Daphnia. Furthermore, although P12 demonstrated adverse effects on Daphnia and green algae, it did not exhibit any detrimental impact on fish. In conclusion, the efficiency of CIP mineralization could be enhanced by extending the reaction duration, thereby mitigating system toxicity, indicating that the catalyst is eco-friendly in the CIP degradation process.

Fig. 8
figure 8

Evaluation of acute and chronic toxicity of CIP and its degradation intermediates towards three aquatic organisms using EPI Suite software

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4 Conclusion

In this work, a novel catalyst CoS@MLBC with carbon nitride structure was synthesized by a one-step HTC method using invasive plants as biomass feedstock, which can efficiently activate PMS to degrade CIP in water. Various characterizations and performance analyses revealed the successful synthesis of carbon nitride structures and the good efficiency of CoS@MLBC for the degradation of CIP. Meanwhile, the advantage of invasive plants for Co-modified biochar synthesis was illustrated through comparative analyses with that prepared using common biomass feedstocks (corn cop and wheat straw). The quenching experiments and EPR analysis were utilized to identify the main active species as ·OH, SO4·− and 1O2 towards CIP degradation. In addition, CoS@MLBC showed good stability and practicality through cycling experiments and actual water testing. Finally, the degradation pathways and intermediates of CIP were analyzed by HPLC–MS, and their ecotoxicity was predicted. In conclusion, this work provides a new perspective on the preparation of biochar from invasive plants for the efficient degradation of antibiotics in water.