Influence of Changes in OM Content and Form on the Adsorption of Two Typical Antibiotics to SPM

doi:10.21203/rs.3.rs-2830296/v1

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Research Article

Influence of Changes in OM Content and Form on the Adsorption of Two Typical Antibiotics to SPM

https://doi.org/10.21203/rs.3.rs-2830296/v1

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published 13 Oct, 2023

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Antibiotics are commonly detected in natural waters. Suspended particulate matter (SPM) in surface water has a good adsorption capacity for antibiotics, and the organic matter (OM) in SPM is the key factor for this adsorption. We investigated the contribution of OM content and form to the adsorption of tetracycline (TC) and norfloxacin (NOR) in the SPM of Taihu Lake. To change the content and form of OM in SPM, we burned it at 505 ℃ and oxidised it with H₂O_2, respectively. The OM in the sample that was combusted (expressed as SPM-505 ℃) was almost entirely eliminated, and the adsorption amount of TC decreased by 14.8% when compared to the original sample, while the adsorption amount of NOR decreased by 64.8%, indicating that OM content contributes to the adsorption of antibiotics. Although the OM content in the oxidised sample (expressed as SPM-H₂O₂) decreased, the adsorption amounts of TC and NOR increased by 31.7% and 2.8%, respectively, compared with the original sample. The infrared spectrum and elemental analysis results showed that the morphology of the OM changed, suggesting that the structure of the OM contributes substantially to the adsorption of antibiotics. The contribution of OM to NOR adsorption was greater than that of TC. In conclusion, the results verify the importance of OM in adsorbing antibiotics onto SPM, which may provide basic data for antibiotic migration in surface water.

suspended particulate matter

organic matter

adsorption of antibiotics

natural aquatic environment

The detection of antibiotic residues in the environment due to antibiotic abuse has raised widespread concern. Antibiotics can enter the environment through human and animal excretion, antibiotic production, discharge of production waste, and disposal of expired drugs. Waterbodies serve as pools from which antibiotics can be collected. Concentrations of antibiotic residues are in the ng/L-µg/L range in surface waters and the ng/g-µg/g range in sediments (Loos et al. 2013; Wei et al. 2011; Xu et al. 2014). Once antibiotics are introduced into aquatic environments, they may undergo one critical process: adsorption. Suspended particulate matter (SPM) in surface water, which is the first material to adsorb antibiotics and other pollutants, has been given attention. SPM is an important component dispersed in water bodies. SPM is widely distributed, although the individual particle size is small. It has a strong adsorption effect on pollutants in the aqueous phase. For example, Aminot (2018) suggested that SPM can moderate the sorption of wastewater-derived pharmaceuticals in estuarine waters. Wang et al. (2016) found that the distribution and interaction of parent and alkyl polycyclic aromatic hydrocarbons among multiple phases were mainly determined by SPM. Wang et al. (2022) suggested that partitioning processes into SPM might affect the distribution of detected antibiotics in rivers.

Furthermore, the organic matter (OM) of SPM is important for illustrating differences in the adsorption behaviour of antibiotics (Zhang et al. 2019). Lou et al. (2017) demonstrated that SPM with a high OM content in swine wastewater can adsorb large quantities of tetracycline (TC). Luo et al. (2019) discussed the influence of particle size on TC and found that norfloxacin (NOR) adsorption was mainly related to the physicochemical properties of SPM, such as OM. These studies indicated that OM is a key factor in SPM adsorption.

Therefore, further study of OM in SPM, such as the influence of OM content and form on contaminant adsorption behaviour, will be helpful. The effects of the OM content on adsorption were evaluated experimentally by OM removal from the samples. Chemical oxidation and combustion are commonly used to remove OM from soil (Zhao et al. 2004; Kumar et al. 2019). Additionally, the OM of SPM in water bodies comes mainly from dead algae or aquatic plants; therefore, SPM is likely to contain a large amount of refractory OM (such as lignin) and decomposed and mature OM (Maxson et al. 2013; Reimers et al. 2021). Chemical oxidation typically uses dichromate, H₂SO₄, or H₂O₂ to oxidize soil OM to CO₂ and can effectively remove decomposed OM while receiving modified OM (Eusterhues et al. 2005). Combustion at 500–600 ℃, which is a simple, fast, and effective OM removal method that does not introduce other impurities (Wang et al. 2003), was intended to measure the contribution of OM to contaminant adsorption. The SPM after chemical oxidation and combustion can be expressed as different factors of OM.

Therefore, the objective of this study was to develop a better understanding of the effects of OM on the antibiotic adsorption behaviour of SPM using different treatments, including chemical oxidation and combustion. The specific objectives were to (1) explore the effects of different types and contents of OM in SPM on antibiotic adsorption, (2) evaluate the contribution of OM in SPM to antibiotic adsorption, and (3) compare two methods of removing OM from antibiotic adsorption. We used two typical antibiotics (TC and NOR) based on their widespread presence in surface water and strong affinity for SPM. SPM, a carrier of antibiotics in surface water, was collected from a eutrophic lake. Scanning electron microscopy (SEM), Fourier-transform infrared (FTIR) spectroscopy, X-ray photoelectron spectroscopy (XPS), Brunauer–Emmett–Teller (BET) surface area analysis, and elemental (C, H, N, and O) analyses were conducted to explain the differences in the OM composition of the SPM samples. Experiments on the effects of OM are invaluable for supporting environmental risk assessments of residual antibiotics in surface waters.

2.1 Collection and treatment of samples

Samples were collected from Zhushan Bay (31.45001 °N, 120.04217 °E) of Taihu Lake, China, from 4 to 17 December 2017, captured by a series of 16 channel SPM traps (the inner diameter of each column was 84 mm and the length was 350 mm). The collection period was 15 days.

After freeze-drying, grinding, and passing through a 300-mesh sieve, debris and plant residue were removed from the sample. The residues remaining on the sieve are hereafter referred to as SPM. SPM samples were then treated using different methods. SPM-505°C samples were obtained by burning a portion of the SPM samples in a muffle furnace for 6 hours. In a 40°C constant temperature water bath, 20 mL of 30% H₂O₂ was added several times to a 20 g SPM sample, and the SPM-H₂O₂ sample was obtained after stirring, foaming, centrifuging, washing, and drying. Finally, part of the SPM-H₂O₂ sample was burned at 505°C in a muffle furnace for 6 hours to obtain the SPM-H₂O₂-505°C sample. Four samples of the adsorption material were investigated in this study: SPM, SPM-505°C, SPM-H₂O₂, and SPM-H₂O₂-505°C.

2.2 Kinetic study of adsorption

Four samples (0.1 g) were placed in 250 mL conical flasks, respectively and 150 mL of TC solution (20 mg/L) was added. The conical flasks were sealed with a plastic film and placed in a shaker at 120 rpm at 20 ℃. TC samples (1 mL) were withdrawn at 0, 15, 30, 67, 90, 120, 180, 252, 360, 510, 600, 720, 840, 1440, and 1560 min using a syringe, passed through a 0.22-µm filter, and then collected in a brown sampling bottle. All experiments were performed in triplicate, and a blank experiment was conducted simultaneously. The kinetics of the NOR adsorption were studied in a similar manner. The samples weighed 0.06 g, and the concentration of the NOR solution was 16 mg/L. Sampling times: 0, 5, 15, 30, 60, 90, 180, 240, 360, 960, and 1080 min.

2.3 Isothermal adsorption study

Four samples (0.017–0.025 g) were placed in 100 mL conical flasks respectively, and 50 mL of TC solution (5–25 mg/L) was added to the flasks, which were then sealed with plastic film. The conical flasks were placed in a shaker and oscillated at 120 rpm for 34 h at 20 ℃. After attaining adsorption equilibrium, a 1 mL aliquot of the mixture was withdrawn using a syringe, passed through a 0.22-µm filter, and injected into a brown sampling bottle. The experiments were conducted in triplicate, and blank experiments were carried out simultaneously. An isothermal adsorption experiment with a NOR solution was carried out in a similar manner with 0.04–0.06 g of sample and 50ml of a NOR solution of 2–20 mg/L.

2.4 Antibiotic detection and sample identification

A high-performance liquid chromatography (HPLC) system (Agilent 1260; Agilent Technologies, Inc., Santa Clara, CA, USA) equipped with a symmetry C18 column (150 mm × 2.1 mm, 3.5 µm) and a protection column (4.0 mm × 3.0 mm) was used to quantify antibiotic concentrations. The column temperature was 40°C and optimal separation was accomplished via isocratic elution using 87% formic acid (0.05%) and 13% acetonitrile. The pH of the mobile phase was adjusted to 3.2 with triethylamine. The injection volume was 15 µL, and the flow rate was set at 1 mL/min. TC and NOR were detected at 350 nm and 278 nm, respectively. The detection limits for both antibiotics were at the microgram level and the measurement accuracy satisfied the requirements of this study.

The surface characteristics of the four samples were observed using scanning electron microscopy (SEM; JSM-7800F; JEOL Ltd., Akishima, Tokyo, Japan). Furthermore, the C, H, O, and N levels in the samples were determined using an elemental analyser (Elementar Vario EL Cube; Elementar, Langenselbold, Germany). All measurements were performed in duplicate. The H/C, O/C, and (O + N)/C atomic ratios of each sample were calculated using the average values. Infrared spectra of the samples were obtained using a FTIR spectrometer (NICOLET IS5; Thermo Fisher Scientific; Waltham, MA, USA). The Brunauer-Emmett-Teller (BET) surface areas of the samples were measured using a BET-N₂ surface area analyser (NOVA; Quantachrome Instruments; Boynton Beach, FL, USA). The elemental composition of the sample surface and the valence states of Fe, Mn, and Al on the SPM surface were analyzed using an X-ray photoelectron spectrometer (EscaLab 250XI; Thermo Fisher Scientific; Waltham, MA, USA).

2.5 Calculations of the formulae of adsorption

Pseudo-first-order (PFO) model:

$$\text{ln}\left({q}_{e}-{q}_{t}\right)=\text{ln}{q}_{e}-{k}_{1}t$$

Pseudo-second-order (PSO) model:

$$\frac{t}{{q}_{t}}=\frac{1}{{k}_{2}{q}_{e}^{2}}+\frac{1}{{q}_{e}}t$$

where k₁ and k₂ are the rate constants of the pseudo-first- and pseudo-second-order kinetic equations, respectively. q_t and q_e are the amounts of antibiotics per unit mass of suspended particles when the adsorption process reaches t min and equilibrium, respectively. The units are mg/g.

The expressions for q_t and q_e are as follows:

$${q}_{t}=\frac{({C}_{0}-{C}_{t})V}{m}$$

$${q}_{e}=\frac{({C}_{0}-{C}_{e})V}{m}$$

In the above formulae, C₀ represents the initial concentration of antibiotics. C_t is the concentration of antibiotics in the aqueous solution at t min. C_e is the equilibrium concentration of antibiotics in aqueous solution. All units are expressed in mg/L. m is the mass of the particles (g). V represents the volume of added antibiotic solution (L).

Freundlich model:

$${q}_{e}={K}_{F}{C}_{e}^{\frac{1}{n}}$$

Langmuir model:

$${q}_{e}={q}_{m}\frac{{K}_{L}{C}_{e}}{1+{K}_{L}{C}_{e}}$$

In the Freundlich model, K_F is the Freundlich coefficient, and n is the Freundlich empirical constant. When n > 1, antibiotics can be easily adsorbed onto the suspended particles. In the Langmuir model, q_m is the adsorption capacity of monolayer adsorption in mg/g. K_L is the surface adsorption affinity constant.

3.1 Analysis of the physicochemical properties of the four samples

3.1.1 Scanning electron microscopic (SEM) and Brunauer–Emmett–Teller (BET) surface area analyses of samples

The SEM images of the four samples are presented in Fig. 1. The surface of the SPM was relatively smooth, but after combustion and treatment with H₂O₂, the surface morphology underwent considerable changes. The surface of SPM-505 ℃ sample is riddled with large holes. Furthermore, the surface of the SPM-H₂O₂ sample has several irregular holes, while both large irregular holes and small holes occur on the surface of the SPM-H₂O₂-505 ˚C sample.

Table 1 lists the BET surface areas of the four samples. The BET surface areas are in the following order: SPM < SPM-505 ˚C < SPM-H₂O₂ < SPM-H₂O₂-505 ˚C. The BET surface area of SPM-505 ˚C is larger than that of SPM because at high temperatures, the OM in SPM decomposes and generates gaseous volatiles. This led to the formation of several macro-hole structures on the surface of the SPM, which in turn increased the BET surface area. Moreover, the BET surface areas of SPM-H₂O₂ are much larger than those of SPM. This is primarily because numerous dense bubbles are generated during the H₂O₂ soaking process, forming many small holes in the SPM and freeing the adsorption sites at which inorganic minerals interact with OM. This also explains the presence of small holes on the surface of SPM-H₂O₂-505 ˚C.

Table 1

Brunauer–Emmett–Teller (BET) surface areas of the four samples obtained using different treatment methods
Sample	BET surface area (m²/g)
SPM	13.5799
SPM-505 ˚C	19.9517
SPM-H₂O₂	23.0513
SPM-H₂O₂-505 ˚C	25.2256

3.1.2 Elemental composition and concentration

The elemental compositions of the four samples are presented in Table 2. The H/C ratio can be used as an indicator of aromaticity. A smaller H/C ratio indicates that the sample is highly aromatic and that the SPM structure is highly stable. Furthermore, the hydrophilicity of the SPM was indicated by its O/C value, with a higher value indicating greater hydrophilicity. The greater the (O + N)/C value, the more polar the functional groups, such as -NH₂, -OH, and-COOH. As shown in Table 2, the N and C content of the SPM samples decreased after treatment, whereas their O content increased. While SPM-H₂O₂-505 ˚C has a slightly higher H concentration than SPM, H levels in the other two samples are lower than in SPM. Combustion increases the loss of N, C, and H from the SPM. This may be attributable to the breaking of weak chemical bonds in the SPM. Furthermore, the number of oxygen-containing functional groups increases during combustion, and inorganic minerals, such as Fe, Mn, Si, Ca, and Mg, are formed through sintering and fusion. The H/C value of SPM was lower than that of SPM-H₂O₂, indicating that the former has stronger aromaticity and better stability. The O/C and (O + N)/C ratios of SPM-H₂O₂ were higher than those of SPM. This implies that SPM-H₂O₂ has stronger hydrophilicity and polarity.

Table 2

Elemental composition (%) of the SPM samples obtained through different treatment methods
Element	Index	SPM	SPM-505˚C	SPM-H₂O₂	SPM-H₂O₂-505˚C
N	w (N, %)	0.60	< 0.3	< 0.3	< 0.3
C	w (C, %)	2.66	< 0.3	0.89	< 0.3
H	w (H, %)	0.82	0.34	0.74	0.39
O	w (O, %)	21.44	27.79	28.08	31.14
H/C	Atomic molar ratio	3.68	-	10.03	-
O/C	Atomic molar ratio	6.25	-	23.86	-
(O + N)/C	Atomic molar ratio	6.06	-	23.80	-

3.1.3 Infrared spectrum analysis and XPS analysis

The peak positions and intensities of the infrared absorption curves were used to determine the presence of certain functional groups. Figure 2 and Table 3 show the infrared spectra of the four samples with adsorbed TC, where the positions of the peaks correspond to specific functional groups. Although the four samples have similar compositions, the concentrations of their functional groups differ. The intensity of each peak corresponds to the level of the relevant functional group, illustrating that the surfaces of the samples obtained after treatment had relatively low concentrations of functional groups. At 3602–3604 cm^− 1, a vibration absorption peak of -OH on an aluminium octahedron in the inner cavity is observed. Furthermore, the infrared spectrum of SPM shows a strong and broad stretching vibration peak attributable to -OH at 3400–3410 cm^− 1, indicating that this substance is the primary structural feature of SPM. The absorption peak of the N-H stretching vibration of amide appears at 3410 cm^− 1 in the infrared spectrum recorded after adsorption, indicating that TC was adsorbed (Lou et al. 2017). The more intense absorption peak around 1654 cm^− 1 is due to the stretching vibration of C = C in olefins, which indicates that C = C is also a major structural feature of SPM. The absorption peak of the stretching vibration of C-H in aliphatic CH₃ and CH₂ or the asymmetric stretching vibration of -COOH was observed at 1450 cm^− 1. The intensity of the peak at 3400–3410 cm^− 1 increased after TC adsorption (Fig. 2), indicating an increase in -OH content on the surface and the introduction of the N-H group. The disappearance of the vibration peak near 1654 cm^− 1 after adsorption indicates C = C bond cleavage and the formation of C = O, -COOH, or C-H groups. It is reported that the ketene structure of TC can be used as a π-electron acceptor, while the C in the aromatic group can be used as a π-electron donor, which promotes the adsorption performance (Dai et al. 2020). In addition, NOR contains fluorine groups with strong electron absorption ability and can also be used as a strong π receptor, with π - π electron donor-receptor interactions for NOR adsorption of highly aromatic biochar (Wu et al. 2022).

Table 3

Attribution of characteristic absorption bands in the infrared spectra
Absorption band position (cm^− 1)	Attribution
3621 − 3410	-OH stretching vibrations
3410	N-H stretching vibration of amide
1650 − 1600	C = C stretching vibration in aromatic groups or C = O stretching vibration in aldehydes and ketones
1450 − 1400	C-H vibrations in aliphatic CH₃ and CH₂ or asymmetric stretching vibrations in -COOH

Figure 3 and Table 4 show the surface element composition and relative content of the four samples. The main metal elements present on the surface of the four samples are Fe, Al, and a small amount of Mn. The XPS spectrum of C 1s in the original sample can be roughly divided into three peaks, around 285 eV, around 286 eV, and between 288–289 eV. These three peaks correspond to sp³C, unsaturated carbon such as C = O, and π - π* satellite peaks, respectively (Reiche et al. 2014). After H₂O₂ treatment, the peak areas of these three peaks decreased significantly, while for the sample after combustion, the peak areas near 285 eV decreased significantly, even lower than that of SPM-H₂O₂, and the remaining two peaks have disappeared. This indicates that both H₂O₂ treatment and combustion treatment can significantly reduce the content of C element, and combustion treatment can destroy unsaturated carbon bonds to a greater extent. The above results are consistent with the elemental analysis results (Table 2).

Table 4

Surface chemical composition and relative contents of the four samples
Sample	$\text{F}\text{e}$					$\text{M}\text{n}$ Area	$\text{A}\text{l}$ Area
Sample	${Fe}^{2+}$ Area	Binding energy (eV)	${Fe}^{3+}$ Area	Binding energy (eV)	$\frac{{Fe}^{2+}}{{Fe}^{2+}+{Fe}^{3+}}$	$\text{M}\text{n}$ Area	$\text{A}\text{l}$ Area
SPM	25052.3	712.4	23750.1	714.2	0.5	849.0	22197.5
SPM-505˚C	28923.8	712.0	24714.8	714.0	0.5	500.10	22944.0
SPM-H₂O₂	25770.9	712.3	33236.2	714.1	0.4	503.30	21253.7
SPM-H₂O₂-505˚C	25064.2	712.2	18891.7	714.1	0.6	539.4	21164.5

The Fe XPS spectrum was divided into two peaks. Owing to the spin quantum in the 2p orbital, the Fe 2p spectrum has two main characteristic peaks, indicating that Fe exists as both Fe²⁺ (712 eV) and Fe³⁺ (714 eV). H₂O₂ is a strong oxidant that can oxidise Fe²⁺ to Fe³⁺, resulting in increased Fe³⁺ levels in the SPM after soaking in H₂O₂. The Mn was changed by combustion and H₂O₂ treatment. The Al content remained constant.

3.2 Kinetics studies

The adsorption kinetics of the fitting results are shown in Fig. 4. TC adsorption onto the samples was rapid from 0 to 300 min, after which the adsorption rate began to decrease. Reaction equilibrium was achieved within 1560 min. For NOR, adsorption began to slow after 150 min and reached equilibrium after 400 min. The initial fast adsorption may be attributed to the adsorption of hydrophobic organics on the hydrophobic surface of the samples, whereas the slow adsorption during the later stages may be caused by the diffusion of pollutants into the samples, where they were adsorbed onto sites within the lattice that would otherwise be difficult to reach (Weissenfels et al. 1992).

Table 5 lists the regression coefficients (R²), adsorption capacities (q_e), and rate constants (k₁ and k₂) obtained through kinetic studies of TC and NOR adsorption on the four SPM samples. The theoretical value of q_e calculated using the PSO model for the adsorption of TC and NOR is closer to the actual equilibrium value of q_e than the PFO model, indicating that the former can better explain the adsorption process than the latter. The TC and NOR adsorption capacities of SPM-H₂O₂ was the strongest, with equilibrium adsorption capacities of 27.9 and 34.8 mg/g, respectively. However, the adsorption capacities of SPM-505 ˚C and SPM-H₂O₂-505 ˚C for TC and NOR were poor. The equilibrium adsorption capacities of SPM-505 ˚C and SPM-H₂O₂-505 ˚C for TC were 18 and 12.9 mg/g, respectively, while those for NOR were 12 and 12.4 mg/g, respectively. The rate constant k₁ of the first-order kinetic equation for TC and NOR adsorption on SPM was always greater than k₂. Furthermore, the rate constant k₁ of SPM-H₂O₂ with adsorbed TC and NOR was the largest, while the rate constant k₁ of SPM-H₂O₂-505 ˚C with adsorbed TC and NOR was the smallest. The second-order kinetic constant k₂ for TC adsorption was the maximum for SPM-H₂O₂, and the minimum for SPM. For the adsorption of NOR, the k₂ value of SPM-505 ˚C was the largest, while that of SPM-H₂O₂-505 ˚C was the smallest.

Table 5

Adsorption kinetics of TC and NOR on four samples
	Sample			PFO			PSO
	Sample	q_e	q_e,cal	k₁	R²	q_e,cal	k₂	R²
TC	SPM	21.19	17.73	0.00185	0.9926	24.11	0.00014	0.9673
	SPM-505˚C	18.04	15.17	0.00195	0.9792	20.57	0.00017	0.9892
	SPM-H₂O₂-505˚C	12.87	10.79	0.00163	0.9979	14.81	0.00019	0.9514
	SPM-H₂O₂	27.92	19.73	0.00271	0.9716	29.28	0.00028	0.9890
NOR	SPM	33.85	22.65	0.00573	0.9140	33.01	0.00092	0.9190
	SPM-505˚C	11.96	9.18	0.00378	0.9654	11.83	0.00183	0.7146
	SPM-H₂O₂-505˚C	12.39	7.93	0.00373	0.7579	12.27	0.00091	0.8372
	SPM-H₂O₂	34.79	21.75	0.00675	0.9444	33.38	0.00138	0.8395

3.3 Isothermal adsorption studies

Figure 5 shows the adsorption isotherms of TC and NOR on the four samples. The isothermal adsorption results were consistent with the adsorption kinetics results. The TC adsorption capacities of the four samples are in the order SPM-H₂O₂ > SPM > SPM-505 ˚C > SPM-H₂O₂-505 ˚C, while the NOR adsorption capacities are in the order SPM-H₂O₂ > SPM > SPM-H₂O₂-505 ˚C > SPM-505 ˚C.

The fitting results for the Freundlich and Langmuir models are listed in Table 6. For TC adsorption, the R² values for the Langmuir model were between 0.5492 and 0.9751, while R² values for the Freundlich model were between 0.5185 and 0.9454. This indicates that the Langmuir model fits the TC adsorption isotherm better. For the adsorption of NOR, the fitting results for SPM-H₂O₂ and the original SPM samples were more consistent with the Freundlich model, with R² values ranging between 0.9747 and 0.9890, which were higher than the corresponding R² values for the Langmuir model (between 0.9244 and 0.9849). The strong TC adsorption capacity is related to the abundance of polar/ionic functional groups (such as -CONH₂, -OH, -C = O, and -NH(CH₃)₂) in the TC molecule, (Figueroa et al. 2004). In addition, TC contains many polar functional groups that may be adsorbed via hydrogen bonding with polar components on the surface of the particulates (Gu et al. 2007). Studies have shown that -COOH significantly contributes to the adsorption of quinolone antibiotics, such as NOR (Zhang et al. 2007). The adsorption capacities of the same soil for quinolone antibiotics before and after the removal of -COOH differ by two orders of magnitude (Nowara et al. 1997).

Table 6

Parameters of the adsorption isotherms of TC and NOR fitted to the Freundlich and Langmuir models
	Sample		Freundlich			Langmuir
		K_F	N	R²	q_m	K_L	R²
TC	SPM	1.64	0.84	0.9331	29617.65	0.00008	0.9132
	SPM-505˚C	4.65	1.91	0.9454	29.23	0.11822	0.9751
	SPM-H₂O₂-505˚C	1.46	1.10	0.9309	114.47	0.01157	0.9305
	SPM-H₂O₂	17.08	2.18	0.5185	61.57	0.28498	0.5492
NOR	SPM	12.45	2.48	0.9747	39.91	0.34485	0.9849
	SPM-505˚C	4.56	2.97	0.9529	11.52	0.55195	0.9978
	SPM-H₂O₂-505˚C	6.51	3.63	0.8894	11.96	1.43472	0.9607
	SPM-H₂O₂	26.32	9.22	0.9890	34.21	2.86151	0.9244

4.1 Effects of OM on the adsorption of TC and NOR by different SPM treatments

The ratio of antibiotic adsorption amount at equilibrium to SPM is obtained by comparing the q_e data of three treated SPM samples with the original SPM sample (Table 7). The adsorption capacity of SPMs treated with H₂O₂ oxidation for both antibiotics was increased, while in the combustion treatment groups the adsorptions were reduced. Since the carbon content in SPM-505 ℃ and SPM-H₂O₂-505 ℃ samples was lower than the detection limit (< 0.3%), the decrease in adsorption after combustion compared with the original SPM sample can be regarded as the contribution of OM to the adsorption of antibiotics. In TC adsorption, OM contributes 14.8–39.3%, and in NOR adsorption the contribution reaches 63.4–64.8%.

Table 7

The ratio of antibiotic adsorption amount at equilibrium of three samples to SPM
Antibiotics	The ratio of adsorption amount at equilibrium to SPM
Antibiotics	SPM-505℃	SPM-H₂O₂	SPM-H₂O₂-505˚C
TC	85.2%	131.7%	60.7%
NOR	35.2%	102.8%	36.6%

Nearly all the OM was removed in the combustion group, and the adsorption amount was the lowest. However, the OM content of the H₂O₂ treatment group (carbon content of 0.89%) was also much lower than that of the original SPM, yet its adsorption capacity was significantly higher than that of the combustion-treated SPM, indicating that the OM content is not the only determining factor.

The SPM-H₂O₂ sample had a higher antibiotic adsorption capacity than the other samples, possibly because of changes in the BET surface area, metal adsorption point, and OM form.

(1) H₂O₂ treatment increased the Brunauer-Emmett-Teller (BET) surface area of the suspended particulate matter samples by nearly 65%. The BET value of the SPM-H₂O₂-505 ˚C sample was the highest among all groups, but the adsorption capacity was the lowest. It therefore appears that BET surface area has little contribution to the adsorption capacity of SPM for antibiotics.

(2) We suppose that H₂O₂ treatment of SPM may expose more clay particles and iron oxide on the surface of inorganic minerals, thereby providing more adsorption sites. The XPS analysis showed that the Fe³⁺ content of the SPM-H₂O₂ group increased by nearly 40%, which supports this speculation.

(3) The OM form of SPM changed after H₂O₂ treatment. The OM form is expressed as the elemental ratio and functional groups. Figure 2 and Fig. 3 illustrate that some infrared absorption peaks of the SPM treated with H₂O₂ have relatively obvious shifts. For example, the original sample has an absorption peak at 3404 cm^− 1, while the SPM-H₂O₂ sample has a peak at 3410 cm^− 1. There were also some peaks that significantly increased after H₂O₂ treatment (such as the peaks between 3604 − 3602 cm^− 1). This indicates that the functional groups changed after H₂O₂ treatment. In addition, the peak near 1653 cm^− 1 is significantly enhanced after adsorption, proving that the adsorption of tetracycline has a strong π-π conjugation effect. The (O + N)/C ratio of SPM-H₂O₂ was much higher than that of the original sample, indicating that the surface polar functional groups of SPM-H₂O₂ were more abundant, thus improving their adsorption capacity. The original SPM sample had a lower H/C value than the SPM-H₂O₂ sample, as shown in Table 7. The aromaticity of the original sample was stronger and more stable than that of the SPM-H₂O₂ sample. The O/C and (O + N) /C values of the SPM-H₂O₂ sample were higher than those of the original sample, indicating that its hydrophilicity and polarity were stronger than those of the original sample. The surface polarity of soil or sediment treatment increases after H₂O₂ (Cabrera et al. 2014), which is consistent with the elemental analysis results. There is typically less adsorption of hydrophobic organic pollutants by suspended particles when hydrophilicity and polarity are strong because water molecules are actively wrapped around the suspended particles and occupy the adsorption points on their surface. However, the adsorption capacity of the SPM-H₂O₂ sample for antibiotics is higher than that of the original SPM sample. Therefore, we believe that the polar functional groups on SPM play a significant role in promoting the adsorption of hydrophobic organic pollutants. The adsorption capacity for this promotion was greater than the inhibition caused by the competitive adsorption of water molecules. As shown in Table 2, a portion of the carbon remains after H₂O₂ treatment, and from the above discussion, the remaining carbon has a stronger adsorption capacity because of its changed form. The carbon slag obtained after H₂O₂ treatment of the original sample is similar to the black carbon (BC) obtained from the carbon slag after heat treatment and/or chemical treatment (Brändli et al., 2009). BC is a highly aromatic organic carbon that is widely present in natural water sediments and has a strong ability to absorb pollutants (Cornelissen et al. 2005; You et al. 2023). The adsorption capacity of BC in sediments is several orders of magnitude higher than that of other natural organic matter in sediments (Cheng et al. 2021). Studies have shown that a large number of oxygen-containing functional groups in BC, such as -C-O, -C = O, and -COOH, play a significant role in the adsorption of antibiotics such as TC. With an increase in the number of these functional groups, the adsorption capacity of BC can be significantly improved (Huang et al. 2022). After treatment with strong oxidants such as H₂O₂, the acidic functional groups on the surface of BC increase, thus enhancing the adsorption capacity of pollutants (Lim et al. 1996; Liu et al. 2013). Therefore the obtained residual organic matter has changed the form after treatment with hydrogen peroxide, which has a strong adsorption capacity for TC and NOR.

According to the above discussion, the OM content and form of SPM affect the adsorption of TC and NOR by SPM. Moreover, the data for the SPM-H₂O₂ group in Table 7 show that after removing most of the OM, the adsorption capacity increased. Thus, the contribution of the OM form to adsorption was greater than that of its content.

4.2 Differences between the adsorption of TC and NOR

The adsorption capacities of SPM-505 ℃ and SPM-H₂O₂ samples for TC adsorption are 31.7% higher and 14.8% lower than that of the original SPM sample, respectively, while the adsorption capacities for NOR adsorption are 2.8% higher and 74.8% lower, respectively, indicating that the adsorption contribution of OM in SPM to TC and NOR are different. It is clear that the contribution of OM to the TC adsorption process is not as significant as that of NOR adsorption. This may be related to the different adsorption modes of TC and NOR in SPM.

The reason for the different adsorption modes may be that the process of TC adsorption is dominated by pore filling (Cheng et al. 2021), and according to Table 1, combustion can significantly increase the BET surface area of the SPM. It has also been reported that the TC adsorption capacity of biochar is positively correlated with its specific surface area and total pore volume. Therefore, the contribution of the BET surface area to TC adsorption was greater than that of NOR adsorption to a certain extent, although it contributed less to the adsorption process than that of OM.

Finally, studies have pointed out that metal ions such as Fe and Al have a significant effect on the adsorption of TC in soil, such as Fe and Al (Nowara et al. 1997). Therefore, the Fe content (Table 4) combined with OM had an impact on the adsorption of TC. The XPS analysis showed that the content of Fe³⁺ increased by nearly 40% after H₂O₂ treatment, which may be one of the reasons for the increased TC adsorption by H₂O₂.

In this study, we investigated the effects of OM on the adsorption of two typical antibiotics, TC and NOR, onto SPM via H₂O₂ oxidation and combustion. This study showed that both the OM content and form had significant effects on antibiotic adsorption. Combustion removed most OM, which decreased the adsorption capability by 14.8–64.8%, while H₂O₂ oxidation removed the amorphous OM and changed the OM form, including functional groups and the structure of carbon, and enhanced the adsorption capacity of antibiotics. Moreover, the effect of OM on the SPM adsorption of antibiotics was different for TC and NOR, with a larger fluctuation in adsorption of NOR. This study provides new ideas for exploring the migration and transformation of antibiotics via SPM in lakes.

OM, organic matter; SPM, suspended particulate matter; TC, tetracycline; NOR, norfloxacin; SEM, scanning electron microscopy; FTIR, Fourier transform infrared; XPS, X-ray photoelectron spectroscopy; BET, Brunauer-Emmett-Teller; PFO, pseudo-first-order; PSO, pseudo-second-order.

Acknowledgments

This study was supported by the National Natural Science Foundation of China (42207275).

Acknowledgements We would like to thank the reviewers and editor for suggestions to help us improve the article.

Author contribution All authors contributed to the study's conceptionand design. Validation, formal analysis, visualization, software, and writing - original draft by Zihao Shen and Yali Sun. Resources and writing - review & editing by Xiaolan Zheng and Ye Yang. Validation, formal analysis, and visualization by Jingge Shang. Resources, writing - review & editing, supervision, and data curation by Ruixin Guo, Yanhua Liu and Jianqiu Chen.Conceptualization, methodology, software, investigation, writing - original draft, supervision, and data curation by Qianjiahua Liao.And all authorscommented on previous versions of the manuscript. All authors readand approved the final manuscript.

Funding This study was supported by the National Natural Science Foundation of China (42207275).

Availability of data and materials Not applicable.

Ethics approval and consent to participate Not applicable.

Consent for publication Not applicable.

Competing interests The authors declare no competing interests.

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Influence of Changes in OM Content and Form on the Adsorption of Two Typical Antibiotics to SPM

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Abstract

Figures

1 Introduction

2 Experimental materials and methods

2.1 Collection and treatment of samples

2.2 Kinetic study of adsorption

2.3 Isothermal adsorption study

2.4 Antibiotic detection and sample identification

2.5 Calculations of the formulae of adsorption

3 Results

3.1 Analysis of the physicochemical properties of the four samples

3.1.1 Scanning electron microscopic (SEM) and Brunauer–Emmett–Teller (BET) surface area analyses of samples

3.1.2 Elemental composition and concentration

3.1.3 Infrared spectrum analysis and XPS analysis

3.2 Kinetics studies

3.3 Isothermal adsorption studies

4 Discussion

4.1 Effects of OM on the adsorption of TC and NOR by different SPM treatments

4.2 Differences between the adsorption of TC and NOR

5 Conclusion

Abbreviations

Declarations

References

Supplementary Files

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Sample	\(\text{F}\text{e}\)					\(\text{M}\text{n}\) Area	\(\text{A}\text{l}\) Area
Sample	\({Fe}^{2+}\) Area	Binding energy (eV)	\({Fe}^{3+}\) Area	Binding energy (eV)	\(\frac{{Fe}^{2+}}{{Fe}^{2+}+{Fe}^{3+}}\)	\(\text{M}\text{n}\) Area	\(\text{A}\text{l}\) Area
SPM	25052.3	712.4	23750.1	714.2	0.5	849.0	22197.5
SPM-505˚C	28923.8	712.0	24714.8	714.0	0.5	500.10	22944.0
SPM-H₂O₂	25770.9	712.3	33236.2	714.1	0.4	503.30	21253.7
SPM-H₂O₂-505˚C	25064.2	712.2	18891.7	714.1	0.6	539.4	21164.5