Journal of Hazardous Materials 417 (2021) 125908

Available online 21 April 20210304-3894/© 2021 Published by Elsevier B.V.

Accumulation of sulfamethazine and ciprofloxacin on grain surface decreases the transport of biochar colloids in saturated porous media

Kang Zhao a, Ling Gao a, Qianru Zhang b, Jianying Shang a,*

a College of Land Science and Technology, China Agricultural University, Beijing 100193, China b Key Laboratory of Nonpoint Source Pollution Control, Ministry of Agriculture and Rural Affairs, Institute of Agricultural Resources and Regional Planning, Chinese Academy of Agricultural Sciences, Beijing, China

A R T I C L E I N F O

Editor: Dr. Rinklebe Jorg

Keywords: Sulfamethazine Ciprofloxacin Biochar colloids Transport

A B S T R A C T

The increasing amount of antibiotics entering the environment through manure usage and sludge application from wastewater treatment plants (WWTP) attracts much concern due to their potential threat to ecological security and human health. When biochar, a soil and water amendment, is introduced into the soil for reme-diation, the antibiotics are usually co-present with the biochar colloids (BC) or pre-accumulated in soils. How-ever, little is known about the effect of antibiotics on the behavior of BC. Column experiments were conducted at three different pH values to study the effect of sulfamethazine (SMT) or ciprofloxacin (CIP) on BC transport. Under certain conditions (co-present in the influent and pre-sorbed on quartz sand), large numbers of cation and zwitterion forms of the less mobile CIP at pH 5 and 7 led to less negatively-charged surface of BC and quartz sand, resulted in higher BC retention compared to the highly mobile SMT. The decrease in BC transport became more significant with a higher amount of SMT or CIP pre-sorption. Therefore, when biochar is applied into soils polluted by antibiotics, the pH-dependency and the loading amount of antibiotics in soil matrix should be paid attention to as they might affect the transport of BC and the related facilitated-contaminants transport.

1. Introduction

Antibiotics have been widely used in pharmaceutical industry to prevent human and livestock diseases, and their production and con-sumption are still increasing (Sarmah et al., 2006; Ben et al., 2008). Antibiotics cannot be adsorbed or metabolized completely in-vivo, and most of them are directly excreted through feces and urine, which are then used as agricultural manure or discharged into wastewater treat-ment plants (WWTP) (Bendz et al., 2005; Jjemba, 2006; Hu et al., 2010; Li et al., 2013). The contents of antibiotics in manure range from μg/kg to mg/kg, while high levels of antibiotics are detected in vegetable fields after repeated manure applications (Hu et al., 2010; Ji et al., 2012). In WWTP, the most common method for antibiotics removal is through sorption and biodegradation by activated sludge (Li et al., 2013; Michael et al., 2013). Sludge becomes a major reservoir of antibiotics, and the concentration of antibiotics in sludge can be more than several ppm levels (Kinney et al., 2006; Gao et al., 2012; Li et al., 2013). It has been recognized that the disposal of sludge (including agricultural applica-tion and landfill) can result in the accumulation and release of antibi-otics into the environment. Moreover, some antibiotics can even move

into groundwater during irrigation and precipitation, resulting in water pollution (Hu et al., 2010; Gottschall et al., 2012; Frey et al., 2015; Gothwal and Shashidhar, 2015; Yao et al., 2017). Hu et al. (2010) showed that the ciprofloxacin residue was above 40 ng/L in 40 m deep groundwater samples. The occurrence of antibiotics in soils and aquatic environments is a serious concern since antibiotics might be toxic to the microorganisms and cause a dramatical increase in the antibiotic resis-tance genes and bacteria (Thiele-Bruhn, 2003; Guo et al., 2018; Wang et al., 2020).

It is to be noted that various antibiotics are present in the natural environment, sulfamethazine (SMT) and ciprofloxacin (CIP) are two commonly used antibiotics detected in different environmental systems, e.g., soils, sediments, rivers, and groundwater (Vasudevan et al., 2009; Awad et al., 2014; Hanna et al., 2018; Zhi et al., 2019). Their unique properties, such as their functional groups and water solubility, deter-mine their distinct leaching and sorption in soils. After entering the soil and aquatic systems, SMT with high mobility can leach with water through the vadose zone into groundwater, posing a potential threat to human health (Chen et al., 2011; Vithanage et al., 2014; Xing et al., 2015; Zhi et al., 2019). On the contrary, CIP has a high affinity towards

* Corresponding author. E-mail address: jyshang@cau.edu.cn (J. Shang).

Contents lists available at ScienceDirect

Journal of Hazardous Materials

journal homepage: www.elsevier.com/locate/jhazmat

https://doi.org/10.1016/j.jhazmat.2021.125908 Received 24 January 2021; Received in revised form 23 March 2021; Accepted 14 April 2021

Journal of Hazardous Materials 417 (2021) 125908

2

soil matrix and thus is accumulated in soils, resulting in microbial toxicity and uptake by plants (Vasudevan et al., 2009; Xing et al., 2015, 2020; Zhi et al., 2019).

Biochar is a carbon-rich material produced from the pyrolysis of biomass and wastes under little or no oxygen conditions (Lehmann, 2007). It can improve soil structure, enhance fertilizer retention, and immobilize contaminants as soil amendment due to its unique properties (Major et al., 2010; Sohi, 2012; Rajapaksha et al., 2015). After biochar is applied into soil, it is broken down into fine particles by biological and non-biological activities, which can cause small biochar particles to transport into deeper soils (Ameloot et al., 2013; Heymann et al., 2014). The fine biochar particles can range from nanometer to micrometer in size (Ameloot et al., 2013; Yang et al., 2017; Fang et al., 2020). Although nanoscale and colloidal biochar account for a relatively small part of the total biochar mass, they have a strong adsorption capacity of pollutants due to their large specific area and could facilitate the transport of contaminants (e.g., naphthalene) in the soil and aquatic systems due to their high mobility (Xing et al., 2016; Yang et al., 2017, 2020). Thus, understanding the fate and transport of biochar colloids (BC) is of sig-nificance when their associated risks in facilitating the transport of environmental contaminants in the subsurface environments are considered.

Because of the excessive antibiotics in soil and subsurface environ-ments, BC can interact with antibiotics after biochar is used for reme-diation. On the one hand, BC co-present with antibiotics could result in the adsorption of antibiotics on BC, which has been extensively inves-tigated, and the related studies show that the stability and aggregation of BC are affected by antibiotics after sorption on colloids’ surface (Teixido et al., 2011; Ahmed et al., 2017; Fang et al., 2018; Qi et al., 2018; Yang et al., 2020; Zhang et al., 2020). It is expected that the transport of BC is also possibly affected by the co-presence of antibiotics. On the other hand, in the soil matrix with large amounts of pre-accumulated antibiotics, the transport behavior of BC might also be affected. Little research has been conducted on this issue. Chen et al. (2015) studied the transport of kaolinite and montmorillonite in porous media after CIP pre-sorption, and although only a slight difference was observed between their transport behaviors, this might be due to the relatively low CIP pre-sorption on porous media. Due to the continuous application of antibiotics, their contents in the soil are increasing, while BC has higher surface activity, i.e., large surface area, high hydropho-bicity, and aromaticity. Thus, the transport behavior of BC in antibiotics pre-accumulated soil may be changed.

Various antibiotics are accumulated in the soil due to manure usage or sludge application from WWTP, and they are either co-present with BC or are pre-accumulated in soil matrix when biochar is applied into soil for remediation. However, it is still unclear that how different an-tibiotics (such as SMT and CIP) affect the mobility of BC when they are co-present with BC or are pre-sorbed on soil matrix. This is of great environmental significance to evaluate related contamination risks. In this study, co-transport experiments were conducted by simultaneously injecting antibiotics (SMT or CIP) and BC into the column to study the effect of antibiotics co-presence on BC transport. In addition, column experiments were conducted by injecting BC after various amounts of antibiotics (SMT or CIP) were pre-sorbed on quartz sand to study BC transport in antibiotic pre-accumulated soils. The zeta potentials of sand surface were measured by electrokinetic analyzer to study the effect of SMT or CIP sorption on the surface charge of sand. The specific objec-tives of this study were to: (1) identify BC transport with co-present SMT/CIP; (2) explore the mechanism of SMT/CIP pre-sorption influ-encing on BC transport; (3) compare the distinct effect between SMT and CIP on BC transport under various pH conditions.

2. Material and methods

2.1. Biochar colloids, sulfamethazine, and ciprofloxacin

Biochar was prepared with wheat straw (Zhengzhou, Henan prov-ince, China) under anaerobic condition in a muffle furnace at 600 ◦C pyrolysis temperature. The biochar was gently grounded into powder form by a ball mill (MM400, Verder Shanghai Instruments, Shanghai, China). The obtained biochar powder was added to ultrapure water, stirred, and then sonicated in a water bath for 30 min for dispersion. The suspension was transferred to a 1 L measuring cylinder and was settled for 24 h. The upper suspension was collected by siphon method to obtain colloidal biochar according to the Stokes’ law (Qu et al., 2016; Yang et al., 2017). The morphology of BC was determined with transmission electron microscopy (TEM, JEOL JEM-1230, JPN). Fourier transform infrared (FTIR) spectra of BC were recorded between 400 and 4000 cm− 1 with a Spectrum Spotlight 200 FTIR microscopy system (Perki-nElmer, USA).

Sulfamethazine (SMT, 99% purity, Aladdin Industrial, Shanghai, China) was used to prepare 1 g/L SMT stock solution in methanol (Fisher Chemical, Shanghai, China), and then refrigerated at − 20 ◦C for further use. CIP stock solution with 1 g/L concentration was prepared by cip-rofloxacin (CIP, 99% purity, Dalian Meilun Biotechnology, Dalian, China) in ultrapure water, and then refrigerated at 4 ◦C temperature. SMT has two dissociation constants (pKa1 = 2.3, pKa2 = 7.4) and exists as cation (SMT+), neutral molecule (SMT0), and anion (SMT-) (Teixido et al., 2011). CIP shows the dissociation constants (pKa1 = 6.1, pKa2 =

8.7) and its main forms are cation (CIP+), zwitterion (CIP±), and anion (CIP-) (Xing et al., 2020). The fractions of non-ionized and ionized forms of SMT or CIP at different pH values affect their sorption to BC and quartz sand. For a given pH, the fraction of different forms of SMT or CIP can be estimated by the following equation (Lertpaitoonpan et al., 2009):

α− =[A− ]

CT,A=

[A− ]

[HA] + [A− ]=

11 + 1

10(pH− pKa )

(1)

where “[A-]” represents concentration of anionic SMT or CIP, “CT,A” represents mole number of species containing “A” (SMT or CIP), and “[HA]” represents concentration of non-ionized SMT or CIP.

2.2. Porous media

Quartz sand with 0.425–0.600 mm diameter was used as porous medium for the column experiments. The sand was soaked in 2 M HCl solution for 24 h to remove organic matters and metal oxides from sand surface (Chen et al., 2018). The sand was then thoroughly washed with deionized water and then was dried in an oven at 105 ◦C before use.

2.3. Measurements of zeta potentials and hydrodynamic diameters

The zeta potentials and hydrodynamic diameters of 100 mg/L BC suspensions in the absence/presence of 1 mg/L SMT/CIP at different pH values (5, 7, and 10) were measured with a Zetasizer Nano ZS90 (Mal-vern Instruments Ltd., Malvern, UK). The SurPASS 3.0 electrokinetic analyzer (Anton Paar GmbH, Austria) was used for the measurement of the zeta potentials of sand surfaces (including clean sand and sand after SMT/CIP pre-sorption) based on the streaming potential method (Pet-rinic et al., 2018). The Derjaguin-Landau-VerweyOverbeek (DLVO) theory was used to calculate the total interaction energy between BC and sand (Supplementary Material).

K. Zhao et al.

Journal of Hazardous Materials 417 (2021) 125908

3

2.4. Column experiments

Column experiments were conducted with a stainless steel column (12 cm long, 2.5 cm inside diameter). The vertically-oriented columns were wet packed with clean sand in an increment of 2 cm layers. The experiments were run in an upward water flow direction, and the water velocity was maintained at 1.0 mL/min with a peristaltic pump. During the experiments, the column was equilibrated by more than 5 pore volumes (PVs) of the background solution (1 mM NaCl) to establish steady-state saturated flow condition.

For the co-transport experiments (SMT/CIP co-present in the in-fluents), 3 PVs of 100 mg/L BC suspensions with 1 mg/L SMT/CIP were injected into the column, followed by flushing with 5 PVs of 1 mM NaCl background solution. For the transport experiments with SMT/CIP pre- sorption on the quartz sand, the details were as follows: (i) injecting 3/ 15 PVs of 1 mg/L SMT/CIP; (ii) flushing with 3 PVs of 1 mM NaCl background solution to flush unattached SMT/CIP; (iii) injecting 3 PVs of 100 mg/L BC suspension followed by flushing with 5 PVs of 1 mM NaCl background solution.

The effluents were collected by a fraction collector with glass tubes every 5 min. Following transport experiments, the columns were sectioned into 12 layers of 1 cm segments. To determine the retention profiles (RPs) of BC, 20 mL ultrapure water was added into each fraction of the quartz sand to liberate BC particles. The concentrations of BC and CIP in the effluents were determined using UV–VIS at the wavelengths of

790 nm and 275 nm, respectively. The concentration of SMT was measured with high phase liquid chromatography (HPLC, Ultimate 3000, Thermo Fisher Scientific, USA), equipped with a UV2000 ultra-violet detector. A Venusil MP C18 stainless steel column (250 mm × 4.6 mm, 5 µm) was used and the temperature was set at 28 ◦C. The elution condition was methanol / 0.08% acetic acid solution (i.e., volume ratio = 40/60) and the flow rate was 1 mL/min. The UV detection wavelength was 265 nm and the retention time of SMT was 8 min.

3. Results and discussion

3.1. Characteristics of biochar colloids

The TEM image of BC indicated an irregular surface morphology, and the particle diameter was 446 ± 18 nm (Fig. 1a). The particles size distribution of BC obtained from Zetasizer was ranged from 164 to 712 nm, and the average hydrodynamic diameter was 320 nm (Fig. 1b). The FTIR analysis showed the O-H stretching at 3354 cm− 1 (Kolodynska et al., 2017), -COOH at 1687 cm− 1 (Gamiz et al., 2019), aromatic C˭C at 1570 cm− 1 (Lyu et al., 2018), and aromatic C-H bending at 872 cm− 1, 746 cm− 1, and 620 cm− 1 (Jia et al., 2018; Lyu et al., 2018) (Fig. S1).

Fig. 1. TEM image (a) and particle size distributions (b) of biochar col-loids (BC).

Fig. 2. Zeta potentials of biochar colloids (BC) in the absence and presence of sulfamethazine (SMT) and ciprofloxacin (CIP) (a), zeta potentials of quartz sand with/without 3 and 15 pore volumes (PVs) of SMT or CIP pre-sorption on sand at different pH values (5, 7, 10) (b). “BC” represents biochar colloids; “BC + SMT”/ “BC + CIP” represents BC with SMT/CIP in the influent; “0PV” represents no antibiotics pre-sorption on sand surface; “3PVs SMT”, “3PVs CIP”, “15PVs SMT”, and “15PVs CIP” represent 3 PVs/15 PVs SMT or CIP pre- sorption on sand surface.

K. Zhao et al.

Journal of Hazardous Materials 417 (2021) 125908

4

3.2. Effect of SMT and CIP on zeta potentials of BC and quartz sand

The zeta potentials of BC in the absence of antibiotics decreased from − 32.0 ± 0.4 mV to − 38.2 ± 0.9 mV with solution pH value changing from 5 to 10 (Fig. 2a). This result was consistent with previous studies reported that BC showed more negative surface charge with the increase in pH (Wang et al., 2019; Yang et al., 2020). At pH 5 and 7, the zeta potentials of BC became less negative with the addition of SMT, and this was attributed to the sorption of SMT on BC. The main form of SMT was neutral molecule, SMT0, with 99.8% and 71.5% of SMT existing as SMT0

at pH 5 and pH 7, respectively (Table 1). The SMT0 was adsorbed on BC surface through H-bonding and electron donor-acceptor (Fig. 3), thus shielding part of the negative surface charge on BC (Teixido et al., 2011; Ahmed et al., 2017). After the addition of CIP, the zeta potentials of BC

also became less negative at both pH 5 and 7 (from − 32.0 ± 0.4 mV without CIP to − 30.0 ± 0.5 mV with CIP at pH 5, and from − 36.6 ± 0.7 mV to − 34.8 ± 1.0 mV at pH 7) (Table 2). The zeta po-tentials of BC with SMT was less negative than those of BC with CIP (− 26.5 ± 0.5 mV > − 30.0 ± 0.5 mV at pH 5, and − 32.6 ± 0.6 mV > − 34.8 ± 1.0 mV at pH 7). This might be attributed to the fact that unlike SMT, CIP could be adsorbed on the BC surface not only through H-bonding and electron donor-acceptor, but also through electrostatic attraction (Fig. 3). There were more CIP existed as cation at pH 5 and 7 (Table 1), which resulted in less negative zeta potentials of BC with CIP than that with SMT.

At pH 10, the zeta potentials of BC (− 38.2 ± 0.9 mV) showed little change after the addition of SMT and CIP (− 38.0 ± 1.6 mV for BC with SMT and − 38.8 ± 0.8 mV for BC with CIP) (Table 2). The primary forms

Fig. 3. Schematic of adsorption mechanisms for sulfamethazine (SMT) and ciprofloxacin (CIP) sorption to biochar colloids (BC) at pH 5 and 10.

Table 1 Structures and basic properties of sulfamethazine and ciprofloxacin.

Antibiotics Molecular weight

log Kow

Structure Antibiotics speciation

pH 5 pH 7 pH 10

Sulfamethazine (SMT) C12H14N4O2S

278.3 0.80 Anion 0.2% Neutral 99.8%

Anion 28.5% Neutral 71.5%

Anion 99.7% Neutral 0.3%

Ciprofloxacin (CIP) C17H18FN3O3

331.4 0.28 Cation 92.6% Zwitterion 7.4%

Cation 11.2% Zwitterion 88.8%

Anion 95.2% Zwitterion 4.8%

The pKa values are from Yang et al. (2020) and Xing et al. (2020); The log Kow values are from Xing et al. (2015, 2020).

K. Zhao et al.

Journal of Hazardous Materials 417 (2021) 125908

5

of SMT and CIP at strong alkaline condition were SMT- (99.7% of total SMT) and CIP- (95.2% of total CIP) (Table 1), respectively, and thus the electrostatic repulsion between antibiotics and BC weakened the sorp-tion of antibiotics on BC surface. This contributed to little change in zeta potentials of BC before and after the addition of SMT or CIP at pH 10.

Fig. 2b shows the zeta potentials of quartz sand surface with/without SMT or CIP pre-sorption at different pH values of 5, 7, and 10. Without SMT or CIP pre-sorption, the zeta potentials of sand surface became more negative (− 21.2 ± 0.7 mV at pH 5 to − 33.3 ± 1.4 mV at pH 10) with the increase in pH, which was due to the stronger deprotonation at the alkaline condition. At pH 5 and 7, the sand surface became less negative after 3 PVs of SMT or CIP pre-sorption than that without SMT or CIP pre-sorption. This was because the sorption of neutral molecule SMT0, cation CIP+, and zwitterion CIP± on quartz sand resulted in the charge shielding on sand surface. Moreover, the zeta potentials of sand surface became less negative after 15 PVs of SMT or CIP pre-sorption than those after 3 PVs pre-sorption because of the higher sorption amount of antibiotics on quartz sand after 15 PVs (Table 3). At pH 5 and 7, the sand surface after 3 PVs and 15 PVs of CIP pre-sorption showed less negative charge than that after SMT pre-sorption (Fig. 2b). This was because CIP induced the larger sorption amount and stronger charge shielding effect on sand surface than SMT. The pre-sorption of cationic and zwitterionic CIP on sand surface was larger than that of the SMT which existed as neutral molecule (for example, 0.475 ng/g after 3 PVs of CIP pre-sorption >0.103 ng/g after 3 PVs of SMT pre-sorption at pH 5) (Table 3). At pH 10, however, the zeta potentials of sand surface showed little change after 3 PVs of SMT or CIP pre-sorption because the anion SMT- or CIP- were difficult to adsorb on negatively charged sand

surface due to electrostatic repulsion (Fig. 3 and Table 3).

3.3. Transport of SMT and CIP in column experiment

The breakthrough curves of SMT and CIP in the column experiments with 3 PVs and 15 PVs injection at pH 7 are shown in Fig. 4a and b. After injecting the antibiotics in the column, the breakthrough concentrations of SMT quickly increased and remained at a peak value. The normalized peak concentration (C/C0) of SMT was almost 1.0, which meant that SMT had high mobility in porous media. This was also consistent with previous studies which reported that SMT had strong mobility in sand and soil media (Stoob et al., 2007; Chen et al., 2011). It was because the anion SMT- (28.5% of the total SMT) and neutral molecule SMT0 (71.5% of the total SMT) were difficult to be adsorbed on the sand surface. However, the breakthrough concentrations of CIP with 3 PVs injection showed a gradually increasing trend and the C/C0 reached a constant level at about 13 PV with further injection. This indicated that the mobility of CIP was weaker than that of SMT in porous media.

We also analyzed the transport of SMT or CIP with 3 PVs injection at different pH values of 5, 7, and 10 (Fig. 4b and c). The transport of SMT or CIP increased with the increase in pH, and this was related to the more negatively-charged anion forms of SMT or CIP under high pH condition. Mass balance calculations at pH 5 and 7 also showed that the amounts of CIP retained in sand column were larger than those of SMT (Table 3). This might be because the cation and zwitterion forms of CIP caused the larger amounts of retained CIP in the column. In addition, the logarithm of octanol-water partition coefficient (log Kow) also affected the adsorption of SMT and CIP on porous media, which was usually used

Table 2 Hydrodynamic diameters and zeta potentials of biochar colloids (BC) in the absence/presence of SMT/CIP, DLVO interaction energies between BC and sand at different pH values (5, 7, and 10).

BC SMT CIP pH Da ζBCb Φmax

c Φmind Meff

e Mretf

mg/L mg/L mg/L nm mV kT kT % %

100 0 0 5 432 ± 22 -32.0 ± 0.4 117.12 -0.68 49.56 41.62 7 415 ± 14 -36.6 ± 0.7 205.66 -0.57 61.20 30.28 10 399 ± 2 -38.2 ± 0.9 259.55 -0.52 64.43 26.52

1 0 5 441 ± 19 -30.0 ± 0.5 108.77 -0.71 44.82 47.49 7 420 ± 11 -34.8 ± 1.0 193.99 -0.58 58.83 30.33 10 396 ± 1 -38.0 ± 1.6 255.20 -0.51 63.51 27.13

0 1 5 471 ± 18 -26.5 ± 0.5 95.57 -0.79 19.65 70.50 7 431 ± 18 -32.6 ± 0.6 183.11 -0.61 51.86 42.31 10 390 ± 10 -38.8 ± 0.8 257.74 -0.50 67.30 27.18

a “D” represents hydrodynamic diameters of BC. b “ζBC” represents zeta potentials of BC. c,d “Φmax and Φmin” represent maximum primary energy barrier and secondary energy minimum between BC and quartz sand. e,f “Meff and Mret” represent percentages of BC recovered from effluent and retained in column experiments.

Table 3 The SMT/CIP pre-sorption amount, zeta potentials of sand, and DLVO interaction energies between biochar colloids (BC) and sand after SMT/CIP pre-sorption at different pH values (5, 7, and 10).

BC SMT CIP pH SMT/CIP injection SMT/CIP pre-sorptiona ζsandb Φmax

c Φmind Meff

e Mretf

mg/L mg/L mg/L PV ng/g mV kT kT % %

100 1 0 5 3 0.103 -18.8 ± 1.0 97.05 -0.70 43.82 48.61 7 3 0.035 -26.2 ± 1.2 185.38 -0.58 51.77 38.22 10 3 0.024 -32.0 ± 1.4 246.07 -0.52 57.36 27.63 7 15 0.132 -17.1 ± 1.1 98.35 -0.67 34.02 53.95

0 1 5 3 0.475 -14.9 ± 0.5 65.34 -0.77 21.66 62.48 7 3 0.371 -24.5 ± 0.5 168.44 -0.60 39.98 47.62 10 3 0.016 -32.6 ± 1.4 251.46 -0.52 60.75 28.68 7 15 0.600 -13.2 ± 0.2 63.46 -0.74 15.55 70.53

a “SMT/CIP pre-sorption” represents the SMT/CIP pre-sorption amount on quartz sand. b “ζsand” represents zeta potentials of sand. c,d “Φmax and Φmin” represent maximum primary energy barrier and secondary energy minimum between BC and sand. e,f “Meff and Mret” represent percentages of BC recovered from effluent and retained in column experiments.

K. Zhao et al.

Journal of Hazardous Materials 417 (2021) 125908

6

to describe the water solubility and hydrophilicity of organic contami-nants. The log Kow value of SMT (0.80) was higher than CIP (0.28) (Table 1), and this meant that CIP was more hydrophilic and more CIP could be adsorbed on the hydrophilic sand surface. Yao et al. (2017) also showed that the fluoroquinolones (FQs) and tetracycline (TC) with lower log Kow values were more soluble in water and showed less sorption on sand surface than the macrolides (MLs) and sulfonamides (SAs) with higher log Kow values. At pH 10, however, little difference was observed in the amounts of SMT and CIP retained in the column and this was because the anion forms of SMT- and CIP- were unlikely to be adsorbed on sand surface. These results showed that the different species of SMT and CIP at different pH values played a critical role in affecting their transport in porous media.

3.4. BC transport in the presence of SMT and CIP

The amounts of antibiotics in soils are continuously increasing, owing to their widespread use. They are co-present with BC in the soil when biochar is introduced for soil remediation. Biochar colloids can adsorb SMT or CIP on their surface and the characteristics of BC may be modified during this interaction, which might influence BC transport. The breakthrough curves and retention profiles of BC in the absence of SMT or CIP under different pH conditions are shown in Fig. 5. The breakthrough curves are shown with the normalized effluent concen-tration (C/C0) of BC as a function of PVs. The retention profiles are

exhibited with the mass ratio (N/Ni) of BC as a function of distance from the column inlet. The mass percentages of BC recovered from the effluent (Meff) were larger than 49% (Table 2), indicating that a large amount of BC could transport through the columns. This was because both BC and sand surface were negatively charged, indicating an unfa-vorable attachment condition for BC (Fig. 2a and b). However, more than 30% BC was still retained in the column, which could be explained by the attachment due to the secondary minima of the DLVO interaction, and surface roughness and heterogeneity of the quartz sand (Fang et al., 2016; Yang et al., 2017). In addition, the mass percentage of BC recovered from the effluent (Meff) became higher with the increase in pH, because the zeta potentials of BC and sand surface became more negative at higher pH and led to a stronger electrostatic repulsion be-tween BC and quartz sand (Fig. 2), and similar phenomena were shown in the previous studies (Ryan and Elimelech, 1996; Wan and Tokunaga, 2002; Xu et al., 2018).

The breakthrough curves and retention profiles of BC in the presence of SMT or CIP under different pH conditions are shown in Fig. 5. At pH 5 and 7, the presence of SMT or CIP decreased BC transport in porous media. The plateau of BC breakthrough curves at pH 7 decreased from 0.77 to 0.70 and to 0.64 with the addition of SMT and CIP, respectively. After sorption onto the BC surface, SMT and CIP shielded part of nega-tive charges and resulted in the less negative zeta potentials of BC. The DLVO theory also indicated that the energy barrier between BC and sand surface was decreased in the presence of SMT or CIP (Fig. S3 and

Fig. 4. Breakthrough curves of sulfamethazine (SMT) and ciprofloxacin (CIP) with 3 PVs (a) and 15 PVs (b) injection at pH 7, and breakthrough curves of SMT and CIP with 3 PVs (c and d) injection at different pH values (5, 7, and 10). “3PVs” and “15PVs” represent 3 PVs and 15 PVs of antibiotics injection; “C” represents the concentration of antibiotics in the effluent; “C0” represents the concentration of antibiotics in the influent.

K. Zhao et al.

Journal of Hazardous Materials 417 (2021) 125908

7

Table 2), which further resulted in more BC retained in porous media. It was noted that the transport of BC with CIP at pH 5 and 7 was weaker than that with SMT. For example, the plateau of BC breakthrough curves at pH 5 decreased from 0.64 to 0.25 with CIP and to 0.52 with SMT. This was consistent with the results that BC with CIP had less negative charge than that with SMT. The lower energy barrier between BC with CIP and sand surface resulted in more BC retention in the column than that with SMT (Fig. S3 and Table 2). At pH 10, the mass percentage of BC

recovered from the effluent (Meff) decreased from 0.81 to 0.77 and to 0.79 in the presence of SMT and CIP, respectively. The result indicated that SMT or CIP showed little effect on BC transport in the column at pH 10, which might be because the zeta potentials of BC showed little change with SMT- or CIP-. These results indicated that the inhibition effect of SMT or CIP was significant under acid and neutral conditions and unimportant under alkaline condition.

Fig. 5. Breakthrough curves (a, c, and e) and retention profiles (b, d, and f) of biochar colloids (BC) in the absence/presence of sulfamethazine (SMT) and cipro-floxacin (CIP) at different pH values (5, 7, and 10). “BC” represents biochar colloids without antibiotics in the influent; “SMT + BC” / “CIP + BC” represents the biochar colloids with SMT/CIP in the influent. “C” represents the concentration of BC in the effluent; “C0” represents the concentration of BC in the influent.

K. Zhao et al.

Journal of Hazardous Materials 417 (2021) 125908

8

3.5. Effect of SMT and CIP pre-sorption on BC transport

Since the co-presence of SMT or CIP with BC decreased the transport of BC (especially at low pH), the accumulated SMT or CIP in the soil might also interact with BC and affect its transport when biochar is applied into soil for remediation. BC transport experiments were con-ducted in quartz sands with 3 PVs and 15 PVs of SMT or CIP pre-sorption before BC injection. At pH 5 and 7, the mass recovery of BC in the effluent after 3 PVs of SMT or CIP pre-sorption was lower than that without SMT or CIP pre-sorption (Fig. 6). This indicated that more BC was retained in porous media due to SMT or CIP pre-sorption. The inhibited transport of BC was consistent with the finding from Chen et al. (2015). Their result exhibited that the transport of montmorillonite was inhibited by CIP sorption, because montmorillonite showed a stronger sorption ability for CIP than sand. However, they did not consider the change in zeta potentials of sand. It was observed in this study that after 3 PVs of SMT or CIP pre-sorption, the zeta potentials of sand became less negative because of the charge shielding effect of pre-sorbed SMT or CIP on the negatively charged surface of sand (Fig. 2b). Thus, the electro-static repulsion between BC and sand surface with SMT/CIP pre-sorption was weaker than without SMT/CIP pre-sorption, which resulted in higher BC retention in the porous media with SMT/CIP pre-sorption.

At pH 7, the BC retention in porous media after 15 PVs of SMT or CIP pre-sorption was stronger than that after 3 PVs of SMT or CIP pre-

sorption (Fig. 7), because after 15 PVs pre-sorption, the larger amount of SMT or CIP was adsorbed on the sand surface than that after 3 PVs pre-sorption (0.132 ng/g > 0.035 ng/g of SMT and 0.6 ng/ g > 0.371 ng/g of CIP) (Table 3). The zeta potentials of quartz sand were less negative and the electrostatic repulsion between BC and sand became weaker after 15 PVs pre-sorption than those after 3 PVs pre- sorption (Fig. 8). The DLVO interaction calculation also showed that the energy barriers between BC and quartz sand were lower after more SMT or CIP pre-sorption at pH 5 and 7 (Fig. S4 and Table 3). The mass percentages of BC in the effluent after CIP pre-sorption at pH 5 and 7 were always smaller than those after SMT pre-sorption. This was consistent with the result that the zeta potentials of quartz sand after CIP pre-sorption were less negative than that after SMT pre-sorption at pH 5 and 7. At pH 10, SMT or CIP pre-sorption had little effect on BC transport due to the little change in the zeta potentials of sand surface.

The retention profiles confirmed that the higher amount of BC was deposited near the column inlet (0-2 cm) when more SMT or CIP was accumulated in sand (Fig. 7). The BC retention amount after SMT or CIP pre-sorption showed a decreasing trend: 15 PVs > 3 PVs > 0 PV. Espe-cially, the deposited BC after 15 PVs of CIP pre-sorption decreased hyper-exponentially with increasing column depth (Fig. 7d). This was attributed to a large amount of CIP accumulated on the sand surface, which could provide the sorption site for BC retention near the column inlet. This trend can not be explained by physical straining which could

Fig. 6. Breakthrough curves (a and c) and retention profiles (b and d) of biochar colloids (BC) after 3 PVs of sulfamethazine (SMT) or ciprofloxacin (CIP) pre-sorption on sand surface at different pH values (5, 7, and 10).

K. Zhao et al.

Journal of Hazardous Materials 417 (2021) 125908

9

induce more colloid deposition, because the diameter ratio of colloid to collector (dp/dc) in this study was smaller than 0.002 (Bradford et al., 2003), indicating that straining was less important in this study.

4. Conclusions

Antibiotics have been widely used to treat livestock diseases, resulting in their entry into the environment through manure usage or sludge application from wastewater treatment plants. This has caused the accumulation of antibiotics in the subsurface. Because of the excel-lent adsorption ability, biochar has attracted extensive interest as an effective agent for removing pollutants in aqueous systems and immo-bilizing contaminants in soils. Antibiotics are usually co-present with BC or are pre-accumulated in soil matrix when biochar is applied into soils for remediation. This study investigated the effect of co-presence of SMT and CIP with biochar as well as pre-accumulation of SMT and CIP on BC transport under different pH conditions. The results indicated that: (1) in the co-presence of SMT or CIP, BC transport through the soil was decreased under acid and neutral conditions (pH 5 and 7) due to the charge shielding effect by sorption of SMT or CIP on BC surface; (2) after SMT or CIP pre-sorption, zeta potentials of sand surface became less negative and BC transport was decreased under acid and neutral con-ditions, and this effect was stronger after increasing SMT or CIP pre- sorption; (3) because of the cation and zwitterion forms of CIP under acid and neutral conditions, it resulted in less negative surface of BC and

quartz sand than SMT did, and thus more BC was retained in the column whether CIP was co-present with BC or pre-sorbed on the quartz sand. Our results demonstrated the importance of antibiotics pollution was pH-dependent in controlling the fate and transport of BC through soil after the biochar was used for remediation of soils with antibiotics pollution.

While this study focused on the transport of biochar colloids in the presence of a large amount of antibiotics, our results may have profound implications on the transport of soil colloids (e.g., clays and Fe/Al ox-ides), which are ubiquitous in soils. The effect of antibiotics on colloid transport can affect the related risks in colloids-facilitated contaminants transport in the subsurface environments. In addition, the soil envi-ronments under the field condition are usually more complex, and many factors or conditions (e.g. water content, preferential flow, and natural organic matters) were not considered in this work. Thus, further efforts should be focused on the long-term observation of the fate of colloids and antibiotics in field sites polluted by antibiotics.

CRediT authorship contribution statement

Kang Zhao: Data curation, Formal analysis, Investigation, Method-ology, Writing - original draft. Ling Gao: Data curation, Investigation, Software. Qianru Zhang: Writing-revision, Funding acquisition. Jia-nying Shang: Conceptualization, Writing - review and editing, Super-vision, Funding acquisition.

Fig. 7. Breakthrough curves (a and c) and retention profiles (b and d) of biochar colloids (BC) after different PVs (0, 3, and 15) of sulfamethazine (SMT) or cip-rofloxacin (CIP) pre-sorption on sand surface at pH 7.

K. Zhao et al.

Journal of Hazardous Materials 417 (2021) 125908

10

Declaration of Competing Interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

Acknowledgements

This work was supported by the National Natural Science Foundation of China (41771255), Agricultural Science and Technology Innovation Programm of Chinese Academy of Agricultural Sciences (CAAS- ZDRW202110), and the Program of “1000-Talents Plan” for young researchers.

Appendix A. Supplementary data

Details of the DLVO interaction, Fourier transform infrared (FTIR) spectra of biochar, biochar colloids transport without antibiotics.

Appendix A. Supporting information

Supplementary data associated with this article can be found in the online version at doi:10.1016/j.jhazmat.2021.125908.

References

Ahmed, M.B., Zhou, J.L., Ngo, H.H., Guo, W.S., Johir, M.A.H., Sornalingam, K., 2017. Single and competitive sorption properties and mechanism of functionalized biochar for removing sulfonamide antibiotics from water. Chem. Eng. J. 311, 348–358.

Ameloot, N., Graber, E.R., Verheijen, F.G.A., De Neve, S., 2013. Interactions between biochar stability and soil organisms: review and research needs. Eur. J. Soil Sci. 64 (4), 379–390.

Awad, Y.M., Kim, S.C., El-Azeem, S.A.M.A., Kim, K.H., Kim, K.R., Kim, K., Jeon, C., Lee, S.S., Ok, Y.S., 2014. Veterinary antibiotics contamination in water, sediment, and soil near a swine manure composting facility. Environ. Earth Sci. 71 (3), 1433–1440.

Bendz, D., Paxeus, N.A., Ginn, T.R., Loge, F.J., 2005. Occurrence and fate of pharmaceutically active compounds in the environment, a case study: Hoje River in Sweden. J. Hazard. Mater. 122 (3), 195–204.

Ben, W.W., Qiang, Z.M., Adams, C., Zhang, H.Q., Chen, L.P., 2008. Simultaneous determination of sulfonamides, tetracyclines and tiamulin in swine wastewater by solid-phase extraction and liquid chromatography-mass spectrometry. J. Chromatogr. A 1202 (2), 173–180.

Bradford, S.A., Simunek, J., Bettahar, M., van Genuchten, M.T., Yates, S.R., 2003. Modeling colloid attachment, straining, and exclusion in saturated porous media. Environ. Sci. Technol. 37 (10), 2242–2250.

Chen, H., Gao, B., Li, H., Ma, L.Q., 2011. Effects of pH and ionic strength on sulfamethoxazole and ciprofloxacin transport in saturated porous media. J. Contam. Hydrol. 126 (1–2), 29–36.

Chen, H., Gao, B., Yang, L.Y., Ma, L.Q., 2015. Montmorillonite enhanced ciprofloxacin transport in saturated porous media with sorbed ciprofloxacin showing antibiotic activity. J. Contam. Hydrol. 173, 1–7.

Chen, C., Shang, J.Y., Zheng, X.L., Zhao, K., Yan, C.R., Sharma, P., Liu, K.S., 2018. Effect of physicochemical factors on transport and retention of graphene oxide in saturated media. Environ. Pollut. 236, 168–176.

Fang, J., Cheng, L.L., Hameed, R., Jin, L., Wang, D.J., Owens, G., Lin, D.H., 2020. Release and stability of water dispersible biochar colloids in aquatic environments: effects of pyrolysis temperature, particle size, and solution chemistry. Environ. Pollut. 260, 114037.

Fang, J., Shen, B., Cheng, L.L., Wang, M.H., Zhang, L.Q., Lin, D.H., 2018. Oxytetracycline increases the mobility of carbon nanotubes in porous media. Sci. Total Environ. 628–629, 1130–1138.

Fang, J., Zhang, K.K., Sun, P.D., Lin, D.H., Shen, B., Luo, Y., 2016. Co-transport of Pb2+

and TiO2 nanoparticles in repacked homogeneous soil columns under saturation condition: effect of ionic strength and fulvic acid. Sci. Total Environ. 571, 471–478.

Frey, S.K., Topp, E., Khan, I.U.H., Ball, B.R., Edwards, M., Gottschall, N., Sunohara, M., Lapen, D.R., 2015. Quantitative Campylobacter spp., antibiotic resistance genes, and veterinary antibiotics in surface and ground water following manure application: influence of tile drainage control. Sci. Total Environ. 532, 138–153.

Gamiz, B., Velarde, P., Spokas, K.A., Celis, R., Cox, L., 2019. Changes in sorption and bioavailability of herbicides in soil amended with fresh and aged biochar. Geoderma 337, 341–349.

Fig. 8. Schematic of BC transport after low and high SMT or CIP pre-sorption on media surface at neutral condition.

K. Zhao et al.

Journal of Hazardous Materials 417 (2021) 125908

11

Gao, L.H., Shi, Y.L., Li, W.H., Niu, H.Y., Liu, J.M., Cai, Y.Q., 2012. Occurrence of antibiotics in eight sewage treatment plants in Beijing, China. Chemosphere 86 (6), 665–671.

Gothwal, R., Shashidhar, T., 2015. Antibiotic pollution in the environment: a review. CLEAN Soil Air Water 43 (4), 479–489.

Gottschall, N., Topp, E., Metcalfe, C., Edwards, M., Payne, M., Kleywegt, S., Russell, P., Lapen, D.R., 2012. Pharmaceutical and personal care products in groundwater, subsurface drainage, soil, and wheat grain, following a high single application of municipal biosolids to a field. Chemosphere 87 (2), 194–203.

Guo, T., Lou, C.L., Zhai, W.W., Tang, X.J., Hashmi, M.Z., Murtaza, R., Li, Y., Liu, X.M., Xu, J.M., 2018. Increased occurrence of heavy metals, antibiotics and resistance genes in surface soil after long-term application of manure. Sci. Total Environ. 635, 995–1003.

Hanna, N., Sun, P., Sun, Q., Li, X.W., Yang, X.W., Ji, X., Zou, H.Y., Ottoson, J., Nilsson, L. E., Berglund, B., Dyar, O.J., Tamhankar, A.J., Lundborg, C.S., 2018. Presence of antibiotic residues in various environmental compartments of Shandong province in eastern China: its potential for resistance development and ecological and human risk. Environ. Int. 114, 131–142.

Heymann, K., Lehmann, J., Solomon, D., Liang, B.Q., Neves, E., Wirick, S., 2014. Can functional group composition of alkaline isolates from black carbon-rich soils be identified on a sub-100 nm scale? Geoderma 235–236, 163–169.

Hu, X.G., Zhou, Q.X., Luo, Y., 2010. Occurrence and source analysis of typical veterinary antibiotics in manure, soil, vegetables and groundwater from organic vegetable bases, northern China. Environ. Pollut. 158 (9), 2992–2998.

Jia, M.Y., Wang, F., Bian, Y.R., Stedtfeld, R.D., Liu, G.X., Yu, J.P., Jiang, X., 2018. Sorption of sulfamethazine to biochars as affected by dissolved organic matters of different origin. Bioresour. Technol. 248, 36–43.

Ji, X.L., Shen, Q.H., Liu, F., Ma, J., Xu, G., Wang, Y.L., Wu, M.H., 2012. Antibiotic resistance gene abundances associated with antibiotics and heavy metals in animal manures and agricultural soils adjacent to feedlots in Shanghai; China. J. Hazard. Mater. 235–236, 178–185.

Jjemba, P.K., 2006. Excretion and ecotoxicity of pharmaceutical and personal care products in the environment. Ecotoxicol. Environ. Saf. 63, 113–130.

Kinney, C.A., Furlong, E.T., Zaugg, S.D., Burkhardt, M.R., Werner, S.L., Cahill, J.D., Jorgensen, G.R., 2006. Survey of organic wastewater contaminants in biosolids destined for land application. Environ. Sci. Technol. 40 (23), 7207–7215.

Kolodynska, D., Krukowska, J., Thomas, P., 2017. Comparison of sorption and desorption studies of heavy metal ions from biochar and commercial active carbon. Chem. Eng. J. 307, 353–363.

Lehmann, J., 2007. A handful of carbon. Nature 447, 143–144. Lertpaitoonpan, W., Ong, S.K., Moorman, T.B., 2009. Effect of organic carbon and pH on

soil sorption of sulfamethazine. Chemosphere 76 (4), 558–564. Li, W.H., Shi, Y., Gao, L.H., Liu, L.M., Cai, Y.Q., 2013. Occurrence, distribution and

potential affecting factors of antibiotics in sewage sludge of wastewater treatment plants in China. Sci. Total Environ. 445–446, 306–313.

Lyu, H.H., Gao, B., He, F., Zimmerman, A.R., Ding, C., Huang, H., Tang, J.C., 2018. Effects of ball milling on the physicochemical and sorptive properties of biochar: experimental observations and governing mechanisms. Environ. Pollut. 233, 54–63.

Major, J., Lehmann, J., Rondon, M., Goodale, C., 2010. Fate of soil-applied black carbon: downward migration, leaching and soil respiration. Glob. Change Biol. 16 (4), 1366–1379.

Michael, I., Rizzo, L., Mcardell, C.S., Manaia, C.M., Merlin, C., Schwartz, T., Dagot, C., Fatta-Kassinos, D., 2013. Urban wastewater treatment plants as hotspots for the release of antibiotics in the environment: a review. Water Res. 47 (3), 957–995.

Petrinic, I., Buksek, H., Luxbacher, T., Pusic, T., Bischof, S., 2018. Influence of the structure of polymer fiber composites on the analysis of the zeta potential. J. Appl. Polym. Sci. 46227, 1–8.

Qi, N., Wang, P.F., Wang, C., Ao, Y.H., 2018. Effect of a typical antibiotic (tetracycline) on the aggregation of TiO2 nanoparticles in an aquatic environment. J. Hazard. Mater. 341, 187–197.

Qu, X.L., Fu, H.Y., Mao, J.D., Ran, Y., Zhang, D.N., Zhu, D.Q., 2016. Chemical and structural properties of dissolved black carbon released from biochars. Carbon 96, 759–767.

Rajapaksha, A.U., Vithanage, M., Ahmad, M., Seo, D.C., Cho, J.S., Lee, S.E., Lee, S.S., Ok, Y.S., 2015. Enhanced sulfamethazine removal by steam-activated invasive plant- derived biochar. J. Hazard. Mater. 290, 43–50.

Ryan, J.N., Elimelech, M., 1996. Colloid mobilization and transport in groundwater. Colloids Surf. A Physicochem. Eng. Asp. 107, 1–56.

Sarmah, A.K., Meyer, M.T., Boxall, A.B.A., 2006. A global perspective on the use, sales, exposure pathways, occurrence, fate and effects of veterinary antibiotics (vas) in the environment. Chemosphere 65 (5), 725–759.

Sohi, S.P., 2012. Carbon storage with benefits. Science 338 (6110), 1034–1035. Stoob, K., Singer, H.P., Mueller, S.R., Schwarzenbach, R.P., Stamm, C.H., 2007.

Dissipation and transport of veterinary sulfonamide antibiotics after manure application to grassland in a small catchment. Environ. Sci. Technol. 41 (21), 7349–7355.

Teixido, M., Pignatello, J.J., Beltran, J.L., Granados, M., Peccia, J., 2011. Speciation of the ionizable antibiotic sulfamethazine on black carbon (Biochar). Environ. Sci. Technol. 45 (23), 10020–10027.

Thiele-Bruhn, S., 2003. Pharmaceutical antibiotic compounds in soils - a review. J. Plant Nutr. Soil Sci. 166 (2), 145–167.

Vasudevan, D., Bruland, G.L., Torrance, B.S., Upchurch, V.G., MacKay, A.A., 2009. pH- dependent ciprofloxacin sorption to soils: interaction mechanisms and soil factors influencing sorption. Geoderma 151 (3–4), 68–76.

Vithanage, M., Rajapaksha, U.A., Tang, X.Y., Thiele-Bruhn, S., Kim, K.H., Lee, S.E., Ok, Y. S., 2014. Sorption and transport of sulfamethazine in agricultural soils amended with invasive-plant-derived biochar. J. Environ. Manag. 141, 95–103.

Wan, J.M., Tokunaga, T.K., 2002. Partitioning of clay colloids at air-water interfaces. J. Colloid Interface Sci. 247 (1), 54–61.

Wang, J.L., Chu, L.B., Wojnarovits, L., Takacs, E., 2020. Occurrence and fate of antibiotics, antibiotic resistant genes (ARGs) and antibiotic resistant bacteria (ARB) in municipal wastewater treatment plant: an overview. Sci. Total Environ. 744, 140997.

Wang, Y., Zhang, W., Shang, J.Y., Shen, C.Y., Joseph, S.D., 2019. Chemical aging changed aggregation kinetics and transport of biochar colloids. Environ. Sci. Technol. 53 (14), 8136–8146.

Xing, Y.N., Chen, X., Wagner, R.E., Zhuang, J., Chen, X.J., 2020. Coupled effect of colloids and surface chemical heterogeneity on the transport of antibiotics in porous media. Sci. Total Environ. 713, 136644.

Xing, Y.N., Chen, X.J., Chen, X., Zhuang, J., 2016. Colloid-mediated transport of pharmaceutical and personal care products through porous media. Sci. Rep. 6, 35407.

Xing, Y.N., Chen, X.J., Zhuang, J., Chen, X., 2015. What happens when pharmaceuticals meet colloids. Ecotoxicology 24 (10), 2100–2114.

Xu, N., Cheng, X.Y., Wang, D.J., Xu, X.T., Huangfu, X.X., Li, Z.L., 2018. Effects of Escherichia coli and phosphate on the transport of titanium dioxide nanoparticles in heterogeneous porous media. Water Res. 146, 264–274.

Yang, W., Feng, T.T., Flury, M., Li, B.G., Shang, J.Y., 2020. Effect of sulfamethazine on surface characteristics of biochar colloids and its implications for transport in porous media. Environ. Pollut. 256, 113482.

Yang, W., Wang, Y., Shang, J.Y., Liu, K.S., Sharma, P., Liu, J., Li, B.G., 2017. Antagonistic effect of humic acid and naphthalene on biochar colloid transport in saturated porous media. Chemosphere 189, 556–564.

Yao, L.L., Wang, Y.X., Tong, L., Deng, Y.M., Li, Y.G., Gan, Y.Q., Guo, W., Dong, C.J., Duan, Y.H., Zhao, K., 2017. Occurrence and risk assessment of antibiotics in surface water and groundwater from different depths of aquifers: a case study at Jianghan Plain, central China. Ecotoxicol. Environ. Saf. 135, 236–242.

Zhang, H.J., Lu, T.T., Zhang, R.Y., Wang, M.J., Krishnan, S., Liu, S.H., Zhou, Y.M., Li, D. L., Qi, Z.C., 2020. Effects of clay colloids on ciprofloxacin transport in saturated quartz sand porous media under different solution chemistry conditions. Ecotoxicol. Environ. Saf. 199, 110754.

Zhi, D., Yang, D.X., Zheng, Y.X., Yang, Y., He, Y.Z., Luo, L., Zhou, Y.Y., 2019. Current progress in the adsorption, transport and biodegradation of antibiotics in soil. J. Environ. Manag. 251 (109598), 1–7.

K. Zhao et al.