Humic Acid Promotes the Adsorption of Lead onto PSMPs: Site Energy Distribution Theory and Fluorescence Quenching Analysis

Xiaotian Lu Sun Yat-sen University School of Chemistry Feng Zeng Sun Yat-Sen University School of Chemistry Shuyin Wei Sun Yat-Sen University School of Chemistry Rui Gao Sun Yat-sen University School of Chemistry Abliz Abdurahman Sun Yat-sen University School of Chemistry Hao Wang Sun Yat-sen University School of Chemistry Weiqian Liang (  liangwq3@mail2.sysu.edu.cn ) Sun Yat-Sen University School of Chemistry https://orcid.org/0000-0002-2329-3761


Introduction
Microplastics (MPs) are de ned as plastic fragments or particles with size less than 5mm by the National Oceanic and Atmospheric Administration (Arthur et al. 2009). MPs consist of plastic microbeads that directly discharged into the environment (primary sources), and plastic fragments derived from the degrading of large plastics (secondary sources) due to weathering processes (e.g., UV photodegradation, mechanical abrasion, biodegradation, etc.) (Alimi et al. 2018). MPs always persist in the environment for a long time as the degradation processes are extremely slow, especially in water, where degradation can take place for decades (Desforges et al. 2014, Hidalgo-Ruz et al. 2012). In addition, MPs are easily driven by wind and water ow to acquire the ability of long-distance diffusion and migration to reach different regions, so they are widely distributed in aquatic environment such as the surface runoff, rivers and lakes as well as Marine (Tang et al. 2019, Tarnminga et al. 2016. MPs can accumulate in organisms through food chains, and cause a lot of adverse effects on aquatic and terrestrial organisms, including inhibited growth and development, endocrine disruption, immunity and neurotransmission dysfunction and so on (Lei et al. 2018, Lu et al. 2016, Rodriguez-Seijo et al. 2016. Furthermore, MPs can serve as vectors to accumulate and transport heavy metal pollutant due to their characteristics of small volume and large speci c surface area, causing the bioaccumulation of contaminations and toxicants in aquatic environments (Brennecke et al. 2016, Hodson et al. 2017. The process and mechanism of the adsorption of heavy metals onto MPs has been studied in recent years (Gao et al. 2019, Wang et al. 2020a. For example, Gao et  Lead (Pb(II)) is one of the most representative heavy metal pollutants owing to its persistence in the environment, bioaccumulation and thus toxicity (Jozef et al. 2020). As a highly toxic metal, Pb(II) is seriously harmful to plant, animal and human healthy. Especially in contaminated water environments, Pb(II) can accumulate in various tissues of aquatic animals and plants that exposed to Pb(II) contamination, thereby in uences morphological, physiological and biochemical processes (Bellinger and Needleman 2003, Guo. 2002, Lee et al. 2019, Li et al. 2019a. Several studies have investigated the adsorption behavior of Pb(II) on MPs, as well as the effect factors of adsorption such as pH and ionic strength (Ahechti et al. 2020, Fu et al. 2020, Gao et al. 2019, Zou et al. 2020. Zou et al. (Zou et al. 2020) studied the sorption of three model heavy metals (i.e., Cu 2+ , Cd 2+ , and Pb 2+ ) on CPE, PVC, LPE and HPE, and found that pH can signi cantly affect the sorption of metals on MPs, but ionic strength exerted a relatively slight effect on this process. Ahechti et al. (Ahechti et al. 2020) evaluated the adsorption capacity of PE and PP for different metals (Cu, Cd, Pb and Zn) depending on the physicochemical conditions of the aquatic environments (exposure time, pH, salinity).
Their results showed that the adsorption percentages increased as pH increase and salinity decrease.
Dissolved organic matter (DOM)in aquatic environment is also an important factor affecting the adsorption process. Godoy et al. investigated the adsorption of Cd, Co, Cr, Cu, Ni, Pb and Zn by ve different types of microplastics was performed in Milli-Q water and natural waters (seawater, urban wastewater and irrigation water), and found that an enhancement of metal adsorption in waters with high dissolved organic matter content as urban wastewater and irrigation water (Godoy et al. 2019). Humic Acid (HA) is a kind of representative dissolved organic matter (DOM) and widely exists in the water environment. HA contains a large number of oxygen-containing functional groups such as carboxyl and hydroxyl groups, which will interact with Pb(II) ). Our previous study has shown that HA can be adsorbed onto PSMPs in the aquatic environment through hydrophobic interaction and π-π electron donor acceptor interaction (Abdurahman et al. 2020). When MPs, Pb(II) and HA coexist in solution, HA has different interactions with MPs and Pb(II), indicating its potential possibility to affect the adsorption of Pb(II) on microplastics. However, there is still a lack of associated mechanisms understanding on the effects of HA on the binding properties of Pb(II) to MPs.
In this study, polystyrene microplastics (PSMPs) were used as the model microplastics  to study the interaction of Pb 2+ with PSMPs and the effects of HA on the adsorption process. Adsorption kinetic and isotherm were conducted at different condition to researched the adsorption characteristic of Pb 2+ adsorption. The site energy distribution theory and uorescence quenching analysis of HA were used to explore the effect of HA on adsorption potentials of PSMPs to Pb 2+ . The results of this study helped to further understand the adsorption characteristics and mechanism of Pb 2+ onto MPs, as well as providing more information for the evaluation of environmental behavior and toxicological effects of microplastics in aquatic environments.
All chemicals were of A.R. grade.

Sample preparation
The PSMPs were prepared into 50 mg/L suspension with ultrapure water (18.2 MΩ). HA solution, obtained by dissolving HA sodium salt in 0.10 mol/L NaOH solution and stirred overnight at 27.0 ℃, was adjusted to pH 7 and then ltered through 0.45-µm cellulose acetate lter paper (Millipore, Billerica, MA, USA). The ltrate was dialyzed with a dialysis membrane (500D) and nally stored at ~4.0 ℃ in the dark. The relevant characterization of HA is shown in the Supporting Information. Pb 2+ stock solution (500 mg/L) was obtained by dissolving a known quantity of Pb(NO 3 ) 2 in distilled water. All the solution pH was adjusted using 0.10 mol/L HNO 3 or 0.10 mol/L NaOH and measured by an Orion pH/ISE meter (Model 710 A, Thermo Fisher Scienti c). The ionic strength was adjusted by adding NaNO 3 or Ca(NO 3 ) 2 solution, respectively.

Adsorption experiments
Batch adsorption experiments were employed as described previously with minor modi cations (Abdurahman et al. 2020). The adsorption kinetic experiments of Pb 2+ uptake on PSMPs were carried out by adding PSMPs suspension, Pb 2+ stock solution and HA solution into 300 mL conical ask.
The initial Pb 2+ concentration was 5.00 mg/L and the HA concentrations were 0.00, 1.00, 2.50, 5.00 mg·C/L, respectively. The suspensions were equilibrated on a reciprocating shaker (Shanghai Tensuc Ltd., China) at 27.0 ℃ (room temperature) in the dark and sampled at different time within 0-4.0 h. For the adsorption isotherm experiment, the initial Pb 2+ concentration was 0.50-15.0 mg/L, and the equilibrium time set at 4.0 h based on preliminary kinetic experiments resulted. The experiment pH value was selected for 3.0 and 6.0, and the ionic strength was set with 0.01, 1.00 and 10.0 mmol/L for NaNO 3 , and 0.03, 0.33 and 3.33mmol/L for Ca(NO 3 ) 2 . At the end of the equilibration time, the suspensions were collected and ltered with a 0.45-µm lter membrane, which part of the ltrate is used for Pb 2+ concentration determination with AAS (Z-2000, Hitachi, Japan), and the other part is used for HA uorescence detection (RF-5301PC, Shimadzu, Japan). The adsorption amounts of Pb 2+ adsorbed on PSMPs were calculated from the differences between the initial and nal Pb 2+ concentrations in solutions; mass losses for control samples were negligible (< 1%).
The adsorption process of Pb 2+ on HA was described in Supplementary data.
The experiments for each condition were performed in triplicate and took the average.

Analytical method
The zeta potential of PSMPs at different background ionic conditions (Na + or Ca 2+ ) were analyzed by Zeta potential analyzer (BI-PALS, Brookhaven, American) at the range of pH 2.0-10.0. The elemental analysis (C H N O S) of HA was characterized using Elemental analyzer (Vario EL cube, Elementar, Germany). The functional groups (carboxyl and hydroxyl) of HA were determined using Ma's work (Ma et al. 2001). The concentration of HA was measured by TOC analyzer (TOC-L CPH, Shimadzu, Japan). Fluorescence excitation (Ex)-emission (Em) matrix (EEM) spectra of HA was measured using a uorescence spectrometer (RF-5301PC, Shimadzu, Japan).
Details of the data analysis for Pb 2+ adsorption onto PSMPs are given in the Supplementary data.

Results And Discussion
3.1. Adsorption kinetics of Pb 2+ onto PSMPs in the presence and absence of HA The adsorption kinetic curves of Pb 2+ onto PSMPs at different initial HA concentrations with varying solution conditions (i.e., pH and ionic strength) were illustrated in Fig. S1 and S2. The adsorption of Pb 2+ onto PSMPs under different conditions showed similar kinetic behaviors, i.e., the uptake of Pb 2+ is rapid in the time range of 0~60 min, then the adsorption decreases gradually and nally reach equilibrium. At the initial adsorption stage, Pb 2+ occupied the abundant adsorbing sites of PSMPs. With prolonged contact time, most of the adsorption sites were occupied, and the adsorption rate of Pb 2+ became slower as well as the adsorption reaches equilibrium gradually . Therefore, the adsorption equilibrium time was set to 240 min, which was su cient for Pb 2+ adsorption.
The pseudo-rst-order and pseudo-second-order kinetic models were used to t the experimental data. Kinetic parameters were listed in Table S1 and S2. The R 2 values for the pseudo-second-order kinetic model were higher than that for the pseudo-rst-order model, indicating that the pseudo-second-order kinetic model was better tted the experimental data than the pseudo-rst-order kinetic model in different experiment conditions, and the chemical sorption may be rate-limiting step of Pb 2+ adsorption mechanism (Tang et al. 2019). The Q e and initial adsorption rate V 0 increased with the increase of pH and decrease of ionic strength, showing that the higher pH value and lower ionic strength would promote the adsorption of Pb 2+ onto PSMPs. Similar trends were previously reported for the adsorption of Pb 2+ or other heavy metal on kinds of adsorbent materials (Fu et al. 2020, Zheng et al. 2018, Zou et al. 2020. At the identical pH and ionic strength condition, the presence of HA signi cantly increased the adsorption amount of Pb 2+ on PSMPs. For example, at the condition of pH 6.0 and 0.10 mmol/L NaNO 3 , the Q e of Pb 2+ increased from 0.302 to 0.888, 1.41 and 1.76 mg/g at the concentration of 1.00, 2.50 and 5.00 mg·C/L of HA, respectively. HA adsorbed on PSMPs would introduce more oxygen-containing functional groups such as hydroxyl and carboxyl groups, which were available for the formation of strong complexes with Pb 2+ , so that more Pb 2+ was adsorbed onto PSMPs (Yang et al. 2011). 3.2. Adsorption isotherms of Pb 2+ onto PSMPs in the presence and absence of HA The isothermal adsorption curves reveal the equilibrium state of adsorbate in solution and adsorbent, and further illustrate the adsorption mechanism(Shen et al. 2021). Fig. S3 and S4 showed the adsorption isotherm process of Pb 2+ onto PSMPs in the absence and presence of HA. It could be seen that the equilibrium adsorption capacity of PSMPs for Pb 2+ rapidly rose at rst due to the relatively strong driving force at high initial concentrations (Qiao et al. 2016). Then the upward trend slew down with the initial Pb 2+ concentration increased. This was because a large number of adsorption sites are already occupied by Pb 2+ .
In order to forecast adsorption behavior, the Langmuir and Freundlich isotherm models were used to t the adsorption curve, respectively, and the tting parameters were listed at Table S3 and S4. Both two isotherm models could t the adsorption process well, and the R 2 values for the Langmuir model were higher than Freundlich model. Therefore, the Langmuir model could be better employed for characterizing equilibrium adsorption of Pb 2+ on PSMPs with/without addition of HA. In addition, the tting results implied that the chemisorption and monolayer adsorption played a signi cant role in the Pb 2+ removal (Jinren et al. 2014).
K L is related to the adsorption a nity between adsorbate and adsorbent. In Table S3 and S4, the signi cantly increased of K L indicated that the presence of HA enhanced the adsorption a nity and capacity of PSMPs for Pb 2+ ). The change of K L also indicated that the adsorption of Pb 2+ onto PSMPs was in uenced by the solution conditions such as pH and ionic strength (Tang et al. 2019). As for Freundlich model, the n values of all experiment conditions were higher than 1, suggesting that the adsorption of Pb 2+ onto PSMPs was greatly nonlinear. The distribution of adsorption sites on the PSMPs is heterogeneous, and electrostatic interactions was major adsorption mechanisms (Xu et al. 2018, Sun et al. 2010). These results indicated that the adsorption mechanism of heavy metals onto PSMPs was mainly includes both physical adsorption and chemical bonding between anion and cation (Luo et al. 2022 The interaction between PSMPs and HA was researched in our reported works (Abdurahman et al. 2020). The results showed that the adsorption of HA on PSMPs conformed to pseudo-second-order kinetic model and Freundlich model, indicating that HA adsorbed on PSMPs surface through hydrophobic and π-π interaction.
The adsorption process was low pH-dependent, and the adsorption capacity of PSMPs increased with the increase of ionic strength. Previous studies showed that HAs is an important natural ligands in regulating the speciation, bioavailability, and ultimate fate of trace metal element in the environment (Kováik et al. 2018, Prado et al. 2006. This was because of the numerous functional groups (-COOH and -OH) in HA molecules, which could combine with heavy metal ions such as Pb 2+ to form stable compounds by means of complexation, ion exchange and electrostatic interaction (Bradl 2004, Yang et al. 2015. The elemental analysis and determination of carboxyl and hydroxyl groups of HA were determined and the results were summarized at Table S5, which was similar to others works . Then the adsorption experiments of Pb 2+ on HA were conducted and the adsorption kinetic and isotherm curves were shown in Fig. S5 and S6, as well as the tting parameters were shown in Table S6 and S7. The adsorption of Pb 2+ on HA was tted to pseudo-second-order kinetic model and Langmuir model. According to the different kinetic adsorption processes of Pb 2+ on PSMPs and HA (Fig. S1, S2 and S5), the adsorption equilibrium of Pb 2+ on HA was faster than that onto PSMPs. When PSMPs, HA and Pb 2+ coexisted in the solution, Pb 2+ was more prone to combine with HA to achieve adsorption equilibrium than PSMPs. Then HA-Pb 2+ complex adsorbed onto PSMPs through the interaction between HA and PSMPs, which led to an indirect adsorption of Pb 2+ onto PSMPs. Besides, the free Pb 2+ without binding to HA also adsorbed directly onto PSMPs due to electrostatic interaction, until the concentration of Pb 2+ in the two phases reached equilibrium.
Another possibility was that part of HA molecules rst adsorbed on the surface of PSMPs, which introduced more functional groups (-COOH and -OH), and functioned as anchors and aid adsorption in the adsorption of

Effect of pH on Pb 2+ adsorption onto PSMPs in the presence and absence of HA
The solution pH can affect the surface charge of adsorbents, the structure of HA and the ionic species of metals, thereby in uence the underlying mechanisms involved in the interaction among different substances in the adsorption process (Yang et al. 2014). Therefore, the adsorption of Pb 2+ on PSMPs with or without HA at different pH value were studied. As shown in Fig. 1, the Q m of Pb 2+ adsorption on PSMPs increased with the increasing solution pH value whether HA existed or not. At the condition of 0.10 mmol/L NaNO 3 and 5.00 mg·C/L HA, the equilibrium adsorption amount of Pb 2+ increased from 1.94 to 2.13 mg/g when pH value was from 3.0 to 6.0, indicating that a higher pH was bene cial to the adsorption of Pb 2+ on PSMPs.
The effect of pH on adsorption is related to the surface charge of PSMPs. Fig. S7 showed that the zeta potential of PSMPs gradually decreased with the increase of pH in the range of pH 2.0-10.0. The pH pzc of PSMPs were 1.06, 1.51 and 2.18 at the conditions of 0.10, 1.00 and 10.0 mmol/L NaNO 3 , as well as 1.47, 1.72 and 2.19 at the conditions of 0.03, 0.33 and 3.33 mmol/L Ca(NO 3 ) 2 , respectively (Table S8). That was to say, the pH pzc of PSMPs increased with the increase of ionic strength.
In the binary system without HA, PSMPs was negatively charged and it was easy to attract positively charged Pb 2+ through electrostatic interaction at the experimental pH conditions (pH 3.0 and pH 6.0) . The negative charges of PSMPs increased as pH increased, which led to the enhancement of electrostatic interaction between PSMPs and Pb 2+ correspondingly, and enhanced the adsorption capacity of PSMPs (Tang et al. 2019). The decrease in pH also causes competitive adsorption. There was a large amount of hydronium ion H 3 O + in the solution under low pH conditions, which competed with Pb 2+ for the adsorption sites on the surface of PSMPs, and inhibited the adsorption of Pb 2+ (Bardestani et al. 2019, Wang 2011. Similar trends were reported for the adsorption of metal ions on other microplastics, as well as some kind of nanomaterials (Jinren et al. 2014, Zhao et al. 2011. When HA was present in solution, the pH value in uenced the adsorption capacity of Pb 2+ on PSMPs and the binding characteristics of HA and Pb 2+ . The adsorption of HA onto PSMPs was little affected by pH, but the adsorption of Pb 2+ on HA increased with the increase of pH value. The molecular structure of HA was affected by the pH value of solution, i.e., with the decreasing of pH, the stretched linear HA structure gradually curled and became a compacted form. Therefore, the exposed functional groups of HA were reduced and the binding with Pb 2+ was weakened under low pH condition (Gezici et al. 2007). In ternary systems, the presence of HA further increased the effect of high pH to promote the adsorption of Pb 2+ onto PSMPs. In general, whether HA is present or not, the increase of pH is favorable to the adsorption of Pb 2+ on PSMPs.

Effect of ionic strength on Pb 2+ adsorption onto PSMPs in the presence and absence of HA
The ionic strength of aquatic environmental can affect the adsorption behaviors of Pb 2+ onto PSMPs, whether HA exist or not. Fig. 1 showed that the adsorption capacities of Pb 2+ onto PSMPs decreased with the increase of ionic strength at different HA concentration. For instence, the Q m of Pb 2+ decreased from 2.13 to 1.57, and from 1.97 to 1.42 mg/g as the concentration of Na + and Ca 2+ increased under the condition of pH 6.0 and 5.00 mg·C/L HA, showing that the presence of background ions is not conducive to the adsorption of Pb 2+ . The results were similar to the previous studies (Fu et al. 2020, Holmes et al. 2014, Zou et al. 2020. Figure S7 showed that the pH pzc of PSMPs increased with the increase of ionic strength, because the positively charged background ions (Na + or Ca 2+ ) would shiel or neutralize the negative surface charges of PSMPs (Wijesena et al. 2020). What's more, the effect of Ca 2+ on the pH pzc of PSMPs is greater than that of Na + under the same ionic strength condition. This is because the charge shielding or neutralization effect of divalent positive ions is stronger than that of mono-valent positive ions. When the ionic strength increased, the surface negative charge of PSMPs would reduce, which resulted in the weakening of electrostatic interaction between PSMPs and Pb 2+ . In addition, according to the theory of DLVO, increasing the ionic strength of the solution will compress the electric double layer and reduce the electrostatic repulsion, resulting in an increase in the aggregation of PSMPs and a decrease in the effective adsorption sites (Alimi et al. 2018. A similar phenomenon has been found in the adsorption of metals by carbon-based nanomaterials such as graphene oxide and carbon oxide nanotubes (Yang et al. 2011). Besides, background electrolyte ions (Na + and Ca 2+ ) can also compete with Pb 2+ for speci c available adsorption sites on PSMPs, which may be one of the reasons for the weakened adsorption of Pb 2+ (Wang et al. 2016, Zhao et al. 2011 3.4. Effects mechanisms of HA on Pb 2+ adsorption onto PSMPs

Site energy distribution analysis
The SEDT (site energy distribution theory) provides relevant information on the energy distribution of the adsorbent surface site, the average site energy and the energy distribution heterogeneity, which can further explain the mechanism of adsorption behavior (Carter et al. 1995). According to the SEDT, the energy distribution of adsorption sites on the surface of adsorbents is heterogeneous, and sites with higher adsorption energy are more likely to be occupied by adsorbents. Fig S8 and S9 showed the effect of HA on the site energy E*, and Fig. 2 and 3 showed the site energy distribution function F(E*) of Pb 2+ adsorption on PSMPs under different condition based on Langmuir model. It can be seen from the Fig S8 and S9 that the E* values gradually decreased with the increase of equilibrium adsorption amount of Pb 2+ on PSMPs, indicating that the surface energy distribution of PSMPs is heterogeneous, and the amount of high energy sites is limited. In the adsorption process of ternary systems, the high-energy adsorption sites on PSMPs were rstly occupied by Pb 2+ or HA-Pb 2+ , then the low-energy adsorption sites (Shi et al. 2013). This is consistent with the adsorption of Cr(VI) on engineered silicate nanoparticles (Liao et al. 2020). In addition, with the increase of HA concentration, the E* values increased, illustrated that the presence of HA enhanced the adsorption site energy on the PSMPs surface, thus promoting the adsorption of Pb 2+ .
From Fig. 2 and 3, the F(E*) curves of Pb 2+ adsorption on PSMPs were all unimodal and quasi-Gaussian (Huang et al. 2018). Relevant parameters such as Em*, F(Em*), µ(E*) and σ e * were shown in the Table 1 and 2. It was reported that the higher the value of the µ(E*), the higher the adsorption a nity (Carter et al. 1995). With the increasing pH value and decreasing ionic strength, the parameters of Em*, F(Em*) and µ(E*) were slightly increased, which was consistent with the adsorption kinetics and isothermal results. The results again re ected the effect of pH and ionic strength on the adsorption of Pb 2+ to PSMPs. When the HA concentration increased, the parameters increased signi cantly, indicating that the presence of HA enhanced the adsorption a nity of PSMPs for Pb 2+ . Evidenced by the site energy distribution (Fig. 2 and 3), the site energy heterogeneity was revealed for adsorption of Pb 2+ on PSMPs and could be characterized by the standard deviation σ e * of the distribution (Yan et al. 2017). The σ e * value was 10.2 under different conditions, indicating that the surface of PSMPs was with high adsorption heterogeneous, and the surface heterogeneity of PSMPs were close, that is, the addition of HA did not change the surface structure of PSMPs.   Figure S10 showed the 3DEEM (three-dimensional uorescence excitation-emission matrix) spectra of HA, indicating that the texted HA molecules had two main uorescence peaks, i.e., peak A (Ex/Em: 275/481 nm) and peak B (Ex/Em: 455/516 nm). Peak A with high intensity was the uorescence peak of terrestrial humiclike, while peak B had a lower intensity, which might be related to microbial metabolism (Coble. 1996). Peak A with higher intensity was selected for uorescence quenching analysis.
In binary HA-Pb 2+ system and ternary PSMPs-HA-Pb 2+ system, the uorescence quenching curves of HA at different conditions were shown in Fig. S11 to S18. The decrease in uorescence intensity with increasing Pb 2+ concentrations revealed the formation of chemical complexes between HA and Pb 2+ . It was also observed that with the increase of Pb 2+ concentration, the maximum emission wavelength of HA shifted toward the lower wavelength (i.e. blue shift), suggesting the possibility reducing of conjugated bonds in the chain structure, or the occurrence of π-π* transition in the reaction process (Zhang et al. 2017). The initial uorescence intensity of HA at pH 3.0 was lower than that at pH 6.0, illustrating that HA has more uorophore at pH6, which would help to adsorption more Pb 2+ (Gezici et al. 2007).
The linear Stern-Volmer equation was applied to reveal the binding behaviors of HA with Pb 2+ , as shown in Fig. 4 and 5, and the model parameter K SV was shown in Table 3. In the ternary system (PSMPs-HA-Pb 2+ ), the uorescence quenching of HA increased with the increase of pH and the decrease of the ionic strength (both Na + and Ca 2+ ), which was consistent with the trend in the binary system (HA-Pb 2+ ), re ecting in that the addition of PSMPs did not change the binding mode of HA and Pb 2+ . Compared to binary system (HA-Pb 2+ ), the uorescence quenching of HA in ternary system (PSMPS-HA-Pb 2+ ) is stronger due to the higher K sv value at the same HA concentration (5.00 mg·C/L). In combination with the increase of adsorption of Pb 2+ on PSMPs in the presence of HA mentioned above, the results again indicated that part of Pb 2+ is indirectly adsorbed on PSMPs with the form of HA-Pb 2+ compound (Fu et al. 2020). In the ternary system, the K SV value decreased with the decrease of HA concentration. The initial uorescence intensity of HA was weak at low concentration, and the change of uorescence quenching was relatively minor with the increase of Pb 2+ concentration. When the concentration of Pb 2+ was the maximum, the nal uorescence intensity of HA under pH 3.0 was always higher than that under pH 6.0, indicating that the binding of Pb 2+ to HA is reduced at low pH due to less exposure of uorescent groups of HA. There for, the possibility of indirect adsorption of Pb 2+ on PSMPs through HA was lower under pH 3.0 condition than that under pH 6.0.

Conclusion
In this article, the adsorption of Pb 2+ onto PSMPs was studied using batch experiments under different pH and ionic strength condition, and the effect of HA on the adsorption process was discussed. Under experimental conditions, the adsorption kinetics and isothermal equations of Pb 2+ on PSMPs conform to the pseudo-second-order kinetics model and Langmuir model, respectively. The increase of pH value and decrease of ionic strength was bene cial to the adsorption of Pb 2+ on PSMPs regardless of HA presence.
The presence of HA enhanced the adsorption capacities of PSMPs for Pb 2+ . Pb 2+ can adsorb onto PSMPs directly through electrostatic interactions, and can also adsorb indirectly through HA.
The site energy distribution of Pb 2+ adsorption onto PSMPs under experimental conditions showed that Pb 2+ /HA-Pb 2+ rstly occupied the high energy adsorption sites of PSMPs and then diffused to the low-energy adsorption sites. The addition of HA increases the site energy and distribution frequency of Pb 2+ on PSMPs.
The heterogeneity of PSMPs at different pH and ionic strength were similar, indicating that the change of solution conditions did not affect the structure of PSMPs. The uorescence quenching of HA in ternary systems (PSMPs-HA-Pb 2+ ) was stronger than that in binary systems (HA-Pb 2+ ), which proved that HA encourages the adsorption of Pb 2+ on PSMPs. These ndings are expected to promote the understanding of the interaction between Pb 2+ , HA and PSMPs for further evaluating the environmental risks of microplastics. The linear Stern-Volme tting curves of HA uorescence quenching process (pH 3.0) (a. 5.00mg·C/L HA; b.