Microplastics influence the adsorption and desorption characteristics of Cd in an agricultural soil

Journal Pre-proof Microplastics influence the adsorption and desorption characteristics of Cd in an agricultural soil Shuwu Zhang, Bin Han, Yuhuan Sun, Fayuan Wang

PII:

S0304-3894(19)31729-7

DOI:

https://doi.org/10.1016/j.jhazmat.2019.121775

Reference:

HAZMAT 121775

To appear in:

Journal of Hazardous Materials

Received Date:

10 October 2019

Revised Date:

14 November 2019

Accepted Date:

27 November 2019

Please cite this article as: Zhang S, Han B, Sun Y, Wang F, Microplastics influence the adsorption and desorption characteristics of Cd in an agricultural soil, Journal of Hazardous Materials (2019), doi: https://doi.org/10.1016/j.jhazmat.2019.121775

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Microplastics influence the adsorption and desorption characteristics of Cd in an agricultural soil Shuwu Zhanga,b, Bin Hana, Yuhuan Suna, Fayuan Wanga,b* a

College of Environment and Safety Engineering, Qingdao University of Science and Technology,

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Qingdao, Shandong Province, 266042, P.R. China Shandong Key Laboratory of Water Pollution Control and Resource Reuse, School of Environmental

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Science and Engineering, Shandong University, Qingdao, 266237, China

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*Corresponding author: e-mail: [email protected]; [email protected] (Fayuan Wang) Phone: +86 532 84022617

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Graphical abstract

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Highlights 

MPs impacts on Cd adsorption and desorption in a farmland soil were first studied.



Adding MPs decreased soil adsorption capacity for Cd, but increased Cd desorption.



MPs impacts depended on MPs dose and particle size, and solution pH.



Addition of MPs may increase the mobility of Cd in soil, resulting in additional risks.

Abstract

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Microplastics (MPs) in terrestrial ecosystems particularly agroecosystem are attracting increasing

attention worldwide. However, the influences of MPs on adsorption and desorption of contaminants in

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agricultural soils remain unknown. Here, batch experiments were conducted to study the effects of

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polyethylene MPs on Cd adsorption and desorption in a farmland soil under varying conditions. Both Cd adsorption and desorption in soils with or without MPs reached equilibrium within 120 min. Cd

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adsorption kinetics followed the pseudo-second order model, and the adsorption isotherm fitted to the Langmuir model more precisely than the Freundlich model. Overall, addition of MPs decreased Cd

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adsorption but increased desorption, and the effects varied with MPs dose and particle size, and solution pH. MPs-induced decrease in Cd adsorption and increase in Cd desorption were more pronounced at higher MPs dose and larger particle size, but varied differently from solution pH. EDS

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analysis confirmed Cd adsorption on MPs surface. Both MPs before and after Cd adsorption showed similar XRD patterns, indicating MPs maintained a high crystallinity and no new crystalline phases formed. In conclusion, the input of MPs into soil might enhance the mobility of Cd via mitigating soil adsorbing capacity, thereby posing additional risks of Cd to agroecosystem.

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Keyword: Microplastics; Cadmium; Adsorption; Desorption; Soil contamination

1. Introduction Microplastics (MPs), defined as plastic debris with size smaller than 5 mm (Thompson et al., 2004), are considered emerging contaminants posing threats to various ecosystems (Cole et al., 2011; 2

Nizzetto et al., 2016b; de Souza Machado et al., 2018a, 2019). MPs can enter agroecosystems via various pathways, such as applications of biosolids (Nizzetto et al., 2016a), organic fertilizers (Weithmann et al., 2018), and plastic mulch films (Ng et al., 2018). The MPs released annually to farmlands through biosolids in Europe and North America were estimated to reach 63000–430000 and 44000–300000 tons, respectively, exceeding the estimated global burden of MPs in oceans (Nizzetto et al., 2016b). Recent findings have confirmed the presence of diverse MPs in farmland soils (Liu et

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al., 2018; Zhang and Liu, 2018). The abundance of MPs was 78.0  and 62.5 items kg−1 in shallow and deep layers, respectively (Liu et al., 2018). In another study, MPs in cropland soils can reach 7100–

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42960 particles kg-1 (Zhang and Liu, 2018). Due to their smaller size, MPs (especially nanoplastics <

100 nm) can be taken up by biota and produce more serious environmental and health risks than their

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bulk counterparts (Cole et al., 2011; Rillig, 2012; Rillig et al., 2017; de Souza Machado et al., 2018a). Unsurprisingly, for the reason of safe food production, the effects of MPs in agroecosystems require

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special attention.

MPs have different properties from soil particles, and can affect soil properties, further mediating

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soil structure and function (de Souza Machado et al., 2018b, 2019). Due to their light density, MPs generally lead to lower soil bulk density (de Souza Machado et al., 2018b). Polyester microfiber

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reduced soil aggregate stability in the presence of soil biota (Lehmann et al., 2019). MPs can also affect soil aggregation, water holding capacity and water availability, and microbial activity, but the effects vary with MPs type and concentrations (de Souza Machado et al., 2018b, 2019). Putatively, MPs may act on contaminants via changing soil properties (especially sorption capacity). For

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example, adding PE (10%, w/w) to the soil reduced the adsorption of two organic contaminants, i.e. atrazine and 4-(2,4-dichlorophenoxy) butyric acid, thus increasing their mobility in soil (Hüffer et al., 2019). However, MPs impacts on soil properties and contaminant behaviors still require more investigation.

In addition to their direct threats, MPs generally have remarkable binding capacity, and can sorb organic and metallic contaminants on their surfaces, thus serving as vectors for these contaminants (Browne et al., 2013; Brennecke et al., 2016; Koelmans et al., 2016; Hodson et al., 2017; Zhang et al., 3

2018; Wang et al., 2019). The adsorption of contaminants onto MPs and the subsequent desorption behaviors may alter the toxicity of both contaminants and MPs, causing varied risks to biota. A previous study found the Zn adsorbed by HDPE MPs could be more bioavailable than that adsorbed to soil particles, implying their potential risks to soil ecosystems (Hodson et al., 2017). In agricultural fields, if the MPs loaded with contaminants are taken up by crops, they may cause threats to both yields and quality.

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Heavy metals are considered common contaminants in agricultural soils (Cabrera et al., 1998; Nicholson et al., 2003; Wei and Yang, 2010; Yang et al., 2018). However, bioavailability and toxicity

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of heavy metals, as well as their sorption behaviors, is influenced by soil properties such as clay

content and cation change capacity (Naidu et al., 1998; Finzgar et al., 2007; Cesar et al., 2012; Jardine

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et al., 2013). Considering the wide occurrence of heavy metals and MPs, these two groups of contaminants probably interact each other and consequently result in varied environmental behaviors

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in agricultural fields. Cadmium (Cd) is one of the most common metallic contaminants in various environments (Cabrera et al., 1998; Yang et al., 2018). For example, Cd ranks the first among the

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metallic contaminants in China’s farmlands (Zhao et al., 2014). Our previous study has shown HDPE MPs can directly adsorb Cd on their surface from aqueous solutions (Wang et al., 2019). However,

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whether MPs affect the adsorption/desorption behaviors of heavy metals particularly Cd in soils still remains unknown.

Based on the above mentioned context, we hypothesize that the presence of MPs will change the adsorption of Cd in soil and the subsequent desorption of the adsorb Cd. To verify this assumption,

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batch experiments were performed to study the adsorption and desorption characteristics of Cd in a farmland soil with or without MPs, as well as the influences of reaction time, MPs dose and particle size, solution pH, and initial Cd concentration. The kinetics and equilibrium isotherms were fitted to understand the adsorption mechanisms. Our present study may provide evidence for understanding MPs impacts in agroecosystems.

2. Materials and methods 2.1. Soil collection and preparation 4

The soil used was sampled from a typical farmland, located in the southeastern part of the Shandong Peninsula and belongs to the Jimo District of Qingdao, Shandong Province (36°26'53.89′′N, 120°08′55.22′′E). The area is a temperate maritime monsoon climate, with annual average temperature of 12.7 °C, and annual average precipitation of 662.1 mm. The soil type is alfisols (US soil taxonomy). The cultivated crops are mainly vegetables, peanuts and corn. Five 5×5m soil sampling plots were selected at the sampling site, and one soil sample at 0-15 cm depth was taken using a

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shovel at each plot. All the soil samples were mixed sufficiently and air-dried naturally, and sieved through a 200-mesh sieve for analysis of soil properties and batch experiments. The physicochemical

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properties of the soil used are shown in Table 1. Table 1. Physicochemical properties of the test soils

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Cation exchange

Organic

Total Cd

Particle distribution (%)

capacity (cmol/kg)

matter (g/kg)

(mg/kg)

2.0-0.05 mm

0.05-0.002 mm

<0.002 mm

7.53

13.1

0.16

68.6

22.9

8.5

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Sandy loam

pH

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Soil texture

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2.2. MPs and reagents

High-density polyethylene (HDPE) is the most common plastic in agricultural mulch films. Thus, HDPE MPs were selected for the present experiments. The virgin HDPE MPs were purchased from

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Dongming Plastic Material Co. Ltd, China, with a density of 0.940-0.976 g/cm3, the degree of crystallinity about 80-90%, and softening point 125-135 ºC. The characteristics of the MPs were presented in Table S1, and the SEM images of MPs with different particle sizes were shown in Fig. S1. MPs were divided into four groups with different particle sizes, that is, 48-58μm, 100-154μm, 0.6-

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1.0mm, and 1.0-2.0mm, by manually sieving through sieves with different pore sizes. To remove potential heavy metals on the surface, MPs were washed with 0.1 M HCl, followed by tap water, and finally by distilled water. Cd was selected as the target contaminant, because it is one of the most common metallic contaminants in soil (Cabrera et al., 1998; Zhao et al., 2014; Yang et al., 2018). Cd (NO3)2·4H2O (guaranteed reagent grade) was used to prepare stock solution with 100 mg/L Cd in distilled water. Work solutions with desired Cd concentrations were obtained by diluting the stock solution with 5

distilled water. 2.3. Soil incubation Considering the diverse contents of MPs in soils (de Souza Machado et al., 2018a; Ng et al., 2018), a wide range of concentrations were designed. MPs were mixed into soil to obtain a series of target concentrations, that is, 0.01%, 0.1%, 1%, and 10% (w/w). The control treatment without MPs was also included. The mixed samples were placed in glass beakers sealed with gas permeable, water-

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tight exchange membranes, and incubated in an incubator in darkness for one week at 25 ± 1.0 ℃. The beakers were weighed every two days to check the water loss, and deionized water was added

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accordingly to maintain a soil moisture of 16% (w/w). After the incubation, the soil was sampled and air-dried for further batch experiments.

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2.4. Adsorption and desorption experiments

All the batch experiments were performed at room temperature (25 ± 1 °C) on a mechanical

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shaker at 270 rpm using 50 mL centrifuge tubes (Falcon). One gram of soil with or without MPs and 25 mL Cd solution were placed into the tubes and shaken for 120 min (equilibrium time). For each

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experiment, three replicates and a blank treatment without soil were included. The solution after shaking was filtered with 0.45 μm filter paper for analysis of Cd concentration. The MPs left on the

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filter paper were separated, washed with deionized water for three times and then air-dried for further analysis.

Cd adsorption capacity (qe and qt) by soil and Cd removal percentage (A) were calculated from the difference between the total Cd amount added and the amount retained in the solution using the

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following equations:

𝑞e = 𝑞t = 𝐴=

(𝐶0 −𝐶e )×𝑉 𝑊 (𝐶0 −𝐶t )×𝑉 𝑊

(𝐶0 −𝐶e )×𝑉 𝐶0

(1) (2) (3)

where qt and qe (mg/kg) is the amount of Cd adsorbed at time t and at equilibrium, respectively; C0, Ct and Ce (mg/L) is the concentration of Cd in solution before adsorption, at time t, and at equilibrium, respectively; V (mL) is the volume of solution added and W (g) is the weight of soil. 6

Desorption experiments were conducted immediately after the adsorption experiments, according to the method described by Khan et al. (2018). The supernatant solution in the tubes was decanted and the soil retained was used for desorption experiments. Twenty-five mL NaNO3 solution (0.1 M) was added into the tubes and shaken at 270 rpm at 25 ± 1 °C for 2 h. The solution after shaking was passed through 0.45 μm filter paper for determination of Cd concentration. The amount of Cd desorbed (qde, mg/kg) was calculated by the Eq. (4) (Khan et al., 2018). qde =｛CNa × (25 + W2 – W1) – Ce × (W2 – W1)｝× 1000

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(4)

where CNa (mg/L) is the concentrations of Cd in the equilibrium solution after desorption, W1 (g)

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is the weight of tube and soil, and W2 (g) is the weight of the tube, soil sample and the residual solution after adsorption.

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Cd desorption percentage (Ade, %) can be calculated based on the Eq. 5.

2.5. Characterization of MPs and Cd analyses

(5)

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Ade = qde / qe × 100

A scanning electron microscope (SEM) (S-4800, HITACH, Japan) was used to analyze the

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surface characteristics of MPs. Energy dispersive X-ray spectroscopy (EDS, EDAX Inc. Genesis XM) was used to analyze the elements on MPs surface. The crystalline compositions of the MPs were

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determined using X-ray diffractometer (XRD) (XRD-7000, SHIMADZU, Japan). The samples were scanned over the range of 5–90° of 2θ at a rate of 1° min−1. The Cd in solutions was measured using flame atomic absorption spectrometry (FAAS, AA 7000, SHIMADZU, Japan). 2.6. Quality control and data analysis

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To ensure quality and accuracy of the data, reagent blanks and standard solutions were included with each batch experiment. All experiments were performed in triplicates and the data were expressed as mean with standard deviations. The graphs were generated using Origin 2017 and Excel 2010. Statistical analysis was performed using SPSS 22.0.

3. Results 3.1. Effect of reaction time on adsorption and desorption Soils with or without 1% MPs (size 1-2mm) and Cd solution with initial concentration of 60 mg/L 7

were used to determine the impact of reaction time and equilibrium time on adsorption and desorption of Cd. Solutions were sampled at 5, 10, 30, 60, 90, 120, 180, and 240 min, respectively. The adsorption of Cd was very quick initially, and more than 98% of the total Cd was adsorbed within the first 10 min (Fig. 1a). Figure 1 Adsorption (qt) increased slowly as the contact time increased, and there was no significant change

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in the concentration after 120 min, which was taken as the equilibrium time. Therefore, a reaction time of 120 min was set for the following tests to ensure adsorption equilibrium. Similarly, desorption of

(Fig. 1 b). The desorption equilibrium also reached within 120 min.

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Cd (qde) from soil was also relatively rapid, and more than 70% of Cd was desorbed in the first 30 min

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As shown in Fig. 1, Cd showed similar trends in adsorption and desorption processes in both soils. However, compared to the soil without MPs, soil with 1% MPs had a slight but significant lower

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adsorption capacity for Cd, and a higher desorption capacity. These findings preliminarily suggest MPs decrease soil adsorption for Cd, but increase Cd desorption capacity.

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The adsorption kinetics can be described by the widely-used models, i.e. pseudo-first order model (Eq. 6), pseudo-second order model (Eq. 7), and intraparticle diffusion model (Eq. 8):

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ln(𝑞e − 𝑞𝑡 ) = 𝑙𝑛𝑞e − 𝑘1 𝑡

(6)

𝑡⁄𝑞t = 1⁄𝑞e 2 𝑘2 + 𝑡⁄𝑞e

(7)

𝑞t = 𝑘p 𝑡 1⁄2 + 𝑥𝑖

(8)

where 𝑘1 (min−1) and 𝑘2 [g·(mg·min)−1] are the first-order and second-order equilibrium rate

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constants, respectively; t is the contact time (min), 𝑘p (g·mg−1·min−0.5) is the rate constant of the intraparticle diffusion model, and 𝑥𝑖 is a number related to the thickness of the interface. The results of pseudo‐ first order, pseudo‐ second order and intraparticle diffusion model were

summarized in Table 2. The values of R2 from the fitting of the pseudo-second order model are much higher than those obtained from the pseudo-first order model. Moreover, the experimental qe (qe,exp) approaches the calculated qe (qe,cal) obtained from the pseudo-second-order model. Therefore, the adsorption fitted more precisely to pseudo‐ second order model. In addition, intraparticle diffusion 8

model well fitted the data of Cd adsorption (0.830 < R2 < 0.870), indicating the intraparticle diffusion is crucial for Cd adsorption by soil samples.

Table 2. Kinetic parameters of Cd adsorption by soils obtained from the pseudo-first order, pseudo-second order and intraparticle diffusion models Pseudo‐ first order model

Pseudo‐ second order model

Intraparticle diffusion model

Soil

𝑞e,exp

Soil

990.12

35.53

0.0388

0.775

990.10

0.0189

0.869

Soil (1% MPs)

985.53

14.45

0.0299

0.425

988.34

0.0241

0.827

𝑘1

R2

𝑞e,cal

𝑘2

[g·(mg·min)−1]

R2

𝑘p (g·mg−1·min−0.5)

R2

983.92

0.4608

0.870

981.23

0.3288

0.830

𝑥𝑖

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𝑞e,cal

(min−1)

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3.2. Effect of MPs size

Cd adsorption and desorption was investigated using soils containing 1% of MPs with different

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particle sizes (48-58 μm, 100-154 μm, 0.6-1mm, and 1-2 mm) and 60 mg/L Cd solution (Fig. 2). As the particle size increased, qe decreased from 1473.21 to 1437.79 mg/kg, whereas qde increased

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gradually from 175.67 to 325.62 mg/kg; accordingly, removal percentage (A) decreased from 98.2% to

Figure 2 3.3. Effect of MPs dose

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95.8%, while desorption percentage (Ade) increased from 11.9% to 22.7%.

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Five doses (0, 0.01%, 0.1%, 1%, and 10%, w/w) of MPs (size 1-2 mm) and 60 mg/L Cd solution were used to test the effect of MPs dose on Cd adsorption and desorption (Fig. 3a). With the increase in MPs ratio, both qe and A gradually decreased, ranging from 1469.6 mg/kg to 1460.0 mg/kg, and from 97.93% to 93.33%, respectively, and the most significant effects occurred at 10%, following by

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1%. Accordingly, as MPs dose increased from 0 to 10%, qde significantly increased from 179.90 to 241.48 mg/kg, and Ade from 12.33% to 16.54% (Fig. 3b). Figure 3

3.4. Effect of solution pH The effects of solution pH on Cd adsorption and desorption by soils with or without 1% MPs (size 1-2 mm) were investigated using 60 mg/L Cd solutions with a series of pH values ranging from 4 to 10. As the pH increased, qe and A of both soils gradually increased, accompanying decreasing qde 9

and Ade (Fig. 4a). However, compared to the soil without MPs, soil with 1% MPs had lower qe but higher qde. At the pH of 10, qe of soil with 1% MPs reached a maximum of 1493.8 mg/kg, and A was 99.59%. However, as the pH increased from 4 to 10, qde greatly decreased from 301.69 mg/kg to 74.88 mg/kg, and Ade decreased from 19.98% to 5.01% (Fig. 4b). Figure 4 3.5. Effect of initial Cd concentration

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The effects of initial Cd concentration on Cd adsorption and desorption by soils without or with 1% MPs (size 1-2 mm) were investigated with varying solution concentrations (20, 40, 60, 80, and 100

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mg/L) (Fig. 5). With increasing concentration of solution, qe increased, whereas A decreased.

However, both qde and Ade greatly increased. In most cases, the soil with 1% MPs had lower qe and A

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but higher qde and Ade than the soil receiving no MPs. 3.6. Adsorption isotherm

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To further explain the adsorption mechanism of Cd onto soil as influenced by MPs, isotherm

expressed as Eq. 9: Ce 1 Ce = + qe bqm qm

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models were used to fit the Cd adsorption isotherm data. The linearized form of Langmuir isotherm is

(9)

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where qm (mg/kg) is the maximum absorption capacity, and b (L/mg) is the affinity constant. The empirical non-linear equation of Freundlich isotherm can be expressed as Eq. 10: 𝑞e = 𝐾f 𝐶e 1⁄𝑛

(10)

where Kf [(mg/g) (L/g)1/n] and n are the adsorption equilibrium constants.

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The isotherm parameters for both models were shown in Table 3, which confirms that Cd

adsorption in the two soils followed both models well, but fitted to the Langmuir isotherm better than the Freundlich isotherm, based on the comparison of R2 coefficients. The Langmuir model shows that qm of two soils was higher than 2600 mg/kg. However, the soil with 1% MPs showed lower qm than the soil without MPs. Table 3. Langmuir and Freundlich parameters for Cd sorption by soil with or without 1% MPs

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Langmuir

Soil Soil (1% MPs)

Freundlich

b (L/mg)

qm (mg/kg)

R2

Kf (mg/kg)·(L/kg)1/n

1/n

R2

0.7369

2624.67

0.9973

0.4763

0.4530

0.9033

0.6678

2604.17

0.9985

0.4731

0.4496

0.9516

3.7. EDS analysis MPs (1-2 mm, 10%) in soils before and after adsorption of Cd were isolated for analysis of EDS

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patterns. The EDS patterns shows that the element Cd was not observed before adsorption, but

occurred after adsorption (Fig. 6), suggesting that the Cd on MPs surface was originated from Cd

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solution.

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Figure 6 3.8. XRD analysis

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To determine the influences of HDPE crystallinity on the adsorption behavior of Cd, the XRD spectra of MPs (1-2 mm, 10%) before and after adsorption were obtained (Fig. 7). The XRD results

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showed that no significant changes in XRD patterns were detected, indicating a similar crystallinity before and after the adsorption of Cd. Figure 7

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4. Discussion

Here we first evaluated the effects of MPs on soil adsorption for Cd, and found that the tested HDPE MPs decreased soil adsorption capacity but increased desorption of Cd. Adsorption behaviors of heavy metals onto soils and soil constituents have been widely recognized (Bradl, 2004), and a high

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soil adsorption capacity for Cd was observed in our present experiment (Fig. 3, Table 3). Furthermore, MPs can also adsorb heavy metals such as Zn (Hodson et al., 2017) and Cd (Wang et al., 2019), but their adsorption capacity was generally much lower compared to soil. In our previous study, the estimated maximum adsorption capacity of HDPE MPs for Cd was 30.5 mg/kg, which was extremely lower than that of the soil used in the present study (2624.67 mg/kg). Soil contains a variety of inorganic and organic components including minerals and organic matters, some of which possess 11

excellent adsorption capacity for heavy metals (Bradl, 2004). However, the virgin HDPE MPs we used have a highly hydrophobic surface with relatively simple surface properties. This may partly explain why the tested MPs have a low adsorption capacity and addition of them lowers soil adsorption capacity. Particle size is one of key factors influencing soil sorption for metallic contaminants, and soil fractions with smallest sizes such as clay generally possess the largest maximum adsorption capacity

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(Li and Zhou, 2007; Huang et al., 2014; Liu et al., 2017). Similarly, due to higher specific surface area and the availability of more sorption sites, MPs with smaller particle sizes usually possess higher Cd

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sorption capacity (Wang et al., 2019). This is why the MPs with different particle sizes produced different impacts on soil adsorption capacity, and the largest size (1-2 mm) showed the most

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pronounced decreasing effects (Fig. 2). These findings suggest size-dependent effects of MPs on soil adsorption for Cd. In a recent study, MPs with similar size and shape to natural soil particles elicited

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smaller impacts on soil structure and properties (de Souza Machado et al., 2019). Those tiny MPs such as nanoplastics (<100 nm) may act as both adsorbent and adsorbate, and exert more profound impacts

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than bulk plastics, deserving further investigation.

We also found a dose-dependent effect on soil adsorption (Fig. 3a): increases in MPs dose led to

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decreasing trend in Cd adsorption capacity of soil, and the addition of 10% MPs produced a significant reduction. Pearson correlation analysis showed a negative correlation (r = -0.885) between MPs dose and soil adsorption capacity. Due to the relatively low adsorption capacity, addition of MPs into soil resulted in a “dilution effect”, which may partly account for the reduction in soil adsorption.

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However, the reduction percentage in soil adsorption capacity was really smaller compared to MPs dose (10%), which can be ascribed to more adsorption sites occupied by Cd when the quantity of soil particles becomes less. Solution pH generally exerts a remarkable influence on the sorption behaviors of heavy metal

ions onto soil (Bradl, 2004), MPs (Wang et al., 2019), and other adsorbents (Wang et al., 2010; AlQodah et al., 2017; Wang and Zhuang, 2017). Overall, Cd adsorption capacity of both pure soil and soil with MPs increased with increasing solution pH, but the former increased more sharply than the 12

latter (Fig. 4). A higher pH can induce the surfaces of soil and MPs negatively charged, which favors adsorption of Cd ions with positive charges. However, MPs have relatively simple surface structure and properties, whereas soil contains various constituents with variable charges, such as clays, metal oxides, hydroxides, organic colloidal matters, whose charges vary with the pH of soil solution (Bradl, 2004). This may determine why soil with MPs respond less sensitively to pH changes than pure soil. Previous studies have found both Langmuir and Freundlich isotherm can describe sorption

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behaviors of Cd in soils (Bradl, 2004), and by MPs (Holmes et al., 2012, 2014; Wang et al., 2019). Our present study showed similar findings, but the Langmuir isotherm fitted more precisely (Table 2).

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However, the values of qm and Kf in previous studies on Cd adsorption by MPs are quite low (Holmes et al., 2012, 2014; Wang et al., 2019). This can explain the fact that soil added with MPs had lower qm

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and Kf than pure soil.

Similarly to our previous study (Wang et al., 2019), the EDS analysis confirmed Cd occurrence

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on the surface of MPs isolated from soil after adsorption, suggesting that the effects of MPs are not just a simple “dilution effect”, but also direct interaction with Cd. However, both MPs before and after

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Cd adsorption showed similar XRD patterns (Fig. 6), indicating during the adsorption process MPs maintained a high crystallinity and no new crystalline phases formed. High crystallinity of MPs

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generally results in a low adsorption capacity (Liu et al., 2019), which can account for the low adsorption capacity of MPs in our current study. No formation of new crystalline phases suggests physical interaction may dominate the sorption of Cd by MPs, and the adsorbed Cd can be easily exchanged by Na+ when NaNO3 was used as desorbent.

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Another important finding of the present study is that, although most Cd was irreversibly adsorbed by soil, Cd desorption percentages increased with the addition of MPs, and correlated positively with MPs dose and particle size, but negatively with solution pH. In general, the degree of desorption varies with the strength of binding force between the adsorbent and the adsorbate. Due to the weak interaction forces between MPs and the adsorbed Cd onto them, the Cd was more easily desorbed from the MPs than from soil, leading to a high Cd desorption rate (> 90%) (Wang et al., 2019). Thus, it stands to reason that soil containing more MPs possesses higher Cd desorption 13

capacity. Furthermore, larger MPs particles may have weaker binding capacity for Cd, which is consequently more desorbable than that bound to smaller MPs. Finally, at lower pH, more H+ ions in solution favor the exchange of Cd from the MPs. However, when pH increases to 10, Cd adsorption will become irreversible due to the formation of Cd(OH)2 precipitate, which can explain the less influences of MPs at higher solution pH. To conclude, MPs impacts on Cd desorption are dependent on variations in MPs dose, particle size, and solution pH.

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Bioavailability and fate of Cd in soil can be greatly controlled by soil particle adsorption and desorption (Rashti et al., 2014). Previous studies have shown that MPs influence soil properties, such

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as soil bulk density, aggregation stability, water holding capacity and water availability (de Souza Machado et al., 2018b, 2019), which may subsequently change the adsorption and desorption

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behaviors of Cd, as well as Cd bioavailability and toxicity. Our present results confirm that addition of MPs decreased soil Cd adsorption but increased desorption capacity, consequently reducing soil

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retention capacity and increasing the mobility of Cd. Thus, when MPs and Cd release into agricultural field, the Cd may become more exchangeable and toxic to plants and soil biota, and increase the risks

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of metals accumulating by crops and entering groundwater thereby presenting additional environmental and health risks. In fact, MPs influence not only soil properties, but also plant

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performance, including plant biomass and tissue elemental composition (de Souza Machado et al., 2019). Considering the pervasive occurrence of MPs in agoecosystem and their potential impacts in soil-plant system, future work should be focused on MPs influences on environmental behaviors and fate of soil contaminants and even essential nutrients for plants such as macronutrients N, P and K,

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and micronutrients Cu and Zn.

Finally, we used only virgin MPs in the present study. Ageing, particularly photooxidation

ageing, generally leads to the formation of oxygen-containing functional groups (C=O, C–O and OH) on MPs surface (Bandow et al., 2017), consequently altering MPs’ polarity and hydrophobicity. Therefore, aged MPs may have different particle shapes and sizes, and surface properties, which may further alter not only their ability to sorb contaminants, but also their impacts on soil sorption capacity. More efforts should be made on aged and secondary MPs in soil in future research. 14

5. Conclusions Here, we emphasized MPs impacts on adsorption and desorption behaviors of Cd in a farmland soil. Batch experimental results showed that the adsorption and desorption of Cd reached equilibrium within 120 min. The adsorption process followed the pseudo - second order model and fitted to the Langmuir model better than the Freundlich model. Generally, addition of MPs decreased Cd adsorption capacity of the soil, but increased desorption of the adsorbed Cd. MPs impacts on both

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adsorption and desorption depended on MPs dose and particle size, and solution pH. Higher MPs dose and larger particle size produced stronger inhibiting effects on adsorption and promoting effect on

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desorption. However, more Cd desorbed at lower pH especially for soil added with MPs. In

conclusion, addition of MPs reduced soil retention capacity for Cd and increased the mobility of Cd,

and presenting additional environmental and health risks.

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Author contributions

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thereby increasing the risks of toxic metals like Cd accumulating by crops and entering groundwater,

Fayuan Wang and Shuwu Zhang designed the study. Bin Han and Yuhuan Sun performed experiments.

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Bin Han, Shuwu Zhang, and Fayuan Wang wrote the manuscript.

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The authors declare that they have no conflict of interest

Acknowledgements

This work was supported by the Shandong Key Laboratory of Water Pollution Control and Resource

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Reuse (Grant No. 2019KF15), the National Natural Science Foundation of China (41471395), the Key Research and Development Program of Shandong Province (2019GSF109008), and the Doctoral Foundation of QUST (0100229003).

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Fig. 1. Effects of reaction time on adsorption (a) and desorption (b) of Cd in soil with or without 1%

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MPs (1-2 mm).

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Fig. 2. Effects of MPs particle size on Cd adsorption (a) and desorption (b). The different letters below

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the bars mean significant difference in qe or qde among different treatments (P<0.05).

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Fig. 3. Effects of MPs dose on adsorption (a) and desorption (b). The different letters below the bars

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mean significant difference in qe or qde among different treatments (P<0.05).

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Fig. 4. Effects of solution pH on adsorption (a) and desorption (b). The asterisk indicates significant

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difference in qe or qde between soil with or without 1% MPs (P<0.05).

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Fig. 5. Effect of initial Cd concentration on adsorption (a) and desorption (b). The asterisk indicates

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significant difference in qe or qde between soil with or without 1% MPs (P<0.05).

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(a)

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90.9

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Weight (%)

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Fig. 6. EDS analysis of MPs before (a) and after (b) Cd adsorption.

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Fig. 7. The XRD patterns of MPs before and after Cd adsorption.

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