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Article

Transport of Cu2+ in Unsaturated Porous Medium with Humic Acid/Iron Oxide Nanoparticle (Fe3O4) Amendment

1
School of Environment, Northeast Normal University, No. 2555, Jingyue Street, Changchun 130117, China
2
Henan Key Laboratory of Groundwater Pollution Prevention and Rehabilitation, Zhengzhou 450001, China
*
Author to whom correspondence should be addressed.
Water 2021, 13(2), 200; https://doi.org/10.3390/w13020200
Submission received: 16 December 2020 / Revised: 7 January 2021 / Accepted: 13 January 2021 / Published: 15 January 2021
(This article belongs to the Section Hydrology)

Abstract

:
Humic acid (HA) and iron oxide (such as Fe3O4) nanoparticles are widely distributed in soil, and their complex embedded in soil might affect the transport and fate of Cu2+ in the vadose zone, while Cu2+ is a serious threat to the underlying groundwater. In this study, we synthesized a composite of Fe3O4 nanoparticles coated with HA (HA@NPs) using as an amendment a packed sand matrix in the transport column experiments. The impacts of HA content and ion strength (IS) on Cu2+ transport in the unsaturated columns were investigated. The results showed that HA exhibited a stronger inhibition effect on Cu2+ transport, and a higher IS enhanced the mobility of Cu2+ in an unsaturated porous medium in the presence of HA@NPs. The recovery ratio (Rr) of Cu2+ breakthrough in the column decreased from 66.56% to 3.94% while the mass concentration ratio CHA/CNPs increased from 0 to 50 in the HA@NPs complex. The Rr increased by 1.64 times while the IS increased from 0 to 100 mM. Batch adsorption experiments, kinetics and isotherm models, and Fourier transform infrared (FTIR) spectra analysis were implemented to elucidate the underlying mechanism. It was found that HA embedded in the sand matrix could bind Cu2+ by forming stable chelate, while the IS-dependent Cu2+ transport could be attributed to the competitive adsorption between Na+ and Cu2+. Our study demonstrates that the physicochemical environment, as well as the presence of iron oxide nanoparticles and natural organic matter, can significantly impact Cu2+ transport in unsaturated porous medium.

Graphical Abstract

1. Introduction

Copper (Cu) pollution poses a dangerous threat to the environment. In particular, China has the most widespread copper soil pollution problem [1,2]. According to the China Soil Pollution Survey in 2014, 2.1% of the contaminated sites in tests had soil polluted by copper [3]. Ionic copper (Cu2+), namely the soluble state of copper, is one of the key soil pollutants that must be removed because of its biotoxicity and mobility [4]. Copper-bearing wastewater from smelting plants, chemical mechanical planarization, and mine drainage could break through the vadose zone and infiltrate into groundwater. This can pose serious risk to the environment and human health, as groundwater supply accounts for 17.5% of the total water supply in China [5]. Understanding the factors affecting transport and fate of Cu2+ in unsaturated porous media is thus essential. In particular, while humic acid (HA) and iron oxide (such as Fe3O4) particles are widely distributed in the soil, there have been few studies that examine how the embedded Fe3O4 particles and HA affect Cu2+ transport in unsaturated porous medium [6,7].
HA is the main organic component in the soil [8], and represents approximately 60% of soil organics [9]. A large number of branched hydrophilic and hydrophobic moieties, as well as functional groups (i.e., carboxy, carbonyl, and hydroxyl), are linked to each other by the backbone of the aliphatic or aromatic units in HA [10]. This complex chemical structure determines the sensitivity and the physicochemical properties of HA to any given environmental factors [11]. HA becomes a repository for many pollutants due to its strong surface adsorption caused by the complexation and chelation between the functional groups of HA and the environmental pollutants [12]. The interactions between HA and heavy metals are mainly chemical complexation, cationic proton exchange, electrostatic adsorption, and flocculation [8,13]. The number of oxygen functional groups on the surface of the HA particles is a crucial factor that determines the nature of its interaction with heavy metals [14,15]. Hydrophilic and hydrophobic organics usually enhanced the mobility of heavy metals in clean sand, since the adsorption of heavy metals on the sand surface was inhibited after they were coated by these organics. The mediated transport depended on the inherent mobility of the organics in sand, and usually followed the order of hydrophilic matters > hydrophobic matters [16]. However, if the dissolved organic matter (DOM) was injected continuously, then ternay complexes (metals–DOM–sand medium) might form and substantially inhibit the mobility of heavy metals [17]. Ionic strength (IS) is recognized to compress the double electric layer and consequently influence the electro kinetic characteristics of macromolecular substance [18]. Thus, the reaction between HA and Cu2+ may be significantly affected by IS. The impacts of IS on the transport of mobile HA and colloid/nanoparticles (NPs) through porous medium have been reported [19]. Because of increased charge screening and charge neutralization effects, the increase of IS can weaken electrostatic repulsion between suspended NPs and between NPs and the medium particles, typically resulting in NPs aggregation and strong deposition at porous medium surfaces [18,19]. However, the role of IS in the transport of Cu2+ in porous medium with embedded HA/ Fe3O4 NPs is not well understood.
Moreover, it is important to understand the transportation and interaction between NPs and heavy metals in unsaturated porous medium because the vadose zone represents both the medium that is polluted with Cu2+ and the natural barrier that prevents Cu2+ from being transported into groundwater [20]. According to survey work done by Corapciogl et al. (1999) [21], there is an average of 108–1017 colloidal particles (biocolloids, inorganic colloids, and organic colloids) per liter in the groundwater. This indicates that colloids/NPs can carry pollutants such as heavy metals into the groundwater [22]. Most studies focus on the transportation of colloids in a saturated porous medium, where the colloids are adsorbed across the solid-water interface (SWI) and deposited in the medium through multi-particle bridging and the physical interception of steric hindrance [23,24]. Bradford et al. showed that the colloid adsorption across an air-water interfaces (AWI) in an unsaturated porous medium is both strong and irreversible because of the negative charge of the gas–liquid interface [25,26]. Processes such as the deposition and transportation of colloids and heavy metals are controlled by the chemical and hydrodynamic conditions of the porous medium, such as ionic strength, pH, moisture content, and flow rate [26,27]. Few studies have examined the heavy metal transport in an unsaturated porous medium embedded with NPs coated with organic material. In particular, Fe3O4 NPs usually appear in soil accompanied by Cu2+ in mining areas [28,29], or magnetic Fe3O4 NPs are often added artificially for soil remediation because they can be easily separated from the soil matrix through an external magnet to avoid secondary pollution [30,31].
The objective of this work was thus to examine the influence of the presence of HA-coated Fe3O4 NPs (HA@NPs) embedded in porous medium and solution IS on Cu2+ transport through the vadose zone. The approach focused on a systematic identification of the mechanism that mediated Cu2+ transport in unsaturated porous medium with HA@NPs amendment. The Cu2+ transport column experiment results were interpreted through monitoring of breakthrough curves (BTCs), a batch adsorption experiment, and surface property measurements of targeted compounds.

2. Materials and Methods

2.1. Porous Media

For the porous medium in our experiments, we collected natural river sand (sifted through 1 mm sieve) from the Yitong River in Changchun, China. Sands with an average diameter of 0.53 mm represent the medium in the vadose zone overlying the local aquifer. The sand was air dried and washed with deionized (DI) water five times to remove impurities. The sand was then immersed in 4% HCl for 24 h, rinsed with deionized water, soaked in a 1 M NaOH solution for another 24 h, and finally washed repeatedly with deionized water until the filtrate had a pH of 5.5 [32,33]. The pH 5.5 used herein aimed to avoid precipitation of Cu. The sand was then dried at 105 °C overnight.

2.2. Preparation of Humic Acid (HA)-Coated Iron Oxide (Fe3O4) Nanoparticles (NPs)

The Fe3O4 NPs were purchased from the Shanghai Macklin Biochemical Technology Co., Ltd., China with a diameter of 20 nm (purity ≥ 99.5%). We added 3.4 g of the NPs to 1.0 L of ultra-pure water, mixed the solution in a sonicator bath for 6 h, and finally deflocculated the solution by heating it at 70 °C with sufficient stirring until the slurry changed to be a stable suspension. In order to avoid the particle aggregation, the NPs were ultrasonicated for 3 h (Fuyang F-060S, 450 W, 40 KHz) again prior to being coated with HA, the average size of NPs was 584 nm measured by ZetaSizer Nano ZS (Malvern Instruments, Worcestershire, UK) (Figure S1 in Supplementary Material). The stock suspension was stored at 4 °C.
We dissolved 1.7 g of HA (acquired from the Tianjin Guangfu Fine Chemical Research Institute, Tianjin, China) in 100 mL of 1 M NaOH under constant stirring at 25°C until the HA had dissolved into the solution. We then introduced ultrapure water to create a final solution with mass concentration ratio CHA/CNPs values of 10, 20, 30, and 50. Every mixture of HA and Fe3O4 NPs was stirred for 50 h; we obtained HA@NPs by injecting that mixture with 0.01 M HCl and 0.01 M NaOH until the pH reached a value of 5.5. The suspension of HA@NPs was then mixed with the cleaned sand at 25 °C until the HA@NPs fully coated the surface of the sand particles.

2.3. Characterization of HA, Fe3O4 NPs, and HA-Coated Fe3O4 NPs

Fourier transform infrared (FTIR) spectra of sand, HA, Fe3O4 NPs, and HA@NPs were recorded in the range of 600–4000 cm−1 using a 330 FTIR microscope (Thermo Fisher Scientific, Waltham, MA USA) [33]. The zeta potential (ZP) and hydrodynamic diameter of the aforementioned four mediums was measured using ZetaSizer Nano ZS (Malvern Instruments, Worcestershire, UK) [34]. We measured the magnetic properties of the Fe3O4 NPs and HA@NPs using a vibrating sample magnetometer (Magnetic Property Measurement System XL-7, Quantum Design, San Diego, CA USA) following the process described by Ghosh et al. (2011) [35].

2.4. Column Transport Experiments

Transport experiments were conducted by pumping Cu2+ solution (50 mg L−1) through an acrylic column packed with a sand matrix (Figure 1). The acrylic column was 8 cm in diameter and 50 cm in height; porous circular plates were installed at the top and bottom of the column to prevent particles from escaping. The column was packed with 3.4 kg of granular media with a density of 1.36 g/cm3. The porosity of the packed medium was 0.378 on average, and the hydraulic dispersion coefficient was estimated to be 0.178 cm2 min−1. Two peristaltic pumps at the inlet and outlet of the column were used to regulate the moisture content and to control the downward flow rate. The target moisture content (80% water saturation) was obtained according to the method developed by Esfandyari Bayat et al. 2015 [36]. The tensiometers (Meter T5x, München, Germany) at both ends read the same values, indicating uniform flow conditions.
There are two groups of column experiment in this study. (1) Group one was performed to examine the impact of the presence of HA coated Fe3O4 NPs (HA@NPs) embedded in porous medium on Cu2+ transport. Six types of packed medium in the columns were sand only (sand), sand containing Fe3O4 NPs (NPs + sand), 1HA@NPs + sand, 2HA@NPs + sand, 3HA@NPs + sand, and 5HA@NPs + sand, wherein the coefficients of HA, I = 1, 2, 3, and 5, mean i g HA embedded in 1 kg sand. The content of Fe3O4 NPs in the sand was 0.1 g kg−1. Accordingly, the CHA/CNPs was raised from 0 to 10, 20, 30, and 50, respectively. The solution IS was 5 mM and pH = 5.5. (2) In group two, three types of packed medium were used: sand, NPs + sand, and 1HA@NPs + sand. The Cu2+ breakthrough curves under IS 0, 5, 10, 50, 100 mM were investigated in each medium to indicate the role of IS. The solution pH was 5.5.
For each column experiment, 10 pore volumes (PVs) of the background electrolyte solution of interest were first passed through the column to ensure that the column was fully equilibrated with this solution. This process also aimed to remove all possible re-mobile particles in the fixed-bed matrix in the column, and the spatial distribution of embedded particles in the column was determined immediately after this elution process. The mechanism for the stability of the packed medium was elucidated by Derjaguin–Landau–Verwey–Overbeek (DLVO) interaction energy calculations [33].
Then, suspension of Cu2+ within the same background electrolyte composition was injected into the packed column for 16.5 PVs (phase 1), followed by elution with Cu2+-free solution again (phase 2). The outflow from the columns was connected by an auto fraction sampler (Huxi CBS-A 100, Shanghai, China), and the effluent Cu2+ concentration was monitored using an inductively coupled plasma optical emission spectrometer (ICP-OES, Avio 200, PerkinElmer, Waltham, MA, USA). The effluent mass recovery rate (Rr) was calculated by integrating the BTCs.

2.5. Batch Adsorption–Desorption Experiments

Both batch adsorption and desorption experiments were carried out to explain the Cu2+ transport behaviors in different packed mediums and under different IS. In the adsorption experiments, three types of medium, sand, NPs + sand, and 1HA@NPs + sand were used as adsorbents. We measured the Cu2+ sorption kinetics by mixing 20 g of targeted adsorbent with 100 mL of Cu2+ solution (50 mg/L) with a pH of 5.5 and IS of 0, 5, 10, 50, and 100 mM in Teflon centrifuge tubes. The time required for equilibrium was set as 5 h following a previous kinetic study [37], and the differences of the concentration values of Cu2+ were actually less than 1% after 3 h indicating the feasibility of 5 h. The adsorption dynamics of Cu2+ in each adsorption system was obtained by measuring Cu2+ concentration at time interval of 5–60 min. The tubes at specific sampling time were centrifuged for 30 min at 12,000 rpm, and the supernatant was filtered by a 0.45-μm membrane, followed by Cu2+ measurement by ICP-OES. The Cu2+ desorption kinetics were then measured by mixing the Cu2+-rich porous medium with 100 mL of a 0.01 M NaCl solution and shaking the solution at room temperature (23–25 °C) for 5 h. The Cu2+ concentrations of the solutions were determined at the same time interval as the aforementioned adsorption experiment. The mechanisms underlying the adsorption process were further investigated using the pseudo-1st order and pseudo-2nd order rate equations to fit the adsorption experimental data as described by Chen et al. [38]. The adsorption equilibrium data were fitted using the Langmuir and Freundlich isotherms to provide quantitative information for adsorption isotherms, respectively [39].

3. Results and Discussion

3.1. Surface Properties of HA, Fe3O4 NPs, HA@NPs, and Sand

FTIR spectra were used to identify the surface functional groups for the HA, Fe3O4 NPs, 1HA@NPs, and sand (Figure 2a). For sand, the main peaks around 998 and 773 cm−1 are attributed to symmetric Si-O-Si vibrations [40]. The FTIR spectrum of HA reveals a broad band at 3255 cm−1, attributed to O–H and trace N–H stretching [41]. The peak at 1565 cm−1 and 1378 cm−1 are identified as N–H bending vibration and C–N stretching vibrations while peaks around wavelengths 1007 cm−1 and 912 cm−1 correspond to C–H stretching [42]. The band located at 1039 cm−1 in the FTIR spectrum of Fe3O4 NPs indicates the symmetric stretching vibration of the Fe–O stretching. While the NPs are coated by HA, the main peak is assigned to C–H stretching. Both the surface hydroxyls (Si–OH) on sand and (3623 cm−1) abundant –OH on HA can provide adsorption sites for Cu2+.
Figure 2b,c shows the zeta potentials and hydrodynamic diameters of different samples when IS varied from 0 to 100 mM. The electrokinetic measurements showed that the HA, 1HA@NPs and sand were negatively (approximately –25 mV) charged over the range of solution conditions investigated, and the Fe3O4 NPs is positively charged (approximately 25 mV). This indicates considerable electrostatic attractive forces between Cu2+ and these compounds. The particles size increased obviously with IS, while the zeta potentials became less negative. Increasing the IS results in compression of diffuse double layers around charged surfaces, thus reducing the repulsive forces between the particles and decreasing their stability to form aggregates [32]. By contrast, the HA coating increased the stability of Fe3O4 NPs (Figure 2c). As shown in Figure 2d, the hysteresis loops of Fe3O4 NPs and 1HA@NPs are standard S-shaped curves without remanence, indicating that the coated HA did not destroy the intrinsic superparamagnetism. Adding the HA coating to the NPs decreased the saturation magnetization value because HA has diamagnetic resistance, and a magnetically dead layer was formed with oxygen and other elements on the surface of the NPs [35]. Some studies demonstrated that the magnetic properties of NPs were attenuated through interaction with the HA surface groups [43,44,45], our magnetic coercivity values for NPs and 1HA@NPs are similar to those of previously published values [46]. Thus, HA reduced the magnetic force between the HA@NPs particles and improved their stability. Moreover, the decreased magnetization still could be observed after rinsing of the HA@NPs, it thus also proved the stability of the HA–NP complex.

3.2. Stability of HA-Fe3O4 NP-Sand Fixed-Bed in the Column

One important premise of the transport experiments is the stability of the fixed-bed matrix in the column. The breakthrough curves of Cu2+ should not be disrupted by the possible presence of mobile particles. Herein, we discuss the possible stabilization mechanisms for the HA-Fe3O4 NP-sand system. Calculated DLVO interaction energies (Φ) between sand and the NPs or 1HA@NPs are plotted as functions of separation distance in Figure 3, and the data are depicted in terms of kT, where k is the Boltzmann constant and T is Kelvin temperature. For NPs, as shown in Section 3.1, the positively charged surface made this scenario as a “favorable condition” for attachment on negative sand surface. NPs were readily deposited into the primary energy minimum (Φprimary-min) (Figure 3a). For 1HA@NPs, while the presence of the high Φmax would prevent the deposition of particles into the primary energy minimum (Figure 3b), they could be immobilized within the secondary energy minimum (Φsecond-min) [33]. Indeed, the calculated interaction energy profiles showed the presence of the Φsecond-min and the depth of the Φsecond-min increased with IS, which indicated that it became more likely for HA@NPs to be retained with sand in the column under high IS. The release of NPs and 1HA@NPs from the column during sequential elution experiment and retention profiles of the two particles after 10 PVs elution finished at IS = 0 mM are shown in Figure 3c,d. The release of immobilized particles could be negligible after 8 PVs elution (near 0 mg L−1), and the distributions of particles along the column were uniform.

3.3. Effect of the HA-Coated Fe3O4 NPs Embedded in Sand Matrix on Cu2+ Transport

As shown in Figure 4, the normalized effluent Cu2+ concentration C/C0 was plotted as a function of PV. The deposition of Cu2+ increased (C/C0 decreased) with increasing HA content in all packed media (Figure 4a and Table 1). The maximum breakthrough concentrations, (C/C0)max, were decreased from 0.857 to 0.810, 0.647, 0.560, and 0.170 finally when CHA/CNPs was raised from 0 (NPs + sand) to 10 (1HA@NPs + sand), 20 (2HA@NPs + sand), 30 (3HA@NPs + sand), and 50 (5HA@NPs + sand), respectively. The corresponding Rr values were 66.56%, 47.40%, 33.72%, 19.21%, and 3.94%, respectively. As a control, the (C/C0)max and Rr in pure sand were 0.859and 66.96%. These results indicate that HA impedes the transportation of Cu2+.
As shown in Equations (1) and (2), Cu2+ was first adsorbed on the surface of HA@NPs and sand by electrostatic attraction. Then, as shown in Equation (3), Cu2+ complexed with the functional groups (–COOH, –C=O) on the surface of HA@NPs and sand, which was controlled by proton exchange between Cu2+ and H+ [34,47]. This HA@NPs that occurred in the potential layer of HA, is irreversible for Cu2+. Additionally, some Cu2+ attached on HA@NPs to produce the stable chelate, and thus Cu2+ was immobilized by the HA@NPs.
Sur(H+/Na+)m + Cu2+Sur(H+/Na+)m-r…Cu2+ + r(H+/Na+),
Sur-O + Cu2+Sur-O…Cu2+,
Sur-(OnHm)-n+m + Cu2+Sur-(OnHm-Cu)-n+m-r+2 + rH+,
where “Sur” is the particle surface, “...” is an electrostatic action, and n, m, and r are constants. The Cu2+ ratio is stable in Equations (1)–(3) [48].
HA@NPs either adhered to the sand surface or accumulated in the porous medium AWI. According to Akbour et al. (2013) [10], the adsorption of HA on the surface of positively charged sand is primarily dependent on electrostatic interaction, and secondarily dependent on the interaction between the non-polar sites of the sand and the hydrophobic groups of the HA. Even if HA carried the same charge as the sand, HA could also be adsorbed onto the sand surface, and the amount of adsorption would surpass the amount of desorption because the HA function groups undergo ligand exchange with the sand [12,49]. The HA@NPs adhering to the sand might be caused by the similarity in the surface chemistries as HA. For the SWI and the AWI in the unsaturated porous medium, some studies indicate that the increased hydrodynamic diameter of HA@NPs could result in physical interception such as multi-particle bridging, steric hindrance, or surface deposition [23]. When the hydrodynamic diameters of the particles exceeded the AWI thickness, HA@NPs were more easily captured by the nearby AWI [26]. Additionally, the oscillating ring produced by water absorption between the particles hindered the migration of HA@NPs [50].

3.4. Effect of Ion Strength (IS) on Cu2+ Transport in Diverse Unsaturated Porous Media

Figure 4b–d shows the transport of Cu2+ in three types of porous medium with different IS values. The (C/C0)max values of Cu2+ in each BTC and Rr are shown in Table 1. Taking the medium HA@NPs + sand as an example, when the value of IS increased from 0 to 5, 10, 50, and 100 mM, the (C/C0)max values of Cu2+ increased from 0.756 to 0.810, 0.830, 0.889, and 0.910, respectively, while the corresponding Rr = 44.83%, 47.40%, 50.70%, 65.12%, and 73.53%, respectively. This IS-dependent Cu2+ transport can be attributed to the competitive adsorption between Na+ and Cu2+. As the IS of the solution increased, a large amount of Na+ ions occupied the negatively charged adsorption sites on the surface of the sand and HA [51]. Those Na+ ions neutralized the negative charges on the surface of the sand and HA, which reduced both the electrostatic force between the Cu2+ and the negative ions adsorbed at the electrical adsorption sites and the total number of Cu2+ ions on the surface of sand and HA. Based on the limit equation and the Bader–Shock theory, IS is negatively correlated with the activity coefficient of ions; the higher IS value corresponds to a lower initial activity of Cu2+ in the solution [52]. As a result, more Cu2+ were transported through the column.

3.5. Adsorption Characteristics of Cu2+ on Different Types of Media

Figure 5 depicts the time-dependent Cu2+ adsorption and desorption capacity of three media across varying IS (0, 5, 10, 50, and 100 mM), pure sand, NPs + sand, and 1HA@NPs + sand. In all adsorption media, the Cu2+ concentration in the liquid phase exhibited a rapid decrease within the initial 1–1.5 h which contributed to > 90% of the total adsorption, followed by a gradual decrease. Compared with the adsorption processes, the desorption of Cu2+ was relatively weak (<6 mg/L after 300 min). The aqueous Cu2+ concentration in the desorption system was an order of magnitude lower than that in the adsorption system. This is in accord with the observation in transport experiments. The C/C0 of Cu2+ quickly reached 0 after elution by 4–5 PVs Cu2+-free solution (phase 2), suggesting that the release of pre-deposited Cu2+ could be negligible under the same IS.
Table 2 and Figure 6a–c present the data fitting results of the pseudo-2nd-order kinetic model, which delineates that the entire adsorption process is generally controlled by chemisorption [39]. The results demonstrate the superior performance of the 2nd-order model with correlation coefficients R2 > 0.98 and modeled qe values closer to the measured results. The R2 values of the pseudo-1st-order model were all less than 0.89 (data not shown here). This indicates the greater suitability of the pseudo-2nd-order model in describing the adsorption mechanism and consequently illustrates the predominance of chemical interactions thereof. The fitting results of the Cu2+ adsorption isotherm data indicated that the Langmuir model exhibited a stronger correlation (R2 = 0.987–0.997) than the Freundlich model (Table 3 and Figure 6d). This suggests that the binding of Cu2+ on the porous medium in testing is a monolayer sorption.
The influences of IS and HA coating on Cu2+ adsorption thus supports the results of Cu2+ transport in the columns. The increasing HA concentration was linked with an enhancement in Cu2+ adsorption on the media, which was confirmed by the increase in k2 and KL (Table 3). Electrolytes inhibited Cu2+ adsorption and accelerated Cu2+ desorption on the surface of mediums. The scenario with higher IS displayed lower k2, which thus led to more Cu2+ breakthrough in the column. This phenomenon could still be explained by either the competition for adsorption sites [53] or the competition for complexation sites on the outer-sphere surface between Na+ and Cu2+ [54,55]. An X-ray energy dispersive spectrometer mapping graph was provided in Figure S2 to confirm the Cu adsorption.

4. Conclusions

In this study, we conducted laboratory experiments in order to investigate the transport of Cu2+ in unsaturated columns with humic acid/iron oxide NP (Fe3O4) amendment under different IS conditions. We found that the presence of HA stabilized Fe3O4 NPs in a sand matrix and was effective in restraining the transport of Cu2+. The mobility of Cu2+ in the packed sand columns was enhanced at higher IS values. Complexation, electrostatic attractive force, hydration, oxygen-containing functional groups, surface attachment, and AWI attachment are all mechanisms or features that may be responsible for the retention and transport of Cu2+ in the unsaturated porous media embedded with HA@NPs. These conclusions were supported by the adsorption tests and modeling analyses. The presence of HA and NPs in the medium, and the IS change altered the fate, transport and diffusion risk of Cu2+ contamination during unsaturated seepage. Note that only monovalent electrolyte solution (NaCl) was used for IS modulation in the current study. We suggested that studies of the influences of cation composition in divalent solutions (Ca2+ or Mg2+) or in mixed systems (Na-Ca-Mg) are necessary, because Na+, Ca2+ and Mg2+ generally coexist in natural groundwater and their relative composition varies with hydro-geochemical process. In addition to Fe3O4 NPs, many natural organic/inorganic colloids and engineered nanomaterials are present in the soil environment, which also tend to associate with dissolved organic materials and heavy metals. Our findings thus provide a reference for delineating the ternary interactions and the consequences of these interactions, which will be an improved comprehension of the environmental risk of heavy metals in complex systems containing organic materials and colloids/NPs.

Supplementary Materials

The following are available online at https://www.mdpi.com/2073-4441/13/2/200/s1, Figure S1: Size distribution of Fe3O4 nanoparticles, Figure S2: X-ray energy dispersive spectrometer mapping to confirm Cu adsorption.

Author Contributions

Conceptualization, S.L. and G.Z.; Data curation, M.S.; Formal analysis, M.S.; Funding acquisition, S.L. and W.F.; Investigation, M.S., Q.W. and J.Y.; Methodology, M.S.; Project administration, S.L. and W.F.; Resources, S.L.; Software, X.L.; Supervision, W.F.; Validation, Q.W. and J.Y.; Visualization, W.F.; Writing—original draft, S.L. and W.F.; Writing—review and editing, S.L., X.L. and W.F. All authors have read and agreed to the published version of the manuscript.

Funding

This work was supported by the National Natural Science Foundation of China [NSFC No. 41772236 and 51978135]; and the Science and Technology Department of Changchun City [17SS027].

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

The data presented in this study is contained within the article and supplementary materials.

Conflicts of Interest

The authors declare no conflict of interest.

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Figure 1. Schematic of the experimental set-up.
Figure 1. Schematic of the experimental set-up.
Water 13 00200 g001
Figure 2. (a) Fourier transform infrared (FTIR) spectra of humic acid (HA), iron oxide (Fe3O4) nanoparticles (NPs), 1HA@NPs, and sand. (b) Influence of IS on the zeta potential of HA, Fe3O4 NPs, 1HA@NPs, and sand. (c) Influence of ion strength (IS) on the hydrodynamic diameter of Fe3O4 NPs and 1HA@NPs. (d) The hysteresis loops of HA, Fe3O4 NPs and 1HA@NPs.
Figure 2. (a) Fourier transform infrared (FTIR) spectra of humic acid (HA), iron oxide (Fe3O4) nanoparticles (NPs), 1HA@NPs, and sand. (b) Influence of IS on the zeta potential of HA, Fe3O4 NPs, 1HA@NPs, and sand. (c) Influence of ion strength (IS) on the hydrodynamic diameter of Fe3O4 NPs and 1HA@NPs. (d) The hysteresis loops of HA, Fe3O4 NPs and 1HA@NPs.
Water 13 00200 g002
Figure 3. Calculated DLVO interaction energy plotted as a function of separation distance between the sand surface and NPs (a) or 1HA@NPs (b) surface at IS = 0 mM and 100 mM. Release of NPs and 1HA@NPs from the column during sequential elution experiment (c) and retention profiles of the two particles after 10 pore volumes (PVs) elution finished (d) (IS = 0 mM).
Figure 3. Calculated DLVO interaction energy plotted as a function of separation distance between the sand surface and NPs (a) or 1HA@NPs (b) surface at IS = 0 mM and 100 mM. Release of NPs and 1HA@NPs from the column during sequential elution experiment (c) and retention profiles of the two particles after 10 pore volumes (PVs) elution finished (d) (IS = 0 mM).
Water 13 00200 g003
Figure 4. (a) Breakthrough curves of Cu2+ in columns packed with different medium, sand, NPs + sand, 1HA@NPs + sand, 2HA@NPs + sand, 3HA@NPs + sand, and 5HA@NPs + sand; the IS was 5 mM and pH = 5.5; (b), (c) Breakthrough curves of Cu2+ in columns packed with sand (b), NPs + sand (c), and 1HA@NPs + sand (d). The IS varied from 0 to 100 mM, while pH = 5.5.
Figure 4. (a) Breakthrough curves of Cu2+ in columns packed with different medium, sand, NPs + sand, 1HA@NPs + sand, 2HA@NPs + sand, 3HA@NPs + sand, and 5HA@NPs + sand; the IS was 5 mM and pH = 5.5; (b), (c) Breakthrough curves of Cu2+ in columns packed with sand (b), NPs + sand (c), and 1HA@NPs + sand (d). The IS varied from 0 to 100 mM, while pH = 5.5.
Water 13 00200 g004
Figure 5. The adsorption and desorption of Cu2+ on (a) pure sand (b) NPs + sand (c) 1HA @NPs + sand at different IS values; (d) Adsorption isotherms of Cu2+ on the aforementioned mediums, and the dotted lines were the Langmuir isotherm fitting lines.
Figure 5. The adsorption and desorption of Cu2+ on (a) pure sand (b) NPs + sand (c) 1HA @NPs + sand at different IS values; (d) Adsorption isotherms of Cu2+ on the aforementioned mediums, and the dotted lines were the Langmuir isotherm fitting lines.
Water 13 00200 g005
Figure 6. Pseudo-2nd order kinetic plot for Cu2+ adsorption processes on (a) pure sand, (b) NPs + sand, and (c) 1HA @NPs + sand at different IS values; (d) fitting results of Langmuir models for Cu2+ adsorption.
Figure 6. Pseudo-2nd order kinetic plot for Cu2+ adsorption processes on (a) pure sand, (b) NPs + sand, and (c) 1HA @NPs + sand at different IS values; (d) fitting results of Langmuir models for Cu2+ adsorption.
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Table 1. Quantifying the breakthrough curves of Cu2+ under different experimental conditions.
Table 1. Quantifying the breakthrough curves of Cu2+ under different experimental conditions.
MediumIS (mM)pH(C/C0)maxRr
Sand55.50.85966.96%
NPs + sand55.50.85766.56%
HA@NPs + sand55.50.81047.40%
2HA@NPs + sand55.50.64733.72%
3HA@NPs + sand55.50.56019.21%
5HA@NPs + sand55.50.1703.94%
Sand05.50.81162.98%
55.50.85966.96%
105.50.87866.71%
505.50.93278.85%
1005.50.96787.39%
NPs + sand05.50.83265.71%
55.50.85766.56%
105.50.86370.21%
505.50.91879.75%
1005.50.92685.92%
HA@NPs + sand05.50.75644.83%
55.50.81047.40%
105.50.83050.70%
505.50.88965.12%
1005.50.91073.53%
Table 2. Fitted parameters of the adsorption kinetics models for the adsorption of Cu2+ in different media, with the IS of 0–100 mM.
Table 2. Fitted parameters of the adsorption kinetics models for the adsorption of Cu2+ in different media, with the IS of 0–100 mM.
Porous MediumParametersIS (mM)
051050100
Sandqe,exp (mg kg−1)0.06980.05380.04580.03080.0243
qe,cal (mg kg−1)0.07060.05580.04640.03120.0255
k2 (min−1)0.00180.00170.00160.00150.0011
R20.9960.9950.9910.9840.979
NPs + sandqe,exp (mg kg−1)0.06730.06430.06250.03230.0215
qe,cal (mg kg−1)0.06910.06660.06430.03940.0230
k2 (min−1)0.00220.00210.00210.00200.0018
R20.9970.9980.9960.9860.955
1HA@NPs + sandqe,exp (mg kg−1)0.13800.11000.10300.09900.0895
qe,cal (mg kg−1)0.13900.11100.10000.10500.0895
k2 (min−1)0.00280.00220.00210.00210.0018
R20.9990.9990.9990.9990.999
Table 3. Fitted parameters of the isotherm models for the adsorption of Cu2+ in different media, with an IS of 5 mM.
Table 3. Fitted parameters of the isotherm models for the adsorption of Cu2+ in different media, with an IS of 5 mM.
ModelsParametersPorous Medium
SandNPs + Sand1HA@NPs + Sand
Langmuirqm (mg kg−1)79.36590.909130.208
KL (L mg−1)0.0750.1210.133
R20.9870.9970.997
Freundlich1/n0.4900.3640.337
KF (L mg−1)10.13220.69631.414
R20.9240.9220.985
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Lin, S.; Shi, M.; Wang, Q.; Yang, J.; Zhang, G.; Liu, X.; Fan, W. Transport of Cu2+ in Unsaturated Porous Medium with Humic Acid/Iron Oxide Nanoparticle (Fe3O4) Amendment. Water 2021, 13, 200. https://doi.org/10.3390/w13020200

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Lin S, Shi M, Wang Q, Yang J, Zhang G, Liu X, Fan W. Transport of Cu2+ in Unsaturated Porous Medium with Humic Acid/Iron Oxide Nanoparticle (Fe3O4) Amendment. Water. 2021; 13(2):200. https://doi.org/10.3390/w13020200

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Lin, Shanshan, Mengdi Shi, Qi Wang, Junlin Yang, Gubin Zhang, Xiangru Liu, and Wei Fan. 2021. "Transport of Cu2+ in Unsaturated Porous Medium with Humic Acid/Iron Oxide Nanoparticle (Fe3O4) Amendment" Water 13, no. 2: 200. https://doi.org/10.3390/w13020200

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