*Article* **Adsorption of** *p***-Arsanilic Acid on Iron (Hydr)oxides and Its Implications for Contamination in Soils**

**Yifan Yang 1,2,3, Shiyong Tao 1,2,3 , Zhichun Dong 1,2,3 , Jing Xu 1,2,3 , Xiang Zhang 1,2,3, \* and Guoyan Pan 1,2,3, \***


**Abstract:** Because of the diversification of industries in developing cities, the phenomenon of the simultaneous contamination of various kinds of pollutants is becoming common, and the environmental process of pollutants in multi-contaminated environmental mediums has attracted attention in recent years. In this study, *p*-arsanilic acid (ASA), a kind of organic arsenic feed additive that contains the arsenic group in a chemical structure, is used as a typical contaminant to investigate its adsorption on iron oxides and its implication for contaminated soils. The adsorption kinetics on all solids can be fitted to the pseudo-second-order kinetic model well. At the same mass dosage conditions, the adsorption amount per unit surface area on iron oxides follows the order α-FeOOH > γ-Fe2O<sup>3</sup> > α-Fe2O<sup>3</sup> , which is significantly higher than that for actual soil, because of the lower content of iron oxides in actual soil. Lower pH conditions favor ASA adsorption, while higher pH conditions inhibit its adsorption as a result of the electrostatic repulsion and weakened hydrophobic interaction. The presence of phosphate also inhibits ASA adsorption because of the competitive effect. Correlations between the amount of ASA adsorption in actual soil and the Fe2O<sup>3</sup> content, total phosphorus content, arsenic content, and organic matter content of actual soil are also investigated in this work, and a moderate positive correlation (*R* <sup>2</sup> = 0.630), strong negative correlation (*R* <sup>2</sup> = 0.734), insignificant positive correlation (*R* <sup>2</sup> = 0.099), and no correlation (*R* <sup>2</sup> = 0.006) are found, respectively. These findings would help evaluate the potential hazard of the usage of organic arsenic feed additives, as well as further the understanding of the geochemical processes of contaminants in complicated mediums.

**Keywords:** *p*-arsanilic acid; iron oxides; soil contamination; adsorption

#### **1. Introduction**

Industries in developing cities are diversified, including agriculture, manufacturing, mining industry, etc. This diversification could lead to a multitude of sources and types of pollution to the environment. The simultaneous contamination of nutrients, heavy metals, toxic organics, Pharmaceutical and Personal Care Products, etc., in water and surface soil has been widely reported [1–3]. Such simultaneous contamination is the joint result of domestic sewage discharge, industrial wastewater discharge, livestock and poultry industrial wastewater discharge, and agricultural non-point source pollution.

The geochemical behavior of pollutants in a multi-contaminated environment has attracted attention in recent years. Both competitive and cooperative effects on contaminants' adsorption in the complex system are reported [4–10]. The competitive effect, which is mainly caused by limited surface sites for adsorption, is reported to be relatively more, while the cooperative effect, which is mainly caused by bridge ions or compounds, is relatively less. Because of the intricacy of contamination in the urban environment, studying

**Citation:** Yang, Y.; Tao, S.; Dong, Z.; Xu, J.; Zhang, X.; Pan, G. Adsorption of *p*-Arsanilic Acid on Iron (Hydr)oxides and Its Implications for Contamination in Soils. *Minerals* **2021**, *11*, 105. https://doi.org/10.3390/ min11020105

Received: 5 December 2020 Accepted: 19 January 2021 Published: 22 January 2021

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**Copyright:** © 2021 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (https:// creativecommons.org/licenses/by/ 4.0/).

the transportation behavior of pollutants on actual environmental mediums is of great importance.

Organic arsenic feed additives, 4-aminobenzenearsenic acid (*p*-arsanilic acid (ASA)) and 4-hydorxy-3-nitrobenzenearsenic acid (roxarsone, ROX), have been used in poultry production for decades. They are types of antibiotics that can promote protein synthesis and animal growth, and can prevent the growth of parasites and microorganisms [11,12]. Organic arsenic feed additives tend to be excreted by animals with no significant chemical structural change, and the waste of these animals is often used as a fertilizer in nearby farms, which could lead to the potential contamination of agricultural fields [13]. The occurrence of organoarsenicals in the surface water, soil, and sediment surrounding swine farms has been reported [14]. These organoarsenicals are made up of the total arsenic in the environmental mediums. Although these two organoarsenicals are less toxic than inorganic arsenic, they can turn into inorganic arsenic through the biological or chemical process during their long-term existence in the environment, which will eventually cause damage to the soil and harm human health [15–18].

Studies about the adsorption of arsenic species on solids have been done for decades, the majority of which focus on the behavior of inorganic arsenic species, because of the higher percentage in the determined total arsenic species [19]. These studies have reported the adsorption behavior of inorganic arsenic on pure iron oxides and collected actual soil or sediment, both in the presence or absence of other contaminants. The results show that the adsorption kinetics and adsorption amount of inorganic arsenic were highly affected by the presence of other contaminants, especially phosphate, due to its similar chemical structure to arsenic [20–22]. Tofan-Lazar and Al-Abadleh compared the adsorption kinetics of phosphate on the surface of iron oxides at various conditions, and found that the adsorption rate was the fastest on freshly prepared iron (oxyhydr)oxide and slowest on arsenate-covered iron (oxyhydr)oxide [23]. Although aromatic organoarsenicals are also important as arsenic pollution sources, their adsorption behavior was relatively less studied previously, and has only received attention in recent years. Those works concerning organoarsenical adsorption mainly report on the behavior and mechanisms of organoarsenicals on iron oxides because of the abundance of iron species in soil [24–27], while the information of their adsorption on actual soil, especially contaminated soil, remains unclear.

Because of the complexity of the environmental medium in urban areas, investigations into the sorption pattern of pollutants on actual soil are important in order to understand the geochemical behavior of pollutants. In this work, the adsorption of ASA on both pure iron oxides (α-Fe2O3, γ-Fe2O3, and α-FeOOH) and collected actual surface soil is investigated. In pure iron oxide systems, the adsorption kinetics are investigated at various dosage conditions, and the effects of pH and the presence of phosphate on the ASA adsorption amount were also studied. The actual soil samples were collected from a multi-industrial city (including the agriculture and mining industries) in Hubei Province, China, in order to illustrate the relationship between the adsorption amount/rate and the chemical properties of solids. Describing the environmental transportation behavior of organoarsenicals would help with evaluating the potential hazards associated with the usage of organic arsenic feed additives, and further the understanding of the geochemical behavior of pollutants in multi-contaminated mediums.

#### **2. Materials and Methods**

#### *2.1. Chemicals*

The ASA (98%) was purchased from Aladdin Co. (Shanghai, China) and was used without further purification. α-Fe2O<sup>3</sup> and γ-Fe2O<sup>3</sup> were purchased from Aladdin Co. (Shanghai, China), and α-FeOOH was purchased from Sigma-Aldrich Co. (St. Louis, MO, USA). All of the other reagents were analytically pure and were purchased from Sinopharm Chemical Reagent Co., Ltd. Ultrapure water (18.2 MΩ, obtained through a water purification system, Ming-Che 24UV, Millipore, France) was used for the reagent

preparation and experiments. All of the prepared solutions were stored in polypropylene plastic bottles (Nalgene, Rochester, NY, USA) avoiding light.

#### *2.2. Soil Collection and Pretreatment*

The soils were collected from Jingmen City (Hubei Province, China)—an important agricultural and industrial area with many farms and mining factories. Both the wastewater and waste generated from farms and factories in Jingmen City are extensive and could lead to contamination of the surrounding surface soils. The ecological and environmental protection in the Yangtze River Basin has attracted attention in recent years, and the government of Hubei Province has launched a pollution survey project to control the contamination situation and annual pollution loads, including for Jingmen City, for the purpose of the environmental management and restoration. The sampling sites in this work are shown in Figure 1. The soil samples were collected 10–15 cm below the surface, in order to avoid collecting anthropogenic impurities. The fresh soil samples were preserved at a low temperature and kept out of light before pretreatment.

**Figure 1.** Map showing the locations of the sampling sites.

The collected soil samples were dried at room temperature and were protected from light. Then, the dried solids were ground and sieved using a 100-mesh sieve. The sieved samples were transferred into sample bags and preserved at 4 ◦C.

#### *2.3. Adsorption Experiments*

The adsorption experiments were conducted in a Nalgene bottle with continuous stirring using a magnetic stirrer at a speed of 750 r/min. The mixed solution, containing 20 µM ASA and 10 mM NaCl, was prepared and adjusted to the desired pH value with diluted NaOH/HCl before adding the solids. Samples were taken at interval times and were filtered with a 0.22 µm polyethersulfone filter membrane for further analysis. In the kinetic adsorption experiments, the solution volume was 200 mL, and the pH value was controlled by re-adjusting several times throughout the entire adsorption period. In the pH effect experiments, the solution volume was 100 mL, the pH value was also re-adjusted several times, and the precise pH value was recorded. In the phosphate competition experiments, the solution volume was also 100 mL, a certain amount of phosphate was also added to the mixed solution, and the rest procedures were the same as described above. Here, >90% of ASA was be desorbed by the co-presence of phosphate and alkaline (pH > 12, 2 mM PO<sup>4</sup> <sup>3</sup>−), confirming that no degradation occurred during adsorption.

#### *2.4. Analytical Methods*

The concentration of ASA was analyzed as described in our previous works [12,18,28]. High-performance liquid chromatography (HPLC; a 20ADVP pump, a DAD-20AVP detector, Shimadzu Instrument Co. Ltd., Kyoto, Japan) with a C18 column (Supelco Discovery, 4.6 mm × 250 mm, 5 µm) was used for the analysis. The mobile phase was a mixture of a 2.5% formic acid and methanol solution (95:5, *v*/*v*). The flow rate was set to 1 mL/min and the detection wavelength was set to 254 nm.

The BET surface area of the iron oxide solids was analyzed using a surface area and porosimetry analyzer (V-Sorb 2800P, Gold APP Instrument Co., Beijing, China). The total phosphorus content of the collected soil samples was analyzed according to the national standard using an ICAP6300 Plasma Emission spectrometer (Thermo Fisher, Waltham, MA, USA). The element content (Fe2O<sup>3</sup> and As) of the soils was analyzed using an Xray fluorescence spectroscopy (XRF, Explorer 9000, Jiangsu Skyray Instrument Co., Ltd., Kunshan, China). The content of organic matter (OM) was estimated by measuring the ignition loss. The pH of the solutions was determined using a pH meter (F2-Meter, Mettler Toledo, Greifensee, Switzerland).

#### *2.5. Statistical Analysis*

The amount of ASA (*q*<sup>t</sup> , <sup>µ</sup>mol·m−<sup>2</sup> ) adsorbed was calculated using the difference between their concentrations at the initial time and at time t. A pseudo-second-order kinetic model was used to fit the kinetic data for ASA, which can be expressed as follows:

$$\frac{t}{q\_t} = \frac{1}{q\_\varepsilon^2 \ast k\_2} + \frac{1}{q\_\varepsilon}t \tag{1}$$

where *k*<sup>2</sup> (m<sup>2</sup> ·µmol−<sup>1</sup> ·h −1 ) is the pseudo-second-order rate constant and *q*<sup>e</sup> is the amount of adsorption at equilibrium time. *k*<sup>2</sup> and *q*<sup>e</sup> can be obtained from the slope and y-intercept of the plots of *t*/q<sup>t</sup> vs. *t*. The adsorption percentage was calculated by the difference between the concentrations at the initial and ending times, which can be expressed as follows:

$$\text{adsorption percentage (\%)} \quad = \left(1 - \frac{\text{C}\_{\text{f}}}{\text{C}\_{0}}\right) \times 100\text{\%} \tag{2}$$

where *C*<sup>0</sup> and *C*<sup>t</sup> are the concentrations of ASA at the initial and ending times, respectively.

#### **3. Results and Discussion**

#### *3.1. Adsorption Kinetics of ASA on Iron Oxides*

The adsorption kinetics of ASA on three iron oxides is investigated at various iron oxide dosages at pH 5. The adsorption kinetics and equilibrium time of ASA on three iron oxides were close. As shown in Figure 2a–c, after adding iron oxides into the mixture solution, the concentrations of ASA decreased significantly in the early stage (within 4 h), while the changes gradually became slow in the later stage, and reached equilibrium at around 24 h. Among the three iron oxides, α-Fe2O<sup>3</sup> had the fastest adsorptive removal efficiency; more than 95% of ASA was be removed from the solution within 30 min at 1 g·L <sup>−</sup><sup>1</sup> dosage, whereare α-FeOOH was the slowest, with the removal efficiency decreasing to only 49% at the same condition. The BET surface areas of α-Fe2O3, γ-Fe2O3, and α-FeOOH were 125.04, 53.39, and 9.50 m<sup>2</sup> ·g −1 , respectively. The significant difference in adsorptive removal efficiency among each of the iron oxides could be caused by the highly different surface areas. The calculated adsorption amount of ASA, normalized to the surface area, is shown in Figure 2d–f. The obtained parameters for the pseudo-secondorder kinetic model are shown in Table 1, and the obtained calculated kinetics from the parameters are also given in Figure 2d–f.

**Figure 2.** Adsorption kinetics of *p*-arsanilic acid (ASA) on α-Fe2O<sup>3</sup> , γ-Fe2O<sup>3</sup> , and α-FeOOH. The solid points are experimental data. The polylines in (**a**–**c**) are for the visual guides. The curves in (**d**–**f**) are the fitted results according to the pseudo-second-order kinetic model. ASA = 20 µM, NaCl = 10 mM, and pH = 5.

**Table 1.** Fitted parameters for ASA adsorption on iron oxides.


As can be seen from Table 1, *q*<sup>e</sup> and *k*<sup>2</sup> show an opposite trend to each other. With *q*<sup>e</sup> for α-Fe2O3, γ-Fe2O<sup>3</sup> decreases dramatically when increasing the solid dosages (i.e., total surface area) by an order of magnitude, and the difference can be as high as 5.5 times for α-Fe2O3. However, the change is less significant for α-FeOOH at the same condition. Such results are thought to be caused by the valid adsorption sites on iron oxide surface [29]. At limited surface area conditions, the adsorption sites on the surface of the iron oxides could be nearly fully occupied by ASA, while at abundant surface area conditions, the adsorption sites are also abundant, and thus would not be completely occupied. At the same solid dosage conditions, the surface areas of α-Fe2O<sup>3</sup> and γ-Fe2O<sup>3</sup> were much larger than that of α-FeOOH; therefore, the two former iron oxides had more vacancy sites. Although the dosage of α-FeOOH increased from 0.1 g·L −1 to 1.0 g·L −1 , the amount of adsorption sites were still limited compared with the dosed ASA, thus the change in *q*<sup>e</sup> was less significant than that of α-Fe2O<sup>3</sup> and γ-Fe2O3. In order to avoid the extreme adsorption circumstances and to obtain a better observation, the dosage of iron oxides in the following experiment was 0.2 g·L −1 .

#### *3.2. Effect of pH*

Previous works have reported the obvious effect of pH on organic compounds' adsorption behavior, which is caused by the joint effect of compounds and solids [30,31]. Here, experiments were therefore conducted at various pH conditions in order to investigate the adsorption behavior of ASA on the three iron oxides. First, 10 mM NaCl was introduced to the mixed solution in order to eliminate the ionic strength effect caused by the pH. As can be seen from Figure 3a, among the three iron oxides, α-Fe2O<sup>3</sup> shows a relatively higher adsorptive removal percentage for the entire investigated pH range, and that of α-FeOOH is the lowest at the same mass dosage. This trend was reversed when calculating the adsorption amount normalized to the surface area (Figure 3b): α-FeOOH showed the highest *q*<sup>e</sup> for the overall investigated pH conditions. For all three oxides, ASA adsorption showed a decrease trend with an increase of pH. The adsorption percentage of ASA on α-Fe2O<sup>3</sup> showed a sharp decrease from 93% to less than 10% when the pH increased from 4.08 to 11.88, and that for α-FeOOH also showed a gradual decrease from 31% to less than 10% when the pH increased from 4.00 to 11.48.

**Figure 3.** Effect of pH on ASA adsorption on iron oxides. (**a**) Data shown as adsorption percentage (%) and (**b**) data shown as adScheme 2. Solid points are experimental data; polylines are for the visual guide. ASA = 20 µM, iron oxdies = 0.2 g·L −1 , and NaCl = 10 mM.

Such a trend is similar to the previous reported works. Bell-shaped adsorption curves with pH have been widely reported for the adsorption of organic compounds on (hydr)oxides [5,31,32], and both bell-shaped and cliff-shaped curves have been found for the adsorption of inorganic arsenic species on oxides or minerals [10,33,34]. The adsorption behavior of ASA is also highly affected by its arsenic group in its chemical structure. ASA has 3 acidity coefficients—2.00, 4.02, and 8.92. Therefore, at neutral and basic conditions, ASA exists in anion form [18,35]. In the meantime, the point of zero charge for α-Fe2O3, γ-Fe2O3, and α-FeOOH were at a circumneutral pH [36–38], also showing a negative charge at high pH conditions. Therefore, the higher pH would exacerbate the electrostatic repulsion and weaken the hydrophobic interaction between the oxyanion compounds and solids.

#### *3.3. Competing Ion (Phosphate) Effect on Equilibrium Adsorption*

Because of the similarity in the chemical structure between the phosphate and arsenate groups, the effect of phosphate on the chemical behavior of arsenic is often considered when investigating the adsorption of arsenic species. Cheng et al. studied the effect of various anions (CO<sup>3</sup> <sup>2</sup>−, SiO<sup>3</sup> <sup>2</sup>−, Cl−, F−, SO<sup>4</sup> <sup>2</sup>−, NO<sup>3</sup> <sup>−</sup>, and HPO<sup>4</sup> <sup>2</sup>−) on the adsorption behavior of As(III) in iron-containing materials, and found that HPO<sup>4</sup> <sup>2</sup><sup>−</sup> has the highest inhibition effect [39]. The removal efficiency by adsorption decreased from >90% to only about 60% in the presence of HPO<sup>4</sup> <sup>2</sup>−. Wang et al. reported the obvious competitive effect between the inorganic arsenic adsorption and phosphorus release by sediments [21]. Lin et al. proposed that the use of phosphorus fertilizers could enhance the mobility of arsenic towards groundwater in arsenic-contaminated aquifers [40]. The inhibition effect of phosphate on organic arsenic adsorption has previously been reported [25,41].

In this work, the effect of phosphate on the adsorption of ASA on α-Fe2O3, γ-Fe2O3, and α-FeOOH were investigated at a wide phosphate concentration range. The inhibition effect could be observed over the whole investigated pH range (4–12) for all the three iron oxides (Figure 4). A significantly higher inhibition effect was observed when the concentration of phosphate increased from 0.02 mM to 2 mM. This strong competitive effect reduced the adsorption ability of ASA on phosphorus abundant soil, thus leading to an enhancement in ASA mobility. Despite this, the high amount of phosphate did not show the complete inhibition of ASA, especially in acid conditions. Liu et al. used the concentrated phosphate solution (0.5 M H3PO4) to extract arsenic species from contaminated soils, and the extraction efficiency for ASA was relatively low (67%) with a high error bar, while that for As (V) was close to 100% [14]. They suggested that the degradation of ASA might have occurred during the long time extraction process (16 h), which, in fact, might be also caused by an improper acid extraction condition (pH < 2 for 0.5 M H3PO4).

#### *3.4. Actual Soil Adsorption of ASA*

The content of actual soil is complex, and contains metal oxides, nutrients, and inorganic and organic contaminants. Such compounds might all have implications on the contaminant adsorption behavior and mobility, resulting in a positive or negative effect. Table 2 shows the contents of Fe2O3, total phosphorus (TP), As, and organic matter (OM) of the nine collected soils. The BET surface area and soil pH are also represented in Table 2. The correlation of each of the parameters is investigated. As our parameters and compound contents were moderately/strongly skewed, we computed their log-transformation. Interestingly, a slight correlation between these parameters could be observed. Figure 5a shows a slight positive correlation between Fe2O<sup>3</sup> and As (*R* <sup>2</sup> = 0.278). These results are not surprising, because the adsorption of arsenic species is highly related to the presence of iron oxides [42]. As the sampling area has many mining activities, it could lead to the contamination of heavy metals, including arsenic species. The higher content of iron oxides in actual soil could therefore immobilize more heavy metals on the surface soils. Nevertheless, the content of As is a jointly affected by the surrounding human activities and the adsorption ability of the soil. Therefore, the current As content/contamination level did not show a strong positive correlation with the iron content. Hafeznezami et al. also reported an increased trend in the adsorption ability of As at a higher content of amorphous Fe, although the statistical correlation was not significant [43]. Surprisingly, opposite to As, a negative correlation between Fe2O<sup>3</sup> and TP was observed (Figure 5b, *R* <sup>2</sup> = 0.505), which was unexpected, as the presence of iron oxides should enhance the

retention ability of phosphorus in soil [44,45]. This phenomenon might be attributed to the proportion of amorphous species. There was no correlation between As content and TP content (*R* <sup>2</sup> = 0.010, results not shown), indicating that the available adsorption sites were still abundant for the pollutant sorption, and the competitive effect between As and TP was insignificant for the current contamination situation. The contributions of each composition to the surface area of the soil samples were also investigated. A slight positive correlation between Fe2O<sup>3</sup> content and the surface area can be observed in Figure 5c (*R* <sup>2</sup> = 0.329), indicating a relatively important contribution. TP showed a moderate negative correlation with the surface area (*R* <sup>2</sup> = 0.602, data not shown), which might be caused by its relationship with the Fe2O<sup>3</sup> content. The content of OM and As showed unimportant contributions to the surface area (*R* <sup>2</sup> = 0.016 for OM and *R* <sup>2</sup> = 0.006 for As).

**Figure 4.** Effect of phosphate on ASA adsorption. Solid points are experimental data; polylines are for the visual guide. ASA = 20 µM, iron oxdies = 0.2 g·L −1 , and NaCl = 10 mM. (**a**) α-Fe2O<sup>3</sup> , (**b**) γ-Fe2O<sup>3</sup> , (**c**) α-FeOOH.


**Table 2.** Properties of the collected actual soil.

\* determined after the ASA adsorption reaction.

**Figure 5.** Correlation between the (**a**) Fe2O<sup>3</sup> content and As content, (**b**) Fe2O<sup>3</sup> content and total phosphorus (TP) content, (**c**) Fe2O<sup>3</sup> content and surface area, and (**d**) organic matter (OM) content and surface area for the collected actual soil.

The adsorption kinetics of ASA for the nine collected soil samples were then investigated. The strong interactions between arsenic species and iron oxides have been widely reported, while the interactions between arsenic species and pure silica (which is the main composition of soil) seem to be less important for the adsorption efficiency of arsenic species. Although the surface area of the soil samples was close to pure iron oxides (γ-Fe2O<sup>3</sup> and α-FeOOH), the valid adsorption sites on the solid surface were be much lower. In order to ensure experimental accuracy (i.e., a relatively obvious adsorption percentage), a high solid dosage (5 g·L −1 ) was used in this section. No pH adjustment was done before or during the adsorption reaction, as the high dosage of soil could form a buffer system, although the final pH was determined. The pseudo-second-order kinetic model was also used to fit the adsorption kinetic results. The experimental results and fitted results are shown in Figure 6, and the calculated parameters for the pseudo-second-order kinetic model are shown in Table 3. The obtained *q*<sup>e</sup> varied significantly with the different soil samples, with SP8 showing the highest result (0.0309 <sup>µ</sup>mol·m−<sup>2</sup> ), which was ~8.4 times as high as the lowest (SP1, 0.0037 <sup>µ</sup>mol·m−<sup>2</sup> ), while all of them were obviously lower than that of the pure iron oxide surface (α-Fe2O3, γ-Fe2O3, and α-FeOOH), indicating invalid adsorption sites on the solid surface. Such a big difference would be mainly caused by the

content of iron oxides. The adsorption rate constants were mainly opposite to the trend of the adsorption amount.

**Figure 6.** Adsorption kinetics of ASA on the collected actual soil. Solid points are experimental data, and curves are fitted results by pseudo-second-order kinetic model. ASA = 20 µM, NaCl = 10 mM, and soil dosage = 5 g·L −1 .

**Table 3.** Fitted parameters for ASA adsorption of soil by a pseudo-second order kinetic model.


Previous works have reported that the adsorption of arsenic species is positively correlated with the metal oxide content and is negatively correlated with the phosphorus content [46,47]. The correlations of the ASA adsorption parameters with the Fe2O<sup>3</sup> and TP content in the actual soils were fitted in this work. As can be seen from Figure 7a, a moderate positive correlation between the *q*<sup>e</sup> and Fe2O<sup>3</sup> content can be observed, as expected. As iron oxides have strong interactions with ASA, the higher Fe2O<sup>3</sup> content would obviously facilitate the adsorption of ASA. However, the positive correlation was not very strong in the limited soil samples, indicating the effect of other physical and chemical parameters on ASA adsorption, which was similar to the situation of the arsenate adsorption on sandy sediments [43]. In contrast, the correlation between *q*<sup>e</sup> and TP content is strongly negative (see Figure 7b). The strong negative correlation between the ASA adsorption and TP content would be caused by the competitive effect, as discussed in Section 3.3, which therefore decreased the statistical correlation between the *q*<sup>e</sup> and Fe2O<sup>3</sup> content. The correlation between the *q*<sup>e</sup> and As content was also investigated, and an insignificant positive correlation was found, indicating potential relevance between ASA adsorption ability and the current As contamination level. Such results also reveal abundant available adsorption sites on the surface of the soil, although these soil samples have been contaminated by As to a certain degree. It seems that the OM content did not affect *q*<sup>e</sup> (Figure 7d), probably because of the joint reason of its low content and its weak sorption ability. The final pH after 48 h of the adsorption reaction was very close between the nine samples; therefore, the statistic correlation between *q*<sup>e</sup> and pH was insignificant (*R* <sup>2</sup> = 0.005, data not shown). In general, the adsorption results on the collected actual soils indicate

that predicting the adsorption behavior was not possible because of single or very few physical and chemical parameters.

**Figure 7.** Correlation fitting between ASA adsorption amount (*q*e) and (**a**) Fe2O<sup>3</sup> content, (**b**) TP content, (**c**) As content, and (**d**) OM content.

#### **4. Conclusions**

The adsorption behavior of ASA on three kinds of pure iron oxides and nine collected actual soil samples are studied in this work. The adsorption kinetics on all solids are well fitted to the pseudo-second-order kinetic model. The dosage of iron oxides, pH conditions, and the concentration of co-present phosphate could all affect the adsorption amount of ASA on iron oxides. The parameters of the actual soil that can affect ASA adsorption are more complicated than those of pure iron oxides. Although the adsorption amount shows a moderate positive correlation with the Fe2O<sup>3</sup> content, strong negative correlation with the TP content, insignificant positive correlation with the As content, and no correlation with the OM content, it is still difficult to obtain an accurate prediction model because of the complexity of the physical and chemical properties of those soils. Investigating the transportation of contaminants in multi-contaminated environmental mediums helps to further understand their geochemical processes and helps with formulating a remediation strategy for the contaminated area.

**Author Contributions:** Conceptualization, X.Z.; data curation, Z.D.; formal analysis, Y.Y. and G.P.; funding acquisition, J.X.; investigation, Y.Y., S.T., and Z.D.; methodology, G.P.; resources, J.X. and X.Z.; supervision, J.X.; validation, Z.D.; visualization, Y.Y. and S.T.; writing—original draft, Y.Y.; writing—review and editing, S.T., X.Z., and G.P. All authors have read and agreed to the published version of the manuscript.

**Funding:** National Natural Science Foundation of China: 21707106. National Natural Science Foundation of China: 42077350. Strategic Priority Research Program of the Chinese Academy of Sciences: XDA23040300. China Postdoctoral Science Foundation: 2016M602358.

**Institutional Review Board Statement:** Not applicable.

**Informed Consent Statement:** Not applicable.

**Data Availability Statement:** Data is contained within the article.

**Acknowledgments:** We thank T. Chen and T. Luo for their help in the actual soil sampling. Comments from the anonymous reviewers are also appreciated.

**Conflicts of Interest:** The authors declare no conflict of interest.

#### **References**


## *Article* **Describing Phosphorus Sorption Processes on Volcanic Soil in the Presence of Copper or Silver Engineered Nanoparticles**

**Jonathan Suazo-Hernández 1,2 , Erwin Klumpp 3 , Nicolás Arancibia-Miranda 4,5 , Patricia Poblete-Grant 2 , Alejandra Jara 2,6 , Roland Bol <sup>3</sup> and María de La Luz Mora 2,6, \***


**Abstract:** Engineered nanoparticles (ENPs) present in consumer products are being released into the agricultural systems. There is little information about the direct effect of ENPs on phosphorus (P) availability, which is an essential nutrient for crop growth naturally occurring in agricultural soils. The present study examined the effect of 1, 3, and 5% doses of Cu <sup>0</sup> or Ag <sup>0</sup> ENPs stabilized with L-ascorbic acid (suspension pH 2–3) on P ad- and desorption in an agricultural Andisol with total organic matter (T-OM) and with partial removal of organic matter (R-OM) by performing batch experiments. Our results showed that the adsorption kinetics data of H2PO<sup>4</sup> − on T-OM and R-OM soil samples with and without ENPs were adequately described by the pseudo-second-order (PSO) and Elovich models. The adsorption isotherm data of H2PO<sup>4</sup> − from T-OM and R-OM soil samples following ENPs addition were better fitted by the Langmuir model than the Freundlich model. When the Cu <sup>0</sup> or Ag <sup>0</sup> ENPs doses were increased, the pH value decreased and H2PO<sup>4</sup> − adsorption increased on T-OM and R-OM. The H2PO<sup>4</sup> − desorption (%) was lower with Cu <sup>0</sup> ENPs than Ag 0 ENPs. Overall, the incorporation of ENPs into Andisols generated an increase in P retention, which may affect agricultural crop production.

**Keywords:** adsorption; engineered nanoparticles; organic matter; phosphorus; nutrients; pollution; volcanic soil

### **1. Introduction**

In the past decade, the incorporation of engineered nanoparticles (ENPs) into consumer products [1,2] has led to a significant increase in their turnover from \$250 billion in 2009 to \$3 trillion in 2020 [3]. Two of the most widely used ENPs in consumer products are metallic copper (Cu 0 ) and silver (Ag 0 ), due to their antibacterial properties. Cu <sup>0</sup> ENPs are added to biocides, electronics, paints, cosmetics, agrochemicals, ceramics, and film [1,3,4], whereas Ag <sup>0</sup> ENPs are used in textiles, air filters, bandages, paints, food storage containers, agrochemicals, deodorants, toothpaste, and household appliances [5]. Thus, as a consequence of extensive and diverse commercial applications, these ENPs can be released into the environment. Soil is the main sink of disposal for most of the released ENPs [6]. Adverse effects on human health and ecosystems may be expected, making it necessary

**Citation:** Suazo-Hernández, J.; Klumpp, E.; Arancibia-Miranda, N.; Poblete-Grant, P.; Jara, A.; Bol, R.; de La Luz Mora, M. Describing Phosphorus Sorption Processes on Volcanic Soil in the Presence of Copper or Silver Engineered Nanoparticles. *Minerals* **2021**, *11*, 373. https://doi.org/10.3390/ min11040373

Academic Editors: Ana Romero-Freire and Hao Qiu

Received: 5 March 2021 Accepted: 29 March 2021 Published: 1 April 2021

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**Copyright:** © 2021 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (https:// creativecommons.org/licenses/by/ 4.0/).

to improve our current understanding of environmental risks, fate, transformations and aggregation behaviors of metallic ENPs [7].

The geochemistry of metallic Cu<sup>0</sup> and Ag<sup>0</sup> ENPs in soils is complex, due to their chemical transformation between Cu<sup>0</sup> , Cu<sup>+</sup> and Cu2+ as well as between Ag<sup>0</sup> and Ag<sup>+</sup> , respectively [1,4], also due to their strong binding capacity to various soil components like clay minerals, organic matter, microorganisms, among others. Transformations of metallic ENPs in soil include oxidation, dissolution, and sulfidation. Over time, Cu<sup>0</sup> ENPs can be oxidized in the soil to form CuO (tenorite) and Cu2O (cuprite) nanoparticles with a core-shell structure. Any of these, both forms of copper oxide nanoparticles, can dissolve and release cuprous and/or cupric ions into solution [8]. Meanwhile, the Ag<sup>0</sup> ENPs show a slow oxidation process, which can be promoted in acid soils. The metallic ENPs oxidation in soils can be diminished when organic molecules are used as stabilizing agents [9]. Transformation on metallic ENPs is an important consideration to developing risk assessments of ENPs [4,9].

Several studies have intended to determine the effects caused by ENPs on soil properties. In these studies, it has been shown than due to metallic Cu<sup>0</sup> and Ag<sup>0</sup> , ENPs are characterized by a high surface area and chemical reactivity, variable surface charge and chemical transformation [10]. Once in contact with soil, ENPs may therefore modify their structural and physico-chemical properties such as pH, electric conductivity, redox potential, porosity, and hydraulic conductivity [10–12]. This could affect reactions and processes of elements in soil, such as precipitation, dissolution, co-precipitation, complexation, oxidation/reduction, plant uptake, and ad- and desorption. Particularly, ad- and desorption are important because they control the availability and mobility of contaminants and nutrients [10]. In this context, Taghipour and Jalali [13] reported that metal oxide ENPs (Al2O<sup>3</sup> and TiO2) caused immobilization of phosphorus (P) in calcareous soils from Hamadan, Western Iran, and reduced the bioavailability of P.

In volcanic soils (Andisol and Ultisol), P is an essential crop macronutrient and this soil contains between 1000 and 3500 mg·kg−<sup>1</sup> [14]. However, P availability for plant growth is limited because it can form inner-sphere complexes by ligand exchange with surface -OH and -OH<sup>2</sup> <sup>+</sup> groups of soil components like ferrihydrite, imogolite, allophane, and Al(Fe) humus complexes [15–17]. Numerous studies have focused on P availability in volcanic soils considering the effects on soils of fertilizers [18], liming [19], microorganisms [20,21], enzymes [22], inorganic/organic ligands [23], specific surface area [24], surface charge [25], organic matter content [26], and pH and mineralogy [27].

In relation to effects caused by ENPs in volcanic soils, no studies have assessed the influence of metallic ENPs on the adsorption of nutrients. In this context, the aim of this research was to evaluate the effect of Cu<sup>0</sup> or Ag<sup>0</sup> ENPs on phosphorus sorption processes in volcanic soils and its relationship with organic matter content. Overall, the results provide new information about the implication of ENPs for nutrient availability in soils.

#### **2. Materials and Methods**

#### *2.1. Chemicals Used*

The reagents used were CuCl2·2H2O, AgNO3, L-ascorbic acid, KH2PO4, KCl, HCl, and KOH (analytical grade, Merck) and double-distilled water. The pH electrode (Orion Star A211 pH Benchtop Meter, Thermo Fischer Scientific Beverly, Waltham, MA, USA) was calibrated using standard buffers of 4.01, 7.01, and 10.01 (Hanna, Woonsocket, RI, USA).

#### *2.2. Synthesis of Cu<sup>0</sup> and Ag<sup>0</sup> ENPs*

CuCl2·2H2O and AgNO<sup>3</sup> were used for the formation of Cu<sup>0</sup> , and Ag<sup>0</sup> ENPs, respectively, and L-ascorbic acid was added as a reducing and capping agent [28]. Cu<sup>0</sup> ENPs (or Ag<sup>0</sup> ENPs) was synthesized by mixing 10.0 mmol·<sup>L</sup> <sup>−</sup><sup>1</sup> CuCl2·2H2O (or 10.0 mmol·L <sup>−</sup>1AgNO3) in 50 mL double-distilled water. An Erlenmeyer flask (100 mL), containing the CuCl2·2H2O (or AgNO3) solution, was heated in a water bath at 80 ◦C with magnetic stirring; 50 mL of L-ascorbic acid (1.0 mol·L −1 ) was added dropwise into the flask

while stirring. The aqueous dispersion of stabilized Cu<sup>0</sup> ENPs (or Ag<sup>0</sup> ENPs) obtained was kept at 80 ◦C for 24 h and it was finally saved to ambient conditions for later research.

#### *2.3. Soil Samples*

The soil used was an Andisol belonging to Santa Barbara series from Southern Chile (36◦50′ S; 71◦55′ W). The soil was collected from the top 20 cm depth of the soil horizon. The soil was passed through a <2 mm mesh sieve and freeze-dried (total organic matter soil sample = T-OM). For partial removal of organic matter (OM), the T-OM soil sample was treated several times with H2O<sup>2</sup> until adding did not result anymore in air bubbles emanating from the aqueous solution and maintained at 40 ◦C in a thermoregulated bath [29]. The resulting sample was then washed four times with double-distilled water (partial removal of OM soil sample = R-OM). Finally, both soil samples were freeze-dried and stored at 4 ◦C.

#### *2.4. Characterization of Ag<sup>0</sup> and Cu<sup>0</sup> ENPs*

The synthetized Cu<sup>0</sup> and Ag<sup>0</sup> ENPs were characterized using transmission electron microscopy (TEM) on a Hitachi model HT7700 (Hitachi, Tokyo, Japan) with Olympus camera (Veleta 2000 × 2000) using high resolution mode at 120 kV. The TEM images obtained were analyzed manually to calculate the particle size with the ImageJ program (version 1.50i, Wayne Rasband, National Institute of Health, Bethesda, MD, USA). The ultraviolet-visible (UV–Vis) spectra was recorded with a double-beam Rayleigh UV-2601 spectrophotometer (BRAIC Co. Ltd., Beijing, China) using 1 cm path length glass cell. The zeta potential (ZP) of Cu<sup>0</sup> and Ag<sup>0</sup> ENPs (25 mg) was measured in the presence of 10 mL KCl 0.01 M using a Nano ZS apparatus (Malvern Instruments, Worcestershire, UK) at 20 ◦C and the isoelectric point (IEP) was obtained from graphs of ZP versus pH. The Fourier-transform infrared spectroscopy (FT-IR) were recorded with a 1 mL of ENPs suspension. FT-IR analysis was realized using a Cary 630 spectrometer (Agilent Technologies, Santa Clara, CA, USA). The transmission spectrum was acquired with 4 cm−<sup>1</sup> resolution and the operating range was 600 cm−<sup>1</sup> to 4000 cm−<sup>1</sup> at atmospheric pressure and 20 ◦C. The pH of the suspensions of ENPs was measured with 10 mL using a pH Meter.

#### *2.5. Characterization of Soil Samples*

The morphological characteristics of both soil samples were obtained by scanning electron microscopy with a STEM SU-3500 transmission module (Hitachi, Tokyo, Japan) and the QUANTAX 100 energy-dispersive X-ray spectrometer detector (EDX), (Bruker, Berlin, Germany) was used for the semi-quantitative analysis of the elemental composition (Al, Si, and Fe content). 20 mg of each soil sample were deposited onto 300-mesh Formvar/carboncoated grids and were inspected under a high-vacuum. Confocal analysis was performed by laser scanning confocal microscopy (LSCM) using the Olympus Fluoview1000 (Olympus Optical Co., Melville, New York, NY, USA). 50 µL of the suspensions were collocated on a microscope slide with a micropipette and the sample was dried on a stove at 40 ◦C. The total organic carbon (TOC) of T-OM and R-OM soil samples was calculated using a Shimadzu TOC-V CPH instrument (Shimadzu, Tokyo, Japan). The TOC was transformed into soil organic matter content using the conversion factor of 1.72 [30]. The specific surface area of R-OM and T-OM soils was obtained using the Brunauer, Emmett and Teller (BET) theory. Approximately 200 mg of soil sample was degassed for 2 h at 105 ◦C and then was conducted using N<sup>2</sup> gas at −196 ◦C in the relative pressure range (P/P0) of 0.05–0.4. Surface area measurements were made with a Quantachrome Nova 1000e analyzer (Quantachrome Instruments, Boynton Beach, FL, USA). The average pore volume and size were obtained using the Barrett-Joyner-Halenda (BJH) model. For the FT-IR absorption spectrum, soil samples were dried at 50 ◦C for 12 h to eliminate the interference produced by the absorption of the water molecules. To determine the functional groups in both soil samples, the analysis was performed under similar conditions to the ENPs. Soil pH was determined in 1:2.5 soil: double-distilled water ratio after 5 min shaking and 120 min

resting, using the same pH Meter used for ENPs determination. Total P was extracted from the soil samples by alkaline oxidation with sodium hypobromite (NaBrO) [31]. After each extraction, the supernatant was filtered (5C, Advantec) and then the concentration of total P in the supernatant was determined using a spectrophotometer Rayleigh UV-2601 with a wavelength of 880 nm [32]. Exchangeable Al was extracted with KCl (1 M) and measured using a Unicam model Solaar 969 atomic absorption spectrophotometer (AAS) (Unicam Ltd, Cambridge, UK). Exchangeable base cations (Na, K, Mg and Ca) in soils were extracted using NH4Ac (1 M, pH 7.0) and were measured by AAS [33]. Effective cation exchange capacity (ECEC) was calculated as the sum of exchangeable Al plus the exchangeable base cations [33].

The ZP and IEP of the soil samples were determined pre- and post-adsorption of H2PO<sup>4</sup> − on T-OM and R-OM soil samples in the absence and presence of 5% Cu<sup>0</sup> or Ag<sup>0</sup> ENPs using the high point adsorption isotherms similar to the procedure followed by ENPs.

#### *2.6. Adsorption Experiments*

Batch experiments were conducted to investigate the adsorption of phosphate (indicated as H2PO<sup>4</sup> −) on T-OM and R-OM soil samples in the absence and presence of 0, 1, 3, and 5% Cu<sup>0</sup> or Ag<sup>0</sup> ENPs doses (% *w/w*). Cu<sup>0</sup> or Ag<sup>0</sup> ENPs doses were added to 0.5 g (dry weight) of soil samples in polyethylene tubes and mixed with 20 mL H2PO<sup>4</sup> − solution. The adsorbed amounts of H2PO<sup>4</sup> − (qt , mmol·kg−<sup>1</sup> ) were determined as the difference between initial concentration and final concentration of H2PO<sup>4</sup> − in the solution (Equation (1)).

$$\mathbf{q}\_{\rm t} = \frac{(\mathbf{C}\_0 - \mathbf{C}\_{\rm t})\mathbf{V}}{\mathbf{w}} \tag{1}$$

where, C<sup>o</sup> is the initial concentrations of H2PO<sup>4</sup> − and C<sup>t</sup> is the concentrations of H2PO<sup>4</sup> − at time t or the equilibrium concentration (mmol·L −1 ), w the weight (kg) of the soil and V is the volume (L).

To evaluate the pH effect on the adsorption of H2PO<sup>4</sup> − onto T-OM and R-OM soil samples, stock solutions of 6.47 mmol·L <sup>−</sup><sup>1</sup> of H2PO<sup>4</sup> − were prepared with double-distilled water at pH ranging from 4.5 to 8.5 by adding 0.1 M HCl or KOH and ionic strength 0.01 M KCl (background electrolyte). The H2PO<sup>4</sup> − solutions were added to soil samples with and without ENPs and were stirred at 200 rpm for 24 h at 20 ± 2 ◦C.

For the kinetic study, the initial solution of 6.47 mmol·L <sup>−</sup><sup>1</sup> of H2PO<sup>4</sup> − was adjusted to pH 5.5 ± 0.2 by adding 0.1 M HCl or KOH at ionic strength 0.01 M KCl and 20 ± 2 ◦C. Samples were taken from the suspension at 2.5, 5, 10, 30, 45, 60, 120, 180, 360, 720, and 1440 min, and H2PO<sup>4</sup> − was determined in solution. Furthermore, the initial pH (pH<sup>i</sup> ) and the final pH (pH<sup>f</sup> ) were measured after H2PO<sup>4</sup> − solution was added to soil samples (time 0 min) and after H2PO<sup>4</sup> − adsorption (1440 min), respectively.

Adsorption isotherms were obtained by varying the initial H2PO<sup>4</sup> − concentrations from 0.016 to 9.71 mmol·L <sup>−</sup><sup>1</sup> and were initially adjusted to pH 5.5 <sup>±</sup> 0.2 and ionic strength 0.01 M KCl. The suspensions were stirred at 200 rpm in an orbital shaker at 20 ± 2 C for 24 h. To determine the effect of copper (Cu2+) or silver cations (Ag<sup>+</sup> ) or L-ascorbic acid on H2PO<sup>4</sup> − adsorption onto T-OM and R-OM soil samples, adsorption isotherms were made in the presence of 3% Cu2+ or Ag<sup>+</sup> or L-ascorbic acid (% *w/w*) under the aforementioned experimental conditions.

The desorption experiment was performed once the adsorption isotherm procedure had ended by adding 20 mL of double-distilled water three times, and the samples were then stirred at 200 rpm in an orbital shaker at 20 ± 2 ◦C for 24 h. The desorption percentages (%) were calculated by the equation used by Silva-Yumi et al. [34] All the samples of the adsorption experiments were first centrifuged at 10,000 rpm for 10 min, using a centrifuge RC-5B Plus (Sorvall, Newtown, CT, USA) and then filtered through 0.22 µm syringe filters. In all experiments, the concentration of H2PO<sup>4</sup> − in the supernatant was determined according to the procedure followed for total P. To minimize manipulation errors in the analysis, the adsorption experiments were performed in triplicate.

#### *2.7. Data Analysis*

The kinetics adsorption (e.g., pseudo-first-order, pseudo-second-order, and Elovich) and isotherm (e.g., Langmuir and Freundlich) models used in this study are presented in Tables 1 and 2, respectively.

**Table 1.** The kinetic models used for the description of phosphate adsorption.


\* From PSO initial adsorption rate (h), can be calculated by multiplying k2q 2 t (mmol·kg−<sup>1</sup> ·min−<sup>1</sup> ).



The data were evaluated through the Chi-square (χ 2 ), adding the coefficient of determination (r<sup>2</sup> ) (Equations (2) and (3)). The lowest χ <sup>2</sup> and highest r<sup>2</sup> values were used as the best fit [37]. The statistical analysis of the adsorption data was conducted using Origin Pro 8.0.

$$\chi^2 = \sum \frac{(\mathbf{q}\_{\text{e,exp}} - \mathbf{q}\_{\text{e,cal}})^2}{\mathbf{q}\_{\text{e,cal}}} \tag{2}$$

$$\mathbf{r}^2 = \sum \frac{(\mathbf{q}\_{\text{e,mean}} - \mathbf{q}\_{\text{e,cal}})^2}{(\mathbf{q}\_{\text{e,cal}} - \mathbf{q}\_{\text{e,mean}})^2 + (\mathbf{q}\_{\text{e,cal}} - \mathbf{q}\_{\text{e,exp}})^2} \tag{3}$$

where, qe,mean is the average value of experimental adsorption capacity (mmol·kg−<sup>1</sup> ), <sup>q</sup>e,cal is the equilibrium capacity from a model (mmol·kg−<sup>1</sup> ) and qe,exp is the experimental adsorption capacity.

#### **3. Results**

#### *3.1. Characterization of Cu<sup>0</sup> and Ag<sup>0</sup> ENPs and Soils*

The size, morphology, surface charge and the presence of functional groups on the surface of prepared ENPs were determined by TEM images, UV-Vis, ZP and FT-IR analyses. TEM images showed that both ENPs had spherical morphology (Figure S1a,b in Supplementary Materials). Cu<sup>0</sup> ENPs had a diameter between 8 and 29 nm, whereas Ag<sup>0</sup> ENPs showed a diameter between 7 and 27 nm (Figure S2a,b). The UV-Vis spectra of Cu<sup>0</sup> and Ag<sup>0</sup> ENPs showed an extended peak in the range of 342–512 and 337–474 nm, respectively (Figure S3). The FT-IR spectra for pure L-ascorbic acid showed a band

corresponding to a stretching vibration carbon–carbon double bond at 1674 cm−<sup>1</sup> and the peak of enol hydroxyl at 1322 cm−<sup>1</sup> (Figure S4a). After the reduction of Cu2+ and Ag<sup>+</sup> by L-ascorbic acid, the peaks disappeared and new peaks at 3481 cm−<sup>1</sup> and 1636 cm−<sup>1</sup> were observed (Figure S4b,c), which were associated with the conjugated hydroxyl and carbonyl groups, respectively. The pH of Cu<sup>0</sup> and Ag<sup>0</sup> ENPs suspension was 2.46 and 2.35, respectively. The IEP of Cu<sup>0</sup> ENPs was 2.7, whereas Ag<sup>0</sup> ENPs had a negatively charged surface in the studied pH range (Figure S5).

Physico-chemical properties of the soils untreated (T-OM) and treated with H2O<sup>2</sup> (R-OM) are shown in Table 3. The T-OM and R-OM were a typical Andisol exhibiting acidic characteristics showing pH values of 5.40 (strongly acidic) for T-OM and 6.20 (slightly acidic) for R-OM. Total P and OM in T-OM were 1.8 and 3.1 times higher as compared to R-OM, whereas the Al and Fe contents for R-OM were 1.2 and 1.4 times higher than T-OM. The SEM images revealed a decreased number of aggregates in R-OM compared to T-OM (Figure 1a,b). The contrasting OM content was also indicated in confocal images (Figure 1c,d) by a higher green fluorescence intensity for T-OM as compared to R-OM images. The IEP for T-OM was 3.2, while it was 5.7 for R-OM. Furthermore, the BET-specific surface area and pore volume increased 1.4 and 11.5 times for R-OM in comparison to T-OM. The FT-IR analysis (Figure S6) showed that R-OM had bands at 1003 cm−<sup>1</sup> and 913 cm−<sup>1</sup> corresponding to alumina and silica-rich allophane, respectively, while T-OM only showed the band at 1003 cm−<sup>1</sup> [29]. T-OM had more effective cation exchange capacity (ECEC) than R-OM (Table 3).

**Table 3.** Physico-chemical properties of soil with total organic matter (T-OM) and with partial removal of matter (R-OM).


\* ECEC: Effective cation exchange capacity.

#### *3.2. H2PO<sup>4</sup>* <sup>−</sup> *Adsorption on Soils with and without Cu<sup>0</sup> or Ag<sup>0</sup> ENPs* 3.2.1. Effect of pH Solution

Figure 2 shows the effect of the H2PO<sup>4</sup> <sup>−</sup> pH solution between 4.5–8.5 on H2PO<sup>4</sup> − adsorption on T-OM and R-OM soil samples in the absence and presence of ENPs. The H2PO<sup>4</sup> − adsorbed on T-OM decreased slightly with increasing pH without and with ENPs. When Cu<sup>0</sup> or Ag<sup>0</sup> ENPs content increased, the H2PO<sup>4</sup> − adsorption on T-OM was 1.4–1.8 times higher than without ENPs (Figure 2a,c). In addition, the H2PO<sup>4</sup> − adsorption on R-OM increased with increased Cu<sup>0</sup> ENPs doses, but with Ag<sup>0</sup> ENPs showed no changes (Figure 2b,d).

**Phosphate adsorbed (mmol·kg−1**

**)**

**a)**

**Figure 1.** SEM analysis to soil with (**a**) total organic matter (T-OM) and (**b**) partial removal of organic matter (R-OM) and confocal images to soil with (**c**) T-OM and (**d**) R-OM. <sup>r</sup><sup>ଶ</sup> = (qୣ,୫ୣୟ୬ − qୣ,ୡୟ୪) ଶ (qୣ,ୡୟ୪ − qୣ,୫ୣୟ୬) <sup>ଶ</sup> + (qୣ,ୡୟ୪ − qୣ,ୣ୶୮) ଶ

−

− **Figure 2.** Initial pH effect of the solution on the adsorption of H2PO<sup>4</sup> <sup>−</sup> in the presence of Cu<sup>0</sup> ENPs on soil with (**a**) total organic matter (T-OM) and (**b**) partial removal of organic matter (R-OM) and Ag<sup>0</sup> ENPs on soil with (**c**) T-OM and (**d**) R-OM.

**R-OM**

#### 3.2.2. Adsorption Kinetics

The kinetic studies are shown in Figure 3. We observed that increasing in contact time at pH 5.5 as well as in the presence of Cu<sup>0</sup> or Ag<sup>0</sup> ENPs there was a subsequent increase in the adsorption of H2PO<sup>4</sup> − in T-OM and R-OM soil samples. It was also shown that adsorption comprised a fast initial phase at 45 min, followed by a slower rate stage until equilibrium was reached at 360 min for T-OM and at 720 min for R-OM, whereas in the presence of ENPs for most systems it was reached at 720 min. Based on the Table 4, in the absence of ENPs after H2PO<sup>4</sup> − adsorption on T-OM and R-OM soil samples, the final pH (pH<sup>f</sup> ) showed an increase in relation to the initial pH (pH<sup>i</sup> ). A similar tendency was obtained with increasing the ENPs doses and the pH<sup>i</sup> and pH<sup>f</sup> values were lower compared with systems without ENPs.

**Figure 3.** Phosphate adsorption kinetics at pH 5.5 <sup>±</sup> 0.2 of the solution in the presence of Cu<sup>0</sup> ENPs on soil with (**a**) total organic matter (T-OM) and (**b**) partial removal of organic matter (R-OM) and Ag<sup>0</sup> ENPs on soil with (**c**) T-OM and (**d**) R-OM modelled by the Elovich model.

**Table 4.** pH changes associated to H2PO<sup>4</sup> <sup>−</sup> adsorption in the absence and presence of different doses of Cu<sup>0</sup> or Ag<sup>0</sup> ENPs and two levels of soil organic matter content (total organic matter, T-OM and partial removal of organic matter, R-OM). Experimental conditions: 6.47 mmol·L <sup>−</sup><sup>1</sup> H2PO<sup>4</sup> <sup>−</sup> solution at pH 5.5, 0.01 M KCl at 20 ± 2 ◦ C. Initial pH (pH<sup>i</sup> ) and final pH (pH<sup>f</sup> ) were measured after H2PO<sup>4</sup> <sup>−</sup> solution added to soil samples (time 0 min) and after H2PO<sup>4</sup> − adsorption (1440 min), respectively.


To determine the kinetic constants and understand the adsorption mechanisms, the experimental kinetics data were modeled by the pseudo-second-order (PSO) Elovich (Table 5) and pseudo-first-order (PFO) (Table S1) models. PSO and PFO models describe the kinetics of the adsorbate on an adsorbent based on chemical-adsorption and physical-adsorption, respectively, with respect to the adsorbent capacity [36]. On the other hand, the Elovich model describes the sorption of adsorbate onto a heterogeneous surface [38,39].

Based on the higher r<sup>2</sup> and the lower χ <sup>2</sup> values, the PSO model fitted to the adsorption kinetics data better than the PFO model. According to the PSO model, the amount of H2PO<sup>4</sup> <sup>−</sup> adsorbed at equilibrium (qe,cal) in T-OM and R-OM soil samples increased with ENPs contents and it was higher in R-OM than T-OM, except for 3 and 5% Ag<sup>0</sup> ENPs doses. The kinetic rate (k2) did not show a clear trend at low ENPs contents. However, it increased in T-OM with 5% Ag<sup>0</sup> ENPs and with 3 and 5% Cu<sup>0</sup> ENPs in R-OM as compared to the soils without ENPs. Similar behavior was observed for the initial adsorption rate (h) in the presence of ENPs leading to increases by adding 3% Cu<sup>0</sup> and 5% Ag<sup>0</sup> ENPs for T-OM and R-OM soil samples and 5% Cu<sup>0</sup> for R-OM and 3% Ag<sup>0</sup> ENPs for T-OM.

Experimental kinetic data at pH 5.5 in T-OM and R-OM soil samples without and with increasing Cu<sup>0</sup> or Ag<sup>0</sup> ENPs content also adequately fitted the Elovich model (r<sup>2</sup> = 0.927 <sup>−</sup> 0.998 and <sup>χ</sup> <sup>2</sup> = 9 <sup>−</sup> 279). This means that the H2PO<sup>4</sup> − adsorption happened on a heterogeneous substrate [38]. The initial rate (α) and the surface coverage (β) obtained from this model showed a similar tendency to h and k2, respectively, calculated from the PSO model. Thus, both PSO and Elovich models were capable of describing the kinetics of H2PO<sup>4</sup> − adsorption on volcanic soils. Similar results have been obtained by other researchers for an acid soil [40] and for adsorbents such as biochar [38] and chitosan [41].

#### 3.2.3. Adsorption Isotherms

The isotherm adsorptions at pH 5.5 (Figure 4) showed that the amount of H2PO<sup>4</sup> − adsorbed was slightly higher in R-OM than T-OM and H2PO<sup>4</sup> − adsorption increased with increasing ENPs contents. In general, all adsorption isotherm described curves type L [42]. This means that a high affinity of H2PO<sup>4</sup> − anions exist in both soils. In particular, in T-OM samples, the curve reached a strict asymptotic plateau, while in R-OM the curve did not reach it. This difference indicated that the number of adsorption sites in the T-OM sample for H2PO<sup>4</sup> − is limited; on the contrary, the R-OM sample had a greater number of adsorption sites for H2PO<sup>4</sup> <sup>−</sup>. At the same time, by increasing the Cu<sup>0</sup> or Ag<sup>0</sup> ENPs content, the curves showed a much less strict plateau for both soil samples, suggesting that the number of available adsorption sites for H2PO<sup>4</sup> − increased [42,43].

142


**Table 5.** Pseudo-second-order and Elovich parameters (± standard error) obtained from H2PO4<sup>−</sup> adsorption kinetics at pH 5.5 ± 0.2 for the soil with total organic matter (T-OM) and with partial removal of organic matter (R-OM) in the absence and presence of different doses of Cu<sup>0</sup> or Ag<sup>0</sup> ENPs.

**Figure 4.** Phosphate adsorption isotherms at pH 5.5 <sup>±</sup> 0.2 of the solutions in the presence of Cu<sup>0</sup> ENPs on soil with (**a**) total organic matter (T-OM) and (**b**) partial removal of organic matter (R-OM) and Ag<sup>0</sup> ENPs on soil with (**c**) T-OM and (**d**) R-OM modelled by the Langmuir and Freundlich models.

qୣ = q୫ୟ୶KCୣ 1 + KCୣ − − The adsorption isotherm data were fitted by Langmuir and Freundlich models (Table 6), which have been frequently used to explain H2PO<sup>4</sup> − adsorption on different soils [44,45].

−

− − − qୣ = KCୣ ଵൗ୬ − **Table 6.** Langmuir and Freundlich parameters (± standard error) obtained from H2PO<sup>4</sup> − adsorption isotherms at pH 5.5 ± 0.2 and desorption (%) for the soil with total organic matter (T-OM) and with partial removal of organic matter (R-OM) in the absence and presence of different doses of Cu<sup>0</sup> or Ag<sup>0</sup> ENPs.


The Freundlich model fitted the experimental data of T-OM and R-OM soil samples better than the Langmuir model (Table 6). However, in the presence of ENPs in T-OM and R-OM soil samples, the Langmuir model, except for R-OM—1% Cu<sup>0</sup> ENPs and T-OM—1% Ag<sup>0</sup> ENPs systems, showed a better fit to the experimental data (r<sup>2</sup> = 0.926 <sup>−</sup> 0.982 and χ <sup>2</sup> = 151 <sup>−</sup> 1000). According to the Langmuir model, the maximum H2PO<sup>4</sup> − adsorption capacity (qmax) in R-OM and T-OM soils increased with ENPs contents, and it was higher on R-OM than T-OM, except for 5% Ag<sup>0</sup> ENPs dose, in contrast to the affinity coefficient (KL).

#### 3.2.4. Desorption

The desorption (%) depends on the chemical nature and energy of the bonds between soil components and phosphate [46]. In this sense, after the soil samples were treated with double-distilled water repeatedly (three times), H2PO<sup>4</sup> − desorption was about 3.2 times higher from T-OM than R-OM (Table 6). In the presence of ENPs, the desorption from R-OM and T-OM soils decreased with increasing Cu<sup>0</sup> ENPs doses as well as from T-OM with 3 and 5% Ag<sup>0</sup> ENPs. In contrast, with increasing Ag<sup>0</sup> ENPs content, desorption from R-OM was greater than without ENPs.

#### **4. Discussion**

#### *4.1. Characterization of Cu<sup>0</sup> and Ag<sup>0</sup> ENPs and Soil Samples Studied*

The particle size average of Cu<sup>0</sup> (19 nm) and Ag<sup>0</sup> ENPs (17 nm) was low due to L-ascorbic acid coating, which provides colloidal stability to the nanoparticles by electrostatic repulsion. The stability effect of the L-ascorbic acid coating could be attributed to the presence of a polyhydroxyl structure on the surface of both nanoparticles [28]. This was supported by the high negative values of ZP, which is normally related to the negatively charged groups of the capping agents [28,47]. Similar results using organic molecules as reducing and capping agents for the preparation of ENPs have been reported previously [28,47–50].

The organic matter in volcanic soils is highly stabilized [51], whereby after repeated treatment with H2O2, only a part of the OM was removed from soil, accounting a 14.1% of OM (T-OM), obtaining a soil sample with 4.6% of OM (R-OM) (Table 3). The partial removal of OM significantly changed the aggregate structure of the soil because OM acts as a binding agent [52]. In addition, T-OM had more aggregates, a higher P concentration and an effective cation exchange capacity (ECEC) as compared to R-OM. In this sense, it is knowing that the functional groups of OM such as carboxyl, alcoholic hydroxyl, and phenolic hydroxyl contribute to the aggregation of soil particles, formation of humic (organic matter)-Al (Fe)-phosphate complexes and cations adsorption [52,53]. Likewise, R-OM samples had a higher IEP and BET-specific surface area than T-OM. This can be explained by the exposure of ≡Fe-OH and ≡Al-OH active sites from amorphous components of the soil, which decreased the negative charges of the surface and increased BET-specific surface area [34]. In general, allophane and ferrihydrite minerals can interact with negatively charged ENPs through attraction (Van der Waals) forces contributing to their retention in the soil [54].

#### *4.2. Ad- and Desorption of Phosphate on Soils*

The phosphate adsorption isotherms on T-OM and R-OM soil samples in the absence of ENPs were best fitted to the Freundlich model (Table 6), which reflected the heterogenic nature of soil components. The intensity of adsorption (n) and relative adsorption capacity (KF) for R-OM were higher than T-OM. The difference between K<sup>F</sup> and n for two soil samples may be due to the higher OM content of T-OM, since OM could block adsorption-specific sites leading to a lower availability of surface-reactive sites and weak interaction with H2PO<sup>4</sup> − [55]. The OM can act by preventing the irreversible retention of H2PO<sup>4</sup> − and increasing the nutrient recovery. We found that, after partial OM removal, the H2PO<sup>4</sup> − desorption from R-OM was lower than from T-OM (Table 6), indicating a strong interaction between the phosphate and mineral components of R-OM [15,16,23]. These

results are supported by the higher BET-specific surface area and lower negative surface charge of R-OM as compared to T-OM. Similar results were obtained by Zeng et al. [56] for H2PO<sup>4</sup> − desorption in volcanic soils exhibiting contrasting OM contents. However, these findings were in contrast to the results reported by Debicka et al. [57] by removing the OM from sandy soil resulted in decreases of K<sup>F</sup> and n values. Contrasting results could be attributed to the particularly components in each soil. According to the FAO-WRB soil classification, sandy soils such as Brunic Arenosols are mainly characterized by minerals such as hematite, goethite, and maghemite [57,58]. On the contrary, Santa Barbara soil is formed by minerals such as allophane (>50%), followed by 1–5% halloysite and vermiculite [59]. In this context, Parfitt [60] found that phosphate was adsorbed in the order hematite ~ goethite < ferrihydrite < allophane. Moreover, H2PO<sup>4</sup> − can be rapidly and strongly adsorbed on the most reactive aluminol (≡Al-OH) groups of the allophane by ligand exchange forming monodentate or/and binuclear complexes.

According to the PSO model, the higher H2PO<sup>4</sup> <sup>−</sup> adsorption (qe,cal) was in the R-OM as compared to T-OM (Table 5), which could due to the destruction of OM in T-OM, leading to a larger pore volume and BET-surface area. In addition, R-OM improved the accessibility to active sites for H2PO<sup>4</sup> − according to the higher values of α and h obtained for R-OM (Table 5) [57]. The h parameter can be associated to the chemical and/or hydrogen bonding interaction between H2PO<sup>4</sup> − and surface hydroxyls in soil samples at the initial adsorption process [16]. Moreover, considering the Elovich model and increase in pH<sup>f</sup> values after H2PO<sup>4</sup> <sup>−</sup> adsorption with respect to pH<sup>i</sup> (Table 4), we might suggest that H2PO<sup>4</sup> − adsorption in T-OM and R-OM soil samples was performed mainly through ligand exchange (chemi-adsorption) onto Fe/Al (hydr)oxides forming monodentate or bidentate complexes. The pH changes were consistent with the studies carried out by Vistoso et al. [24], who reported that H2PO<sup>4</sup> − was adsorbed through ligand exchange mechanism in volcanic soils with contrasting properties.

The H2PO<sup>4</sup> − adsorption on T-OM was pH-dependent in contrast to R-OM (Figure 2). In this context, the IEP of T-OM was 3.2 whereas it was 5.7 for R-OM. Therefore, in acidic pH H2PO<sup>4</sup> − solution the surface hydroxyl (–OH) groups in R-OM were more protonated than in T-OM, causing a favorable effect on electrostatic interaction and ligand exchange [61]. However, at alkaline pH H2PO<sup>4</sup> − solution, mainly for T-OM, there was a decrease in the ligand exchange and an increase in electrostatic repulsion due to deprotonation from soil superficial groups. Likewise, at a higher pH, the competition between OH− and H2PO<sup>4</sup> − on the T-OM surface would also reduce the H2PO<sup>4</sup> − adsorption [62].

#### *4.3. Ad- and Desorption of Phosphate on Soils in the Presence of Cu<sup>0</sup> or Ag<sup>0</sup> ENPs*

The increasing phosphate adsorption with increasing ENPs content in soils indicated that in the presence of ENPs, the number of adsorption sites increased. Although, there was a decrease in the initial adsorption rate (h) with 1% ENPs content, which implied that during the first few minutes ENPs compete with H2PO<sup>4</sup> − for the adsorption sites of the soil surface. Additionally, h strongly increased with 3 and 5% ENPs content, suggesting that ENPs also contributed to new adsorption sites for H2PO<sup>4</sup> − [63,64]. Accordingly, Duncan and Owens [63] indicated that CeO<sup>2</sup> ENPs can be adsorbed on soil adsorption sites before Pb(II) and Sun et al. [64] determined a similar trend for h with increasing carbon nanotubes (CNTs) content after studying the effects of CNTs with outer diameter of 25 nm and inner diameter of 5 nm on Cd(II) adsorption in sediments.

The adsorption isotherms of H2PO<sup>4</sup> <sup>−</sup> on T-OM and R-OM following Cu<sup>0</sup> or Ag<sup>0</sup> ENPs addition fitted to the Langmuir model (Table 6). Similarly, Sun et al. [64] found that in the presence of CNTs the isotherms for Cd(II) on sediment showed a better fit to the Langmuir than the Freundlich model; however due to the adsorption sites of sediments with CNTs are heterogeneous, they used the Freundlich to describe their results. Therefore, the fit of adsorption data to the Langmuir model in the presence of ENPs should be more studied.

Adsorption enhancement was larger through Cu<sup>0</sup> than Ag<sup>0</sup> ENPs. According to Afshinnia and Baalousha [65], the decrease in the zeta potential after H2PO<sup>4</sup> − adsorption

on T-OM and R-OM soil samples with ENPs could be associated with H2PO<sup>4</sup> − adsorption/complexation onto the ENPs surface (Figure S7). In this context, Niaura et al. [66] indicated that H2PO<sup>4</sup> <sup>−</sup> was adsorbed through monodentate surface coordination on Cu<sup>0</sup> ENPs, while on Ag<sup>0</sup> ENPs it was performed through hydrogen bonding [66,67]. Although both coated ENPs had a low rate of oxidation and dissolution [68], it was probable that these processes could be favored by an acidic soil pH as well as a consequence of the ionic exchange between H2PO<sup>4</sup> − and L-ascorbic acid on the surface of the ENPs, being similar to the mechanism observed for citric acid [50]. Under such conditions, Cu<sup>0</sup> could be oxidized to Cu2+ (E◦ Cu2+/Cu0 = 0.337 V) and the amount of phosphate adsorbed in T-OM and R-OM soil samples increased (Figure S8) because Cu2+ could be linked to H2PO<sup>4</sup> − and hydroxyl groups of OM via a cation bridge [69]. Furthermore, this could be attributed to the formation of complexes between Cu2+ and H2PO<sup>4</sup> <sup>−</sup> and the precipitate of Cu3(PO4)<sup>2</sup> (Ksp = 2.07 <sup>×</sup> <sup>10</sup>−33) [70]. Meanwhile, in the case that Ag<sup>+</sup> ions were released from Ag<sup>0</sup> ENPs into solution (E◦ Ag+/Ag0 = 0.799 V), the formation of AgCl precipitate was more favorable (Ksp = 1.77 <sup>×</sup> <sup>10</sup>−10) than a Ag3PO<sup>4</sup> formation (Ksp = 8.89 <sup>×</sup> <sup>10</sup>−17) [71,72].

On the other hand, the presence of L-ascorbic acid free in soil solution slightly competes with H2PO<sup>4</sup> − for available adsorption sites, decreasing H2PO<sup>4</sup> − adsorption on T-OM and R-OM soil samples (Figure S8). However, as a consequence of the addition of Cu<sup>0</sup> or Ag<sup>0</sup> ENPs suspensions to soil samples, the pH<sup>i</sup> values decreased, being less acidic in T-OM as compared to R-OM (Table 4), which was consistent with the buffering capacity of OM [73]. An acid pH can be associated with a decrease in the electrostatic repulsion between H2PO<sup>4</sup> − and the negatively charged surface of the organic matter (-COOH, -OH) due to a decrease in the number of deprotonated surface groups [74]. Furthermore, the protonation of surface hydroxyl groups of Fe/Al (hydr)oxides might be favored by acid pH values, promoting the H2PO<sup>4</sup> − adsorption through a ligand exchange [24,75,76]. In the same way, it has been reported that below 4.5 of pH values the mineral dissolution is favored, promoting the precipitation reactions between H2PO<sup>4</sup> − and cations in solution (Al3+ and Fe3+) [77], and to form H2PO<sup>4</sup> −-cation-organic matter complexes [53].

The increase of the H2PO<sup>4</sup> − adsorption at a low pH has been demonstrated on pillared bentonites [75], AgNPs-tea activated carbon [76], sediments [78] and in Andisol soils [24]. Future research should be addressed to corroborate whether, in the presence of both ENPs, one of these mechanisms was prevalent for H2PO<sup>4</sup> − adsorption on T-OM and R-OM soil samples, or whether several mechanisms acted together.

The H2PO<sup>4</sup> − adsorption in the presence of ENPs through chemical interactions onto a heterogeneous surface was indicated by the adequate fits of the kinetic data to the PSO and Elovich models (Table 5). In addition, the desorption behavior supported the adsorption mechanisms proposed in the presence of ENPs. With Cu<sup>0</sup> ENPs, the desorption of H2PO<sup>4</sup> − from T-OM and R-OM soil samples was smaller than Ag<sup>0</sup> ENPs. These results can be supported by a chemisorption-like interaction between H2PO<sup>4</sup> <sup>−</sup> and Cu<sup>0</sup> ENPs. Similarly, desorption studies of U(VI) on the soil in the presence of nano-crystalline goethite showed that U(VI) was more resistant to released due to an increase in the inner-sphere complexes on the soil surface [79]. In addition, Elkhatib et al. [80] revealed that sorption of Hg(II) on arid soils in the presence of water treatment residual nanoparticles occurred mainly through inner-sphere complexes, which enhanced Hg immobilization in the arid soils. The high desorption of H2PO<sup>4</sup> <sup>−</sup> in R-OM following Ag<sup>0</sup> ENPs addition needs further investigation. One possible explanation for this is that the Ag<sup>0</sup> ENPs were attached to the potential H2PO<sup>4</sup> − adsorption sites, such as allophane and Fe oxides, leading to a blocking effect for H2PO<sup>4</sup> − on this soil with lower levels of OM. Then, the H2PO<sup>4</sup> − physi-adsorbed (through hydrogen bonding) on the surface of the attached Ag<sup>0</sup> ENPs was more desorbable.

#### **5. Conclusions**

Our study demonstrated that the phosphate adsorption process in the presence of ENPs was dependent on the amount of ENPs and soil organic matter content. The addition of Cu<sup>0</sup> caused a higher increase in phophate adsorpion on T-OM and R-OM as compared to the Ag<sup>0</sup> ENPs. The Elovich and pseudo-second-order (PSO) models correctly described the kinetic adsorption of phosphate on T-OM and R-OM soil samples without and with ENPs.

The phosphate adsorption with both ENPs was better described by the Langmuir isotherm model than the Freundlich model. According to the Langmuir model, by increasing the ENPs content from 0 to 5%, the maximum adsorption capacity (qmax) of H2PO<sup>4</sup> − for T-OM ranged from 216.1 to 316.4 mmol·kg−<sup>1</sup> following the Cu<sup>0</sup> ENPs addition and to 332.8 mmol·kg−<sup>1</sup> using Ag<sup>0</sup> ENPs. Meanwhile, with the increase from 0 to 5% of ENPs, the qmax of H2PO<sup>4</sup> <sup>−</sup> for R-OM ranged from 224.7 to 440.2 mmol·kg−<sup>1</sup> with Cu<sup>0</sup> ENPs and to 301.4 mmol·kg−<sup>1</sup> with Ag<sup>0</sup> ENPs. Phosphate desorption in T-OM and R-OM soils following Cu<sup>0</sup> ENPs addition was lower than Ag<sup>0</sup> ENPs. In the future, more attention should be pointed globally to management agriculture practices based on nanotechnology, because the incorporation of ENPs into the soil have the potential to reduce the already limited crop phosphorus availability.

**Supplementary Materials:** The following are available online at https://www.mdpi.com/article/10 .3390/min11040373/s1, Figure S1: TEM images L-ascorbic acid-stabilized (a) Cu<sup>0</sup> and (b) Ag<sup>0</sup> ENPs, Figure S2: Histograms with the corresponding particle size distribution for L-ascorbic acid-stabilized (a) Cu<sup>0</sup> and (b) Ag<sup>0</sup> ENPs, Figure S3: UV-Vis absorption spectra for L-ascorbic acid-stabilized Cu<sup>0</sup> and Ag<sup>0</sup> ENPs, Figure S4: FT-IR spectra of (a) Pure L-ascorbic acid, (b) L-ascorbic acid-stabilized Cu<sup>0</sup> ENPs and (c) L-ascorbic acid-stabilized Ag<sup>0</sup> ENPs, Figure S5: Zeta potential of L-ascorbic acidstabilized Cu<sup>0</sup> and Ag<sup>0</sup> ENPs in 0.01 M KCl, Figure S6: FT-IR spectrum for soil samples with (a) total organic matter (T-OM) and (b) partial removal of organic matter (R-OM), Figure S7: Zeta potential curves in the presence of 9.71 mmol·L <sup>−</sup><sup>1</sup> H2PO<sup>4</sup> <sup>−</sup> and 5% Cu<sup>0</sup> or 5% Ag<sup>0</sup> ENPs at constant ionic strength (0.01 M KCl) for soil with (a) total organic matter (R-OM) and (b) partial removal of organic matter (R-OM), Figure S8: Adsorption isotherm curves of H2PO<sup>4</sup> − on (a) total organic matter (T-OM) and /9b) partial removal of organic matter (R-OM) in the presence of 3% L-ascorbic acid and Cu2+ and Ag<sup>+</sup> . Reaction conditions: Concentrations from 0.016 to 9.71 mmol·L <sup>−</sup><sup>1</sup> H2PO<sup>4</sup> − on 0.5 g soil in 0.01 M KCl at 20 ± 2 ◦ C and pH 5.5, Table S1: Pseudo-first-order parameters (± standard error) obtained from H2PO<sup>4</sup> <sup>−</sup> adsorption kinetics in the absence and presence of different doses of Cu<sup>0</sup> and Ag<sup>0</sup> ENPs at pH 5.5 <sup>±</sup> 0.2 for soil with total organic matter (T-OM) and with partial removal of organic matter (R-OM).

**Author Contributions:** Conceptualization, E.K., M.d.L.L.M. and A.J.; methodology, J.S.-H.; software, J.S.-H.; validation, M.d.L.L.M., E.K., N.A.-M. and R.B.; formal analysis, J.S.-H.; P.P.-G. and A.J.; investigation, J.S.-H. and P.P.-G.; resources, M.d.L.L.M.; data curation, J.S.-H.; writing—original draft preparation, J.S.-H., N.A.-M. and E.K.; writing—review and editing, E.K., R.B., N.A.-M. and A.J.; visualization, J.S.-H. and R.B.; supervision, M.d.L.L.M. and N.A.-M.; project administration, M.d.L.L.M.; funding acquisition, M.d.L.L.M. and J.S.-H. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was funded by the Fondo Nacional de Desarrollo Científico y Tecnológico (FONDECYT) projects Nº 1181050 and 1191018 and by the Agencia Nacional de Investigación y Desarrollo (ANID) Ph.D. scholarships Nº 21171685.

**Data Availability Statement:** Data are contained within this article.

**Acknowledgments:** Jonathan Suazo-Hernández acknowledges to Daniela Vergara, the FONDE-CYT project Nº 3210228, the Technological Bioresource Nucleus (BIOREN-UFRO) and the Soil and Plant Laboratory.

**Conflicts of Interest:** The authors declare no conflict of interests.

#### **References**

