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

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 Cu⁰ or Ag⁰ 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 H₂PO₄⁻ 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 H₂PO₄⁻ from T-OM and R-OM soil samples following ENPs addition were better fitted by the Langmuir model than the Freundlich model. When the Cu⁰ or Ag⁰ ENPs doses were increased, the pH value decreased and H₂PO₄⁻ adsorption increased on T-OM and R-OM. The H₂PO₄⁻ desorption (%) was lower with Cu⁰ ENPs than Ag⁰ ENPs. Overall, the incorporation of ENPs into Andisols generated an increase in P retention, which may affect agricultural crop production.

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 (Cu⁰) and silver (Ag⁰), due to their antibacterial properties. Cu⁰ ENPs are added to biocides, electronics, paints, cosmetics, agrochemicals, ceramics, and film [1,3,4], whereas Ag⁰ 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 to improve our current understanding of environmental risks, fate, transformations and aggregation behaviors of metallic ENPs [7].

The geochemistry of metallic Cu⁰ and Ag⁰ ENPs in soils is complex, due to their chemical transformation between Cu⁰, Cu⁺ and Cu²⁺ as well as between Ag⁰ and Ag⁺, 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, Cu⁰ 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 Ag⁰ 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 Cu⁰ and Ag⁰, 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 [10,11,12]. 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 (Al₂O₃ 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⁻¹ [14]. However, P availability for plant growth is limited because it can form inner-sphere complexes by ligand exchange with surface -OH and -OH₂⁺ groups of soil components like ferrihydrite, imogolite, allophane, and Al(Fe)-humus complexes [15,16,17]. 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 Cu⁰ or Ag⁰ 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 Cu⁰ and Ag⁰ ENPs

CuCl₂·2H₂O and AgNO₃ were used for the formation of Cu⁰, and Ag⁰ ENPs, respectively, and L-ascorbic acid was added as a reducing and capping agent [28]. Cu⁰ ENPs (or Ag⁰ ENPs) was synthesized by mixing 10.0 mmol·L⁻¹ CuCl₂·2H₂O (or 10.0 mmol·L⁻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 Cu⁰ ENPs (or Ag⁰ 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 H₂O₂ 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 Ag⁰ and Cu⁰ ENPs

The synthetized Cu⁰ and Ag⁰ 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 Cu⁰ and Ag⁰ 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⁻¹ resolution and the operating range was 600 cm⁻¹ to 4000 cm⁻¹ 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/carbon-coated 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 N₂ 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 H₂PO₄⁻ on T-OM and R-OM soil samples in the absence and presence of 5% Cu⁰ or Ag⁰ 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 H₂PO₄⁻) on T-OM and R-OM soil samples in the absence and presence of 0, 1, 3, and 5% Cu⁰ or Ag⁰ ENPs doses (% w/w). Cu⁰ or Ag⁰ ENPs doses were added to 0.5 g (dry weight) of soil samples in polyethylene tubes and mixed with 20 mL H₂PO₄⁻ solution. The adsorbed amounts of H₂PO₄⁻ (q_t, mmol·kg⁻¹) were determined as the difference between initial concentration and final concentration of H₂PO₄⁻ in the solution (Equation (1)).

$${\mathrm{q}}_{\mathrm{t}}=\frac{\left({\mathrm{C}}_0-{ \mathrm{C}}_{\mathrm{t}}\right)\mathrm{V}}{\mathrm{w}}$$

(1)

where, C_o is the initial concentrations of H₂PO₄⁻ and C_t is the concentrations of H₂PO₄⁻ 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 H₂PO₄⁻ onto T-OM and R-OM soil samples, stock solutions of 6.47 mmol·L⁻¹ of H₂PO₄⁻ 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 H₂PO₄⁻ 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⁻¹ of H₂PO₄⁻ 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 H₂PO₄⁻ was determined in solution. Furthermore, the initial pH (pH_i) and the final pH (pH_f) were measured after H₂PO₄⁻ solution was added to soil samples (time 0 min) and after H₂PO₄⁻ adsorption (1440 min), respectively.

Adsorption isotherms were obtained by varying the initial H₂PO₄⁻ concentrations from 0.016 to 9.71 mmol·L⁻¹ and were initially adjusted to pH 5.5 ± 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⁺) or L-ascorbic acid on H₂PO₄⁻ adsorption onto T-OM and R-OM soil samples, adsorption isotherms were made in the presence of 3% Cu²⁺ or Ag⁺ 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 H₂PO₄⁻ 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

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

Kinetic Equations	Expression Formula	Parameters	References
Pseudo-first-order (PFO)	${\mathrm{q}}_{\mathrm{t}}={\mathrm{q}}_{\mathrm{e}}\left(1-{\mathrm{e}}^{-{\mathrm{k}}_1 \mathrm{t}}\right)$	qt = amount of anion adsorbed at any time (mmol·kg−1). qe = amount of anion adsorbed at equilibrium (mmol·kg−1). k1 = PFO rate constant (min−1). k2 = PSO rate constant (kg·mmol−1·min−1). t = time (min)	[35,36]
Pseudo-second-order (PSO) *	${\mathrm{q}}_{\mathrm{t}}=\frac{{\mathrm{k}}_2{\mathrm{q}}_{\mathrm{e}}^2\mathrm{t}}{1+{\mathrm{k}}_2{\mathrm{q}}_{\mathrm{e}}\mathrm{t}}$
Elovich	${\mathrm{q}}_{\mathrm{t}}=\frac{1}{\mathsf{\beta }}\ln \left(1+\mathsf{\alpha }\mathsf{\beta }\mathrm{t}\right)$	$\mathsf{\alpha }$ = initial rate constant (mmol·kg−·min−1). $\mathsf{\beta }$ = number of sites available for the sorption and desorption constant (mmol·kg−1).

* From PSO initial adsorption rate (h), can be calculated by multiplying k₂${\mathrm{q}}_{\mathrm{t}}^2$ (mmol·kg⁻¹·min⁻¹).

and

Table 2. The isotherm models used for the description of phosphate adsorption.

Isotherm Equations	Expression Formula	Parameters	References
Langmuir	${\mathrm{q}}_{\mathrm{e}}=\frac{{\mathrm{q}}_{\mathrm{m}}{\mathrm{K}}_{\mathrm{L}}{\mathrm{C}}_{\mathrm{e}}}{1+{\mathrm{K}}_{\mathrm{L}}{\mathrm{C}}_{\mathrm{e}}}$	qe = amount of adsorbed anion per unit mass of the adsorbent at equilibrium (mmol·kg−1). qmax = maximum adsorption capacity (mmol·kg−1). Ce = concentration of anion at equilibrium in the solution (mmol·L−1). KL = constant related to the affinity (L·mmol−1). KF = freundlich adsorption coefficient ((mmol·kg−1) (L·kg−1)1/n). n = adsorption intensity (1 < n < 10).	[34,37]
Freundlich	${\mathrm{q}}_{\mathrm{e}}={\mathrm{K}}_{\mathrm{F}}{\mathrm{C}}_{\mathrm{e}}^{\raisebox{1ex}{$1$}\!\left/ \!\raisebox{-1ex}{$\mathrm{n}$}\right.}$

, respectively.

The data were evaluated through the Chi-square (χ²), adding the coefficient of determination (r²) (Equations (2) and (3)). The lowest χ² and highest r² values were used as the best fit [37]. The statistical analysis of the adsorption data was conducted using Origin Pro 8.0.

$${ \mathsf{\chi }}^2=\sum \frac{{({\mathrm{q}}_{\mathrm{e},\exp }-{\mathrm{q}}_{\mathrm{e},\mathrm{cal}})}^2}{{\mathrm{q}}_{\mathrm{e},\mathrm{cal}}}$$

(2)

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

(3)

where, q_e,mean is the average value of experimental adsorption capacity (mmol·kg⁻¹), q_e,cal is the equilibrium capacity from a model (mmol·kg⁻¹) and q_e,exp is the experimental adsorption capacity.

3. Results

3.1. Characterization of Cu⁰ and Ag⁰ 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). Cu⁰ ENPs had a diameter between 8 and 29 nm, whereas Ag⁰ ENPs showed a diameter between 7 and 27 nm (Figure S2a,b). The UV-Vis spectra of Cu⁰ and Ag⁰ 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⁻¹ and the peak of enol hydroxyl at 1322 cm⁻¹ (Figure S4a). After the reduction of Cu²⁺ and Ag⁺ by L-ascorbic acid, the peaks disappeared and new peaks at 3481 cm⁻¹ and 1636 cm⁻¹ were observed (Figure S4b,c), which were associated with the conjugated hydroxyl and carbonyl groups, respectively. The pH of Cu⁰ and Ag⁰ ENPs suspension was 2.46 and 2.35, respectively. The IEP of Cu⁰ ENPs was 2.7, whereas Ag⁰ 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 H₂O₂ (R-OM) are shown in

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

Parameter	T-OM	R-OM
pH (H2O)	5.4 ± 0.0	6.2 ± 0.0
Total P (mg·kg−1)	1766.4 ± 27.0	996.6 ± 15.0
Si (%)	15.9 ± 3.5	16.3 ± 2.9
Al (%)	11.7 ± 1.1	14.1 ± 1.5
Fe (%)	7.5 ± 0.6	10.5 ± 1.2
OM (%)	14.1 ± 0.1	4.6 ± 0.1
ECEC (cmol(+) kg−1) *	8.8 ± 0.5	7.8 ± 0.0
Isoelectric point	3.2	5.7
BET- specific surface area (m2·g−1)	17.4	24.4
Average pore volume (cm3·g−1)	0.002	0.023
Average pore diameter (A)	10.7	10.4

* ECEC: Effective cation exchange capacity.

. 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⁻¹ and 913 cm⁻¹ corresponding to alumina and silica-rich allophane, respectively, while T-OM only showed the band at 1003 cm⁻¹ [29]. T-OM had more effective cation exchange capacity (ECEC) than R-OM (Table 3).

3.2. H₂PO₄⁻ Adsorption on Soils with and without Cu⁰ or Ag⁰ ENPs

3.2.1. Effect of pH Solution

Figure 2 shows the effect of the H₂PO₄⁻ pH solution between 4.5–8.5 on H₂PO₄⁻ adsorption on T-OM and R-OM soil samples in the absence and presence of ENPs. The H₂PO₄⁻ adsorbed on T-OM decreased slightly with increasing pH without and with ENPs. When Cu⁰ or Ag⁰ ENPs content increased, the H₂PO₄⁻ adsorption on T-OM was 1.4–1.8 times higher than without ENPs (Figure 2a,c). In addition, the H₂PO₄⁻ adsorption on R-OM increased with increased Cu⁰ ENPs doses, but with Ag⁰ ENPs showed no changes (Figure 2b,d).

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 Cu⁰ or Ag⁰ ENPs there was a subsequent increase in the adsorption of H₂PO₄⁻ 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. pH changes associated to H₂PO₄⁻ adsorption in the absence and presence of different doses of Cu⁰ or Ag⁰ 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⁻¹ H₂PO₄⁻ solution at pH 5.5, 0.01 M KCl at 20 ± 2 ^°C. Initial pH (pH_i) and final pH (pH_f) were measured after H₂PO₄⁻ solution added to soil samples (time 0 min) and after H₂PO₄⁻ adsorption (1440 min), respectively.

			Cu0 ENPs						Ag0 ENPs
	0%		1%		3%		5%		1%		3%		5%
	pHi	pHf	pHi	pHf	pHi	pHf	pHi	pHf	pHi	pHf	pHi	pHf	pHi	pHf
T-OM	5.4 ± 0.1	5.7 ± 0.0	5.0 ± 0.1	5.4 ± 0.0	4.3 ± 0.1	4.7 ± 0.0	4.0 ± 0.0	4.2 ± 0.0	5.0 ± 0.1	5.6 ± 0.2	4.5 ± 0.1	4.9 ± 0.0	4.2 ± 0.1	4.7 ± 0.0
R-OM	5.2 ± 0.0	5.6 ± 0.1	4.7 ± 0.1	5.3 ± 0.1	4.2 ± 0.0	4.4 ± 0.0	3.8 ± 0.0	3.9 ± 0.0	4.6 ± 0.1	5.0 ± 0.1	4.2 ± 0.0	4.4 ± 0.0	3.9 ± 0.0	4.0 ± 0.0

, in the absence of ENPs after H₂PO₄⁻ adsorption on T-OM and R-OM soil samples, the final pH (pH_f) showed an increase in relation to the initial pH (pH_i). A similar tendency was obtained with increasing the ENPs doses and the pH_i and pH_f values were lower compared with systems without ENPs.

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. Pseudo-second-order and Elovich parameters (± standard error) obtained from H₂PO₄⁻ 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 Cu⁰ or Ag⁰ ENPs.

			Pseudo-Second-Order					Elovich
	ENPs Doses (%)	qe,exp (mmol·kg−1)	qe,cal (mmol·kg−1)	k2 (×10−4 kg·mmol−1· min−1)	h (mmol·kg−1·min−1)	r2	χ2	α (×103 mmol·kg−1·min−1)	β (×10−2 kg·mmol−1)	r2	χ2
T-OM	0	177.9 ± 2.4	164.3 ± 5.1	16.6 ± 1.9	45.6 ± 0.0	0.940	157	1.9 ± 0.3	6.4 ± 0.4	0.985	40
R-OM		200.4 ± 4.2	174.4 ± 6.1	16.9 ± 2.5	50.5 ± 0.0	0.924	226	2.0 ± 0.3	5.9 ± 0.3	0.993	21
			Cu0
T-OM	1	203.1 ± 5.1	177.4 ± 6.8	14.0 ± 4.0	44.0 ± 0.0	0.915	271	1.0 ± 0.2	5.4 ± 0.3	0.986	43
R-OM		227.1 ± 2.8	206.5 ± 4.5	10.9 ± 1.7	46.5 ± 0.0	0.973	114	1.3 ± 0.4	4.8 ± 0.4	0.972	118
T-OM	3	234.2 ± 3.9	202.4 ± 8.4	14.0 ± 4.4	57.4 ± 0.0	0.900	430	1.7 ± 0.4	4.9 ± 0.3	0.989	45
R-OM		234.5 ± 1.5	221.4 ± 3.5	26.4 ± 4.0	129.4 ± 0.0	0.980	86	590.5 ± 15.2	7.3 ± 0.5	0.970	123
T-OM	5	251.1 ± 7.2	214.9 ± 9.0	9.4 ± 2.8	43.4 ± 0.0	0.910	454	0.7 ± 0.1	4.2 ± 0.0	0.998	9
R-OM		244.0 ± 4.3	232.9 ± 4.5	28.8 ± 5.6	156.2 ± 0.0	0.970	139	1518.7 ± 101.6	7.4 ± 0.8	0.980	91
			Ag0
T-OM	1	220.9 ± 4.8	183.4 ± 6.5	9.0 ± 3.5	30.1 ± 0.0	0.941	280	0.4 ± 0.1	4.6 ± 0.2	0.988	43
R-OM		215.9 ± 6.0	185.8 ± 7.7	8.1 ± 2.3	28.1 ± 0.0	0.933	306	0.3 ± 0.2	4.6 ± 0.5	0.927	279
T-OM	3	228.1 ± 4.1	201.7 ± 7.9	13.6 ± 4.0	55.3 ± 0.0	0.927	373	1.7 ± 0.3	5.0. ± 0.2	0.995	21
R-OM		224.0 ± 5.2	187.9 ± 8.7	7.9 ± 2.5	27.9 ± 0.0	0.905	386	0.2 ± 0.1	4.2 ± 0.3	0.968	132
T-OM	5	239.6 ± 3.6	213.2 ± 5.6	18.1 ± 3.9	82.3 ± 0.0	0.952	186	11.2 ± 1.0	6.6 ± 0.4	0.974	112
R-OM		232.0 ± 4.5	194.2 ± 8.8	13.7 ± 4.7	51.7 ± 0.0	0.942	263	1.1 ± 0.2	4.9 ± 0.4	0.979	84

) 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 r² and the lower χ² values, the PSO model fitted to the adsorption kinetics data better than the PFO model. According to the PSO model, the amount of H₂PO₄⁻ adsorbed at equilibrium (q_e,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% Ag⁰ ENPs doses. The kinetic rate (k₂) did not show a clear trend at low ENPs contents. However, it increased in T-OM with 5% Ag⁰ ENPs and with 3 and 5% Cu⁰ 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% Cu⁰ and 5% Ag⁰ ENPs for T-OM and R-OM soil samples and 5% Cu⁰ for R-OM and 3% Ag⁰ ENPs for T-OM.

Experimental kinetic data at pH 5.5 in T-OM and R-OM soil samples without and with increasing Cu⁰ or Ag⁰ ENPs content also adequately fitted the Elovich model (r² = 0.927 − 0.998 and χ² = 9 − 279). This means that the H₂PO₄⁻ 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 H₂PO₄⁻ 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 H₂PO₄⁻ adsorbed was slightly higher in R-OM than T-OM and H₂PO₄⁻ adsorption increased with increasing ENPs contents. In general, all adsorption isotherm described curves type L [42]. This means that a high affinity of H₂PO₄⁻ 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 H₂PO₄⁻ is limited; on the contrary, the R-OM sample had a greater number of adsorption sites for H₂PO₄⁻. At the same time, by increasing the Cu⁰ or Ag⁰ ENPs content, the curves showed a much less strict plateau for both soil samples, suggesting that the number of available adsorption sites for H₂PO₄⁻ increased [42,43].

The adsorption isotherm data were fitted by Langmuir and Freundlich models (

Table 6. Langmuir and Freundlich parameters (± standard error) obtained from H₂PO₄⁻ 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 Cu⁰ or Ag⁰ ENPs.

			Langmuir				Freundlich				Desorption
	ENPs Doses (%)	qe,exp (mmol·kg−1)	KL (L·mmol−1)	qmax (mmol·kg−1)	r2	χ2	KF ((mmol·kg−1) (L·kg−1)1/n)	n	r2	χ2	(%)
T-OM	0	213.6 ± 0.6	3.6 ± 1.5	216.1 ± 20.0	0.934	396	143.5 ± 5.0	3.4 ± 0.3	0.982	111	31.6
R-OM		256.4 ± 1.4	20.1 ± 7.3	224.7 ± 15.6	0.944	437	181.0 ± 9.0	3.6 ± 0.4	0.961	299	9.7
			Cu0
T-OM	1	225.0 ± 1.1	7.8 ± 3.2	216.2 ± 16.6	0.943	388	152.1 ± 10.6	3.0 ± 0.4	0.928	488	24.0
R-OM		301.0 ± 0.7	7.3 ± 2.8	280.4 ± 23.2	0.942	600	216.3 ± 10.2	2.6 ± 0.2	0.966	352	9.5
T-OM	3	262.0 ± 0.2	7.2 ± 1.6	255.4 ± 14.3	0.968	274	175.0 ± 13.6	2.6 ± 0.3	0.923	738	10.7
R-OM		318.4 ± 1.4	3.6 ± 0.8	382.7 ± 31.5	0.963	434	266.8 ± 25.8	1.9 ± 0.3	0.870	1550	6.4
T-OM	5	299.2 ± 0.5	5.4 ± 1.5	316.4 ± 26.1	0.945	604	216.0 ± 22.0	2.3 ± 0.4	0.854	1590	9.5
R-OM		343.3 ± 2.4	3.0 ± 0.9	440.2 ± 55.3	0.926	1000	298.2 ± 33.5	1.9 ± 0.3	0.836	2208	4.5
			Ag0
T-OM	1	213.8 ± 3.0	8.1 ± 2.7	200.5 ± 13.3	0.953	290	143.4 ± 7.0	3.3 ± 0.3	0.964	219	31.5
R-OM		261.3 ± 0.3	5.2 ± 0.8	264.0 ± 11.4	0.982	151	171.3 ± 13.3	2.5 ± 0.3	0.921	724	16.8
T-OM	3	236.2 ± 1.9	3.2 ± 0.8	249.9 ± 19.2	0.948	361	139.9 ± 14.2	2.3 ± 0.4	0.870	946	29.4
R-OM		292.7 ± 2.1	7.3 ± 1.5	281.4 ± 15.5	0.970	302	202.9 ± 13.0	2.5 ± 0.3	0.940	604	21.0
T-OM	5	284.1 ± 2.8	2.2 ± 0.6	332.8 ± 30.8	0.951	500	178.6 ± 18.1	2.0. ± 0.3	0.871	1327	22.2
R-OM		297.5 ± 1.4	4.7 ± 1.2	301.4 ± 21.4	0.958	434	201.5 ± 15.6	2.2 ± 0.3	0.924	881	22.1

), which have been frequently used to explain H₂PO₄⁻ adsorption on different soils [44,45].

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% Cu⁰ ENPs and T-OM—1% Ag⁰ ENPs systems, showed a better fit to the experimental data (r² = 0.926 − 0.982 and χ² = 151 − 1000). According to the Langmuir model, the maximum H₂PO₄⁻ adsorption capacity (q_max) in R-OM and T-OM soils increased with ENPs contents, and it was higher on R-OM than T-OM, except for 5% Ag⁰ ENPs dose, in contrast to the affinity coefficient (K_L).

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), H₂PO₄⁻ 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 Cu⁰ ENPs doses as well as from T-OM with 3 and 5% Ag⁰ ENPs. In contrast, with increasing Ag⁰ ENPs content, desorption from R-OM was greater than without ENPs.

4. Discussion

4.1. Characterization of Cu⁰ and Ag⁰ ENPs and Soil Samples Studied

The particle size average of Cu⁰ (19 nm) and Ag⁰ 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 [28,47,48,49,50].

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 (K_F) for R-OM were higher than T-OM. The difference between K_F 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 H₂PO₄⁻ [55]. The OM can act by preventing the irreversible retention of H₂PO₄⁻ and increasing the nutrient recovery. We found that, after partial OM removal, the H₂PO₄⁻ 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 H₂PO₄⁻ 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_F 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, H₂PO₄⁻ 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 H₂PO₄⁻ adsorption (q_e,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 H₂PO₄⁻ 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 H₂PO₄⁻ and surface hydroxyls in soil samples at the initial adsorption process [16]. Moreover, considering the Elovich model and increase in pH_f values after H₂PO₄⁻ adsorption with respect to pH_i (Table 4), we might suggest that H₂PO₄⁻ 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 H₂PO₄⁻ was adsorbed through ligand exchange mechanism in volcanic soils with contrasting properties.

The H₂PO₄⁻ 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 H₂PO₄⁻ 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 H₂PO₄⁻ 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 H₂PO₄⁻ on the T-OM surface would also reduce the H₂PO₄⁻ adsorption [62].

4.3. Ad- and Desorption of Phosphate on Soils in the Presence of Cu⁰ or Ag⁰ 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 H₂PO₄⁻ 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 H₂PO₄⁻ [63,64]. Accordingly, Duncan and Owens [63] indicated that CeO₂ 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 H₂PO₄⁻ on T-OM and R-OM following Cu⁰ or Ag⁰ 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 Cu⁰ than Ag⁰ ENPs. According to Afshinnia and Baalousha [65], the decrease in the zeta potential after H₂PO₄⁻ adsorption on T-OM and R-OM soil samples with ENPs could be associated with H₂PO₄⁻ adsorption/complexation onto the ENPs surface (Figure S7). In this context, Niaura et al. [66] indicated that H₂PO₄⁻ was adsorbed through monodentate surface coordination on Cu⁰ ENPs, while on Ag⁰ 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 H₂PO₄⁻ and L-ascorbic acid on the surface of the ENPs, being similar to the mechanism observed for citric acid [50]. Under such conditions, Cu⁰ 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 H₂PO₄⁻ and hydroxyl groups of OM via a cation bridge [69]. Furthermore, this could be attributed to the formation of complexes between Cu²⁺ and H₂PO₄⁻ and the precipitate of Cu₃(PO₄)₂ (Ksp = 2.07 × 10⁻³³) [70]. Meanwhile, in the case that Ag⁺ ions were released from Ag⁰ ENPs into solution (E°_Ag+/Ag0 = 0.799 V), the formation of AgCl precipitate was more favorable (Ksp = 1.77 × 10⁻¹⁰) than a Ag₃PO₄ formation (Ksp = 8.89 × 10⁻¹⁷) [71,72].

On the other hand, the presence of L-ascorbic acid free in soil solution slightly competes with H₂PO₄⁻ for available adsorption sites, decreasing H₂PO₄⁻ adsorption on T-OM and R-OM soil samples (Figure S8). However, as a consequence of the addition of Cu⁰ or Ag⁰ ENPs suspensions to soil samples, the pH_i 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 H₂PO₄⁻ 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 H₂PO₄⁻ 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 H₂PO₄⁻ and cations in solution (Al³⁺ and Fe³⁺) [77], and to form H₂PO₄⁻-cation-organic matter complexes [53].

The increase of the H₂PO₄⁻ 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 H₂PO₄⁻ adsorption on T-OM and R-OM soil samples, or whether several mechanisms acted together.

The H₂PO₄⁻ 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 Cu⁰ ENPs, the desorption of H₂PO₄⁻ from T-OM and R-OM soil samples was smaller than Ag⁰ ENPs. These results can be supported by a chemisorption-like interaction between H₂PO₄⁻ and Cu⁰ 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 H₂PO₄⁻ in R-OM following Ag⁰ ENPs addition needs further investigation. One possible explanation for this is that the Ag⁰ ENPs were attached to the potential H₂PO₄⁻ adsorption sites, such as allophane and Fe oxides, leading to a blocking effect for H₂PO₄⁻ on this soil with lower levels of OM. Then, the H₂PO₄⁻ physi-adsorbed (through hydrogen bonding) on the surface of the attached Ag⁰ 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 Cu⁰ caused a higher increase in phophate adsorpion on T-OM and R-OM as compared to the Ag⁰ 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 (q_max) of H₂PO₄⁻ for T-OM ranged from 216.1 to 316.4 mmol·kg⁻¹ following the Cu⁰ ENPs addition and to 332.8 mmol·kg⁻¹ using Ag⁰ ENPs. Meanwhile, with the increase from 0 to 5% of ENPs, the q_max of H₂PO₄⁻ for R-OM ranged from 224.7 to 440.2 mmol·kg⁻¹ with Cu⁰ ENPs and to 301.4 mmol·kg⁻¹ with Ag⁰ ENPs. Phosphate desorption in T-OM and R-OM soils following Cu⁰ ENPs addition was lower than Ag⁰ 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.