Chemistry of soil-type dependent soil matrices and its influence on behaviors of pharmaceutical compounds (PCs) in soils

Abstract

Behaviors of pharmaceutical compounds (PCs) in soil are usually determined by experimental extrapolation of results from separate constitutes to the soil, or from a special soil to other regional soil conditions. However, such extrapolation is problematic due to variations in soil clay mineral and organic matter (OM) compositions with soil types, which dominate the interaction mechanisms of PCs in soil. It is essential to review current literature to enhance our understanding of the soil-type dependent surface chemistry of soil matrices and the environmental behavior of PCs in different soil types. Major types of soils occur globally in parallel to the latitudinal or altitudinal zonation due to regional climate conditions with distinct clay mineral and OM compositions. The soil-type dependent surface chemistry results in variations in retention, distribution, transport, and transformation PCs in soil. The mixture of PCs of different classes usually exhibited enhanced sorption due to the cooperative multilayer sorption on soil constituents, and that of the same class often caused differential adsorption capacity compared to the sorption from single compound due to competitive sorption. PCs preferentially adsorb to a soil component, or to a special soil type, and exhibit notably soil-type dependent sorption affinity, mobility, and dissipation. The soil-dependent surface chemistry of soil is critical to predict the persistence and bioavailability of PCs in soil. In the future, more detailed studies of influence of individual soil factor on the behaviors of PCs and especially the practical field site investigation are required to better understand the sorption, transport, transformation, and ecotoxicology of PCs in typical soil types.

1. Introduction

Pharmaceutical compounds (PCs) have been used extensively for treating bacterial infections in both humans and animals, and even as additives into the animal feed in sub-therapeutic dose in order to prevent diseases and promote their growth [1,2]. Due to wastewater irrigation and soil application with animal manures or biosolids, PCs directly enter to agroecosystems and are present widely in surface water, groundwater, soils, and sediments [[3], [4], [5], [6], [7], [8], [9]]. In sludge matrices, the commonly quantified compounds reached 305 residues with concentrations up to several mg/kg in dry weight, in which antibiotics, stimulants, and antidepressants were detected in the highest concentrations up to 232 mg/kg, and diuretics, anticoagulants or anti-anxieties in the lowest concentrations reaching to 686 μg/kg [10]. Recently, PC pollutants have become one of the new environmental threats to the continued capacity of soil to function as a sustainable vital living ecosystem for plants, animals, and humans due to the development and occurrence of antibiotic-resistance genes and antibiotic-resistant bacteria in soil [[11], [12], [13]]. Adsorption and degradation are two main environmental passways of PCs in soil, which control the transport, fate, bioavailability, and ecotoxicological risks of PCs in soil environments [[14], [15], [16], [17]]. The highly mobile compounds are prone to leach to the groundwater or migrate to drainage water and surface water and, thus, cause environmental water contamination, while those strongly adsorbed PCs may accumulate in the upper soil layer and pose ecotoxicological risks in soil [9,18,19].

Sorption of PCs on the soil materials affects the mobility of PCs and even their degradation and bioavailability in the soil environment [6,16,20,21]. The sorption behaviors of PCs in soils are not only a function of their physicochemical properties, but are also particularly controlled by soil properties including soil pH, organic matter (OM) content, cation exchange capacity (CEC), and presence of cations [15,[22], [23], [24]]. Different clay mineral species and soil OM components display variable charges due to the presence of distinct active groups on structures of the materials [[25], [26], [27]]. Sorption of PCs onto clays and soil OM is generally governed by the surface charge properties of the materials as well as the complexation between the sorbates and the specific surface functional groups [28,29]. PC sorption in soil is usually determined by experimental extrapolation, and the obtained sorption affinity K_d (L/kg) or K_f (the Freundlich constant, mmol¹⁻ⁿLⁿ/kg) is commonly used to assess the sorption behaviors of PC pollutants in soil [16]. The prediction of environmental behaviors of PCs requires knowledge of combined contributions from various components of soils and their sorption behaviors of different species [30,31]. However, extrapolation of results from separate constitutes to the soil, or from a special soil to other regional soil conditions is problematic due to variations in soil properties under the climate [32]. Most studies of fate and transport of PCs in soils in the recent decade focused on distribution and occurrence of PCs in various sites via case studies, or via modeling their sorption behaviors of separate clay species, oxide minerals, and soil OM [11,12,33]. Instead, those obtained in the soil and pedogenic soil type are still very limited [11,23,[34], [35], [36]].

Soil particle-size class only characterizes the grain-size composition of the whole soil. It does not include soil OM, nor types of clay mineral species, while soil taxonomy clearly follows the different pedogenesis conditions and, thus, characterizes the soil substrates and the overall soil physicochemical properties, such as mineral species of clays and Fe-oxides, organic compounds of soil OM, pH, and coexisting ions in soils [26,37]. Different clay mineral species and soil OM components between different types of soils will lead to significant variations in surface chemistry and the resultant interaction mechanisms of PCs with solid soil matrices, which finally controls the behaviors of PCs in soils [16,38,39]. As such, this review is intended to address the present literature concerning the specific surface chemistry of major solid soil matrices of various pedogenic soil types, as well as influence of soil property on persistence, transport, and degradation of PCs in soils, with an aim to better understand what will happen to PCs in the specific soil conditions and the intrinsic interaction mechanisms affecting the environmental behavior of PCs in different soil types, and to better establish their predictable persistence and bioavailability in various soil environments.

2. Methodology

Comprehensive literature of the relevant topic was searched in Web of Science and ScienceDirect, by using the selected keywords such as soil, soil type, adsorption, degradation, clay mineral, organic matter, surface chemistry, antibiotics, and pharmaceuticals based on certain combinations. Such search allowed collection of literature from the entire world of science community on this topic, which was collected in the database and published in years since 1975. References were also collected according to the relevance to the subject and their relevant literature was also accessed for significance of information. In addition, special keywords such as chemistry, kaolinite, goethite, and organic matter were also used for searching the more specific subjects such as the surface chemistry of kaolinite, goethite, and organic matter.

3. The soil-type dependent surface chemistry of soil matrices

3.1. Chacarteristic clay mineral phases and OM in major soils

Soil classification is based on soil chemical, physical, and biological properties such as the contents of OM, Fe(Al)-oxides, silicate clays, pH, and salts, and the groups of soils are mostly similar in their genesis [40,41]. The development of soil and formation of clay minerals are strongly dependent on climate, principally the moisture and temperature conditions. Thus, most types of soils show a global distribution that is parallel to the latitudinal or altitudinal zonation, controlled by annual temperature and precipitation, although other factors such as lithography, landforms, vegetation and fauna, and time also play important roles on the formation of soils and clay phases (Fig. 1, Fig. 2).

Fig. 1.

A scheme showing the main factors (parent material, climate, and biological activity) influencing the formation of soil [40,42].

Fig. 2.

A generalized map showing the global soil type distribution. Modified from Ref. [43].

In combination with previous results of nature and properties of soils, the characteristic clay minerals of different soil types under different climate conditions are summarized in Table 1.

Table 1.

Characteristic clay minerals of typical soil types formed under different climate conditions.

Climate [41]		Soil types [40]	Characteristic clay phases [44]	pH [40]
Tropical to subtropical	Humid	Ferralsols; Lixisols; Acrisols	Kaolinite (or halloysite), Fe-, Al-oxyhydroxides	3–5
	Contrasted wet-dry	Lixisols; Acrisols; Luvisols, Retisols; Vertisol	Kaolinite, smectite, illite, vermiculite, Fe-, Al-oxyhydroxides	4–6
	Semi-arid to arid	Vertisol; Gypsisols; Calcisols	Illite, vermiculite, mixed-layers, chlorite	6–8
Temperate rainfall (variable temperature); monsoon climate		Andosols	Halloysite, kaolinite/smectite, Fe-hydroxy-oxides	4–5
Temperate	Warm, or humid	Podzols; Umbrisols, Luvisols	Illite, vermiculite, I/S minerals, Fe-oxides	3–4
	Cool, dry	Solonetz; Vertisols; Phaeozems; Fluvisols	Illite, illitic I/S minerals	6–7
	Cool	Phaeozems; Umbrisols; Fluvisols	Vermiculite and hydroxy-interlayered (HI) minerals	6–9
Arid to semiarid		Solonchaks; Gypsisols; Calcisols	Sepiolite, palygorskite, I/S, carbonates (smectite, illite, chlorite, kaolinite)	7–9
Dry continental zone		Phaeozems, Chernozems, Kastanozems	Illite, smectite, vermiculite, chlorite	6–7
Very cold	Glacial and peri-glacial climate	Cryosols, Leptosols	Illite, chlorite	5–7

The formation and preservation of soil OM are intimately dependent on their pedogenic process, which result in the presence of distinct mineral composition, chemical and physical properties of soils together with the unique combination with the co-occurrence of OM [27]. Soil OM is generated from plants or microbial molecules or their monomers. The plant OM includes primarily neutral sugar polymers (celluloses and hemicelluloses), phenolic compound polymers (lignins), polymers containing charged sugars (pectins), proteins, polyphenolic compounds, tannins, complex sugars, and organic acids (released from the roots), and lipids (decomposed from waxes, cuticles, bark, and root cortexes) [45]. The microbially-derived OM is generally composed of the same molecules except for celluloses and lignins, which consist mainly of polysaccharides, lipids, proteins, amino-saccharides, nucleic acids, and a very diverse range of metabolites in comparison with plant OM [45,46].

The ¹³C NMR signals of OM from bulk soils are significantly similar, indicating the general soil OM components with typical molecular structures. OM in bulk soil is composed of mainly O/N-alkyl structure C (45 %), alkyl C (25 %), aromatic C (20 %), and carboxyl and amide C (10 %) [47]. However, the composition and property of soil OM vary with different soils [27]. The O/N-alkyl part of soil OM includes dominantly polysaccharides and certain amounts of cellulose, hemicellulose, proteins, and peptides. The alkyl part of soil OM consists mainly of methylenic C in long-chain aliphatic compounds (fatty acids, lipids, cutin acids). The aromatic (aryl) part of soil OM represents aromatic compounds derived from lignin, tannins, and charcoal of plant origin. While the carboxyl and amide parts of soil OM contain mainly uronic acids and amino acids with most of proteinaceous material [46].

3.2. Surface chemistry of soil clay minerals

Soil consists of dominantly solid phases (minerals and organic matter) and even fluid phases (solutions and/or gas) and biota, which are significantly complex because of their high degree of heterogeneity [40]. To totally define the nature of the soil system, all its components should be known including the chemical composition and the relative quantities of various phase components [27,48,49], since PC sorption onto clay minerals surfaces is dependent largely on the surface charge properties of clay species and the surface functional groups [16,29]. The crystal structures of clay minerals are remarkably polymeric sandwiches of tetrahedral and octahedral sheets with various manners of coherently stacking of the lattice structure in the a-b plane. Different groups of clay minerals are subdivided based on difference in combination of tetrahedral and octahedral sheets and stacking coherences, and various clay members are derived from different types and amounts of substitutions in the brucite-like layer and the tetrahedral and octahedral positions of the 2:1 mica-like basic unit [50].

Kaolinite group has the 1:1 layer type structure with the basic unit consisting of one tetrahedral sheet combining with an octahedral sheet and has the layer charge usually <0.01 eq/mol, while illite, vermiculite, and smectite groups have the 2:1 layer type structure with the basic unit composing of two tetrahedral sheets fusing to an octahedral sheet, having the generally decreasing layer charges from 1.4 to 2.0 to 1.2–1.8, and to 0.5–1.2 eq/mol (half unit-cell) respectively [51]. The layer charge for cations is derived from the isomorphic replacement of tetrahedral Si⁴⁺ by Al³⁺, which is required to balance the negative structural charge by cation sorption on or near the siloxane layer and is believed to account for the CEC of smectite [52]. Chlorite consists of a brucite-like or gibbsite-like sheet sandwiched between illite-like trilayers. The mica-like face carries a permanent negative charge, since in the tetrahedral sheet about 25 % of the tetrahedral sites are occupied by Al³⁺ atoms, and the charge deficiency is compensated by the positively charged brucite-like or gibbsite-like sheet [50,53].

The 1:1 (kaolinite) and 2:1:1 (chlorite) layer-type clay minerals consist of heterogeneous surfaces. The basal surface atoms of these species have one side of oxygen atoms of the Si–O tetrahedra sheets and another side of OH groups from the Al–O or Mg–O octahedra on the Al- or Mg-side of the layer, while the 2:1 layer type clay minerals (smectite and illite) have oxygen atoms of the Si–O tetrahedra sheets [50,52]. Thus, there are two surfaces for these clay minerals, i.e., the Si–O–Si basal surface and the edge surface with SiOH and AlOH groups. The basal surface is pH-independent for the presence of permanent negative charges, while the edge surface is pH-dependent due to protonation/deprotonation of the SiOH and AlOH groups [54].

For soil clay minerals, the Si–O tetrahedral side of 1:1 kaolin group displays the neutral siloxane surface, and the oxygen atoms on the neutral siloxane surface can only act as relatively weak electron donors (Lewis bases), but not strongly interacting with water molecules [52,54]. The 2:1 layer-type smectite displays intermediate to strong surface hydrophilicity, depending mainly on its surface charge, exchangeable cation, and ion adsorption characteristics [25,36]. Smectite has relatively low surface charge density, and usually adsorbs weakly hydrated exchangeable inorganic cations on the basal surface due to charge compensation. Therefore, between the sites of exchangeable inorganic cations, the non-polar regions exhibit still hydrophobic character, and adsorption of non-polar organic molecules onto these surface sites is dominantly controlled by the surface charge density of the clay species and the nature of the hydrated exchangeable cations (effective ionic radius, hydration energy, hydrolysis constant) [55,56].

The interlayer regions of 2:1 smectite and vermiculite species exhibit an overall hydrophilic nature due to the occurrence of hydrated cations, and hydration of these cations increase the spacing between the stacked layers and thus cause swelling of clays. Computational chemistry simulations showed that accumulation of hydrated layers in the interlayer space forms quasi-crystal structures, which produce a stepwise increased spacing in the interlayer [57,58]. The hydrated cations are coordinated to their nearest-neighbor water molecules, and the water molecules interact with the interlayer siloxane surfaces by hydrogen bonds. Charge on the edges is thought to be due to the protonation/deprotonation of terminal hydroxyl groups and, thus, depends on the solution pH. The point of zero charge (PZC) of smectite is at a pH of ∼8.0. At pH below the PZC, the cation exchange occurs simultaneously at layer permanent-charge sites and protonation of edge sites (>A1OH groups), and when the pH reaches to the PZC, deprotonation of surface hydroxyl groups (>SiOH and >A1OH) causes the exposure of these edge sites of smectie and results in an overall negative charge [36,59].

Illite is a dioctahedral, 2:1 layer-type clay species. It contains varying degrees of lattice substitution of Al³⁺ for Si⁴⁺ in the tetrahedral sheet and, thus, tends to attract cations to the surface to compensate the negative surface charges [50]. As a result, cations in solution can form either surface complexation at the hydroxyl group sites or replace the charge-balanced surface cations on the planar sites of illite [60]. Change in surface charge with pH for the illite edge is attributed to deprotonation or protonation of the amphoteric SiOH and AlOH groups on the edge surfaces. The surface charge of edge surface is more strongly dependent on pH relative to its basal surface, since the permanent charge on the surface is derived from isomorphic substitution [50]. The PZC of edge surface of illite occurs at pH of 2.5. The edge surface of illite becomes more negative with increasing pH, when the pH increases to 10.0, the AlOH group starts to deprotonate and the edge surface continues to be negatively charged [61].

The hydroxyl groups on the basal surfaces of the 2:1:1 chlorite and the 1:1 kaolin species, as well as goethite, gibbsite, and other oxides such as gibbsite and brucite, are bonded with metal atoms having complete coordination environment and have only minimal chemical reactivity [53]. However, they are different for hydroxyl groups originating from the edge surfaces, i.e., the edge planes (110) and (010), and broken or defect sites of clay minerals as well as the related oxides. These OH groups are under-coordinated terminal OH groups, which usually carry either a positive or a negative charge on the basis of the type of metal atom and the solution pH, and are more reactive than the charge-neutral OH group on basal surface [52,62]. Terminal OH groups of these mineral surfaces can potentially chemisorb certain types of ions and show high sorption affinity for the ions. The PZC value of kaolinite is usually at pH of 5–6, which is closely related to the Al/Si ratio of kaolinite composition [63]. In case of pH values higher than the PZC value, the kaolinite surface will have a net negative charge and tend to interact with cationic species, whereas at pH values lower than the PZC, the mineral surface will have a net positive charge and favor adsorption of anionic species [64].

3.3. Surface properties of soil organic matter and Fe-, Al-(hydr)oxides

Soil OM consists generally of a complex mixture of intact and partly degraded biopolymers (e.g., proteins, carbohydrates, aliphatic biopolymers, and lignin) and their decomposed biomolecular fragments (dissolved OM and humic substances), which is the major source of variable charge in OM-bearing soils due to the presence of different functional groups on their molecular structures [28] (Table 2, Table 3). Most functional groups of soil OM contain oxygen and the functional groups are either acidic (negatively charged via proton dissociation) or neutral (unionized but probably polar). Although the nature of OM is very heterogeneous and difficult to characterize, the chemistry of soil OM depends directly on the quantities and types of organic functional groups and structural components of the molecules [65]. The acidic functional groups include dominantly the carboxyl in the aliphatic and aromatic molecule structures. The neutral functional groups and structures of soil OM containing oxygen include carbonyl (esters, aldehydes, anhydrides, and ketones), aliphatic-OH (alcoholic-OH), aliphatic and aromatic ethers, and quinones (aromatic carbonyl). The basic organic functional groups and compounds (positively charged via proton complexation) are mainly the nitrogen-containing groups (Table 2).

Table 2.

Chemical properties of soil clay minerals, Fe-, Al- (hydr)oxides, and organic matter.

Species	Layer charge (eq/mol; half unit-cell)	Surface charge density (C/m2)	CEC (cmol/kg)	PZC	Surface functional groups	References
Clay minerals
Smectite	0.5-1.2 [51]	0.12 [66]	70–150 [67]	∼8.0 [52]	> X– (structural charged sites on basal and interlayer surfaces) > SiOH; > AlOH (silanol groups at edge)	[52,59,68]
Vermiculite	1.2-1.8 [51]	0.248 [69]	100-150 [67]	8.5 [68]
Illite	1.4-2.0 [51]	0.49 [66]	10–40 [67]	2.5 [60]	> XNa+ (cation exchange on the planar sites) > SiOH; > AlOH (surface complexation reactions)	[60]
Chlorite	1.2–2.6 [51]	0.56 [66]	10–40 [67]	<3 [70]	> Si2O (cation exchange on the illite-like face) > Al2–OH (brucite-like or gibbsite-like sheet) > SiOH; > Al(Mg)–OH (silanol groups at edge)	[70]
Kaolinite	<0.01 [51]	0.064 [66]	3–15 [67]	5-6 [54]	> SiOH; > AlOH (edge surface) > Al2–OH (gibbsite-like sheet)	[54]
(Hydr)oxides		Stern layer capacitance (F/m2)
Goethite		1.7 [71]		8.5 [72]	> FeOH (surface groups) > Fe3OH (Fe at double octahedra chains or at junction of two double Fe octahedra chains)	[72]
Hematite		1.3 [73]		8.1 [74]	> Fe–O (Lewis acid Fe surface sites)	[74,75]
Ferrihydrite		1.17 [71]		8.2 [76]	> FeOH (surface groups) > Fe3O (Fe octahedra at corner of two faces)	[76]
Gibbsite		1.4 [71]		5.7 [77]	> AlOH (edge surface groups)	[77]
Organic matter
Acidic functional groups					R−COOH → R−COO− + H+ (carboxyl, R: aliphatic or aromatic carbon structure) Ar−OH → Ar−O− + H+ (phenolic-OH, Ar: aromatic structure)	[65]
Neutral functional groups and structures					Carbonyl (in esters, ketones, aldehydes, and anhydrides) Aliphatic-OH (alcoholic-OH) Aliphatic and aromatic ether Quinone (aromatic carbonyl)	[65]
Basic compounds by proton complexation					R–NH20 + H+ → R−NH3+ (R: amide, imine, aromatic ring nitrogen)	[65]

Notes: >X^– represents the permanent structural charged sites, and the ion exchange reaction is described by: >XH + Na⁺ >XNa + H⁺.

Table 3.

Examples of frequently studied pharmaceutical compounds and their basic properties.

Compound	CAS	Formula	Solubility	Solubility in water	pKa
Sulfathiazole	72-14-0	C9H9N3O2S2	Slightly soluble in ethanol (0.5 mg/ml)	Insoluble (<1 mg/ml)	7.2
Sulfamethoxazole	723-46-6	C10H11N3O3S	Soluble in ethanol or acetone	Insoluble (0.1 mg/ml)	5.60
Sulfameter	651-06-9	C11H12N4O3S	Sparingly soluble in methanol	Insoluble (0.47 mg/ml)	7.02
Sulfapyridine	144-83-2	C11H13N3O2S	Slightly soluble in methanol	Insoluble (<1 mg/ml)	8.48
Sulfadimethoxine	122-11-2	C12H14N3O2S	Soluble in NH4OH (50 mg/ml)	Insoluble	5.94
Sulfamethazine	57-68-1	C12H14N4O2S	Soluble in dilute acid or base	Soluble (1.5 mg/ml)	7.4/2.65
Sulfadiazine	68-35-9	C10H10N4O2S	Soluble in dilute acid or base	Soluble (13 mg/100 ml)	2.21
Sulfamerazine	127-79-7	C11H12N4O2S	Slightly soluble in DMSO, methanol	Slightly soluble (0.2 mg/ml)	2.29
Sulfamonomethoxine	1220-83-3	C11H12N4O3S	Soluble in ethanol or methanol	Soluble (10 mg/ml)	5.94
Sulfachloropyridazine	80-32-0	C10H9N4CLO2S	Soluble in DMSO (56 mg/ml), in 0.5 M NaOH (50 mg/ml)	Insoluble	6.10
Sulfanilamide	63-74-1	C6H8N2O2S	Soluble in acetone	Soluble (7.5 mg/ml)	10.6
Norfloxacin	70458-96-7	C16H18FN3O3	Soluble in dilute acid or base	Slightly soluble (0.28 mg/ml)	6.34/8.75
Ciprofloxacin	85721-33-1	C17H18FN3O3	Insoluble in DMSO (<1 mg/ml)	Soluble (30 mg/ml)	6.43/8.68
Enrofloxacin	93106-60-6	C19H22FN3O3	Soluble in dilute acid or base; Very slightly soluble in DMSO (1 mg/ml)	Slightly soluble	6.43
Ofloxacin	82419-36-1	C18H20FN3O4	Soluble in acetic acid, in 1 M NaOH (50 mg/ml)	Slightly soluble	5.19
Levofloxacin	100986-85-4	C18H20FN3O4	Soluble in glacial acetic acid or dichloromethane	Slightly soluble	5.19
Nalidixic acid	389-08-2	C12H12N2O3	Soluble in DMSO, in chloroform (20 mg/ml), in 0.5 M NaOH (50 mg/ml)	Insoluble (0.1 mg/ml)	6.11
Tetracycline	60-54-8	C22H24N2O8	Soluble in ethanol (＞20 mg/ml)	Soluble (1.7 mg/ml)	8.3/10.2
Oxytetracycline	79-57-2	C22H24N2O9	Soluble in alcohol or ether	Insoluble (0.2 mg/ml)	3.27/7.32/9.11
Chlortetracycline	57-62-5	C22H23ClN2O8	Soluble in DMSO, ethanol, methanol	Slightly soluble (0.63 mg/ml)	3.3
Tylosin	1401-69-0	C46H77NO17	Soluble in DMSO (25 mg/ml), in ethanol (30 mg/ml)	Soluble (50 mg/ml)	7.73
Erythromycin	114-07-8	C33H67NO13	Soluble in ethanol	Soluble (2 mg/ml)	8.8
Clarithromycin	81103-11-9	C41H76N2O15	Soluble in DMSO	Insoluble (0.1 mg/ml)	8.99
Amoxicillin	26787-78-0	C16H25N3O8S	/	Soluble (4 mg/ml)	2.4
Chloramphenicol	56-75-7	C11H12Cl2N2O5	Soluble in alcohol (5–20 mg/ml)	Insoluble	11.03
Thiamphenicol	15318-45-3	C12H15Cl2NO5S	Soluble in DMSO (66 mg/ml), in ethanol (50 mg/ml)	Insoluble	11.05
Clindamycin	18323-44-9	C18H33ClN2O5	Soluble in DMSO (≥15.3 mg/ml)	Slightly soluble （0.3 mg/ml）	7.6
Trimethoprim	738-70-5	C14H18N4O3	Soluble in DMSO (40 mg/ml)	Insoluble (<1 mg/ml)	6.6
Monensin	17090-79-8	C36H62O11	Soluble in organic solvent	Slightly soluble	4.26
Atenolol	29122-68-7	C14H22N2O3	Soluble in DMSO (53 mg/ml), in ethanol (53 mg/ml)	Slightly soluble (0.3 mg/ml)	9.6
Metoprolol	37350-58-6	C15H25NO3	soluble in methanol (>500 mg/ml)	Soluble (>1000 mg/ml)	/
Nalidixic acid	389-08-2	C12H12N2O3	Soluble in chloroform (20 mg/ml)	Insoluble (0.1 mg/ml)	6.11
Carbamazepine	298-46-4	C15H12N2O	Soluble in DMSO (47 mg/ml), in ethanol (18 mg/ml)	Slightly soluble (<1 mg/ml)	13.94
Bisphenol A	80-05-7	C15H16O	Soluble in methanol, acetone	Insoluble	10.29
17β-Estradiol	50-28-2	C18H24O2	Soluble in acetone	Insoluble	10.71
azithromycin	83905-01-5	C38H72N2O12	Soluble in anhydrous ethanol and in methylene chloride	Insoluble	8.74
Citalopram	59729-33-8	C20H21FN2O	Soluble in DMSO	Slightly soluble	9.38
Fexofenadine	83799-24-0	C32H39NO4	Soluble in DMSO (12 mg/ml)	Insoluble	/
Irbesartan	138402-11-6	C25H28N6O	Soluble in DMSO (>25 mg/ml)	Insoluble	4.16

Notes: Data from "Chemical Book" (https://www.chemicalbook.com/ProductIndex_EN.aspx; Solubility of <1 mg/ml refers to the product slightly soluble or insoluble.

It is generally considered that OM contains polar functional groups (e.g., carboxyl, phenolic OH and alcoholic OH), which can usually form stable complexes with cations. Some OM functions are also pH dependent. For examples, the carboxyl group is mainly in the form of COOH at pH < 4 but in COO⁻ at pH > 5, and the OM has non-polar groups (e.g. aromatic, aliphatic) that can interact with hydrophobic phases [78]. Therefore, organic constituent in soil can strongly influence the speciation and mobility of cations and organic pollutants [45,78]. All soils have generally moderate to weak acidic chemical properties with pH values from 3 to 7, except for Vertisols having weak acidic to weak alkaline conditions with pH of 6–8 [27]. Organic matter can also aggregate into labile supramolecular structures through hydrogen bonding, cation bridging, and hydrophobic interaction. The intrinsic nature of OM is the total characters of their constituent biomolecular fragments, bound cations, and the higher-order aggregates [79].

Soil OM is intimately associated with soil minerals and plays important roles in maintenance of aggregate stability, water retention capacity, and buffering of soil solution pH. The high specific surface areas of clay minerals and Fe-, Al-(hydr)oxides in soils make them dominating the solid–water interface, and their surface charge and reactive surface functional groups facilitate their effective sequestration of inorganic ion species and OM [48,49]. The coating OM on soil mineral surfaces can significantly alter the interfacial chemistry relative to that of the underlying mineral phases, although they are only present at trace levels [80]. The mineral-stabilized OM can change the interfacial behaviors of associated minerals and produce hydrodynamically rough and chemically heterogeneous surfaces in soil solutions, which can retard the advective/diffusive transport processes and provide heterogeneous surfaces for precipitation of substances from solutions [[81], [82], [83], [84]]. Most of soil OM (>90 %) occur in close association with clay minerals, and the amphiphilic property of OM is considered to be fundamental to the formation of mineral–organic complexes [48,85].

Quartz particles in soils usually exhibit only weak bonding affinity for OM due to very low surface charge. The soil mineral phases have positively to negatively charged surfaces, although the negatively charged OM compounds are generally repelled from negative soil surfaces. Adsorption and retention of OM on clay mineral surfaces occur when polyvalent cations are present on the exchange complexes due to the widely isomorphic substitution in the clay mineral structure, because polyvalent cations can simultaneously bind to the negatively charged soil surfaces (e.g., clay minerals) and the acidic functional groups (e.g., COO⁻) of the OM and, hence, act as a bridge between the two negatively charged sites [27]. The 1:1-type kaolinite-associated soil OM is mainly polysaccharide products, while the 2:1-type smectite-associated OM is dominantly aromatic compounds [86]. The organic head functional groups determine largely the relative affinity of organic molecules for kaolinite surfaces, polar compounds adsorb preferentially on Al- or Mg-octahedra side, and alkanes adsorb preferentially on siloxane surface [87]. Decane and protonated decanoic acid molecules exhibit stronger affinity for the siloxane surface, and decanoate anions preferentially adsorb onto the octahedra side via an anion exchange mechanism. In particular, decanamine can adsorb onto the siloxane surface via van der Waals interaction, but also onto the octahedral side via hydrogen bonding and water bridging interactions, and protonated decanamine can still adsorb onto both the siloxane and octahedra surfaces respectively through the ionic bridging mechanism [88].

4. Sorption and retention of PCs by soil

4.1. Influence of chemical structure of PCs on sorption to soil

Adsorption is the important process affecting the ultimate fate and bioavailability of PCs in soil matrices [9,17]. The accumulation of PCs in soil may cause long exposure and increase their toxic effects, and elimination of PCs in soil is the result of its fate and degradation pathways, which is closely dependent on soil properties including species and amount of clay minerals, OM content, ion exchange capacity, and pH [15]. Soil OM, metal oxides, and clay minerals can act as the cation exchangers in soil environment. If the PC has a charged hydrophilic head group and a large hydrophobic moiety, the neutral part of the compound can sorb to a nonpolar phase and concurrently the head group can bond with the polar phase [16]. Variability in sorption characteristic exerts greatly on the degradability and persistence of the respective compounds in soil, which needs to be evaluated in relation to the stability and persistence of a particular PC in the soil matrices, even the effect of environmental temperature [89]. The surface chemical properties of soils (surface charge and zeta potential) and the mineralogy of soil clay fraction are notably different between soil types developed in various geographical locations (Table 1, Table 2).

PCs occur usually as negative, neutral, zwitterionic, and positively charged species depending on the nature of the PCs and the specific soil condition (Table 3). Thus, PCs can interact with broken-bond edge surface, with basal oxygen and hydroxyl planes of 1:1 type clay minerals, with interlayer surface of swelling 2:1-type clay minerals, and with natural OM in soil [9,52,[90], [91], [92]]. The amphoteric compounds may exist as anions, cations, and/or zwitterions according to the pH of the soil solution. Interactions between PCs and soil components involve a number of sorption mechanisms including cation exchange, electrostatic interactions, hydrogen bonding, van der Waals forces, cation bridging, hydrophobic interactions, and surface complexes, and different sorption mechanisms of PCs to soil matrices display significant difference in their binding intensity [16,37]. In general, strongly basic and amphoteric PCs can form complexes to varying degrees with smectite, illite, and kaolinite clays, whereas acidic and neutral compounds can only sorb to smectite in small amounts. Accordingly, strong basic PCs almost do not release from smectite, vermiculite, or illite clays, while certain amounts of strong basic PCs may release from kaolinite and, in particular, all amphoteric PCs can release from all common soil clay minerals [9,92].

As most soil surfaces are negatively charged, cationic PCs are prone to react with soil matrices due to the attraction of a cation to a negatively charged site on the soil particle surface, which results in the exchange of one cation for another at the reactive site and bind strongly to the negatively charged exchangeable sites of soil components [16,54]. Although anionic species are generally repulsed from the negative soil surfaces, anion and zwitterion PCs can complex cations from soil solution and, thus, adsorb to soil matrix, since the cations can bind to the negative soil surfaces through electrostatic attraction or surface-bridging complexation [30]. Cation bridging interactions form usually due to the formation of an inner-sphere complex on the surface of soil matrix, in which cationic molecule is adsorbed by the negatively charged surface of solid phases (clay minerals, Fe, Al-oxides, and OM). The anionic PCs can bind to the positively charged inner-sphere complex by cation bridging interaction and result in the adsorption of anions [9,93]. In addition, their polar functional groups can interact with soil matrices by polar interactions such as hydrogen bonding, hence, neutral PCs can adsorb to the soil matrices via hydrophobic interactions [16]. Electrostatic interactions occur readily between soil particles with negative charge and positively charged PCs, which plays a critical role for retention of PCs in ionic state [38,84].

Interactions between PCs and the soil matrices depend on the polarity and functional group characters of the compounds and substrates [91,94,95]. Soil OM contains typically hydrophobic domains, which act as important sorption sites for PCs and is the most important sorbent for those nonionic and less polar organic pollutants in soils [96,97]. Zwitterionic PCs display varying sorption and desorption interactions in different soil environments including various positions of a soil profile due to their different chemical conditions [37]. In addition, soil OM such as humic acid (HA) molecules can ionize under different pH conditions, the carboxyl groups of HA molecules are dissociated and negatively charged at basic pH environment, whereas they are neutral or low negatively charged at acid pH [30,98]. Hydrogen-bonding interaction is regarded as an important sorption process of PCs on soil matrices, which dominantly occurs between aromatic carboxyl or hydroxyl groups of soil OM and O atoms in the carbonyl group of PCs, and it is also formed between the hydroxyl (-OH) groups of clay minerals and PCs [9,34,99]. van der Waals force contributes mainly interactions between nonionic PCs and the soil solid phases, sorption of nonionic PCs in soils is believed to be controlled by the weak van der Waals forces to soil OM or negatively charged silica [30,49,100].

PCs are usually classified to be ionized, amphiphilic or amphoteric with regard to the various structural classes, and their physicochemical behaviours in soil depend on their molecular structures. The same pharmaceutical class compounds display generally similarity in the sorption behavior on soils and on pure phase minerals [90,95]. For instance, adsorption of fluoroquinolone (FQ) carboxylic acid derivatives in soils with varied clay mineral and OM contents of a wide range of soil types showed that different compounds in the same soil had similar K_d value due to their similar molecule structure of the carboxylic group, and other different substituents such as piperazine ring and chlorine in the FQ structures exerted no essential influence on their sorption capacities [90]. On the contrary, PCs of highly different chemical structures display notably different sorption capacity even dominated by the same sorption mechanism on the same sorbent. As an example, oxytetracycline (OTC) and ciprofloxacin (CIP) zwitterion sorption onto the same soil showed that sorption of both the compounds was pH-dependent, OTC exhibited greater sorption in most types of soils but CIP only showed greater sorption in limited soils [38,101,102].

Many PCs are polar and ionizable with pKa values in the pH ranges of natural soils. For polar PCs or ionic PCs, electrostatic (ionic and/or covalent) interactions and hydrogen bonding will play critical roles in their sorption to soil solid phases, while for neutral PCs, sorption to soils will be dominated by van der Waals and hydrophobic interaction [95]. Sulfonamides (SAs) contain the aminophenylsulfonylamide core structure with >S O and >N–H functional groups. These compounds are in the neutral form under neutral or slightly acidic conditions, but in anionic or cationic species depending on the solution pH [9,21,91,103,104]. FQs contain carboxylic acid and ketone groups and are ionic compounds, existing as cationic, anionic, or zwitterionic forms with varying pH [14,21,102,105]. Macrolides comprise highly substituted monocyclic lactone with glycosidical saccharides of hydroxyl group and are hydrophilic ionic compounds, which are cationic or neutral, but the functional group is protonated in neutral to acidic solutions [39,106]. β-Lactams consist of the sulfur-bearing ring fused to a four membered β-lactam ring, and are zwitterionic due to the coexistence of carboxylates and ammonium groups [107,108]. Aminoglycosides are composed of amino sugars shared by a glycoside linkage to a hexose nucleus of the compounds, and are basic and strongly polar polycationic molecules [109].

SAs are predominantly in the neutral form in general soil environments. Adsorption of SAs on soil occurs likely on the (001) surfac of clay minerals and even on the interlayer space of swelling clay phases through an exothermic process [58,92,110], or on specific molecular sites and functional groups of soil OM [9,103]. Thus, the removal of OM in calcined soil decreased the adsorption of sulfamethoxypyridazine (SMP) and enhanced obviously desorption remarkably [31]. In the five Iowa sandy loam and loam agricultural soils (pH: 5.4–8.2; OM: 0.1–3.8 %; CEC: 10.7–23.3 cmol/kg), the sorption capacity of sulfamethazine (SMZ) correlated positively with soil OM content but negatively with soil pH value. The influence of soil pH on sorption behavior was attributed to the variable ionization states of SMZ with pH, since hydrophobic interactions dominated in pH < 7.4 due to the non-ionized form of SMZ and surface sorption dominated in pH > 7.4 due to the anionic form of SMZ [111]. The sorption potential of amphoteric sulfamethoxazole (SMX), sulfadimethoxine (SDT), and SMZ in the agricultural sand soils from South Korea increased as pH reducing from 8.0 to 4.0, attributing to the resultant electrostatical interactions between the negatively charged soil surface and the neutral and positive-charged SAs when pH is reduced [112].

TCs have generally similar structure, pK_a, and solubility, comprising ionic/polar groups such as hydroxyl, amino, and ketone. These compounds behave as amphoteric compounds and exhibit hydrophilic behavior and high affinity for soil OM via cation bridging and cation exchange interactions [14,113]. Adsorption of TCs to soil is largely affected by soil OM and pH [14,114]. Different compounds of TCs class exhibited various affinities for soils, which were attributed to their different molecular characteristics among the species [115]. In the structure of TC molecules, the steric hindrance between the hydroxyl group at the C5 position and the protonated dimethylamino group result in the lower OTC adsorption compared to TC and chlortetracycline (CTC), and the notably higher mean molecular polarity of CTC molecule relative to that of TC and OTC causes the higher affinity for soils. As a result, CTC showed higher affinity for soils relative to TC and OTC, and the adsorption affinity for soils exhibited the sequence as: CTC > TC > OTC [116,117].

FQs (CIP, enrofloxacin: ENR, and levofloxacin: LEV) have a fluorine atom in the C6 position of the quinolone ring and behave neutral, charged, or zwitterion forms in soil [24,105,118]. Sorption of CIP onto soil matrix can take place via both surface complexation (–COOH group) and cation exchange (–NH₃⁺ group). The deprotonated –COOH group favored the formation of surface complexation with soil components, and the protonated –NH₃⁺ group facilitated the cation exchange reaction with soil matrix, which resulted in relatively greater sorption capacity of CIP in soil compared to most of other organic pollutants [119]. Adsorption of CIP, ENR, and LEV multiple compounds in the Kentucky (USA) silt loam agricultural soil (Typic Hapludalf) showed that FQs at environmentally relevant concentrations sorbed quickly to agricultural soils, due to their amphoteric property and different speciation with various functional groups including carbonyl groups, carboxylic acids, and amides [119].

Sorption of macrolides TYL on goethite was attributed to the surface complexation between the reaction between their active functional groups (–OH) or the formation of carboxyl dentate complex on the surface of goethite [120], while sorption of TYL on HA was influenced by ion exchange, surface complexation, and hydrophobic interactions, with a significantly high K_d value of 386 L/kg [121]. ERY sorption to loess soil was mainly controlled by electrostatic and cation exchanged interactions with surfaces of soil caly minrals, which was facilitated under weak acid (pH 5–6) condition [122]. Moreover, PCs often contain multiple functional groups or pH-dependent functional group character, the wide diversity of functional groups present in PCs suggest that interactions between PCs and the soil solid phases may simultaneously involve multiple sorption mechanisms. For example, perfluorooctane sulfonate (PFOS) adsorption is likely affected by hydrophobic interaction in addition to electrostatic interaction, due to the presence of the sulfonate group of PFOS and the hydrophobic nature of its perfluoroalkyl tail [94].

4.2. Synergic effect of PCs in soil

Different PCs such as those in the groups of FQs, TCs, and SAs were frequently observed in environments especially in soils with manure-based organic fertilizers [4,12,123]. Adsorption of these compounds to soil is one of the main control factors influencing their persistence and potential effects on biogeochemical cycles in soil. The adsorption process may differ when presents in a soil as a single compound or in a soil with other pharmaceuticals due to the synergic effect [37,124]. Study of behaviors of six common pharmaceuticals carbamazepine (CBZ), citalopram, clindamycin (CDM), fexofenadine, irbesartan, and SMX with three different forms (cationic, zwitter-ionic or neutral, anionic) in seven soil types (Stagnic Chernozem, Haplic Chernozem, Greyic Phaeozem, Haplic Luvisol, Haplic Cambisol, Dysric Cambisol, Arenosol Epieuric), showed that these compounds exhibited synergic sorption effect in certain soils but was antagonistic in other soils, dependent closely on the chemical property of PC molecules [125]. Fexofenadine and irbesartan (zwitter-ionic or neutral) exhibited increasing sorption in Stagnic Chernozem soil due to synergic effect, the sorption enhanced moderately when co-existence with CBZ (neutral) in Greyic Phaeozem and Haplic Chernozem and, in particular, the sorption capacity increased most strongly when presence of citalopram (cationic) in Dysric Cambisol [125]. In crop soils from NW Spain, TC adsorption was almost total and irreversible, while SDZ adsorption was lower and was more easily released. However, SDZ was more adsorbed and suffered lower desorption when co-existence with TC, indicating a synergic effect of TC on SDZ adsorption [126]. The increasing sorption of PCs was probably attributed to the cooperative multilayer sorption on soil constituents, since sorption of positively charged PCs onto the negatively charged soil surfaces may reduce the repulsion between the soil components and the negatively charged PC molecules, and ionization of PCs due to dipole-induced dipole interactions between non-polar and polar molecule can increase the sorption of neutral compounds [124].

The sorption capacity of a compound depends on its chemical property, interactions with molecules of other compounds, and the specific soil properties, i.e., CBZ on OM content, citalopram and CDM on the base cation concentration, SMX on hydrolytic acidity, and fexofenadine and irbesartan on complexation, respectively [37,125]. In some cases PC mixture of different classes can also cause adverse impact on their sorption in soil. For examples, sorption of both CDM (cation and neutral) and SMX moderately decreased when co-existence in Haplic Luvisol, Haplic Cambisol, and Arenosol Epieuric [125]. In different fractions of Zibo (eastern China) farmland soil, the effective adsorption sites of soil fractions facilitated adsorption of SDZ, TC, and norfloxacin (NFX), and the adsorption capacity of these compounds exhibited the order as: TC > NFX > SDZ due to the competitive effect, which was dependent on the individual chemical property of each compound [127].

PCs with similar structure of the same class often showed competitive adsorption effect in soils [128,129]. Sorption capacity of mixed FQs (CIP, ENR, and LEV) in soil was observed in order as: CIP > ENR > LEV due to competitive adsorption [130]. TCs (TC, OTC, and CTC) showed competition for adsorption sites in six Galicia (NW Spain) agricultural soils with different pH (4.49–7.06) and OM content (1.07–10.92 %). Sorption of TCs depended mainly on the soil OM content and behaved very similarily in ternary systems with low concentrations, but at high dose conditions a favorable sorption for CTC compared to CT and OTC was observed due to the competition for adsorption sites [128]. In general, the adsorption capacity of amphenicols in soils is significantly weak, since the adsorption of amphenicols on the soil surface is controlled by auxiliary hydrogen bonds and weak hydrogen bonds [129,131]. However, sorption behaviors of amphenicols chloramphenicol (CAP), thiamphenicol (TAP), and florfenicol (FF) in different agricultural soils (clay, loam, and sandy loam) showed that amphenicols exhibited different adsorption capacity in single compound from that in ternary amphenicols. The adsorption capacity in single compound followed the order of TAP > FF > CAP, while in the ternary compounds the order changed to: CAP > FF > TAP, suggesting that TAP and FF will readily migrate to surface water from soil when presence of CAP due to competitive adsorption [129].

4.3. Preferential sorption of PCs by the soil components

Sorption of PCs onto mixture of soil OM and soil minerals deviates largely from the sum of their separated components due to the complex interactions between PCs and soils [49,84,100]. Soil OM has a notably lower ponit of zero net point charge (PZNPC) value of <3, as the carboxyl group of OM is much acidic than the simple carboxylic acid. In comparison, PZNPC values of Fe, Al-(hydr)oxides in monovalent electrolyte solutions are usually >7. With a decrease in ionic strength and pH > PZNPC, the surface-bound OM may expose additional ionizable functional groups due to the conformational change and result in complexing with more cations, while at low pH, cation adsorption may take place by variable-charged surface functional groups such as soil OM [30]. In natural mineral soils, soil OM and clay minerals may counteract their adsorption capacities due to the combination between the charges and functional groups of OM and the reactive sites on surface of clay minerals, which consequently reduce the amounts of reactive sites and decrease the PC sorption capacity of the mixture of clay minerals and humic acid compared to that of each of the components.

The mineral particles such as clay minerals and Fe, Al-(hydr)oxides are often coated by negatively charged soil OM, which leads to their negative zeta potentials even with a very low soil OM coverage. The variable surface charges of different soil matrices may adjust their surface charge on adsorption, which act seemingly as a new phase for surface complexation with PCs by electrostatic potential and site density, and are significantly different from those of their discrete components [84,132]. For example, sorption of OTC, TYL, and sulfachloropyridazine (SCP) on Dutch soils with a wide range of properties i.e., soil pH, OM content, clay content, CEC, Fe, Al-oxyhydroxide content showed significant variations in sorption capacity due to the complex interactions between the compounds and soils, which can not be explained according to a separate soil property [22,31,122]. In addition, the decreasing adsorption capacity of smectite and OM aggregates for TCs compared to smectite and OM separates was attributed to the hindering of entry of TC into interlayers of smectite, as TCs may also sorb in the interlayer region of smectite [92,133].

PCs exhibit usually preferential sorption to different soil components due to their differential affinity and sorption capacity, as interactions between PCs and soil matrices are mainly controlled by chemistry of PCs and the surface chemistry of soil matrices (clay minerals, OM, and Fe-oxides) of the soil. For examples, adsorption and transport of LEV and ofloxacin (OFL) in the mixed goethite and soil OM matrix showed that both LEV and OFL preferentially adsorbed to soil OM than to goethite [134]. The sorption amount of OTC by OM-removed phaeozem soil (black soil dominated by illitic minerals) exhibited no significant difference from that of the untreated phaeozem soil, while the sorption capacity of OTC onto the Ultisol soil (red soil dominated by kaolinitic minerals) increased notably in the OM-removed Ultisol soil [135]. Sorption and desorption of OTC increased in the Fe-, Al-oxides removed phaeozem soil, but sorption of OTC onto the Fe-, Al-oxides removed Ultisol soil showed negligible change, and its desorption capacity decreased. The behaviors of OTC sorption and desorption in the Ultisol soil and the Phaeozem soil were significantly different due to the distinct physical and chemical properties of the two soils. The Phaeozem soil is dominated by illitic minerals, while the Ultisol soil is consisted mainly of kaolinitic minerals [135]. Soil OM and kaolinite are relatively more important than Fe, Al-(hydr)oxides in the variable-charge properties of these soils [136]. For clays, OM, and clay-OM aggregates separated from the smectitic Vertic Endoaquoll and loamy Aquic Hapludoll agricultural soils (USA), sorption of TC and CTC displayed the strongest sorption to clays, followed by OM, and then clay-OM materials [133]. OTC sorption by different fractions of OM from the Weihe River sedimentary soil and forest soil in Shanxi (NW China) showed that dissolved organic matter (DOM) in soil inhibited the sorption of OTC, but OTC sorption by different OM fractions exhibited distinct characteristic. The sorption capacity of OTC for the OM-rich components was positively correlated with the aromaticity of OM, and for the inorganic mineral-rich component the sorption capacity was influenced by aromaticity of OM and the mineral content as well as the mineral species [35].

Soil CEC is largely dependent on compositions of soil matrices, which is considered as one of the key factors influencing the sorption capacity of PCs. For high CEC (>20 cmol/kg) soils at natural soil pH, cation exchange was the dominant mechanism of CIP sorption due to the availabe surface sites, although cation bridging and surface complexation mechanisms could also potentially contribute to CIP sorption to soils; for soils with low CEC (<20 cmol/kg) but more Fe-,Al-oxides, sorption of CIP to soils was partially attributed to the specific contribution from surface complexation with surfaces of Fe-,Al-oxides at high pH conditions [137]. Sorption of ENR, decarboxylated ENR, CIP, LEV, and the fluorochloroquinolone carboxylic acid derivative in soils with different properties showed that all chemicals were strongly adsorbed in soils and the sorption was related to the presence of clay minerals, due to the preferable adsorption in the interlayers of swelling clay and in the outer surface via Coulombic interactions [90]. OTC and CIP sorption onto Alfisol, Inceptisol, Spodosol, Ultisol, and Vertisol suggested that OTC zwitterions sorbed to a greater extent than CIP zwitterions in soils with moderate-to-low ECEC values (<10 cmol/kg) and Fe-hydroxides, due to greater surface complexation to soil metal oxides and aluminosilicate edge sites via adjacent pairs of hydroxyl groups on the OTC molecule (absent in CIP). Inversely, CIP sorbed to a higher extent than OTC in soils with relatively higher ECEC values (>10 cmol/kg) and aluminosilicates, attributing to the greater distance between the anionic and cationic groups on the compound and thus the reducing Coulombic attraction to the surface [24,138].

OM component of different soil types exhibits significantly variable chemical property and thus shows different affinity and binding intensity in complexing PCs. In the tropical Brazilian soils, molecular interactions between OTC and HA extracted from soil were relatively weak, and the sorption mechanism involved no hydrophobic interaction or overlapping π-bonding interaction, unlike those HAs in temperate soils, which showed higher sorption capacity and favored the retention of OTC in the soils [139]. The variable chemistry of OM from different soils was also observed in CIP sorption by the Elliott soil HA, the Pahokee peat HA, and the Suwannee river HA. The sorption capacity of CIP in soil HA and peat HA was notably higher relative to the aquatic HA at pH > 6 due to their generally high aromaticity of soil and peat OMs, as the aromatic structure and H–bond donating moiety in soil and peat humic substances could also facilitate the van der Waals interactions, except for the electrostatic interactions between CIP and the humic substances [140]. In the Spain loamy agricultural soils with addition of different amounts of HA, FQs sorption increased by orders of magnitude when addition of HA to soils, suggesting that soil OM played the critically role in the environmental retention, reactivity, and bioavailability of these compounds [105,141]. Therefore, FQs preferably persisted in the organic-rich horizon of soil profile due to the strong interactions between soil OM and FQs, which has been regarded as the major controlling factor influencing the persistence of FQs in soils [96,102,105,140].

Soil texture affects the relative contents of soil components of clay minerals, OM, and Fe-oxides and, consequently, exerts on the sorption capacity of PCs in the soil. Adsorption of azithromycin (polar compound) in four agricultural soils (Inceptisols on alluvial sediments) with different texture and OM content at the Bogotá savanna, Colombia showed that the retention capacity of azithromycin in soils was highly changeable and was intimately related to the clay content other than the soil OM [142]. In the Danish temperate agricultural soils with various clay contents from sand soil to sandy loam soil and varying CEC (6.7–35.3 cmol/kg), OTC sorbed strongly in different types of soils from the sand soil to sandy loam soil, with notably higher K_d values of 417–1026 L/kg and no significant desorption, and TYL sorption seemed to only positively correlate with soil clay content, with K_d values increasing from 8.3 to 128 L/kg. Whereas, metronidazole and olaquindox (neutral compounds) were only weakly adsorbed in these soils and showed highly mobile [143]. For 13 Turkey agricultural soils with soil texture from clay loam, sandy, sandy clay loam, clay, loamy, and sandy loam, the PC sorption was dominantly dependent on the metal content rather than the OM content and CEC of soils, the sorption capacity of TC was markedly enhanced by increasing amounts of polyvalent metals, but the presence of polyvalent metals reduced the SA sorption in the soils [144].

5. Soil-type dependent behaviors of PCs in soil

5.1. Soil-type dependent sorption behaviors of PCs

Sorption and affinity of PCs for soils are notably soil-type dependent [[6], [15], [89]], as different types of soils form under different climatic conditions and comprise distint clay mineral and OM components, which determine the sorption behaviors of PCs in the soil environment (Table 4). For example, due to the specific soil OM composition of mainly alkylic and aromatic components in the spodic horizon of the Northwest German Podzols soils, organic pollutants were obviously adsorbed by direct binding as polar species at the aromatic structure, because the alkylic OM components interacted with the stiff-benzene rings to a flexible, sponge-like molecular system, and organic pollutants might be physically trapped in these voids [26,27,79]. Also, spodic horizons often contain more amorphous Fe-, Al-hydroxides with a variable surface charge, which changed significantly with solution pH and thus efficiently fixed organic pollutants under soil environment [145]. Sorption of PCs to different types of soils (inceptisol, ultisol, alfisol, spodosol, and vertisol) from the eastern United States was largely influenced by the soil-type dependent property [38]. Soils with different soil property exhibited changeable sorption capacity of PCs, and the predicted sorption capacity from soil property of a soil type can be applied for assessing PC mobility in soil with similar soil property. Therefore, the modeling sorption behaviors in combination with half-lives of PCs and soil texture can be proposed to assess the persistence and bioavailability of PCs in soils [89].

Table 4.

Characteristic soil components of some major soil groups and behaviors of frequently investigated PC classes in the soils.

Soils	Soil matrix [40,44]	Orgarnic matter and functional groups [46,47,65]	Sulfonamides (amphoteric compounds)	Fluoroquinolones (ionic compound)	Tetracyclines (ionic/polar groups)	Macrolides (methylamine functional group)
Ferralsols; Lixisols; Acrisols	Kaolinite (or halloysite), Fe-, Al-oxyhydroxides	O-alkyl and alkyl (polysaccharide) compounds; hydroxyl; carboxyl	Hydrophobic; mobile; Kd(SMZ):1.37 [14] Kd(SDZ):8.4–11.1 [146] Kd(SDZ):0.8–14.3 [147] Kd(SMX):0.7–28.5 [147] Kd(SCP):0.7–70.1 [147] Kd(SDI):1.0–32.0 [147] Kd(STA):1.01–67.1 [147]	Cation exchange; immobile; Kd(ENR):3040 [148] Kd(ENR):3037 [90] Kd(NOR):591 [14] Kd(CIP):85 [149] Kd(NOR):2500 [150] Kd(ENR):544–251430 [147] Kd(DAN):847–255643 [147] Kd(CIP):726–1277873 [147] Kd(NOR):999–335633 [147]	Cation exchanged and complexation; immobile; Kd(TC):1093 [14] Kd(TC):130.92 [26] Kd(OTC):650–2191 [151] Kd(OTC):1435–3135 [101]	Cation exchange; moderate mobile; Kd(ERY):130 [14] Kd(TYL):6.0–22.6 [81]
Vertisols	Smectite, vermiculite, mixed-layers, illite, chlorite	Aliphatic compounds, and minor peptides, heterocyclic N forms, and carbohydrates; hydroxyl; carboxyl; aromatic; amide	Hydrophobic; highly mobile	Cation exchange; immobile; Kd(CIP): 26000–36000 [38]	Cation exchange, complexation; immobile Kd(OTC):1751–4377 [101]	Cation exchange; immobile; Kd(TYLA):5520 [152]
Andosol	Kaolinite/smectite, kaolinite (or halloysite), Fe-hydroxy-oxides	Polysaccharide, chitin, aromatic, lignin, carboxylic compounds; hydroxyl; aromatic; carboxyl	Hydrophobic; mobile; Kd(SMX):1.12–3.09 [153] Kd(SCP):1.87–10.59 [153] Kd(SMZ):0.96–5.08 [153] Kd(SMP):0.9–26.0 [154]	Cation exchange and hydrophobic; immobile; Kd(CIP):128 [149] Kd(ENR):121–2345 [154]	Cation exchange and complexation; immobile	Cation exchange; immobile; Kd(TT):1.13–1150 [153]
Podzols	Smectite, I/S, I/V minerals, Fe oxides	Lignins, tannins, aliphatic waxes, and branched alkyl components; aromatic; hydroxyl; carboxyl	Hydrophobic; highly mobile; Kd(SSX):0.27 [155]	Cation exchange and hydrophobic; immobile; Kd(ENR): 1230 [90] Kd(CIP): 150–990 [38]	Cation exchange and complexation; immobile; Kd(OTC):486–2519 [101]	Cation exchange; immobile
Luvisols	Illite, vermiculite, I/S minerals, kaolinite, Fe oxides	Carboxyl, aliphatic, and lignin components; hydroxyl; aromatic; carboxyl	Hydrophobic; highly mobile; KF(SMX):1.28 [89]	Cation exchange; immobile; Kd(CIP): 2420 [130] Kd(ENR): 1345 [130] Kd(LVF):1117 [130]	Cation exchange and complexation; immobile; Kd(OTC): 769 [151] Kd(OTC):1454–4897 [101]	Cation exchange; moderate mobile; KF(CLA):436.3 [89]
Phaeozems, Chernozems, Kastanozems	Illite, smectite, vermiculite, chlorite	Saccharides, lignins; aromatic; hydroxyl	Hydrophobic; mobile; Kd(SPY): 3.47 [98] KF(SMX):0.57–1.65 [89] Kd(SDZ):80.0 [146]	Cation exchange and hydrophobic; immobile	Cation exchange and complexation; immobile; Kd(TC):5879.84 [26]	Cation exchange; moderate mobile; KF(CLA):159.9–601.6 [89]
Paddy soils	Kaolinite, illite, Fe-, Al-oxyhydroxides	Aromatic compounds (lignins); aromatic; hydroxyl	Hydrophobic; mobile; KF(STZ):22.1 [156]	Cation exchange and hydrophobic; immobile; Kd(NOR):8700–8900 [150]	Cation exchange and complexation; immobile; KF(OTC):3012.4 [156] Kd(TC):790.38 [26]	Cation exchange; moderate mobile; Kd(TYL):6.19 [157]
Preferential sorption to soil component			OM > Clay	OM > Clay	Clay > OM > oxides	Clay > OM

Notes: Sorption coefficient K_d in L/kg and K_F (the Freundlich constant) in mmol¹⁻ⁿLⁿ/kg.

Soil type influences significantly the sorption behavior and durability of PCs in soil environment. For instance, sorption behavior of TC onto different soils (the Fujian lateritic soil, the Sichuan paddy and purplish soils, the Lanzhou cultivated loessial soil, the middle-China cinnamon soil, and the Harbin black soil) from China showed that the TC sorption capacity varied significantly with soil type due to diverse soil property, the Harbin black soil (OM-rich Mollisol) had the highest TC sorption capacity, while the Lanzhou cultivated loessial soil had the lowest. In particular, the TC sorption in acid soils was predominantly attributed to hydrophobic interaction, and in alkaline soils it was derived from cation exchange [26]. A field study of TCs and SAs in a temperate northern Germany sandy soil showed that TCs tended to persist and accumulate in the sandy soil with repeated fertilizations, no leaching of TCs was observed into deeper soil segments or groundwater, while SAs occurred only in very low concentration in the cultivated soil layer due to continuously leaching from soil into groundwater. The different environmental behaviors between the two groups of PCs were attributed to their different sorption capacities in the soil and thus the different mobility in the ecosystem [37,158]. In subtropical vegetable fields at the Pearl River Delta, South China, TC, OTC, and CTC were detected in 85 % of soil samples with a mean concentration of 78.05 μg/kg and few samples of ∼100 μg/kg due to manure application, and accumulation of amphoteric TCs in soils was attributed to their strong complexation with the kaolinitic soil matrix [123].

Soil texture varies greatly with soil type and significantly influences the sorption behaviors of PCs. Sorption of OTC, amoxicillin (AMX), and sulfathiazole (STZ) in silt loam paddy soil (OM, 1.83 %; CEC, 10.35 cmol/kg) and reclaimed sandy loam tideland soil (OM, 0.82 %; CEC, 5.2 cmol/kg) from South Korea showed that the silt loam soil exhibited higher sorption capacity for these compounds than the sandy loam soil, and the sorption capacity enhanced with increasing soil OM content but decreasing soil pH [156]. Similar sorption behaviors of other selected PCs including bisphenol A, CBZ, gemfibrozil, octylphenol, and triclosan were also observed in three different American soils, i.e., sand, clay, and loam types with gradually increasing OC content, octylphenol and triclosan were retained readily by soils, while CBZ and gemfibrozil exhibited poor sorption to soils [20].

The soil-type dependent clay mineral species in soils exert greatly on adsorption of PCs in soils. Unlike the temperate soils, tropical soils are characterized by kaolinitic clays and more Fe, Al-oxides instead of smectite, illite, and mixed-layer species, and have relatively lower values of CEC, pH, and OM content [147,159]. Also, tropical soils form under climate conditions of higher temperature and annual precipitation, which may influence intensively on biodegradation, accumulation, and transportation of organic pollutants in these soils [146,160]. Kaolinite agricultural soils from Brazil and France exhibited considerably lower sorption capacity for the FQ carboxylic acid derivatives than those of Germany, Sweden, and Phillippines smectite-bearing soils [90]. The illite-dominated soils (Ultisol, Alfisol) exhibited higher sorption capacity for OTC, but the interaction intensity between soil matrices and OTC was relatively weak and the bioavailability of sorbed OTC increased due to its higher release from soils. Conversely, OM- and kaolinite-dominated soils (Oxisol, Ultisol) had lower sorption capacity, but the release of sorbed OTC from soils decreased due to the stronger interactions between OTC and the soil OM, kaolinite, and Fe-, Al-oxides [151]. A recent invstigation of CIP sorption in Chile Ultisol and Andisol soil also indicated the different adsorption kinetics in the soils, sorption of CIP in the Ultisol occurred on heterogenous sites as multilayers, while on the Chile Andisol by monolayer, and the Andisol soil had a higher sorption capacity compared to the Ultisol soil [149].

For the same soil-type groups, soils formed under slightly different temperature and precipitation conditions or from various evolution processes may also display certain difference in soil property. For example, although Ultisols at Guizhou, Hunan, Jiangxi, and Guangxi in South China were consisted of mainly kaolinite and gibbsite, the soil samples from Guizhou, Hunan, and Jiangxi Province contained certain amounts of 2:1 type clay minerals (illite and vermiculite) and goethite, while the Hunan, Jiangxi, Guangxi Ultisols also comprised hematite and magnetite. The positive surface charge, the zeta potential and isoelectric point, and the net surface charge of these soils showed slightly different, with a geographical northward decreasing order as: Guangxi > Guizhou > Hunan, Jiangxi, consistent with their small difference in weathering intensity of the soils [161].

5.2. Transport behaviors of PCs in various soil types

5.2.1. Variable mobility of PCs in a soil

The sorption and desorption processes of PCs in soil influence on either of being retained in the topsoil or leaching into the lower strata of the soil. In particular, the sorption intensity of PCs in soil affects their availability and the residual concentrations within the soil environment [6]. The mobility of a pharmaceutical compound through the lower soil horizon is correlated to the rate of sorption by the soil matrices, which depends mainly on interactions between PC and soil matrices [21,126,162]. For instance, in the field fertilized with liquid manure TCs were found persisting at soil up to 30 cm depth over long time periods and accumulated in the soil environment due to repeated fertilizations. However, these sorbed compounds were also released from the soil matrices and further mobilized into ground water or surface water due to leaching or erosion [124]. As a result, OTC and CTC concentrations were high in agricultural lands, and abundances of CIP, NOR, and TC were relatively lower [6].

In a certain soil environment, sorption and transport of PCs are mainly dependent on their molecular structures and differ significantly with regard to various structural classes [15]. Sorption of OTC, olaquindox, metronidazole, and TYL to the sandy loam and sand columns showed that OTC was strongly adsorbed to the soil matrix and did not transport into deeper soil horizons, TYL was retained in different horizons due to the relatively weak sorption strength and certain extent of transport, while olaquindox and metronidazole migrated trough the soil columns due to their lower sorption strength to soils [143]. Soil pH can largely influence the sorption behaviors of PCs in soils by changing their chemical states, especially for the amphoteric compounds. For example, in the sandy loam, loam, loamy sand soils (USA) with various pH (7.5, 4.9, 5.3) and CEC values (23.7, 12.7, 6.8 cmol/kg), sorption and mobility of SMZ and STZ depended dominantly on pH, soil charge density, and contact time [163]. Although LEV was efficiently adsorbed (adsorption ratio >90 %) by the silty clay soil from the temperate North China Plain, the sorption strength of LEV was relatively weak via a physical interaction, which was readily influenced by solution pH, cations (Ca²⁺, Na⁺, NH₄⁺), and soil OM, LEV was observed penetrating through the silty clay and completely released to the groundwater after 100 days [23].

DOM is strongly mobile in soil and thus plays a role in PCs sorption and mobility. The concentrations of TCs and SAs in soils were positively correlated with the soil OM, with notably higher R² values of 0.93 and 0.86, respectively [164]. In particular, DOM contains usually polar functional groups (carboxyl and hydroxyl), which can form stable complexes with PCs and potentially increase their mobility, while the non-polar groups (aromatic and aliphatic) in DOM can solubilize hydrophobic PCs and facilitate their uptake into groundwater. The migration and transformation of macrolides antibiotic erythromycin (ERY) in the surface water of soil were significantly related to DOM due to the complexation of ERY with DOM functional groups [165]. The presence of DOM was observed to decrease the sorption of OTC on smectite and resulted in an increase in its mobility in the soil environment [82,134]. In addition, OM coating on soil minerals altered their surface charge characteristics, and/or competed directly for potential sorption sites between OM and TC molecules, hence, both dissolved and mineral-bound OM potentially increased the transport of TC in soil environment [166,167]. TCs were strongly sorbed and fixed to the soil column of the Colorado sandy clay loam agricultural soil and were rarely observed in the leachate, while SMZ was mostly observed in the leachate from the soil column. Transport of these compounds was considered to be closely related to mobile colloids such as DOM especially for SMZ [168].

Organic pollutants often exhibit variable affinity and mobility in different soil environments due to discrepancy in soil property. For the thirteen highly-weathered tropical Brazilian soils including Hapludox, Rhodic Eutrudox, Arenic Hapludult, Hapludalf, Udordent, Dystrochept, Albaqult, and Argiudoll, sorption of FQs was notably strong (K_d ≥ 544 L/kg), while sorption of SAs was relatively weak (K_d = 0.7–70 L/kg) [147]. The higher retention of SAs in soils was observed positively correlating with OM, and the affinity sequence of the study PCs in the Brazil soils was in the order: sulfadiazine (SDZ) < SMZ < SCP, SDZ < SMX < SMZ < SCP < STZ, but for some acidic Brazil soils exhibited the order: SDZ < SMZ < SDT < SCP < sulfaquinoxaline [117]. In temperate Brazil soils with different texture (sandy, sandy-clay, and clay), SDM, SQX and SMZ all had a low sorption potential on soils and were highly mobile at natural pH, and SAs sorption increased with the increasing OM content. Clay soils had relatively higher sorption capacities compared to the sandy soil, with a sorption capacity sequence as: SQX > SDM > SMZ. The generally low sorption capacity of SAs indicated the notably weak interactions between the neutral molecules and the binding sites of soil matrices, and SAs tended to leach from sandy soils with low OM and low cation exchange capacity [169]. However, in temperate Germany soils with nearly neutral pH, the affinity sequence displayed different order as: SAA (sulfanilamide) < SDZ < SDT < SMZ < SPY (sulfapyridine) [98].

Field studies of the clay loam soil and the sandy loam soil in temperate United Kingdom showed that SCP was highly mobile and was rapidly transported to surface waters. Despite the high concentration of SCP in drain flow and also the influence of soil pH on sorption of SCP, the ecotoxicological effect of this compound was negligible due to the rapid degradation of SCP in soil [117,170]. However, this kind of PC exhibited notably different persistence and accumulation in the tropical Vietnam soil. In the silty clay loam (3.7 % sand, 59.4 % silt, 38.5 % clay, 18.5 g/kg OM, 10.8 cmol/kg CEC) in tropical fluvisol topsoil at the Red River Delta, Vietnam, SAs including sulfadiazine (SDZ), SMZ, SMX, trimethoprim (TMP) persisted and accumulated in the fooded soil [159]. Although the mineral components of the mineral soils are unclear, the temperate UK soils are expected to consist usually of 2:1 clay species such as vermiculite, illite, smectite, and mixed-layer clays, while the tropical Vietnam soil will contain mainly kaolinite and Fe-, Al-oxides minerals (Table 1). The strong surface complexation by π-π, hydrogen, and electrostatic interactions between SAs and surfaces of kaolinite, Fe-oxides, and soil OM instead of 2:1 clay species is probably responsible for the distinct behavior in the Vietnam soil.

5.2.2. Influence of soil property on PC mobility in soil

PCs are generally ionizable or are polar organic compounds with low water solubility, which can adsorb quickly to soil matrices and are strongly retarded in soils. Hence, the mobility and accumulation of PCs in soil are significantly related to properties of soil minerals and organic matter, as adsorption of PCs to soil depend mainly on specific molecular sites and functional groups of organic pollutants, as well as surface chemistry of soil matrices [9,171]. Also, differential sorption intensities of PCs to soil OM and nano-size soil minerals such as Fe-, Al-oxyhydroxides can cause fractionation of different PC species due to their transport in soil [134]. The soil-type related pH can be an important factor affecting the sorption and migration of PCs, as in different soil pH, PCs may present different states of cations, anions, or zwitterions, and change their interaction mechanism with soil matrices [15].

Pharmaceuticals showed generally lower average dissipation half-lives and variability in Chernozem soils compared to Cambisol (Inceptisol) soils, CDM, clarithromycin (CLA), and atenolol seemed immobile in soil environment due to their generally higher sorption affinity and the dissipation rate, and the potential mobile species due to their great persistency and low sorption affinities in soils were in order as: CBZ > SMX > TMP > metoprolol [32]. FQs can be highly sequestrated even for soils with rich sand and poor OM, and in the highly weathered tropical soils sorption of FQs is mainly influenced by cation exchange, while sorption of SAs is mainly controlled by hydrophobic interaction as well as non-hydrophobic interactions with organic and/or mineral surfaces, depending largely on the soil texture and cation exchange capacity [147]. In the six New Zealand agricultural soils with different CEC, OM and clay contents, sorption of SAs was primarily influenced by hydrophobic interactions and could potentially penetrate to groundwater due to their low sorption affinity, and TYL was strongly adsorbed in smectitic and kaolinitic loams and could likely exhibit non-mobile in the soils [153]. SAs were substantially retained in soils at low pH condition due to the increasing fraction of cationic and neutral forms of the SAs, but at high pH condition the mobility of SAs was enhanced since SAs presented predominantly in the anionic form, and the repulsive interaction between the compounds and the negatively charged soil surface reduced their sorption capacity onto soils [22,37,125,163].

Variations in soil structure and soil component between different soil types and different soil horizons of a soil profile impact significantly on water flow, adsorption extent, and contaminant transport in soils [172]. The transport of sulfentrazone in soil horizons A, B, and C of an Ultisol and an Oxisol in Viçosa-MG, Brazil showed that sulfentrazone leached heterogeneously in Ultisol and Oxisol profiles, with leaching order in Ultisol as: A < B < C, and in the Oxisol as: A < C < B [173]. Different from the Haplic Cambisol, chlorotoluron was highly mobile under saturated conditions in deeper soil horizons of Czech Republic Haplic Luvisol and Greyic Phaeozem. The highly infiltration and transport of chlorotoluron in the Bt₁ horizon of Haplic Luvisol were facilitated by the well-developed prismatic structure of the soil, and the decreasing infiltration in the Ap horizon this soil was attributed to the low aggregate stability of the initially well-aggregated soil. On the contrary, the infiltration and transport of chlorotoluron were most slow in the Bw horizon of Haplic Cambisol, since this soil horizon had poorly-developed capillary pores and gravitational pores [172]. In the Chongming (Eastern China) agricultural soil profile with soil texture of loamy sand, sandy loam, and silty loam for different soil layers, the sorption capacity of the compounds in all soil layers followed the order as: TMP > SPD > SDM > SMZ, and the difference in sorption and transport behaviors of TMP, SPY, sulfameter, and SDT between the soil layers was closely dependent on chemical property of the PC and soil properties including porosity, texture, soil pH, and OM content [174].

Soil texture and soil type highly influenced the behavior of estrone and 17β-estradiol in soils, those with large particle, low OM, and presence of macropore and other heterogeneities resulted in less contact time between soil matrices and the compounds and facilitated their transport through the soils. As a result, estrone and 17β-estradiol were highly retained in the Oxisol and Mollisol soils compared to those in cinder [175]. In comparison with the three arid soils, i.e., Arizona clay loam (Typic Natrargid) with low OM, Utah silt loam (Aquic Natrixeroll) with high OM and biosolid amendments, and Utah silt loam with no biosolid amendment, sorption capacity of lincomycin (LIN) showed the order as: the unamended Utah silt loam > the biosolid-treated Utah silt loam > the Arizona clay loam. The difference in LIN adsorption between the soils was attributed to soil pH, since LIN existed mainly in cationic species in pH < 7.5–7.8, and sorption by cation exchange decreased due to the high cation concentration (Ca²⁺ and Na⁺) in the arid soils [176]. Behaviors of SMX, TMP, and LIN in silt loam to loam soils in Penn State (USA) soils changed with soil properties such as pH, OM content, soil texture, and CEC, TMP adsorbed strongly to all soils with limited transport, while SMX and LIN interacted weakly with the soil matrices and showed greater mobility throughout the soil profile [37,176].

Longer contact time increases interactions of PCs with the soil matrices and thus reduces their mobility in soils, since the sorption process will change from the initial boundary layer diffusion or external surface to the intraparticle diffusion or pore diffusion stage, and then to the equilibrium stage of sorption on the interior surface of the sorbent, as the reaction time proceeding [121,163]. The bioaccessibility and bioavailability of PCs will decrease with increasing contact time due to the diffusion into micro and/or nanopores in soil. However, the sorption and fixation of PCs do not necessarily eliminate their antimicrobial activities, since PCs are mostly polar molecules [58]. For instance, the soil-bound TC and TYL remain active, and show still antimicrobial effects that can exert on the selection of antibiotic resistant bacteria in the soil environment [177]. Only when diffusing into rigid components (interlayer spaces of swelling clays, nanopores within Fe, Al-oxides, and soil OM) and finally forming non-extractable molecules through the formation of covalent bonds with the solid matrices, the antimicrobial activity can be eliminated due to inhibition of releasing back into the soil solution [58,92,178,179]. In general, the sequestration process can only reduce the acute toxicity of PCs through transient storage in a form of not bioavailable, but it extends the residence time of PCs in soil, as the sequestered PCs may subsequently release back into groundwater in a bioaccessible form [6,180].

5.3. Degradation behaviors of PCs in soil

Prolonged contact time between PCs and soil matrices can result in “aging” effect and thus diminish their bioavailability in soils [18,163,179]. This process involves numerous molecular-scale mechanisms including pore and surface diffusion to sorption sites and evolution in bonding structure. For example, PC compounds can be partially degraded and catalyzed either by enzymes or mineral surfaces and thus produce reactive intermediates, which form covalent bonds with soil OM [181]. Kinetics which are crucial in the elimination of PCs from the environment include generally biological processes (biological redox and co-metabolism), chemical processes (hydrolysis, oxidation, and photodegradation), and physical processes (adsorption, desorption, and sedimentation). However, the adsorption/desorption related transport and degradation are considered as the more efficient processes in eliminating PCs from soils [24]. The dissipation half-lives of PCs in soils are quite variable, depending largely on their sorption, sequestration, molecule structure, and soil environment conditions, some of these compounds seem to persist for quite long time but others degrade very fast [182,183]. For examples, the dissipation half-lives of extractable and unextractable SAs in soil were 2–10 and 330 days respectively [183]. TCs persisted in soil for more than one year, and the half-life of macrolides varied significantly from 5 to 65 days [179]. β-lactams were biodegraded within hours to a few days [184], and dissipation of the neutral compound CBZ occurred in a wide range of half-life from 74 to 533 days [185,186].

Degradation of PCs in soil is mainly brought about by microbial activity especially enzymatic reactions, which results in transformation of the parent compound via hydroxylation and oxidative decarboxylation. Photodegradation may not play a major role since the influence of light is reduced when PCs are protected in soil [187]. The degradation is highly dependent on the overall soil quality such as microbial species, anaerobic and aerobic conditions, property of substrates, and soil temperature [6,24,188]. TCs were likely sequestered into a dormant undegradable phase in sandy loam agricultural soils, higher water and OM contents in soils facilitated the dissipation of TCs with a half-life time order as: OTC > TC > CTC [3]. OTC dissipation in three South China soils with various soil pH (4.3–8.5), OM content (0.35–2.57 %), CEC value (2.90–13.99 cmol/kg), and different redox conditions showed that the half-life of OTC in soil under aerobic condition was 29–56 days and under anoxic condition was 43–62 days. The slightly rapid dissipation rate under aerobic conditions was attributed to the increasing biodegradation in soils and to abiotic factors such as limited hydrolysis due to strong adsorption onto soil matrices, which depended mainly on soil pH, OM content, and redox conditions [189]. The half-lives of aerobic incubation of ceftiofur in sand, clay loam, and silty clay loam soils varied slightly from >49, 22, and 41 days respectively due to the different soil property [182]. Degradation of SDZ and SMX in the clay loam agricultural soil with low OM (0.50 %) and low CEC (7.0 cmol/kg) occurred at a lower rate under the anaerobic environment, with a degradation ratio of <75 % and a half-life of >20 days in sterile and non-sterile soils, influenced markedly by the soil physicochemical properties especially the microorganisms induced biodegradation [190]. Aerobic biodegradation of sarafloxacin in loam, silt loam, and sandy loam soils exhibited notably low degradation rate, with the extent of around 0.5 % mineralization of the compound in the three soils in 80 days, which was regarded as the strong binding of sarafloxacin to soil and thus, protected from the availability to microorganisms [191].

PCs in sterilized soil or soil with higher OM content show relatively longer half-life time, since soil OM can reduce bioavailability of organic pollutants and protect them from degradation [20,31]. In sterilized soils both FF and CAP almost did not degrade and only small portion (14.8 %) of TAP was degraded after 21 days. However, in unsterilized soil degradation of these compounds occurred quickly in 10 days, indicating that dissipation of FF, TAP, and CAP in soils was largely facilitated by microbial processes [129]. In the Hong Kong agricultural soil (clay loam with 0.80 % OM) and its sterilized sample (treated by autoclaving), PCs of different structure classes TC, SMZ, norfloxacin (NOR), erythromycin (ERY), and chloramphenicol (CAP) all exhibited moderate to high persistence under both aerobic and anaerobic conditions with a half-life time of 40.8–86.6 days in sterilized soil and the degradation order under aerobic condition as: NOR > ERY > CAP > SMZ > TC [14]. Hence, TC adsorbed most strongly and degraded most slowly in soils, and was more persistent in the surface soil and posed higher environmental risks to organisms in the topsoil. On the contrary, SMZ exhibited poor sorption and slow degradation, which transported readily in soils and thus posed greater risks to groundwater. Degradation of organic pollutants was mainly influenced by soil microbial activity and oxygen status and, particularly, aerobic condition facilitated the degradation of the pollution compounds compared to anaerobic condition [14,24,170,192]. The half-life time of monensin in clay loam soil (9.38 % OM and 33 % clay) was 22 days, notably longer than 4 days for loam soil (3.80 % OM and 15 % clay), indicating that microbiological activity played an important role in degradation of the compound, since the biodegradable organic pollutants served as a substrate for microbes, and microbiological activity depended intimately on the soil OM and moisture [6,188,193]. The microbe-induced degradation of PCs (CIP, ENR, and LEV) was also revealed by a field investigation in the silt loam agricultural soil (Kentucky, USA) under greenhouse conditions, these compounds were efficiently biodegraded in the presence of degrading bacteria and wheat rhizospheres [194].

The reactive clay-size components in soils such as Fe/Mn oxides and clay minerals, in particular, Fe(III)-bearing clays and pristine phyllosilicates can induce PC decomposition via hydrolysis or redox pathways in certain environmental conditions [36,195]. Under limited surface moisture conditions (2–11 %), soil clay phases possess high surface acidity and can potentially mediate the abiotic transformation of PCs [195]. Cation (Fe³⁺ and Al³⁺)-exchanged smectites show catalytic performance of hydrolysis of CAP due to the wetting-sensitive surface Brønsted- and Lewis-acid properties despite of the presence of soil OM, which occurs at a wider moisture range (10–100 %) and is considered as an effective pathway of abiotic transformation to eliminate PCs in soils [36]. Iron phases such as maghemite, hematite, and goethite occur widely in soils and are also important clay-size solid soil components [44]. Under the water-limited environment with an atmospheric relative humidity of 33–76 %, all these iron oxides exhibit strong surface chemical activity, and surface hydrolytic reaction of PCs on surfaces of iron oxide minerals may occur through H-bonding or Lewis acid interaction [196,197]. The markedly short half-life of <6 days for CAP degradation in the presence of iron oxide minerals has been attributed to the efficient hydrolysis of CAP due to catalysis of the iron phases, which was regarded as a potential pathway for degradation of PCs in natural soil [197].

6. Implications and perspectives

The characteristic clay mineral and OM compositions of the soil group determine the overall soil chemistry due to their different surface chemistry of the soil components, which largely determine the interaction mechanisms between PCs and soils, and exert distinctly on sorption, transport, and persistence of PCs in the type of soil (Table 1, Table 2, Table 4; Fig. 3). The mobility of PCs in soil and their potential pollution in groundwater are largely soil-type dependent, as the typical soil group exhibit a similar trend of adsorption-desorption mechanisms of PCs [6,15]. The analogous K_d value of a PC in the same soil group from different studies is also in line with the soil-type dependent property (Table 4). For example, acidic tropical soils such as Ferralsols, Lixisols, Acrisols, and Luvisols consist mainly of kaolinitic clays and minor Fe(Al)-oxides, with only trace OM of polysaccharide compounds, the mobility and potential pollution of PCs are largely dependent on both chemical properties of the pharmaceutical species and surface chemistry of kaolinite and Fe-,Al-oxides [160]. Thus, amphoteric SAs are generally mobile in these soils due to their dominant hydrophobic interactions with solid soil components, and ionic FQs display usually immobile behavior due to their mainly cation exchanged interactions with soils.

Fig. 3.

Change in surface chemistry of soil matrices with soil type and its influence on PCs behaviors in some selected soils [40,41,54,65,72,69].

Variations in sorption mechanisms of PCs and soil matrices determine greatly the degradability and persistence of the respective compounds in soil environment, which depend mainly on clay minerals, OM, and Fe-, Al-(hydr)oxide components of soils (Fig. 3). Sulfonamides have the potential to transport via groundwater and will increase possible exposure to aquatic organisms. Fluoroquinolones can be strongly combined with soil matrices by cation exchange and hydrophobic interactions, while tetracyclines interact strongly with soil matricesby cation exchange and complexation. These two groups of chemicals are highly immobile and will accumulate in the top soil layer, which accordingly increase their availability to soil-dwelling organisms. Adsorption of macrolides in soils is mainly controlled by cation exchange interaction via the methylamine functional group. Macrolides are generally in mobile to moderate mobile in Paddy soils, Phaeozems, Chernozems, Kastanozems, Luvisols, Ferralsols, Lixisols, and Acrisols, but are immobile in Vertisols, Andosols, and Podzols. With regard to the behavior and ecotoxicology of PC in soil, information about the general properties of various soil types will help to predict the persistence and bioavailability of PCs in soil environment.

Degradation and catalysis of PCs is highly dependent on the overall soil properties including microbial community, chemistry of soil components, as well as soil temperature and oxygen status. Lower soil OM and pH usually correspond to a lower microbial activity and thus yield higher durability of PCs, and the dissipation half-life of a PC in soil varies significantly with its sorption, sequestration, molecule structure, and soil environment conditions [6,32,198]. The degradation processes can generally eliminate toxicity of PCs, as the final degradation products are essentially nontoxic. However, some intermediates produced from the degradation processes can be toxic [188,[199], [200], [201]]. In the future for a better understanding of the behavior and fate of PCs in soil, more detailed studies are required on their degradation processes, as well as the behavior and toxicity assessment of the transformation products of PCs.

Experiments on mobility, bioavailability, and degradation of PCs were almost undertaken in laboratory investigations, and most factors affecting the PC behavior in the experiments were inconsistent with those in the field sites, especially the soil structure and the variable moisture and temperature of the employed soil samples [23,31]. In particular, the adsorption isotherm of a PC in soil is nonlinear due to changeable surface charges of clay minerals and soil OM with different solution pH and variations in sorption capacity, the adsorption coefficient will decrease with increasing PC concentration in soil solution. Since the concentrations of PC in soil is usually lower than those in experimental studies, the partitioning coefficient (K_d) of a PC in soil may be higher than it from laboratory study, and the risk assessment of PC based on value of K_d should consider the practical PC concentration in the field soil [37]. The variable texture and soil compositions as well as the different groundwater flow and adsorption capacity between soil horizons of a soil profile also influence the transport of PCs in soil. In order to establish a model predicting PC sorption and transport in soils based on typical soil types, knowledge of influence of each individual environmental factor on the behaviors of PCs is needed and especially a field site investigation is required to understand the sorption-desorption, transformation, and ecotoxicology of PCs in the typical soil type.

Data availability statement

Data will be made available on request.

CRediT authorship contribution statement

Hanlie Hong: Conceptualization, Methodology, Writing - original draft. Chen Liu: Visualization. Zhaohui Li: Writing - review & editing.

Declaration of competing interest

The authors declare no competing financial interests.