# Standing on the shoulders of giants – what we should learn from past soil chemists to address the current challenge of “legacy-phosphorus”

David Yalin, Mikhail Borisover, Tamar Lavi, Frédéric Gérard

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David Yalin, Mikhail Borisover, Tamar Lavi, Frédéric Gérard. Standing on the shoulders of giants – what we should learn from past soil chemists to address the current challenge of “legacy-phosphorus”. Plant and Soil, 2025, 516, pp.1 - 15. 10.1007/s11104-025-07751-6 . hal-05495071

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# Standing on the shoulders of giants – what we should learn from past soil chemists to address the current challenge of “legacy-phosphorus”

David Yalin ◦ · Mikhail Borisover · Tamar Lavi · Frédéric Gérard

Received: 23 February 2025 / Accepted: 17 July 2025 / Published online: 13 August 2025
© The Author(s) 2025

**Abstract** Decades of excessive phosphorus (P) fertilizer application in intensive agriculture have left significant residues of plant-unavailable soil P, commonly referred to as “legacy-P.” There is an urgent need to utilize legacy-P in agriculture, both because fossil P reserves are dwindling and because P leakage from agricultural systems into the environment already exceeds planetary boundaries. In this opinion paper, we advocate against acidification as a “one-size-fits-all” solution for releasing legacy-P, because legacy-P can be bound in various chemical forms that respond differently to this strategy. To guide future plant physiologists and soil microbiologists in developing site-specific solutions, we collate practical methodologies to distinguish between soil systems dominated by adsorption-desorption compared to precipitation-dissolution processes, and propose potential strategies to promote P release in each system. We also aim to rekindle interest in mechanistic studies of P in soils, paying tribute to Jim N. Barrow

and Uzi Kafkafi—two key figures in advancing soil P chemistry who passed away in 2024. We argue that the phenomenon of P adsorption-desorption hysteresis, which they both explored, remains enigmatic, with practical implications for how we study and model legacy-P behavior today. Furthermore, we highlight that knowledge gaps remain even regarding the role of soil pH in P dynamics, and outline a path forward, including practical directions to address the open questions. Above all, we wish to emphasize that advancing the sustainable management of legacy P will benefit from interdisciplinary collaboration between soil chemists, microbiologists, and plant physiologists, drawing on insights gained through decades of rigorous research.

**Keywords** Legacy phosphorus · Soil inorganic phosphorus · Phosphate · Adsorption desorption · Hysteresis phenomena · Precipitation dissolution · Interdisciplinary nutrient management

Responsible Editor: Hans Lambers.

D. Yalin (✉) · M. Borisover · T. Lavi
The Agricultural Research Organization (Volcani Institute), Rishon LeZion, Israel
e-mail: Yalin@volcani.agri.gov.il

T. Lavi
The Hebrew University of Jerusalem, Rehovot, Israel

F. Gérard
INRAE, UMR Eco&Sols, Montpellier, France

## A challenging legacy

Phosphorus (P) is an essential element for plants, recognized as such by Justus von Liebig over 150 years ago (Von Liebig 1855). With the advent of modern agriculture, P fertilizers have been increasingly used to supply production requirements. Nowadays, rock phosphate is the main source of P fertilizer with about 220 Mt of rock phosphate being mined annually
worldwide, to yield approximately 18 Mt P fertilizer per year (Lu and Tian 2017; Illakwahhi et al. 2024). This extensive mining has led to some warnings regarding the depletion of global P reserves within the coming century (Cordell et al. 2009). Despite more reassuring estimates of P reserves from other sources (Illakwahhi et al. 2024), the concentration of 85% of these reserves in only five countries still raises concerns regarding food security and economic disparities for countries lacking fossil-P reserves (Brownlie et al. 2021). Environmental concerns also arise from P use in agriculture, because P leakage into the environment has already exceeded the planetary boundaries to levels estimated to be of high risk (Richardson et al. 2023). Paradoxically, despite these issues and the costs of excessive P use for farmers, the P use efficiency (PUE, Eq. 1) in agriculture is typically low, leading to the build-up of the unused P in soils, as what is referred to as legacy-P (Sharpley et al. 2013) or residual-P (Zhou and Margenot 2023; Turner and Kim 2024; Shober et al. 2024).

$$PUE (\%) = \frac{\text{Amount of P taken up by crops (Mg P ha}^{-1}\text{)}}{\text{Amount of P applied as fertilizer (Mg P ha}^{-1}\text{)}} \times 100\% \quad (1)$$

Recent work by Zou et al. (2022) provided global estimates of PUE and the corresponding amounts of P retained in soils. Trends in PUE for the top five global consumers of P and the leading European consumer, France (Lu and Tian 2017; Eurostat 2025; Fig. 1), reveal a decreasing PUE in India and China between 1965–2015, so that as of 2015 the average PUE in these countries was ~50%. The driver behind the low PUE in these countries, as well as in Brazil, is the highly reactive nature of P in soils, which limits the availability of applied

fertilizer to crops. Recent work in Brazil, for example, showed that even after 30 years of extensive P application, soils still had a high P affinity, effectively capturing half of the applied P fertilizers (Roy et al. 2017). Accessing legacy-P in such P-limiting soils therefore poses a substantial agronomic challenge and will require the development and application of novel strategies (Pavinato et al. 2020). France, the USA, and Canada represent contrasting conditions, with an increasing PUE to above 100% by 2015 (i.e., null excess P and some tapping into legacy-P). This improved PUE can be ascribed to improved agronomic practices in these countries. However, it may also be related to the history of heavy P applications leading to soil P saturation (Van Doorn et al. 2024). In P-saturated soils, tapping into legacy-P reserves may be less challenging, at least in the first years (Koopmans et al. 2004). Zou et al. (2022) estimated based on annual P budgets that in France ~1200 kg P ha⁻¹ was accumulated due to excess application between 1965–2015. This could potentially supply ~50 years of crops at current P uptake rates. However, this does not account for soil processes which may inhibit P release to crops. McDowell and Haygarth (2025) analyzed global soil survey data and estimated plant-unavailable P reserves—defined as P not extractable by the Olsen method (Olsen et al. 1954)—to be sufficient to sustain crop production for 148–186 years at current P uptake rates. According to their work Olsen-extractable P composed only ~4% of total P, emphasizing the fact that utilizing these reserves poses a significant challenge for the future.

Notably, legacy P can also accumulate as the result of intensive application of animal manures to

<figure><img src="image_2.png" /><figcaption>**Fig. 1** PUE trends from 1965 to 2015 for the five countries with the highest global P consumption and the leading European consumer, France (classification as of 2013, following Lu and Tian 2017; Eurostat 2025). Data from Zou et al. (2022)</figcaption></figure>
agricultural fields in countries with large industrial
livestock sectors, such as the Netherlands (Schipanski
and Bennett, 2012). In these cases, organic P forms
can accumulate in soils (Stutter et al. 2015). While
this is also of great importance, here we largely limit
our discussion to inorganic P (IP) fertilizer applica-
tion and the reactions of IP in soils.

One size cannot fit all (soils)

Some of the leading strategies proposed by plant and
soil microbiology experts to improve PUE include the
development and wider adoption of P-efficient crops
and the engineering of soil microbiomes through
inoculation with P-solubilizing bacteria (PSB) or
mycorrhizal fungi, as discussed in recent reviews
(Menezes-Blackburn et al. 2018; Hallama et al. 2019;
Cheng et al. 2023). A wide array of mechanisms can
be harnessed to increase legacy-P availability in soils
(Textbox 1); however, much of the work cited in these
reviews focused on plant and microbial strategies that
rely on acidification of the soil matrix. Acidification
may have positive effects on plant P uptake (Bar-
row 2017), which is advantageous in itself. However,
the effects of acidification on P solubility are soil-
dependent and may, in certain contexts, even reduce
P levels in solution — as demonstrated for some
sandy soils in Gustafsson et al. (2012). Erel et al.
(2017) showed that variations in soil constituents,
such as lime content, even within a single site, can
significantly influence the effectiveness of plant strat-
egies in mobilizing soil P. We therefore advocate for
plant physiologists and soil microbiologists to move
beyond this‘one-size-fits-all‘approach and develop a
broader suite of tools tailored to diverse soil environ-
ments. To that end, in this section we briefly describe
the main interactions between P and various soil con-
stituents, referred to (often) synonymously in the lit-
erature as binding, fixation, or sorption processes (the
latter also used to describe the specific adsorption
to surfaces). In Textbox 1, we then propose simple
accessible methods to identify the dominant P-bind-
ing interactions in soils and suggest mechanisms
effective at counteracting these binding interactions.

Soluble IP ions, and especially the H₂PO₄⁻ species, are the primary form of IP that can be taken up by plants (Barrow 2017). For this active transport into the root to occur a critical IP concentration is

required at the root surface. This value differs accord-
ing to plants and is typically above 1 µM (Lambers
2022). Notably, to sustain P uptake, these soluble IP
levels must be replenished by a sufficient IP flux to
the root surface (Lambers 2022). One of the major
limiting factors for this flux in soils is the transi-
tion of IP into soluble forms (Hinsinger et al. 2011;
Gerke 2024). We make the distinction here between
two basic physico-chemical processes which hinder
IP solubilization: adsorption and precipitation (Penn
and Camberato 2019). These are expected to prevail
to different extents depending on soil pH (Fig. 2).

Adsorption is a relatively rapid binding process to
surfaces of metal (hydr)oxides (in soils, mainly Fe
and Al) and clay minerals, such as kaolinite, illite,
and montmorillonite (Arai and Sparks 2007; Gérard
2016). Adsorption involves electrostatic interac-
tions of IP with mineral surfaces, ligand exchange
and formation of inner-sphere complexes (Fig. 2).
The higher affinity of IP to the particle surface and
involvement in the formation of inner-sphere com-
plexes, as compared to chloride ions for example, that
form outer-sphere complexes, leads to the use of the
terms specific-sorption or chemisorption to describe
this process. However, other inorganic and organic
anions such as sulfate, silicate, oxalate and malate
can also compete for the same surface sites (Nagara-
jah et al. 1970; Lopez-Hernandez et al. 1986; Kafkafi
et al. 1988).

The concentration of soil-adsorbed IP typi-
cally follows a saturation-like curve, suggesting
a finite adsorption capacity of the soil, and is often
described by a Langmuir-type equation. The Freun-
dlich model is also used in the literature, although
it does not assume a saturation limit. Soils saturated
with IP show reduced binding of added IP, poten-
tially increasing P plant-availability, but also the
risk of IP leaching into the environment (Van Doorn
et al. 2024). The phosphate binding capacity (PBC,
the maximum phosphate a soil can adsorb) varies
between adsorbing minerals and also changes with
pH (Gérard 2016). The changes in PBC with pH
(especially in the case of clay minerals) are non-triv-
ial due to variations in the surface charge and IP spe-
ciation (Barrow 2017). A “sorption envelope” is typi-
cally used to describe the integrated effect of these
two processes, typically characterized by minimal
PBC values at high and low pHs and maximal PBC
values at middle pH ranges (Barrow 1984; Goldberg
<figure><img src="image_3.png" /></figure>

**Fig. 2** Conceptual illustration of the abiotic processes controlling soluble inorganic P (IP) concentrations in soils. The bottom graph depicts the pH-dependent distribution of H₂PO₄⁻ and HPO₄²⁻ species. At lower pH values (left), H₂PO₄⁻ is the major soluble IP species. In soils with inherently low pH, precipitation of Fe/Al-phosphate minerals is expected to predominate, illustrated here with a simplified representation of Fe and P atoms in the strengite mineral structure. In soils with pH values between ~4 and 8 (center), adsorption of phosphate to metal (hydr)oxides and clay minerals is expected to be the

dominant process. This ligand exchange mechanism, represented here by a schematic metal hydroxide (Me–OH) surface, involves displacement of surface hydroxyl groups by phosphate, a process favored under lower pH conditions. At higher pH values (right), HPO₄²⁻ becomes the dominant soluble IP species. In soils with inherently higher pH, calcite is often present, and precipitation of Ca-phosphate minerals is expected to dominate, represented here with a simplified image of the hydroxyapatite mineral structure

and Sposito 1984; Bar-Yosef et al. 1988). Manipulating soil pH towards the points of maximal PBC values is expected to increase P binding to soil and decrease P availability to crops, and therefore can be counterproductive to the release of legacy P.

Slower processes within the soil matrix can further reduce soluble IP concentrations – inner particle diffusion has been advocated by Barrow as the primary mechanism for long-term decreases in soluble

IP (Barrow 2023). This implies that IP migrates from particle surfaces into the mineral matrix either by solid-state diffusion or through the mineral pores.

Precipitation describes salt-forming interactions between IP and cations such as Ca, Mg, Fe, and Al. For precipitation processes to occur a prerequisite is oversaturation, as indicated by a positive saturation index (SI), described in Eq. 2 where Q is the ion activity product and K is the thermodynamic

<figure><img src="image_4.png" /></figure>
equilibrium constant. Many computer codes such as PHREEQC (USGS), Visual Minteq (Gustafsson 2019), and MIN3P (https://www.min3p.com/) are available to calculate SI based on solution chemistry, an ion activity model, and thermodynamic equilibrium constants for mineral reactions. Notably, precipitation processes do not follow a saturation curve thus precipitation is expected to continue as long as oversaturation (SI > 0) persists.

$$SI = \log(Q) - \log(K) \quad (2)$$

Precipitation is thought to be the primary cause of P fixation in calcareous soils where oversaturation with respect to Ca-P minerals can occur even at low IP concentrations in soil solutions (Lindsay and Moreno 1960). In acid soils, precipitation of Al-P and Fe-P has also been suggested as a limiting factor of P availability (e.g., Gustafsson et al. 2012 for Al-P).

We believe it is essential to distinguish between soils where P dynamics are dominated by adsorption-desorption processes and those where precipitation-dissolution processes dominate, as each responds differently to management strategies (Textbox 1). The response of these processes to pH, for example, is markedly different as demonstrated in Fig. 3,

where we compare adsorption on montmorillonite as the dominant process controlling IP concentrations, with precipitation of Ca-P minerals. In this example, considering a calcareous soil, reducing the pH in a montmorillonite-dominated soil from pH ~ 8 to pH ~ 5 would decrease soluble IP concentration by 10%, whereas in the soil dominated by Ca-P precipitation, it would have an opposite effect, increasing IP concentration by several orders of magnitude. However, the occurrence of precipitation in soils has long been a contentious issue in the soil chemistry community. Barrow, for example, argued fervently against the possibility of P precipitation in soils (Barrow et al. 2021; Barrow 2021a, b). His position was based on extensive work on carbonate-poor soils, where he demonstrated that adsorption-desorption models adequately describe IP dynamics across a wide range of pH values, application rates, and temperatures (Barrow 1980, 2021b). Moreover, several of his studies found that the metal:P release ratios from soils did not correspond to those of known metal-P minerals, suggesting that dissolution was not the source of P release. Against these valid arguments, counter-evidence exists: a recent study identified precipitation as the only plausible mechanism explaining P dynamics

<figure><img src="image_5.png" /><figcaption>**Fig. 3** Effect of pH on dissolved IP concentrations (normalized to the maximum IP concentration for each line). Blue circle symbols represent results measured in an adsorption experiment performed with Na-montmorillonite (Manning and Goldberg, 1996). The maximum concentration of solution IP (i.e., a minimal adsorption) is observed around pH = 3. The black continuous line stands for the solubility of hydroxyapatite at 25 °C in the presence of dissolved calcium (2.06 mM)</figcaption></figure>

corresponding to equilibrium with calcite at pH = 8 and at atmospheric CO₂ partial pressure (pCO₂ = 10⁻³.⁵ atm). Solubility calculations were performed with the MIN3P model (e.g., Mayer et al. 2012; Gérard et al. 2017) for the range above pH = 5 (because below this pH due to high ionic strength dissolved IP values predicted exceed the realistic values for soil, e.g., > 14 mM)
in a calcisol at P concentrations above 40 mg P L⁻¹ (Raymond et al. 2023). Direct evidence for the existence of metal-P mineral phases in soils is now being supplied with an increasing body of work using X-ray absorption near-edge structure (XANES) spectroscopy, reporting distinct signals of Al/Fe/Ca-P in soils (Liu et al. 2015; Andersson et al. 2016; Eriksson et al. 2016; Lucas et al. 2022; Luo et al. 2017; Mosesso et al. 2024; Szerlag et al. 2022; Szerlag et al. 2024). Above all, we cannot see why precipitation processes that are so well-established in aqueous systems and

routinely exploited in P fertilizer production, should not also occur in soils.

Notwithstanding, we agree with Barrow's warning (Barrow et al. 2021) against the erroneous interpretation of sequential extraction methods (e.g. Hedley et al. 1982) as a sole means to distinguish between adsorbed and precipitated P. XANES is costly and not widely available, and therefore we present a combination of accessible methods that can be used to promote better distinction of dominant soil IP binding processes (Textbox 1).

### Textbox 1 Mechanisms to promote P solubilization as a function of the dominant fixation processes in soil

As soils differ in their composition, soluble IP concentrations can either be dominated by adsorption-desorption or by precipitation-dissolution processes. In Table 1, we propose simple indicators to distinguish between the two processes and highlight mechanisms that should be effective to enhance IP release in each case. The eventual choice of strategy should integrate a broad set of considerations including physiological effects: Si additions, for example, proposed as a method to increase IP availability via competitive sorption have been shown to have antagonistic effects on P uptake possibly due to hindrance of root permeability (Hu et al. 2020).

<table><caption>Table 1 Summary of key differences between adsorbed and precipitated IP soil forms</caption><thead><tr><th>Dominant processes of IP in soil</th><th>Indicators of presence</th><th>Efficient mechanism to release</th></tr></thead><tbody><tr><td><em>Adsorption-desorption</em></td><td><ul><li>SI &lt; 0 with respect to common soil P minerals</li><li>Release of substantial parts of soil-IP by alkalinization (with NaOH for example)</li><li>Decreasing P buffering with increasing IP inputs</li></ul></td><td><ul><li>Decreasing the sorption capacity of Al/Fe (hydr) oxides and clay minerals by changing pH according to the sorption envelope (e.g., Gérard 2016)</li><li>Dissolution of adsorbing clay minerals and Al/Fe (hydr)oxides via pH change (e.g., Hermanska et al. 2023 and Oelkers and Addassi 2025, respectively)</li><li>Reductive dissolution of adsorbing Fe (hydr)oxides (e.g., Zhang et al. 2010)</li><li>Dissolution of adsorbing phases via chelation, for instance by citrate (e.g., Henintsoa et al. 2017)</li><li>Introduction of competing anions (such as humic substances) for IP adsorption (e.g., Weng et al. 2011)</li><li>Decrease of synergistic cations (such as Ca²⁺) for IP adsorption (e.g., Devau et al. 2010 for clay minerals, and Weng et al. 2011 for Fe (hydr)oxides)</li></ul></td></tr><tr><td><em>Precipitation-dissolution</em></td><td><ul><li>SI &gt; 0 with respect to common soil P minerals</li><li>Release of substantial parts of soil-IP by acidification (in clay-rich calcareous soils)</li><li>Stoichiometric relations of P to corresponding cations in solution</li></ul></td><td><ul><li>Dissolution of P minerals via pH changes (e.g., Tu et al. 2014)</li><li>Reductive dissolution of Fe–P minerals (e.g., Wang et al., 2023)</li><li>Introduction of organic anions such as oxalate to promote P mineral dissolution (e.g., Wei et al. 2011)</li><li>Dissolution driven by complexation of dissolved Ca, such as by polyphosphates (Erel et al. 2023)</li></ul></td></tr></tbody></table>

## Needed – soil chemists (open questions and issues)

During 2024, we lost two influential figures in soil P chemistry: Jim N. Barrow, a researcher at the University of Western Australia and Uzi Kafkafi, a researcher at the Hebrew University of Jerusalem,

Israel, who both brought about some of the seminal experimental work on P adsorption processes and subsequent modelling efforts (Kafkafi et al. 1967, 1988; Barrow 1979, 2015). In this section we discuss their work and argue that despite the years that have passed and the exhaustive efforts by those scientists
and others (Arai and Sparks 2007; Gerke 2024), there are fundamental questions regarding the interactions of IP in soil which are still unresolved.

## Is P fixation a reversible process?

It is clear that for legacy-P to be released, P-fixation must be reversible. However, very few studies have assessed the true reversibility of P-fixation using plants (e.g., Koopmans et al. 2004). On a mechanistic level, according to chemical conventions, reversibility means that at equilibrium, each elementary process is equilibrated by its reverse process (Gorban and Yablonsky 2011). In regard to the IP sorption process (a term used here to describe fixation via processes other than precipitation) – it is expected that in batch experiments being at equilibrium, the reversibility should be manifested by coincidence between sorption and desorption isotherms. However, in many studies this is not the case, and an apparent hysteresis is observed (also described in the literature for anions other than IP, Barrow 1986; Chen et al. 2009). Figure 4, for example, describes a batch experiment with two types of kaolinite and two types of montmorillonite clays differing in their innate pH. Incubation for 48 h increased IP sorption as compared to 24 h in the two clays at pH~7 but had no continued effect in pH~9. The hysteresis in the clays with lower pH was not lessened even when incubation with the P-void solution was carried out for 168 h. This hysteretic behavior may have practical implications as it implies that P release into a P-void solution such as irrigation/rain water may be negligible. The literature (mostly from the past century), brings different explanations for the hysteretic behavior which highlight opposing views regarding the chemical equilibration of P sorption processes with implications to the possibility of releasing legacy-P.

One explanation for the hysteretic behavior developed and advanced by Barrow (2023) is the “slow” intra-solid diffusion of IP, where the degree of hysteresis reflects the magnitude of this diffusion process. This theory referred IP-soil interactions to an empirical time-dependent sorption isotherm (Eq. 3):

$$S = aC^{b1}t^{b2} \qquad (3)$$

where S and C are the sorbed and solution concentrations of IP in coexisting soil and solution

phases, respectively; t is the time, and a, b1 and b2 are the empirical coefficients. Sorption of IP was stated "reversible": a continuous curve could be drawn to connect sorption isotherms determined on IP-preloaded (and incubated) soil samples when new IP portions were added, and desorption isotherms when that IP-preloaded/incubated soil was mixed with the aqueous (calcium chloride) solution in variable soil-to-solution ratios (Barrow 1983). Nevertheless, it should be noted that such "a combined sorption isotherm" is not reversible in a thermodynamic sense because it does not represent a sequence of equilibrium states reached at varying IP concentrations. This conclusion has important implications for the interpretation of experimental data – as very often batch experiments are only 24–48 h long (Gu et al. 2025; Saidi et al. 2025; Sharma and Kaur 2025). In these cases, the obtained "sorption isotherms" may not necessarily represent reversible changes, and a straightforward application of equilibrium models to these sorption data could be erroneous.

A second explanation for the hysteretic behavior in batch experiments was noted in work by Muljadi et al. (1966), who concluded that the sorption equilibrium on K-kaolinite is reached after 24 h, and mentioned that the hysteresis observed at low IP concentrations could be related to a phase change during IP desorption. Kafkafi et al. (1967) addressed the phase-change hypothesis in more detail by using isotopic tools, measuring adsorption/desorption and re-sorption in different modes, and concluding that: (i) sorption isotherms on kaolinite measured in regular batch experiments were found at equilibrium after 24 h, and (ii) the desorption procedure was the apparent cause of irreversibility and the loss of the ability to undergo isotopic exchange for adsorbed IP. The latter conclusion was further affirmed by Bar-Yosef and Kafkafi (1978): when examining IP desorption on kaolinite by diluting clay mineral suspensions, upon reducing the magnitudes of IP concentration changes, desorption isotherms given sufficient time tended to coincide with sorption isotherms.

Thus, these earlier studies represented two essentially different views of the IP sorption hysteresis, i.e., non-completed sorption kinetics vs time-independent sorption isotherms with apparently slower rates or phase/sorbent matrix changes during desorption. It is worth noting that the b2 coefficient in Eq. 3 can be as small as 0.08 (for phosphate sorption on kaolinitic
<figure><img src="image_6.png" /><figcaption>**Fig. 4** Batch experiment with four standard clay samples (Wyoming Montmorillonite from the Clay Mineral Society, Montmorillonite #31, Kaolinite #4 and Kaolinite #9 from Wards Natural Science Establishment, Rochester New York). Adsorption was performed on a 1% clay suspension with a background of 10 mM KCl and sampled at 24 and 48 h. After</figcaption></figure>

48 h adsorption, single step desorption into KCl 10 mM was performed and sampled at 20 and 168 h. Averages of 3 replicates (error bars are not presented for clarity), illustrate the extent of hysteresis at different P concentrations and different pH ranges

soil, Barrow 2023). This indicates that varying the sorption time by an order of magnitude causes only a 20% change in IP adsorbed concentrations. Therefore, slow IP sorption onto soils could be missed in most studies since its detection may demand experiments extending beyond the commonly used 24 h.

Examining desorption should be more commonly integrated into routine sorption experiments: if no

hysteresis would occur, this would serve as evidence of equilibrium occupation of sorption sites; alternatively, strong hysteresis exposes the possible lack of equilibrium across sorption isotherms. Respectively, demonstrating the lack of hysteresis may be crucial when a fast IP sorption is believed to be separated, in a time scale, from slow processes in soil solution, such as intra-solid diffusion or precipitation.

<figure><img src="image_7.png" /></figure>
Textbox 2 summarizes some experimental tests and expected observations helping to shed light on reversibility of sorption isotherms.

## Textbox 2 Verifying reversibility scenarios: experimental observations and interpretations

<table><thead><tr><th>#</th><th>Observation</th><th>Interpretation</th></tr></thead><tbody><tr><td>1</td><td><ul><li>Sorption isotherms obtained by incremental increases in solution IP concentration coincide with those determined in a series of solutions significantly differing by IP concentrations</li></ul></td><td><ul><li>Attaining equilibria across the isotherm does not depend on the way a certain state is reached thus indicating that sorption is thermodynamically reversible</li></ul></td></tr><tr><td>2</td><td><ul><li>IP concentration positively affects the attainment of sorption equilibrium</li><li>No desorption hysteresis is observed in the linear portion of sorption isotherms</li></ul></td><td><ul><li>Microscopic reversibility is reached in sorption isotherms</li><li>Hysteresis is caused by slower desorption rates compared to sorption, due to finite concentration drop</li><li>In a linear portion of the sorption isotherm, no differences in attainment of equilibrium are expected for sorption and desorption associated with diffusion into or out of a solid matrix and reversible first-order adsorption reactions (Crank 1975)</li></ul></td></tr><tr><td>3</td><td><ul><li>Hysteresis does not disappear in the linear portion of IP sorption isotherms (as in Muljadi et al. 1966)</li><li>Hysteresis caused by a finite drop in solution IP concentrations tends to disappear as the concentration drop is reduced (e.g., as in Bar-Yosef and Kafkafi, 1978)</li></ul></td><td><ul><li>Desorption hysteresis is not related to concentration-induced retardation of kinetics</li><li>Matrix changes occur due to finite changes in IP concentrations during desorption</li></ul></td></tr><tr><td>4</td><td><ul><li>Resorption isotherm initiated at the end of a desorptive sequence coincides with the desorption isotherm (as in Okajima et al. 1983)</li></ul></td><td><ul><li>Matrix changes lead to desorption isotherm revealing "reversibility", thus suggesting the applicability of an equilibrium approach to the metastable state of a sorbent</li></ul></td></tr></tbody></table>

Irreversible changes in the sorbent matrix during desorption can indeed lead to IP entrapment preventing its exchange with the environment (observed by Kafkafi et al. 1967). The changes in the sorbent matrix induced by elevated sorbate concentrations and persisting during desorption could also lead to the formation of a sorbent metastable state (Borisover 2019), that would allow IP to exchange and equilibrate with solution on newly formed surfaces and pores. The latter vision associated with irreversible pore deformation and non-elastic swelling was related to hysteretic sorption of organic compounds by natural organic matter, clay minerals, polymers and charcoals (Bai et al. 2024; Braida et al. 2003; Sander and Pignatello 2005; Thommes et al. 2015; Xing and Pignatello 2005). In this case, hysteresis is observed albeit no entrapment of a solute is expected (Sander and Pignatello 2005). Hence, a potentially reversible successive desorption isotherm associated with a metastable sorbent state could be identified, differing from the equilibrium sorption isotherm (as observed by Okajima et al. 1983 for IP sorption onto a soil). Textbox 2 further suggests expected observations that can help elucidate irreversible changes in a sorbent following IP desorption. Both

scenarios associated with irreversible changes in sorbent matrix are of importance for understanding whether "poorly" desorbed soil IP will be available for plants: (i) the entrapped (indeed fixed) IP is prevented from its release whereas (ii) the IP adsorbed onto newly modified (metastable) surfaces may be released, yet showing stronger interactions with a sorbent compared to an adsorption isotherm. It is of interest that one possible mechanism of formation of such metastable states in soil during sorption of IP and persisting during its desorption could be explained based on the idea (Hiemstra and Zhao 2016) that metal (hydr)oxides in soil may transform during IP sorption into smaller clusters capable of binding elevated IP amounts. Upon reducing adsorbed IP concentration, these clusters with higher surface area would most probably remain in their metastable state and eventually lead to IP sorption hysteresis.

### How pH affects adsorption-desorption

Barrow (1984) rigorously investigated the effect of pH on IP sorption in soils. He proposed a single site adsorption model to infer that the overall decrease of P sorption with pH is a result of the build-up of
surface negative charges (i.e., decrease of surface potential). Since then, Weng et al. (2011) investigated a wide range of soils and pH values and successfully simulated the observations using surface complexation modelling with provisions for electrostatic interactions by the competition between dissolved organic matter (DOM) and IP on some representative Fe (hydr)oxides. Devau et al. (2009, 2011) modelled similar observations in near-neutral soils also using surface complexation with provisions for electrostatic interactions and the inclusion of clay minerals and gibbsite but did not regard DOM. More recent modelling work has showed the intricate pH-dependent competitive effect of DOM on P sorption (Bai et al. 2024). Interestingly, recent experimental work has also suggested increased IP adsorption in the presence of DOM (Freiberg et al. 2022, 2023). From these it arises that in real-world soils, a combination of all these processes would be required because: (i) competitive or cooperative sorption of IP in the presence of DOM cannot be ruled out as it notably depends on DOM properties/concentration/speciation, and reactivity to mineral surfaces (e.g. Kaiser and Guggenberger 2003); (ii) clay minerals such as kaolinite and illite cannot be ruled out as they do have the capacity to bind comparable IP amounts to Fe (hydr)oxides (Gérard 2016), as confirmed by XANES in clay-rich soils (e.g., Andersson et al. 2016; Antonangelo et al. 2020; Prietzel et al. 2024).

Mineral precipitation during IP adsorption experiments is a matter that has often been neglected in adsorption studies. Recent work assessed the SI of Ca-P minerals (the most stable/least soluble being apatite) and Al- and Fe-bearing P minerals such as variscite and strengite (e.g., Devau et al. 2011; Weng et al. 2011; Gustafsson et al. 2023). Weng et al. (2011) for instance found that the apparent increase of P adsorption observed in some soils below pH=4 and above pH=8 may be due to the precipitation of Al-P and Ca-P minerals, respectively. Conversely, the occurrence of mineral P dissolution during desorption experiments has not been widely explored so far, although this is likely to occur at some rate in most soils, especially between mildly acidic and near alkaline pH conditions, i.e., where the reacting solutions are less favorable to the formation of Ca-P and Al/Fe-P minerals, respectively (Fig. 2). Moreover, P mineral dissolution rate increased with acidity in the acid range due to the catalytic effect of H⁺. The rate is

constant at near-neutral pH and further increases with pH in the alkaline range due to the catalytic effect of OH⁻ (e.g. Oelkers and Addassi 2025). The dissolution of IP sorbents such as clay minerals and Fe/Al (hydr)oxides can certainly be an important process for the solubilization of legacy-P (Brand-Klibanski et al. 2007), but has been seldom considered in IP sorption experiments. An increase in concentrations of dissolved Si, Al, Fe concomitantly to IP strongly suggests that this process occurs when citrate is added to soil (Henintsoa et al. 2017). Dissolution rates of clay minerals and Al (hydr)oxide are influenced by pH in a similar manner to P minerals (Heřmanská et al. 2023; Oelkers and Addassi 2025). Accordingly, the same mechanistic rate law is used to express the dissolution kinetics of all these minerals:

$$r_d = k_i a_i^n (1 - (Q/K)^m) \quad (4)$$

where K and Q are as defined in Eq. 2, r_d is the dissolution rate (mol/s), k_i the rate constant (mol/s), and a_i, the activity of aqueous species i, n_i a corresponding experimental constant (with i corresponding to H⁺, H₂O or OH⁻ depending on pH range), and m is another experimental constant (e.g., Heřmanská et al. 2023).

## Lessons for the road ahead

Overall, this work wishes to convey that addressing the challenge of mobilizing legacy-P for crop production would benefit from interdisciplinary collaboration among experts in plant physiology, soil microbiology, and soil chemistry. Such collaborations can generate a diverse toolkit adaptable to various soil conditions. In textbox 1, we aimed to trigger such collaborative work by presenting simple methods, which are currently little in use, and can be useful for detecting dominant processes controlling IP availability in specific soils.

Our brief overview of the works of Barrow and Kafkafi nourished by other studies stresses that explicitly addressing and explaining hysteresis is important for the mechanistic interpretation of macroscopic adsorption-desorption data. Furthermore, their work highlights that classical chemical methodology can produce insights on the mechanisms and processes likely to control P reactions in soils. Notwithstanding, we see the great potential in advancing this knowledge also by modern techniques used in environmental science. For
example, we mentioned how XANES can be used to unravel the form under which P is fixed during long-term P sorption experiments. Such experimental data can support the development of the next generation of reactive transport models to simulate P availability in soils systems using recently developed mechanistic tools (e.g., Bingham et al. 2020). Some current models are already quite realistic and relevant for P sorption as they can cope with ion-specific diffusion coefficients (Rasouli et al. 2015) and multi-porosity diffusion problem in the presence of charged surfaces (Tertre et al. 2024).

Another direction to pave the path forward for better understanding of P interactions with soil and more efficient P use, relies on computational chemistry using atomistic (and quantum-chemical) modeling (as reviewed by Kruse et al. 2015, and more recently by Ahmed et al. 2022). For example, Ahmed et al. (2024) performed molecular dynamics modeling of inorganic (and organic) phosphates on goethite surfaces in the presence of model organic compounds (citric acid and histidine), using force-field molecular mechanics previously validated against corresponding quantum mechanics/molecular mechanics simulations: inhibiting effects of organic anions on P sorption on mineral surfaces were revealed. Other studies have been performed to explain the competitive effects of organic anions (presumed to represent soil organic matter components, Ohno and Kubicki 2020) and silicates (Chen et al. 2022) on phosphate sorption.

Much of this atomistic modelling was centered on interaction of IP with metal oxyhydroxides and oxides, with little focus on clay minerals. In light of multiple experimental studies both mentioned in this opinion and present in the literature that demonstrate the significance of clay minerals in soil IP sorption, it seems that more work is desirable on atomistic modeling of IP interactions with clay minerals, in particular, including those with different swelling ability and accounting for the presence in soil solutions of multivalent cations capable of bridging inorganic phosphate anions to negatively charged clay mineral surfaces. A further challenge is to gain more insight into the role of organic matter in IP sorption in soils. Small molecules, such as aliphatic, aromatic and amino acids, explicitly considered in atomistic simulations, albeit providing useful insights into the possible role of organic molecules in IP sorption in soils may be too simplified to represent water-soluble soil organic matter, which is composed of

multiple molecules of variable size and chemical composition. One way of filling this critical gap can be use of approaches similar to that presented in the Vienna Soil Organic Matter Modeler (VSOMM), allowing to systematically generate multiple structures of humic substances of variable size and composition from the library of building blocks (Escalona et al. 2021).

If integrated properly, future models have the potential not only to advance legacy-P utilization but also to improve P use efficiency and inform policy-making at regional and global scales. This has particular importance in assessing the implications of climate change and guiding necessary adaptations to P fertilization for future climates.

**Authors contribution** DY, MB, and FG have all contributed to the conception and design of this paper; TL has contributed to data collection and analysis.

**Funding** Open access funding provided by The Agricultural Research Organization of Israel. The authors declare that no funds, grants, or other support were received during the preparation of this manuscript. The authors have no relevant financial or non-financial interests to disclose.

### Declarations

All authors read and approved the final manuscript. The datasets generated and/or analyzed during the current study are available from the corresponding author on reasonable request.

**Open Access** This article is licensed under a Creative Commons Attribution 4.0 International License, which permits use, sharing, adaptation, distribution and reproduction in any medium or format, as long as you give appropriate credit to the original author(s) and the source, provide a link to the Creative Commons licence, and indicate if changes were made. The images or other third party material in this article are included in the article's Creative Commons licence, unless indicated otherwise in a credit line to the material. If material is not included in the article's Creative Commons licence and your intended use is not permitted by statutory regulation or exceeds the permitted use, you will need to obtain permission directly from the copyright holder. To view a copy of this licence, visit http://creativecommons.org/licenses/by/4.0/.

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