From Reactive Remediation to Predictive Prevention: A Systems Synthesis of Global Soil Salinization and Sodicity

ABSTRACT

Soil salinization and sodicity are accelerating worldwide, reducing crop productivity across more than 160 million hectares of farmland and causing production losses approaching USD 27 billion annually. Despite decades of research documenting these impacts, the extent of salt‐affected soils continues to expand. Salt‐affected soil degradation emerges from feedbacks among physical, biological, and socioeconomic processes that operate across scales and resist intervention when addressed in isolation. This review synthesizes current understanding of these coupled feedbacks. We trace how ion accumulation, soil structural decline, microbial community disruption, and plant stress interact with irrigation practices, land use decisions, and governance systems. These interactions create threshold dynamics in which degradation can appear stable for years before rapid decline. Evidence from diverse regions reveals that effective restoration depends on coordinated interventions matched to the complexity of the system itself. Four research priorities emerge for advancing prediction and prevention. First, quantifying the thresholds that separate reversible from recalcitrant degradation across soil types and climates. Second, expanding predictive capacity through integrated sensor networks and data platforms that enable early warning. Third, improving understanding of how institutional structures influence adoption of restoration practices. Fourth, integrating biophysical and governance models to enable scaling from field trials to basin‐wide implementation. By coupling mechanistic understanding with monitoring infrastructure and adaptive governance, reactive remediation can give way to proactive prevention.

Salty soils are spreading across the world's farmland, often building up unnoticed for years before crop yields fall sharply and repair becomes slow and costly. Our review shows that this damage comes from linked problems in soil chemistry, water management, and policy, so fixing any one of them alone rarely works. Watching for early warning signs and acting before soils cross a tipping point can protect farmland at far lower cost than restoring it later.

1. Introduction: Salinization as an Emergent Systems Crisis

Soil salinization is degrading cropland productivity worldwide, threatening farmer livelihoods and food security across about 10% of global croplands (FAO ²⁰²⁴) (Figure 1). The problem is accelerating, and climate change is projected to expand salt‐affected area further across many drylands (Hassani et al. 2021). Soil salinization has many origins including seawater intrusions, mobilization of subsurface salts, and intensification of surface salts due to irrigation and land use changes. Salinity is particularly prevalent in irrigated regions, affecting an estimated 20%–33% of the 320 million hectares under irrigation (Qadir et al. 2014; Mohanavelu et al. ²⁰²¹). Farms in these regions typically produce high yields and economic returns because irrigation alleviates crop water stress. However, salt accumulation in the root zone creates a paradoxical water stress as it generates osmotically driven water deficits near the roots (Muhammad et al. 2024). Soil salinization in irrigated systems causes direct crop production losses of about $27.3 billion annually, with substantially higher total economic costs due to remediation expenses, environmental externalities including health impacts from dust emissions, land value depreciation, and social displacement (Qadir et al. 2014).

FIGURE 1.

Global map of soil salinity. Values shown are mean modeled topsoil electrical conductivity values in the period from 1980 to 2024. Data are based on multiple global soil profile datasets using soil observations from the upper 0–30 cm and excluding deeper samples. Observational coverage was sparse in northern Africa, the Arabian Peninsula, and parts of the Middle East; modeled values in these regions therefore carry greater uncertainty. The scale reflects electrical conductivity in decisiemens per meter (dS m⁻¹). Reproduced from Wang et al. (2025) with permission from Elsevier. Map lines delineate study areas and do not necessarily depict accepted national boundaries.

Conventional responses to salinization remain largely reactive, treating symptoms rather than underlying causes. Flushing salts from the root zone or applying chemical amendments can provide temporary relief but rarely address the hydrological and management imbalances that drive salt accumulation (Devkota et al. 2022). These isolated interventions fail because salinization emerges as a property of coupled human–natural systems where management decisions at one scale propagate through hydrological feedbacks to create degradation at other scales. Interactions among physical, biological, and socioeconomic drivers create threshold dynamics in which systems can appear stable for years before crossing tipping points.

The urgency of the salinization crisis stems from these hidden dynamics, as salinity often develops gradually and remains undetected for years, masked by short‐term management responses such as increased irrigation or crop substitution. Productivity can then decline rapidly once ion concentrations exceed plant physiological thresholds or cause soil structure collapse (Demo et al. 2025; Zörb et al. ²⁰¹⁹). Early detection enables interventions before thresholds are crossed, whereas delayed action leads to structural collapse requiring decade‐long recovery timelines that exceed most farmers' investment horizons.

1.1. Coupled Drivers Operating Across Scales

The drivers of contemporary salinization operate as tightly coupled processes across multiple scales (Figure 2). At the local scale, inefficient irrigation and inadequate drainage deliver excess water that raises water tables (Mohanavelu et al. 2021). Water drains from one field, flows downstream, and is reused by another, progressively concentrating salts with each cycle. In closed basins that lack surface drainage outlets, salts accumulate indefinitely in low‐lying areas. As agricultural intensity increases, salt accumulation can exceed soils' natural leaching capacity.

FIGURE 2.

Cross‐scale mechanistic pathways leading to, and reinforcing, salt‐affected soil degradation. Salt‐affected soil degradation emerges from nested mechanisms spanning microscale soil chemistry through field and landscape hydrology to regional and global drivers. The semicircular layers illustrate processes at each scale. Large red arrows represent self‐reinforcing feedbacks that amplify degradation (sequential poor drainage, capillary rise, salt accumulation, and structural degradation). Small blue arrows indicate water movement, small orange arrows indicate salt mobilization, and purple dashed arrows indicate cross‐scale coupling. The orange dashed line marks the critical threshold separating reversible from difficult‐to‐reverse degradation. The far‐right side illustrates seawater intrusion. N₂O, nitrous oxide. The cross‐scale nature of these processes illustrates why single‐scale interventions fail and cross‐scale coordination is necessary.

Beyond surface irrigation dynamics, subsurface processes, particularly groundwater extraction, also amplify salt mobilization. In coastal areas, removing freshwater from shallow aquifers reduces hydraulic pressure, allowing seawater intrusion into those aquifers and beyond, rendering groundwater and surface water unsuitable for irrigation (Shokri et al. 2024). Inland, over‐extraction of groundwater initially lowers water tables, temporarily reducing capillary rise of salts to the surface. However, this apparent benefit is short‐lived because a deeper water table also means less water moving through the soil profile to leach salts downward and away. As a result, salts delivered by irrigation, whether sourced from surface water or imported supplies, accumulate near the surface rather than being flushed to depth, intensifying salinization. These feedbacks demonstrate how water management transforms local hydrology with cumulative impacts on soil function.

Land‐use change compounds water‐driven processes by disrupting the hydrological balance that once buffered salt accumulation. Deep‐rooted perennial vegetation maintains low water tables through high evapotranspiration, keeping salts sequestered in the subsoil. When replaced with shallow‐rooted annual crops, evapotranspiration declines and less water is extracted from depth, allowing water tables to rise (Barrett‐Lennard ²⁰⁰²; Cañedo‐Argüelles et al. ²⁰¹³). Rising water tables mobilize salts stored in the subsoil and transport them upward through capillary action to the root zone. This mechanism explains widespread salinization following conversion of native grasslands and woodlands to cropland in dryland systems (Stavi et al. 2021). In irrigated systems, bare fallow periods create a similar risk: water extraction by crops ceases while recharge continues, raising water tables and mobilizing deep salts.

Climate change further amplifies these anthropogenic pressures (Shokri et al. 2024; Kramer et al. ²⁰²⁵). Rising temperatures increase evapotranspiration and promote upward movement of saline groundwater (Corwin 2021). Altered precipitation patterns—longer dry periods and more intense rainfall events—diminish natural leaching effectiveness and accelerate soil structure degradation (Hassani et al. 2021; Mukhopadhyay et al. ²⁰²¹). In coastal regions, sea‐level rise and storm surges drive saltwater farther inland, contaminating estuaries and low‐lying agricultural zones (Shokri et al. 2024). Combined warming, altered rainfall, and sea‐level rise expand both the spatial extent and persistence of salinization (Hassani et al. 2021), with rates projected to rise by up to 10% annually (FAO ²⁰²⁴).

Regional manifestations of climate‐driven processes vary in intensity but converge to amplify salinization risk globally. In South Asia, rising evaporative demand and erratic monsoons are accelerating salt accumulation across the Indo‐Gangetic Plains, where 30% of Pakistan's agricultural land now exhibits salinization and sodicity (FAO ²⁰²⁴). Australia's Murray‐Darling Basin is facing a projected hotter, drier, and more variable climate, with increased drought frequency projected to elevate salinity in the western basin despite decades of active management (Evans et al. 2021; Hart et al. ²⁰²⁰). Projected end‐of‐century precipitation declines of up to 40% in southern and western Australia, together with winter declines along Mediterranean coasts, would amplify salt accumulation in already water‐stressed agricultural systems (Hassani et al. 2021). Global projections identify southern and western Australia, Mexico, the southwestern United States, and South Africa as hotspots where climate‐induced salinization will intensify most rapidly through mid‐century (Hassani et al. 2021).

Global syntheses of salinization are biased by an uneven evidence base. Mapping and monitoring data are densest where long‐running programs and field networks exist, including Australia, South Asia, China, and parts of the United States and Mediterranean Europe, and sparsest in several highly salinization‐prone regions. This uneven coverage propagates into mapping uncertainty, with global hotspot identification and area estimates more reliable in well‐sampled regions than in data‐poor ones. Conclusions drawn at the global scale should therefore be treated as provisional where ground data are scarce, and expanded monitoring in under‐represented regions should be prioritized to improve the reliability of global assessments.

Together, irrigation intensification, groundwater depletion, sea‐level rise, land‐use conversion, and climate shifts are transforming salinization from a local agronomic problem into a global environmental crisis that demands proactive, model‐informed interventions integrating land management, hydrology, and climate dynamics (Hassani et al. 2021; Kramer et al. ²⁰²⁵).

This review advances beyond existing syntheses of soil salinity and sodicity in three respects. First, rather than cataloguing remediation measures, it integrates mechanisms across scales, linking microbial physiology and plant‐microbe interactions to field‐scale reclamation and, in turn, basin‐scale hydrology and governance. Second, it organizes this evidence around thresholds, feedbacks, and early warning, providing a framework for transitioning from reactive remediation to predictive prevention. Third, it brings the biological, engineering, remote‐sensing, and policy dimensions into a single framework. Existing reviews are organized around single disciplinary axes: irrigation‐induced salinity and sodicity hazards (Mohanavelu et al. 2021), sensor‐based assessment of climate‐driven salinity change (Corwin 2021), and the physics of salinization from global to pore scale (Shokri et al. 2024). Shokri et al. (2024) integrate across spatial scales but remain within process hydrology, whereas the integration provided here is across disciplines. Together, these elements position the review as a systems‐level synthesis intended to inform anticipatory, model‐guided management.

We use “salt‐affected soils” as the umbrella term throughout this review. “Salinity” refers to elevated soluble salts and the associated osmotic and specific‐ion stress, whereas “sodicity” refers to excessive exchangeable sodium and the associated structural degradation. Saline‐sodic soils exhibit both conditions, and alkalization refers here to the high‐pH, carbonate‐dominated expression of sodicity. Table 1 summarizes the diagnostic criteria that distinguish these classes, and we apply these definitions consistently in the sections that follow.

TABLE 1.

Diagnostic classification of salt‐affected soils.

Soil class	ECe (dS m−1) a	SAR a	ESP (%) a	pH	Dominant ions and diagnostic features
Saline	≥ 4	< 13	< 15	< 8.5	Neutral soluble salts (chlorides and sulfates of Na+, Ca2+, Mg2+). Structure generally remains flocculated; plant growth is limited mainly by osmotic and specific‐ion stress.
Sodic	< 4	≥ 13	≥ 15	8.5–10	Exchangeable Na+ with carbonates and bicarbonates. Structure is poor, with clay dispersion and low infiltration; plants experience Na+ toxicity and nutrient imbalance.
Saline–sodic	≥ 4	≥ 13	≥ 15	usually ≤ 8.5	High soluble salts together with exchangeable Na+. Structure may remain flocculated while electrolyte concentration is high, but can disperse when salts are leached without an adequate calcium source.
Alkaline (carbonate‐dominated)	variable	≥ 13	frequently ≥ 15	> 8.5 (often 9–10)	Extreme end of sodic soils, dominated by sodium carbonate and bicarbonate. Structure is severely degraded; plant growth is limited by restricted availability of P, Fe, and Zn due to very high pH.

Note: Criteria used in this review to distinguish saline, sodic, saline–sodic, and alkaline soils, with the dominant ions and diagnostic features of each class.

^a ECe, electrical conductivity of the saturation extract; SAR, sodium adsorption ratio; ESP, exchangeable sodium percentage. Boundary values are diagnostic conventions that follow the USDA Salinity Laboratory classification (Richards 1954) and FAO guidance.

1.2. Reframing Salinization as a Systems Problem

Understanding why isolated interventions fail requires examining the feedbacks that connect physical, biological, and socioeconomic components of the coupled system. In this review, we first characterize the scope and cascading impacts of salinization across scales (Section 2), then analyze the mechanistic pathways through which soil structure, microbial function, and plant physiology interact to generate threshold dynamics (Section 3). We then evaluate solution pathways that span field‐level interventions, landscape‐scale management, and governance frameworks (Section 4), before identifying the research priorities necessary to shift from reactive remediation to predictive prevention (Section 5). The challenge is not simply generating more knowledge within each discipline but building integrated frameworks that identify where and how to intervene.

2. Scope and Impacts of Salinization

Salinization manifests as impacts that cascade from field to global scales. At the field scale, salinity directly reduces crop productivity through osmotic and ionic stress (Zörb et al. 2019), the physiological mechanisms of which are detailed in Section 3.3. As stress intensifies, plants allocate more energy to maintaining osmotic homeostasis, leaving less energy for growth and reproduction (Munns and Gilliham 2015). Micronutrient availability is also disrupted: under saline‐sodic conditions with elevated pH, manganese and iron become less available due to reduced solubility, while in poorly drained saline soils with low redox potential, anaerobic conditions can mobilize these same micronutrients to potentially toxic concentrations (Mishra et al. 2023). In irrigated systems, yield losses are compounded by reduced water and nutrient use efficiency, creating a cycle of declining profitability and increasing input dependence.

Beyond individual fields, salinization alters landscape‐scale soil integrity. The structural degradation described mechanistically in Section 3.1 reduces rainfall absorption, promotes runoff, and increases erosion (Zaman et al. 2018; Demo et al. ²⁰²⁵). Sediments and salts carried by overland flow accumulate in drainage systems and surface waters, degrading water quality and aquatic habitats. Vegetation decline reduces organic matter inputs, further suppressing nutrient cycling rates and microbial carbon processing efficiency (Haj‐Amor et al. ²⁰²²; Rietz and Haynes ²⁰⁰³). Ultimately, salinization can convert productive land into sparsely vegetated or barren surfaces that generate dust and heat, intensifying local aridity and contributing to dust‐related health problems (Papatheodorou and Achilleos 2026).

At regional and national scales, the economic and social consequences of salinization are profound. Declining yields undermine farm income, reduce food security, and erode the economic viability of irrigated agriculture (Qadir et al. 2014). Farmers often respond by intensifying irrigation or expanding into marginal lands, inadvertently accelerating degradation in those marginalized lands. Remediation is constrained by high costs, limited access to technology, and insecure land tenure. The cumulative result is a loss of production capacity that reduces export revenues and rural employment while increasing dependence on food imports. In severely affected regions, persistent salinization drives rural displacement and land abandonment, reinforcing cycles of poverty and ecological decline.

The ecological effects of salinization extend beyond croplands. Rising salt concentrations in groundwater and surface water threaten biodiversity in wetlands, estuaries, and riparian zones (Shokri et al. 2024). Salt‐tolerant species often replace native vegetation, simplifying community structure and altering ecosystem processes. Reduced plant cover diminishes carbon inputs to soil and accelerates mineralization, weakening terrestrial carbon storage (Li et al. 2023). In coastal systems, saline intrusion interacts with sea‐level rise and nutrient pollution, amplifying eutrophication and habitat loss. Across climatic zones, these processes collectively degrade the capacity of ecosystems to buffer hydrological extremes and support biological productivity.

At the Earth system scale, widespread soil salinization alters climate‐relevant surface properties. Vegetation loss increases albedo and shifts surface energy balance, intensifying local warming. Exposed saline surfaces generate dust plumes that degrade air quality and alter radiative forcing. Degraded soils lose the capacity to regulate water and carbon fluxes, reducing both agricultural resilience and climate mitigation potential through diminished carbon sequestration. Although direct greenhouse gas emissions from salinized soils remain poorly quantified, the functional loss of these ecosystems clearly weakens terrestrial buffering of climate variability.

What makes salinization particularly intractable is that degradation is cryptic, often masked by short‐term management responses until reinforcing feedbacks are firmly established. The socioeconomic drivers that accelerate it, including water subsidies, insecure tenure, and fragmented governance, are precisely the factors that prevent an effective response. The mechanistic analysis that follows examines how specific processes at the soil, plant, and microbial scales generate the feedbacks that determine whether systems degrade or recover.

3. Mechanistic Pathways of Soil and Plant Response

Salinity and sodicity disrupt the physical, chemical, and biological processes that sustain plant growth and ecosystem productivity through interconnected pathways operating from molecular to field scales.

3.1. Soil Physical and Chemical Transformations

Physical degradation differs between saline and sodic soils. Salinity primarily imposes osmotic and specific‐ion stress on plants, whereas sodicity alters cation exchange equilibria and drives the structural changes described here. Specifically, as exchangeable sodium accumulates on cation exchange sites, it displaces calcium and magnesium, weakening the electrostatic bonds holding clay particles together (Mishra et al. 2023). The resulting dispersion of clay particles reduces aggregate stability, decreases average pore size, blocks macropores, and diminishes hydraulic conductivity (Stavi et al. 2021). Moreover, in sodic soils, elevated pH enhances organic matter dissolution and reduces its stabilization, leading to loss of soil carbon and reduced nutrient retention capacity (Shokri et al. 2024). Under these conditions, infiltration slows, waterlogging increases, and aeration declines, creating anaerobic microsites that inhibit root respiration and alter redox conditions, mobilizing iron and manganese to potentially toxic concentrations. Reduced permeability also limits leaching, trapping salts in the upper soil horizon and intensifying surface salinity through capillary rise and evaporation. Salinity, in turn, imposes ion‐specific effects that compound soil degradation. Chloride and sulfate contribute to osmotic stress by increasing total solute concentration, while sodium and bicarbonate alter soil pH and ion balance. Soils with high exchangeable sodium percentage (ESP) typically exhibit elevated pH. The combined effects of poor structure, high ionic strength, and altered pH create a self‐reinforcing cycle in which physical and chemical deterioration accelerate one another.

3.2. Microbial and Biogeochemical Responses

Microbial communities respond rapidly to salinity, with cascading effects on nutrient cycling and organic matter dynamics. High salinity disrupts membrane potential, protein folding, and enzyme activity. Microbial growth and respiration rates decline under elevated soil salinity and respond more strongly than microbial biomass (Rath et al. 2019). These declines reflect a fundamental resource allocation trade‐off: under osmotic stress, microbes redirect carbon and nitrogen from growth to synthesizing compatible solutes and stress‐protective molecules, a physiological cost large enough to consume 3%–6% of ecosystem net primary production during a single stress episode (Schimel et al. 2007). Sensitive taxa decline while halotolerant bacteria and archaea proliferate, altering community composition and function (Li et al. 2021). High salinity also decreases microbial activity (Rietz and Haynes 2003), and together these shifts reduce decomposition efficiency and slow nutrient mineralization, thereby decreasing the availability of nitrogen and phosphorus to plants.

Under prolonged salinity, microbial networks become less diverse and more functionally constrained (Zheng et al. 2017), potentially affecting greenhouse gas emissions. Denitrification increases in poorly aerated microsites, producing nitrous oxide, while nitrification and symbiotic nitrogen fixation decline (Li et al. 2021). Prolonged salinity can also disrupt microbial phosphorus mobilization by inhibiting microbial processes that support organic phosphorus mineralization and inorganic phosphate solubilization (Mishra et al. 2023). The decline in microbial processes contributing to soil aggregation, such as microbial activity and biofilm formation, further weakens soil structure, linking biological deterioration to the physical and chemical pathways described above. The net effect of microbial community shifts depends on the severity of salinity and the management context. In severely saline soils (EC > 12 dS m⁻¹), detrimental community‐wide disruptions typically overwhelm any beneficial effects.

3.2.1. Microbial Mechanisms for Salinity Mitigation

Halotolerant Plant Growth‐Promoting Rhizobacteria (PGPR) and Arbuscular Mycorrhizal Fungi (AMF) counteract salinity stress through “induced systemic tolerance,” a suite of mechanisms operating across molecular, cellular, and whole‐plant scales (Bhardwaj et al. 2025; Kumawat et al. ²⁰²³). At the molecular level, many halotolerant microbes produce ACC deaminase, which degrades the ethylene precursor 1‐aminocyclopropane‐1‐carboxylic acid. Because ethylene inhibits root elongation under saline conditions, this enzymatic intervention sustains root growth precisely when plants need expanded soil exploration capacity (Orozco‐Mosqueda et al. ²⁰²⁰; Mahmud‐Ur‐Rahman et al. ²⁰²²). These same microorganisms synthesize indole‐3‐acetic acid (IAA), stimulating lateral root development and increasing absorptive surface area despite osmotic constraints (Khan et al. 2019; Haroon et al. ²⁰²¹). Halotolerant PGPR also produce compatible solutes (proline, trehalose, and glycine betaine) that they provide to host plants through secretion and cell lysis while simultaneously triggering endogenous solute synthesis in plant tissues (Gupta et al. 2022; Kumar et al. ²⁰²¹). This dual provisioning mechanism operates across diverse crop species, including wheat, soybean, cucumber, and tomato, maintaining cellular turgor and protecting enzymatic machinery from osmotic damage (Vaishnav et al. 2015; Upadhyay and Singh ²⁰¹⁵; Saberi‐Riseh et al. ²⁰²⁰; Orozco‐Mosqueda et al. ²⁰¹⁹).

Ion homeostasis is maintained through complementary rhizosphere and intracellular pathways. In the rhizosphere, PGPR secrete extracellular polysaccharides (EPS) that chelate sodium ions before they reach root surfaces, thus physically reducing uptake (Bhagat et al. 2021; Tewari and Arora ²⁰¹⁴; Ashraf et al. ²⁰⁰⁴). Within plant tissues, microbial volatile organic compounds, particularly acetoin and 2,3‐butanediol, modulate expression of ion transporters, including HKT1, shifting selectivity toward potassium over sodium and maintaining the high K⁺/Na⁺ ratios essential for metabolic function (Zhang et al. 2008; Sunita et al. ²⁰²⁰). Together, these microbial mechanisms address salinity stress at multiple intervention points: preventing sodium entry, promoting potassium acquisition, and sequestering toxic ions into vacuoles and non‐sensitive tissues.

3.2.2. The Role of AMF in Salt Stress Mitigation

AMF link plant physiology and soil physical processes in saline environments (Pooja et al. 2025; Boyno et al. ²⁰²⁵). AMF hyphae extend the effective root system and mobilize inorganic phosphate that would otherwise remain immobilized in saline soils, transporting it directly to the plant host and bypassing the restricted P mobility that typically limits uptake under salt stress (Abbaspour et al. 2021). Beyond nutrient acquisition, AMF modulate plant aquaporin expression, maintaining cellular hydration even when high salt concentrations depress soil water potential (Liu et al. 2023; Liang et al. ²⁰²²). AMF colonization upregulates plant antioxidant enzymes, including superoxide dismutase and peroxidase, which neutralize the reactive oxygen species (ROS) generated by ionic stress (Afshar and Abbaspour 2023; Zhang et al. ²⁰²⁴). Beyond these plant‐level effects, AMF also modify soil physical properties through extracellular protein production.

AMF produce glomalin‐related soil proteins (GRSP), which overlap with the operationally defined autoclaved citrate extractable (ACE) protein fraction routinely measured in soil health assessments (Naasko et al. 2024; Amsili et al. ²⁰²⁵). The original concept of glomalin as a single AMF‐specific glycoprotein has been abandoned: GRSP/ACE fractions co‐extract heat‐stable proteins of plant and bacterial origin along with lipids and humic materials (Son et al. 2024). These fractions are associated with soil aggregation and carbon retention (Banegas et al. 2022; Son et al. ²⁰²⁴) and with organic nitrogen supply capacity, and they respond to management across a wide range of long‐term experiments (Naasko et al. 2024; Amsili et al. ²⁰²⁵).

Under saline conditions, GRSP promotes aggregate stability by counteracting the structural collapse caused by sodium‐induced clay dispersion. Its N‐glycosylated functional groups enhance cation sorption capacity for calcium and phosphorus (Zhang et al. 2017; Son et al. ²⁰²⁴). Where GRSP‐mediated aggregation operates, improved porosity facilitates salt leaching from the root zone, creating a positive feedback between biological activity and physical remediation (Singh et al. 2022). Salinity suppresses the AMF activity that produces GRSP; however, resulting in lower concentrations that could offset per‐unit gains in binding capacity under high salt loads (Zhang et al. 2017). Direct causal evidence for this pathway in strongly saline or sodic field systems remains limited. Most documented benefits arise through AMF inoculation rather than protein amendments. Controlled field trials across salinity gradients are needed to determine whether GRSP amendments provide structural benefits independent of living AMF networks.

3.3. Plant Physiological and Developmental Responses

Plants experience salinity as distinct but interacting osmotic and ionic stresses. Osmotic stress manifests rapidly because soil salinity reduces the osmotic gradient that normally promotes water uptake into roots. Even under high soil moisture content, salinity reduces water uptake and turgor pressure (Alkharabsheh et al. 2021). Ionic stress develops gradually as sodium and chloride accumulate in tissues, impairing photosynthesis and metabolic processes (Muhammad et al. 2024). Sodium disrupts potassium homeostasis, interferes with protein synthesis, and generates ROS that damage cell membranes and proteins.

Plants counter osmotic and ionic stress through complementary cellular and whole‐plant mechanisms. At the cellular level, plants compartmentalize sodium into vacuoles to prevent cytoplasmic toxicity and synthesize compatible solutes to maintain osmotic balance without disrupting metabolism (Muhammad et al. 2024). Whole‐plant strategies include restricting sodium uptake at the root surface, preferentially loading potassium into xylem, and increasing tissue succulence to dilute salt concentrations.

These adaptations enable short‐term survival but impose substantial energetic costs (Munns and Gilliham 2015). Synthesizing compatible solutes, maintaining ion gradients across membranes, and repairing oxidative damage all divert resources from growth and reproduction. Under prolonged salinity, these allocation trade‐offs reduce root biomass and lifespan, alter exudate composition, and decrease canopy cover. These changes feed back into soil processes. Diminished root activity reduces microbial carbon inputs and alters plant‐microbiome interactions, decreased canopy cover increases soil surface evaporation, and shortened root lifespans accelerate organic matter turnover, reinforcing physical and biological soil degradation.

3.4. Cross‐Scale Feedbacks and Threshold Dynamics

When salt inputs exceed leaching and exchangeable sodium exceeds the soil's structural buffering capacity, degradation can cross from predominantly osmotic stress into difficult‐to‐reverse hydraulic and structural decline. Threshold identification requires distinguishing between universal reference values derived from controlled experiments and context‐specific values that vary with local conditions. The ESP at which soil structure fails varies with clay type: smectitic (montmorillonitic) soils can disperse at ESP values as low as about 6%, the threshold commonly used to define sodic soils in Australia, whereas kaolinitic soils tolerate considerably higher sodium saturation (Rengasamy and Olsson 1991). This difference reflects clay mineralogy: expandable clays disperse at lower sodium saturation because swelling pressure overcomes interparticle attraction more readily than in non‐expandable clays. Hydraulic conductivity declines nonlinearly with exchangeable sodium and falling electrolyte concentration, and the magnitude of the decline depends strongly on clay mineralogy: soils dominated by 2:1 layer silicates lose conductivity most readily, whereas kaolinitic and sesquioxide‐rich soils are comparatively insensitive (McNeal and Coleman ¹⁹⁶⁶). Well‐aggregated soils maintain function longer than poorly structured soils because organic matter and calcium bridges stabilize aggregates against sodium‐induced dispersion.

Crop‐response thresholds vary widely among species and cultivars. Strawberry fruit yield begins to decline at an ECe near 1 dS m⁻¹ and is predicted to reach 50% of its nonsaline value near 2.5 dS m⁻¹, whereas bread‐wheat grain yield begins to decline near 6 dS m⁻¹ and reaches 50% near 13 dS m⁻¹ (Maas and Hoffman 1977; Maas and Grattan ¹⁹⁹⁹). Halophytes tolerate substantially greater salinity and are commonly defined as plants capable of completing their life cycle at approximately 200 mM NaCl or more (Flowers and Colmer 2008). Plant and structural thresholds interact in ways that depend on electrolyte concentration. In saline‐sodic soils, a high concentration of soluble salts can hold clays in a flocculated state despite elevated exchangeable sodium, so structural degradation may remain latent while salinity is high. Dispersion frequently becomes most acute when soluble salts are leached without first supplying a calcium source, at which point osmotic relief and structural collapse can occur together and accelerate degradation through positive feedbacks. Further feedbacks include reduced plant cover increasing surface evaporation, which leads to further concentration of salts, reduced infiltration, and diminished leaching (Demo et al. 2025). Vegetation loss also decreases evapotranspiration and deep‐water uptake, allowing water tables to rise and further mobilize salts from the subsoil.

4. Solution Pathways and Restoration Strategies

Addressing salinization requires matching intervention strategies to the multi‐scale nature of the system itself. Because physical, biological, and socioeconomic drivers interact across scales, solutions must simultaneously target soil structure, microbial function, and governance frameworks that align incentives for sustained recovery.

4.1. Field‐Level Interventions

4.1.1. Engineering and Soil‐Amendment Approaches

Engineering and amendment strategies must be matched to the process limiting recovery rather than applied generically. In saline soils the objective is a net downward flux of water carrying soluble salts below the root zone. In sodic soils, exchangeable sodium must first be displaced and hydraulic function restored sufficiently to permit leaching. Saline‐sodic soils require both, in sequence, because leaching soluble electrolytes without supplying calcium can expose or intensify sodium‐induced dispersion. Selection therefore depends on salinity and sodicity, mineralogy and permeability, irrigation‐water quality, water‐table depth, and access to a drainage outlet (Table 2).

TABLE 2.

Field and landscape interventions for salt‐affected soils.

Strategy	Mechanism	Target soil type	Relative cost a	Evidence base b	Long‐term stability	Key risks and trade‐offs	Scalability
Leaching with adequate drainage	Dissolves and transports soluble salts below the root zone	Saline; saline–sodic after Ca2+ amendment	$$	Established (Ayers and Westcot 1985)	Temporary unless drainage and water‐table control are maintained	Salt export to drainage water and downstream salinization; high water use	High where water and drainage exist
Irrigation scheduling and leaching‐fraction management	Reduces salinity through controlled irrigation and leaching, limiting ion accumulation	All classes under irrigation	$	Established	Requires continuous management	Under‐leaching accumulates salt; over‐leaching raises water tables	High
Subsurface drainage	Lowers water table and removes displaced salts	Saline; saline–sodic; waterlogged	$$$	Established (Ayers and Westcot 1985)	Long‐lived with maintenance	Effluent disposal to sensitive reservoirs; trace‐element loading (e.g., Se, B)	Moderate; capital‐intensive
Drainage‐water reuse and disposal	Redirects effluent to progressively more tolerant crops or a managed terminal sink	Saline	$–$$	Established but site‐specific	Requires a terminal sink to remain viable	Progressive concentration of salts and trace elements can cause downstream impacts	Moderate
Cyclic or blended brackish‐water irrigation	Uses saline water selectively by crop stage, salinity level, or blending with freshwater, while relying on freshwater or rainfall to maintain the root‐zone salt balance.	Saline	$	Established in selected irrigated systems but strongly site‐dependent (Sharma and Rao 1998; Liu et al. 2024; Jiang et al. 2026)	Conditional on irrigation‐water salinity, application frequency, drainage, and periodic freshwater or rainfall leaching.	Progressive accumulation of ions if export falls below input	High where brackish water is available
Gypsum and calcium amendments (CaSO4·2H2O)	Ca2+ displaces exchangeable Na+, restoring soil aggregation and permeability	Sodic; saline–sodic	$$	Established (Rasouli et al. 2013)	Durable if reclamation completed and Na+ re‐accumulation prevented	Benefit requires subsequent leaching of added electrolytes; sulfate in drainage	Moderate to high where amendment supply, drainage, and leaching water are available.
Elemental sulfur and acid‐forming amendments	Microbial oxidation acidifies soil, dissolving native carbonates to release Ca2+	Calcareous sodic	$$	Context‐dependent	Slow; depends on oxidation conditions	Over‐application risks acidification and solute imbalance	Moderate
Organic amendments (compost, manure, vermicompost)	Improves aggregation, porosity, cation exchange, and biological activity	Saline; sodic as adjunct	$–$$	Established as adjunct (Chaganti et al. 2015; Kim et al. 2017; Huang et al. 2022)	Moderate; requires repeated application	Microbial contaminants in some feedstocks and salinity in some manures	Moderate; constrained by feedstock volume, quality, and transport distance
Biochar	Improves bulk density, porosity, hydraulic conductivity, and cation exchange	Saline; sodic or saline–sodic, depending on biochar properties and site conditions	$$–$$$	Emerging; supported by meta‐analysis and limited long‐term field evidence (Wang et al. 2024; Jin et al. 2024)	Persistence demonstrated for six years in one field experiment (Jin et al. 2024)	High‐ash feedstocks can raise EC; high application rates may suppress yield	Low to moderate; constrained by production volume and application rate
Straw return and residue incorporation	Moderates soil evaporation and crop salt stress via residue inputs and surface cover.	Saline under irrigation	$	Moderate; regionally concentrated evidence (Jiang et al. 2026)	Requires annual repetition	Nitrogen immobilization	High
Halophyte cultivation and phytoremediation	Promotes salt removal via uptake and extraction, rhizosphere effects	Saline; saline–sodic	$–$$	Emerging	Gradual; modest salt‐removal rates	Potential invasiveness of some species; limited profile‐scale removal	Moderate
Salt‐tolerant cultivars and breeding	Raises crop tolerance threshold without altering the soil	All classes (crop adaptation)	$	Established	Durable as a genetic strategy; requires continued access to seed	Does not reclaim soil; salt load continues to accumulate	Very high
Biostimulants (silicon, PGPR, AMF)	Modify plant physiology, ion homeostasis, and nutrient uptake under stress	Saline (crop adaptation, not reclamation)	$	Emerging; largely pot and greenhouse evidence	Low to moderate; field persistence uncertain	Does not reclaim soil; limited field validation at agronomic rates	High if formulations prove robust

Note: Strategies are compared by mechanism, suitable soil class (as defined in Table 1), relative cost, strength of evidence, long‐term stability, principal risks and trade‐offs, and scalability.

^a Relative cost: $ low, $$ moderate, $$$ high. These ratings compare strategies within this table and are qualitative author assessments informed by the reviewed literature; they are not absolute cost estimates.

^b Evidence base: “established” indicates support from long‐term field studies or operational programs; “emerging” indicates promising approaches with limited long‐term field validation. Ratings are a qualitative synthesis intended to orient comparison rather than precise estimates.

Leaching removes soluble salts only where drainage water can leave the profile. The additional water required depends on irrigation‐water salinity, crop tolerance, rooting depth, and hydraulic properties rather than a universal leaching fraction. Where natural drainage is insufficient, subsurface drains, open drains, or pumped wells are needed to prevent water‐table rise. These systems relocate salts rather than eliminate them, exporting them to drainage channels, evaporation basins, aquifers, or downstream ecosystems, sometimes with elevated trace elements, so effluent volume and composition must be part of the design rather than an afterthought (Ayers and Westcot 1985).

Calcium amendments are required where sodicity has degraded aggregate stability. Gypsum (CaSO₄·2H₂O) supplies Ca²⁺ to replace exchangeable Na⁺, which must then be leached. Rates should be calculated from initial and target exchangeable sodium percentage, cation‐exchange capacity, purity, treatment depth, and reaction efficiency rather than taken from a generic range. In a three‐year field experiment on a sodic soil in semi‐arid Iran, applications equivalent to 50 to 100% of the calculated gypsum requirement lowered the sodium adsorption ratio and pH and increased hydraulic conductivity, while also increasing ECe (Rasouli et al. 2013). The increased ECe reflects the addition and mobilization of soluble electrolytes; sustained reclamation therefore requires subsequent removal of displaced sodium and soluble salts through drainage and leaching. Elemental sulfur can mobilize native carbonates in calcareous soils, but depends on microbial oxidation, temperature, aeration, and time.

Controlled incubation and leaching studies show that organic amendments can improve aggregation, porosity, cation‐exchange properties, and permeability, thereby assisting salt and sodium removal, particularly when paired with gypsum, where exchangeable sodium is high (Chaganti et al. 2015; Sundha et al. ²⁰²⁰; Liu et al. ²⁰²⁰). Long‐term experiments demonstrate the value of these interventions but reveal that recovery is often slow. After 10 years of leaching, neither an untreated control nor rice‐straw compost alone reached salinity and sodicity classification limits, whereas gypsum plus compost doubled aggregate mean weight diameter (Kim et al. 2017). Over 10 years in Northeast China, tillage and rice cultivation alone did not significantly reduce electrical conductivity (EC1:5) at all sampled depths, while adding phosphogypsum or farmyard manure lowered sodium, carbonate, chloride, pH, conductivity, and ESP (Huang et al. 2022). Biochar can also improve the physical and chemical properties of salt‐affected soils, but responses are highly conditional. Across 660 paired observations from 99 publications, biochar reduced ECe by 13.2% and raised cation‐exchange capacity by 17.0% without altering pH, with the reduction in ECe significant in severely salinized soils but not in the slight or moderate salinity classes (Wang et al. 2024). Moreover, a 6‐year field trial found sustained structural and chemical improvement, though rice yield peaked at an intermediate rather than the highest application rate (Jin et al. 2024). Feedstock, salt content, and dose therefore require characterization before field use.

Where freshwater is scarce, brackish or saline drainage water can supplement irrigation as a water‐allocation strategy rather than a reclamation treatment, provided annual salt export equals or exceeds salt input, and exceeds it during active reclamation. A meta‐analysis of 2265 wheat observations from northern China identified an irrigation‐water salinity threshold (about 5 g L⁻¹) beyond which yield and water‐productivity losses became severe and found that alternating saline and freshwater irrigation recovered part of the loss (Jiang et al. 2026). The same meta‐analysis found that straw residue incorporation into the soil partially mitigated wheat yield and water‐productivity losses under saline irrigation, although that evidence is geographically concentrated and does not demonstrate profile‐scale reclamation (Jiang et al. 2026). Multi‐year field studies support conditional use of saline or brackish irrigation water, with preplant freshwater, monsoon leaching, and subsurface drainage preventing progressive salinity accumulation over seven years in semi‐arid India (Sharma and Rao 1998). Also, no apparent salt accumulation was observed in the upper 1 m after eight years under a single seasonal irrigation with water up to 6.3 dS m⁻¹ in the North China Plain. However, a slight increase in salinity occurred at 1–2 m, and risks increased with higher salinity or repeated saline irrigation (Liu et al. 2024).

The evidence supports a sequential approach: diagnose whether salinity, sodicity, or waterlogging dominates; restore permeability; leach mobilized salts; maintain water‐table and drainage control; and monitor the profile and exported drainage water over years, since apparent topsoil recovery can conceal salt accumulation at depth. No amendment substitutes for an explicit water and salt balance, and results from pots or short columns should not be extrapolated to profile or basin scale.

4.1.2. Biological and Crop‐Based Approaches

Interventions such as organic amendments and microbial inoculants can benefit soil structure and biology. These can help rebuild functional diversity, enhance nutrient cycling and stabilize aggregates (Li et al. 2023). However, success rates vary widely. For example, well‐drained sandy soils may recover within 2–3 years, whereas heavy clay soils with high ESP may require a decade or more despite intensive intervention. Crop and residue management can enhance remediation efficiency. Deep‐rooted crops such as barley, alfalfa, or salt‐tolerant grasses increase evapotranspiration and lower the water table, reducing capillary rise (Devkota et al. 2022). They also improve infiltration, cation exchange capacity, and microbial activity, accelerating the recovery of soil structure (Li et al. 2023).

Halophytes such as Atriplex or Salicornia can help extract salts and stabilize bare lands. Although these species rarely restore full productivity in highly saline environments (EC > 12 dS m⁻¹) without complementary management approaches (Navarro‐Torre et al. ²⁰²³), cultivating economically productive halophytes represents an emerging strategy for maintaining revenue from salt‐affected lands during restoration or when conventional desalination is economically infeasible. Commercial halophytic crops include Salicornia spp. as vegetables and oilseed (tolerating EC 20–40 dS m⁻¹), Atriplex spp. as livestock forage, salt‐tolerant quinoa ( Chenopodium quinoa ) as nutritious grain (moderate tolerance, EC 8–12 dS m⁻¹), and Distichlis palmeri (saltgrass) for grain production. Although individual yields may be lower than those of conventional crops, economic viability is possible due to reduced input costs, premium prices for specialty products, and potential carbon credits for restoration (Navarro‐Torre et al. ²⁰²³). Strategies to integrate halophyte production with conventional cropping include spatial mosaics (placing halophytes in the most affected areas) and temporal mosaics (planting halophytes during restoration, transitioning to conventional crops as salinity declines). The development of salt‐tolerant crop varieties and the incorporation of microbes that further enhance halophyte performance also warrant greater research attention as adaptive strategies for both current conditions and climate change scenarios.

In moderately affected soils (EC 4–8 dS m⁻¹), inoculation with halotolerant plant growth‐promoting rhizobacteria (particularly ACC deaminase‐producing strains), AMF, or microbial consortia has shown yield improvements of 15%–40% in field trials (Bhardwaj et al. 2025). Critical uncertainties remain regarding persistence of beneficial effects without repeated inoculation; effectiveness across diverse soil types, crops, and management systems; and scalability to commercial agriculture. Whether microbial inoculation becomes a viable component of integrated salinity management or remains a niche application depends on resolving these uncertainties through long‐term, multi‐site field trials. The integration of halotolerant microbes, appropriate crops, and management strategies could offer sustainable solutions.

Improving irrigation efficiency is an important preventive measure. Converting from flood to drip or sprinkler systems can reduce percolation losses, limiting salt accumulation (FAO ¹⁹⁹⁷). Scheduling irrigation based on soil moisture and salinity sensors maintains optimal leaching while conserving water. For instance, real‐time feedback systems now enable adjustments in water and fertilizer application, aligning inputs with plant demand. These technologies require substantial initial capital investment, with reported costs varying widely by system type, field size, and region, but provide long‐term gains in productivity and salinity control.

4.2. Landscape‐Scale Interventions

Because salinization often originates from groundwater rise and regional hydrological imbalance, field‐level actions must be reinforced by coordinated landscape management. Subsurface drainage remains a highly effective engineering control, with pipe systems installed at 0.9–1.6 m depth holding water tables below the root zone through the recharge season and progressively desalinizing the profile over subsequent years (Sharma and Gupta 2006; Ritzema et al. ²⁰⁰⁸). However, its installation and maintenance costs are substantial and highly site‐dependent, and disposal of saline effluent poses environmental challenges. Where drainage is infeasible, bio‐drainage using deep‐rooted perennial vegetation to intercept and transpire groundwater provides a lower‐cost alternative. Several studies indicate that Populus and Eucalyptus plantations can reduce groundwater recharge and lower water tables under certain conditions (e.g., Li et al. ²⁰¹⁴; Vance et al. ²⁰¹⁴), though the magnitude and timeframe of these effects vary by site, species, and climate. However, bio‐drainage requires dedicating a portion of the landscape to plantations, typically 5–10 ha per 100 ha of cropland, and this reduces cultivable area and is most effective where water tables are within 3–4 m of the surface.

Restoration of natural vegetation and wetland systems can also stabilize hydrology. Re‐establishing deep‐rooted or high‐transpiration perennials, such as lucerne and saltbush, reduces recharge and enhances evapotranspiration, helping maintain water tables below critical depths and improving aeration (Barrett‐Lennard ²⁰⁰²). Integrated catchment programs combining vegetation reintroduction, improved irrigation scheduling, and managed aquifer recharge have demonstrated success in reducing regional groundwater salinity over decadal timescales. In arid basins, managed aquifer recharge using treated wastewater or stormwater can dilute groundwater salinity and improve irrigation water quality.

Ecological restoration complements hydrological control by rebuilding soil function. Introducing salt‐tolerant native species accelerates soil‐carbon recovery and microbial recolonization (Bhardwaj et al. 2025). Mixed‐species plantings provide multiple functional traits, including deep rooting, litter quality, and rhizosphere activity, that enhance both physical stabilization and biogeochemical cycling. Over time, the return of vegetation cover reduces erosion, moderates microclimate extremes, and supports gradual re‐establishment of non‐halophytic flora once salinity declines.

4.3. Governance, Economic, and Policy Frameworks

The persistence of salinization reflects institutional and economic barriers as much as biophysical constraints. In many regions, water remains underpriced or unmetered, removing incentives for efficient use. This creates a paradox: while inefficient irrigation increases drainage volumes that can help leach salts, it also raises water tables and mobilizes subsurface salts. In well‐drained alluvial systems the leaching benefit can dominate, but in closed basins or areas with shallow water tables, net salinization typically results. For example, in northwest India, policies providing highly subsidized electricity enabled on‐demand groundwater withdrawal, contributing to rapidly depleting water tables. In Punjab and Haryana, which together produce over 60% of India's grain, the result has been rapidly depleting water tables (Bhattarai et al. 2021). Conversely, market‐based instruments and volumetric water pricing, such as in Australia's Murray–Darling Basin, incentivized efficiency improvements, but the effects were complex and moderated by factors like drought and shifts in crop types (Goesch et al. 2020).

Land tenure security directly affects investment in remediation. Studies in Pakistan's Punjab indicate that farmers with secure legal land rights are more likely to adopt long‐term water‐saving technologies compared to tenants with insecure leases (Khan et al. 2022). In Türkiye, the Alanya water‐user association reduced its unit management, operation, and maintenance costs by 40%–70% over 2013–2020 despite rising competition for water from tourism (Alcon et al. 2024). Decision‐support platforms that integrate real‐time salinity sensors, satellite imagery, and economic models now guide targeted interventions in Australia and California, prioritizing areas where recovery may exceed costs. At national scales, integrating salinity into climate‐adaptation frameworks, as Morocco and Tunisia have done, mobilizes international funding while framing restoration as a strategy for both agricultural productivity and carbon sequestration.

Knowledge transfer and capacity building are essential complements to these policy frameworks. Participatory extension programs that involve farmers in monitoring and adaptive management consistently outperform top‐down interventions. Digital platforms for sharing soil and water data enable near‐real‐time tracking of salinity trends, thereby improving transparency and accountability. Policy effectiveness ultimately depends on feedback between science, governance, and practice, reflecting the same adaptive coupling that characterizes the soil system itself.

5. Research Priorities for Predictive Prevention

The transition from reactive remediation to proactive prevention of soil salinization (Figure 3) will require increased research in many areas. Here, we highlight research priorities to address key deficiencies, including the need to understand the threshold dynamics that separate reversible from recalcitrant degradation, ensuring an adequate data infrastructure to support early warning, and overcoming socioeconomic, scaling, and governance challenges to enable adaptive management.

FIGURE 3.

Management paradigms for salt‐affected soils: Comparisons of reactive remediation versus predictive prevention. (a) Reactive remediation cycle in which crossing a threshold of salt‐affected soil degradation triggers costly interventions with uncertain outcomes, driven in part by self‐reinforcing degradation feedbacks. (b) Predictive prevention pathway in which coordinated planning, policy, and incentives aim to keep the system below context‐specific thresholds and increase the likelihood of durable management outcomes. The color gradient distinguishes the nested spatial and organizational scales, and the dashed yellow boundary denotes maintenance of the system below the critical threshold. The large central arrow illustrates the necessary paradigm shift from reactive remediation to predictive prevention. Economic assessments indicate that allowing salt‐induced degradation to persist imposes large long‐term production losses and that investments in prevention, management, or remediation can be cost‐effective, although outcomes vary with degradation severity, local conditions, and the intervention used (Qadir et al. 2014; Nkonya et al. ²⁰¹⁶).

5.1. Quantifying Biophysical Thresholds and Recovery Dynamics

Translating the mechanistic understanding of sodium‐clay interactions into field‐level predictions requires longitudinal threshold data that do not yet exist for most pedoclimatic contexts. Laboratory percolation studies show that structural thresholds depend strongly on clay mineralogy, with smectitic soils losing hydraulic conductivity at much lower sodium saturation than kaolinitic or sesquioxide‐rich soils (McNeal and Coleman ¹⁹⁶⁶). This variability reflects interactions among clay type, organic matter content, initial aggregate stability, and the ratio of sodium to calcium plus magnesium in solution. Despite this mechanistic understanding, systematic datasets relating EC, ESP, and hydraulic conductivity across threshold regions remain scarce. Most published studies report single‐point observations at fixed salinity levels rather than the trajectories through threshold regions that separate reversible from recalcitrant degradation.

Current monitoring programs that track only surface EC miss the deeper salt stores below 1 m that influence long‐term recovery potential, since leaching must displace salts below the root zone to be effective. Temporally, annual observations miss the transient wetting‐drying cycles that drive capillary rise and surface salt accumulation. Filling these spatial and temporal gaps is essential across the diverse pedoclimatic contexts in which salinization occurs.

Early detection of approaching thresholds requires monitoring resilience loss before structural collapse occurs. Theory from dynamical systems indicates that systems approaching critical transitions exhibit characteristic signatures including increased variance, slower recovery from perturbations, and increased temporal autocorrelation (Scheffer et al. 2009). These signatures have not yet been systematically tested in salinizing soils, and the lead times at which they would become detectable remain unquantified. Validating these early‐warning signatures across diverse contexts requires controlled experiments that deliberately transition soils through threshold boundaries while monitoring structural, hydraulic, chemical, and biological indicators simultaneously. Such experiments need sustained observation over 5 to 10 years across diverse soil textures, clay mineralogies, and climatic conditions. A coordinated network of reference sites, each tracking EC, ESP, hydraulic conductivity, water table depth, and microbial composition, would enable cross‐site synthesis to derive context‐specific thresholds and the lead times at which different indicators signal approaching collapse. Because degradation accelerates nonlinearly once thresholds are crossed, this investment in monitoring infrastructure would represent a fraction of current remediation expenditures, yet could substantially reduce long‐term restoration needs.

5.2. Building Predictive Monitoring Infrastructure

Early detection of salinization before visible symptoms and economic losses is critical for timely intervention while restoration remains economically viable. The recent convergence of satellite remote sensing, machine learning, and cloud computing has made continuous landscape‐scale monitoring operationally feasible (Wang, Shi, et al. ²⁰²⁰; Wang, Ding, et al. ²⁰²⁰; Ma et al. ²⁰²³).

Multispectral satellites (Landsat 8–9 at 30 m resolution, Sentinel‐2 at 10 m for visible‐NIR bands) detect salinity through two complementary pathways: vegetation stress indices and soil surface spectral signatures. Normalized Difference Vegetation Index (NDVI) declines from 0.6 to 0.8 in healthy crops to 0.2–0.4 under moderate salinity stress (EC 6–10 dS m⁻¹), while salinity indices exploit increased reflectance in visible wavelengths from surface salt crusts (Khan et al. 2005; Ivushkin et al. ²⁰¹⁹). Machine learning algorithms (Random Forest, XGBoost) that combine multiple spectral indices with topographic, climate, and management variables achieve prediction accuracy of R ² = 0.75–0.90 when calibrated with 50–100 ground measurements per 10,000 ha (Wang, Shi, et al. ²⁰²⁰; Wang, Ding, et al. ²⁰²⁰; Dehkordi et al. ²⁰²⁴). This accuracy approaches electromagnetic induction ground surveys (R ² = 0.90–0.95) while covering 100–1000 times larger spatial extents at dramatically lower cost. These results demonstrate the operational viability of satellite‐based early detection systems.

Accuracy improves when several Sentinel‐2 images acquired near the sampling date are combined rather than interpreted singly; in coastal farmland in China, combined‐temporal models raised validated R ² to 0.62–0.72, gains of 0.02–0.36 over single‐date models (Duan et al. 2025). Extending temporal analysis to multi‐decadal scales, Wang et al. (2025) used annually composited Landsat indices with random forest models to map global soil salinity and sodicity at 1 km resolution from 1980 to 2024, revealing significant increasing trends across more than 20% of global land area for salinity and nearly 50% for sodicity. These results demonstrate that multi‐temporal remote sensing combined with machine learning can achieve coefficients of determination of 0.65–0.75 at global scales, though most studies predict continuous salinity levels rather than explicitly forecasting the critical thresholds at which management intervention is needed.

Google Earth Engine, a cloud computing platform, democratizes analysis by hosting petabyte‐scale satellite archives with free access, thereby eliminating local hardware requirements (Gorelick et al. 2017). This capability enables processing of continental‐scale datasets that previously required specialized infrastructure. Several long‐running programs show how monitoring at these scales can move management from reaction toward prevention. In Australia's Murray‐Darling Basin, three decades of coordinated salinity management have combined in‐stream salinity monitoring and salt‐interception schemes within an accountability framework of basin‐wide salinity targets and a salinity register of credits and debits; this coordinated approach contributed to bringing river salinity largely under control and met the Basin Salinity Target for the first time in 2010, although climatic variability also influenced observed salinity trends (Hart et al. 2020). In the upper Indus Basin of Pakistan, a district‐scale study in Faisalabad combined multispectral satellite imagery with field measurements to map temporal and spatial patterns of salt‐affected soils. Supervised classification achieved approximately 90% overall accuracy, illustrating the potential of remote sensing to provide rapid spatial information for targeting field assessment and management (Abbas et al. 2013).

Farmer‐facing applications can translate monitoring data into field‐specific guidance, delivered through advisory services, mobile apps, or SMS alerts. In Vietnam's Mekong Delta, a statistical model using El Niño–Southern Oscillation and upstream‐streamflow predictors forecasts exceedance of critical salinity thresholds at a coastal monitoring station with lead times of several months and strong model fitting metrics, providing a potential basis for earlier cropping and water‐management decisions (Apel et al. 2020). Such anticipatory forecasting exemplifies the shift from reactive response toward predictive prevention, and because it draws on widely available climate and streamflow data, it is readily transferable to other monitoring sites.

Emerging technologies offer additional early‐warning potential but require further development before operational deployment. For instance, hyperspectral imaging (400–2500 nm) can detect pre‐visual stress and discriminate ion‐specific spectral signatures, although current satellite systems remain experimental. Uncrewed aerial vehicle (UAV)‐based multispectral sensors provide sub‐meter resolution for within‐field heterogeneity mapping at intermediate scales between ground sampling and satellite coverage.

Data standardization and sharing infrastructure represent critical but under‐invested needs. Interoperability across national and regional monitoring systems would enable continental‐scale trend analysis and facilitate model validation across diverse contexts. Successful examples include ISRIC's SoilGrids global soil database and FAO's Global Soil Information System (GLOSIS), which demonstrate the feasibility of standardized, open‐access platforms. Extending these frameworks to include dynamic salinity monitoring with standardized measurement protocols would accelerate knowledge synthesis and enable early detection of emerging hotspots. Table 3 summarizes these monitoring and governance measures and their role in enabling and coordinating the field and landscape interventions in Table 2.

TABLE 3.

Monitoring and governance measures for salt‐affected soils.

Strategy a	Mechanism	Application context	Relative cost	Evidence base	Long‐term stability	Key risks and trade‐offs	Scalability
Ground‐based monitoring (ECe, SAR, ESP)	Direct measurement of salinity and sodicity through the profile	All salt‐affected soil classes	$–$$	Established	Enabling activity rather than a soil change	Requires high sampling density; often expensive; depth coverage often inadequate	Moderate
Remote sensing and mapping	Maps salinity extent and change to spatially target interventions	Regional mapping and field targeting	$	Established for mapping	Enabling activity	Accuracy depends on ground calibration; sparse in data‐poor regions	Very high
Seasonal forecasting and early warning	Anticipates intrusion or accumulation, providing lead time to adjust management	Coastal and irrigated systems with predictable seasonal salinity risk	$	Emerging	Depends on sustained data and institutional uptake	False alarms erode trust; requires responsive management capacity	High
Policy and economic instruments	Coordinate basin‐scale prevention via salinity targets, credit and debit registers, and incentives	Basins or jurisdictions requiring coordinated salt management	$–$$	Established in some basins	High where institutions endure	Potential distributional inequities; requires enforcement capacity	High; jurisdiction‐dependent
Data standardization and sharing infrastructure	Enables interoperable protocols for cross‐region synthesis and model validation	Cross‐region monitoring and model development	$	Emerging	Potentially high, but dependent on continued stewardship, interoperability, and funding.	Requires sustained funding and governance of shared platforms	High

Note: These measures direct, sustain, or coordinate the field and landscape interventions described in Table 2; relative cost and evidence‐base ratings follow the definitions given for Table 2. ECe, electrical conductivity of the saturation extract; SAR, sodium adsorption ratio; ESP, exchangeable sodium percentage.

^a These measures do not themselves alter soil properties. They determine whether the interventions described in Table 2 are applied before or after threshold crossing.

5.3. Institutional, Economic, and Integration Challenges

Understanding of governance structures and behavioral dynamics influencing restoration outcomes remains limited despite evidence that institutional factors are as critical as technical interventions (Section 4.3). The institutional features that enable collective drainage investment, such as secure water rights, enforcement capacity, and social capital within water‐user associations, have been documented in case studies (Alcon et al. 2024). However, we lack a systematic cross‐regional comparison to distinguish generalizable design principles from context‐dependent factors. Comparative institutional analysis across at least 15–20 water‐user associations spanning diverse political systems, water scarcity levels, and land tenure arrangements would reveal which governance structures function universally and which require local adaptation. Similarly, the relationship between property‐rights regimes, land‐tenure security, and willingness to invest in long‐term soil recovery remains poorly quantified across institutional contexts. Panel data tracking investment decisions across land tenure gradients over 10–15‐year periods would quantify how tenure security affects adoption rates and the duration of sustained practices, isolating tenure effects from confounding factors such as credit access and technical knowledge.

Rigorous experimental comparison of alternative governance approaches is largely absent from the literature. The comparative effectiveness of participatory monitoring versus top‐down enforcement in sustaining the adoption of improved practices is based on anecdotal evidence and lacks a rigorous experimental comparison. Field experiments embedding different governance structures (farmer cooperatives, government enforcement, hybrid models) within comparable biophysical settings could isolate the effects of institutional design on long‐term adoption and restoration outcomes. Such studies would require multi‐year commitments spanning at least one full cropping cycle to capture delayed responses and adaptive learning.

Economic instrument design requires deeper analysis to distinguish interventions that reduce salt loads from those that merely displace problems spatially or temporally. Integrated models linking farmer decision‐making with hydrological solute transport and soil physicochemical processes could reveal which combinations of incentives and regulations achieve durable salt mass reduction rather than temporary symptom relief or problem displacement. Specifically, coupled models should quantify downstream salinity impacts of upstream efficiency improvements that reduce return flows, temporal lags between subsidy implementation and water table responses across hydrogeological settings, and threshold subsidy levels that trigger adoption without generating rebound effects through intensification. Agent‐based models coupled with biophysical simulators, such as the MP‐MAS framework for agricultural systems (Schreinemachers and Berger 2011), offer methodological templates adaptable to salinity contexts but require calibration against longitudinal datasets that track both farmer decisions and salt fluxes across basin scales over 10–20 year periods.

The effectiveness of capacity‐building remains poorly quantified across delivery models and cultural contexts. Extension models that successfully translate technical knowledge into farmer adoption vary widely, from traditional top‐down technology transfer to participatory farmer field schools to digital decision‐support platforms. Randomized controlled trials comparing adoption rates, sustained practice change, and cost‐effectiveness across these models within the same region would identify which approaches merit scaled investment. The integration of indigenous knowledge with scientific understanding represents a particularly under‐explored opportunity. Systematic documentation and experimental validation of traditional salt‐management practices in regions with centuries of irrigation experience (Central Asian oases, Middle Eastern qanat systems, South Asian karez networks) could identify low‐cost, locally adapted strategies transferable to similar agroecological zones.

The central challenge in salinity management is integration: linking biophysical understanding with socioeconomic dynamics, and scaling from controlled experiments to heterogeneous regional contexts. Most salinity models emphasize either biophysical processes (e.g., HYDRUS and SWAP for solute transport, APSIM for crop response) or economic decision‐making (e.g., farm‐level optimization models, basin‐scale water allocation models), with limited integration between domains despite growing recognition of their interdependence. Developing coupled frameworks that integrate soil physics, microbial ecology, plant physiology, and farmer behavior within a single model structure would enable evaluation of combined interventions. Such models could, for instance, test whether subsidizing drainage infrastructure is more effective than adjusting water pricing, or whether bio‐drainage performs better than gypsum amendment under specific climatic and economic conditions. Comparative policy analysis in related agricultural domains, such as nitrogen pollution control through combinations of regulations and economic incentives, provides methodological frameworks adaptable to salinity (Oenema et al. 2011).

Even the most sophisticated coupled models require validation through long‐term field experiments. Currently, experiments integrating biophysical treatments (amendments, drainage, halophyte rotations) with varied governance structures (subsidies, cooperatives, regulations) across salinity gradients are limited to a handful of sites globally, including the Australian Cooperative Research Center for Plant‐Based Management of Dryland Salinity sites and select International Center for Agricultural Research in the Dry Areas (ICARDA) trials in Central Asia. A minimum decade‐long observation period is necessary to capture slow hydrological feedbacks and verify sustained improvement beyond transient responses. Expanding this network to represent diverse soil‐climate‐institutional contexts across major affected regions (South Asia, North Africa, Central Asia, Mediterranean, North American Great Plains) would enable identification of context‐specific versus universal restoration principles.

Equally important are learning networks that link researchers, practitioners, and policymakers, allowing continuous exchange of data and experience. Such networks transform restoration from isolated interventions into adaptive systems of innovation and feedback. The Consultative Group on International Agricultural Research (CGIAR) research programs and networks, such as the Global Soil Partnership, provide organizational templates, although salinity‐specific learning platforms remain fragmented. Establishing a global salinity monitoring and learning network that builds on existing remote sensing infrastructure could accelerate knowledge synthesis and enable rapid identification and replication of successful innovations.

Scaling challenges extend beyond scientific understanding to implementation capacity. Even where effective technologies exist, barriers include a lack of technical expertise, insufficient extension services, inadequate financing mechanisms, and policy misalignments. Addressing these constraints requires integrated socio‐technical strategies that combine technology development with institutional innovation, capacity building, and enabling policies. Success depends on recognizing that salinity management is fundamentally a coupled human‐natural system problem requiring integrated intervention across scales and domains.

Closing these knowledge gaps will require sustained investment that connects monitoring infrastructure to long‐term field experiments embedded within contrasting governance frameworks. Equally important is the development of integrated modeling platforms that couple biophysical and socioeconomic dynamics with learning networks that enable adaptive management. Progress across these fronts would enable a shift from reactive remediation to proactive prevention, guided by early‐warning systems and governance structures that align private incentives with public benefits.

6. Conclusion

Soil salinization arises from interactions among physical, biological, and socioeconomic feedbacks operating across scales. At the biophysical level, shifts in soil structure, microbial function, and plant physiology create threshold behavior and accelerate degradation once critical limits are exceeded. At the socioeconomic level, decisions about irrigation and land use both shape and respond to these biophysical dynamics. Effective management therefore requires interventions aligned with enabling institutions, incentive structures, and data systems (Figure 4).

FIGURE 4.

Intervention strategies for salt‐affected soils by spatial scale and stage of degradation. Strategies are organized by the stage of degradation at which they are applied (columns: Prevention before threshold approach, early response during declining resilience, and remediation after functional or structural decline) and by the spatial scale at which they operate (rows, from microscale soil processes to institutional governance). Placement indicates the stage and scale at which each strategy is typically most useful. Many strategies span stages and scales, and selection should be based on diagnosis and local context. The matrix emphasizes that earlier, coordinated intervention generally preserves a broader range of feasible responses.

The next decade offers a critical window to translate scientific understanding into predictive management. Emerging technologies such as real‐time sensor networks, remote sensing platforms, and coupled process–decision models create unprecedented opportunities to forecast salinity dynamics and evaluate the system‐wide effects of interventions. Achieving this transformation will require open data infrastructures, long‐term experimental networks, and collaborative institutions capable of integrating knowledge across disciplines and jurisdictions.

Restoring more than 160 million hectares of salt‐affected cropland is essential for food security, climate adaptation, and biodiversity (FAO ²⁰²¹). Progress requires treating restoration as a systems problem linking soil physics to governance design and local management to planetary resilience. Without this shift, continued degradation will further erode global food production capacity, undermine climate adaptation efforts, and compound socioeconomic vulnerability in regions already stressed by water scarcity and environmental change.

Author Contributions

Matthew D. Wallenstein: conceptualization, writing – original draft, visualization, writing – review and editing. Gwyn A. Beattie: writing – original draft, writing – review and editing, visualization. Amy M. Grunden: conceptualization, writing – original draft, writing – review and editing. Joana Falcao Salles: conceptualization, writing – original draft, writing – review and editing.

Funding

This work was supported by the National Institute of Food and Agriculture, IOW04108, Iowa Soybean Research Center, and Novo Nordisk Fonden, NNF19SA0059360.

Conflicts of Interest

M.D.W. is employed by Syngenta, which develops and markets agricultural products including those related to soil health and crop protection. The remaining authors declare no conflicts of interest.

Data Availability Statement

The authors have nothing to report.