Potential management practices of saltwater intrusion impacts on soil health and water quality: a review

Most of the large cities globally are located along the coastlines and more than 40% of the world population lives within 100 km of the coastline, while 10% of the population lives in coastal areas that are less than 10 m.a.s.l. (UN 2017). Overall, there is strong evidence that changes in the sea level will impact a significant portion of the global population around the world, especially those in coastal regions (Neumann et al. 2015; Kummu et al. 2016). By the year 2050, the global population residing in low elevation coastal zones is expected to increase significantly, which represents the most vulnerable population to sea-level rise (SLR) (Kummu et al. 2016; UN 2017; Abadie 2018; Kulp & Strauss 2019). The most densely populated countries in coastal regions, such as Bangladesh with a large population in coastal areas, are extremely vulnerable to an accelerated rate of SLR (Kulp & Strauss 2019). A report by Weeman & Lynch (2018) based on the findings of a study at the University of Colorado Boulder that evaluated 25 years of NASA and European satellite data concluded that the rate of global SLR in recent years is becoming alarming. The same study estimated that the SLR is expected to reach 0.65 m by 2100 if it continues at the current rate (Weeman & Lynch 2018). Projected accelerations of SLR highlight the contribution of climate change and variability as the driving factors (Horton et al. 2014; Olsson et al. 2015). The rate of climate variability during the recent years has been accelerated due to increasing concentrations of greenhouse gases (GHGs) in the atmosphere that contributed to an increase in the air and water temperatures, which in turn results in an increased rate of ice melting in Greenland and Antarctica thus contributing to accelerated SLR (MacCracken 2001; Singh & Singh 2012; Mimura 2013; Pattyn et al. 2018; Overland et al. 2019). The accumulation of GHGs traps heat causing warming of the earth (US EPA 2015; Dahlman 2019; Butler & Montzka 2020). Oceans absorb more than 90% of the trapped heat (Cheng et al. 2017). However, with the increase in temperature due to climate change and variability water volume in oceans rises, which is referred to as ‘thermal expansion’, contributing to increases in the rate of SLR (Milly et al. 2003; Vermeer & Rahmstorf 2009; Meehl et al. 2012; Mimura 2013). Recent studies have shown that the rate of thermal expansion increased from 0.4 to 0.8 mm/year from 1970 to 2010 (Rhein et al. 2013).

On the other hand, the carbon content in the ocean is about 50-fold greater than that in the atmosphere (Sarmiento & Gruber 2002) and absorbs nearly 30% of anthropogenic CO₂ emissions (Cicerone et al. 2004; Bates et al. 2012). Therefore, the ocean plays a major role in slowing down the rate of climate change (Raven & Falkowski 1999; Bigg et al. 2003; UN 2017). The ocean also plays a major role in the global water cycle by representing the majority of the total water exchange, as evaporation and precipitation, between the atmosphere and the earth's surface (Schmitt 1995; Bengtsson 2010; Charette & Smith 2010). Climate change and variability are accelerating the rate of SLR, which in turn causes storm surge and flooding that impacts the livelihood (Awal et al. 2016) and salinization of water and soils (Figure 1).

Currently, most of the focus and efforts are concentrated on mitigating the impacts of climate change on coastal regions due to storm surge, land subsidence, and loss of aquatic habitat (Craft et al. 2009; Ardon et al. 2016; Rosen 2017). Increases in SLR, due to climate change and variability, and associated risks of flooding, saltwater intrusion (SWI), and salt buildup in soils and water are major concerns for human habitat, sustainable development as well as soil health, and crop productivity (Craft et al. 2009; Singh & Singh 2012; Mimura 2013; Rosen 2017). The net negative effects of SWI are on food security due to the salinization of water and soils (Pezeshki et al. 1990; Arslan & Demir 2013; Dasgupta et al. 2015; Szabo et al. 2016; Litalien & Zeeb 2020). Salinization of water and soils affects the availability of freshwater and suitability of arable land for crop production as most crops are sensitive to high salt contents in irrigation water and/or in soils (Habibullah et al. 1999; Wichelns & Qadir 2015; Butcher et al. 2016; Szabo et al. 2016; Khanom 2016; Chen & Mueller 2018; Majeed & Muhammad 2019). In addition, SWI can mobilize nutrients from the soil by influencing soil chemistry, which in turn become sources of eutrophication of the adjacent freshwater bodies as well as coastal waters impacting marine life (DeLaune & Smith 1985; Portnoy & Giblin 1997; Williams et al. 2014; Neubauer et al. 2019; Tully et al. 2019). Guha & Panday (2012) used the groundwater model to predict the impact of SLR on salinity and groundwater levels, chloride concentrations, and groundwater discharge into canals in Miami-Dade and Broward counties. Their predictions showed that SLR of 0.6–1.2 m resulted in a groundwater head increase in some areas of 4–15%; however, the effect was more significant on the average relative chloride concentrations which increased 100–600% in some wells.

Currently, an estimated 1.1 × 10⁹ ha land, which accounts for 7% of the world's total land area, is impacted by soil salinity and/or sodicity (Yensen 2008; Wicke et al. 2011). Poor management of irrigation, particularly using brackish water with inadequate freshwater for leaching salt buildup from the root zone, is contributing to increasing soil salinity (Connor et al. 2008; Wichelns & Qadir 2015). About 33% of globally irrigated land is impacted by salinization (Munns 2005; Chaves et al. 2009). In addition, salinity negatively impacts plant physiological processes, such as photosynthesis and respiration, which in turn reduces plant growth and production (Minhas et al. 2020; Tomaz et al. 2020).

This review paper mainly focuses on available research on the impacts of soil and water salinization due to SWI on soil health indicator parameters, nutrient transformation and transport, and GHG emissions. In addition, this review summarizes the potentials of selected soil amendments as potential mitigation strategies of soil and water salinity in agricultural soils.

In coastal areas, under normal conditions, freshwater moves towards the sea; therefore, it prevents SWI of coastal aquifers (Barlow & Reichard 2010; Beebe et al. 2016; Memari et al. 2020). The interface between the freshwater and saltwater is a diffuse layer, generally referred to as the ‘zone of dispersion’ where freshwater and saltwater are mixed in the aquifer deep below the land surface (Abd-Elhamid et al. 2020; Sowe et al. 2020) (Figure 2).

Major factors causing SWI are (i) excessive pumping of groundwater; (ii) reduced recharge of groundwater; (iii) changes in land use such as man-made canals for draining the land for urban and/or agricultural uses; (iv) SLR; and (v) increased frequency of storm surges, due to climate change, that floods the land surface with the seawater part of it may be discharged into the canal and eventually drains into the fresh aquifer system (Chen et al. 2020; Maliva 2020; Siegel 2020; Zhu et al. 2020).

As SLR continues to accelerate during the recent years, much of the attention has been on SWI impacts on groundwater and the threat it poses to the freshwater resources in coastal regions. However, limited attention is given to the impact of the landward movement of saltwater by the surface and/or groundwater connections, which can influence the salinity of the agricultural soils and, in turn, the impact of crop production in the coastal zones (Herbert et al. 2015; Bhattachan et al. 2018). The impact of SLR is also on the salinization of drainage ways, including natural as well as man-made canals and ditches, which in turn impact the soils in areas adjacent to these waterways as well.

Land-use changes such as the conversion of wetlands for cultivation often result in lowering the elevation of the landscape and more inland water connection through drainage ditches to maintain the soil suitable for cultivation (Alshammari et al. 2020; Sathiya Bama et al. 2020). These changes will contribute to an increased risk of salinization of inland surface water and the soil. This impact will accelerate as the climate crisis enhances the risk of drought and storms in the coastal zone and inland regions where an extensive network of artificial drainage canals exists as a part of land-use change and development.

SWI is a critical issue in Florida, which heavily depends on freshwater aquifers for drinking water and other uses including agriculture and landscaping. Due to high chloride concentration in seawater (∼19,000 mg/L), SWI of the slightest scale could result in higher chloride concentration in freshwater greater than the EPA's 250 mg/L recommendation for drinking water use (Stumm & Morgan 1981). South Florida is surrounded by seawater on three sides, where seawater is found in deep rock layers below the freshwater aquifer. The impact of SWI is very significant for freshwater management as pumping increases to meet the water needs for the increasing population and other water uses.

SLR results in two most important factors in coastal forested wetlands, i.e. soil moisture or wet/dry cycles and salinity. These changes, in turn, affect the biogeochemistry of the soil, as well as GHG emissions. Liu et al. (2017) examined the changes in soil moisture levels and salt concentration (0, 1, and 5 parts per thousand as sodium chloride (NaCl)) on GHG emissions and dissolved organic carbon (DOC) characteristics of forested wetlands. Their results showed that carbon (C) and nitrogen (N) cycling in these soils were impacted more by water levels than salinity. Wet and dry cycles of the soil decreased the DOC production and total methane (CH₄) emissions. CO₂ emission was linearly related to DOC; therefore, it indicates that DOC is the main factor that determines the availability of CO₂ which can be subject to emission.

Other studies have also shown that an influx of seawater into wetlands increased CO₂ production (Marton et al. 2012; Chambers et al. 2014). Similarly, Steinmuller & Chambers (2018) reported that 5 and 15 parts per thousand salt content in water increased CO₂ emission from wetland soils.

However, the mechanisms and magnitude of the salt-induced loss of soil carbon (C), as CO₂, CH₄, or DOC, from wetland soils remain less understood.

Maucieri et al. (2017) investigated the effects of salinity, biochar amendment, and different soil wetting levels on GHG emission from a semi-arid Australian soil. They reported that the major effect was associated with soil wetting with an increase in water holding capacity from 25 to 100% resulting in a 170% increase in CO₂ emission, while CO₂ emission was decreased by 19 and 28% at 5 and 10 dS (decisiemens) per meter salinity, respectively, compared with that in the control, i.e. no salinity treatment.

Nitrous oxide (N₂O) emission from agricultural soils accounts for 30% of global anthropogenic N₂O emissions (Mosier 1994). However, some estimates place this proportion as high as 80% (Beauchamp 1997) with a high level of caution on lack of reliable data for high precision estimates N₂O emission from agricultural soils. Soil water management appears to be the main factor determining the N₂O emission from agricultural soils (Cao et al. 2019). However, the precise relationship between water management, irrigation water salinity, and N₂O emission is not clearly defined and hence needs more research.

Salinization has shown a varied impact on GHG emissions. Increasing salinity has been shown to increase N₂O emission (Low et al. 1997) and decrease CH₄ emission (Marton et al. 2012), while Kontopoulou et al. (2015) reported no salinity effects on CH₄, CO₂, and N₂O emissions.

Wei et al. (2018) reported that N₂O emission was greater from soil irrigated by 5 mg/L saline water when compared with irrigation water with 2 and 8 mg/L salt. They studied N₂O emission at 0 and 120 kg N/ha treatments. Across both N rates, N₂O emission decreased by 22–40% with irrigation using saline water at 2 or 8 mg/L, when compared with that irrigation using freshwater. For a similar comparison using saline irrigation water at 5 mg/L, N₂O emission increased by 58–88%. Surprisingly, a substantial increase in N₂O emission occurred at the intermediate salinity level (5 mg/L) compared with that of low and high salinity levels (2 and 8 mg/L, respectively) in irrigation water.

A salinity level of 0.4–8.0 dS/m in irrigation water increased N₂O emission from a cotton field without nitrogen fertilizer under drip irrigation (Zhang et al. 2016). Other researchers (Inubushi et al. 1999) reported that N₂O emission and irrigation water salinity were variable in different soils, indicating that soil properties influence determination of N₂O emission and irrigation water salinity interactions. Yet, other studies have shown that salinity decreased N₂O emission (Tsuneda et al. 2005; Maucieri et al. 2017).

Methane is the most potent GHG; therefore, it plays a significant role in the climate crisis. The global warming potential of CH₄ is 25-fold greater than that of CO₂ (IPCC 2007). A better understanding of factors impacting CH₄ emission is important to develop management practices to mitigate CH₄ emission that would slow the rate of the climate crisis. CH₄ emission from soils is primarily influenced by the depth of water table, temperature, salinity, and soil organic carbon. Olsson et al. (2015) reported that CH₄ emission increased at the high and constant water table and at high temperatures. Therefore, any factors, such as rising sea level and seawater intrusion inland, which influence the rising water table, increase CH₄ emission. However, the CH₄ emission rate decreased as the salinity level increased above 18 parts per thousand.

Human activities, including expansion in industrialization and intensive agricultural production systems, will continue to impact the eutrophication of water bodies. Furthermore, climate change and variability may continue to play a major role in eutrophication due to greater water temperature, increased outflow of freshwater, and nutrients (Rabalais et al. 2009).

SWI may impact soil biogeochemical processes that lead to nutrient export into coastal zones, which in turn contribute to eutrophication. Steinmuller & Chambers (2018), in their study on three wetland soils from Florida, have shown that 5 and 15 parts per thousand salt content in water increased CO₂ production as well as P and NH₄ concentrations in porewater. This suggests increased risk of phosphorus (P) and nitrogen (N) transport from coastal soils into the sea, which contributes to increased eutrophication. Herbert et al. (2015) also reported that SWI impacted soil biogeochemical processes as well as vegetation communities, habitat, and microbial community.

Some of the reasons for increased nutrient transport from coastal soils into marine ecosystems are: (i) loss of the wetland's capacity to retain or denitrify N due to the introduction of brackish water into the wetlands (Craft et al. 2009; Chambers et al. 2011); (ii) as the wetlands become brackish, plant uptake of N decreases due to reduced plant growth as a result of sulfide toxicity and/or increased salt stress; therefore, increased residual N can be subject to transport (Cormier et al. 2013); (iii) the influx of seawater with high concentrations of cations such as Na⁺, Mg²⁺, and Ca²⁺ exchanges NH₄⁺ from the exchange sites in wetlands (Giblin et al. 2010; Weston et al. 2010), which makes NH₄⁺ subject to transport from inland into sea resulting in eutrophication; (iv) an increase in SO₄²⁻ due to SWI promotes organic matter mineralization in wetland sediments resulting in an increased release of NH₄⁺ from organically bound N (Lamers et al. 1998); soil water intrusion to inland wetlands can increase dissolved NH₄⁺ concentration as a result of microbial processes that inhibit nitrification (Joye & Hollibaugh 1995) and/or reduction of NO₃ to NH₄⁺ (Gardner et al. 2006). However, these findings are based on the studies conducted on soil cores or small plots; therefore, there is a need for landscape- and ecosystem-scale studies to understand fully the above processes and determine the mechanisms and extent of nutrient transformation and transport from inland to the sea.

Along the western and eastern coasts of Florida, there is clear evidence of SWI gradually impacting the freshwater aquifers. Nearly 90% of the drinking water supply in South Florida comes from underground aquifers. The soils in South Florida are extremely porous due to limestone subsoils, therefore, they are favorable for an increased rate of SWI. Most crops are sensitive to high salinity leading to a decline in crop growth and production.

The data from a sea-level monitoring station in Key West have shown that the mean sea level has risen about 228 mm over the last 100 years. This is slightly greater than the global rise of sea level, which is 170–210 mm (Hagemann 2016). Future sea-level projections compared with the base level in 1992 show that the average sea-level rise of 154–254 mm is expected by 2030, 356–864 mm by 2060, and 787–2,057 mm by 2100 (Hagemann 2016). The rising sea-level increases flooding risks to coastal regions, which also includes agricultural lands. To some extent, this risk is managed by an extensive network of canals to provide adequate drainage. The canal water levels are monitored regularly to evaluate the potential for rising groundwater levels as a result of SLR.

The Central and South Florida Water Management system was built in the 1950s without the knowledge of current and future accelerated rates of SLR. However, since 2008 SFWMD has been rigorously working on developing short- and long-term strategies to manage the impacts of accelerated SLR.

As of 2011, about 1,200 km² of the mainland part of the Biscayne Bay was impacted by SWI (Hagemann 2016). In an area of 24 km², the saltwater front was mapped further inland than it was in 1995 (Hagemann 2016). The studies have shown that saltwater intruded along the base of the aquifer.

Blanco et al. (2013) evaluated temporal and spatial changes in chloride concentrations in shallow and deep groundwater in Southwest (Collier and Lee counties) and Southeast (Broward and Miami-Dade counties) Florida between 1985 and 2000. Their study provided clear evidence of the impact of SWI on the quality of freshwater aquifer in South Florida. The same study reported that chloride concentration in Southwest Florida was increased from 132 to 230 mg/L for shallow wells (0–43 m), while for deep wells (greater than 43 m) the increase was from 392 to 447 mg/L (Figures 3 and 4).

Similarly, in Southeast Florida, chloride concentrations in shallow wells (less than 33 m) increased from 159 to 470 mg/L and in deep (greater than 33 m) wells from 1,360 to 2,050 mg/L during the same period (1985–2000) (Figures 5 and 6). Based on the US EPA guidelines on allowable chloride concentration in drinking water, the authors estimated that between 1985 and 2000 Southwest and Southeast Florida lost about 12–17% of the freshwater resource due to SWI.

In Florida, with the accelerated rate of SLR in recent years, in part due to climate change and variability, and increased pumping of groundwater for irrigation and meeting other competing needs due to the growing population, addressing the impact of SWI on freshwater resources and on soil health is a pressing priority.

The most effective technique to mitigate salt buildup in the soil from any source is to drain the salts by using fresh quality water so that the excess salts are drained below the rooting zone to overcome negative effects of salt concentration on root growth, which in turn affects the above-ground portion of the plants. This may sound a very simple and effective technique but may not be feasible in all cases. The availability of good quality freshwater in the vicinity of saline soils is often a major bottleneck, because over time the available water in the saline soil zone may also be impacted by salt; therefore, using brackish water for leaching the saline soil will not be effective in leaching salts. In addition, some of the soils in the saline soil zone may not be well drained, which will limit the leaching of salts from the root zone. This is the limitation in soils with hardpan and/or shallow water tables. In this review paper, we summarize the potential and limitations of selected soil amendment techniques in mitigating effects of soil salinity.

Several studies in different parts of the world have documented the beneficial effects of gypsum amendment in amelioration of soil salinity, thereby increased a yield response of different crops (Shainberg et al. 1989; Ilyas et al. 1997; Hanay et al. 2004; Silveira et al. 2008; Gharaibeh et al. 2009; Niazi et al. 2009; Makoi & Verplancke 2010) or gypsum plus manure (Smith 2009; Cha-um et al. 2011; Kahlon et al. 2012; Murtaza et al. 2013). In addition to its positive impact on soil salinity, gypsum amendment of agricultural soils has been demonstrated to have multiple benefits, such as (i) improving soil structure and water infiltration; (ii) mitigating aluminum toxicity; (iii) decreasing swelling of clays; (iv) excellent source of plant-available calcium and sulfur to those soils deficient in those nutrients; (v) decreasing bulk density; (vi) helping to mitigate phosphorus runoff from soils; (vii) preventing soil erosion, and (viii) increasing N fertilizer efficiency and availability of soil potassium (K).

Gharaibeh et al. (2009) conducted a study in Jordan to demonstrate the reclamation of a sandy clay loam soil using gypsum that was a byproduct of phosphate fertilizer, generally called ‘phosphogypsum’ (PG), which has very similar properties to those of the mined gypsum. They used moderately saline (EC = 2.16 dS/m) irrigation water for leaching of Na⁺ from gypsum-amended soil. They used PG rates at 5, 10, 15, 20, 25, and 32 Mg/ha. Their results showed that 32 Mg/ha PG amendment resulted in a 96% reduction in total Na⁺ in the soil and increased soil hydraulic conductivity (6.8 mm/h) compared with that of non-gypsum-amended soil (5.2 mm/h).

The application of PG at the rate of 32 Mg/ha is expected to require more cost for fuel and labor for transportation and spreading the material. However, PG is a waste byproduct during phosphate fertilizer production from rock phosphate, and the phosphate fertilizer industry is interested in the potential use of this byproduct because of extra costs and environmental concerns involved in storing PG. However, it is necessary to ensure that the specific PG to be used as a soil amendment is free of other contaminants. In addition, gypsum is also reported to have some negative impacts on nutrient solubility and leaching. Gypsum could negatively affect P, Fe, and Mn solubility; and nutrient uptakes by plant roots and thereby plant growth (Elrashidi et al. 2010). Similarly, gypsum-amended soils showed a reduction in exchangeable K and Mg (Syed-Omar & Sumner 1991).

The ultimate mechanism of mitigation of salinity between sulfuric acid and elemental sulfur is the same. However, elemental sulfur has a lag period of up to 4–6 weeks or longer in cooler soils for the microbial transformation of elemental sulfur into sulfuric acid, which is also dependent on soil temperature, humidity, and soil aeration. The additional benefit of this amendment is the reduction of soil pH, which promotes plant uptake of several nutrients.

Unlike the other amendments described above, there is a need to find low cost locally available materials especially for use in developing countries, which have seen an increase in soil salinity problems due to poor irrigation management. Soil mulching is an age-old technique in practice for a very long time in agriculture for improving soil's physical and chemical properties, minimizing soil erosion, reducing extreme fluctuations in soil temperature, and improving soil nutrient availability. Mulching also reduces soil evaporation losses, thereby improving water-use efficiency, reducing excess salt buildup, and minimizing salinity effects on plant growth and production (Zhang et al. 2008; Abou-Baker 2011; Dong 2012; Shirish et al. 2013).

Moniruzzamani & Shamim (2015) used rice straw to evaluate the effects of different mulching techniques to mitigate soil salinity and sweet gourd yield response in Bangladesh. They compared: (i) mulching at the bottom of the plow layer of the pit; (ii) top of the soil surface of the pit; (iii) combination of (i) and (ii); and (iv) no mulch. The EC measurement 75 days later was significantly lower in mulching at the bottom and top of the soil surface of the pit (4.98 dS/m) when compared with that of no mulch (8.82 dS/m). The former treatment resulted in significantly greater yield (18.5 Mg/ha) compared with that in the rest of the treatments. Yield from no mulch treatment was only 10.6 Mg/ha. Their results distinctly showed the beneficial effects of mulching to reduce salinity, thereby increasing the yield.

Al-Dhuhli et al. (2010) evaluated the effectiveness of date palm leaves and plastic mulch in reducing soil evaporation and maintaining favorable soil temperature for sorghum production in Oman using two irrigation water salinity levels (3 and 6 dS/m). They concluded that date palm leaves and mulch were more effective than plastic mulch in reducing salt accumulation in the soil, soil water loss by evaporation, and soil temperature, and in turn resulted in greater sorghum yield. This study also confirmed that salt accumulation can be reduced by mulching by minimizing the evaporation water loss from the soil.

Taufiq et al. (2017) investigated the effects of rice straw mulch (3.5 Mg/ha) and other amendments, including 5 Mg/ha gypsum, or manure and combination of gypsum and manure, on amelioration of soil salinity and peanut yield response in Indonesia on soil with 12 dS/m salinity. Mulching and 5 Mg/ha gypsum decreased soil EC, but peanut yield did not increase. They also reported favorable responses with mulching plus 750 kg/ha sulfur amendment.

Biochar is a charcoal-type material produced by ‘pyrolysis’ of biomass, of different origin, i.e. heating at a temperature typically between 300 and 700 ^oC under oxygen-deprived conditions. Although biochar can be produced from biomass of any type, the focus is on the use of agricultural wastes such as straw, rice husk, peanut hull, yard clippings, or tree trimmings. This provides an economically and environmentally sound way to recycle these wastes. Biochar is high in C content; therefore, it has been investigated as an amendment on agricultural soils for carbon sequestration and to improve soil properties. Recently, there is also an interest in exploring the benefits of biochar amendment to alleviate the negative effects of saline soils to improve crop production and quality.

High-specific surface area, CEC, and microporosity are the properties of biochar which are responsible for enhancing the soil productivity when biochar is amended to agricultural soils (Thies & Rillig 2009). These properties of biochar are responsible for increased retention of water and nutrients in poor soils amended with biochar. Furthermore, biochar can also absorb heavy metals, pesticides, and other contaminants (Rhodes et al. 2008; Beesley et al. 2010; Kookana 2010; Beesley et al. 2011, p. 201; Buss et al. 2012).

Jin et al. (2018) conducted a pot experiment to evaluate the effects of biochar amendment (15, 30, and 45 g/kg) on a saline-sodic soil (pH 8.53; CEC = 23.91 μS/m; clay content = 35.6%) and evaluated sodium accumulation in rice plants as well as yield and quality. Biochar amendment significantly decreased Na⁺ accumulation, while increasing rice yield even at the lowest rate of biochar amendment.

She et al. (2018) reported that the biochar amendment reduced the negative effects of saline irrigation water; therefore, it was beneficial to increase tomato production by mitigating the negative effects of salt stress. They also reported the highest rates of photosynthesis and respiration at the highest rate of biochar amendment but using non-saline irrigation water. Biochar amendment adsorbed Na ions, while released Ca, K, and Mg ions into the soil solution hence increased their availability to plants under saline conditions. Once again, the question is the economic feasibility of the biochar amendment. The lowest rate they studied was 45 Mg/ha.

Thomas et al. (2013) investigated the effects of two rates (5 and 50 Mg/ha) of biochar, made from Fagus grandifolia sawdust prepared at 378 °C, on the response of two herbaceous plants Abutilon theophrasti and Prunella vulgaris under induced salt stress condition created by adding 30 g/m² NaCl. The salt-induced mortality of seedlings was completely mitigated only at 50 Mg/ha biochar treatment, but not at the lower rate (5 Mg/ha). However, biochar treatments did not affect photosynthetic carbon gain, water-use efficiency, or chlorophyll fluorescence in both plant species compared with the control. The authors claim that the mitigation of salt stress by the biochar amendment was due to salt sorption. However, this favorable response was only observed at a 50 Mg/ha application rate, which is quite high and could be less feasible for large-scale field applications due to high input costs, and thus net returns may be very marginal or not possible. Therefore, it is important to assess the economic feasibility of using biochar as a soil amendment given the costs of biochar production through pyrolysis as well as labor and fuel costs for its application.

Hydrogels are synthetic polymers, produced by chemical stabilization of hydrophilic polymers in a tri-dimensional network (Montesano et al. 2015), with the ability to absorb a large amount of water, and therefore called ‘superabsorbers’. In the past, agricultural application has been mostly as a soil amendment to absorb and retain water and subsequently make it available to plants, therefore facilitating increased water conservation and water-use efficiency (Horie et al. 2004; Zohuriaan-Mehr & Kabiri 2008). Recently, attention has been expanded to investigate the role of hydrogels to mitigate salt stress on various crops (Sayed et al. 1991, p. 199; Szmidt & Graham 1991; Wang et al. 2004; Dorraji et al. 2010). The major limitation concerning the use of synthetic polymers as soil amendments has been lack of and/or slow degradation of these polymers resulting in prolonged-lasting of these residues in the soil, raising concerns about negative environmental and health impacts (Mooney 2009; Timilsena et al. 2015). To overcome the above limitation and to maintain sustainable agricultural practices, i.e. to minimize waste residues in our soil and water resources, the development of polymers using biodegradable biomass-based raw materials has been investigated. The emphasis has been on using lignin or agricultural residues to develop water-insoluble hydrogels. Lignin is abundantly available as a residue of biorefinery processes or pulp industry (Wang et al. 2016). Agricultural biomass residues that can be used include bagasse or sorghum, which are abundantly available in most places. Agricultural residues can be used to produce a hydrogel with high cationic exchange capacity and swelling characteristics that could reduce salinity risks. Furthermore, this hydrogel and gypsum can be used together to achieve maximum absorption of Ca²⁺ and reduce the impacts of Na⁺ and Cl⁻ ions. Further research is needed to provide additional data on the appropriate rate of hydrogel amendment to mitigate the negative effects of salinity on plant growth and production under a range of salinity stresses.

Soil salinization due to poor irrigation management is a global issue. In addition, soil salinity SWI due to SLR is becoming a major threat that impacts agricultural production and natural resources. Addressing soil salinization is of significant importance to global food and water security. However, there is still a critical knowledge gap on how soil salinity alters the chemistry, soil health, nutrient transport, and GHG emissions in agricultural soils. Such knowledge is extremely critical to develop strategies to mitigate these negative impacts.

The major knowledge gap related to climate change and SWI is a lack of clear understanding with some degree of certainty on the extent of the problem. Current and future projections both in areal coverage and severity are not well known.

The availability of reliable data on soil health indicator parameters, water quality, and other ecosystem processes remains a major issue. For example, data on chloride concentrations, which is an indicator of water salinity, are scarcely available. Chloride concentration data are required to predict the trends in salinity changes in groundwaters. Guha & Panday (2012), from their modeling work on SWI, indicated that the lack of accurate and adequate measured chloride data needed to calibrate and validate models was a major bottleneck.

Salt accumulation can be leached below the crop root zone using freshwater (rainfall or irrigation). However, climate change and variability cause extreme weather events such as cycle droughts and flooding. Erratic and often deficit rainfall in some areas is inadequate to leach salts below the root zone and thus groundwater is used to meet freshwater needs. Therefore, a better understanding of the mechanisms and negative impacts of SLR and SWI on the freshwater resources, soil health, nutrient transformation and transport, and GHG emissions is a global priority to develop best management practices to mitigate these negative impacts to safeguard food and water security. However, our understanding of the potential of management practices such as soil amendments, leaching the salts from the crop root zone, also needs to be enhanced so that alternative management strategies are in place to mitigate the negative impacts of SWI on agricultural production and soil health.

Another area where the knowledge gap exists is the possibility of genetically improving crops for increased salt tolerance to keep salt-affected lands under production. Similarly, there is only a limited understanding of hydraulic and physical management strategies to mitigate SWI.

With projected increases in global SLR, much of the attention with regards to its impact has been on the SWI of groundwater and the threat it poses to livelihoods and freshwater resources coastal regions. However, limited attention is given to the impact of the landward movement of saltwater, through the surface and/or groundwater connections, on the salinity of the soil, overall soil health, and agricultural productivity. Land-use changes such as the conversion of wetlands to agricultural areas due to population pressure and urbanization will result in a lower groundwater level which in turn would result in more water flowing inland from the sea. These changes will contribute to increased risk of salinization of inland surface water and the soil, as the SLR is expected to continue at an accelerated rate into the future. This impact will be severe as the climate crisis enhances the risk of drought and storms in the coastal and inland regions with increased groundwater pumping to meet crop water requirements. The impact of soil salinity on soil health and water quality, and overall agricultural productivity and thereby food and water security could be significant. More research is needed to better understand the scale of the impacts and develop integrated management practices to mitigate the negative impacts. Research should involve different scales and employ multi-pronged approaches including monitoring of changes in major indicators of soil health such as soil organic matter, microbial biomass, soil bulk density, water infiltration, aggregate stability, soil pH, and status of major nutrients such as N and P. Furthermore, there is a compelling need to understand the role and mechanism of different management strategies to mitigate the negative effects of SWI on soil salinity and crop productivity. The potential management strategies include soil amendments such as engineered biochar, salt repelling hydrogel, gypsum, and organic mulch. Improved understanding of the processes and impact of soil salinity on soil health will assist in designing best management practices to mitigate the impacts of soil salinity on agricultural soils and crop productivity.

This material is based upon work that is supported by the National Institute of Food and Agriculture, U.S. Department of Agriculture, under award no. 2020-67019-31163. Any opinions, findings, conclusions, or recommendations expressed in this publication are those of the authors and do not necessarily reflect the view of the U.S. Department of Agriculture.

All relevant data are included in the paper or its Supplementary Information.