It Is Time to Develop Sustainable Management of Agricultural Sulfur

Globally, sulfur (S) applications to croplands result in S inputs that often exceed historical atmospheric deposition. Sulfur is applied to crops as a fertilizer, fungicide, soil conditioner, pH regulator, and carrier for other elements. However, excess S in soils and aquatic ecosystems can have detrimental ecological and biogeochemical consequences, including soil base cation depletion, surface water acidification, hydrogen sulfide toxicity, and increased production of methyl mercury. The dichotomy between S benefits to crops and environmental consequences parallels that of nitrogen and phosphorus; however, there has not yet been a focus on developing sustainable S management plans in agriculture. We review the current literature on S cycling in agricultural systems and propose solutions that reduce S inputs, losses, and ecological consequences, including field applications of organic matter, adaptation of precision agriculture, and implementation of total maximum daily loads. We suggest opportunities for technological innovation, including analysis of remote sensing imagery to identify location and timing of S deficiencies and stresses, crop genetic modification to reduce S requirements, inoculation of plants with arbuscular mycorrhizal fungi to enhance plant S acquisition, and remediation of wetlands and other anoxic environments with high S loads. We conclude with areas for continued research on S biogeochemistry.

• It is critical to address optimal S applications in agricultural systems to meet crop needs while reducing environmental consequences of excess S • Field additions of organic matter, adaptation of precision agriculture, and robust surface and groundwater monitoring plans improve agricultural S management • Precision agriculture using satellite imagery and crop genetic modifications may help manage crop S needs; remediation can reduce excess S

Supporting Information:
Supporting Information may be found in the online version of this article. 10.1029/2023EF003723 2 of 18 Following the United States' passage of the Clean Air Act and its Amendments in the 1990s and similar legislation in Europe, atmospheric S deposition drastically decreased (Aas et al., 2019;Baumgardner et al., 2002;Benish et al., 2022;Crippa et al., 2016).In China and India, while atmospheric S deposition almost doubled from the 1980s-2000s with industrialization (Hoesly et al., 2018), regulations have more than halved SO 2 emissions from 2010 to 2017 (Zheng et al., 2018).Global projections for the future (2095-2099) suggest that atmospheric S deposition across the Northern Hemisphere will decrease by 70%-90% compared to 2005-2009 values (Feinberg et al., 2021).Consequently, a combination of decreased atmospheric S deposition globally, intensification of agriculture, and increased risk of crop disease has led to increased S applications to croplands to ensure optimal crop yield and quality in the Global North (Camberato et al., 2012;Jez, 2015;McGrath et al., 2003), Asia (Fan & Messick, 2005;Khurana et al., 2008;Sahrawat et al., 2007), Africa (Friesen, 1991;Kang & Osiname, 1976), Oceania (Edmeades et al., 2005), and South America (de Castro Pias et al., 2019;Fageria & Nascente, 2014).
As a result, the use of S amendments in croplands has become increasingly common in the last few decades (Haneklaus et al., 2008;Schnug, 1991) and is likely to continue into the future as atmospheric S deposition declines.
The increased use of S in agricultural systems results from a shift in human manipulation of the S cycle from emissions associated with industrial activities to inputs associated with agricultural activities (Hinckley & Driscoll, 2022).Sulfur amendments to agricultural fields (Hinckley et al., 2020;Hinckley & Matson, 2011;Schueneman & Sanchez, 1994) are often on par with average agricultural N inputs in the U.S. (Basso et al., 2019) and far exceed global P inputs (Alewell et al., 2020).Yet there is a notable gap regarding the sustainable management of S in croplands-that is, both optimizing inputs for crop needs and reducing the export of excess S to adjacent ecosystems, which could have negative consequences.Sulfur deficiency will continue to increase in the future as yields increase and historic S adsorbed in soils is released into surface waters (Hinckley & Driscoll, 2022).
In this article, we begin with a brief introduction to the negative ecological and biogeochemical effects of excess S in the environment.We then discuss management strategies for S use in agricultural systems that meet the needs for plant growth and crop yield while reducing the environmental consequences.We draw upon strategies used in crop N and P management, highlighting where S practices differ, and suggest opportunities for co-management of these elements.We conclude with a discussion of remaining questions that can guide future research to inform implementation of sustainable S management.

Consequences of Excess Sulfur in the Environment
Historic and legacy effects of excess S from acid deposition provide important lessons regarding the likely environmental consequences of high S inputs to croplands (Hinckley et al., 2020).High levels of acid deposition led to depletion of soil base cations (Bailey et al., 2005;Likens et al., 1996), reduced forest health (Driscoll et al., 2001), and acidified soils and surface waters (Likens et al., 1996).In general, differences in soil buffering capacity (i.e., low in granite and igneous non-calcareous sandstones vs. high in calcite and carbonate bedrocks) controlled the extent of ecosystem degradation (Longhurst et al., 1993).As a result, the negative effects of acid deposition were more pronounced in the northeastern U.S. and Europe, where there are younger, rocky soilsand thus little soil S adsorption-compared to the older, more weathered, clay soils in the southeastern U.S. that have higher anion exchange capacity (Rice et al., 2014).Soil acidification led to acid stress on tree roots, which reduced nutrient and water uptake, along with root growth and mortality, and led to forest growth depressions and dieback in Europe and throughout the U.S. (Long et al., 2009;Pitelka & Raynal, 1989;Ulrich, 1990).
The consequences of historic acid deposition in forest ecosystems extended beyond trees.Soil macroinvertebrate abundance, fungal biomass, and nematode biodiversity decreased with acid deposition, altering rates of organic matter turnover and food availability for birds (Kuperman, 1996;Ruess et al., 1996).Declines in snail community richness and abundance, as well as salamander species composition (Beier et al., 2012), were similarly observed.
In aquatic ecosystems, those streams draining younger soils experienced elevated sulfate and aluminum (Al) concentrations and decreased pH.These biogeochemical effects led to a cascade of ecological changes, including declines in zooplankton, macroinvertebrate, and fish biodiversity and abundance in northern Europe and the northeastern U.S. (Driscoll et al., 2001;Likens et al., 1996;Longhurst et al., 1993).Today, streams in the southeastern U.S. continue to have high sulfate concentrations despite more than 80% reductions in atmospheric S deposition.This phenomenon indicates that there are legacy effects of high S inputs even in ecosystems that 10.1029/2023EF003723 3 of 18 responded minimally during the peak of acid deposition (Galloway, 1995;Rice et al., 2014).In locations such as the northeastern U.S., base cation depletion of soils has caused lasting consequences for forest stress tolerance, reproduction, growth, and stand health, as well as increased sensitivity to acid inputs and impaired surface water quality.These effects persist even with regulation of SO 2 emissions (Lawrence et al., 2015;Likens et al., 1996).
While acid deposition has decreased dramatically over the past few decades, there has been a concomitant increase in S losses from agriculture.Agricultural additions of S to soils are associated with enhanced sulfate leaching into groundwaters and overland flow into nearby surface waters (David et al., 2016;Hermes et al., 2021;Hinckley et al., 2020;Scherer, 2001).For example, California vineyard soils have two to four times more S than nearby grasslands as a result of agricultural S additions; however, S concentrations in leachate during rainfall events can be 16-20 times higher in vineyards than grasslands.Indeed, the rapid oxidation of applied elemental S to sulfate leads to the loss of nearly all S applied annually during normal water years (Hermes et al., 2021;Hinckley et al., 2011;Hinckley & Matson, 2011).While S leaching rates in California likely differ from other locations based on soil types, high leaching rates are particularly concerning given that excess freshwater sulfate could negatively affect aquatic organisms and/or mobilize heavy metals (Davies, 2007;Soucek, 2007).The spatial distribution of S use in agriculture differs from where acid deposition was historically elevated and where the effects were most pronounced.While remote forests and streams were highly impacted by acid deposition, excess agricultural S impacts affect ecosystems near croplands that are already suffering the consequences of high N and P loading.Yet, unlike highly regionalized acid deposition, agricultural S use is widespread and could exacerbate legacy soil acidification and base cation depletion, similar to N fertilization (Barak et al., 1997;Lucas et al., 2011;Schroder et al., 2011;Zhou et al., 2014).While most studies on watershed response to S inputs have focused on historic acid deposition from industrial processes, we expect some of the same patterns when the S source is agricultural activity.
What are the ultimate fates of agricultural S? Sulfate can be reduced to sulfide when it leaches from agricultural soils into sediments or wetlands with high organic matter and low oxygen content (Jez, 2015;Wang & Chapman, 1999).Sulfide is toxic to many aquatic macrofauna (Van Der Welle et al., 2006;Wang & Chapman, 1999;Zak et al., 2021).In macrophytes, elevated levels of sulfide can decrease nutrient uptake, aboveand below-ground productivity, rates of photosynthesis, and root respiration, while increasing blockage of gas pathways, radial oxygen loss, and root die-off (Armstrong et al., 1996;King et al., 1982;Koch et al., 1990;Koch & Mendelssohn, 1989;Lamers et al., 1998Lamers et al., , 2013;;Smolders & Roelofs, 1996).In fact, sulfide is one of the top three most impactful chemical stressors in freshwater anoxic sediments (Zak et al., 2021).Rice fields are particularly vulnerable to the toxic effects of sulfide due to the anoxic conditions that promote sulfate reduction, resulting in reduced root respiration and lower yields (Allam & Hollis, 1972;Bell, 2008).When rice straw is added to rice paddies, sulfide production is enhanced, and yields are further reduced (Gao et al., 2004;Kuo & Mikkelsen, 1981).Sulfate loading from agriculture into anoxic environments is compounded by salinization of freshwater ecosystems.Studies of saltwater intrusion have found that elevated levels of sulfide associated with salinization inhibit nitrification (Joye & Hollibaugh, 1995), even when there is only a short exposure of microbial communities to hydrogen sulfide (Joye & Hollibaugh, 1995).Elevated sulfide under saline conditions can also cause incomplete denitrification of (Brunet & Garcia-Gil, 1996;Senga et al., 2006).Alternatively, freshwater ecosystems can be recipients of excess sulfide if upgradient terrestrial ecosystems support reducing conditions that cause sulfate reduction.Sulfide in soil can bind to iron (Fe), mobilizing P, and resulting in subsequent pulses of P into nearby aquatic ecosystems (Caraco et al., 1989;Hinckley et al., 2020;Smolders & Roelofs, 1996;Zak et al., 2006).This internal eutrophication may be responsible for biodiversity losses in wetlands (Smolders et al., 2010).
High levels of sulfate reduction can likewise have ramifications for the bioavailability of toxic trace elements.For example, the conversion of inorganic mercury (Hg) into the neurotoxic and bioavailable form of methyl mercury (MeHg) is driven by anaerobic microbes, including sulfate reducing bacteria (Choi & Bartha, 1994;Compeau & Bartha, 1985).Therefore, increased inputs of sulfate to wetlands and anoxic sediments could increase the production, as well as the bioaccumulation, of MeHg within the food web (Jeremiason et al., 2006).Studies from Florida (U.S.) show that high levels of S from agricultural activity in the Everglades Agricultural Area enter nearby waterways in the Everglades Water Conservation Area (Bates et al., 2002).For example, Orem et al. (2011) and Bates et al. (2002) demonstrated that excess agricultural runoff is linked to elevated MeHg production.However, many questions still remain regarding the link between S use in agriculture and MeHg production, particularly in upland agricultural systems (Hinckley et al., 2020).

The Need to Develop Sustainable Sulfur Management Plans
The use of S in agricultural systems has increased since the mid-1900s.Sulfur is often applied as inorganic sulfate or as elemental S, which rapidly oxidizes to sulfate in the soil (Germida & Janzen, 1993;Hinckley et al., 2011).Over time, inorganic sulfate can be transported hydrologically or bind to organic molecules (Hermes et al., 2023).Indeed, more than 95% of soil S is found in the form of organic S (Kertesz & Mirleau, 2004;Rehm & Clapp, 2008;Scherer, 2001;Tabatabai, 1984).While organic S is less bioavailable to plants (Kertesz & Mirleau, 2004), it influences soil biogeochemical processes (e.g., DOM reactivity, Hg methylation) (Graham et al., 2017;Hermes et al., 2023;Likens et al., 2002), as well as aquatic photodegradation to sulfate (Ossola et al., 2019;Poulin, 2023).Sulfur amendments to Midwest U.S. corn and soybean average ∼40 kg S ha −1 yr −1 , with even higher rates for other S-intensive crops across the U.S. and elsewhere (Hinckley et al., 2020).Sulfur application rates as high as 100-300 kg S ha −1 yr −1 are common in California vineyards to prevent powdery mildew infection (Hinckley & Matson, 2011), while over 500 kg S ha −1 yr −1 is often used in the Everglades Agricultural Area to reduce soil pH 10.1029/2023EF003723 5 of 18 for vegetables and to liberate P in sugarcane fields (Schueneman & Sanchez, 1994).These S application rates are on par with average N inputs of ∼160 kg N ha −1 yr −1 applied to corn in the Midwestern U.S. (Basso et al., 2019) and exceed maximum global averages of 19 kg P ha −1 yr −1 applied to agricultural fields in China (Alewell et al., 2020).Unfortunately, data are not currently available at the country or global level to determine total agricultural S loads, but these individual case studies suggest that S inputs are high and continuing to increase.
An effective management plan for S must address several important factors.First, many countries have established protocols for sustainably managing N and P in agricultural fields (Drohan et al., 2019;Garske et al., 2020;Parry, 1998); these protocols could be adapted to address S as well, although the multiple uses of S in croplands (beyond as a fertilizer to meet crop nutritional requirements, see Supporting Information S1) must be considered.Similarly, emerging strategies targeted at reducing S losses from fields often simultaneously manage N and P both directly (i.e., the management strategy reduces N and P losses) and indirectly (i.e., retention of S leads to increased N use efficiency and reduced P mobilization).Second, an effective S management plan must consider the dynamic ways in which climate change and groundwater withdrawal are affecting the terrestrial S cycle.As climate change alters temperature and precipitation regimes (IPCC, 2022), sulfate mobilization and reduction rates will likely increase.Changes in temperature and precipitation patterns will also affect the distribution of pests and pathogens (Caffarra et al., 2012;IPCC, 2022;Skendžić et al., 2021;Tang et al., 2017), shifting the requirements for S use as a fungicide and pesticide.Increased evapotranspiration associated with climate change and lowered water tables due to groundwater withdrawal (IPCC, 2022) will enhance salinization of coastal areas (Thorslund et al., 2021), further increasing S inputs to soils.Thus, there is a need for the development of a holistic approach to S management that considers coupled elemental cycling and a suite of approaches that allows for adaptation to climate change and local-to-global scale water pressures.

Approaches to Agricultural S Management
Effective S management must target a reduction in soil S losses, which would decrease the need for agricultural S inputs as well as reduce environmental impacts.This goal is similar to that of many N and P management plans, which seek to reduce nutrient losses.Nitrogen and P management strategies use erosion control, surface runoff reduction, and sediment impoundments to prevent nutrient runoff (Sharpley et al., 2003(Sharpley et al., , 2006;;Xia et al., 2020).While these strategies can also be applied to S management, most S is generally dissolved and thus not bound to sediment and soil particles when transported (Scherer, 2009), and these strategies are less effective.There are three main pathways by which S can be lost from soils: harvesting of crop tissue, leaching (Hinckley & Driscoll, 2022), and gas evasion.In field crop harvests, S is directly removed with the crop tissue, resulting in the loss of up to 15-30 kg S ha −1 yr −1 , particularly in S-rich crops such as soybean and cotton (Rennenberg, 1984).Since S leaching rates are regulated by soil characteristics, altering organic matter composition can reduce these losses and provide the co-benefit of increased soil S mineralization rates, improved soil structure, and increased water holding capacity (Dong et al., 2022;Lucas et al., 2014;Tejada & Gonzalez, 2007;Xia et al., 2020), especially in sandy soils.Organic additions include biochar, biosolids, manure, compost, sewage sludge, and green manure (Dick et al., 2008;Eriksen, 2008).Soil organic matter content can further be increased through the return of crop residues to the soil, which simultaneously reduces harvest losses of S (Dick et al., 2008, Franzen & Grant 2008).In either form, these organic amendments increase soil C content, which enhances soil capacity to build organic S stocks and reduces S leaching (Eriksen, 2008).The use of organic amendments generally meets the S needs of cereal crops (Badawy et al., 2011;Eriksen, 2008).Although organic additions are not sufficient for S-demanding crops such as oilseed rape, grass ley, kale, and onion, they significantly reduce S fertilizer requirements in these crops, and, consequently, overall S losses.For example, Larson et al. (1972) showed that cornfields planted in silty loam soil that had alfalfa residue or cornstalks returned to the soil after harvesting experienced a 45% increase in soil organic S and reduced S fertilizer requirements.
Note that while the use of organic matter additions rather than S fertilizers is promising for meeting plant S nutritional requirements, it is only effective as a nutritional supplement if growing seasons are long enough for available S to be mineralized (Eriksen, 2008).During sufficiently long growing seasons, these organic amendments can effectively meet crop nutritional S demands, increase the time duration for abiotic and biotic S-induced stress reduction in situ and in vivo, regulate soil pH, and decrease drought conditions.Furthermore, organic additions supply N and P to plants (Eriksen, 2008;Garske et al., 2020), increase N mineralization rates, and reduce erosion and runoff losses of C, N, and P (Sharpley et al., 2006).These co-nutrient benefits result in reduced need for N and P fertilizers and the presence of S amendments to serve as a carrier for N and P. Thus, organic additions have holistic benefits for soil and crop health that include increasing S, C, N, and P availability within soils and reducing losses of all four elements.
Another strategy for S management that uses tools adapted for N and P management is precision agriculture, which aims for optimum crop health and yield while reducing negative environmental consequences.This practice applies resources (i.e., nutrients and water) specifically at required levels when they are needed by the crop, thereby increasing nutrient and/or water use efficiency, reducing the build-up of excess resources, and minimizing gaseous and aqueous losses (Tilman et al., 2002).In some crops, the timing and amount of S amendments for fungicidal spraying are determined using proximal and remote sensors, which allow for variable spraying (Hedley, 2015;Sharpley et al., 2006).Optimizing amendment timing and amount for S nutrition can occur via soil testing for S concentration and forms (Dick et al., 2008), alteration in amendment type to promote slow-releasing elemental S applications rather than ionic S applications (Santoso et al., 1995;Scherer, 2001), and the use of S fertilizers in solid and spray form (Hinckley & Matson, 2011;Santoso et al., 1995).Often, the result is an overall reduction in the amount of fertilizer and irrigation used, as well as diminished losses from the field.For example, nutritional S needs are often most pronounced at mid-vegetative and reproductive periods (Zuber et al., 2013) and in growing seasons that occur prior to organic matter mineralization (Eriksen, 2008).Fertilizer recommendations can thereby be established for farmers based upon the S needs of their crops, which will have both environmental and economic (i.e., cost reduction) benefits, and education programs can be provided regarding these best management practices.Irrigation management, an important component of precision management, similarly has direct effects on S losses.Because sulfate dissolves readily in water (Hinckley et al., 2011;Scherer, 2009), reducing the amount of excess water added through irrigation prevents the loss of sulfate leached into groundwater and reduces the frequency at which S applications are required for all crop S purposes.Practitioners could therefore target form and dose of S across space and time to meet the demands of their crops for all S needs, beyond nutritional requirements.
Since S is often used as a carrier ion for N and P and these nutrient cycles are linked, precision agriculture could target how, when, and where to apply S-based NPK fertilizers to optimize plant nutrient uses and diminish losses.For example, a pot experiment with P and S fertilizers found that the retention of S fertilizer was determined both by soil properties and by the proximity of S to P fertilizers (Santoso et al., 1995).Co-locating these amendments or using S-coated P fertilizers could improve management of both nutrients.Accordingly, selecting forms of S that also provide other macronutrients is beneficial to crops.Precision agriculture has been applied successfully to manage N (Cui et al., 2010;Diacono et al., 2013;Spiertz, 2010) and P (Drohan et al., 2019;Garske et al., 2020;Sharpley et al., 2003Sharpley et al., , 2006)), increasing nutrient use efficiency by over 350%, decreasing fertilizer use by up to 80%, and saving 5-60 USD per hectare (Diacono et al., 2013).The use of precision agriculture can be expanded to include S, thereby allowing for the collective management of N, P, and S deficiencies and coupled biogeochemical cycles, especially given the potential of S amendments to increase N use efficiency and thus ultimately to have impacts on N retention and leaching.There is a need for co-management of nutrients in agricultural fields, since N, K, and magnesium (Mg) deficiencies can affect uptake of S from soils and S, in turn, can influence the uptake of these other important nutrients (Courbet et al., 2019).
From a regulatory standpoint, total maximum daily loads (TMDLs) could be a good model for encouraging precision agriculture, particularly in regard to non-point agricultural inputs of S to surface waters (Sharpley et al., 2006).The TMDL framework was created as part of the 1972 U.S. Clean Water Act and has since been adapted in other countries to determine the maximum pollutant amount allowed for a water body to meet water quality standards (Parry, 1998).Policies are then created to reduce the influx of these point and non-point pollutants into surface waters to ensure that concentrations remain below a set threshold (Elshorbagy et al., 2005).Lemly (2002) describes seven steps to creating an environmentally-safe TMDL for contaminants: (a) delineate the water body of interest; (b) determine concentrations of the relevant contaminant and assess the hazard to biota; (c) calculate the load from all sources, concentrations, and volumes of the contaminant entering the waterbody; (d) estimate the retention capacity of the contaminant within the water body; (e) calculate the allowable daily load of the pollutant and determine how the current load could meet this allowable load; (f) allocate this allowable load between the different sources of the contaminant; and (g) monitor to determine whether the reduction in contaminant load has the desired water quality effects.TMDLs have been created for N and P-due to their role in eutrophication and dead zones-as well as a suite of other elements (Elshorbagy et al., 2005), but currently there are no TMDL levels for S. Establishing TMDLs for S will encourage farmers to adopt strategies 10.1029/2023EF003723 7 of 18 that reduce leaching losses attributed to different agricultural S uses.It will also have co-benefits for reducing the mobilization of other associated elements, such as N and Al 3+ .

Integration of Technology and Remediation Into Agricultural Sulfur Management
The development of new technologies allows farmers to incorporate precision agriculture approaches more effectively.Satellite imagery, including hyperspectral imagery, is a powerful tool for diagnosing crop S deficiency (Diacono et al., 2013;Mahajan et al., 2017), indicating a need for S fertilizer application only when deficiency is detected.These same technologies, along with deep learning approaches, can be applied to the detection of pathogen presence (Cséfalvay et al., 2009;Gutierrez et al., 2021;Oerke et al., 2016).When infected areas are found, proximal locations can be targeted for S fungicidal additions while more distant areas can remain free from spraying.Precise quantity and timing of S fertilizer and pesticide applications can occur through variable rate technologies when sensors and computers are mounted on tractors.Modified S inputs can then occur as the tractor moves through the field.Both satellite imagery and tractor sensors have already been applied for N, P, and pesticide delivery decisions globally (Chen et al., 2018;Fulton et al., 2001;Matese & Di Gennaro, 2015;Mirzakhaninafchi et al., 2022)-and some tractor companies are building these technologies directly into the tractor design-allowing the same technology to be useful for simultaneous management of N, P, and S. In addition, the use of variable rate technology in irrigation systems could also be effective for S management and has already been applied for N and P management (Hedley & Yule, 2009;Xia et al., 2020).These irrigation technologies use data from moisture sensors and models to determine different water application rates required across cropland rows.They can be expanded to fields receiving S additions to manage all three macronutrients together and would potentially reduce crop fertilizer requirements.While advanced technologies for precision agriculture are promising, recent studies show that farmers often do not use them to guide fertilizer programs due to lack of outreach regarding emerging methods, technical issues with equipment, and limited software access (Arbuckle & Rosman, 2014;Robertson et al., 2012).Partnerships with agribusinesses and other crop advisors are needed to support farmers to use these tools (Arbuckle & Rosman, 2014).
An emerging aspect of S management is the use of biotechnological approaches, which can be applied to alter plant S requirements through genetic modifications and crop breeding programs.Modification of crop genotypes would allow for the creation and selection of crops with lower S nutritional requirements.For example, gene manipulation could enhance plant rates of sulfate uptake and biosynthesis of S and S-rich seed storage proteins; this genetic approach could be paired with crop breeding programs that produce optimized varieties with increased sulfate use efficiency (Hoefgen & Hesse, 2008;Khan & Hell, 2008;Williams & Cooper, 2004).Genetic manipulation can also develop crops with enhanced resistance to diseases and abiotic stresses including water stress, such as through increased antimicrobial protein and metabolite production (e.g., glucosinolates, thionins, and defensins); this, in conjunction with crop breeding programs, can dramatically reduce the need for S additions for pathogen and stress resistance (Kruse et al., 2005;Wittstock & Halkier, 2002).While these genetic modifications are only beginning to target crop S demands (Khan & Hell, 2008), they have previously been applied to crop N (Hirel et al., 2011;Li et al., 2020;Mcallister et al., 2012) and P needs (Bovill et al., 2013;Shenoy & Kalagudi, 2005) and have been implemented in conjunction with plant breeding programs (Herridge & Rose, 2000).Future biotechnological approaches that simultaneously target enhanced N, P, and S use efficiency and biochemical utilization could be particularly effective to accelerate optimization of management approaches.
The second novel S biotechnological management approach is the use of arbuscular mycorrhizal fungi (AMF) inoculations to enhance plant S acquisition.AMF are symbiotic fungi that can associate with plants, allowing them to grow in low nutrient conditions.Arbuscular mycorrhizal fungi colonize plant roots, increasing soil S uptake into plant tissues (Narayan et al., 2021(Narayan et al., , 2022)).In return, AMF receive sugars from the plant (Parniske, 2008).When plants have depleted sulfate in the soil, creating a sulfate depletion zone around the roots, extra-radicular hyphae of AMF-which have elevated sulfonate desulphurizing bacterial communities compared to surrounding soil-can upregulate plant sulfate transporters and bring in sulfate from surrounding areas that plants are unable to access (Gahan & Schmalenberger, 2015;Giovannetti et al., 2014;Kertesz et al., 2007;Narayan et al., 2021).For example, barrel clover (Medicago truncato) (Sieh et al., 2013), carrot (Daucus carota) (Allen & Shachar-Hill, 2009), and trefoil (Lotus japonicus) (Giovannetti et al., 2014) plants inoculated with AMF exhibit increased sulfate uptake, particularly in S-limited soils.Not only are AMF beneficial for plant S uptake efficiency, but some AMF are similarly important for the acquisition of other essential nutrients (e.g., N (Govindarajulu et al., 2005;Guether et al., 2009;Kobae et al., 2010), P (Smith et al., 2011), Mg (Prasad et al., 2019), and Fe (Verma et al., 2022)); inoculation with AMF would thereby benefit plants in the acquisition of a suite of macronutrients and decrease overall fertilizer needs.The benefits of S-targeted AMF also include increased drought and salt tolerance, pathogen protection, soil aggregation, and surface area of interaction with beneficial soil microbes, while reducing stresses from toxic metals (Ruiz-Lozano et al., 1995;Waller et al., 2005;Wu et al., 2018).Since AMF would therefore perform many of the same roles as S amendments without exacerbating S losses, crop inoculation with AMF could be an important alternative to inorganic S applications in agricultural fields.
Another solution targets a reduction in the environmental impact of agricultural S losses through biological and chemical remediation of wetlands and other anoxic environments that accumulate high levels of S. Bioremediation includes constructed wetlands (Chen et al., 2016;Di Luca et al., 2011;Wu et al., 2011Wu et al., , 2012)), permeable reactive barriers (Benner et al., 1999;Jeen et al., 2014), the use of S-oxidizing bacteria to convert sulfide into sulfate (Nguyen et al., 2022;Yang et al., 2021), and the introduction of plants with high S requirements to riparian buffers.These approaches have generally been applied to contaminated groundwater (Jeen et al., 2014;Wu et al., 2012), industrial wastewater (Chen et al., 2016;Di Luca et al., 2011), and mine drainage (Barton & Karathanasis, 1999;Benner et al., 1999), but their efficacy in removing sulfate could be applied for agricultural purposes if placed between crop fields and receiving waterbodies.Bioremediation could reduce sulfate contamination by up to 70% through the formation of insoluble metal sulfide (e.g., Fe sulfide), calcium sulfate, and Fe hydroxy sulfate precipitates (Benner et al., 1999;Chen et al., 2016;Jeen et al., 2014;Wu et al., 2012), while also reducing concentrations of nitrates, P, and organic carbon (Chen et al., 2016;Wu et al., 2012).The cultivation of plants with high S requirements (e.g., Tropaeolum and Brassicacea) within constructed wetlands or riparian buffers could also either be used as a source of high S organic matter fertilizer on fields or a supplemental S source for livestock, reducing the need for S fertilizer applications to support high S field crops as animal feed (Bloem et al., 2008;Haneklaus et al., 2008) and providing important ecosystem services (Anbumozhi et al., 2005;Cole et al., 2020;Hansson et al., 2005;Liu et al., 2012).
In addition to bioremediation approaches, chemical remediation could be applied to reduce the effects of high S runoff to surface waters.This could occur through redox manipulation (e.g., aeration, oxygenation, nitrate additions) that would reduce sulfate reduction.For example, summer nitrate additions to the industrially polluted and seasonally stratified Onondaga Lake in New York resulted in enhanced nitrate reduction and decreased pelagic MeHg concentrations (Matthews et al., 2013;Todorova et al., 2009).Yet while this chemical manipulation was effective at controlling redox processes, nitrate additions are only appropriate when large watershed additions of N are not present, and the ecological consequences must be carefully considered.

Research Needs
Given the widespread and important use of S in agricultural activities for crop quality and yield (Hinckley & Driscoll, 2022), there is a need for S management to begin immediately, and for sufficient data collection to guide initial management plans.Early S management strategies can be informed by N and P management schemes.However, inputs of S to croplands for purposes beyond plant nutritional requirements (e.g., as pesticides, fungicides, and soil conditioners) highlight the importance of new management strategies that reduce S inputs and losses.Emerging tools such as S stable isotope biogeochemistry permit tracing S pathways and biogeochemical cycling across the landscape (Hermes et al., 2021(Hermes et al., , 2022)), thus helping to further refine management plans.The process of creating S management plans begins with better reporting of global agricultural uses of S in terms of forms, amount, and timing.It continues with consistent monitoring, research, and agronomic trials to guide where and when S is needed, especially in those countries and regions where data are currently scarce, and that involves farmers in educational programs on best management strategies.The effects of elevated S in the environment, including the impacts on production of MeHg, necessitate that S be a standard component of soil monitoring and water quality plans.
We call for the investigation of current gaps in the biogeochemical cycling of S to better inform biogeochemical models and decisions related to the rate, type, and timing of S applications across a variety of soil types, textures, landscape positions, and organic matter compositions (Figure 2).Specifically, research priorities include addressing: What factors control soil S mineralization rates, and how do they vary across heterogenous landscapes?How much S is currently stored in soils across various soil characteristics and climatic zones, and what is the rate of its release over time?What are country-level and global-scale S loads from agricultural activity?How and in what forms is S mobilized across different lands uses, and what are the fluxes to surface water and groundwater?How can remote sensing, hyperspectral imagery, and other emerging technologies act as early sentinels of S deficiency in plants, and how can we apply them in real-time to guide the quantity and timing of S additions?What combinations of AMF and crops are most effective, and are there additional co-benefits or unanticipated negative consequences of these inoculations?Can bioremediation and chemical remediation approaches reduce the harms of S in the environment?Specific questions also remain regarding the effects of S used in agricultural activity, including: Is sulfate from agriculture entering nearby wetlands and anoxic sites that promote hydrogen sulfide and MeHg production?Does S fertilizer use in croplands have consequences that mimic those of historic acid deposition, including base cation and Al mobilization from soils into nearby surface waters?Is there a potential for bioremediation of S-contaminated sites?
Looking toward the future of S use in agriculture, it is important to consider how S biogeochemical processes will be altered by climate change globally.Most of the anticipated climatic changes will likely lead to decreased efficacy and increased environmental consequences associated with S additions.For example, we anticipate that warming associated with climate change will increase soil S mineralization rates since soil temperature is generally the master variable in this process (Germida & Janzen, 1993;Nor & Tabatabai, 1977;Schoenau & Malhi, 2008;Watkinson & Lee, 1994), thereby leading to increased sulfate availability, as well as the potential for leaching from soils.Higher temperatures additionally increase sulfate desorption from mineral and organic soils (Zak et al., 2021), further enhancing leaching losses.Increased rates of evapotranspiration will place additional water stress on plants; this may increase plant irrigation needs, further exacerbating leaching losses from soils, while requiring additional S inputs to regulate stomatal opening and reduce plant water stress.Higher air temperatures, alongside more frequent wetting-drying cycles, can also stimulate Fe sulfide oxidation in soils and sulfide production in shallow waters (Lamers et al., 2013;Smolders et al., 2006), increasing the likelihood of sulfide toxicity and the production of toxic MeHg.Climate change is likewise expected to increase pressures from pests and pathogens by changing geographic distributions and altering the timing of their growth stages (Caffarra et al., 2012;IPCC, 2022;Skendžić et al., 2021;Tang et al., 2017).This shift may lead to additional S use as a fungicide and pesticide, although it should be noted that some pathogens (e.g., powdery mildew in Italian vineyards, see Caffarra et al., 2012) are predicted to decrease in severity.These climate change-related impacts on S biogeochemical cycling are all occurring within systems that will be receiving 70%-90% less atmospheric S Figure 2. Interventions and research needs to reduce the consequences of sulfur (S) from agricultural activity.Sulfur management plans can be targeted at various points on the landscape to reduce S inputs, losses, and environmental consequences.These include precision agriculture that targets S inputs, inoculation with arbuscular mycorrhizal fungi (AMF), genetic modifications of crops, additions of organic matter, total maximum daily load (TMDL) targets, bioremediation, and chemical remediation.However, there are many gaps in our understanding of S fluxes and pools across agricultural landscapes, including S inputs via acid deposition and agricultural activity (S in ), S uptake by crops (S uptake ), leaching of S into soils (S leach ), storage of S in soils (S pool ), release of S from soil pools (S release ), overland flow of S in surface runoff (S overland ), sequestration of S in precipitates (FeS), and release of S from aquatic environments (H 2 S).deposition to agricultural soils (Feinberg et al., 2021).Sulfur management must therefore consider strategies that can effectively meet crop S demands within the context of the changing climate and socioeconomic conditions.

Conclusion
It is time to create sustainable S management plans that meet crop health, yield, and quality needs, while also reducing the ecological harms caused by excess S in the environment (Haneklaus et al., 2008;Schnug, 1991).Achieving this goal will require partnerships amongst researchers, agricultural extension agents, farmers, and regulators, along with the incorporation of S into soil and water monitoring programs.It will also require an investment in research in data-scarce countries.As we work toward these solutions, we can apply lessons already learned from research on the efficacy and adoption of N and P management schemes at local, regional, and national levels and can develop integrated management plans that collectively consider a suite of nutrients together.No one solution will work across both space and time, especially given the wide range of uses in which S is applied in agriculture; rather, a portfolio of management tools is necessary.While further research is needed to understand S fate, transport, and transformations across agricultural landscapes and into nearby aquatic environments, we have the preliminary knowledge necessary to begin S management now.

Figure 1 .
Figure 1.Sulfur use in agricultural areas.(a) Application of sulfur via spraying in vineyards in Napa Valley, California.(b) Powdery mildew on grapes (Vitis) and (c) grape leaves in Napa Valley, California.Photo credit: E.S. Hinckley (a) and M. Cooper (b, c).Photos included with permission.
Figure2.Interventions and research needs to reduce the consequences of sulfur (S) from agricultural activity.Sulfur management plans can be targeted at various points on the landscape to reduce S inputs, losses, and environmental consequences.These include precision agriculture that targets S inputs, inoculation with arbuscular mycorrhizal fungi (AMF), genetic modifications of crops, additions of organic matter, total maximum daily load (TMDL) targets, bioremediation, and chemical remediation.However, there are many gaps in our understanding of S fluxes and pools across agricultural landscapes, including S inputs via acid deposition and agricultural activity (S in ), S uptake by crops (S uptake ), leaching of S into soils (S leach ), storage of S in soils (S pool ), release of S from soil pools (S release ), overland flow of S in surface runoff (S overland ), sequestration of S in precipitates (FeS), and release of S from aquatic environments (H 2 S). Figure created in BioRender.