Effects of climate change on river and groundwater nutrient inputs to the coastal ocean

Abstract

Rivers and groundwater are major sources of nutrients to the global coastal ocean. Climate change is expected to impact nutrient fluxes from river basins and coastal aquifers through alterations to both hydrological and nutrient cycling processes. In this Review, we identify and summarize how climate change impacts, such as changes in precipitation, increased cryosphere melt, and sea level rise, will affect water discharge and nutrient concentrations in rivers and coastal groundwater, which ultimately control nutrient inputs to the coastal ocean. We also document key limitations in the current understanding of climate-related changes to nutrient fluxes, especially in coastal groundwater basins. The impacts of climate change will interact with local human impacts, highlighting the need for studies spanning local to global scales to better understand and improve predictions of future nutrient fluxes from these hydrological pathways.

Nutrient fluxes from rivers and groundwater flowing into the ocean are impacted by climate change impacts such as precipitation changes, cryosphere melt, and sea level rise.

Introduction

Rivers and groundwater discharging into the ocean account for a large fraction of global coastal nutrient budgets. These land-sea fluxes of nutrients (carbon, nitrogen, phosphorus, silicon, and other micronutrients) sustain coastal primary productivity and can affect carbon (C) sequestration, air-sea gas exchange, oxygen balance, and acidification¹. For instance, rivers add 40-70 Tg of nitrogen (N) per year to coastal ecosystems, about two-thirds of the total N input to the near shore area^{2, 3–4}, and roughly 43 Tg of the 430 Tg of C transported by rivers annually ends up buried in coastal marine sediments⁵. Coastal groundwater nutrient fluxes can exceed riverine contributions for dissolved silicon (DSi, Si(OH)₄)^6,7 and some forms of N and phosphorous (P)^8,9 (Table 1), mainly owing to the large volume of seawater recirculating through coastal aquifers and interacting with aquifer materials to liberate nutrients. Despite the relatively small global fresh groundwater flux to the coastal zone, nutrient inputs from fresh coastal groundwater can be disproportionately important for some nearshore ecosystems, such as coral reefs, estuaries, and salt marshes¹⁰.

Climate change is rapidly transforming environmental systems and processes around the world^11,12. Marine and terrestrial impacts of climate change, commonly referred to as climatic impact-drivers (CIDs)¹³, are expected to alter water and chemical balances in rivers and groundwater, both at the coast and farther inland^{14, 15–16}. These climate change impacts will affect nutrient exports from rivers and groundwater to the coastal zone, with consequences for nearshore ecosystems and biogeochemical processes.

Local anthropogenic activities (fertilizer use, mining, land use changes, wastewater inputs, etc.) will interact with the impacts of climate change, further affecting nutrient input from hydrological pathways. For instance, anthropogenic N sources in rivers are expected to increase and account for up to 80% of total riverine N inputs to the ocean by 2050 under most climate change scenarios, largely due to increases in agricultural use of fertilizers needed to keep pace with demand for food production¹⁷. In another example, groundwater extraction can cause seawater intrusion and salinization of coastal aquifers, a process exacerbated, in some regions, by climate change induced sea level rise (SLR) and increasing aridity¹⁸. Nearly 80% of coastal areas below 60°N are expected to experience seawater intrusion by 2100 from the combined impacts of recharge changes and SLR¹⁹. This salinization will shift coastal groundwater nutrient chemistry, affecting fluxes to the ocean^15,20.

This review summarizes current knowledge on how climate change is expected to modify nutrient inputs from rivers and coastal groundwater from a global perspective. First, we provide a brief overview of the processes controlling global nutrient fluxes from rivers and coastal groundwater. Then, we assess how climate-induced changes in water and nutrient balances will affect nutrient delivery to the coastal ocean. The input of nutrients from land to sea is a function of two parameters: (1) the flux of water from rivers and groundwater into coastal areas, and (2) the concentrations of nutrients in this water. Both parameters are impacted by climate and other anthropogenic changes. We first discuss how changes in water balance due to climate change could alter hydrologic nutrient inputs to the coast and then consider how climate change might influence biogeochemical processes affecting nutrient concentration and speciation in rivers and coastal groundwater.

We focus on CIDs previously identified with medium to high confidence in their likelihood of affecting at least 30% of global regions^15,21. These CIDs include, but are not limited to: increasing air and sea surface temperatures, increasing intensity of tropical cyclones, increases in heavy precipitation events and flooding, changes in mean annual rainfall, changing aridity and droughts, changes in mean wind speed, thawing of permafrost, accelerated melting of snow, glacier, and ice sheets, SLR, enhanced coastal erosion, increased fire weather days, as well as changes in ocean chemistry, such as shifting salinity, ocean acidification, and deoxygenation²¹ (Fig. 1, Supplementary Fig. 1). These CIDs are expected to alter both dissolved and particulate nutrient fluxes from rivers and coastal groundwater to the marine environment. Here, we broadly consider dissolved inorganic (ammonium, nitrate/nitrite, phosphate, and carbon) and organic (carbon, nitrogen, and phosphorus) forms of nutrients as well as particulate organic carbon and nitrogen. Because many CIDs are understudied, we highlight key examples for which data exist and are expected to drive shifts in nutrient fluxes under changing climate conditions (Supplementary Table 1).

Fig. 1 Overview of climate change impacts. [Images not available. See PDF.]

CIDs with medium to high likelihood of affecting at least 30% of global regions²¹ and expected to affect rivers and coastal aquifers. The up and down arrows indicate increase and decrease, respectively. The magnitude and direction of impact of many of these CIDs vary spatially. These impacts will also generally co-occur, compound/mitigate each other, and interact with other local anthropogenic stressors, making changes and associated impacts often site-specific. Supplementary Fig. 1 presents an overview of projected changes in CIDs grouped by continent and/or region.

Table 1. Global estimates of water and nutrient fluxes from rivers and coastal groundwater

			Fresh Water	Fresh and Saline Water	TP	DIP	DOP	PP	POP	Fe-P	TN	DIN	DON	PN	TC	TOC	DOC	DIC	POC	PIC	Si
Rivers
265GRDC	2021		39,693	na
266Fekete et al.	1999		39,476	na
267Meybeck	1979																				217
268Meybeck	1982				9													380
269Meybeck	1988	natural				0.4						4.64	6.8				220		180
270Meybeck	2003																				163
271Meybeck & Vörösmarty	1999																	385
272Treguer et al.	1995	diss.																			168 ± 18
7Treguer et al.	2021	diss. & amorph.																			227 ± 56
273Ludwig et al.	1996															380
274Hedges et al.	1997																250		150
275Cauwet	2002																176
276Green et al.	2004	exorheic									38.6
277Boyer et al.	2006	exorheic									48
2Seitzinger et al.	2005	exorheic			11						66					367
261Seitzinger et al.	2010	for 1970			7.6	1.1	0.6	5.9			36.7	14.0	10.3	12.4			161		127		141
261Seitzinger et al.	2010	for 2000			8.6	1.4	0.6	6.6			43.2	18.9	10.8	13.5			164		140		144
261Seitzinger et al.	2010	for 2050 GO			8.5	2.3	0.6	5.6			47.5	24.4	11.5	11.6			160		120		136
261Seitzinger et al.	2010	for 2050 AM			8.6	2.0	0.6	6.0			42.0	18.5	11.1	12.4			160		129		138
278Mayorga et al.	2010	exorheic			8.6	1.4		6.6			43.2	18.9		13.5		303	163		140
278Mayorga et al.	2010	total			9.0						44.9					317
279Beusen et al.	2009	modern																			178 ± 21
280Dürr et al.	2011	modern																			173
30Galy et al.	2015	biosph.& petrog.																	200 ± 135/75
281Li et al.	2017																240	410	240	170
282Fabre et al.	2019																131.6
283Lacroix et al.	2020	nat./bioav.			3.7	0.5	0.1		2.2	0.8		27			603	237	134	366	103		158
284Tian et al.	2023															416	262 ± 33	453 ± 94	154 ± 25
285Liu et al.	2024														1020 ± 220		300 ± 140	520 ± 170	180 ± 40	30 ± 2
Coastal groundwater
10Luijendijk et al.	2020		224
24Zhou et al.	2019		489 ± 337
25Cho et al.	2016			44,000 ± 12,000
9Cho et al.	2018					1.8 ± 0.6						32 ± 8									107 ± 28
6Rahman et al.	2019																				103.9
29Zhang & Planavsky	2020																	89-996
286Santos et al.	2021					1.86						32.2
7Treguer et al.	2021																				65 ± 31
8Wilson et al.	2024	mean (range)				5.6 (2.2-15.5)						75.6 (28-210)	36.4 (14-92.4)

TP total phosphorus, DIP dissolved inorganic phosphorus, DOP dissolved organic phosphorus, PP particulate phosphorus, POP particulate organic phosphorus, Fe-P Fe-bound phosphorus, TN total nitrogen, DIN dissolved inorganic nitrogen, DON dissolved organic nitrogen, PN particulate nitrogen, TC total carbon, TOC total organic carbon, DOC dissolved organic carbon, DIC dissolved inorganic carbon (Zhang & Planavsky, 2020, reflects HCO₃⁻), POC particulate organic carbon, Si dissolved silicon (Beusen et al., 2009, reflect average and uncertainy of range 158–199 Tg Si yr⁻¹). Treguer et al. (2021) include amorphous Si. If fluxes are based on global average concentrations in the original publication, global fluxes are calculated or recalculated assuming a global river water flux of 40,000 km³ yr⁻¹ (e.g., Meybeck, 1979, 1988, 2003; Treguer et al., 1995). Lacroix et al. (2020) model results are based on a global river water flux of 23,496 km³ yr⁻¹. Seitzinger et al. (2010): GO - Global orchestration, AM - Adapting mosaic model, 1970 - Model reconstruction.

 All water fluxes are in km³ yr^-1, and all nutrient fluxes are in Tg (C, N, P, Si) yr^-1.

Current global hydrologic nutrient flux estimates

Annually, rivers deliver substantial amounts of C, N, P, and Si in their various forms to the coastal ocean (Table 1), though these estimates exclude modifications in estuaries, which can significantly alter nutrient composition^22,23. Submarine groundwater discharge (SGD), which is a mix of freshwater and recirculated seawater, contributes additional nutrients. While the fresh component of SGD is relatively small compared to rivers, generally accepted to be around 0.6 to 1% of river discharge^10,24, the saline component is considerable, and the total global discharge (fresh and saline) is similar to global river discharge estimates²⁵.

Due to the magnitude of water fluxes from saline SGD and the reactivity of nutrients in the subterranean estuary, global nutrient fluxes from total SGD often exceed those from rivers (Table 1). For example, dissolved inorganic nitrogen (DIN) and phosphorus (DIP) from total SGD are estimated to be 1.4 to 1.6 times greater than river inputs⁹, with more recent studies suggesting up to 3 to 4 times greater fluxes for DIN, DIP, and dissolved organic nitrogen (DON)⁸. Dissolved Si fluxes from total SGD are estimated to be up to half of riverine fluxes^6,7. While global estimates of dissolved inorganic carbon (DIC) and dissolved organic carbon (DOC) fluxes from total SGD are unavailable, local studies indicate that both DIC and DOC fluxes from SGD can be comparable to or exceed river contributions^{26, 27–28}. Estimates of mean global bicarbonate fluxes (the main component of DIC at circumneutral pH) from fresh coastal groundwater range from 7.4 to 28 Tmol yr⁻¹, potentially exceeding riverine HCO₃⁻ fluxes by up to ~240%²⁹.

Primary nutrient sources and processes

There are several sources and processes that determine the concentration of macronutrients (C, N, P, and Si) delivered by rivers and groundwater to the nearshore environment. Weathering of bedrock contributes riverine C^30,31, N³², P³³, and Si⁷, while atmospheric sources strongly influence C^31,34 and especially N^{35, 36–37}. The natural cycles of the major nutrients have been significantly disturbed by humans, particularly through the industrial-scale utilization of the Haber-Bosch process to fix atmospheric N^38,39, large-scale mining of P deposits for fertilizer applications⁴⁰, river damming modifying the natural cycle of Si through changes in sediment discharge and deposition^41,42, and land use changes⁴³.

Nutrients in coastal groundwater originate from both infiltrating fresh and salt water into the coastal aquifer as well as the biogeochemically active subterranean estuary⁴⁴. Diverse land and sea forcings, such as tides, waves, inland recharge, flow path geometry, and aquifer properties control the relative proportions and fluxes of coastal groundwater^45,46. Many of these properties are regulated by local lithology, which varies widely along the global coastline (Box 1).

Human activities, such as fertilizer use and wastewater leakage, have modified nutrient cycling in groundwater^{47, 48, 49–50}, leading to higher nutrient concentrations in coastal groundwater than what would be found naturally in many regions⁵¹. A synthesis of data from over 200 sites shows higher DIN and DON concentrations in low salinity compared to saline coastal groundwater, with land use being an important predictor⁸. Coastal groundwater also supplies dissolved Si to the nearshore ocean, originating from both lithogenic sources via water-rock interactions and biogenic sources from silicifying organisms^6,52.

Terrestrial DOC contributions to groundwater can occur via surface inputs or subsurface leaching of sediment or soil organic matter^{53, 54–55}. Concentrations and properties of surface dissolved organic matter, sorption capacity in the unsaturated zone, and precipitation patterns influence the hydrological connectivity of DOC between surface water and groundwater, with up to 95% of surface-derived DOC removed before reaching the water table, likely due to adsorption^{56, 57–58}. In contrast, DOC in saline coastal groundwater often originates from seawater and is typically highly labile⁵⁹. In some coastal groundwater systems, marine DOC consumption is substantial and drives high DIC export⁶⁰.

Both subterranean and river estuaries further regulate nutrient fluxes from the land to the sea⁶¹. The subterranean estuary is a biogeochemically reactive zone due to dynamic physical and chemical conditions, which exerts a strong control on nutrient concentrations, forms, and fluxes^8,44. Biogeochemical transformations within subterranean and river estuaries modify nutrient concentrations, with physical and chemical conditions such as residence time, redox, pH, and salinity exerting strong control over nutrient dynamics^51,62. Because the N cycle is microbially mediated, DIN changes within coastal aquifers commonly arise from microbial processes (denitrification, anammox, coupled nitrification/denitrification, or Mn oxidation of ammonium), which can attenuate DIN by up to 65%⁶³. In contrast, dissolved inorganic phosphorus (DIP) is often controlled by physicochemical interactions with sediment. Under oxic/anoxic conditions, dissolved phosphate is readily sorbed/desorbed to/from iron- and aluminum-oxides and/or co-precipitates with other dissolved metals, though dissolved phosphate can be mobilized in coastal groundwater if the sorption capacity of the aquifer is overwhelmed^51,64,65.

Box 1 Global variability in coastal lithology affects groundwater vulnerability to climate change

Lithology is an important control on groundwater flow and solute transport. The global distribution of coastal lithology is highly variable. Dominant lithologies are generally of sedimentary origin. Even within similar lithologic groups, aquifer properties can be highly variable. For instance, intrinsic permeability, a key aquifer parameter that describes the ease of groundwater flow through a material, varies by several orders of magnitude across and within different lithologies²⁸⁷. Unconsolidated sediments, for example, compositionally range from highly permeable coarse sands and gravels to low-permeability silts and clays that restrict groundwater movement.

In coastal aquifers, these land-based lithologic units laterally abut shoreline features, such as cliffs, beaches, mangroves, and muddy coasts, as well as seaward sediments, which further regulate groundwater and solute fluxes across the land-sea interface²⁸⁸. Lateral permeability gradients can be further complicated by vertical changes in subsurface geology.

Together, these landward, shoreline, and seaward lithologic properties will help shape coastal groundwater responses to climate change. High-permeability aquifers might respond more rapidly to CIDs, such as extreme precipitation and coastal flooding, while low-permeability systems are likely to have longer residence times and therefore are, potentially, more influenced by biogeochemical changes within the aquifer. Lithologic differences can also influence solute composition in groundwater²⁸⁹, which can be important in shaping subterranean estuary biogeochemistry and vulnerability to CIDs. For instance, carbonate aquifers are likely more susceptible to increasing ocean acidity²⁹⁰. Such factors will influence how CIDs affect nutrient biogeochemistry. This heterogeneity may also cause nutrient responses to CIDs to deviate from broader regional or global trends in some coastal aquifers.

Global distribution of coastal lithology types. Classifications along the coast were derived from the Global Lithological Map (GLiM)using the Eckert IV equal-area projection²⁹¹. Lithologic categories were grouped into: siliciclastic sedimentary rocks (ss, sm), unconsolidated sediments (su), carbonate and evaporite sedimentary rocks (sc, ev), extrusive igneous rocks (va, vb, vi, and py), intrusive igneous rocks (pa, pb, pi), metamorphic rocks (mt), and water and ice (ig, wb).

Impacts of climate change on the water balance in rivers and coastal aquifers

Climate change will alter key components of the water cycle (precipitation, evaporation, recharge, runoff) and, as a result, water fluxes entering and leaving river basins and coastal aquifers^{15,66, 67, 68, 69, 70–71}. In this section, we discuss how water balances could shift with climate change, and how these changes directly affect water fluxes from rivers and coastal groundwater, thereby altering nutrient input to the coastal ocean. The associated changes in nutrient biogeochemistry will be discussed in “Impacts of climate change on nutrient chemistry in rivers and coastal aquifers”.

Alteration of river nutrient fluxes due to climate-driven water cycle changes

Increasing air temperatures will increase atmospheric water vapor pressure by ~7% per °C⁷² and cause an intensification of the global hydrologic cycle. Changes to precipitation in a warming world will directly and indirectly alter river discharge and these changes will vary spatially. Global model projections suggest that by 2050 average annual discharge will increase over 47% and decrease over 36% of the global land surface⁷³. Global climate models further suggest that nearly 90% of the land area will experience greater than 10% alterations in seasonal flow regimes, and maximum mean monthly river discharge timing is expected to shift by at least one month for more than one third of the global land area⁷⁴. Pronounced increases in discharge are expected in the high northern latitudes, East Asia, South Asia and East Africa, and marked decreases in Southern Africa, South America, and Central and Western Europe⁷⁵. Such changes alter river flow regimes and could even force perennially flowing rivers to become intermittent^74,76. These shifts in precipitation patterns will alter river discharge, directly affecting water and nutrient fluxes to the coastal zone, even in the absence of changes to nutrient concentrations (Box 2).

Climate-induced low-flow conditions might also drive decreases in particulate transport, as has been observed for suspended sediment in the Itata River basin of central Chile⁷⁷. While drought-induced low flow periods can increase concentrations of dissolved nutrients from reductions in dilution capacities¹⁴, overall nutrient fluxes would likely decrease in rivers with lower discharge. In some areas, rain events following drought periods triggers flushing events that drive substantial increases in concentrations of nutrients, for instance, dissolved P^78,79.

Changes in precipitation can also alter patterns of river flooding, increasing flood events in basins, such as the Amazon, Magdalena, Brahmaputra and Niger river basins, while suppressing floods in others, like the Mekong, Mississippi, and Rhine⁸⁰. Flooding can drive large increases in riverine dissolved nutrient fluxes; for example, nitrate and nitrite fluxes were over 100 times higher than average in parts of French Polynesia due to flood-induced nutrient pulses⁸¹. More broadly, heavy precipitation and flooding have been associated with mixed riverine nutrient responses. Particulate nutrient concentrations often increase owing to increased erosion, transport from hillslopes, and re-suspension of in-stream or flooded bank materials¹⁴, while dilution during high flows typically lowers dissolved nutrient concentrations but increases total inputs due to the high water fluxes⁸².

Rivers also interact with groundwater in river basins, with groundwater sustaining river baseflow. Therefore, changes to groundwater balances due to climate change will also indirectly affect river discharge⁸³. For instance, increased forest water use in a warmer climate reduces groundwater recharge and storage, causing annual streamflow losses and intermittent flow conditions during low precipitation periods⁸⁴. In some rivers, the effects of rising aridity are already observable. The Colorado River basin has seen multi-decade flow declines of 13% due to increased evapotranspiration from warming, with additional projected declines of 9.3% per °C of additional warming⁸⁵. Increases in soil temperatures may even outpace increases in air temperature⁷⁸. Rivers also directly face rising temperatures, with current warming averaging 0.05 °C per year between 30 °S and 30 °N⁸⁶, which could enhance evaporation from inland waters and potentially reduce water fluxes and associated nutrient inputs. These climate-induced changes in river water fluxes will result in changes in nutrient fluxes to the coast, even in the unlikely scenario of no changes in nutrient concentrations. Increasing aridity and drought in some systems also drives higher wildfire activity, which can increase particulate nutrient exports from watersheds⁸⁷. In some systems, fire also alters soil hydrology and increases runoff⁸⁸, leading to higher water fluxes that may further enhance riverine nutrient export.

In high-latitude systems, cryosphere thaw and melt as well as climate change-induced transitions from snowfall to rainfall could substantially alter water balances as previously frozen areas become active sources and/or pathways for water movement and storage^{89, 90, 91–92}. Glacial melting has already resulted in a significant reduction of river flows in many glacial-fed streams, owing to earlier and lower spring freshets⁹³. However, ice mass loss in some regions of Canada has increased groundwater storage and river baseflow⁹⁴. While warming-induced melting can initially drive increased runoff, this threshold shifts towards water limitation as glacier recession ultimately reduces meltwater volumes⁹⁵. Other areas, such as high-mountain regions of Asia, which feed the Ganges, Brahmaputra, Indus, Mekong, Yangtze, and Yellow rivers, are more susceptible to snowmelt compared to glacial mass loss, as snowmelt volume is generally a larger contributor to these downslope basins⁹⁶. In these areas, warming is expected to result in significant reductions in future snow meltwater supply, a substantial shortening of the snowmelt season, and changes in the timing of melt observed in rivers and associated fluxes to the coastal ocean⁹⁶. Increased snowmelt and rain-on-snow events can also change the timing and transport of nutrients in snow-covered regions. For instance, 50% of soil N and P pools in the US may end up in rivers and groundwater due to rain-on-snow events⁹⁷.

In the Arctic, permafrost thaw is expected to result in a transition from surface water to groundwater domination⁹⁸. Model simulations indicate that annual total runoff from Arctic rivers could increase by up to 25% due to warming, largely owing to increases in water runoff from subsurface pathways⁹⁹. Such a shift in hydrology could be very impactful to the Arctic Ocean that holds ~1% of the global ocean water but receives 10% of the world’s riverine water discharge⁹⁸. These changes highlight the need for understanding hydrologic changes in polar regions and beyond, particularly in areas where the hydrological cycle was muted due to frozen conditions. Reduction or amplification of surface water fluxes will change the magnitude of river nutrient fluxes. Likewise, shifts in flow paths, from surface to subsurface domination, will also impact nutrient concentrations and speciation. Both changes will contribute to alterations in the delivery of river nutrient fluxes to the coastal ocean.

SLR and coastal flooding will drive seawater intrusion in many coastal systems. Because of the direct connection between groundwater, surface water, and the ocean, seawater intrusion will likely increase the salinity of rivers in many coastal regions¹⁰⁰. For instance, recent estimates suggest that SLR will cause river salinity to increase by nearly 1 ppt as far as 80 km upstream from a river mouth in Bangladesh¹⁰¹. Salinization could be particularly pronounced in coastal watersheds and basins that support ecosystems at their terminus with the sea, such as wetlands, estuaries, or coastal lagoons. While these coastal stressors have less direct effect on the magnitude of river water exported, salinization can persist for long timescales and affect river and estuarine nutrient chemistry abiotically and biotically (see “Alterations in river nutrient fluxes due to climate-driven nutrient chemistry changes“—“Alterations in coastal groundwater nutrient fluxes due to climate-driven nutrient chemistry changes”).

Increased tropical cyclone intensity can drive extreme precipitation in coastal watersheds, which can episodically increase nutrient fluxes to the coastal ocean. In Fiji, tropical cyclone events deliver nearly six times more sediment to the coastal ocean than other precipitation events¹⁰². Heavy precipitation common to cyclones can also enhance recharge¹⁰³ and induce fluvial flooding¹⁰⁴. In parts of the southeast coast of the US, tropical cyclones account for over half of the region’s record floods and nearly one third of the top ten annual flood peaks in rivers¹⁰⁵. An increase in cyclone intensity could therefore result in increases in the magnitude of water flux and episodic nutrient pulses to the coastal zone¹⁰⁶. Increased tropical cyclone intensity will also bring other compounding stressors, such as waves, strong winds, and storm surges, that further contribute to coastal flooding and river discharge¹⁰⁷.

Coastal erosion, which is expected to increase along nearly every global coastline, could reduce the spatial extent of watersheds as a result of shoreline retreat, thereby reducing catchment size. This would reduce drainage area and thus water fluxes. Erosion is particularly impactful for coastal ecosystems, such as marshes, which are eroding at rates up to ~3 m/yr in some areas of the southern US¹⁰⁸. The loss of coastal ecosystems could drive nutrient flux increases as these ecosystems are highly productive environments that can buffer upstream nutrient pollution¹⁰⁹.

The impacts of other CIDs on river water dynamics, such as changing ocean chemistry (salinity, acidity, dissolved oxygen) and increasing sea surface temperatures, are less clear. Changes in ocean chemistry can propagate to rivers directly via surface water mixing and indirectly via groundwater, potentially altering river discharge patterns, estuarine exchange, and the balance between fresh and saline water flows at the land-ocean boundary.

Box 2 Hydrologic controls on dissolved nutrient transport: the role of discharge in a changing climate

The relationship between riverine constituent concentrations (C) and discharge (Q) is generally described by a power-law for straight-line segments in log (C)-log (Q) diagrams²⁹²: C = aQ^b, though more complex formulations have been proposed²⁹³. At very low and very high Q, C-Q relationships approach constant C due to solubility and dilution limits, respectively. The exponent “b” quantifies the sensitivity of changes in C for a given change in Q. Three end-member relations include minimal change in C with changing Q (b~0, “chemostatic”), decrease in C with an increase in Q (b < 0, “pure dilution”), and increase in C with increasing Q (b > 0, “enrichment”). This simple classification was expanded by adding six endmembers that emerge by investigating the full range of discharge, particularly in smaller watersheds that are more strongly affected by hydrologic events²⁹⁴. Emerging C-Q relationships are often segmented at a critical Q value (often near the median Q) that separates regimes with different b values (regimes 1A to 3B, where chemodymamic regimes are shown in red and blue, and chemostatic relationships in green).

In order to show the complexity of C-Q relationships for nutrients, we compare C-Q regimes for Q values that range from small streams (Erlenbach, Switzerland) to large rivers (e.g. Amazon, Brazil). While C-Q regimes for large rivers are often stable, nutrients exported from smaller watersheds that are more strongly affected by hydrologic events can transition between C-Q endmembers²⁹⁵. Others have also shown that the exponent “b” for dissolved nutrients can change with changing climate and hydrological conditions¹⁵². These complexities challenge predictions of future riverine nutrient exports to the global ocean under changing climate conditions.

An analysis of 506 relatively pristine rivers in the United States revealed that Q_mean primarily controls many C_solutes, including nutrients, with reduced C in more humid locations and elevated C in arid climates because such regions are biogeochemical reactors (via in situ production and weathering reactions) that accumulate rather than transport solutes¹⁵³. A warmer climate that reduces Q is therefore expected to result in a tradeoff with higher C, but lower dissolved nutrient fluxes that are reduced at a magnitude related to the decrease in Q. In areas that become wetter, fluxes may increase due to higher Q despite lower C_nutrient. Organic-bound dissolved nutrients are particularly sensitive to changes in Q_mean with b near -1, whereas dissolved Si is much less sensitive (b -0.2), with a doubling of C possible when Q_mean decreases by half.

Diverse concentration-discharge patterns across global rivers.a. An example of classification schemes for C-Q relationships.b. Examples of various river C-Q relationships. References: [1] Amazon River, Brazil²⁹⁶, [2] Erlenbach, Switzerland²⁹⁵, [3] Jinsha/Yalong rivers, China¹⁵², [4] various rivers in France²⁹⁷, [5] various rivers in USA¹⁵³, [6] Mississippi River near St. Francisville, US²⁹⁸, [7] Fraser River, Canada²⁹⁹, and [8] Magdalena River, Colombia; unpublished data.

Alteration of coastal groundwater nutrient fluxes due to climate-driven water cycle changes

Increasing air temperatures will raise shallow coastal groundwater temperatures and change groundwater thermal regimes^110,111. Experiments demonstrate that decreasing contrasts between groundwater temperature and ocean water temperature might reduce SGD and hence could affect associated nutrient fluxes¹¹². Similarly, changes in the global patterns of water vapor saturation and precipitation will impact groundwater fluxes, with rises in groundwater levels projected for areas experiencing precipitation intensification¹¹³. Regions projected to experience increases in mean annual precipitation include small islands in the western Pacific and Indian Ocean, which could see up to a 120% increase in recharge rates¹¹⁴. These characteristics can influence river (fluvial) and land (pluvial) flooding, which will have variable effects on groundwater recharge because the infiltration of floodwater is temporally and spatially variable, as has been observed in areas where floodwater is used to manage aquifer recharge¹¹⁵.

Conversely, aridification is expected to significantly reduce groundwater recharge¹¹⁶. In the continental US, increasing aridity will reduce shallow groundwater storage and lower water tables¹¹⁷, thereby likely lowering baseflow in rivers and discharge at the coasts in affected areas. For example, precipitation declines are expected to lower recharge rates by ~25 to 60% in coastal aquifers in Australia¹¹⁸, the Mediterranean¹¹⁹, Hawai’i¹²⁰, and over 30 other islands¹²¹. These changes in recharge and groundwater availability will directly affect the flux of groundwater to the coast and thus nutrient input. For instance, groundwater DOC concentrations are projected to decrease (increase) by ~10% (~3%) for every 10 mm increase (decrease) in precipitation during the driest (wettest) months of the year, likely due to a dilution affect¹²².

Cryosphere melt will alter water pathways and amounts in high-latitude polar systems. A shift to groundwater domination in the Arctic^98,123 could result in increases in water fluxes from coastal groundwater. The thaw of permafrost, which occurs along more than one third of the global coastline¹²⁴, can impact subsurface properties and activate new conduits for groundwater flow and storage^125,126. In these and similar areas, discharge of fresh groundwater could increase by up to two orders of magnitude by 2100 due to increases in top-down thawing¹²⁷. Ice sheet mass loss can contribute to declines in land-sea hydraulic gradients that might reduce coastal groundwater discharge due to reductions in overlying ice loading that pressurizes groundwater¹²⁸. Shifts in the timing and magnitude of snowmelt will likely affect coastal groundwater fluxes as well since snowmelt can be an important contributor to recharge in some coastal regions¹²⁹. In Japan, a decrease in snowfall and an increase in rainfall has increased contributions of fresh groundwater by 30–35%, resulting in increased DIC fluxes (23–40%) but reductions in nutrient fluxes (20-30%) to the coastal ocean¹³⁰. Nutrient inputs from SGD via glacial^89,131 and ice sheet meltwater¹³² could also rise because water fluxes are expected to increase in line with meltwater contributions. Likewise, previously frozen particulates could become mobilized with melting¹³³, increasing the flux of suspended nutrients to the coast.

SLR and increased coastal flooding will contribute to vertical and lateral seawater intrusion in many global coastal systems. SLR will directly impact fresh and saline water fluxes from coastal groundwater and, in some systems, could entirely stop offshore discharge, redirecting flow inland instead¹³⁴. Lateral seawater intrusion, driven by SLR, will push the freshwater-saltwater interface inland and raise groundwater levels due to lifting from below^134,135 (Box 3). Vertical seawater intrusion from the infiltration of saline flood water can occur due to the individual or combined impacts of SLR, wave overtopping, and storm surges^{136, 137–138}. As a result, fresh coastal groundwater fluxes would decline in response to relative SLR and other drivers of coastal flooding, which can reduce land-sea hydraulic gradients and directly salinize groundwater¹³⁴.

Coastal erosion could reduce the spatial extent of coastal aquifers as a result of shoreline retreat, thereby reducing coastal groundwater storage. Projections from strip island aquifers show that erosion and SLR together reduce the extent of freshwater lens volume by 22%¹³⁹. Coastal groundwater flow can also be sensitive to beach morphology¹⁴⁰, with coastal groundwater levels decreasing in erosion zones¹⁴¹. Both declines in water levels and storage could reduce water and thus nutrient fluxes from coastal groundwater.

The impacts of climate-induced changes in ocean chemistry (salinity, acidity, dissolved oxygen) and increasing sea surface temperatures on coastal groundwater fluxes are poorly documented. Changes to saline coastal groundwater temperatures from warming ocean temperatures might increase tidally driven coastal groundwater discharge, particularly in coastal temperate zones with large thermal contrasts between coastal groundwater and seawater¹¹². Increased groundwater temperatures related to warmer seawater can enhance permafrost thawing in high latitudes, increasing discharge¹⁴². Increasing acidity could facilitate increased groundwater flow in carbonate systems that are more susceptible to dissolution^143,144.

Increased tropical cyclone intensity will increase coastal flooding from high sea levels and waves and can drive seawater intrusion in coastal aquifers, with similar effects on water fluxes as those expected for SLR and other drivers of coastal flooding. Extreme precipitation during tropical cyclones might also lead to pulses of high recharge¹⁴⁵ and thus groundwater discharge at the coast.

The effect of wind speed changes, which show globally mixed responses to climate change, on coastal groundwater is generally understudied, though existing work hints at how wind can be important to groundwater flow in both offshore^146,147 and onshore coastal aquifers^148,149. Even in the Arctic, wind can drive sustained periods of coastal groundwater discharge and aquifer recharge¹⁵⁰. Therefore, the magnitude of nutrient fluxes could be affected by shifts in wind speed.

Box 3 Vulnerability of coastal aquifers to sea level rise

Seawater intrusion in coastal aquifers is primarily influenced by the freshwater boundary condition³⁰⁰, which defines two physical system types: recharge-limited and topography-limited coastal aquifers¹³⁴. Recharge-limited systems are less vulnerable to salinization from SLR because the water table can rise in response to increasing sea levels, as there is adequate space between the land surface and the water table. In contrast, topography-limited systems are particularly vulnerable to salinization because the water table is constrained by the elevation of the land surface, leaving little room for any rise in response to SLR. In topography-limited systems, this lifting effect could result in inland groundwater inundation and seawater intrusion on the order of hundreds of meters to several kilometers, forcing submarine groundwater and associated nutrient fluxes to discharge entirely onshore^134,300. In recharge-limited systems, seawater intrusion will result in inland migration of the freshwater-saltwater interface to a lesser extent¹³⁴, with SLR of 1.5 m projected to result in seawater intrusion of 50 m³⁰⁰. Globally, nearly 70% of the world’s coastal aquifers are thought to be topography-limited, highlighting the vulnerability of global coastal groundwater to SLR-driven salinization¹³⁴. Changes in salinity will also affect coastal groundwater nutrient biogeochemistry (see “Alterations in coastal groundwater nutrient fluxes due to climate-driven nutrient chemistry changes”) .

An example of how recharge-limited (a, c) and topography-limited (b, d) coastal aquifers respond differently to SLR, affecting the extent of salinization due to seawater intrusion. Figure based on and modified from Michael et al.¹³⁴.

Impacts of climate change on nutrient chemistry in rivers and coastal aquifers

In addition to altering water fluxes and the associated transport of nutrients, climate change will also affect nutrient biogeochemistry, and thus the quantity and speciation of nutrients in rivers and coastal groundwater. These changes will modify nutrient inputs to the ocean beyond the effects related to water flux alone.

Alterations in river nutrient fluxes due to climate-driven nutrient chemistry changes

Many watersheds will experience more heatwaves and droughts¹⁵¹ and changes to mean annual rainfall. In a case study, DIC concentrations have increased by one third while fluxes doubled as a high-elevation drainage basin of the Jinsha River has warmed over the last decades due to increases in geogenic and biogenic production of DIC¹⁵². Increases in water residence times in rivers have similarly been shown to increase the concentration of some nutrients^14,153, thereby likely altering inputs to the coast. Other work suggests that warming could prolong the growing season, promoting nutrient uptake by algae and aquatic plants, thus reducing concentrations and hence inputs¹⁵⁴. Changes in the residence time of riverine and estuarine waters due to shifts in discharge can also affect nutrient retention within the soils of catchments, uptake in the water column, and denitrification at the sediment-water interface^155,156.

Increases in precipitation might result in increased nutrient leaching from soils and less retention by soil and sediment at some locations and could therefore contribute to higher nutrient loading¹⁴. In anthropogenically-impacted basins, flooding can increase nutrient loading due to changes in abiotic (discussed above) and biotic processes, such as primary productivity and organic matter decomposition¹⁵⁷, as documented by increased discharge of nutrients from urban runoff in European coastal cities^158,159 and elsewhere¹⁶⁰.

Increasing land temperatures and aridity will directly result in increases in river temperatures and indirectly increase salinity due to evapoconcentration¹⁴. Warming can alter the metabolic balance of streams, with a 1 °C increase in stream temperature projected to make streams 24% more heterotrophic, which will likely affect in situ C concentrations¹⁶¹. Warming can also alter chemical weathering¹⁶² and biogeochemical processes in soils^{163, 164–165}, with consequences for nutrient delivery to rivers and underlying groundwater.

Increasing wildfire activity, another direct consequence of rising temperatures, can alter aerosol nutrient delivery¹⁶⁶, soil biogeochemistry¹⁶⁷, groundwater discharge¹⁶⁸, and river chemistry⁸⁷. For example, fires increase black carbon, a fire-generated form of organic carbon that is recalcitrant^169,170. Over one third of black carbon produced by wildfires ends up in the ocean via rivers¹⁷¹.

Changes in salinity, temperature, and oxygen content in rivers could change river ecology¹⁷² and thus biogeochemical processes, such as microbial function¹⁷³ and algal growth¹⁷⁴, affecting nutrient concentrations both in and near rivers. Lower oxygen levels due to warming are already observable in many rivers¹⁷⁵ and could increase redox sensitive chemical transformations of nutrients (e.g., denitrification). In addition, increasing atmospheric CO₂ might facilitate increased cyanobacterial growth rates^176,177, which will also alter nutrient dynamics in rivers with consequences to the flux of nutrients to the coastal ocean.

Cryosphere melting can impact riverine chemistry in a few ways (Box 4). For instance, permafrost thaw has mixed effects on stream water chemistry that depend on catchment characteristics and thaw processes. In some areas, thaw reduces DOC and increases solute (including some nutrients) concentrations in connected surface waters, owing to increases in travel time as active layers deepen, flow paths lengthen, and transport velocities slow, which facilitate increases in mineralization and DOC adsorption^{98,178, 179–180}. In other regions, DOC and DON export to proximal streams are expected to increase substantially as permafrost thaws^181,182. Dissolved Si and P are widely anticipated to increase as flow paths deepen and entrain more mineral soil contributions, while changes in DIN are mixed⁹⁸. Transitions in permafrost landscapes due to thawing can likewise induce substantial shifts in microbial communities, which will affect nutrient biogeochemistry¹⁸³. Warming is also expected to increase nutrient inputs from glacial meltwaters, which are generally depleted in DON, but represent a source of inorganic N, mostly in the form of nitrate, in the Eastern Canadian Arctic Archipelago¹⁸⁴.

Increases in coastal erosion and flooding are expected to increase along nearly every global coastline and will impact nutrient concentrations. Coastal erosion will provide new sources of dissolved and particulate matter to coastal waters¹⁸⁵. Coastal flooding and SLR are especially relevant to river mouths and estuaries, where seawater intrusion could increase salinity and thus alter biogeochemical processes and associated nutrient transformations^186,187. Salinization of surface waters can impact ionic strength and composition¹⁸⁸; such changes can induce particle flocculation¹⁸⁹ and alter nutrient availability¹⁹⁰. For instance, experimental salinization of wetland sediments resulted in the immediate release of macronutrients (e.g., ammonium) and micronutrients (e.g., dissolved iron), due to cation competition¹⁸⁸. Increases in salinity have been shown to have mixed effects on DOC dynamics in surface water systems. In some regions, salinization increases DOC concentrations due to enhanced rates of sulfate reduction, and in other areas, DOC concentrations are reduced as productivity decreases¹⁹¹.

Changing ocean chemistry could alter the reactions river-borne particulate matter undergoes as it moves through the salinity gradient into the coastal ocean, including inorganic and biologically mediated reactions that affect nutrient cycles^192,193. These dynamic transition zones are also subject to changes in wind speed that can affect nutrient profiles and mixing in rivers. Strong winds mix river plumes vertically or advect the plume downwind, impacting the footprint of river discharge and associated nutrients and sediment loads in the coastal ocean¹⁹⁴.

Box 4 Complex climate change feedbacks on nutrient and discharge dynamics in Arctic rivers

Several regional studies from rivers draining Arctic regions further illustrate the subtleties of climate change impacts on river discharge and nutrient fluxes. Long-term discharge from the six large Arctic rivers that make up about two-thirds of the Arctic Ocean watershed area has been increasing, particularly in the Russian Arctic, where those records go back to the late 1930s³⁰¹. Records for the past 20 years show a slight reversal of this longer-term pan-Arctic trend. However, the combined flux of nitrate from these six watersheds has decreased by one third over the past 20 years, much more drastically than the slight (<1 %) decrease in overall discharge³⁰². The decrease in N fluxes is particularly pronounced in basins that experience stable or slightly decreasing runoff (the Yenisey, Lena, Kolyma and Mackenzie rivers). Only in the Ob’ and Yukon rivers, where river discharge has increased, have nitrate fluxes increased over the last 20 years³⁰².

Complex feedbacks, some reinforcing, some antagonistic, exist between rivers and surrounding landscapes that experience, among other factors, thawing permafrost^98,303, changes in land surface characteristics³⁰⁴, hydrology, soil microbiology, and shrubification of the landscape; all of which might be responsible for the observed changes in river biogeochemistry. For instance, the deepening of the active layer and thermokarst development contribute to heightened hydrologic connectivity, subsequently resulting in higher nitrate and climate-active N₂O fluxes^{305, 306–307}. In contrast, shrubification and enhanced utilization of nutrients by vegetation and soil microbes³⁰⁸ would cause nitrate fluxes to decline. Increased microbial activity in warmer soils intensifies nitrogen cycling³⁰⁹. This could lead to enhanced N uptake and/or fixation³¹⁰, a compensatory process that might counterbalance flux increases caused by hydrologic factors^311,312. Climate scenarios projected for the end of the century also predict regional rises in dissolved Si entering the ocean, such as in the Arctic³¹³ due to increased terrestrial precipitation. Moreover, none of the large Arctic river records address processes in their estuaries³¹⁴, though some recent studies have been done in the Lena and Kolyma estuaries^315,316, which both suggest that processes in the estuary and the delta increase the nutrient flux to the ocean. These complex interactions are just a few examples of how climate change stressors could affect fluxes.

Long-term nitrate fluxes from major Arctic rivers.a. Cumulative annual nitrate fluxes over approximately twenty years from the six major Arctic rivers, andb. individual annual nitrate fluxes for each river. Figure and data adapted from Tank et al.³⁰¹.

Alterations in coastal groundwater nutrient fluxes due to climate-driven nutrient chemistry changes

Changes to coastal groundwater nutrient chemistry will arise from alterations to water balances in coastal aquifers, as discussed above, as well as shifts in groundwater biogeochemical processes. These alterations are shaped by diverse land-sea forcings discussed earlier as well as aquifer properties such as lithology. For instance, change in recharge could indirectly portend changes to coastal groundwater chemistry by altering the timing and extent of saturated conditions in the shallow subsurface, which could affect prevailing biogeochemical reactions that alter the influx and sources of nutrients in fresh coastal groundwater. In many systems, surface water-groundwater connectivity is an important conduit for C, N, and P, so changes in the timing and magnitude of precipitation and hence recharge will affect nutrient speciation, amount, mobility, and, thus, transport. For example, soil leaching of DIC, which will add C to groundwater, is greatest in areas with higher precipitation¹⁹⁵. Similarly, DIC concentrations in some groundwater basins have been increasing non-linearly in response to increasing atmospheric CO₂ concentrations¹⁹⁶. In other areas, rainfall-driven declines in recharge reduce fresh SGD and organic matter remineralization¹⁹⁷. Alterations to recharge could also impact the lithogenic production of dissolved Si in fresh coastal groundwater, as chemical erosion rates respond to precipitation and runoff¹⁹⁸.

Changes to the water balance from shifts in recharge might also impact coastal groundwater residence times. Residence time and the extent of mixing between water masses with distinct oxidation-reduction potentials are important regulators of subterranean estuary nutrient biogeochemistry⁵¹. For instance, short residence times can prevent or limit N transformations, such as heterotrophic or autotrophic denitrification, coupled nitrification: denitrification, anammox, or Mn oxidation of ammonium⁶³. The advective flow of seawater through coastal aquifers or beaches¹⁹⁹ can also accumulate increased levels of dissolved Si due to the dissolution of biogenic Si or minerals. The extent of enrichment depends on how long the seawater resides underground and the type of minerals involved^{200, 201–202}.

Increasing land and sea surface temperatures will impact coastal groundwater in many shallow aquifer systems¹¹⁰ (Box 5). For instance, higher temperatures are associated with higher metabolic rates and enhanced organic matter remineralization in soil²⁰³, which can increase N and P cycling and decrease O₂ concentrations²⁰⁴. Warming coastal groundwater also enhances subterranean estuary biogeochemical reactivity, with higher modeled nitrate removal efficiency and altered solute transport paths²⁰⁵. Changes to coastal groundwater temperature could similarly increase the geogenic production of new dissolved Si as lithogenic silica dissolution is temperature dependent^198,206.

An increase in atmospheric CO₂ can also alter nutrient dynamics. For instance, respiration by plant roots and microbes increases under higher atmospheric CO₂²⁰⁷. Indeed, increases in groundwater DIC concentrations and associated decrease in pH have been tied to higher biomass accumulation in soils and temperature-enhanced soil respiration^204,208. Similarly, heatwaves can drive significant increases, on average of 26%, in soil respiration²⁰⁹, probably affecting DIC leaching to underlying groundwater as well.

Increases in permafrost thaw as well as ice sheet, glacier, and snow melt might likewise drive changes in coastal groundwater chemistry. As regions of the Arctic are expected to shift to groundwater domination, nutrient contributions from SGD, already a measurable source of nitrate and dissolved Si²¹⁰, could increase. For example, permafrost thaw can expose organic matter-rich sediment and release new nutrients^180,211. Thawing of permafrost is expected to increase contributions from supra-permafrost groundwater²¹², which can be a substantial source of DOC and DON to Arctic coastal waters²¹³. These inputs of highly reactive DOC and DON during permafrost collapse might increase microbial reaction rates, potentially modifying nutrient speciation²¹⁴. Characterization of C and N stocks in permafrost across the Canadian Arctic supports that thawing could mobilize considerable amounts of labile DOC and ammonium²¹⁵. In other areas that experience ice sheet and glacial melting, it is unclear how these inputs could shift nutrient biogeochemistry of coastal groundwater. However, it is likely that meltwater will mobilize previously frozen constituents, such as particle-bound and dissolved nutrients²¹⁶, which could alter prevailing microbial processes.

Closer to the coastline, seawater intrusion from increased coastal flooding, tropical cyclones, and SLR could increase the spatial extent of the subterranean estuary, affecting nutrient dynamics as saline water moves into previously fresh areas of coastal aquifers²⁰. Such changes are also anticipated for micronutrients and trace elements²¹⁷. Seawater intrusion into coastal aquifers amplifies fluid and solid reactions in coastal aquifers that can facilitate the desorption of ions, increasing concentrations of nutrients (e.g., ammonium and phosphate), and enhance the oxidation of organic matter, increasing DIC concentrations as electron acceptors, such as SO₄²⁻, become available²⁰. For instance, in the Okatee subterranean estuary of South Carolina (US), sulfate depletion driven by the oxidation of marine organic matter is associated with an order of magnitude increase in ammonium concentrations and a 2:1 increase in dissolved CO₂ concentrations compared to when dissolved oxygen serves as the primary electron acceptor²⁰. However, increasing salinization can also inhibit processes, such as denitrification, nitrification, and methanogenesis, potentially altering the forms and fluxes of nutrients²¹⁸. Increases in the extent of the subterranean estuary similarly implies that more labile marine organic matter will reach aquifer areas previously occupied by fresh water, where increases in remineralization could subsequently release ammonium and phosphate²¹⁹. Episodic coastal flooding and seawater intrusion might likewise facilitate the displacement of ammonium adsorbed on soils by cation exchange with sea salts^63,220. In model simulations of seawater intrusion in a tidal freshwater marsh, salinization weakens denitrification and increases terrestrial nitrate transport to the coast²²¹. Reductions in fresh coastal groundwater caused by seawater intrusion and declines in coastal erosion-driven storage will also decrease lithogenic dissolved Si from water-rock interactions. This could occur directly due to declines in fresh coastal groundwater storage, which limits the Si substrate available for leaching, as well as indirectly, from reductions in dissolution as alkali cations common to seawater (e.g., Na⁺) can decrease the rate of dissolution of some Si-containing minerals²²².

Seawater intrusion in coastal aquifers might also affect coastal groundwater microbial assemblages and therefore nutrient speciation and availability. Prokaryotic and heterotrophic activity in coastal groundwater responds to salinity changes, with 5–11 fold increases in sites where salinity changes from fresh (<1 psu) to saline sites (>33 psu), respectively²²³. Even slight changes in salinity could impact the zonation and metabolic functioning of microbial assemblages as bacterial groups are distinct along the salinity gradient of the subterranean estuary^224,225.

Along with organic matter concentrations, change in dissolved oxygen content affects aquifer redox chemistry and is an important predictor of subterranean estuary microbial community composition^51,226. In some systems, mixing between suboxic to anoxic fresh coastal groundwaters and oxygenated saline coastal groundwater can greatly alter redox-sensitive N-cycling processes. For instance, under reducing conditions, ammonium dominates, whereas nitrate is the main N species under oxygenated conditions²²⁷. Concentrations of dissolved P, which is also sensitive to oxygen content and pH, could shift due to changes in sorption following declines in seawater pH and oxygen content, which are expected in every ocean basin⁶⁵. In addition, declines in ocean pH might also affect amounts and forms of N in saline coastal groundwater, as is expected for marine N responses to ocean acidification²²⁸.

Box 5 Groundwater warming due to climate change

Climate change is driving rising groundwater temperatures globally, with an average increase of 2.1 °C projected by 2100¹¹¹. Coastal regions will experience varying degrees of warming depending on their location. Even modest increases could alter nutrient concentrations and forms. For example, warming can induce shifts in microbial community composition³¹⁷ and alter gas solubility, such as dissolved O₂³¹⁸, both of which can shift nutrient dynamics^204,319.

a. Projected global change in groundwater temperature from 2000 to 2100 under the SSP 5-8.5 climate scenario, relative to 2000.b. Groundwater temperature projections for the eight highlighted locations through 2100. Data from Benz et al.¹¹¹.

Other anthropogenic modifications

Human activities continue to modify nutrient fluxes to the atmosphere, soils, groundwater, and rivers at all spatial scales. Future human actions can exacerbate or reverse trends in nutrient delivery to coastal oceans. Examples of the former include enhanced fertilizer applications²²⁹, expansion of arable land area²³⁰, inadequate wastewater management²³¹, insufficient riparian buffer zones²³², canalization and dredging^233,234, hydropower dam construction²³⁵, deforestation²³⁶, and increased groundwater pumping²³⁷. Conversely, improvements in agricultural and wastewater practices, water management, and conservation measures can reduce nutrient loads^{231,232,238, 239, 240–241}.

One projection is that global river exports of anthropogenic DIN alone will more than double (47.2 Tg N) by 2050, which will result in a ~ 13% increase in nitrous oxide emissions from marine coastal areas²⁴². Fertilizer and wastewater leachate have already elevated nutrient concentrations in coastal groundwater, with long-lasting impacts due to groundwater’s typically long residence time^{50,243, 244, 245–246}. Alterations to land use and land cover, such as deforestation, have also altered the Si cycle, affecting rivers²⁴⁷, and therefore coastal and continental margin zones.

Another major anthropogenic effect on coastal groundwater is salinization via seawater intrusion, which is often driven by local groundwater extraction²⁴⁸ and can alter subterranean estuary biogeochemistry and nutrient dynamics. Salinization is a worldwide issue, documented in hundreds of coastal cities, that will continue to expand in many regions as the impacts of climate and other local anthropogenic changes unfold together^{249, 250–251}.

Infrastructure, such as dams, artificially deepened shipping channels, and impermeable surfaces, can exacerbate or mitigate the impacts of climate change on nutrient loading (Box 6). Dams, in particular, significantly reduce downstream fluxes of global total N, total P, and Si by 7.4%, 12% and 5.3%, respectively²⁵². While many of the reservoirs built in the 1960s and 1970s heydays of dam building will reach their design lifetimes in the coming decades, with efforts already underway to remove obsolete dams in the US and Europe²⁵³, an ongoing phase of dam building with foci in South America, Africa and Southeast Asia^254,255 will likely contribute to decreasing nutrient fluxes from newly dammed rivers for the coming decades. These and other local and regional anthropogenic stressors will have a critical joint role in shaping how the impacts of climate change will alter the hydrologic nutrient flux to the global coastal ocean.

Box 6 Infrastructure investments can shape future nutrient dynamics

Achieving the United Nations’ Sustainable Development Goals demands investments of >$22.5 trillion in global water infrastructure by 2050^320,321. Studies hint at how effective proactive management of nutrients could significantly reduce future nutrient fluxes from these anthropogenic sources²⁶¹ (see also Table 1). However, existing long-term records of nutrient fluxes in rivers also demonstrate that significant increases can occur over a decade (Thames River, England)³²⁰ or two (Mississippi River, US; Rhine, Europe)^{322, 323–324}, emphasizing the importance of proactive management. To highlight just one example, investments of nearly $10 billion into Vancouver’s (British Columbia, Canada) Iona Island Wastewater Treatment Plant are projected to allow tertiary treatment (N and P removal) by 2035 which will reduce the nutrient loading to the receiving coastal environment. Comparable initiatives, such as India’s Second National Ganga River Basin Project, set ambitious goals for reducing pollution, including excess nutrients, from urban, industrial, and agricultural point and diffuse sources³²⁵. In Europe, efforts are focusing on reducing nutrients by changing fertilizer regulations, expansion of wastewater treatment plants, and reduction of phosphate in detergents³²⁶. Future fluxes of nutrients to the coastal ocean will be determined by such infrastructure transformations and future trends in the use of commercial fertilizers and water, as well as the impacts of climate change.

Dissolved nitrate plus nitrite concentrations in the Thames (England), Mississippi (US), and Rhine (Europe) rivers^{294,326, 327, 328–329}.

Implications of changing hydrologic nutrient fluxes for society and ecosystems

Climate change-induced alterations to coastal groundwater and rivers will affect services and functions of hydrologically connected ecosystems, which span terrestrial, subterranean, and marine environments^256,257. For instance, rivers and coastal erosion supply nearly one-third of Arctic Ocean nutrients that support primary production¹⁸⁵, and nearly 72% of marine fish species representing over ¾ of total catch have life cycles linked to river flows²⁵⁸. Coastal groundwater, via nearshore groundwater discharge and SGD, also influences a variety of terrestrial and marine biota^15,259,260. Changing nutrient fluxes to the coastal ocean could therefore impact diverse coastal processes. These ecological impacts will be highly location-dependent due to the regional to local effects of many CIDs and anthropogenic stressors^14,15. For example, in the continental shelf of Southeast Asia, where riverine nutrient inputs are expected to see the largest increase²⁶¹, these increases are of the same order of magnitude as expected declines in primary productivity due to changes in ocean dynamics and thus could help compensate for nutrient limitations¹.

Future directions

The impacts of climate change will include diverse, region-specific shifts in river and coastal groundwater nutrient fluxes to the coastal zone, shaped by interactions between hydrology, biogeochemistry, and other local anthropogenic stressors. However, understanding of how these fluxes will shift is generally unclear due to the limited availability of long-term datasets needed to assess trends and baseline variability, and the complexity of interactions between CIDs. While we have outlined individual mechanisms that can drive shifts in hydrologic and nutrient responses and therefore, nutrient fluxes, many uncertainties remain, particularly in anticipating the cumulative effects of CIDs and how they vary across global regions. Addressing these gaps will require a better understanding of both individual drivers and their interactions across different climate zones and ecosystem types.

Our understanding of nutrient delivery to the coastal ocean could be revolutionized by enhanced capabilities of long-term and in-situ sensor-based measurements at high temporal resolution. This is particularly important in remote areas and in highly dynamic environments, such as tidally influenced systems²⁶². Furthermore, satellite-derived assessments of river discharge and other land surface and ocean surface characteristics now complement gauging station records and other direct measurements, particularly in ungauged remote watersheds and in countries where hydrologic information remains classified^263,264. However, while satellite-based approaches have improved coverage, in situ hydrologic and geochemical monitoring would benefit from expansion in many regions (e.g., South America, Africa, Southeast Asia). Both technological innovations and targeted investments in regional monitoring networks should improve our ability to monitor how human actions affect future changes in hydrologic nutrient fluxes on local, regional, and global scales, which will be critical to anticipating changes in connected coastal ecosystems. For coastal aquifers, even incremental steps towards increasing monitoring and modeling would significantly advance current understanding, particularly in data-scarce regions where few synoptic and long-term comprehensive data sets exist (e.g., Africa, Southeast and Southwest Asia, South America)¹⁵. More work is urgently needed to understand both current baselines and ongoing changes in globally diverse groundwater and river systems, and addressing these gaps will require a combination of modeling, field-based studies, and increased interdisciplinary collaboration, particularly in regions where monitoring capacity has historically been low.

Acknowledgements

Authors are members or collaborators of Working Group 45 on Climate Change and Greenhouse Gas Related Impacts on Contaminants in the Ocean (WG45) of the Group of Experts on the Scientific Aspects of Marine Environmental Protection (GESAMP), supported by the International Atomic Energy Agency (IAEA), and co-sponsored by the United Nations Environmental Program (UNEP), Intergovernmental Oceanographic Commission (IOC-UNESCO), the World Meteorological Organization (WMO), and the International Maritime Organization (IMO). The IAEA is grateful to the Government of the Principality of Monaco for the support provided to its Marine Environment Laboratories.

Author contributions

C.M.R. and B.P.E. led the writing and developed all visualizations. C.M.R. and A.B. coordinated the project. All authors contributed to the conceptualization, writing, and revision of the manuscript.

Peer review

Peer review information

Communications Earth and Environment thanks Andrea Pain and the other, anonymous, reviewer(s) for their contribution to the peer review of this work. Primary Handling Editor: Alice Drinkwater. A peer review file is available.

Data availability

Data sharing not applicable to this article as no datasets were generated or analysed during the current study.

Competing interests

Annie Bourbonnais is a guest Editorial Board Member for Communications Earth & Environment, but was not involved in the editorial review of, nor the decision to publish this article.

Supplementary information

The online version contains supplementary material available at https://doi.org/10.1038/s43247-025-02594-6.

Publisher’s note Springer Nature remains neutral with regard to jurisdictional claims in published maps and institutional affiliations.

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