of the watershed that is not forested. Third, riverine loads from different Susquehanna subwatersheds have generally declined due to reductions in source input over the same period. Finally, dams and reservoirs can strongly modulate the storage and release of particulate constituents and the extent of modulation has varied considerably over time as reservoirs fill. Two decades ago, the Conowingo Reservoir and two others in the lower Susquehanna River trapped about 2%, 45%, and 70% of annual N, P, and SS load, respectively (Langland & Hainly, 1997). Currently, the reservoir system is no longer an effective trap of these constituents (Hirsch, 2012; Langland, 2015; Zhang et al., 2013; Zhang, Hirsch, et al., 2016).
In the NAS watershed, the annual export of nutrient and sediment is directly linked to river flow and can be estimated using Po River loads as a reference (Cozzi & Giani, 2011). However, the short‐term dynamics of riverine loads are complicated during freshets by the effects of flow on the erosion, groundwater inputs, dilution, and biological processes in the river environment (Marchina et al., 2015; Tesi et al., 2013). The N and P loads in the marine environment have long‐term trends consistent with watershed inputs from industrial activities, agriculture, livestock, urban settlements, and atmospheric deposition; dynamics of their delivery are also modulated by the variable retention of these elements in soils and aquifers (Cozzi et al., 2019; Palmeri et al., 2005; Salvetti et al., 2006; Viaroli et al., 2018). In general, retention mechanisms differ considerably between N and P—sedimentation may be the major retention mechanism for particulate P during overland flow conditions (Cirmo & McDonnell, 1997; Hoffmann et al., 2009), whereas sorption/desorption reactions and denitrification are more important for P and N, respectively, during subsurface flow conditions (House, 2003; Withers & Jarvie, 2008).
2.4.3. Watershed Management
For CB, coordinated efforts have been implemented to reduce pollutant inputs since the first Chesapeake Bay Agreement was signed in 1983. Here we provide a brief overview of historical changes in management and practices associated with point, agricultural, and stormwater sources in the Susquehanna River (Zhang et al., 2013). For point sources, the more important historical controls were the P ban in detergents (Litke, 1999) and the adoption of increasingly effective nutrient removal technologies at wastewater treatment plants (Chesapeake Executive Council, 1988). For agriculture nonpoint sources, many strategies have focused on controlling fertilizer and manure applications, including regulations on storage and usage of animal manure and regulations on concentrated animal operations and feeding operations (New York State Department of Environmental Conservation, 2007; Pennsylvania Department of Environmental Protection, 2004; US Department of Agriculture and US Environmental Protection Agency, 1999). For stormwater sources, the USEPA initiated the National Pollutant Discharge Elimination System Phase I regulations in 1990 for Municipalities with Separate Storm Sewer Systems (MS4s) serving populations of 100,000 or more (US Environmental Protection Agency, 2000) and expanded the program in 1999 to include smaller MS4s (US Environmental Protection Agency, 2000). States also took actions to promulgate regulations on stormwater discharges in the late 1990s to early 2000s (New York State Department of Environmental Conservation, 2007; Pennsylvania Department of Environmental Protection, 2004).
For the NAS, efforts were made by Italian regulators to progressively reduce P content in detergents to 1% (6% for automatic laundry detergents) in 1989 (Marchetti et al., 1989; Rinaldi, 2014). In that year, industries also substituted sodium‐triphosphate‐based detergents with zeolite A (Glennie et al., 2002). The control of N loads has been improved recently with the adoption of the European Nitrates Directive (91/676/EEC). However, the reduction of N loads was less effective compared to that of P loads, because N is largely contributed by nonpoint sources (Viaroli et al., 2010, 2018). The creation of designated Authorities for specific hydrographic basins and of Regional Environmental Protection Agencies in 1994 further supported a coordinated monitoring of continental waters in Italy. The main measures adopted to reduce eutrophication from river loads were (a) establishment of spill basins and manure wastes for crop and livestock farming, (b) implementation of good farming practices in vulnerable areas, and (c) implementation of the Urban Waste Water Treatment Directive (91/271/EEC) (Bortone, 2014).
The N and P inputs in the NAS are not limited to riverine inputs. Direct inputs to the NAS occur via groundwater discharge and treatment plants, but are poorly quantified. Groundwater aquifers, particularly in the NE Adriatic region, are very sensitive to external pollution due to their large draining capacity of surface waters and low self‐cleaning potential (The EU.WATER Project, 2010). For treatment plants, several underwater pipelines have been built since the 1980s, especially in the NE Adriatic region. Despite a gradual upgrade of treatment plants from secondary to tertiary treatment, wastewater loads continue to be an important source of nutrients in the NAS (Cozzi et al., 2014; Scroccaro et al., 2010; Sekulić et al., 2004; Volf et al., 2018).
2.5. MAJOR CHALLENGES
2.5.1. Legacy Sources
Many restoration efforts around the world have not yet achieved significant progress in reducing riverine loads of nutrient and sediment due to challenges such as legacy inputs, which accumulate and are stored in groundwater aquifers and sediments. Such effects have been documented for watersheds in North America (e.g., Chesapeake Bay, Mississippi River, and Lake Erie) and Europe (Basu et al., 2010; Jarvie et al., 2013; Sharpley et al., 2013; Van Meter et al., 2017; Van Meter et al., 2017, 2018; Vero et al., 2017). For CB, there is strong evidence for the importance of legacy sources. For example, riverine loads in the Susquehanna remained relatively constant in the past 30 years despite strong reductions in anthropogenic inputs to watersheds (Zhang, Ball, et al., 2016). Such patterns may reflect the effects of legacy sources (Basu et al., 2010; Thompson et al., 2011). Van Meter et al. (2017) reported that N dynamics in the Susquehanna are dominated by groundwater legacies, with 18% of the current annual N input to the river being at least 10 years old. Apparent ages of groundwater in the CB watershed can reach 20 years or more (Focazio et al., 1997) and base flow accounts for a major fraction of riverine N load at many CB sites (Bachman et al., 1998). For this region, the legacy stores are comprised primarily of groundwater for N (Bachman et al., 1998; Sanford & Pope, 2013), surface soils and river sediments for P (Ator et al., 2011; Sharpley et al., 2013), and stream corridors and reservoir beds for sediment (Gellis et al., 2008; Pizzuto et al., 2014; Walter & Merritts, 2008). These results suggest the importance of considering lag time between implementation of management actions and achievement of water‐quality improvement. For the NAS, budget estimates indicate the accumulation in river watersheds of inorganic and organic N and P from anthropogenic sources that still negatively affect the quality of freshwater systems (Giani et al., 2012; Viaroli et al., 2018; Volf et al., 2018) and river‐dominated coastal areas (Alvisi & Cozzi, 2016).
2.5.2. Climate Change
Climate change is another major challenge to ecosystem restoration (Charlton et al., 2018; Forber et al., 2018; Meier et al., 2018; Rankinen et al., 2016; Sinha et al., 2017). In general, climate change is expected to result in increased air and water temperature and an acceleration of the water cycle (Bloschl et al., 2017; Milly et al., 2005; Najjar et al., 2010; Rice & Jastram, 2014; Rice et al., 2017), which can alter the volume transport of freshwater and inputs of nutrients and sediments. For example, Sinha et al. (2017) estimated that climate‐change‐induced precipitation changes alone will substantially increase (19 ± 14%) riverine inputs of TN within the continental United States by the end of the century. In addition, the effects of climate change can differ among seasons. For CB, projected acceleration of the water cycle is expected to increase river runoff and associated inputs of nutrients and sediments during winter–spring and to decrease runoff during summer–fall (Wagena et al., 2018). Thus, management strategies for CB need to account for the impact of projected climate change on water quality. In this context, modeling and assessment is underway in the Chesapeake Bay Program partnership to evaluate the effects of climate change on nutrient export, efficacy of best management practices, and water quality in the estuary.
In contrast, climate‐driven changes in the water cycle in the NAS watershed may tend toward persistent periods of low runoff alternating with episodic events of high discharge. Climate change appears to be increasing