Long-term trends in nitrate and chloride in streams in an exurban watershed

doi:10.21203/rs.3.rs-1220673/v1

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Research Article

Long-term trends in nitrate and chloride in streams in an exurban watershed

https://doi.org/10.21203/rs.3.rs-1220673/v1

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The lines between natural areas and human habitats have blurred as urbanization continues, creating a need for the study of ecosystems at all levels of development. This need is particularly acute for exurban environments, which have low population density but are rapidly changing and have a dynamic mix of natural and human-dominated features. We examined long-term (1998 - 2018) trends in nitrate and chloride concentrations and fluxes in forested and exurban streams in Baltimore County, Md USA. Concentrations and fluxes of nitrate and chloride were an order of magnitude higher in the exurban stream than the forested stream and were increasing even though snowfall and road salt use did not increase over the study period. In the forested stream, concentrations of chloride increased from 1998 to 2008, but decreased from 2008 to 2018 due to unquantified factors. Concentrations of nitrate decreased in the forested stream, likely due to decreases in atmospheric deposition. These decreases in atmospheric deposition, and efforts to reduce fertilizer use by county governments, do not appear to have affected nitrate concentrations in the exurban watershed. Any efforts to reduce the concentrations and fluxes of chloride and nitrate in exurban streams will likely benefit substantially from further understanding of the mechanisms underlying the temporal patterns.

nitrate

chloride

salt

exurban

watershed

long-term

Long-term monitoring of chloride and nitrate in exurban streams suggests that more aggressive efforts are needed to reduce these solutes.

Urban land use affects watershed dynamics and stream water quality in complex ways. The addition of impervious surfaces and drainage infrastructure alters the way water moves from the land to streams, and the use of fertilizers, water softeners, deicing agents and other chemicals alters the chemistry of that water. Urbanization also affects the ability of soils, plants and other ecosystem components to process water and chemicals, creating enormous complexity in the watershed dynamics of cities (Walsh et al. 2005). These effects vary markedly with the density and nature of development within metropolitan areas (Moore et al. 2017; Blaszczak et al. 2019), and with climate variability (Kaushal et al. 2008; Bettez et al. 2015; Hale et al. 2016). Unraveling this complexity has been facilitated by long-term studies that apply the watershed approach (Likens 2013). In the Baltimore Ecosystem Study (BES), a component of the U.S. National Science Foundation funded the Long Term Ecological Research (LTER) network, the watershed approach has been used as a platform to examine differences among forest, agricultural and urban, suburban and exurban areas within the Baltimore, Maryland metropolitan area (Groffman et al. 2019).

Two solutes, nitrate and chloride, are of particular concern to the water quality of urban, suburban and exurban watersheds. Nitrate is the most common and mobile form of reactive nitrogen, acting as a drinking water pollutant and cause of eutrophication in coastal waters, including the Chesapeake Bay (Boesch et al. 2001). The dominant sources of nitrate in urban watersheds are fertilizers and sewage (Hobbie et al. 2017; Byrnes et al. 2020). The dominant source of chloride in watersheds in areas with frozen winter precipitation is the salt used to mitigate icy road conditions in colder months (Kaushal et al. 2018). The chloride from road salt is transported to streams and aquifers by runoff water, harming aquatic life, damaging infrastructure, and contaminating drinking water (Findlay and Kelly 2011).

Previous work in urban watersheds has identified complexities and uncertainties around nitrate dynamics. Results from several cities have found significant retention of nitrogen in urban watersheds, i.e., inputs/sources greater than exports (Baker et al. 2001; Bettez et al. 2015; Hobbie et al. 2017). Nitrogen dynamics in the Baltimore area are further complicated by three co-occurring long-term changes. First, atmospheric nitrogen deposition has been declining in the Chesapeake Bay region since implementation of the 1990 Clean Air Act Amendments, which limits NOx emissions (Eshleman and Sabo 2016). Second, nitrate concentrations in the Baltimore environment are likely affected by efforts to reduce nitrogen fertilizer use and nitrogen exports from septic systems and municipal sewer systems (Birch et al. 2011; Wainger 2012; Reisinger et al. 2018). Third, increases in the volume and intensity of precipitation reduce ecosystem nitrogen retention and result in increased N export to downstream water bodies (Kaushal et al. 2008; Bettez et al. 2015).

Similar to nitrate, research has identified complexities and uncertainties associated with chloride concentrations in urban and suburban streams (Kaushal et al. 2018). While high concentrations of chloride are expected following snow events and the subsequent application of road salt in winter, salty winter snowmelt water recharges groundwater, the primary source of water for low flows (base flow), giving rise to elevated salt concentrations at base flow year round (Gelhar and Wilson 1974; Kaushal et al. 2005; Blaszczak et al. 2019). These complex interactions are further affected by increases in the amount and intensity of precipitation in ways that are not clear. There is an expectation that chloride concentration in streams should decline as climate warming should decrease snow events and the subsequent need for road salt application. In addition, efforts to improve the efficiency of road salt use should reduce overall mass loading of salt, resulting in lower chloride concentrations in streams.

This study focused on changes in nitrate and chloride concentrations from 1999 – 2018 in two long-term study watersheds in BES; Pond Branch (POBR) which is forested and serves as a reference site for BES and Baisman Run (BARN) to which POBR is a tributary. BARN, which is characterized by exurban land use in the upper third of the catchment, and is otherwise forested, is part of the Gunpowder Falls basin, which drains into one of Baltimore’s three major water supply reservoirs (Law et al. 2004). The forested and exurban watersheds are ideal for exploring how effects of climate, land management (fertilizer and road salt application), and atmospheric chemistry interact as these watershed dynamics are not dominated by sewage and stormwater infrastructure, although BARN has 120 houses with septic systems (Law et al. 2004). We tested two hypotheses: 1) Chloride concentrations will decline over time in the exurban watershed as snowfall and associated road salt use will decline as a result of climate warming. 2) Nitrate concentrations will decline in both watersheds over time due to declines in atmospheric deposition as facilitated by the 1990 Clean Air Act Amendmentsand efforts to reduce fertilizer use such as the introduction of the 2013 Maryland’s Lawn Fertilizer Law (https://mda.maryland.gov/Pages/fertilizer.aspx).

Study sites. BARN is a 420 hectare (ha) subwatershed (latitude 39°28′49.1′′, longitude 76°41′15.0′′) of the Gunpowder Falls watershed and is characterized by low density, large lot development on septic systems in the upper third of the watershed (Figure 1). The remainder of the watershed is a forested park. POBR is a forested 40 ha subwatershed of BARN. The entire watershed is underlain by the medium- to coarse-grained micaceous schist of the Loch Raven Formation. Depth to saprolite is highest on ridges, thins (<1m) at steep midslope positions, and is 1–2m in bottomland locations (Cleaves et al. 1970). Soils range from silt clay loam to silt loam in the riparian areas to sandy loam on steeper slopes. Forested areas are dominated by approximately 100 year old Quercus spp. (oaks) and Carya spp. (hickory).

Water Sampling and Analysis. The BES has collected stream chemistry samples at POBR and BARN since 1998. Sites are continuously monitored for discharge by the USGS (BARN: https://waterdata.usgs.gov/usa/nwis/uv?0158358; Pond Branch: https://waterdata.usgs.gov/usa/nwis/uv?0158357) and weekly grab samples are taken for chemical analysis without regard to flow conditions. The weekly stream samples are filtered (0.45 µm) and analyzed for NO₃⁻ and Cl⁻ concentrations by ion chromatography (Groffman et al. 2018).

Hydrologic conditions at each sampling date were calculated using algorithms described by Nathan and McMahon (1990) that are incorporated into the EcoHydRology program (https://cran.r-project.org/web/packages/EcoHydRology/EcoHydRology.pdf). For each sample date, this program produces estimates of: (1) the total discharge (Q) in cubic meters per second (cms), (2) percentage of discharge attributable to base flow, (3) percentage of discharge attributable to stormflow and (4) base flow index, i.e., the ratio of base flow to stormflow (Table 1). The period of analysis for both sites was 1999 to 2018 (a 19-year range) because data on chloride and nitrate concentrations from 1998 were not available for BARN.

Table 1

Hydrologic data from the Baisman Run and Pond Branch watersheds
	Pond Branch (forested)				Baisman Run (exurban)
	Highest discharge (m³/d)	Lowest discharge (cm³/d)	# grab samples	Average base flow ratio on sampling days	Highest discharge (cm³/d)	Lowest discharge (m³/d)	# grab samples	Average base flow ratio on sampling days
1999	2936	24	52	0.68	5627	2691	14	0.84
2000	1248	171	52	0.79	18594	1541	52	0.8
2001	489	73	55	0.76	5382	783	55	0.77
2002	245	2	55	0.66	3205	98	57	0.65
2003	1566	171	50	0.81	20135	2349	51	0.75
2004	2202	220	52	0.84	41592	1639	52	0.82
2005	1076	147	50	0.8	11597	979	53	0.79
2006	1028	98	53	0.78	16172	1248	55	0.79
2007	930	49	55	0.76	9297	367	54	0.74
2008	440	5	51	0.7	8245	269	51	0.76
2009	1688	122	47	0.75	17982	1884	48	0.79
2010	1101	122	49	0.76	11744	856	49	0.78
2011	979	73	51	0.73	13163	979	51	0.73
2012	1713	171	51	0.82	16001	1199	64	0.83
2013	1370	147	50	0.75	14508	1664	50	0.83
2014	8416	220	49	0.74	87098	2153	49	0.74
2015	1028	196	52	0.78	18790	1664	51	0.78
2016	1492	122	49	0.76	19157	1297	50	0.81
2017	3425	49	50	0.68	25689	1125	50	0.8
2018	294	73	13	0.68	17199	1395	53	0.8

Flux calculations were made with Weighted Regressions on Time, Discharge, and Season (WRTDS), a USGS statistical tool (Hirsch et al. 2010), which calculates solute fluxes from stream flow, solute concentration, and time by fitting locally weighted regressions (Bettez et al. 2015; Reisinger et al. 2018). Yields were calculated by dividing flux by watershed area.

Road Salt Data. Baltimore County publishes the mass of road salt applied annually (https://www.baltimorecountymd.gov/Agencies/publicworks/highways/stormrelatedcosts.html). Their data includes a table on Storm Related Costs, including the mass of salt used for each fiscal year (July 1 - June 30), the depth of snowfall, and the number of events requiring salt application, from 2001 to 2019. We assumed that BARN contains roads that are typical of Baltimore County and that county-wide trends were applicable to this watershed.

Fertilizer data. Data on fertilizer usage in Baltimore County were derived from three sources: First, we used a national compilation of county-by-county fertilizer use, split into farm and non-farm use, compiled by the USGS: https://doi.org/10.5066/F7H41PKX.

However, these data are only available until 2012. Second, data on fertilizer applied by county personnel on county owned and managed land are compiled by Baltimore County for permit requirements: https://www.baltimorecountymd.gov/Agencies/environment/npdes/ and are presented in their annual report from 2018: http://resources.baltimorecountymd.gov/Documents/Environment/npdes/2018/fullnpdes2018.pdf. Third, Baltimore County implemented a “lawn fertilizer law” in 2013 (https://mda.maryland.gov/Pages/fertilizer.aspx) that mandates collection of data on fertilizer applied by professional lawn care services. Collection and organization of these data are incomplete and are only available for the years 2015, 2017 and 2018 (Table 2). We assumed trends in fertilizer use in BARN are consistent with the trends in Baltimore County as a whole.

Table 2

Fertilized area and amount of fertilizer applied in Baltimore County in 2015, 2017, 2018. Data compiled from Fertilizer Application Report produced by the Maryland Department of Agriculture as part of “lawn fertilizer law” that was passed in 2013. The data for 2016 are not included due to data quality concerns for that year.
Year	Fertilized area (ha)	Nitrogen applied (kg)
2015	6241	402099
2017	5412	507955
2018	5308	515797

Atmospheric Deposition. Atmospheric nitrogen and chloride deposition are measured by the EPA in Beltsville, MD, approximately 45 miles from BARN and data are available for nitrogen at https://www3.epa.gov/castnet/site_pages/BEL116.html and for chloride at http://nadp.slh.wisc.edu/data/ntn/ntnAllsites.aspx. Data available included wet NO₃⁻ and ammonium (NH₄⁺) deposition as well as “dry deposition” of nitric acid vapor (HNO₃) and particulate NO₃⁻ and NH₄⁺. Values for total annual deposition (wet plus dry) were used here and referenced as Total N Deposition.

Septic Systems. Estimates of nitrate and chloride loading from septic systems were produced by assuming that each of the 120 houses in the BARN watershed has a septic system that produces 650 L/day of effluent that is discharged to groundwater (Rutledge et al. 1993) and that this effluent has a chloride concentration of 400 mg/L (Kochary et al. 2017) and a nitrate concentration of 70 mg N/L (Gold et al. 1990).

Climate. Monthly precipitation and temperature data are available from the Baltimore Washington International airport site from the Climate and Hydrology Database Projects database (http://climhy.lternet.edu/).

Statistical analyses. All analyses were done on annual values for all variables. Comparisons of the two watersheds were made using one-way analysis of variance. Trends through time, and relationships between variables were explored with linear (Pearson) and nonparametric (Spearman) correlations. Relationships were considered statistically significant at p < 0.05. All analyses were done using SAS (version 9.4).

Hydrology and climate

At BARN, discharge ranged from 98 to 87,098 m³/d (0.02 to 20.7 mm/d) over the 18 year data collection period (Table 1). POBR had stream discharge values ranging from 2 to 8416 m³/d (0.02 – 21 mm/d). Average base flow index for all sample dates over the 19 year period was 0.78 ± 0.04 at BARN and 0.75 ± 0.05 at POBR (Table 1). There was no significant trend in discharge over time in either watershed.

Mean annual temperature showed a marginally significant increase (r = 0.40, p < 0.10) from 1998 – 2018 (data not shown) but there was no significant increase in winter temperature over this period. There was no trend in total annual or winter precipitation over the study period although Reisinger et al. (2018) reported a significant increase in winter precipitation between 1999 and 2016. There was also no trend in seasonal snowfall, the number of snowstorms requiring application of road salt, or total salt use from 1999 - 2018.

Chloride

In an analysis over all years, chloride concentrations (Figure 2a) and loads (Figure 2c) were significantly higher (p < 0.001) in the exurban stream, BARN (~ 40 mg/L), than in the forested stream, POBR (~3 mg/L). Concentrations (r = 0.86, p < 0.001) and loads (r = 0.63, p < 0.01) significantly increased from 1999 – 2018 in BARN (from ~25 to ~ 45 mg/L) and showed no significant change over the same period in POBR. However, when the data were split (in response to visually obvious patterns), concentrations in POBR significantly (r = 0.79, p < 0.01) increased from 1999 – 2008 (from ~ 2.5 to ~3.0 mg/L) and significantly (r = -0.78, p < 0.01) decreased from 2008 – 2018 (Figure 2b). These temporal trends in POBR were not significant for fluxes (Figure 2d).

Mean annual chloride concentrations in BARN were significantly correlated with annual precipitation (r = 0.51, p < 0.05) and with the number of storms requiring application of road salt (r = 0.52, p < 0.05), but not with total salt mass applied. Chloride loads in BARN were significantly correlated with winter precipitation (r = 0.65, p < 0.01), snowfall (r = 0.65, p < 0.01), and total salt mass applied (r = 0.49, p < 0.05). Total salt mass applied was significantly correlated with mean annual temperature (r = -0.56, p < 0.05), total annual precipitation (r = 0.54, p < 0.05), total snowfall (r = 0.70, p < 0.01), and the number of storms requiring application of road salt (r = 0.62, p < 0.01). Chloride concentrations in POBR were not correlated with any climate variables, but chloride loads were correlated with winter precipitation (r = 0.60, p < 0.01) and snowfall (r = 0.56, p < 0.05). Atmospheric deposition of chloride ranged from 1.2 to 3.7 kg/ha/y, showed no trend with time, and was not correlated with stream concentrations or fluxes of chloride in either watershed.

Septic systems were estimated to add 11,388 kg of chloride to the watershed, which is equivalent to ~ 21% of typical chloride exports from the watershed. The estimate of septic system chloride is based on literature values and is therefore approximate. The estimate of chloride exports is based on average flows and concentration over the study period.

Nitrate

Nitrate concentrations (Figure 3a) and loads (Figure 3c) were significantly (p < 0.001) higher in BARN than in POBR (Figure 2a). Concentrations increased (from ~1.5 to ~ 1.7 mg N/L) significantly in BARN (r = 0.52, p < 0.05) and decreased (from ~ 0.06 to ~ 0.02 mg N/L) significantly in POBR (r = -0.48, p < 0.05) from 1999 – 2018 (Figure 3b). There were no significant trends in loads with time in either BARN (Figure 2c) or POBR (Figure 3d).

Nitrate concentrations in POBR were negatively correlated with annual (r = -0.56, p < 0.01) and winter (r = -0.41, p < 0.10) precipitation and positively correlated with atmospheric deposition (r = 0.43, p < 0.10) (Figure 4b). Atmospheric deposition markedly decreased over time (r = -0.93, p < 0.001). In contrast to concentrations, nitrate loads in POBR were positively correlated with winter precipitation (r = 0.50, p < 0.05) and snowfall (r = 0.47, p < 0.05), but not with atmospheric deposition (Figure 4d).

Nitrate concentrations in BARN were negatively correlated with atmospheric deposition (r = -0.39, p < 0.10) (Figure 4a). Nitrate loads in BARN were positively correlated with winter precipitation (r = 0.68, p < 0.01) and snow (r = 0.66, p < 0.01). There were no correlations between nitrate loads and atmospheric deposition in BARN (Figure 4c).

There were no trends in non-farm fertilizer usage in Baltimore County from 1999 to 2012. Fertilizer applied by Baltimore county personnel on county owned and managed land showed no trend between 1999 and 2018. There were no relationships between fertilizer usage and nitrate concentrations or loads in BARN (Table 2).

Septic systems were estimated to add 1,993 kg of nitrate-N to the watershed, which is equivalent to ~ 100% of typical nitrate-N exports from the watershed. The estimate of septic system nitrate is based on literature values and is therefore approximate. The estimate of nitrate exports is based on average flows and concentration over the study period.

Long-term data on stream solute concentrations and loads are rare and valuable for understanding complex controls on water quality in both natural and human-dominated ecosystems (Frazar et al. 2019; Seybold et al. 2019). Nitrate and chloride have proven to be challenging to manage, in both urban and agricultural watersheds. The use of road salt and fertilizer is difficult to reduce, and interactions between management, climate change, and variability have reduced the effectiveness of management efforts (Bettez et al. 2015). The results here emphasize these challenges. In the exurban study watershed, stream chloride concentrations have risen markedly over a 20 year period even when snowfall and road salt use by Baltimore County has not increased. Fertilizer usage has not declined despite significant efforts to reduce its use, and marked declines in atmospheric deposition do not appear to have affected nitrate concentrations in this low density, exurban watershed.

Chloride

The consistently high chloride concentrations and loads in the exurban BARN stream compared to the forested POBR stream have been reported earlier (Kaushal et al. 2005). These results are consistent with many studies characterizing increases in salt concentrations in human-affected watersheds around the world (Kaushal et al. 2018). These studies have shown that watersheds with more human activity, especially roads treated with salt during winter, have high chloride concentrations in streams, throughout the year. Other important sources of chloride in exurban watersheds include septic systems, water softeners, fertilizer, and atmospheric deposition (Oberhelman and Peterson 2020; Overbo et al. 2021). None of these sources has increased in BARN over the last 20 years, although there has been some new residential development with septic systems. We estimate that septic systems contribute an amount of chloride that is equivalent to ~21% of exports from the watershed. Septic effluents are added directly to the groundwater and may therefore be more readily transported to the stream than road salt.

There is much current research interest and practical concern about the occurrence of high chloride concentrations in streams during seasons when chloride is not being applied to roads, and persistent increases in concentrations despite heightened awareness and efforts to reduce chloride contamination (Kelly et al. 2019). Our results may add to these concerns as chloride concentrations have increased markedly in BARN from 1999 to 2018, even though snowfall, road salt usage and urban land cover (Bird et al. 2018) have not increased over this time period. A new subdivision (~ 10 houses) was added to the watershed in 2007/2008, but we see no evidence of a change in the rate of increase in either chloride concentrations or loads after that time. The high concentrations in summer, and the persistent increase over the 20 year record are likely driven by storage of chloride in watershed groundwater which creates a lag between changes in road salt use and stream chloride concentrations (Gelhar and Wilson 1974). In an exurban watershed in New York, Kelly et al. (2019) estimated that it will take between 20 and 30 years for decreases in road salt use to be apparent in stream chloride concentrations.

Despite the importance of storage and lag effects, we did observe significant correlations between annual mean chloride concentrations and loads and total annual and winter precipitation, and with the number of snowfall events requiring the use of road salt. The correlations with precipitation are likely caused by increased precipitation flushing chloride across the landscape and into streams. The correlation with the number of snowfall events requiring the use of road salt is straightforward, i.e., we expect to see more salt in the stream in years with more salt application. Further, while snowfall depth did not change over the course of the study, there was a significant negative relationship between road salt usage and mean annual temperature (r = -0.55, p < 0.05), and as expected a significant positive relationship between road salt usage and snowfall (r = 0.82, p < 0.001). These relationships suggest that continued climate warming and reductions in snowfall will ultimately lead to less road salt use. In combination with efforts to increase the efficiency of road salt use, these changes may ultimately result in decreases in chloride concentrations in streams (Kelly et al. 2019).

The complex changes in chloride concentrations in Pond Branch, the forested reference stream, illustrate how natural processes influence concentrations in complex and poorly understood ways. It is not clear why concentrations in this stream increased from 1999 – 2008 and then decreased from 2008 – 2018 and there were no correlations between these concentrations and any climatic variables, or with atmospheric deposition of chloride. Chloride concentrations in forested streams can be influenced by the source of precipitation, e.g., precipitation that originates over the ocean, or by evapotranspiration that removes water and concentrates chloride in the remaining soil solution (Lovett et al. 2005). Further research would likely help to isolate these mechanisms and to determine if they influence concentrations in more urbanized watersheds. It is interesting to note that while temporal patterns in chloride concentrations in POBR were strong and significant, there were no significant trends in loads. Trends in concentrations are likely more indicative of internal watershed processes as most of our samples were taken under base flow conditions (Table 1) and therefore are less affected by hydrologic conditions than load estimates (Wherry et al. 2021).

Nitrate

Along urban to exurban gradients, nitrate concentrations are often relatively high in exurban areas. There are multiple potential sources of nitrate in these areas, including atmospheric deposition, lawn and agricultural fertilizer, septic systems, and “legacy” nitrate in groundwater from past sources (Kaushal et al. 2011). Long-term data from POBR and BARN provide an opportunity to evaluate how these sources interact with climate variation and change. Of particular interest are marked declines in atmospheric deposition that have occurred over the past 20 years due to changes in the Clean Air Act (Sabo et al. 2016; 2019), and efforts to reduce fertilizer use as part of Watershed Implementation Plans to achieve Total Maximum Daily Load goals for reducing nitrogen delivery to the Chesapeake Bay (Birch et al. 2011; Wainger 2012).

While our results suggest that declining atmospheric deposition is driving declines in nitrate concentrations in POBR, it does not appear to be affecting concentrations in BARN. There was a positive correlation between deposition and nitrate concentration in POBR, i.e., concentrations declined along with deposition over the past 20 years. At the same time, there was a negative correlation between deposition and nitrate concentrations in BARN, i.e., concentrations increased as deposition declined over the past 20 years. Interestingly, there was no correlation between deposition and nitrate loads in either stream. As for chloride, analysis of base flow nitrate concentrations may be more useful for assessing the effects of environmental change on internal watershed processes than analysis of loads, which are dominated by hydrologic factors (Wherry et al. 2021).

Clearly, other nitrogen sources, e.g., fertilizer, septic systems, overwhelm any declines in atmospheric deposition in BARN. Indeed, septic systems appear to be the overwhelmingly dominant source of anthropogenic N in BARN as our estimates of septic system nitrate-N additions are roughly equivalent to watershed exports. The increase in nitrate concentrations in BARN over the past 20 years may be due to the addition of a new subdivision (~10 houses) to the watershed, or to storage of nitrate in groundwater similar to what occurs for chloride. The rate of increase in nitrate is much lower than the rate of increase of chloride, reflecting the high biological reactivity and diverse potential sinks of nitrate relative to chloride.

As for chloride, the influence of annual climate was important for nitrate in both watersheds. There were significant negative correlations between annual and winter precipitation and stream nitrate in POBR suggesting that increases in precipitation dilute nitrate concentrations in this watershed. In contrast, there were positive relationships between nitrate loads and winter precipitation and snowfall in both POBR and BARN, suggesting that increased precipitation more effectively flushes this ion out of these watersheds.

The high and stable nitrate concentrations in BARN suggest that efforts to reduce fertilizer usage in this watershed have yet to produce results. It is difficult to track changes in fertilizer usage, especially in specific locations. We see no evidence in national-scale assessments for declines in non-farm fertilizer usage in Baltimore County from 1999 to 2012, and there were no relationships between this usage and nitrate concentrations in BARN. We also see no trends in the amount of fertilizer used by county personnel between 1999 to 2018, but this usage is not particularly relevant or indicative of use in our specific study watershed. Baltimore County passed a new fertilizer law in 2013 and has started to collect more detailed data on fertilizer usage across the county. The limited data available from this collection show a small decline in the area fertilized, but a small increase in the total amount of fertilizer applied between 2015 and 2018. There is a clear need for further analysis of relationships between nitrate concentrations and fertilizer usage in BARN. Further analysis of relationships between nitrate concentrations and fertilizer usage in BARN would likely lead to a better understanding of the anthropogenic effects on water quality.

Given the importance of septic systems as the dominant source of nitrate to BARN, efforts to reduce nitrate exports may need to focus on installation of nitrate reducing septic systems (Withers et al. 2013). As for chloride, accumulation of nitrate in groundwater means that there will be a significant lag between any reductions in fertilizer or septic sources and nitrate concentrations in exurban streams (Wherry et al. 2021). Stream restoration, which has the capacity to increase biological uptake of nitrate, especially in base flow dominated watersheds such as BARN could also decrease nitrate export in this watershed (Craig et al. 2008; Reisinger et al. 2019).

Implications for Management

Water quality in the exurban BARN watershed has clearly not improved over the past 20 years. Chloride concentrations have markedly increased and nitrate concentrations have remained stable, despite great concerns and efforts, including the mandated reductions of these loads in the Chesapeake Bay Watershed, to reduce the delivery of these solutes to receiving waters. Reducing chloride concentrations will require significant reductions in salt usage, and patience as it may take decades for legacy chloride that has accumulated in groundwater to be flushed from the watershed. More research on ways to reduce salt usage and the factors influencing retention and legacy dynamics would help develop solutions to these persistent problems. Reducing nitrate will require significant reductions in fertilizer and/or septic system sources, coupled with practices to increase nitrate retention, and a better understanding of legacy effects. Detailed tracking of response to the new fertilizer law likely be an important part of this assessment.

Improving water quality in exurban streams is not easy. Understanding the multiple factors affecting stream solute concentrations, and the slow and complex response to these factors is greatly aided by long-term data collection. Establishing and maintaining strong linkages between this data collection, models and management programs will be fundamentally important to this challenging endeavor.

Funding

This work was performed as part of the Baltimore Ecosystem Study, a long-term ecological research project funded by the National Science Foundation Long-Term Ecological Research Program (DEB 1027188). The Cary Institute of Ecosystem Studies and the National Science Foundation (DBI 1559769) supported the Research Experience for Undergraduates of 2020 which allowed Emma Castiblanco to conduct research.

Competing Interests

The authors have no financial or non-financial conflicts of interest to disclose.

Availability of data and material

All data for this paper are publicly available. The stream chemistry and flow data are available through the Environmental Data Initiative as cited in the paper. Ancillary data on climate and land management are available on the specific websites cited in the paper

Code availability

There is no code used to report.

Acknowledgments

We thank the numerous field technicians and volunteers who have braved a range of weather conditions to perform stream monitoring for the Baltimore Ecosystem Study. We also thank Lisa Martel and the staff of the Cary Institute Analytical Facility for the analysis, processing, and QA/QC of all of the water quality data. Any use of trade, firm, or product names is for descriptive purposes only and does not imply endorsement by the U.S. Government

Baker LA, Hope D, Xu Y, Edmonds J, Lauver L (2001) Nitrogen balance for the central Arizona-Phoenix (CAP) ecosystem. Ecosystems 4: 582-602.
Bettez ND, Duncan JM, Groffman PM, Band LE, O'Neil-Dunne J, Kaushal SS, Belt KT, Law N (2015) Climate variation overwhelms efforts to reduce nitrogen delivery to coastal waters. Ecosystems 18: 1319 - 1331.
Birch MBL, Gramig BM, Moomaw WR, Doering IIIOC, Reeling CJ (2011) Why metrics matter: Evaluating policy choices for reactive nitrogen in the Chesapeake Bay watershed. Environmental Science & Technology 45: 168-174.
Bird DL, Groffman PM, Salice CJ, Moore J (2018) Steady-state land cover but non-steady-state major ion chemistry in urban streams. Environmental Science & Technology 52: 13015-13026.
Blaszczak JR, Delesantro JM, Zhong Y, Urban DL, Bernhardt ES (2019) Watershed urban development controls on urban streamwater chemistry variability. Biogeochemistry 144: 61-84.
Boesch DF, Brinsfield RB, Magnien RE (2001) Chesapeake Bay eutrophication: Scientific understanding, ecosystem restoration, and challenges for agriculture. Journal of Environmental Quality 30: 303-320.
Byrnes DK, Van Meter KJ, Basu NB (2020) Long‐term shifts in U.S. nitrogen sources and sinks revealed by the new TREND‐nitrogen dataset (1930‐2017). Global Biogeochemical Cycles.
Cleaves ET, Godfrey AE, Bricker OP (1970) Geochemical balance of a small watershed and its geomorphic implications. GSA Bulletin 81: 3015-3032.
Craig LS, Palmer MA, Richardson DC, Filoso S, Bernhardt ES, Bledsoe BP, Doyle MW, Groffman PM, Hassett BA, Kaushal SS, Mayer PM, Smith SM, Wilcock PR (2008) Stream restoration strategies for reducing river nitrogen loads. Frontiers in Ecology and the Environment 6: 529-538.
Eshleman KN, Sabo RD (2016) Declining nitrate-N yields in the Upper Potomac River Basin: What is really driving progress under the Chesapeake Bay restoration? Atmospheric Environment 146: 280-289.
Findlay SEG, Kelly VR (2011) Emerging indirect and long-term road salt effects on ecosystems. Annals of the New York Academy of Sciences 1223: 58-68.
Frazar S, Gold AJ, Addy K, Moatar F, Birgand F, Schroth AW, Kellogg DQ, Pradhanang SM (2019) Contrasting behavior of nitrate and phosphate flux from high flow events on small agricultural and urban watersheds. Biogeochemistry 145: 141-160.
Gelhar LW, Wilson JL (1974) Ground-water quality modeling. Groundwater 12: 399-408.
Gold AJ, Deragon WR, Sullivan WM, Lemunyon JL (1990) Nitrate-nitrogen losses to groundwater from rural and suburban land uses. Journal of Soil and Water Conservation 45: 305-310.
Groffman PM, Band LE, Belt KT, Bettez ND, Bhaskar A, Doheny E, Duncan JM, Kaushal SS, Rosi EJ, Welty C, (2019) Applying the watershed approach to urban ecosystems. In: Pickett, S.T.A., Cadenasso, M.L., Grove, J.M., Irwin, E.G., Rosi, E.J., Swan, C.M. (Eds.), Science for the Sustainable City. Yale University Press, pp. 155-173.
Groffman PM, Rosi EJ, Martel L, Kaushal S (2018) Stream chemistry for core sites in Gwynns Falls: concentration of Cl, NO3, PO4, total N and P, SO4, dissolved oxygen, E. coli, plus temperature, pH, clarity, turbidity, isotopes, and pharmaceuticals ver 600. Environmental Data Initiative. https://doi.org/10.6073/pasta/9a7c95a4b7b12e7d97fc296b050ce8a5. Accessed 2020-05-13.
Hale RL, Scoggins M, Smucker NJ, Suchy A (2016) Effects of climate on the expression of the urban stream syndrome. Freshwater Science 35: 421-428.
Hirsch RM, Moyer DL, Archfield SA (2010) Weighted Regressions on Time, Discharge, and Season (WRTDS), with an Application to Chesapeake Bay River Inputs. JAWRA Journal of the American Water Resources Association 46: 857-880.
Hobbie SE, Finlay JC, Janke BD, Nidzgorski DA, Millet DB, Baker LA (2017) Contrasting nitrogen and phosphorus budgets in urban watersheds and implications for managing urban water pollution. Proceedings of the National Academy of Sciences 114: 4177-4182.
Kaushal SS, Groffman PM, Band LE, Elliott EM, Shields CA, Kendall C (2011) Tracking nonpoint source nitrogen pollution in human-impacted watersheds. Environmental Science & Technology 45: 8225-8232.
Kaushal SS, Groffman PM, Band LE, Shields CA, Morgan RP, Palmer MA, Belt KT, Swan CM, Findlay SEG, Fisher GT (2008) Interaction between urbanization and climate variability amplifies watershed nitrate export in Maryland. Environmental Science & Technology 42: 5872-5878.
Kaushal SS, Groffman PM, Likens GE, Belt KT, Stack WP, Kelly VR, Band LE, Fisher GT (2005) Increased salinization of fresh water in the northeastern United States. Proceedings of the National Academy of Sciences of the United States of America 102: 13517-13520.
Kaushal SS, Likens GE, Pace ML, Utz RM, Haq S, Gorman J, Grese M (2018) Freshwater salinization syndrome on a continental scale. Proceedings of the National Academy of Sciences 115: E574-E583.
Kelly VR, Findlay SE, Hamilton SK, Lovett GM, Weathers KC (2019) Seasonal and long-term dynamics in stream water sodium chloride concentrations and the effectiveness of road salt best management practices. Water, Air, & Soil Pollution 230.
Kochary S, Byl T, Noori B (2017) Modeling the movement of septic water chloride through a soil profile. Sustainability 9: 501.
Law LN, Band EL, Grove JM (2004) Nitrogen input from residential lawn care practices in suburban watersheds in Baltimore County, MD. Journal of Environmental Planning and Management 47: 737-755.
Likens GE, (2013) Biogeochemistry of a Forested Ecosystem, 3rd Edition. Springer-Verlag, New York.
Lovett GM, Likens GE, Buso DC, Driscoll CT, Bailey SW (2005) The biogeochemistry of chlorine at Hubbard Brook, New Hampshire, USA. Biogeochemistry 72: 191-232.
Moore J, Bird DL, Dobbis SK, Woodward G (2017) Nonpoint source contributions drive elevated major ion and dissolved inorganic carbon concentrations in urban watersheds. Environmental Science & Technology Letters 4: 198-204.
Nathan RJ, McMahon TA (1990) Evaluation of automated techniques for base flow and recession analyses. Water Resources Research 26: 1465-1473.
Oberhelman A, Peterson EW (2020) Chloride source delineation in an urban‐agricultural watershed: Deicing agents versus agricultural contributions. Hydrological Processes 34: 4017-4029.
Overbo A, Heger S, Gulliver J (2021) Evaluation of chloride contributions from major point and nonpoint sources in a northern U.S. state. Science of The Total Environment 764: 144179.
Reisinger AJ, Doody TR, Groffman PM, Kaushal SS, Rosi EJ (2019) Seeing the light: urban stream restoration affects stream metabolism and nitrate uptake via changes in canopy cover. Ecological Applications 29: e01941.
Reisinger AJ, Woytowitz E, Majcher E, Rosi EJ, Belt KT, Duncan JM, Kaushal SS, Groffman PM (2018) Changes in long-term water quality of Baltimore streams are associated with both gray and green-infrastructure Limnology and Oceanography 17: 1-17.
Rutledge EM, Teppen BJ, Mote CR, Wolf DC (1993) Designing septic tank filter fields based on effluent storage during climatic stress. Journal of Environmental Quality 22: 46-51.
Sabo RD, Clark CM, Bash J, Sobota D, Cooter E, Dobrowolski JP, Houlton BZ, Rea A, Schwede D, Morford SL, Compton JE (2019) Decadal shift in nitrogen inputs and fluxes across the contiguous United States: 2002–2012. Journal of Geophysical Research: Biogeosciences 124: 3104-3124.
Sabo RD, Scanga SE, Lawrence GB, Nelson DM, Eshleman KN, Zabala GA, Alinea AA, Schirmer CD (2016) Watershed-scale changes in terrestrial nitrogen cycling during a period of decreased atmospheric nitrate and sulfur deposition. Atmospheric Environment 146: 271-279.
Seybold E, Gold AJ, Inamdar SP, Adair C, Bowden WB, Vaughan MCH, Pradhanang SM, Addy K, Shanley JB, Vermilyea A, Levia DF, Wemple BC, Schroth AW (2019) Influence of land use and hydrologic variability on seasonal dissolved organic carbon and nitrate export: insights from a multi-year regional analysis for the northeastern USA. Biogeochemistry.
Wainger LA (2012) Opportunities for reducing total maximum daily load (TMDL) compliance costs: Lessons from the Chesapeake Bay. Environmental Science & Technology 46: 9256-9265.
Walsh CJ, Roy AH, Feminella JW, Cottingham PD, Groffman PM, Morgan RP (2005) The urban stream syndrome: current knowledge and the search for a cure. Journal of the North American Benthological Society 24: 706-723.
Wherry SA, Tesoriero AJ, Terziotti S (2021) Factors affecting nitrate concentrations in stream base flow. Environmental Science & Technology 55: 902-911.
Withers PJA, Jordan P, May L, Jarvie HP, Deal NE (2013) Do septic tank systems pose a hidden threat to water quality? Frontiers in Ecology and the Environment 12: 123-130.

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Reviews received at journal
29 Mar, 2022
Reviewers invited by journal
29 Mar, 2022
Editor invited by journal
13 Jan, 2022
Editor assigned by journal
13 Jan, 2022
First submitted to journal
31 Dec, 2021

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Long-term trends in nitrate and chloride in streams in an exurban watershed

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Version 1

Abstract

Figures

Synopsis

Introduction

Experimental Section

Results

Hydrology and climate

Chloride

Nitrate

Discussion

Chloride

Nitrate

Implications for Management

Declarations

Funding

Competing Interests

Availability of data and material

Code availability

Acknowledgments

References

Status:

Version 1