Atmospheric Deposition of Nitrogen Oxides onto the Landscape

Environ. Sci. Technol. 1997, 31, 1995-2004

Atmospheric Deposition of Nitrogen Oxides onto the Landscape Contributes to Coastal Eutrophication in the Northeast United States N. A. JAWORSKI,* R. W. HOWARTH, AND L. J. HETLING Ecology & Systematic, Cornell University, 202 Wordens Pond Road, Wakefield, Rhode Island 02879

Recently compiled data document a 3-8-fold increase in nitrate fluxes from 10 watersheds in the Northeast United States since the early 1900s. During this period, nitrogen oxide emissions from combustion sources have increased about 5-fold. For 17 large watersheds with relatively minor agricultural or urban influences, riverine nitrogen fluxes from 1990 to 1993 were highly correlated with atmospheric deposition onto their landscapes and also with nitrogen oxide emissions into their airsheds. These relationships provided two methods of estimating riverine nitrogen export directly from either deposition or emission fluxes. For 10 benchmark watersheds with good historical data, about 36-80% of the riverine total nitrogen export, with an average of 64%, was derived directly or indirectly from nitrogen oxide emissions. Atmospheric deposition of nitrogen represented only about 25% of the airshed emissions with the remaining 75% transported out of the airshed. Nitrogen is the element most responsible for eutrophication in coastal waters of this region. Our analysis suggests a strong linkage between the increase in cultural eutrophication of the coastal waters of the Northeast United States and the increase in nitrogen oxide emissions from fossil fuel combustion.

Introduction Eutrophication, caused by excessive inputs of nitrogen to rivers in the coastal zone, is one of the greatest factors altering water quality in estuarine ecosystems in the United States (1). Globally, human activity has accelerated nitrogen cycling and at least doubled the rate of nitrogen fixation over natural levels (2, 3). Several studies have documented increased nitrogen fluxes to estuaries, such as the Mississippi River mouth (4), Narragansett Bay (5), and the Baltic Sea (6). Recent comparisons of nitrogen fluxes from large regions around the North Atlantic Ocean suggested that human activity may have increased nitrogen fluxes to the coastal rivers of the Northeast United States by 5-14 times over natural rates (7). A recent study (8) showed that an increase in nitrogen loading not only caused a loss of terrestrial diversity but also increased the nitrogen export. At regional scales, a recent study (9) tested the hypotheses that riverine export of nitrogen increases as anthropogenic input of nitrogen increases and that net anthropogenic input onto the landscape is a good predictor of river export. Dramatic regional and global shifts in the * Corresponding author telephone: 401-783-7483; fax: 401-7823201; e-mail: [email protected].

S0013-936X(96)00803-6 CCC: $14.00

 1997 American Chemical Society

spatial movement of nitrogen in agriculture has been examined as a result of the rise in cash crop production and intensive animal production (10). In general, while we are reducing much of the uncertainty in estimates of the anthropogenic sources of nitrogen to the landscapes of watersheds, the quantitative linkages of the sources with the increased riverine fluxes remain poorly characterized. Human sewage, animal wastes, and runoff of fertilizer are usually assumed to dominate (4, 11, 12), although a significant contribution from atmospheric deposition has been suggested (13-15). Atmospheric releases and deposition of nitrogen have increased (3). This nitrogen deposition can contribute to downstream nitrogen fluxes in streams and rivers if the input exceeds the uptake needs of vegetation (16) or if the soils are particularly permeable and unable to hold on to the nitrogen (17). For the northeast U.S. region as a whole, atmospheric deposition of nitrogen onto the landscape, which originates mainly from fossil fuel combustion, exceeds all other individual nitrogen inputs, including fertilizer, leguminous crop fixation, and imported food and feeds (7). The regional scale analysis (9) reported that food and feed import was the largest source of nitrogen to the Northeast United States. The major objectives of this paper are to (i) compile and document historical riverine nitrate fluxes for 10 benchmark watersheds that represent the landscape of the northeastern United States; (ii) explore spatial response relationships of riverine export fluxes of nitrogen with atmospheric deposition of nitrogen and with airshed nitrogen emissions for watersheds with minor agricultural and urban influences; (iii) estimate current and historical riverine nitrogen export as a result of nitrogen oxide emissions for the 10 benchmark watersheds directly from deposition and emission fluxes; (iv) compare, on a total nitrogen bases, the landscape anthropogenic inputs, riverine exports and landscape consumption for the 10 benchmark watersheds; and (v) reconstruct the 1900-1993 total nitrogen loading fluxes from atmospheric deposition, wastewater discharges, and agricultural runoff to the coastal waters of the northeastern United States.

Study Area The Northeast United States study area encompasses the watersheds (north to south) of the Gulf of Maine including Massachusetts Bay, Buzzards Bay, Narragansett Bay, Long Island Sound, Hudson-Raritan Estuary, Delaware Bay, and Chesapeake Bay. The watersheds have a combined area of about 475 400 km2 with an estimated 1990 population of 55 million people. About 90% of the 22 billion L/day of municipal and industrial wastewater is discharged directly into tidal coastal waters. Ten large coastal watersheds, with sufficient historical water quality monitoring data, were selected as benchmarks (Table 1) to document the historical trends and to examine current nitrogen landscape loadings and riverine export fluxes. The ten watersheds have a drainage area of about 218,000 km2 and a wastewater discharge flow of about 3.5 billion L/day. These 10 watersheds, which include the Penobscot River flowing into the Gulf of Maine; the Merrimack into Massachusetts Bay; the Connecticut into Long Island Sound; the Delaware and Schuylkill into Delaware Bay; and the Susquehanna, Potomac, Rappahhanock, and James Rivers flowing into Chesapeake Bay, uniformly span the northeastern United States study area. The landscape of the 10 watersheds is about 70% forested, 25% agricultural, and 5% urban. Portions of the Potomac, Susquehanna, and Schuylkill Watersheds support intense crop and animal production resulting in large nutrient runoff fluxes.

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TABLE 1. Basic Data for 10 Benchmark Watersheds, North to South Array watershed Penobscot Merrimack Connecticut Hudson Delaware Schuylkill Susquehanna Potomac Rappahannock James average

USGS station no.

drainage area (km2)

airshed area (km2)

av river discharge (m3 s-2 km-2)

wastewater flows (L/day) (106)

1036390 1100000 1184000 1358000 1463500 1474500 1578310 1646580 1668000 2035000

20 109 12 005 25 019 20 953 17 560 4 903 67 314 29 940 4 134 16 206 21 814

155 500 70 300 86 400 270 680 265 400 170 400 251 500 313 300 41 600 518 600 214 368

0.02053 0.01987 0.02122 0.01723 0.01781 0.01503 0.01638 0.01076 0.01068 0.01415 0.01637

68 492 568 68 454 379 984 322 15 129

The Connecticut, Delaware, Merrimack, and Schuylkill Watersheds have the highest wastewater flows per unit area of watershed.

Data Sources and Methods As part of a regional analysis of nitrogen inputs to the North Atlantic Ocean, data previously were compiled to estimate riverine nutrient and mineral fluxes for 33 major watershed monitoring sites between Maine and Virginia for the period 1990-1993 (7, 18; Jaworski et al., in preparation). Riverine nutrient and mineral export fluxes were estimated by multiplying the average annual concentration of a specific nutrient or mineral by the mean annual river discharge. Except for nitrates, these 4-year average annual estimates were usually within (5% of the estimates in the Chesapeake Bay watersheds that were conducted by the U.S. Geological Survey using a load estimation model (19). While the 4-year average annual estimates for nitrates were usually 5-15% lower than the flux predicted by the USGS model, the estimates were well within annual variations, which were often a factor of 2-3. Nitrate and stream flow data used to calculate the nitrate fluxes for 10 benchmark watersheds (Figures 1a and 2b) come from drinking water monitoring programs (unpublished reports from municipal drinking water authorities), from the Massachusetts State Health Agency monitoring program from 1890 to 1915, and from the published U.S. Geological Survey Water Resource Data annual reports. Both river and drinking water sampling data are included and cover most of this century. The annual summary data reflect mainly monthly sampling; however, some of the more recent sampling has been reduced to bimonthly. To assess the quality of the early data, comparisons were made with aperiodic water quality data obtained by the U.S. Geological Survey (20) between 1905 and 1920 and by the Massachusetts State Health Agency from 1890 to 1915. The recent drinking water data were also compared with recent U.S. Geological Survey data. In general, there was very good agreement where data sets overlapped. For continuous data sets, such as for the Potomac Watershed, the nitrate, chloride, and sulfate concentrations increased steadily over time, further confirming that the increases in chemical concentration observed were not due to major changes in analytical methodology. Large changes in concentration occurred as river discharge increased or, for example, during the 19601970 drought. Chlorides and sulfates, being conservative, went up in concentration while nitrates, being nonconservative, went down in concentration during the drought period. For the ten benchmark watersheds, historical nitrate fluxes (Figures 1a and 2b) were calculated from the nitrate concentration and the historical river discharge data. The current inorganic, organic, and total nitrogen (TN), historical nitrate export fluxes, which were extrapolated to a common historical

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forest (%)

land use agriculture (%)

urban (%)

62 91 78 72 85 48 64 59 54 75 69

21 5 12 23 12 42 32 37 44 23 25

2 4 10 5 3 10 4 4 2 2 5

base year (1900), and total organic carbon (TOC) fluxes are presented in Table 2. In our review of the water quality data from a large number of monitoring sites in the study area, some of these riverine systems had significantly elevated nutrient levels and thus were clearly impacted by large wastewater discharges and/or by extensive agricultural fertilizer and animal waste runoff. To determine whether riverine nitrogen exports were related to atmospheric deposition of nitrogen, watersheds were screened to eliminate those that were heavily affected by wastewater discharges, and agricultural runoff were identified. A total of 55 monitoring sites, for which there were nutrient flux estimates for the 1990-1993 period, were screened. Elevated phosphorus and potassium levels were used as indicators of significant increases in nutrients from wastewater discharges and/or agriculture runoff. High levels of ammonium, which is the unoxdized form of nitrogen, indicate the presence of significant wastewater and/or agriculture runoff that has not been nitrified. Sites that were excluded were those that had phosphorus concentrations greater than 0.075 P mg L-1, potassium concentrations above 2.0 K mg L-1, or ammonium nitrogen concentrations greater than 0.06 N mg L-1. Watersheds that had large impoundments on their main stems were also excluded since such impoundments will lower riverine nitrogen export fluxes by encouraging denitrification (7). This screening left 17 monitoring sites with low phosphorus and potassium concentrations and for which urban wastewater discharges and agricultural runoff should be relatively small. Low to moderate application rates of commercial fertilizer and animal waste in the 17 watersheds, as obtained from a national watershed-based analysis (12), further suggest that the agricultural practices would not be the major source of nitrogen in these watersheds and thus have only a minor affect on the export fluxes of nutrients. While it would be ideal to have all 17 watersheds with the same percent of the landscape in agricultural production so that land use would not be an additional unknown variable, the intensity of crop and animal production and not the percent of agricultural land appears to be the most important factor. The final selected, minimally impacted by urban wastewater and agricultural runoff, 17 monitoring sites had rather large watersheds ranging in size from 120 to 20 109 km2, with an average area of 6286 km2 (Table 3). About 74% of the landscape of the 17 watersheds was forested, 22% was in agriculture, and 4% was urban. While the 10 benchmark watersheds had a slightly higher percent agricultural landscape at 25%, the average riverine TN export for the 10 was significantly higher at 766 kg of N km-2 yr-1 (Table 2) as compared to 422 kg of N km-2 yr-1 for the 17 watersheds (Table 4). Small sub-basins in the 10 benchmark watersheds with intense agricultural practices often had riverine TN export fluxes of 2000-3000 kg of N km-2 yr-1.

FIGURE 1. Rivine nitrate flux trends for the five southern benchmark watersheds (a, top) and the five northern benchmark watersheds (b, bottom).

FIGURE 2. Nitrogen oxide emissions trends for the 10 benchmark watersheds. The first five watersheds in Table 3 (St John, St Criox, Penobscot, Kennebec, and Saco) flow into the Gulf of Maine.

The Pawcatuck, Quinebaug, and Upper Housatonic Watersheds are in the Long Island Sound Basin. The Upper Hudson


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TABLE 2. Riverine Nitrogen and TOC Flux Data for 10 Benchmark Watersheds, North to South Array

watershed

1900 estd riverine NO3-N flux (kg of N km-2 yr-1)

1990-1993 riverine NO3-N flux (kg of N km-2 yr-1)

increase of NO3-N flux since 1900 (%)

1990-1993 riverine NH3-N flux (kg of N km-2 yr-1)

1990-1993 riverine organic N flux (kg of N km-2 yr-1)

1990-1993 riverine TN flux (kg of N km-2 yr-1)

1990-1993 riverine TOC flux (kg of C km-2 yr-1)

Penobscot Merrimack Connecticut Hudson Delaware Schuylkill Susquehanna Potomac Rappahannock James average

15 30 25 50 120 275 135 75 40 25 79

60 201 237 316 672 1443 825 724 240 133 485

298 569 847 532 460 425 511 866 500 431 544

28 91 88 46 29 72 48 20 20 9 45

188 247 221 270 294 294 202 181 240 221 236

276 539 546 632 996 1808 1075 925 500 362 766

7400 3410 2450 2710 1400 2610 2050 2675 2170 2180 2906

TABLE 3. Basic Data for 17 Watersheds Minimally Impacted by Urban Sewage and Agricultural Runoff, North to South Array watershed

USGS station no.

drainage area (km2)

river discharge (m3 s-2 km-2)

forest (%)

land use agriculture (%)

urban (%)

St. John St. Croix Penobscot Kennebec Saco Pawcatuck Quinebaug Upper Housatonic Upper Hudson WB Susuquehanna Y. Woman Creek Maurice Raystown SB Potomac Pamunkey Mattaponi Appomattox average

1015000 1021050 1036390 1049265 1066000 1119040 1127000 1198125 1335770 1553500 1545600 1411800 1562000 1608500 1673000 1674500 2041650

14 672 3 559 20 109 13 994 8 451 764 401 1 205 11 966 17 734 120 290 1 958 3 807 2 800 1 557 3 481 6 286

0.01893 0.01959 0.02053 0.01723 0.02210 0.02190 0.01995 0.02020 0.02116 0.01794 0.01920 0.01356 0.01269 0.01074 0.00947 0.00898 0.00904 0.01666

93 51 62 80 69 60 76 74 82 83 95 75 77 81 63 70 64 74

5 46 21 18 28 28 14 20 15 15 4 22 20 15 35 28 34 22

2 4 2 2 2 3 10 6 3 2 1 3 3 4 2 2 2 3

TABLE 4. 1990-1993 Nitrogen Data for 17 Watersheds Minimally Impacted by Urban Sewage and Agricultural Runoff, North to South Array

watershed

river NO3 as N (kg of N km-2 yr-1)

river inorganic N (kg of N km-2 yr-1)

river organic N (kg of N km-2 yr-1)

river total N (kg of N km-2 yr-1)

wet inorganic N deposition (kg of N km-2 yr-1)

nitrogen oxide emissions (kg of N km-2 yr-1)

St. John St. Croix Penobscot Kennebec Saco Pawcatuck Quinebaug Upper Housatonic Upper Hudson WB Susuquehanna Y. Woman Creek Maurice Raystown SB Potomac Pamunkey Mattaponi Appomattox average

93 56 60 73 82 261 257 281 330 402 309 620 675 213 73 47 51 228

110 83 88 90 98 289 301 305 380 426 357 635 684 221 88 61 67 251

163 202 188 145 166 225 309 185 143 245 143 154 170 93 118 115 111 169

273 285 276 235 264 514 610 441 522 671 520 788 854 313 206 175 178 422

340 345 350 350 360 500 540 580 620 670 670 710 720 550 450 420 400 504

700 800 852 860 1100 1700 2100 2100 1800 2600 2600 2700 2700 2100 1300 1300 1000 1665

is the forested portion of the Hudson Watershed. The West Branch, Young Woman Creek, and Raystown are sub-basins of the Susquehanna. The Maurice flows into Delaware Bay. The last four in Table 3 are part of the lower Chesapeake Bay Basin. Average annual riverine nitrate, inorganic, organic, and total nitrogen fluxes during 1990-1993 for these 17 watersheds

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are presented in Table 4, with the watersheds listed from north to south along the coast. Export fluxes compared very favorably with the recently completed water quality fluxes by the USGS for the Chesapeake Bay (21), the Hudson River (22), and the Connecticut, Housatonic, and Thames River Basins (23). Estimates for atmospheric deposition of wet inorganic nitrogen for the 1990-1993 period were obtained

FIGURE 3. Riverine total N (upper line) and inorganic N (lower line) flux versus wet inorganic N deposition for 17 selected watersheds for the 1990-1993 period.

FIGURE 4. Wet and dry inorganic N deposition (upper line) and riverine inorganic N flux (lower line) versus nitrogen oxide emissions as N for 17 selected watersheds for the 1990-1993 period. from the National Atmospheric Deposition Program (24), and for two sites in Connecticut from the Long Island Atmospheric Project (25). Wet inorganic nitrogen included both the nitrate and ammonium ions. Wet inorganic nitrogen deposition estimates (Table 4) were used since dry deposition data are not available for all sites. Data from Connecticut (25) and analysis of the Chesapeake Bay region deposition data (26) suggest that dry deposition was about 30-40% of the wet deposition for the study area. We increased the wet inorganic deposition by 40% to account for dry deposition. While data on atmospheric nitrogen deposition are only available since 1978, historical changes in deposition were estimated from estimates of emissions of nitrogen oxides to the atmosphere. Historical emission estimates by state from the period 1900-1994 were obtained from the U.S. EPA (27). For the 10 benchmark watersheds of the study area, historical average annual emissions in kg of N km-2 yr-1 were normalized for a watershed’s airshed by dividing the summation of the emissions for the states in or adjacent to the watershed by the area of the states. The historical nitrogen oxide emissions for 10 benchmark watersheds are presented in Figure 2 and for the 1990-1993 emissions into the airsheds of the 17 selected watersheds in Table 4. This airshed normalization procedure is similar to the application of a regional airshed

acid deposition model, such as for the Chesapeake Bay Watersheds (28). Linear regression scatter plots were used to quantify the riverine/deposition flux (Figure 3) and deposition/emissions and riverine/emissions flux (Figure 4) response relationships for the 17 watersheds for the 1990-1993 period (Table 4). For the 10 watersheds, the amounts of riverine total and inorganic nitrogen coming from the airshed emissions (Table 5) were estimated using the riverine/deposition and riverine/emissions flux response relationships in Figure 3 and 4, respectively. For the period 1990-1993, the anthropogenic TN landscape inputs, riverine export, and amount of nitrogen “consumed” (inputs - export) within the landscape and riverine ecosystems for the 10 watersheds were estimated (Table 6). The fertilizer application data were extrapolations of the Tennessee Valley Authority annual Market Reports on the Consumption of Primary Plant Nutrients (12). About 1000 kg of N km-2 yr-1 in feed and food was imported into the study area (9). The amount of imported feed and food for each of the 10 benchmark watersheds was assumed to be proportional to the amount of animal waste generated in each watershed (12). The amount of biological fixation by agricultural biota, estimates ranged from 830 (9) to 900 kg of


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TABLE 5. Estimated Riverine Inorganic and TN Export Fluxes from Air Emissions and Atmospheric Deposition for 10 Benchmark Watersheds, North to South Array

watershed Penobscot Merrimack Connecticut Hudson Delaware Schuylkill Susquehanna Potomac Rappahannock James average

airshed N riverine estd river wet inorganic riverine estd river emissions inorganic inorg. N from inorganic N atmos total N total TN from (kg of N flux (kg of emissions (kg of N from deposition (kg of flux (kg of deposition (kg of TN from N km-2 yr-1) N km-2 yr-1) N km-2 yr-1) emissions (%) N km-2 yr-1) N km-2 yr-1) N km-2 yr-1) deposition (%) 850 1600 1800 1600 2100 2650 2550 2200 1548 1150 1805

88 292 325 362 701 1515 873 744 260 142 530

53 226 273 226 344 473 450 381 224 119 277

61 77 84 62 49 31 52 51 86 84 64

350 490 515 510 520 670 670 540 450 410 513

276 539 546 632 996 1808 1075 925 500 362 766

209 399 433 427 440 644 644 467 345 291 430

76 74 79 68 44 36 60 51 69 80 64

TABLE 6. 1990-1993 TN Data of Landscape Anthropogenic Inputs, Riverine Export, and Landscape Consumed luxes for 10 Benchmark Watersheds


TN TN TN TN biotic TN atmos total input of all TN measd TN consumed wastewater fertilizer feed & food fixation deposition TN sources of river export on landscape TN discharges use (kg of N import (kg of N (kg of N (kg of N landscape (kg of (kg of N (kg of N consumed on (kg of N km-2 yr-1) km-2 yr-1) km-2 yr-1) km-2 yr-1) N km-2 yr-1) km-2 yr-1) m-2 yr-1) landscape (%) km-2 yr-1) 140 350 300 730 800 1500 710 1200 1000 930 766

250 500 500 1000 1000 1500 1500 1500 1000 1000 975

154 385 330 803 880 1650 781 1320 1100 1023 843

486 688 756 714 728 938 938 756 729 707 744

1030 1923 1886 3247 3408 5588 3929 4776 3829 3660 3328

276 539 546 632 996 1808 1075 925 500 362 766

754 1384 1340 2615 2413 3780 2855 3851 3329 3298 2562

73 72 71 81 71 68 73 81 87 90 77

24 137 123 18 187 614 105 56 19 5 129

TABLE 7. 1990-1993 Sources of Riverine Export for 10 Benchmark Watersheds


total input of all TN agric. estd river TN TN measd TN wastewater estd river TN from estd river TN sources on landscape from deposition river export discharges all agric sources from all agric (kg of N km-2 yr-1) (kg of N km-2 yr-1) (kg of N km-2 yr-1) (kg of N km-2 yr-1) (kg of N km-2 yr-1) input sources (%) 544 1235 1130 2533 2680 5150 2991 4470 3100 2953 2679

209 399 433 427 440 644 644 467 345 291 430

N km-2 yr-1 (7) for the Northeast. We used the higher flux and assumed, for this study, and that the fixation of each of the 10 watersheds was proportional to its fertilizer use. The loadings from wastewater discharges are based on current data we have obtained as part of our overall research effort (Jaworski et al., in preparation). By difference, i.e., the wastewater and atmospheric contributions were subtracted form the riverine export with the remainder being from all the agricultural sources, we estimated the total amount that all agricultural sources contributed to the TN riverine export for the 10 benchmark watersheds (Table 7). Finally, an estimate of the 1900-1994 total nitrogen loading fluxes to the coastal waters of the study area (Figure 5) was reconstructed reflecting from atmospheric deposition, wastewater discharges, and agricultural runoff. The reconstruction is based on loading estimates using the above riverine export

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276 539 546 632 996 1808 1075 925 500 362 766

24 137 98 18 187 614 105 56 19 5 126

43 3 15 187 368 550 325 401 136 67 209

7.8 0.2 1.3 7.4 13.7 10.7 10.9 9.0 4.4 2.3 6.8

flux/emission response relationship, fertilizer and imported feed data as indicated earlier with export rates from the Chesapeake Bay watershed model (28), and wastewater loadings from Jaworski et al. (in preparation).

Discussion of Results The temporal plots of nitrate fluxes for the 10 benchmark watersheds (Figures 1a and 2b) suggest that the maximum riverine fluxes occurred during the 1970-1980 period, corresponding to the maximum period of increase of nitrogen oxides emissions (Figure 2). The riverine nitrate fluxes also appear to have leveled off and decreased during the past decade. However, these trends were greatly influenced by the annual river discharge flows, and the nitrate fluxes were especially low during the 1960-1970 drought as can be seen

FIGURE 5. Estimates of total nitrogen loading trends per unit of watershed drainage area to the northeastern U.S. coastal waters by sources. in Figure 1a,b. The use of nitrogen fertilizer also reached a maximum in the period from 1980 to 1985 (10, 12). As summarized in Table 2, there was a 3-8-fold increase in nitrate fluxes since the early 1900s for the 10 benchmark watersheds. The average nitrate riverine flux for the 10 watersheds has increased from about 80 to 480 kg of N km-2 yr-1. The Susquehanna, Potomac, Schuylkill, and Delaware Rivers had the largest 1990-1993 riverine nitrates fluxes and TN fluxes greater than 925 kg of N km-2 yr-1 (Table 2). The average ammonia nitrogen fluxes were, on the average, 45 kg of N km-2 yr-1 , or about 10% of the nitrate fluxes. The organic nitrogen fluxes were surprisingly similar with an average rate for the 10 watersheds of 236 kg of N km-2 yr-1 and about 67 kg of N km-2 yr-1 higher than the organic nitrogen fluxes for the 17 selected watersheds (Table 4). Airshed emissions, atmospheric deposition of nitrogen and riverine export of nitrogen were all higher in watersheds in the middle of the study area, i.e. in the Pawcatuck, Quinebaug, Upper Housatonic, Upper Hudson, WB Susquehanna, Young Woman Creek, Maurice, and Raystown Watersheds, and lower in the northern watersheds in Maine and lower in the southern Chesapeake Bay Watersheds (Table 4). The riverine flux of inorganic nitrogen in the 17 watersheds was highly correlated with the wet inorganic nitrogen deposition (Figure 3). Interestingly, the intercept for the riverine/deposition response relationships, corresponding to a riverine inorganic nitrogen flux approaching zero, was 315 kg of N km-2 yr-1. This suggests that the landscape has a capacity to retain inorganic nitrogen at low levels of atmospheric deposition, providing strong support for the saturation hypothesis (16) which states that forests have large downstream losses of nitrate only when nitrogen inputs are so high as to saturate biological uptake. Nitrogen saturation (30) has been reported to be occurring throughout high elevation catchments of the Colorado Front Range at annual inorganic wet deposition rates of about 400 kg of N km-2 yr-1. While no evidence of nitrogen saturation occurring in the forests of the Chesapeake Bay Watersheds has been found (31), trend information and seasonal patterns suggest Catskill Watersheds in New York may be approaching nitrogen saturation (32). Experimental inducement of nitrogen saturation at the state of Maine’s West Bear Brook Watershed suggests that saturation may be induced at rates of N deposition lower than previously believed (33). At levels of deposition greater than 315 kg of N km-2 yr-1 onto the 17 watersheds, the biological uptake processes appeared to be saturated with inorganic nitrogen. The

riverine inorganic nitrogen export flux increased in a linear fashion with increased deposition (Figure 3, lower trendline). Annual deposition fluxes of 315 kg of N km-2 yr-1 and greater can be viewed, in part as biological, as the deposition saturation flux during the growing periods for the watersheds of the study area. The annual deposition flux can also be physical, due to nitrogen-rich sub-surface runoff during dormant periods and/or nitrogen-rich direct surface flows such as during high flow and/or snowmelt periods. The second view is supported by a study of a watershed in the Susquehanna Basin in which temporal variations in the nitrate, sulfate, and chloride concentrations were primarily hydrologically controlled (34). The slope of the response regression is approximately 1.36, suggesting that the inorganic nitrogen riverine export flux was about 40% larger than the wet inorganic nitrogen atmospheric deposition. While most of this increase was likely a result of dry deposition, which in Connecticut (25) was about 40% of the wet deposition, some of the increase could reflect other anthropogenic sources. The total nitrogen riverine export flux/deposition response relationship (Figure 3, upper trendline) had a slope of 1.4 and a nitrogen deposition saturation flux of approximately 196 kg of N km-2 yr-1. The fact that the slopes of the two are comparable suggests that the riverine organic nitrogen export fluxes were similar for the 17 watersheds in the study area, about 170 kg of N km-2 yr-1. This suggests that the amount of biological processing of the landscape inorganic nitrogen into the organic form in the riverine export was about the same for all 17 watersheds. The watersheds with higher inorganic nitrogen deposition and resulting higher riverine inorganic fluxes did not result in proportionally higher riverine organic fluxes. Deposition/emissions and riverine/emissions flux response relationships for the 17 selected watersheds for the 1990-1993 period are presented in Figure 4. The upper trendline (Figure 4) shows a high degree of correlation between wet and dry deposition and emissions. The slope of the trendline (0.25) suggests that about 25% of the emissions were returned to the watershed as wet and dry inorganic deposition. From the slope of the riverine trendline (Figure 4), about 24% of the emissions in the watershed’s airshed showed up as inorganic nitrogen riverine export fluxes. This also suggests that a large portion, about 75% of the air emissions of nitrogen oxides, were transported out of the airsheds and had no impact on the riverine export. About 50% of emissions in a source apportionment study of nitrogen species was accounted for


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in atmospheric deposition for an air quality basin (35) in southern California. Extending the upper trendine (Figure 4) to a point where the emissions equal zero provides an estimate of nitrogen deposition at zero emissions. The zero deposition-emission flux was 265 kg of N km-2 yr-1 and was probably from a combination of combustion sources outside the airshed area and from non-combustion sources, such as ammonia volatilization of animal waste, within the study area. The zero deposition-emission flux was below saturation flux, previously estimated to be 315 kg of N km-2 yr-1. Thus for these watershed conditions, there would be minimal inorganic nitrogen leakage due only to other sources of atmospheric deposition of nitrogen. The lower trendline in the scatter plot (Figure 4) presents the landscape response of inorganic nitrogen riverine flux to emissions into the airsheds of the 17 watersheds. The riverine nitrogen export versus emissions flux response relationship has the same two interesting features as in Figure 3, i.e., (i) a watershed nitrogen emission saturation flux and (ii) a linear nitrogen riverine export versus nitrogen emission flux response relationship. For the 17 watersheds, the nitrogen emission saturation flux (lower trendline in Figure 4) was about 645 kg of N km-2 yr-1, corresponding to a wet and dry inorganic nitrogen deposition flux of about 475 kg of N km-2 yr-1 (upper trendline in Figure 4) or a wet inorganic nitrogen deposition flux of about 350 kg of N km-2 yr-1. The nitrogen deposition saturation flux as previously quantified from Figure 3 data was similar at 315 kg of N km-2 yr-1. Using two independently determined measurements of combustion sources, i.e., (1) deposition and (2) emissions, yielded nitrogen saturation values that ranged from 315 to 350 kg of N km-2 yr-1. For the 1990-1993 period, emissions in the Susquehanna and Schuylkill Watersheds in the middle of the study area had emission fluxes (Figure 4) about four times greater than the emission saturation flux. The emissions for the James River in the south and for the Penobscot River in the north were 1.8 and 1.2 times, respectively, greater than the saturation flux. Two estimates were made of the riverine export: (1) inorganic nitrogen fluxes from air emissions and (2) total nitrogen fluxes form atmospheric deposition for the 10 benchmark watersheds (Table 5). The inorganic estimates were based on the inorganic riverine/emissions response relationship (lower trendline) in Figure 4, while TN estimates were based on the riverine/deposition relationship (upper trendline) in Figure 3. For the 10 benchmark watersheds, the estimated amount of the inorganic riverine nitrogen that originated from air emissions ranged from 31% to 86% for the 1990-1993 period (Table 5). The estimated amount of the riverine total nitrogen that originated from atmospheric deposition was similar and ranged from 36% to 80%. For the 10 benchmark watersheds, the average inorganic nitrogen riverine flux from emissions ranged from 53 to 473 kg of N km-2 yr-1 with an average of 277 kg of N km-2 yr-1 or 64% of the riverine inorganic nitrogen flux. The average total nitrogen riverine flux from wet inorganic atmospheric deposition from ranged from 209 to 644 kg of N km-2 yr-1 with an average of 430 kg of N km-2 yr-1 or 64% of the riverine TN flux. Most of increase in riverine TN export from fossil fuel combustion was as nitrates except for watersheds like the James in the south and the Penobscot in the north, for which more than 60% of the 1990-1993 TN export was organic nitrogen. A study of the atmospheric input of inorganic nitrogen to the western Mediterranean suggested that the riverine export was about 65% of the atmospheric input (36). Not totally resolved is the amount of deposition nitrogen that flows directly into the rivers via overland surface flows, which is controlled hydrologically, or the amount transported in subsurface flows, which is controlled hydrologically and

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moderated biologically. Using the Potomac as an example, about 50% of the river discharge is via overland surface flow. Therefore, if we assume 50% of the nitrogen deposition will be transported with the overland flow, about 380 kg of N km-2 yr-1 of the 756 kg of N km-2 yr-1 of TN deposition (Table 6) was directly exported out of the watershed via overland surface flow. The 380 kg of N km-2 yr-1 in overland surface flow is about 81% of the 467 kg of N km-2 yr-1 of TN estimated contribution from atmospheric deposition (Table 6) with remaining 87 kg of N km-2 yr-1 transported via subsurface flow. The 380 kg of N km-2 yr-1 could be viewed as a physical or hydrologic nitrogen saturation flux. This estimate of the amount of TN in overland flows strongly supports our predictions based on both deposition and emission fluxes. In Table 6, we compared, for the 10 benchmark watersheds, the four major sources of anthropogenic nitrogen landscape inputs: fertilizer, food and feed import, agricultural biotic fixation, and atmospheric deposition. In this comparison, we assumed that the TN in wastewater comes from a combination of (i) crop or animal production, (ii) from food and feed import, and (iii) biological fixation and that 100% of the TN from wastewater discharges was exported out of the watersheds via the rivers. Therefore, wastewater was not included as an anthropogenic source. Landscape “consumption” equals the sum of inputs; fertilizer use, feed and food import, fixation, and atmospheric deposition minus the outputs, which is riverine export. Of the annual average riverine export of 766 kg of N km-2 yr-1 (Table 6), wastewater discharges contributed 126 kg of N km-2 yr-1 (Table 7). For the entire study, about 280 kg of N km-2 yr-1 (Figure 5) of wastewater nitrogen was exported out of the watersheds via the rivers or in direct discharges into tidal waters, thus removing about 10% of 2679 kg of N km-2 yr-1 of the sum total of agricultural input (Table 7). The average landscape input was about 3330 kg of N km-2 yr-1 with the largest source being food and feed import, about 975 kg of N km-2 yr-1. It has been estimated that 10-40% of the animal feed nitrogen is released by volatilization of ammonia (37). With dairy cattle waste, about 40-50% is loss due to volatilization (38). Using 50%, we estimated that about 485 kg of N km-2 yr-1 of air emissions would occur in response to the volatilization of ammonia of animal waste resulting in 120 kg of N km-2 yr-1 of atmospheric deposition or about 16% of the average annual inorganic deposition for the 10 benchmark watersheds. This assumes that 25% of the emissions are returned, similar to combustion emissions (Figure 4). The highest landscape input fluxes, over 3900 kg of N km-2 -1 yr , of TN were from intense agriculture practices and elevated deposition in the Potomac, Susquehanna, and Schuylkill Watersheds (Table 6). For the Potomac (Table 7), if we subtract from the 925 kg of N km-2 yr-1 of riverine export, 467 kg of N km-2 yr-1 of deposition, and 56 kg of N km-2 yr-1 of wastewater, 401 kg of N km-2 yr-1 or about 15% of the 3600 kg of N km-2 yr-1 (Table 6) TN inputs from fertilizer and food and feed import left the watershed via the river. If we include fixation, about 11% of the three agricultural inputs were exported out of the watershed via the river. The 9-15% export rates are very similar to the nitrogen fertilizer export coefficients used in the Chesapeake Watersheds model (29). If 50% of the animal waste is volatilized, only about 7% of import animal feed was exported out of the watershed via the river. This suggests that biotically fixed nitrogen would be less than 7%. While there appears to be large uncertainties in how much of each of the three agricultural sources leaves the watersheds via the river, the average aggregate for all three agricultural sources for large basins like the 10 benchmarks watersheds was about 209 kg of N km-2 yr-1 or about 7% of the agricultural inputs (Table 7). For the five southern watersheds from the Schuylkill through the James, over 2800 kg of N km-2 yr-1 of TN inputs

was consumed; i.e., (i) converted to gaseous forms through denitrification, (ii) volatilized as ammonia and/or (iii) stored in biomass, groundwaters or soils of the landscape. In the five northern-most watersheds, the consumption fluxes were dramatically lower with a flux of about 754 kg of N km-2 yr-1 for the Penobscot. The open question remaining is the fate of anthropogentic TN consumed on the landscape, although the large-scale regional analysis suggests that denitrification dominates (7). The southern watersheds had the highest per cent of TN consumed while the percent consumed was the lowest in the northern watersheds. Over 87% was consumed in the James and Rappahannock with about 72% for the Penobscot, Merrimack, and Connecticut. While riverine nitrogen export fluxes can be reduced in a specific watershed with many large reservoirs (7), abundant riparian buffer strips (39), and numerous riparian wetlands (40), hydrologic conditions (34) including temperature (41) may also be very important. The annual historical TN export fluxes for the Potomac have often varied by a factor of 2-3, mainly as a result stream flow conditions. As can be seen in comparing the deposition and riverine data of the southern watersheds to the northern watersheds in Table 4, the Pamunkey, Mattaponi, and Appomattox had higher deposition than the St John, St Croix, Penobscot, Kennebec, and Saco; yet the three southern watersheds had significantly lower riverine export fluxes. A longer period of frozen landscapes and a shorter growing season may be a major factor for the higher riverine exports fluxes from the St John, St Croix, Penobscot, Kennebec, and Saco watersheds. The northern watersheds had about twice the river discharge fluxes, about 0.019 m3 s-2 km-2, as compared to southern watersheds with fluxes of 0.009 m3 s-2 km-2 (Table 3). The watersheds in Maine also had much higher TOC fluxes of about 7000 kg of N km-2 yr-1 as compared to about 2100 kg of N km-2 yr-1 for the southern watersheds (Jaworski et al., in preparation). When we examine the data for the seven watersheds in the lower left corner of the lower trendline (Figure 3), it appears that the watersheds are in two clusters, the four to the extreme lower left are in Maine and the next three are in Virginia. The Maine watersheds had nitrogen saturation fluxes of less than 300 kg of N km-2 yr-1 while the Virginia watersheds had about 350 kg of N km-2 yr-1. The Maine watersheds with lower saturation fluxes had greater proportions of the atmospheric deposition exported from its landscape. The data suggest that the northern watersheds, with the higher riverine flow fluxes resulting in shorter watershed hydraulic retention times, have both lower per cent of TN consumed and lower nitrogen saturation fluxes, and therefore less nutrients were retained in the watersheds. It appears that the saturation fluxes in the southern watersheds are influenced more biologically while northern are more physically influenced. Figure 5 shows a reconstruction of the 1900-1994 total nitrogen loading fluxes from atmospheric deposition, wastewater, and agriculture runoff to the coastal waters of the Northeast United States. The estimated combined total nitrogen flux for all three sources has increased from about 200 to just over 1000 kg of N km-2 yr-1 during the period from 1900 to 1994. The largest source of anthropogenic nitrogen loadings into the coastal waters of the northeastern United States was the deposition of the nitrogen oxide emissions to the airsheds of the coastal watersheds.

Summary The nitrogen fluxes of the landscape of the watersheds of the 475 400 km2 of the Northeast United States, which are mosaics of hydrologically connected riverine and terrestrial ecosystems, have been greatly influenced by the nitrogen oxide emissions from fossil fuel combustion into their airsheds. A strong linkage between the increase in cultural eutrophication of the coastal waters of the northeastern United States and the increase in nitrogen oxide emissions has been established.

Since the early 1900s, there has been a 3-8-fold increase in nitrate fluxes from 10 benchmark watersheds. During the 1900-1993 period, nitrogen oxide emissions have increased about 5-fold. For watersheds minimally impacted by human sewage and agricultural runoff, riverine nitrogen fluxes were highly correlated with atmospheric deposition onto their landscapes and also with nitrogen oxide emissions into their airsheds. A major finding of the 1990-1993 spatial riverine export flux/ deposition response relationship for the 17 watersheds was that there was a certain atmospheric deposition flux required to satisfy the landscape nitrogen deposition saturation capacity, about 315 kg of N km-2 yr-1, before there was any significant leakage of inorganic nitrogen. Once the nitrogen deposition saturation flux was exceeded, the watershed inorganic nitrogen export flux was about equal to the wet plus dry atmospheric deposition flux. The spatial riverine flux/emission response relationship for the 17 watersheds has also shown a similar landscape emission saturation capacity flux before an individual watershed has any measurable inorganic nitrogen leakage. Likewise, the riverine export flux/emission response relationship was also linear. This emission saturation capacity flux translated into a deposition saturation flux of 350 kg of N km-2 yr-1. The two independently determined measurements, i.e., (1) deposition and (2) emissions, have yielded nitrogen saturation fluxes ranging from 315 to 350 kg of N km-2 yr-1. This second major finding further indicates that only about 25% of the emission flux within the immediate airshed was accounted for in the wet and dry inorganic nitrogen atmospheric deposition flux. For the 17 mosaic landscapes, about 25% of the airshed emissions were exported from the watershed into the coastal waters of the study area once the emission saturation capacity flux was exceeded. This suggests that about 75% of the air emission of nitrogen oxides was transported out of the airshed. The estimated wet and dry deposition from a combination of combustion sources outside the airshed area and from natural and/or noncombustion source, was 265 kg of N km-2 yr-1 or about 38% of the average deposition for the study area for the 1990-1993 period. For the 10 benchmark watersheds, about 277 kg of N km-2 yr-1 or 64% of the inorganic riverine flux was attributed to emissions of nitrogen oxides from combustion sources. Likewise, about 64% of the total riverine flux was attributed to atmospheric deposition of nitrogen oxides from combustion sources. Of the 1100 kg of N km-2 yr-1 of 1990-1993 total nitrogen loading fluxes from the three major sources, about 43% was from the atmosphere. It appears that the northern watersheds, with the higher riverine flow fluxes resulting in shorter watershed hydraulic retention times, have both lower percent of TN consumed within the watershed and lower nitrogen saturation fluxes, and therefore less TN and TOC were retained in the northern than in the southern watersheds. Moreover, it appears that the saturation fluxes in the southern watersheds were more biologically moderated while in the northern watersheds the fluxes were more hydrologically or physically influenced. During the period from 1900 to 1994, the total nitrogen loading of the coastal waters of the northeastern United States has increased over 5-fold, from about 200 to over 1000 kg of N km-2 yr-1, with the largest source of nitrogen from atmospheric deposition. These analyses aid in reducing the “scientific uncertainty” (42) in estimating how much riverine nitrogen comes from nitrogen oxide emissions from combustion sources. The nitrogen deposition or emission saturation capacity flux should be useful in determining the “critical” nitrogen loadings to our coastal waters (43).


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Acknowledgments Drinking water monitoring data were received from numerous sources. The authors wish to acknowledge specifically the contributions from the Massachusetts Water Resources Authority, the City of Baltimore, the City of New York, the City of Portland, the Washington Aqueduct Division of U.S. Army Engineers, and the Philadelphia Water Department. The authors wish to also acknowledge the support of both the EPA and NOAA Research Laboratories in Narragansett, RI, of the Mellon Foundation, and of the Hudson River Foundation.

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Received for review September 18, 1996. Revised manuscript received March 5, 1997. Accepted March 10, 1997.X ES960803F X

Abstract published in Advance ACS Abstracts, May 15, 1997.

Atmospheric Deposition of Nitrogen Oxides onto the Landscape

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