the Lower Hudson River Estuary 1. The Ecosystems Center, Marine Biological Lab, Woods Hole, MA 02543

Transcription

1 Wastewater and Watershed Influences on Primary Productivity and Oxygen Dynamics in the Lower Hudson River Estuary Robert W. Howarth 1, 2, Roxanne Marino 1, 2 Dennis P. Swaney 2, and Elizabeth W. Boyer 3 1. The Ecosystems Center, Marine Biological Lab, Woods Hole, MA Department of Ecology & Evolutionary Biology, Cornell University, Ithaca, NY College of Environmental Science and Forestry, State University of New York, Syracuse, NY 13210

2 Abstract: Primary productivity in the saline Hudson River estuary is strongly regulated by water residence times in the estuary. Nutrient loads and concentrations are very high, and when residence times are more than 2 days, production is extremely high. When water residence times are less than 2 days, production rates are low to moderate. Residence times are controlled both by freshwater discharge into the estuary and by tidal mixing, so residence times are longest and production is highest during neap tides when freshwater discharge is low. Freshwater discharge was generally high in the 1970s, which kept primary production low. In contrast, freshwater discharge rates were lower in the 1990s, and the estuary became hypereutrophic. Nutrient loading per area of estuary to the saline portion of the Hudson is probably the highest for any major estuary in North America. Approximately 58% of the nitrogen and 81% of the phosphorus comes from wastewater effluent and other urban discharges in the New York City metropolitan area. Some 42% of the nitrogen and 19% of the phosphorus comes from upriver tributary sources. For nitrogen, these tributary inputs are dominated by nonpoint sources, with atmospheric deposition from fossil fuel combustion and agricultural sources contributing equally. Human activity has probably increased nitrogen loading to the Hudson estuary by 12-fold and phosphorus loading by 50-fold or more since European settlement. Nitrogen and phosphorus loadings to the estuary have decreased somewhat since 1970 due to universal secondary treatment of dry-weather wastewater effluents and a ban on phosphates in detergents. The Hudson estuary suffered from low dissolved oxygen concentrations over much of the 20 th century, but by the mid 1990s, all dry-weather discharges from sewage treatment plants in the New York City region received secondary treatment. This greatly reduced labile organic carbon (BOD) loadings, and resulted in dissolved oxygen concentrations that in most recent years have met the New York State standards. Between the1970s and the 1990s, the estuary switched from sewage as the primary input of labile organic matter to phytoplankton primary production as the major input. Further improvement in water quality is desirable, with the goal of reducing the tropic status of the estuary from hypereutrophic to moderately eutrophic. This will require upgrading of sewage treatment plants to nutrient reduction technology, at an estimated cost of $112 to $277 million per year, and a substantial reduction in nitrogen loading from combined sewer overflows and from nonpoint sources.

3 1 Over the past 20 years, the primary concern of water quality management in estuaries has switched from control of discharges of toxic substances and organic matter which contributes to biological oxygen demand (BOD) to control of nutrient pollution and eutrophication (NRC 1993, 2000). In part, this is because of great success in reducing problems with BOD and toxics over this time period, in the Hudson estuary and elsewhere (NRC 1993). In addition, problems from excess nutrient inputs and particularly nitrogen have grown dramatically, to the point where nutrients are now considered the greatest pollution threat to coastal marine ecosystems (NRC 2000; Howarth et al. 2000b). Some two-thirds of the estuaries in the contiguous United States are now moderately to severely degraded from nutrient over-enrichment (Bricker et al. 1999). Estuaries vary in their sensitivity to nutrient pollution, and the Hudson in the past has been considered to be relatively insensitive (Bricker et al. 1999). However, our recent research demonstrates that primary productivity by phytoplankton in the saline regions of the Hudson estuary increased dramatically in the 1990s relative to the 1970s, and the Hudson is now quite eutrophic (Howarth et al. 2000a). The Hudson River estuary is profoundly influenced both by its watershed and by enormous inputs of materials from municipal wastewater treatment plants. The estuary receives runoff from a watershed having an area of 34,600 km 2 (USGS 2002). The freshwater discharge from the watershed is similar to that in several other large estuaries, such as Delaware Bay, San Francisco Bay, and Long Island Sound (Limburg et al. 1986; Howarth et al. 2000a; However, when compared to other well-known estuaries with large drainage basins which are part of the LOICZ project data set, the surface area of the Hudson River estuary is fairly small, and thus the freshwater discharge per area of estuary is quite high (Table 1). This contributes to a rapid flushing of the estuary, with water residence times on the scale of 0.1 to 4 days (Howarth et al. 2000a). By comparison, the water residence times in Delaware Bay, Chesapeake Bay, and Long Island Sound are on the order of 60, 250, and 1100 days, respectively (Nixon et al. 1996; Rapid flushing, particularly when water residence times are less than 1 to 2 days, makes the Hudson less sensitive to nutrient pollution than other estuaries with longer water residence times (Howarth et al. 2000a). Some 4.4 million people live in the watershed (US Census Bureau 2002), giving an average population density of 128 individuals km -2. The majority of the population lives in the New York City metropolitan area, with fewer people living in the mostly forested basin upriver (Howarth et al. 1991; Swaney et al. 1996; Boyer et al. 2002). The Hudson River estuary receives a substantial amount of sewage from the New York City metropolitan area, and as a result, the nutrient loading per area or volume of estuary is the highest for any large estuary in the United States (Nixon and Pilson 1983; NRC 1993). BOD loading is also high. The first sewers in New York City were constructed in 1696, and the sewer system grew steadily in collection capacity until the middle of the 20 th century (Brosnan and O Shea 1996). The first primary sewage treatment plants in the New York City metropolitan area were only built in the 1930 s, and secondary treatment

4 2 plants to reduce BOD loadings were not constructed until after such treatment was mandated by the Clean Water Act in Full secondary treatment of the waste stream was only achieved in the past decade (Brosnan and O Shea 1996). The result has been a dramatic increase in water quality, as measured by dissolved oxygen levels in the estuary. In this paper, we discuss how freshwater discharge and nutrients affect primary productivity in the saline Hudson River estuary, how human activity has altered the inputs of nutrients and labile organic matter to the estuary over time, what regulates oxygen concentrations in the estuary, and what actions might be taken to further improve water quality in the Hudson in the future. Our focus is on the portion of the Hudson River which begins at the Battery at the southern tip of Manhattan and extends northward 66 km to the northern end of Haverstraw Bay (Fig. 1). So defined, the saline Hudson estuary has an area of 149 km 2 and a volume of 1.11 km 3 (estimated from NOAA navigational charts). The estuary can be divided into the saltier, mesohaline region from the Battery north 36 km and the less salty, oligohaline region from river km 36 north to river km 66, an area that includes the Tappanzee and Haverstraw Bay. Primary Productivity: Rates and Controls Nutrient concentrations and loadings in the saline Hudson River estuary are very high, and primary productivity is regulated largely by physical factors (Malone 1977; Howarth et al. 2000a). Perhaps the most important of these physical factors is water residence time. Compared to most large estuaries, the Hudson is rapidly flushed, and water residence times in the surface waters that comprise the photic zone of the mesohaline estuary range from 0.1 to 4 days (Howarth et al. 2000a). Most species of phytoplankton have maximum growth rates that result in a doubling of the population on the time scale of roughly 1 day, with larger species having somewhat longer doubling times ranging from 2 to 15 days (Banse 1976; Malone 1977, 1980; Chan 2001). Thus, much of the time phytoplankton are flushed from the Hudson estuary about as rapidly as they grow, limiting the size of populations and keeping rates of primary productivity fairly low. Both the tidal cycle and freshwater discharge into the Hudson affect water residence times in the estuary. When freshwater discharge at the Green Island gauging station is greater than 300 m 3 sec -1, water residence times in the photic zone are less than 1 day (Fig. 2.A). At lower rates of discharge, water residence times vary greatly, with at least some of this variation related to the tidal cycle; shorter water residence times are more prevalent during the spring-tide period of the month when tidal mixing is greater, and neap tides favor longer water residence times. In a series of cruises in the mid-1990s, we found that rates of gross primary production (GPP) were always less than 2 g C m -2 d -1 and generally less than 1 g C m -2 d -1 when the freshwater discharge (measured at Green Island) was greater than 300 m 3 sec -1, leading to short water residence times (Fig. 2.B). Substantially higher rates of production, up to 18 g C m -2 d -1, were found during some neap tide cycles when

5 3 freshwater discharge was low and water residence times in the photic zone exceeded 2 days. In addition to increased water residence times, greater light penetration into the water column may have favored high rates of production during periods of low discharge and low tidal mixing (Fig. 2.C). The greater light penetration likely resulted from lessened input of sediment from up river and decreased resuspension of bottom sediments. Estuaries have been classified as eutrophic when annual production ranges between 300 and 500 g C m -2 y -1, and as hypereutrophic when annual production exceeds 500 g C m -2 y -1 (Nixon 1995; NRC 2000). This corresponds to a daily production rate of 2 to 3 g C m -2 d -1 on average during the active growing season (NRC 2000; Howarth et al. 2000a). Thus, the saline Hudson estuary would be considered hypereutrophic during neap tide periods when freshwater discharge is low, but not when discharge is high (Fig. 2.B). Prior to our measurements in the 1990s, the only published data on rates of primary productivity in the saline Hudson estuary were from studies in the early1970s. Then rates were generally less than 1 g C m -2 d -1 and were never higher than 2 g C m -2 d -1 (Malone 1977; Sirois and Frederick 1978). These lower primary production values are consistent with our measurements during the 1990s for periods of high discharge, and in fact the decade of the 1970s was a wetter decade with generally higher rates of freshwater discharge (Fig. 3). Average summer discharge rates at Green Island exceeded 200 m 3 sec -1 in most of the years of the early to mid 1970s and exceeded 300 m 3 sec -1 in every year except The resulting high flushing and corresponding low rates of productivity during that time period gave rise to the thought that the Hudson estuary is relatively insensitive to eutrophication despite the very high nutrient concentrations present (Garside et al. 1976; Malone 1977; Bricker et al. 1999). Different methods were used to measure productivity in the 1970s than in the 1990s, which can complicate direct comparison of the data sets. The measurements in the 1990s were made from in-situ changes in dissolved oxygen concentrations over a diel cycle (Swaney et al. 1999), whereas the measurements in the 1970s were made by the C 14 method (O Reilly et al. 1976; Malone 1977) and by the light-dark bottle method (Sirois and Fredrick 1978). The in-situ oxygen technique would be expected to give a more reliable and most likely higher estimate of gross primary production (GPP) than the other two methods (Swaney et al. 1999; Howarth and Michaels 2000). The C 14 method in particular would be expected to give a lower estimate, as it measures a rate of carbon fixation somewhere between net and gross primary production (Howarth and Michaels 2000). Nonetheless, the apparent increase in production between the 1970s and the 1990s is probably real at least in part, as chlorophyll levels were also higher in the 1990s, while rates of production per chlorophyll appear to be similar in both sets of studies (Howarth et al. 2000a). Most of our data were collected between May and October, and we have not previously estimated an annual rate of production. However, for the 6-month period from May through October over several years, we found GPP to range between 300 and 370 g C m -2 in the oligohaline estuary and 500 to 750 g C m -2 in the mesohaline estuary (Swaney et al. 1999, and our unpublished data). By comparison, for the May to October

6 4 period for various studies in , production was roughly 150 g C m -2 in the mesohaline estuary and 180 g C m -2 for the oligohaline estuary (Swaney et al. 1999, using data from Sirois and Fredrick 1978 and Malone 1977). Overall, it would appear that GPP in the 1990s was perhaps twice the rate of production reported for the 1970s in the oligohaline Hudson estuary and 4-times the rate reported for the mesohaline estuary. Annual productivity reported for the saline Hudson estuary in the 1970s was approximately 1.3 times the rate measured over the May to June period in the 1970s (O Reilly et al. 1976; Malone 1977; Sirois and Fredrick 1978; Limburg et al. 1986; Malone and Conley 1996; Swaney et al. 1999). Assuming that this relationship holds for our data, mean annual rates of GPP in the Hudson during the 1990s can be estimated as 850 g C m -2 y -1 in the mesohaline estuary and 450 g C m -2 y -1 in the oligohaline estuary. The freshwater discharge during the 1990s is much more characteristic of the situation for the Hudson over the past 6 decades than is that of the 1970s (Fig. 3), so our estimates are a better reflection of long-term average rates of GPP in the saline Hudson. Future climate change may well lessen freshwater discharge from the Hudson during the summer, continuing a trend of high production (Howarth et al. 2000a; Scavia et al. 2002). For many marine and estuarine ecosystems, the log-transformed rate of primary production (measured by the C 14 method) is a linear function of the log-transformed inorganic nitrogen loading rate (Fig. 4; Nixon et al. 1996). While primary productivity in many estuaries is less than predicted by the regression of Nixon et al. (1996) due to a variety of factors including rapid flushing and light limitation (NRC 2000; Cloern 2001), the regression sets a reasonable upper bound for the relationship between nutrient loading and production in estuaries if these physical factors were not limiting. Nitrogen (N) loading to the Hudson estuary is estimated as 43 x 10 3 tons N y -1 (Table 2, and discussion below in Nutrient Loading in the Past 30 Years ), most of which is as inorganic nitrogen (Malone and Conley 1996). This corresponds to an average loading per area of the estuary of 290 g N m -2 y -1, although in fact the loading in the oligohaline estuary would be somewhat less, as most of the nutrient input enters directly into the mesohaline estuary near Manhattan and some of this is N exported from the estuary rather than being mixed up into the oligohaline estuary. This level of nitrogen loading corresponds to a predicted value for C 14 primary production in the Hudson estuary of 820 g C m -2 y -1 (Fig. 4), which is remarkably close to our roughly estimated annual value of GPP in the mesohaline estuary (850 g C m -2 y -1 ). While we might expect the maximum value of productivity predicted from nutrient loading to be higher than that measured in-situ, C 14 productivity underestimates GPP. We tentatively conclude that the rates of GPP measured during the 1990s reflect the maximum rate that can be obtained in the Hudson under its current nutrient loading regime because of the constraint imposed by short water residence times from advection and tidal mixing. Nutrient Loading over the Past 30 Years: Nutrient inputs to the Hudson estuary are of interest both because they set an upper limit on rates of GPP there and because much of the nitrogen and phosphorus is

7 5 exported from the Hudson to other parts of the harbor of New York City and to the coastal waters of the New York Bight. Primary production in the lower bay of New York and in the plume of the Hudson River in the 1970s and 1980s was reported to be in the range of 600 to 800 g C m -2 y -1 (O Reilly et al. 1976; Malone and Conley 1996), indicating that these systems are quite eutrophic (Nixon 1995; NRC 2000). Chlorophyll concentrations in the plume of the Hudson River on the continental shelf range up to 20 µg l -1 (Malone and Conley 1996), levels which also indicate a high degree of eutrophication (NRC 1993). The apex of the New York Bight (an area of 1,250 km2) becomes hypoxic every year, and a large region of the Bight became anoxic in 1976 (Mearns et al. 1982). Nutrients enter the saline Hudson estuary from wastewater in the New York City metropolitan area, from combined sewer overflows and storm runoff in the metropolitan area, from the freshwater portion of the Hudson River (above river km 66), and from the salt water entering the estuary from the ocean. In this chapter, we only estimate the inputs from sources within the watershed, including the tributary sources coming down the Hudson River. While first-order estimates of nutrient exchange with the sea are possible under the assumption of steady state (Gordon et al. 1996), this term is difficult to assess without detailed hydrodynamic modeling, and generally has not been estimated in nutrient budgets for other estuaries (Nixon et al. 1996; NRC 2000). In the case of the Hudson, it may be large, as the salt water entering the estuary first passes through Raritan Bay and lower New York harbor, and these systems receive substantial nutrient inputs themselves. The saline Hudson estuary receives a daily input of wastewater of approximately 3.4 x 10 6 m 3 d -1 (calculated from data in Clark et al. 1992), or one third of the total wastewater flux for the entire New York City Metropolitan area, an estimated 10 x 10 6 m 3 d -1 (Clark et al. 1992; Brosnan and O Shea 1996). By the early 1990s, all of the dryweather sewage discharges in the metropolitan area received secondary treatment (Brosnan and O Shea 1996). The effluent from the average secondary sewage treatment plant in the United States has a total nitrogen concentration of 19 g N m -3 and a total phosphorus content of 3 g P m -3 (NRC 1993). Assuming that these values apply to the treatment plants in the New York City metropolitan area, we estimate nitrogen and phosphorus loads to the saline Hudson estuary in the 1990s as 24 x 10 3 metric tons N y -1 and 3.7 x 10 3 metric tons P y -1 (Table 2). Another estimate of nutrient inputs from wastewater to this portion of the estuary can be made by scaling the estimate of Brosnan and O Shea (1996) for the entire metropolitan waste flow to the percentage of the wastewater flow that enters the saline Hudson estuary (34%). The nitrogen input to the estuary calculated this way is very similar to that estimated using the NRC concentration data and wastewater flows; however, the phosphorus input is only 2.1 x 10 3 tons P y -1, a value 40% lower. For the entire metropolitan New York City area, Brosnan and O Shea estimate that nitrogen and phosphorus inputs in combined sewer overflows (CSOs) and in storm water runoff are 4.1 x 10 3 tons N y -1 and 0.67 x 10 3 tons P y -1. If we assume that the percentage of these inputs that go directly into the Hudson estuary is one third, as is true

8 6 for wastewater effluent, then we can estimate these other urban inputs to the estuary as 1.4 x 10 3 tons N y -1 and 0.22 x 10 3 tons P y -1 (Table 2). Lampman et al. (1999) have estimated the flows of total nitrogen and phosphorus down the Hudson River as 18 x 10 3 tons N y -1 and 0.9 x 10 3 tons P y -1 at a point 125 km north of the Battery, or 60 km north of beginning of the oligohaline estuary in Haverstraw Bay. This is a conservative estimate of the input of these nutrients to the saline estuary, as additional nitrogen and phosphorus enter the freshwater Hudson from tributaries over this 60 km stretch. Nonetheless, the estimates of Lampman et al. (1999) are the best available estimates for total nutrient loading to the estuary from the upstream Hudson and its tributaries (Table 2). These non-point source estimates of N and P input from the freshwater Hudson, combined with the total urban inputs (wastewater, CSO s and storm water runoff) result in an estimate of the total nutrient load to the saline Hudson estuary as of the mid 1990s of 43 x 10 3 tons N y -1 and 4.8 x 10 3 tons P y -1 (Table 2). For nitrogen, a similar but slightly smaller estimate (38 x 10 3 tons N y -1 ) is obtained from the population density in the watershed and a regression model that relates population density to total nitrogen flux for large regions in the temperate zone (Howarth 1998). Wastewater effluent plus CSO s and storm water runoff contribute 58% of the nitrogen and 81% of the phosphorus, while the upstream tributary sources contribute 42% of the nitrogen and 19% of the phosphorus (Table 2). Nitrogen and phosphorus loading rates expressed per area of the saline Hudson are 290 g N m -2 y -1 and 32 g P m -2 y -1. These are far higher than reported for any other large estuary in the United States (NRC 1993). For comparison, total nitrogen and phosphorus loadings to Chesapeake Bay are estimated to be 20 to 30-fold less than the Hudson (13 g N m -2 y -1 and 1 g P m -2 y -1 ; Boynton et al. 1985), yet the Chesapeake is considered an ecosystem that is highly degraded from nutrient pollution (Bricker et al. 1999; NRC 2000). Nitrogen loadings to Delaware Bay, Narragansett Bay, and Boston Harbor are also all substantially lower than to the saline Hudson, at 27, 27, 130 g N m -2 y -1, respectively (Nixon et al. 1996). Total phosphorus loadings to these three systems are 5, 2.6, and 21 g P m -2 y -1, respectively (Nixon et al. 1996). Even by European standards, the nutrient loading to the Hudson estuary is high. The highly polluted Scheldt estuary in Belgium has a nitrogen loading of 190 g N m -2 y -1 and a phosphorus loading of 33 g P m -2 y -1 (Billen et al. 1985). Of the nutrients entering the estuary from upstream in the Hudson River, a substantial portion likely comes from non-point sources. Boyer et al. (2002) compared the nitrogen cycle of 16 major watersheds in the northeastern United States, including the Mohawk River valley and the basin of the upper Hudson River (above river km 260 where the Mohawk joins the Hudson). Together, these two tributaries comprise 60% of the area of the entire Hudson River basin (Boyer et al. 2002) and contribute 65% of the total freshwater discharge of the Hudson River (Howarth et al. 1996a). For the upper Hudson and Mohawk River valleys combined, nitrogen deposition from the atmosphere contributes 41% of the flux of nitrogen from the landscape into the rivers. Agriculture

9 7 contributes another 39% of the flux, with 28% of the nitrogen in the rivers originating from nitrogen fixation by agricultural crops and 11% from nitrogen fertilizer (Table 3; Boyer et al. 2002). The total nitrogen flux from these two watersheds is 13 x 10 3 tons y -1 (Table 3; Boyer et al. 2002), as compared to an estimated flux of 18 x 10 3 tons y -1 for nitrogen further down the Hudson River at river km 125 (Lampman et al. 1999). We would expect the nutrient flux between river km 125 and 66 to be even greater than this, since the watersheds of the tributaries there are more disturbed. The watersheds of the Mohawk River and upper Hudson are 73% forested on average (Table 3; Boyer et al. 2002), while the lower Hudson is 57% forested (Swaney et al. 1996). Also, atmospheric deposition of nitrogen is much greater closer to urban sources (Holland et al. 1999; NRC 2000). On the other hand, a significant fraction of the N entering the freshwater Hudson is not exported to the saline estuary, due to the in-river processes of denitrification and sedimentation (Lampman et al. 1999). In the two decades between the 1970s and the 1990s, nutrient inputs from sewage decreased, due in part to improved sewage treatment. The improvements were designed to lower BOD loadings, and not nutrient levels, but nonetheless probably resulted in some reduction of nutrient loading. Population in the lower part of the Hudson watershed grew by only 2 % between 1970 and 1990 (US Census Bureau 2002), and so total discharge of wastewater into the estuary likely remained almost unchanged over that time, at 3.4 x 10 6 m 3 d -1. While a major new treatment plant, the North River plant, was built and came on line in the 1980s, the total wastewater flow into the saline estuary was not increased; the plant simply replaced the wastewater volume of raw sewage from 50 individual outlets from Manhattan to the Hudson estuary with the same volume of secondary-treated sewage (Clark et al. 1992; Brosnan and O Shea 1996). As of the early 1970s, 38% of the wastewater discharge into the Hudson estuary was raw sewage, 15% received primary treatment, and 47% received secondary treatment (calculated from data in Clark et al. 1992). Using the average concentration of nutrients in effluents from plants receiving those different levels of sewage treatment in the United States (Table 4; NRC 1993), we estimate that nutrient loads from wastewater plants would have been 30 x 10 3 tons N y -1 and 5 x 10 3 tons P y -1 in the early 1970s. Thus, the improved sewage treatment of the 1990s resulted in a 25% decrease in both N and P loadings to the saline estuary (Table 2). As noted above, our estimate for nitrogen loading from wastewater in the 1990s is in reasonable agreement with that derived from scaling down the estimate of Brosnan and O Shea (1996), but our estimate for phosphorus is higher. Our estimate for phosphorus loading from wastewater in the 1970s (Table 2) is also higher than that suggested by the data of Clark et al. (1992), which lead to an estimate of total phosphorus loading from wastewater of 3 x 10 3 tons P y -1 in the early 1970s. However, their estimate that raw sewage contains only 1.3 g P m -3 seems low in comparison to either their data for treated wastewater or data for average U.S. sewage plants (Table 4; NRC 1993). Based on observations of soluble reactive phosphorus (SRP) in the estuary over time, the assumption that SRP is conservative within the estuary, and a transport model, Clark et al. (1992) concluded that SRP loadings as of the late 1980s were only one third of the loadings in the early 1970s, which is a bigger change that our estimates suggest (66% and

10 8 25% reductions, respectively). We may be underestimating the decrease in phosphorus loading over time, as a ban on phosphates in detergents in New York in the early 1970s led to a 33% reduction in the phosphorus concentration in wastewater effluent from plants in the New York City metropolitan which received no upgrade in treatment technology (Mueller et al and 1978 as cited in Clark et al. 1992), and this is not taken into account using the NRC (1993) sewage plant data. On the other hand, the assumption that SRP is conservative in the Hudson estuary (Clark et al. 1992) may be less valid in the 1980s than in the 1970s, as increased GPP would have assimilated more SRP in the 1980s, and higher oxygen concentrations in the water column may have increased the phosphate adsorptive capacity of bottom sediments as well (Howarth et al. 1995). These changes may have led to Clark et al. (1992) to overestimate the decrease in SRP loading from wastewater sources over the two decades based on in-situ SRP measurements. The upstream tributary sources of nitrogen have probably changed rather little in the Hudson basin since the 1970s (Jaworski et al. 1997). We are aware of no data that would allow us to estimate how the upstream tributary inputs of phosphorus have changed since In any event, it seems likely that the wastewater sources were a greater percentage of the total nutrient input in the 1970s than in the 1990s. Nutrient Loading and GPP in the Pristine Hudson Estuary: Even before European settlement, nutrient loading per area of the estuary was probably high in the Hudson River compared to most estuaries, because of the high ratio of watershed area to estuary area (Table 1). At the scale of large regions, the nitrogen flux into the North Atlantic Ocean per area of watershed in the temperate zone is a linear function of the net anthropogenic inputs of nitrogen per area to the region. The zero intercept of this relationship, corresponding to no human influence, gives an estimate of the riverine nitrogen export off the pristine landscape of approximately 100 kg N km -2 y -1 (Howarth et al. 1996b; NRC 2000). This relationship holds well at a smaller scale as well, as seen for 16 major watersheds in the northeastern United States, including the Mohawk River basin and the upper Hudson River basin (Fig. 5; Boyer et al. 2002). A nitrogen flux of 100 kg N km -2 y -1 from the landscape contributes a total load to the estuary of 3.5 x 10 3 tons N y -1 for a watershed the size of Hudson basin, suggesting that human activity up to the 1990s has increased nitrogen loading to the Hudson estuary by 12-fold (Table 2, 5). Per area of estuary, the pristine nitrogen loading would have corresponded to an input of 23 g N m -2 y -1 (Table 5). Note that this estimated nitrogen loading to the Hudson estuary under pristine conditions is in fact higher than the current loading to Chesapeake Bay and is only slightly less than the current loading to Delaware and Narragansett Bays. Again, this reflects the very high ratio of watershed area to estuarine surface area for the Hudson in comparison to other large estuaries. While the export of nitrogen from the landscape in the temperate zone can be well predicted from the net anthropogenic inputs of nitrogen, export of phosphorus is quite dependent upon the amount of phosphorus in the parent soil, which is highly variable

11 9 across regions. As a result, there is no consistent estimate of a baseline flux of phosphorus from pristine watersheds, as there is for nitrogen (NRC 2000). For the Hudson, we can estimate what the pre-european phosphorus input to the estuary may have been by evaluating changes in erosion. In the pristine landscape (i.e. 100% forested), most of the phosphorus input to the Hudson would likely have been bound to particles. Currently, inputs of sediment to the Hudson estuary from erosion in the watershed are 10-fold higher than if the basin were entirely forested (Swaney et al. 1996). Assuming that the phosphorus content of this eroded sediment has not changed over time, then the pre-european input of phosphorus to the estuary can be estimated as 10% of the current input from present upstream tributary sources (Table 2), or 0.09 x 10 3 tons P y -1. The actual flux was likely less than this, both because the current input from upstream tributary sources includes some wastewater inputs from the tributaries and because the phosphorus content of the soils in the Hudson basin have probably been increased from fertilization with phosphorus. In comparison to the estimated load for the 1990s (Table 2), human activity in the Hudson basin appears to have increased phosphorus loading to the estuary by at least 50-fold. A pristine nitrogen loading of 23 g N m -2 y -1 would predict a rate of primary production of 270 g C m -2 y -1 (Fig. 4; Nixon et al. 1996) if the nitrogen loading were as inorganic nitrogen and if other factors such as short water residence times were not constraining GPP. As discussed above, GPP in the 1990s in the mesohaline estuary was roughly equal to the potential rate of C 14 production estimated from nitrogen loading (Fig. 4). Assuming this was true before European settlement, GPP can be estimated as 270 g C m -2 y -1. The actual value was probably lower, because in fact, most of the nitrogen export from the pristine landscape was probably as organic nitrogen (Howarth et al. 1996b; Lewis 2002), of which the refractory component would be unavailable to phytoplankton, and short water residence times caused by spring tide mixing would have resulted in lower rates of primary production even during periods of low freshwater discharge. However, this exercise suggests that human activity has increased GPP in the mesohaline Hudson estuary by 3-fold or more (Table 5). Dissolved Oxygen: Historical Trends and Controls The Hudson estuary is somewhat protected against low dissolved oxygen events both by the rapid flushing that removes organic wastes and by a rapid mixing over depth that can rapidly replenish oxygen as it diffuses in from the atmosphere (Clark et al. 1995; Swaney et al. 1999). Nonetheless, the estuary has historically had problems with low oxygen (Fig. 6). Much of this can be ascribed to organic loading from sewage effluents (BOD), and oxygen concentrations in the saline Hudson estuary have increased steadily in response to improved sewage treatment (Fig. 6, 7; Suszkowski 1990; Clark et al. 1995; Brosnan and O Shea 1996). The estuary is classified by the State of New York as class I for water quality goals (secondary contact recreation, not primary contact), which sets a limit of 4 mg l -1 for minimum dissolved oxygen concentrations. As a result of reduced BOD loadings from upgrading to secondary sewage treatment throughout the New York City metropolitan area, this goal has been met most of the time since 1990 according to the data collected by the City of New York (Fig. 7; DEP 2001). However, even in

12 10 recent years the Hudson estuary would have difficulty meeting the state standard of 5 mg l -1 for primary contact recreation. In our cruises in 1998, 1999, and 2000, we found dissolved oxygen concentrations in the bottom waters of the estuary in early morning (when concentrations are lowest; Swaney et al. 1999) to be below 5 mg l -1 on roughly half the cruises and below 4 mg l -1 on 2 cruises (out of a total of 20 cruises), in August of 1999 and July of 2000 (our unpublished data). In the saline Hudson estuary, the primary sources of organic matter that fuel respiration leading to oxygen depletion are BOD inputs from sewage and phytoplankton primary production. The estuary also received substantial inputs of organic matter from upriver, estimated as 50 x 10 3 tons C y -1 in the late 1980s (Howarth et al. 1996a). Much of the labile organic carbon that enters the freshwater Hudson is respired in-situ (Howarth et al. 1996a), and as a result most of the organic C that is exported downstream is likely fairly refractory and has little influence on oxygen dynamics in the saline estuary. At the peak of agricultural activity in the Hudson River basin a century ago, the inputs of organic matter may have been 80% greater than at present due to higher erosion, and prior to European settlement in the basin, the flux may have been 40% of the current rate (Swaney et al. 1996). By the early 1970s when the new environmental movement focused attention on water quality resulting in the Clean Water Act of 1972, the organic carbon inputs to the Hudson estuary were dominated by sewage. As discussed above, in the early 1970s 38% of the wastewater discharge into the Hudson estuary was raw sewage, 15% received primary treatment, and 47% received secondary treatment (calculated from data in Clark et al. 1992). Using average values for the United States for the BOD load from treatments plants receiving various levels of treatments (Table 4), we estimate BOD loadings to the saline Hudson estuary in the early 1970s as 49 x 10 3 tons C y -1. By the 1990s, virtually 100% of the wastewater inputs to the Hudson estuary during dry-weather conditions received secondary treatment (Brosnan and O Shea 1996). Using the same approach as for our 1970s estimate, we calculate that the complete conversion to secondary level would have reduced the input of labile organic carbon from wastewater treatment plants to 7.5 x 10 3 tons C y -1 in the 1990s. Scaling the estimates of Brosnan and O Shea (1996) for discharges from CSOs and storm water discharge for the entire metropolitan area to only the area of the saline Hudson estuary, as we did for nutrients above, suggests a further BOD loading of 6 x 10 3 tons C y -1. If we assume that the CSO and storm runoff remained constant over the past several decades, then we estimate that the total BOD from wastewater effluent and other urban sources decreased by 75% between the early 1970s and the mid 1990s, from 55 x 10 3 tons C y -1 to 14 x 10 3 tons C y -1. At the same time as BOD loadings from wastewater treatment plants decreased, rates of GPP increased in the Hudson estuary, probably due to the longer water residence times resulting from the decrease in freshwater discharge between the 1970s and the 1990s. Given a rate of GPP of 200 to 250 g C m -2 y -1 in the early 1970s (O Leary et al. 1976; Malone 1977; Sirois and Fredrick 1978), phytoplankton production would have provided an input of 33 x 10 3 tons C y -1 of labile organic matter to the estuary. If we

13 11 assume that in the 1990s, GPP was on average 850 g C m -2 y -1 in the mesohaline estuary and 450 g C m -2 y -1 in the oligohaline estuary, the total input of organic carbon from GPP to the saline estuary would be approximately 90 x 10 3 tons C y -1. Despite the uncertainty in these estimates, the relative importance of sewage effluent and GPP clearly shifted between the early 1970s and the mid 1990s (Table 5). BOD from sewage sources contributed over 60% of the labile carbon to the Hudson estuary in the early 1970s but only 10% in the 1990s. Surprisingly, the total input of labile organic matter to the estuary actually increased over those two decades due to the large increase in GPP (Table 5). GPP by phytoplankton produces oxygen as well as labile carbon, and so given a comparable input of organic matter from GPP and BOD, the BOD loading will have a much greater negative impact on dissolved oxygen concentrations. However, excess GPP can lead to hypoxia and anoxia in estuaries, particularly when the water column is stratified (NRC 1993, 2000). The saline Hudson estuary is generally stratified, yet significant mixing occurs across the pycnocline, and mixing is in fact rapid compared with gas exchange with the atmosphere (Clark et al. 1995; Swaney et al. 1999). Even in completely mixed water columns, high levels of GPP can lead to hypoxia, as was demonstrated experimentally in the Marine Ecosystems Research Laboratory (MERL) facility at nitrogen loadings comparable to those that occur in the Hudson estuary (Frithsen et al. 1985). Eutrophication leads to anoxic and hypoxic events in estuaries as a result of spatial and/or temporal separation of the production of oxygen associated with GPP and its consumption in respiration. Freshwater discharge can dramatically affect oxygen concentrations in the saline Hudson estuary, with concentrations lower at times of lower discharge (Clark et al. 1995; Brosnan and O Shea 1996). This may result from the slower flushing that accompanies reduced discharges (Fig. 2B; Brosnan and O Shea 1996). The mesohaline Hudson estuary also becomes more stratified at times of lower freshwater discharge, in contrast to the general expectation that stratification lessens in estuaries as discharge decreases (Howarth et al. 2000a). This greater stratification may also contribute to lowered oxygen concentrations. Thus, the lower discharge that increases GPP and so in-situ oxygen production also makes the Hudson more sensitive to the effects of this organic loading on oxygen levels. Further Improvements to Water Quality in the Hudson River Estuary The upgrade of sewage treatment in the New York City metropolitan area to secondary treatment has resulted in marked improvement in water quality (Brosnan and O Shea 1996; DEP 2001) and was highlighted by a 1993 report from the National Research Council as one of the greatest success stories in water quality management in estuaries over the past several decades (NRC 1993). However, while oxygen concentrations in the Hudson estuary have improved and usually meet the New York State standard for secondary contact recreation (4 mg O 2 l -1 ), they still do not reliably meet the standard for primary contact recreation (5 mg O 2 l -1 ), as noted above. Our analysis shows that the Hudson estuary is often hypereutrophic, and despite fairly rapid

14 12 flushing, is more sensitive to nutrient pollution than has been previously assumed. Further, water quality management in estuaries is moving beyond consideration just of dissolved oxygen levels, and must now consider other adverse effects of eutrophication, such as reduced biodiversity, increased incidences and duration of harmful algal blooms, and alteration in food web structure (NRC 1993, 2000; Howarth et al.2000b; EPA 2001). Nutrient pollution from the Hudson estuary also contributes to eutrophication in downstream ecosystems, including the plume of the Hudson River on the continental shelf, where hypoxia is a regular event. The nitrogen and phosphorus loads to the Hudson River estuary and to the downstream ecosystems could be significantly reduced through improved sewage treatment. While the nitrogen in effluent from an average secondary sewage treatment plant in the United States contains 19 g N m -3, plants designed for nutrient removal on average discharge only 3 g N m -3 ; for phosphorus, nutrient removal technology results in an average effluent concentration of 1.5 g P m -3, as compared to 3 g P m -3 for secondary treatment (Table 4; NRC 1993). If all the municipal wastewater plants that discharge into the saline Hudson estuary were to upgrade to this level of treatment, nitrogen loading from the sewage plants would be reduced from a current estimated 23 x 10 3 tons N y -1 to 3.7 x 10 3 tons y -1. Assuming no change in discharges from CSOs and from storm sewers, and no change in the nitrogen coming down the Hudson River from upstream sources, total nitrogen loading to the estuary would be reduced to 24 x 10 3 tons y -1, or 150 g N m -2 y -1 per area of estuary. Similarly, phosphorus loading from sewage plants upgraded to nutrient removal technology would be reduced from a current estimate of 3.7 x 10 3 tons y -1 to 1.9 x 10 3 tons y -1, resulting in a total P loading to the saline estuary of 3 x 10 3 tons y -1 or 20 g P m -2 y -1. The cost of building and maintaining sewage treatment plants that include nutrient-reduction technology in the United States is on average $0.37 per m 3 treated, compared to a cost of $0.28 per m 3 for only secondary sewage treatment (Table 4; NRC 1993). Thus, if the New York metropolitan region had upgraded to nutrient removal technology rather than just to secondary over the past few decades, the incremental cost would have been an estimated $0.09 per m 3 of effluent, or $112 million per year for the plants that discharge into the Hudson estuary. To build new nutrient reduction plants in the future would cost more, and if there were no capital savings from converting secondary plants to nutrient reduction plants, the capital cost would be an estimated $0.22 per m 3 and the increased operating costs over that for secondary plants would be $0.01 per m 3 of effluent (Table 4), or a total cost of $277 million per year. There probably is some saving of capital costs when converting secondary treatment to nutrient reduction treatment, so the actual cost of nutrient reduction technology for the Hudson estuary is probably between $112 and $277 million per year, or between $0.08 and $0.17 per person in the watershed per day, if national average costs apply. Note that these estimates are based on 1990 dollars and do not include land costs, but they are otherwise conservative as they are based on 8% interest rates and 20-year depreciation of plants (NRC 1993).

15 13 The CSO and storm sewer discharges could in theory be eliminated as nitrogen sources to the Hudson estuary, and although the cost would be high, this would be desirable for other water quality reasons as well, such as reducing the pathogen load to the estuary. Most pathogens enter the Hudson estuary from CSOs (DEP 2001). Ending CSO discharges should perhaps be a priority of rebuilding the urban infrastructure of the New York metropolitan area. Reducing the nitrogen from upstream tributaries would also be difficult, but a reduction of 50% or more seems possible through a combination of improved sewage treatment upstream, reduction in nitrogen deposition from fossil fuel pollution, improved farming practices, and other measures such as wetland creation (NRC 2000). With this effort, it seems possible to reduce nitrogen loading to the Hudson estuary to 13 x 10 3 tons y -1, or 87 g N m -2 y -1. The regression illustrated in Figure 4 indicates a maximum potential rate of primary production at this loading rate of 480 g C m -2 y -1. We conclude that, given sufficient public will and effort, the Hudson estuary can be restored to an ecosystem that is only moderately eutrophic rather than hypereutrophic, and where the risk of hypoxic events is greatly lessened (Table 5). Acknowledgements We thank Tom Butler and Jon Cole for useful input. Preparation of this manuscript was supported by a grant from the Hudson River Foundation, a not-for-profit corporation with offices in New York. Additional support for our Hudson research has been provided by endowment given by David R. Atkinson to Cornell University. The views expressed here are those of the authors and not of the Hudson River Foundation or Cornell.

16 14 References Banse, K Rates of growth, respiration and photosynthesis of unicellular algae as related to cell size: a review. J. Phycol. 12: Billen, G., M. Somville, E. De Becker, and P. Servais A nitrogen budget of the Scheldt hydrographical basin. Neth. J. Sea Res. 19: Brosnan, T. M.,and M. L. O Shea Long-term improvements in water quality due to sewage abatement in the lower Hudson River. Estuaries 19: Boyer, E. W., C. L. Goodale, N. A. Jaworski, and R. W. Howarth Anthropogenic nitrogen sources and relationships to riverine nitrogen export in the northeastern USA. Biogeochemistry, in press. Boynton, W. R., J. H. Garber, R. Summers, and W. M. Kemp Inputs, transformations, and transport of nitrogen and phosphorus in Chesapeake Bay and selected tributaries. Estuaries 18: Bricker, S. B., C. G. Clement, D. E. Pirhalla, S. P. Orland, and D. G. G. Farrow National Estuarine Eutrophication Assessment: Effects of Nutrient Enrichment in the Nation s Estuaries. Special Projects Office and the National Centers for Coastal Ocean Science, National Ocean Service, National Oceanic and Atmospheric Administration, Silver Springs, MD. Chan, F. A Ecological controls on planktonic nitrogen-fixation: the roles of grazing and cross-ecosystem patterns in phytoplankton mortality. Ph.D. thesis, Cornell University. Clark, J. F., H. J. Simpson, R. F. Bopp, and B. Deck Geochemistry and loading history of phosphate and silicate in the Hudson estuary. Est. Coast. Shelf Sci. 34: Clark, J. F, H. J. Simpson, R. F. Bopp, and B. L. Deck Dissolved oxygen in lower Hudson estuary: Journal of Environmental Engineering Oct. 1995: Cloern, J. E Our evolving conceptual model of the coastal eutrophication problem. Mar. Ecol. Progr. Ser. 210: Department of Environmental Protection New York City 2000 Regional Harbor Survey. New York City Department of Environmental Protection, NY. Environmental Protection Agency Nutrient Criteria Technical Guidance Manual: Estuarine and Coastal Marine Waters. EPA-822-B , Office of Water, Washington, DC.

17 15 Frithsen, J. B., A. A. Keller, and M. E. Q. Pilsen Effects of Inorganic Nutrient Additions in Coastal Areas: A Mesocosm Experiment Data Report, Vol. 1. Marine Ecosystem Research Laboratory, Report 3, Univ. of Rhode Island, Kingston, RI. Garside, C., T. C. Malone, O. A. Roels, and B. A. Sharfstein An evaluation of sewage-derived nutrients and their influence on the Hudson estuary and New York Bight. Est. Coast. Mar. Sci. 4: Gordon, Jr., D. C., P. R. Boudreau, K. H. Mann, J.-E. Ong, W. L. Silvert, S. V. Smith, G. Wattayakorn, F. Wulff and T. Yanagi, LOICZ Biogeochemical Modeling Guidelines. LOICZ Reports & Studies No 5, IGBP, Stockholm. Holland, E. A., F. J. Dentener, B. H. Braswell, and J. M. Sulzman Contemporary and pre-industrial global reactive nitrogen budgets. Biogeochemistry 46: Howarth, R. W An assessment of human influences on inputs of nitrogen to the estuaries and continental shelves of the North Atlantic Ocean. Nutrient Cycling in Agroecosystems 52: Howarth, R. W., H. Jensen, R. Marino, and H. Postma Transport to and processing of P in near-shore and oceanic waters. Pages in H. Tiessen (ed.), Phosphorus in the Global Environment: Transfers, Cycles, and Management. Wiley, Chichester. Howarth, R. W., R. Schneider, and D. Swaney. 1996a. Metabolism and organic carbon fluxes in the tidal, freshwater Hudson River. Estuaries 19: Howarth, R. W., G. Billen, D. Swaney, A. Townsend, N. Jaworski, K. Lajtha, J. A. Downing, R. Elmgren, N. Caraco, T. Jordan, F. Berendse, J. Freney, V. Kuteyarov, P. Murdoch, and Zhu Zhao-liang. 1996b. Riverine inputs of nitrogen to the North Atlantic Ocean: Fluxes and human influences. Biogeochemistry 35: Howarth, R. W., D. Swaney, T. J. Butler, and R. Marino. 2000a. Climatic control on eutrophication of the Hudson River estuary. Ecosystems 3: Howarth, R. W., D. Anderson, J. Cloern, C. Elfring, C. Hopkinson, B. Lapointe, T. Malone, N. Marcus, K. McGlathery, A. Sharpley, and D. Walker. 2000b. Nutrient pollution of coastal rivers, bays, and seas. Issues in Ecology 7: Howarth, R.W., and A. F. Michaels The measurement of primary production in aquatic ecosystems. Pages in: O. Sala, R. Jackson, H. Mooney, and R. W. Howarth (eds.), Methods in Ecosystem Science, Springer, NY. Jaworski, N. A., R. W. Howarth, and L. J. Hetling Atmospheric deposition of nitrogen oxides onto the landscape contributes to coastal eutrophication in the northeast United States. Environmental Science and Technology 31:

18 16 Lampman, G. G., N. F. Caraco, and J. J. Cole Spatial and temporal patterns of nutrient concentration and export in the tidal Hudson River. Estuaries 22: Lewis, W. M Yield of nitrogen from minimally disturbed watersheds in the United States. Biogeochemistry, in press. Limburg, K. E., A. Moran, and W. H. McDowell The Hudson River Ecosystem. Springer-Verlag, NY. Malone, T. C Environmental regulation of phytoplankton productivity in the lower Hudson estuary. Est. Coast. Shelf Sci. 5: Malone, T. C Algal size. Pages in I. Morris (editor), The Physiological Ecology of Phytoplankton. Blackwell, Oxford. Malone, T. C., and D. J. Conley Trends in nutrient loading and eutrophication: A comparison of the Chesapeake Bay and the Hudson River Estuarine Systems. Pages in K. Sherman, N. A. Jaworski, and T. J. Smayda (eds.), The Northeast Shelf Ecosystems: Assessment, Sustainability, and Management. Blackwell, Oxford. Mearns, A. J., E. Haines, G. S. Klepple, R. A. McGrath, J. J. A. McLaughlin, D. A. Segar, J. H. Sharp, J. J. Walsh, J. Q. Word, D. K. Young, and M. W. Young Effects of nutrients and carbon loadings on communities and ecosystems. In: G. F. Mayer (ed.), Ecological Stress and the New York Bight: Science and Management. Estuarine Research Federation, Columbia, SC. National Research Council Managing Wastewater in Coastal Urban Areas. National Academy Press, Washington, DC. National Research Council Clean Coastal Waters: Understanding and Reducing the Effects of Nutrient Pollution. National Academy Press, Washington, DC. Nixon, S. W Coastal marine eutrophication: a definition, social causes, and future concerns. Ophelia 41: Nixon, S. W., J. W. Ammerman, L. P. Atkinson, V. M. Berounsky, G. Billen, W. C. Boicourt, W. R. Boynton, T. M. Church, D. M. DiToro, R. Elmgren, J. H. Garber, A. E. Giblin, R. A. Jahnke, N. J. P. Owens, M. E. Q. Pilson, and S. P. Seitzinger The fate of nitrogen and phosphorus at the land-sea margin of the North Atlantic Ocean. Biogeochemistry 35: O Reilly, J. E., J. P. Thomas, and C. Evans Annual primary production (nanoplankton, netplankton, dissolved organic matter) in the lower New York Bay. Pager 19 in: Hudson River Ecology, 4 th Symposium. Hudson River Environmental Society. Bronx, NY.

19 17 Scavia, D., J. C. Field, Boesch, R. Buddemeier, V. Burkett, D. Canyan, M. Fogarty, M. A. Harwell, R. W. Howarth, C. Mason, D. J. Reed, T. C. Royer, A. H. Sallenger, J. G. Titus Climate change impacts on US Coastal and marine ecosystems. Estuaries, in press. Sirois, D. L., and S. W. Fredrick Phytoplankton and primary production in the lower Hudson River estuary. Est. Coast. Mar. Sci. 5: Suszkowski, D. J Conditions in New York/New Jersey Harbor Estuary. Pages in Proceedings of Cleaning Up Our Coastal Waters: An Unfinished Agenda. Manhattan College, Riverdale, NY. Swaney, D. P., D. Sherman, and R. W. Howarth Modeling water, sediment, and organic carbon discharges in the Hudson-Mohawk Basin: coupling to terrestrial sources. Estuaries 19: Swaney, D. P., R. W. Howarth, and T. J. Butler A novel approach for estimating ecosystem production and respiration in estuaries: application to the oligohaline and mesohaline Hudson River estuary. Limnol. Oceanogr. 44: US Census Bureau Census of Population and Housing. Economics and Statistics Administration, U. S. Department of Commerce. [online] url: USGS National Water-Quality Assessment (NAWQA) Study-Unit Investigations in the conterminous United States: Watershed Boundaries. [online] url:

20 Figure Legends Fig. 1. Hudson River basin and saline Hudson River estuary. Proposed LTER upper and lower boundaries correspond to the boundaries of the saline estuary as defined in this paper. Fig. 2. Relationship between freshwater discharge and water residence time (A), GPP (B), and light penetration (C) in the estuary during 25 cruises conducted during the spring, summer, and fall of 1994, 1995, and Open symbols represent times when tidal amplitude was <1.15 m; dark symbols represent tides greater than 1.15 m. The dashed line in (A) represent the approximate value for GPP above which an estuary is considered to be hypereutrophic. Discharge data are from the USGS monitoring station at Green Island and represent approximately two-thirds of the total discharge into the estuary. Reprinted from Howarth et al. (2000a). Fig. 3. Average freshwater discharge to the Hudson at the USGS gauging station at Green Island, NY. Upper curve shows annual average flows, lower curve shows average summertime flows. Horizontal lines indicate mean values for annual and summertime discharge. Reprinted from Howarth et al. (2000a). Fig. 4. Primary productivity as function of the input of inorganic nitrogen per area for a variety of marine ecosystems. The open circles are from experimental mesocosm studies at the MERL facility. Dark circles represent natural ecosystems. Reprinted from Nixon et al. (1996). Fig. 5. Export of nitrogen per area of watershed from large watersheds in the northeastern United States as a function of the net anthropogenic nitrogen inputs to the watersheds. Inputs include fertilizer use, nitrogen fixation in agricultural systems, deposition of NO y from the atmosphere, and the net import or export of nitrogen in food and animal feeds. HUD refers to the upper Hudson River basin, and MOH refers to the Mohawk River basin. Data are from Boyer et al. (2002). Fig. 6. Percent oxygen saturation in the Hudson River estuary in top (surface) and bottom waters over time. Data were collected at approximately river km 8. Y indicates significant upgrades to the Yonkers water pollution control plant; NR indicates construction of the North River water pollution control plant. Reprinted from Brosnan and O Shea (1996). Fig. 7. Mean oxygen concentrations in bottom waters of the saline Hudson River estuary during the summer, from 1985 to Bars represent 95% CI; dashed lines indicate the State of New York standards for secondary (4 mg l -1 ) and primary contact recreation (5 mg l -1 ). Reprinted from DEP (2001).