Wastewater and Watershed Influences on Primary Productivity and Oxygen Dynamics in
the Lower Hudson River Estuary
Robert W. Howarth1, 2, Roxanne Marino1, 2
Dennis P. Swaney2, and Elizabeth W. Boyer3
1. The Ecosystems Center, Marine Biological Lab, Woods Hole, MA 02543
2. Department of Ecology & Evolutionary Biology, Cornell University, Ithaca, NY
14853
3. College of Environmental Science and Forestry, State University of New York,
Syracuse, NY 13210
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 20th 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.
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 km2 (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; http://data.ecology.su.se/MNODE/wmap.htm;
http://cads.nos.noaa.gov). 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;
http://data.ecology.su.se/MNODE/wmap.htm; http://cads.nos.noaa.gov). 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 20th
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
2
plants to reduce BOD loadings were not constructed until after such treatment was
mandated by the Clean Water Act in 1972. 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 km2 and a volume of 1.11 km3 (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 m3 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 m3
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
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 m3
sec-1 in most of the years of the early to mid 1970s and exceeded 300 m3 sec-1 in every
year except 1972. 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 C14
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 C14 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
4
period for various studies in 1972-1974, 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 C14 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 103 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 C14 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, C14 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
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 106 m3 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 106
m3 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 103 metric tons N y-1
and 3.7 x 103 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 103 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 103 tons N y-1 and 0.67 x 103 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
6
for wastewater effluent, then we can estimate these other urban inputs to the estuary as
1.4 x 103 tons N y-1 and 0.22 x 103 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 103 tons N y-1 and 0.9 x 103 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 103 tons N y-1 and 4.8 x 103 tons P y-1 (Table 2). For nitrogen, a similar but
slightly smaller estimate (38 x 103 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
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 103 tons y-1
(Table 3; Boyer et al. 2002), as compared to an estimated flux of 18 x 103 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 106 m3 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
103 tons N y-1 and 5 x 103 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 103 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
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. 1976 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 1970. 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 103 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
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 103 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 C14 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
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 103 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 103 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 103 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 103 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 103 tons C y-1 to 14 x 103 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 103 tons C y-1 of labile organic matter to the estuary. If we
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 103 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 O2 l-1), they still do not reliably
meet the standard for primary contact recreation (5 mg O2 l-1), as noted above. Our
analysis shows that the Hudson estuary is often hypereutrophic, and despite fairly rapid
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 103 tons N y-1 to
3.7 x 103 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 103 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 103 tons
y-1 to 1.9 x 103 tons y-1, resulting in a total P loading to the saline estuary of 3 x 103 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 m3 treated,
compared to a cost of $0.28 per m3 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 m3 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 m3 and the increased operating costs over that for secondary plants would be $0.01
per m3 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).
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 103 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.
14
References
Banse, K. 1976. Rates of growth, respiration and photosynthesis of unicellular algae as
related to cell size: a review. J. Phycol. 12: 135-140.
Billen, G., M. Somville, E. De Becker, and P. Servais. 1985. A nitrogen budget of the
Scheldt hydrographical basin. Neth. J. Sea Res. 19: 223-230.
Brosnan, T. M. ,and M. L. O’Shea. 1996. Long-term improvements in water quality due
to sewage abatement in the lower Hudson River. Estuaries 19: 890-900.
Boyer, E. W., C. L. Goodale, N. A. Jaworski, and R. W. Howarth. 2002. 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. 1995. Inputs,
transformations, and transport of nitrogen and phosphorus in Chesapeake Bay and
selected tributaries. Estuaries 18: 285-314.
Bricker, S. B., C. G. Clement, D. E. Pirhalla, S. P. Orland, and D. G. G. Farrow. 1999.
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. 2001. 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. 1992. Geochemistry and loading
history of phosphate and silicate in the Hudson estuary. Est. Coast. Shelf Sci. 34: 213233.
Clark, J. F, H. J. Simpson, R. F. Bopp, and B. L. Deck. 1995. Dissolved oxygen in lower
Hudson estuary: 1978-1993. 1995. Journal of Environmental Engineering Oct. 1995:
760-763.
Cloern, J. E. 2001. Our evolving conceptual model of the coastal eutrophication
problem. Mar. Ecol. Progr. Ser. 210: 223-253.
Department of Environmental Protection. 2001. New York City 2000 Regional Harbor
Survey. New York City Department of Environmental Protection, NY.
Environmental Protection Agency. 2001. Nutrient Criteria Technical Guidance Manual:
Estuarine and Coastal Marine Waters. EPA-822-B-01-003, Office of Water, Washington,
DC.
15
Frithsen, J. B., A. A. Keller, and M. E. Q. Pilsen. 1985. 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. 1976. An evaluation of
sewage-derived nutrients and their influence on the Hudson estuary and New York Bight.
Est. Coast. Mar. Sci. 4: 281-289.
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, 1996. LOICZ Biogeochemical Modeling
Guidelines. LOICZ Reports & Studies No 5, 1-96. IGBP, Stockholm.
Holland, E. A., F. J. Dentener, B. H. Braswell, and J. M. Sulzman. 1999. Contemporary
and pre-industrial global reactive nitrogen budgets. Biogeochemistry 46: 7-43.
Howarth, R. W. 1998. 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: 213-223.
Howarth, R. W., H. Jensen, R. Marino, and H. Postma. 1995. Transport to and
processing of P in near-shore and oceanic waters. Pages 323-345 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: 848-865.
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: 75-139.
Howarth, R. W., D. Swaney, T. J. Butler, and R. Marino. 2000a. Climatic control on
eutrophication of the Hudson River estuary. Ecosystems 3: 210-215.
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: 1-15.
Howarth, R.W., and A. F. Michaels. 2000. The measurement of primary production in
aquatic ecosystems. Pages 72-85 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. 1997. Atmospheric deposition of
nitrogen oxides onto the landscape contributes to coastal eutrophication in the northeast
United States. Environmental Science and Technology 31: 1995-2004.
16
Lampman, G. G., N. F. Caraco, and J. J. Cole. 1999. Spatial and temporal patterns of
nutrient concentration and export in the tidal Hudson River. Estuaries 22: 285-296.
Lewis, W. M.. 2002 Yield of nitrogen from minimally disturbed watersheds in the
United States. Biogeochemistry, in press.
Limburg, K. E., A. Moran, and W. H. McDowell. 1986. The Hudson River Ecosystem.
Springer-Verlag, NY.
Malone, T. C. 1977. Environmental regulation of phytoplankton productivity in the
lower Hudson estuary. Est. Coast. Shelf Sci. 5: 157-171.
Malone, T. C. 1980. Algal size. Pages 433-463 in I. Morris (editor), The Physiological
Ecology of Phytoplankton. Blackwell, Oxford.
Malone, T. C., and D. J. Conley. 1996. Trends in nutrient loading and eutrophication: A
comparison of the Chesapeake Bay and the Hudson River Estuarine Systems. Pages 327349 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. 1982.
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. 1993. Managing Wastewater in Coastal Urban Areas.
National Academy Press, Washington, DC.
National Research Council. 2000. Clean Coastal Waters: Understanding and Reducing
the Effects of Nutrient Pollution. National Academy Press, Washington, DC.
Nixon, S. W. 1995. Coastal marine eutrophication: a definition, social causes, and
future concerns. Ophelia 41: 199-219.
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. 1996. The
fate of nitrogen and phosphorus at the land-sea margin of the North Atlantic Ocean.
Biogeochemistry 35: 141-180.
O’Reilly, J. E., J. P. Thomas, and C. Evans. 1976. Annual primary production
(nanoplankton, netplankton, dissolved organic matter) in the lower New York Bay.
Pager 19 in: Hudson River Ecology, 4th Symposium. Hudson River Environmental
Society. Bronx, NY.
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. 2002. Climate change impacts on US Coastal and marine ecosystems. Estuaries,
in press.
Sirois, D. L., and S. W. Fredrick. 1978. Phytoplankton and primary production in the
lower Hudson River estuary. Est. Coast. Mar. Sci. 5: 57-171.
Suszkowski, D. J. 1990. Conditions in New York/New Jersey Harbor Estuary. Pages
105-131 in Proceedings of Cleaning Up Our Coastal Waters: An Unfinished Agenda.
Manhattan College, Riverdale, NY.
Swaney, D. P., D. Sherman, and R. W. Howarth. 1996. Modeling water, sediment, and
organic carbon discharges in the Hudson-Mohawk Basin: coupling to terrestrial sources.
Estuaries 19: 833-847.
Swaney, D. P., R. W. Howarth, and T. J. Butler. 1999. A novel approach for estimating
ecosystem production and respiration in estuaries: application to the oligohaline and
mesohaline Hudson River estuary. Limnol. Oceanogr. 44: 1509-1521.
US Census Bureau. 2002. 2000 Census of Population and Housing. Economics and
Statistics Administration, U. S. Department of Commerce. [online] url:
http://www.census.gov/dmd/www/2khome.htm.
USGS. 2002. National Water-Quality Assessment (NAWQA) Study-Unit Investigations
in the conterminous United States: Watershed Boundaries. [online] url:
http://water.usgs.gov/GIS/metadata/usgswrd/nawqa.html.
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 1997. 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 NOy
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 2000. 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).
Table 1. Physical characteristics of the saline Hudson estuary in comparison to other
representative estuaries in the temperate zone. Data for all estuaries except the Hudson
are from the LOICZ web site (http://data.ecology.su.se/mnode/index.htm). Only
estuaries larger than 15 km2 are included.
saline Hudson
Area of
estuary
(km2)
Drainage
basin
area
(km2)
149
34,680
233
0.54
4.0
1
134,000
3,500
164,180
560
115,000
3,000,000
118,780
45,000
72,000
62,630
22,000
141,000
8
13
15
35
108
136
173
173
214
368
647
4,273
0.27
0.86
0.35
0.16
0.60
0.23
0.14
0.65
0.07
0.62
0.02
0.01
0.09
0.36
0.16
0.19
2.1
1.0
0.73
3.6
0.48
7.1
0.44
1.3
825
26
230
15
8
44
45
2
14
5
9
17
Gulf of Riga
16,330
Narragansett Bay
264
Chesapeake Bay
11,000
Tomales Bay
16
Mobile Bay
1,060
Rio de la Plata
22,000
Szczecin Lagoon
687
Apalachicola Bay
260
Inner Thermaikos
336
Yalujiang estuary
170
Hawkesbury-Nepean
34
Swan-Canning
33
Ratio of
Riverine
basin area discharge
to estuary per basin
area
area
(m y-1)
Riverine
Mean
discharge
water
per area residence
of estuary
time
(m3 km-2 s-1) (days)
Table 2. Loadings of total nitrogen and phosphorus to the saline Hudson River estuary.
Total nitrogen (103 tons y-1)
Contribution from wastewater plants effluent
Contribution from upriver tributaries
Contribution from CSOs and storm water
Phosphorus (103 tons y-1)
Contribution from wastewater plants effluent
Contribution from upriver tributaries
Contribution from CSOs and storm water
Early 1970s
Mid 1990s
49
43
61%
37%
2%
53%
42%
5%
6.1
4.8
82%
15%
3%
77%
19%
4%
Note: the input from the ocean and downstream aquatic ecosystems are not included; see
text for derivation of estimates.
Table 3. Characteristics of the Mohawk and Upper Hudson River basins (data from
Boyer et al. 2002).
Mohawk River
Basin
Area (km2)
Population density (# km-2)
Land Use
forested
agriculture
urban
8,935
54
Upper Hudson
River Basin
11,942
32
Combined Mohawk
and Upper Hudson
River Basins
20,877
41
63%
28%
5%
81%
10%
3%
73%
18%
4%
Nitrogen export (kg N km-2 y-1)
Nitrogen export (103 tons y-1)
795
7.1
502
6.0
627
13.1
N export from deposition
N export from agriculture
(%fertilizer)
(% agricultural N fixation)
31%
48%
(12%)
(36%)
52%
29%
(10%)
(19%)
41%
39%
(11%)
(28%)
Table 4. Average effluent concentrations and costs for sewage treatment systems in the
United States (data from NRC 1993).
Treatment
System
BOD
(g C m-3)
TN
(g N m-3)
TP
(g P m-3)
Operating
costs
($ m-3)
Capital
costs
($ m-3)
Total
costs
($ m-3)
No treatment
(Raw)
76
30
6
---
---
---
Primary
52
23
4
0.06
0.08
0.14
Secondary
6
19
3
0.14
0.14
0.28
Nutrient
Removal
5.6
3
1.5
0.15
0.22
0.37
NRC (1993) refers to secondary treatment plants as “biological” plants. BOD loads are
converted to units of labile organic carbon by assuming 1 mole of organic carbon
oxidized for every mole of O2 consumed. Costs are based on averages for facilities in the
United States, assuming an 8% interest rate, 20 year design period, and facilities designed
to handle 72.6 x 103 m3 d-1 of effluent. Land costs are not included. “Operating” costs
include maintenance and operating costs. Note that costs are for cumulative level of
treatment; secondary treatment includes primary treatment, and nutrient-reduction
treatment includes both secondary and primary treatment.
Table 5. Summary of nitrogen loading, estimated primary productivity, and inputs of
organic matter from wastewater and upstream tributary sources over time.
TN loading‡
(g N m-2 y-1)
Pre-European
Settlement
GPP †
(g C m-2 y-1)
Organic matter ‡
from CSOs
and wastewater
(g C m-2 y-1)
Total input
of labile
organic C
(g C m-2 y-1)
23
270
---
270
Early 1970s
330
225
370
595
1990s
295
850
94
944
87
480
50
530
§
Potential future
†
GPP estimates are for mesohaline estuary and are based on measured data for 1970s and
1990s and are estimated based on nutrient loading for pre-European settlement and
potential future. ‡TN and organic matter (BOD) loadings averaged over area of entire
saline estuary. §Potential future assumes complete conversion to nutrient-reduction
treatment for sewage treatment, elimination of CSO discharges, and a significant
reduction in nitrogen loading from upriver tributaries. See text for further details on
derivation of estimates.
Water Residence Time (d)
A.
5
4
3
2
1
0
200
400
600
0
200
400
600
800
1000
0
200
400
600
800
1000
0
800
1000
B.
GPP (g C m-2 d-1)
20
15
10
5
0
Depth at 1% light (m)
C.
7
6
5
4
3
2
1
0
Discharge (m3 sec-1)
Riverine N export (kg N km-2 yr -1)
2000
CHA
SCH
1500
BLA
1000
SUS
DEL
POT
MOH
HUD
500
AND
SAC
CON
MER
JAM
RAP
PEN KEN
y = 0.26x + 107; R2 = 0.62
0
0
1000
2000
3000
4000
5000
6000
Net anthropogenic N inputs (kg N km-2 yr -1)
Summer bottom dissolved oxygen (mg l-1)
7
6
5
4
y = 0.08x + 4.02
r 2 = 0.42
3
2
1
0
1985
1990
1995
2000