Figure 7-12 Long-term trends of the spatial distribution of total nitrogen in the Delaware estuary (mean of data from 1968-1970, 1978-1980, and 1988-1990). Source: Marino et al., 1991.
period has been shown to correspond to a concurrent increase in nitrate-N from the mid-1970s through the mid-1980s as a result of nitrification (Santoro, 1998; Marino et al., 1991).
Reflecting deforestation, agricultural practices, fossil fuel combustion, and the increase in human population of an increasingly urbanized drainage basin over the much longer time scale of a century, ambient nitrate and chloride levels (Figure 7-15) have steadily increased by approximately 400 percent to 500 percent since measurements were first recorded in 1905 at a water supply intake near Philadelphia (Ja-worksi and Hetling, 1996). Similar patterns of long-term increasing trends in ambient nitrate and chlorides have also been recorded at other East Coast water supply intakes for the Merrimack, Connecticut, Hudson, Schuykill, and Potomac Rivers (Ja-worski and Hetling, 1996).
Total phosphorus has also declined from peak levels of approximately 0.45 mg P/L during 1968-1970 to much lower levels of approximately 0.15 mg P/L by 1988— 1990 near River Mile 100 (Figure 7-13). An interannual time series of total phosphorus (Figure 7-14) for a station near Marcus Hook (River Mile 78) exhibits a trend similar to that of ammonia-N, with a sharp decline from approximately 0.8 mg P/L in the late 1960s to approximately 0.3 mg P/L by the late 1970s, followed by relatively unchanging concentrations (approximately 0.1 mg P/L) from the mid-1980s through the mid-1990s. The decline of ambient levels of total phosphorus has been attributed to the detergent phosphate ban of the early 1970s (Roman et al., 2000), reductions of effluent loads from wastewater facility upgrades (Sharp, 1988), and changes in partitioning of dissolved and soluble phases of phosphorus and changes in solubility of phosphate (Lebo and Sharp, 1993).
Figure 7-15 Long-term trends of chlorides and nitrate-N at a water supply intake in the tidal Delaware River near Philadelphia. Source: Jaworski and Hetling, 1996. Copyright © Water Environment Federation, Alexandria, VA. Reprinted with permission.
Evaluation of Water Quality Benefits Following Treatment Plant Upgrade
From a policy and planning perspective, the central question related to the effectiveness of the secondary treatment requirement of the 1972 CWA is simply: Would water quality standards for DO be attained if primary treatment levels were considered acceptable? In addition to the qualitative assessment of historical data, water quality models can provide a quantitative approach to evaluate improvements in DO and other water quality parameters achieved as a result of upgrades in wastewater treatment. Since the early 1960s four classes of water quality models, developed from the 1960s through the 1990s, have been applied to determine wasteload allocations for municipal and industrial dischargers to meet the needs for water quality management decisions for the Delaware estuary (Mooney et al., 1998).
During the 1960s, one-dimensional estuarine water quality models of DO and carbonaceous and nitrogenous BOD were developed by Thomann (1963), O'Connor et al., (1968), Pence et al. (1968), Jeglic and Pence (1968), and Feigner and Harris (1970). DRBC used a 1960s era model, known as the Delaware Estuary Comprehensive Study (DECS) model, to establish wasteload allocations for ultimate CBOD and nitrogenous BOD for the six zones of the Delaware estuary.
With funding available from the CWA Section 208 program during the 1970s, Clark et al. (1978) upgraded the kinetics of the DECS water quality model to incorporate nitrification and denitrification in a nitrogen cycle represented by organic nitrogen, ammonia, and nitrate + nitrite as state variables. The oxygen contribution by algal production and respiration was included as an empirical input term dependent on chlorophyll observations. Transport was provided to the water quality model with one-dimensional link-node hydrodynamics, and the 1970s-era model was identified as the Dynamic Estuary Model (DEM) (Mooney et al., 1998).
As a result of industrial and municipal waste treatment plant upgrades from primary to secondary levels of treatment during the late 1970s and early 1980s, the water quality model used for wasteload allocations was once again upgraded to reflect the reduced wasteloads and improvements in water quality conditions (Mooney et al., 1998). The model was upgraded from a one-dimensional (longitudinal) to a two-dimensional (longitudinal and lateral) representation of water quality and transport in the Delaware estuary. A two-dimensional hydrodynamic model was coupled with a water quality model that retained the kinetic framework of the one-dimensional model with kinetic coefficients adjusted to reflect changes in pollutant loading (LTI, 1985). The upgraded 1980s-era two-dimensional model (DEM-2D) was used to conduct a toxics analysis (Ambrose, 1987) and to reevaluate the wasteload allocations developed with the earlier models (DRBC, 1987).
Following the completion of the Delaware Estuary Use Attainability (DEL USA) Project (DRBC, 1989), a technical review of the two-dimensional DEM model recommended that a new time-variable model be developed to incorporate state-of-the-art advances, with a three-dimensional hydrodynamic model coupled to an advanced eu-trophication model framework (HydroQual, 1994). Using revised kinetic coefficients to reflect reductions in waste loads and improvements in water quality, the kinetics of the water quality framework were expanded to include a eutrophication submodel, nitrogen and phosphorus cycles, labile and refractory organic carbon, and particulate and dissolved fractions of organic carbon and nutrients (HydroQual, 1998; Mooney et al., 1999). Unlike the advanced eutrophication model developed for the Chesapeake Bay (Cerco and Cole, 1993), internal coupling of particulate organic matter deposition with sediment oxygen demand and benthic nutrient fluxes was not included in the upgraded framework; benthic fluxes were assigned as model input on the basis of monitoring data (HydroQual, 1998).
To evaluate the incremental improvements in water quality conditions that can be achieved by upgrading municipal wastewater facilities from primary to secondary and better-than-secondary levels of waste treatment, Lung (1991) used the 1970s-era one-dimensional DEM model (Clark et al., 1978) to demonstrate the water quality benefits attained by the secondary treatment requirements of the 1972 CWA. With this model, Lung used existing population and municipal and industrial wastewater flow and effluent loading data (ca. 1976) to compare water quality for summer flow conditions simulated with three management scenarios for municipal facilities: (1) primary effluent, (2) secondary effluent, and (3) existing wastewater loading. Water quality conditions for these alternatives were calibrated (Figure 7-16) using data for 1976, a
year characterized by average summer flow of the Delaware River (see Figure 7-3). Freshwater flow at Trenton, New Jersey, was 7,700 cfs; flow in the Schuykill River, a major tributary to the Delaware estuary, was 1,350 cfs for the 1976 calibration. Flow conditions during the summer of 1976 were 120 percent higher than the long-term (1951—1980) summer (July—September) mean streamflow of 5,986 cfs recorded at Trenton. Upstream of Trenton, flow releases from several impoundments along the free-flowing Delaware River are regulated to maintain the guideline for a minimum summer streamflow of 2,500 to 3,000 cfs at Trenton (Mooney et al., 1998).
Under the primary effluent assumption, water quality is noticeably deteriorated in comparison to the 1976 calibration results. DO concentrations are at a minimum about 35 miles downstream of Trenton, the traditional region of minimum DO levels. Under the primary scenario, an oxygen sag of 2 mg/L is computed by the model under summer (28°C), low-flow 7Q10 conditions (2,500 cfs for the Delaware at Trenton and 285 cfs for the Schuykill River at Philadelphia) (Figure 7-17).
Using the secondary effluent assumption, the reduction in ultimate CBOD loading significantly improves DO downstream of Philadelphia at the critical oxygen sag location (RM 96). In comparison to the primary scenario, minimum oxygen levels increased to almost 4 mg/L from approximately 2 mg/L under the secondary effluent scenario (Figure 7-18). To achieve compliance with a water quality standard of 5 mg/L, advanced waste treatment is required (Albert, 1997). As shown with the historical water quality data sets, the implementation of secondary and better than secondary levels of wastewater treatment has resulted in major improvements in DO, BOD5, ammonia, and total phosphorus conditions of the estuary (Figures 7-8 through 7-14). As demonstrated with the model, better than secondary treatment is required to achieve compliance with the water quality standard of 5 mg/L for DO downstream of Philadelphia. In contrast to the 1950s and 1960s, the historical occurrence of persistent and extreme low DO conditions has essentially been eliminated from the upper Delaware estuary. Improvements in suspended solids, heavy metals, and fecal col-iform bacteria levels have also been achieved as a result of upgrades in municipal and industrial wastewater treatment.
With vast tidal marshes and freshwater tributaries providing spawning and nursery grounds for abundant fishery resources, the coastal plain of the Delaware estuary provided a cornucopia of fishery and waterfowl resources important for sustenance to both Native American villages and colonial settlements. Historically, the estuary produced an enormous quantity of seafood from the early colonial era (ca. 1700s) through the early twentieth century. Colonial-era reports suggest schools of herring and sheepshead thick enough to walk on in a stream (Price et al., 1988). Abundant harvests of American shad and shortnose sturgeon provided important sustenance to the growing population of the Delaware valley for about 200 years.
Since the mid-1900s, however, the abundance of these, and other, species has declined dramatically as a result of urbanization and industrialization of the drainage basin. Deterioration in water quality (e.g., severe oxygen depletion), overfishing,
construction of dams, and habitat destruction have all contributed to the decline of the river's fisheries resources, beginning around the turn of the twentieth century (Majumdar et al., 1988). Massive fish kills were a frequent occurrence along the river from about 1900 through 1970 (Albert, 1988). Former wetlands and tributaries, crit-
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