Environmental Pollution 173 (2013) 138e149
Contents lists available at SciVerse ScienceDirect
Environmental Pollution
journal homepage: www.elsevier.com/locate/envpol
Review
Evaluating nutrient impacts in urban watersheds: Challenges and research
opportunities
Richard O. Carey a, *,1, George J. Hochmuth a, Christopher J. Martinez b, Treavor H. Boyer c,
Michael D. Dukes b, Gurpal S. Toor d, John L. Cisar e
a
Soil and Water Science Department, University of Florida, PO Box 110510, Gainesville, FL 32611-0510, USA
Department of Agricultural and Biological Engineering, University of Florida, PO Box 110570, Gainesville, FL 32611-0570, USA
Department of Environmental Engineering Sciences, University of Florida, PO Box 116450, Gainesville, FL 32611-6450, USA
d
Soil and Water Science Department, Gulf Coast Research & Education Center, University of Florida, 14625 C.R. 672, Wimauma, FL 33598, USA
e
Environmental Horticulture Department, Ft. Lauderdale Research and Education Center, University of Florida, Ft. Lauderdale, FL 33314, USA
b
c
a r t i c l e i n f o
a b s t r a c t
Article history:
Received 18 July 2012
Received in revised form
11 October 2012
Accepted 17 October 2012
This literature review focuses on the prevalence of nitrogen and phosphorus in urban environments and
the complex relationships between land use and water quality. Extensive research in urban watersheds
has broadened our knowledge about point and non-point pollutant sources, but the fate of nutrients is
not completely understood. For example, it is not known how long-term nutrient cycling processes in
turfgrass landscapes influence nitrogen retention rates or the relative atmospheric contribution to urban
nitrogen exports. The effect of prolonged reclaimed water irrigation is also unknown. Stable isotopes
have been used to trace pollutants, but distinguishing sources (e.g., fertilizers, wastewater, etc.) can be
difficult. Identifying pollutant sources may aid our understanding of harmful algal blooms because the
extent of the relationship between urban nutrient sources and algal blooms is unclear. Further research
on the delivery and fate of nutrients within urban watersheds is needed to address manageable water
quality impacts.
Ó 2012 Elsevier Ltd. All rights reserved.
Keywords:
Point source
Non-point source
Reclaimed water
Septic systems
Wastewater treatment facilities
Isotopic analyses
Stormwater
1. Introduction
Urban watersheds are unique environments with characteristic
disturbance gradients that alter natural biogeochemical cycles
(Paul and Meyer, 2001; Beck, 2005; Kaye et al., 2006). Human
activities control major factors driving these cycles, including land
use change and soil variability, atmospheric chemistry, and
hydrologic modifications (Kaye et al., 2006). Nitrogen (N) and
phosphorus (P) are derived from multiple sources and pathways in
urbanized watersheds and thus the relative proportion and spatial
configuration of urban land use affect nutrient inputs to surface and
groundwater (Basnyat et al., 1999; Groffman et al., 2004; Brett et al.,
2005; Carey et al., 2011b). Common pollutant sources include
stormwater runoff, atmospheric deposition, and wastewater
* Corresponding author.
E-mail address: [email protected] (R.O. Carey).
1
Present address: Department of Natural Resources and Environment, Earth
Systems Research Center, Institute for the Study of Earth, Oceans, and Space,
University of New Hampshire, Durham, NH 03824, USA.
0269-7491/$ e see front matter Ó 2012 Elsevier Ltd. All rights reserved.
http://dx.doi.org/10.1016/j.envpol.2012.10.004
treatment systems (Grimshaw and Dolske, 2002; Schueler, 2003;
Andersen et al., 2004). In urban watersheds, impervious surfaces
(e.g., roads, driveways, walkways, etc.), stormwater management
projects, and artificial drainage systems (e.g., canals), disrupt
natural hydrological pathways and may enhance nutrient transport
(Arnold and Gibbons, 1996; Caccia and Boyer, 2007; Bell and Moss,
2008; NRC, 2008). Nitrogen and P are critical to the ecological
health of aquatic ecosystems, but excessive nutrient loading leads
to undesirable consequences such as cultural eutrophication (Smith
et al., 1999; Pinckney et al., 2001; Conley et al., 2009). Characteristic
problems associated with eutrophic systems include the development of toxic and non-toxic algal blooms (Glibert et al., 2006;
Bricker et al., 2008; Heisler et al., 2008). Algal blooms can limit light
to submerged aquatic vegetation, reduce water transparency, and
produce hypoxic or anoxic conditions (i.e., “dead” zones) that have
adverse effects on fish populations (Smith et al., 1999; Heisler et al.,
2008).
Developing strategies to reduce overall nutrient exports from
urban watersheds require an assessment of relative contributions
from nutrient sources and pollutant transport mechanisms
(Basnyat et al., 1999; Carey et al., 2011a, 2011b). Key aspects of
R.O. Carey et al. / Environmental Pollution 173 (2013) 138e149
urban water quality research include analyzing land use-water
quality relationships and identifying specific nutrient sources
(Beaulac and Reckhow, 1982; Wickham and Wade, 2002; Wickham
et al., 2002). Stable isotope analysis has been used to trace pollutants, especially N and P inputs that can potentially impair the
ability of water resources to meet designated uses (e.g., fishing,
recreation, etc.) (Widory et al., 2005; McLaughlin et al., 2006;
Kendall et al., 2007). However, the relationship between nutrient
inputs and water quality impacts is complex; algal blooms may
occur independent of human influences and elevated nutrients
may not produce blooms (Steidinger et al., 1999; Lenes et al., 2001;
Walsh and Steidinger, 2001; Heisler et al., 2008).
The objective of this review is to synthesize current knowledge
on major urban pollutant sources delivering nutrients to aquatic
systems and the challenges of evaluating water quality relationships in urban environments. Carey et al. (2012a) reviewed turfgrass fertilizer management practices and the implications for
urban water quality. Turfgrass fertilization is briefly discussed in
this paper within the context of other nutrient sources. Specific
aspects of urban water quality discussed include: (1) watershed
nutrient sources; (2) isotopic analyses to distinguish nutrient
sources; (3) the effect of landscape composition and configuration
on nutrient exports; and (4) the relationship between watershed
nutrient exports and water quality impacts, using the development
of algal blooms as an example. Research gaps pertaining to these
issues are also identified.
2. Urban watersheds: non-point sources
2.1. Stormwater runoff and leaching
Precipitation events that occur on impervious surfaces, or that
exceed the infiltration or saturation capacity of soils, produce
stormwater runoff. Typical nutrient concentrations in urban
stormwater runoff in the U.S. are 2.0 mg L1 for total N (TN) and
0.26 mg L1 for total P (TP) (Schueler, 2003). A variety of factors in
urban watersheds contribute to nutrient concentrations in runoff.
Limited data exist on nutrient inputs from pet waste, but this can be
substantial (Baker et al., 2001; Groffman et al., 2004; Fissore et al.,
2012). Estimated N inputs from pet waste (17 kg N ha1 yr1) in
a suburban watershed in Baltimore, Maryland exceeded contributions from fertilizers (14.4 kg N ha1 yr1) and atmospheric
deposition (11.2 kg N ha1 yr1) (Baker et al., 2001; Groffman et al.,
2004). Pet waste in the Minneapolis-Saint Paul, Minnesota region
represented 84% of P inputs due to phosphate (PO4) fertilizer
restrictions (Fissore et al., 2012).
Fertilizer restrictions are intended to improve water quality
because fertilizer management practices can lead to nutrient
exports from turfgrass and landscape plants during precipitation
events and melting snow. Management practices that influence
nutrient retention include the type (i.e., soluble or controlledrelease), rate, and timing of fertilization (Engelsjord and Singh,
1997; Easton and Petrovic, 2004; Shober et al., 2010; Carey et al.,
2012a). Established (mature) turfgrass receiving appropriate
fertilizer applications typically leach less than 5% of applied N
(Barton and Colmer, 2006). However, recycling grass clippings
without adjusting (downward) fertilizer rates increases the
potential for nutrient losses (Kopp and Guillard, 2002). Qian et al.
(2003) simulated the long-term effects of returning clippings to
Kentucky bluegrass grown on a clay loam soil while continuously
using a fertilization rate of 150 kg N ha1 yr1. Leaching rates under
this scenario were minimal (<2 kg N ha1 yr1) 20e30 years after
establishment, but gradually increased (50e60 kg N ha1 yr1 after
100 years) as carbon and N sequestration declined. In addition to
fertilization, irrigation practices (e.g., rate, frequency, etc.), species
139
variability (e.g., cool and warm season grasses, nutrient uptake
efficiencies, etc.), and soil characteristics (e.g., texture, structure,
etc.) determine whether turfgrass retains or exports nutrients
(Trenholm et al., 1998; Barton and Colmer, 2006; Bowman et al.,
2002). For example, turfgrass with dense ground cover reduces
sediment and P losses in runoff (Linde et al., 1995).
Urban watersheds now increasingly include green roofs and
fertilizer applications, plant species, rainfall rates, and substrate
characteristics are among several factors that influence pollutant
loads from these systems (Berndtsson, 2010; Rowe, 2011). Green
roofs collect atmospheric pollutants, attenuate heat island effects,
and reduce runoff volumes compared to non-vegetated roofs, but
function as pollutant sources or sinks during precipitation events.
Berndtsson et al. (2006) revealed differences in nutrient exports
from an extensive green roof system (0.95 ha) after comparing
annual TN and TP loads in precipitation (TN: 9.1 kg ha1 yr1; TP:
0.2 kg ha1 yr1) and runoff (TN: 3.8 kg ha1 yr1; TP:
1 kg ha1 yr1). Green roofs located in areas receiving intense
storms are particularly susceptible to nutrient exports (NRC, 2008).
Hathaway et al. (2008) demonstrated that the growth media used
for green roofs contributes to nutrient losses. Green roofs containing 15% compost leached nutrients and contributed to greater
TN and TP outflow concentrations compared to rainfall alone or
non-vegetated roofs. The type of fertilizers used on green roofs
during the initial establishment period also contributes to nutrient
losses (Berndtsson et al., 2006).
Landfills accumulate nutrients from various waste products (e.g.,
lawn and household waste, biosolids, etc.) and are another source of
urban pollutants (Kjeldsen et al., 2002; Louis, 2004; Renou et al.,
2008). Precipitation percolating through waste layers leads to
various physical, chemical, and microbial processes that generate
leachate through interactions between landfill constituents and
groundwater (Christensen et al., 2001). Factors affecting the quantity and quality of landfill leachate flow include landfill age, design
specifications such as liners, the degree of waste compaction,
climate variability, and the inherent properties of waste products
(Renou et al., 2008). Carey et al. (2012b) described regulatory and
resource management practices associated with nutrient exports
from municipal solid waste facilities such as landfills.
Construction activities (e.g., new subdivisions, commercial
centers, highways, etc.) contribute nutrients to urban stormwater
runoff as elevated erosion rates during construction facilitate the
transport of sediment-bound P (Carpenter et al., 1998; Atasoy et al.,
2006). Line et al. (2002) reported sediment export during the
clearing and grading phase of construction (referred to as
a construction-I site) in North Carolina was 10 times greater than
other land uses (single-family residential, golf course, dairy cow
pasture, etc.). However, average nutrient exports from a construction-II site (TN: 36.3 kg ha1 yr1; TP: 1.3 kg ha1 yr1), which
includes drainage infrastructure installation and the housebuilding phase, differed from the construction-I site (TN:
8.3 kg ha1 yr1; TP: 3.0 kg ha1 yr1). Total N exports from
a residential area (23.9 kg ha1 yr1) and golf course
(31.2 kg ha1 yr1) were similar to the construction-II site, but TN
rates from a pasture (6.7 kg ha1 yr1) and wooded area
(11.4 kg ha1 yr1) were similar to the construction-I site (Line
et al., 2002). In another study, sediment export from a developing
subdivision was 95% greater than forested/agricultural areas (Line
and White, 2007). Both TN (15.5 kg ha1 yr1) and TP
(1.3 kg ha1 yr1) loads from the subdivision exceeded rates from
the undeveloped areas (TN: 6.3 kg ha1 yr1; TP: 0.5 kg ha1 yr1).
Local conditions (e.g., climate, soil, and topographic characteristics)
affect sediment export from construction sites. Burton and Pitt
(2001) noted that intense rainfall, soil erodibility, and steep
terrain produced high erosion rates in Birmingham, Alabama,
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R.O. Carey et al. / Environmental Pollution 173 (2013) 138e149
which amplified the impact of even small construction projects.
Stormwater pollutant exports associated with construction sites
may continue for several years until soils become stabilized postdevelopment and then impervious surfaces dictate runoff characteristics (NRC, 2008).
2.2. Atmospheric deposition
Atmospheric deposition can be a source of both N and P to
surface waters, but N deposition is greater in magnitude than P; the
vast majority (90%) of P deposition from air is due to wind-eroded
particles (Smil, 2000). Monitoring programs in the U.S. (e.g.,
National Atmospheric Deposition Program) and Europe (e.g.,
European Monitoring and Evaluation Programme) measure N wet
deposition (precipitation) rates, but dry deposition (gases and
particles) of N continues to be difficult to measure (Holland et al.,
2005). Only a fraction of dry deposition is usually measured, with
some gases (e.g., NO, NO2, NH3, etc.) excluded from dry deposition
monitoring programs (Pyke et al., 2008).
Land use affects atmospheric deposition because agricultural
activities such as livestock operations produce reduced N (e.g.,
NH3eN) emissions. Fossil fuel combustion produces N oxides (NO
and NO2; called NOx), which are converted to nitric acid and nitrate
aerosols. These atmospheric conversion processes enhance
pollutant transport and increase the potential impact radius.
Reduced N (NH3eN) is deposited much more rapidly than oxidized
N, thereby limiting the impact of agricultural emissions to local
areas (Spokes and Jickells, 2005). Catalytic converters in vehicles
also release NH3eN as a byproduct of the process to convert NOx to
N2; NH3eN is released in reducing conditions (e.g., high fuel to air
ratio during acceleration), producing elevated NH3eN deposition
rates along major roads (Kirchner et al., 2005; Bernhardt et al.,
2008). Maestre and Pitt (2005) provided further evidence of the
relationship between automobiles and localized NH3eN deposition
as freeway runoff contained the highest NH3eN concentrations
among various U.S. land uses. Emission rates vary for different types
of vehicles but range from 10 to 155 mg NH3eN km1 of roads
(Emmenegger et al., 2004).
The potential contribution from runoff highlights the fact that
both direct deposition onto water surfaces and deposition within
watersheds affect receiving waters. For estuaries with a surface
water to watershed area ratio greater than 0.2, direct deposition
typically contributes at least 20% of the TN loads. If this ratio is below
0.1, less than 10% of TN loads derive from direct deposition (Valigura
et al., 2000). Poor (2002) estimated 816,466 kg N were directly
deposited onto the surface of Tampa Bay, Florida in 2001. From 1999
to 2003, atmospheric deposition contributed 21% of N loads to the
estuary (Anderson, 2006). In contrast to Tampa Bay, Caccia and
Boyer (2007) estimated only 12% of N loads to Biscayne Bay, Florida were derived from direct atmospheric inputs, with the remaining 88% of N inputs from canals. However, Caccia and Boyer (2007)
did not evaluate the contribution of atmospheric deposits on landscape surfaces to canal loads. In a study conducted in Phoenix, Arizona, Hope et al. (2004) suggested parking lot surfaces could
accumulate dry-deposited nutrients in arid urban watersheds
between rainfall events and contribute to nutrient loading during
runoff, although nitrateeN (NO3eN) and ammoniumeN (NH4eN)
could be transformed prior to entering surface waters. For
example, in the Chesapeake Bay watershed, only 22% of atmospheric
N deposition is transported to the estuary (Castro et al., 2003).
2.3. Septic systems
Onsite sewage treatment and disposal systems, which include
septic systems, have been used in the U.S. since the late 19th
century. The U.S. Census Bureau officially began counting the
number of homes using these systems in 1960, when 14 million
homes were identified (Rome, 2001). Later reports included estimates from 1985 (24.6 million) and 2007 (26.1 million), with the
2007 estimate representing 20% of all U.S. homes (USEPA, 2008).
Rural and suburban areas often lack centralized wastewater treatment, thereby influencing the distribution of onsite treatment
systems. In 2007, 97% of U.S. housing units with septic systems
were located in either rural (50%) or suburban settings (47%)
(USEPA, 2008). Septic system densities can be particularly high in
areas experiencing a surge in new development (Marella, 2004).
Groundwater contamination caused by N loading from septic
systems is a primary concern (Gold et al., 1990; Wernick et al.,
1998). Subsurface septic systems treat wastewater by using septic
tanks to separate liquid and solid components before discharging
effluent to surrounding absorption systems or drainfields containing distribution pipes. Septic tanks discharge approximately
280 L per capita d1 and effluent concentrations (40e80 mg TN L1)
typically contain a mixture of 75% NH4eN and 25% organic-N
(Novotny and Chesters, 1981; Novotny et al., 1989). As effluent
percolates through the soil beneath the drainfield, NH4eN and
organic-N are oxidized in the unsaturated zone to NO3eN. The N
removal efficiency of septic systems determines the potential for
NO3eN loading to groundwater. Conventional systems remove 10e
44% of TN from wastewater, but alternative systems with anaerobic
up-flow filters (40e75%) and recirculating sand filters (60e85%)
remove more TN by enhancing nitrificationedenitrification
processes (USEPA, 1993; NRC, 2000). Soil characteristics (e.g.,
temperature, density, conductivity, etc.) are additional factors
influencing N loading rates to groundwater (NRC, 2000; Bernhardt
et al., 2008).
Septic system failure rates can range from 5 to 40% and include
hydraulic, subsurface, and treatment malfunctions (Swann, 2001).
Hydraulic failures reflect clogged systems, subsurface failures
produce partially treated wastewater plumes, and treatment failures occur when pollutants are not sufficiently removed from
wastewater. Pollutants from septic systems enter receiving waters
as a result of hydraulic failures that cause effluent to emerge onto
the surface of drainfields or through subsurface transport to
groundwater. High water tables may also interfere with the effectiveness of septic systems to treat wastewater because submerged
drainage networks become directly linked to surrounding water
systems (Bocca et al., 2007). Coastal communities with naturally
high water tables consequently have an increased risk of water
quality impairment from septic systems.
If conditions facilitate NO3eN transport, N delivery rates from
septic systems can be similar to inputs from row crop production.
Gold et al. (1990) measured NO3eN in shallow groundwater from
various land uses and reported values for fertilized cornfields
(66 kg ha1 yr1), septic systems (48 kg ha1 yr1), fertilized lawns
(6 kg ha1 yr1), and unfertilized lawns (1.4 kg ha1 yr1). Morgan
et al. (2007) reported NO3eN loadings from on-site wastewater
systems in La Pine, Oregon increased from 1700 kg N yr1 in 1960 to
41,000 kg N yr1 in 2005, due to septic tanks associated with
residential development. Additional residential lots are available
for development in this area and NO3eN loading to groundwater
could increase to 68,000 kg N yr1 by 2019 (Morgan et al., 2007).
Groundwater N derived from septic systems eventually
contributes to surface water nutrient inputs. Wernick et al. (1998)
evaluated water quality in Vancouver, Canada where increased
NO3eN concentrations in streams corresponded with watershed
septic system densities. For estuaries and watersheds, septic
systems can be a major component of overall N budgets (Fig. 1).
Buttermilk Bay, Massachusetts (74%), Buzzards Bay, Massachusetts
(16%), Charleston Harbor, South Carolina (15%), and Narragansett
R.O. Carey et al. / Environmental Pollution 173 (2013) 138e149
Bay, Rhode Island (11%) all receive TN inputs derived from septic
systems (Horsley and Witten, 1994; Castro et al., 2003). In coastal
communities of Taylor County, Florida, where there is a heavy
reliance on septic systems, the occurrence of both elevated bacterial and TN concentrations in coastal waters suggests that septic
systems can also contribute more nutrients than runoff in certain
local areas (Bocca et al., 2007).
Compared to N, P delivery to ground and surface waters is less of
a concern because of the ability of most soils to retain P. Soil
mineralogy determines P removal efficiency as clay minerals and
metal oxides (iron and aluminum) enhance PO4 sorption, while
calcareous soils remove PO4 through precipitation reactions (Jones
and Lee, 1979). Additional factors affecting the potential for P
contamination include soil pH and the distance between septic
systems and the water table. Soils can retain up to 95% of P, with
adsorption and precipitation reactions occurring within a few
meters of the drainfield (Mandel and Haith, 1992). For example,
soils rich in calcium carbonate can limit P migration (Lapointe et al.,
1990; Meeroff et al., 2008). Conditions that may reduce nutrient
retention and enhance pollutant transport include septic systems
that are located immediately adjacent to water bodies, soils with
reduced adsorption capacities (e.g., sandy or P-saturated soils), and
elevated water tables (Jones and Lee, 1979; Swann, 2001; Briggs
et al., 2008; Meeroff et al., 2008).
Products that increase household P outputs, such as laundry and
dishwasher detergents containing PO4, provide more opportunities
for P transport. Early detergent formulations included P builders
141
(e.g., sodium tripolyphosphate) that removed water hardness and
enhanced cleaning, but increased P delivery to aquatic systems
(Jenkins et al., 1973; Kehoe, 1992; Litke, 1999). In 1967, at the peak of
P detergent use in the U.S., manufacturing processes consumed
approximately 220,000 metric tons of P (Litke, 1999). Several states
have since instituted PO4 detergent bans and the industry voluntarily stopped manufacturing domestic laundry detergents with PO4
in the 1990s. Both actions have reduced the contribution of laundry
detergents to urban watershed nutrient budgets (Hoffman and
Bishop, 1994; Litke, 1999). Statewide bans typically target laundry
detergents, but dishwasher detergents and household cleaning
products still contain PO4 (Litke, 1999). The acknowledgment that
PO4 in dishwasher detergents can also reduce water quality has led
to specific PO4 bans for these products in sixteen states (NYS, 2010).
3. Urban watersheds: point sources
3.1. Wastewater treatment plants
Centralized wastewater treatment plants (WWTPs) process
wastewater from multiple sources before discharging effluent to
surface and/or groundwater. Although WWTPs require discharge
permits and monitoring programs ensure compliance with regulatory standards, the nutrient content of discharged effluent can
vary considerably depending on the level of wastewater treatment.
Primary treatment targets large objects and suspended solids;
secondary treatment removes additional organic material that
Fig. 1. Relative nitrogen inputs by source to (a) several estuaries draining urban watersheds (Castro et al., 2003) and (b) receiving waters of the Wekiva Study Area, Central Florida
(Roeder, 2008). The wastewater category in (a) includes effluent from wastewater treatment plants (WWTPs) and septic systems.
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R.O. Carey et al. / Environmental Pollution 173 (2013) 138e149
remains in the effluent after primary treatment; and tertiary or
advanced treatment reduces nutrient and metal concentrations
(Tchobanoglous et al., 2003). Concentration ranges for TN in
untreated wastewater (20e70 mg L1), wastewater receiving
secondary treatment (15e35 mg L1), and wastewater receiving
advanced treatment (1e8 mg L1) reflect the variability of N
concentrations in discharged effluent (Asano et al., 2007;
Tchobanoglous et al., 2003). Total P concentrations in wastewater
receiving secondary treatment (4e10 mg L1) are also greater than
discharged effluent receiving advanced treatment (0.5e2 mg L1)
(Tchobanoglous et al., 2003; Asano et al., 2007).
Effluent from WWTPs can provide consistent nutrient loads to
receiving waters that exceed contributions from non-point sources
(Carey and Migliaccio, 2009). The discharge volume from WWTPs is
an important factor affecting overall nutrient loads. These facilities
can dominate (>70% by volume) downstream flows (Andersen
et al., 2004; Ekka et al., 2006) and deliver large loads even if
effluent contains low nutrient concentrations. During naturally
low-flow periods, effluent concentrations and discharge volumes
become an even greater concern for downstream areas (Andersen
et al., 2004). Important land use and water quality relationships
in urban watersheds may therefore be obscured by the overwhelming influence of WWTP discharges (Miltner et al., 2004). For
example, Andersen et al. (2004) compared stream water quality at
multiple sites in South Carolina and average NO3eN and soluble
reactive phosphorus (SRP) concentrations downstream from two
WWTPs (NO3eN: 50.5 mg L1 and SRP: 3.7 mg L1) were considerably higher than upstream measurements (NO3eN: 1.6 mg L1
and SRP: 0.3 mg L1).
The use of reclaimed (reuse) water from WWTPs may additionally contribute to nutrient exports from urban watersheds
because reuse applications (e.g., landscape irrigation, groundwater
recharge, and non-potable urban uses) apply varying levels of
wastewater and associated nutrients (Swancar, 1996; Asano et al.,
2007; Carey and Migliaccio, 2009). Conventional fertilizers are
often used with reclaimed water to balance nutrient ratios or to
supplement overall nutrient requirements for landscape plants
(Sala and Mujeriego, 2001; Martinez and Clark, 2009). Swancar
(1996) compared six pairs of golf courses irrigated with either
reclaimed water or groundwater and concluded that the type of
irrigation affects shallow groundwater quality due to higher
percentages of constituents such as NO3eN in reclaimed water.
Seasonal differences in nutrient uptake efficiencies may additionally lead to leaching or runoff. Effluent is commonly applied to
bermudagrass in the southern U.S. to utilize its ability to assimilate
nutrients, but Wherley et al. (2009) cautioned against year-round
dispersal due to differences in NO3eN uptake during the active
growth period in summer (>90%), fall and spring transition months
(80e90%), and winter dormancy (10e20%).
Landscape irrigation (e.g., golf courses, parks, residential areas,
etc.) is second to agricultural irrigation in terms of reclaimed water
use in the U.S. Utilizing reclaimed water has been an ongoing
priority in states such as Florida, where the total reuse capacity of
domestic WWTPs has increased from 1.4 Mm3 d1 in 1986 to
5.2 Mm3 d1 in 2006 (SFWMD, 2008). Dual distribution systems
that deliver both potable and reclaimed water have operated in St.
Petersburg, Florida since 1977 and residents were originally
encouraged to use un-metered reclaimed water for irrigation
(Okun, 2000; USEPA, 2004). The majority of operators distributing
reclaimed water now charge a fee to conserve this resource (FDEP,
2010), and this fee influences the volume of reclaimed water used
per customer (USEPA, 2004). Knowledge of the nutrient content of
reclaimed water, nutritional requirements of landscape plants, and
the effect of long-term applications may improve nutrient retention and assimilation when using reclaimed water for irrigation.
4. Distinguishing nutrient sources
Development of effective nutrient control measures requires an
assessment of contributing sources. Stable isotopes can help to
differentiate nutrient sources and explore pollutant pathways that
lead to water quality impairment (Heaton, 1986; Peterson and Fry,
1987; Mayer et al., 2002). Chemical properties for stable isotopes
used in ecological research include low atomic masses and large
relative mass differences for rare and abundant species (Sulzman,
2007). Both N (14N and 15N) and oxygen (16O, 17O, and 18O) stable
isotopes provide clues about nutrient sources. The 14N isotope
represents 99.64% of atmospheric N and 15N represents 0.36%.
Oxygen isotopes share a similar distribution (16O: 99.76%; 17O:
0.04%; and 18O: 0.20%). Isotopic ratios (15N/14N, 17O/16O, or 18O/16O)
for measured samples are reported in terms of d15N, d17O, and d18O
values, which are parts per thousand differences (&) from ratios in
accepted standards (N: atmosphere; oxygen: Vienna Standard
Mean Ocean Water). The following equation is used to determine
isotopic compositions (&) of samples:
d 15 Nsample or d 18 Osample
¼
Rsample Rstandard
.
Rstandard 1000
where R is the isotopic ratio for samples and standards.
Pollutant sources in the environment have characteristic
isotopic composition ranges. Typical d15N values for synthetic
fertilizers (0 4&) are lower than values for human or animal
wastewater (>þ10&) (Table 1) (Gormly and Spalding, 1979;
Heaton, 1986; Lapointe and Bedford, 2007). Synthetic fertilizers and
wastewater can combine to produce mixed-range d15N values (þ4.3
to þ8.7&), as measured in samples from several Florida springs
(Toth, 2003). Differences in d15N values among pollutant sources
arise because fractionation mechanisms change isotopic ratios
(Sulzman, 2007). Commercial fertilizer production occurs with
limited isotopic fractionation because of the industrial fixation of
atmospheric N, creating fertilizers with d15N values approaching
zero. In contrast, fractionation associated with the conversion of
urea to ammonia, as well as ammonia volatilization and coupled
nitrification and denitrification processes, contribute to elevated
d15N values from septic systems and WWTPs (Heaton, 1986).
However, wastes from human and animal sources have similar d15N
values (Table 1) (Curt et al., 2004). Organic fertilizers (e.g., from
animal sources) can have d15N values (þ2 to þ30&) that are
indistinguishable from wastewater (Kendall et al., 2007). Characteristic d15N values for N derived from natural soil organic N (3
to þ5&) and atmospheric deposition (10 to þ8&) also overlap
with values for synthetic fertilizers (Mayer et al., 2002). Relying
solely on N isotopes to identify pollutant sources is consequently
problematic.
Table 1
Sample nitrogen isotopic compositions (d15N) from different sources.
Nitrogen sourcea
Synthetic fertilizers
Ammonium nitrate
Ammonium sulfate
Urea
Animal waste
Swine manure
Poultry manure
Dairy cattle manure
WWTPs
Sludge
Effluent
a
Curt et al. (2004).
Mean d15N (&)
1.46 2.32
1.16 1.46
1.28 0.33
13.82 7.03
10.98 4.44
12.19 2.69
11.42 7.21
11.61 2.71
R.O. Carey et al. / Environmental Pollution 173 (2013) 138e149
Several researchers have used multiple isotopes to investigate N
transport, cycling, and resultant effects on water quality. Widory
et al. (2005) combined N and boron isotopes to trace pollutants
and identified WWTP effluent as the main source of NO3eN in
subsurface alluvial groundwater. Boron helped to identify this
source because of differences between d11B values for cattle
manure (þ11.6& to þ19.2&) and wastewater effluent
(1.8 0.1&) (Widory et al., 2005). A widely used technique for
tracing the origin of NO3eN involves calculating both d15N and d18O
values, due to typical isotopic ranges for NO3eN derived from
fertilizers, atmospheric deposition, and wastewater (Fig. 2). Both
15
N and 18O isotopes are also useful to determine the significance of
N cycling processes (e.g., denitrification or nitrification) on NO3eN
(Aravena et al., 1993; Mayer et al., 2002; Kendall et al., 2007).
Nitrogen cycling can mask nutrient sources by changing isotopic
ratios during processes such as denitrification. Kendall et al. (2007)
discussed several methods to account for both N cycling and mixing
among several pollutant sources, including the use of multiple
isotopes. For example, Aravena et al. (1993) differentiated NO3eN
in septic system plumes from fertilizer-derived concentrations in
groundwater using d15N and d18O. Aravena et al. (1993) also identified nitrification of NH4eN, from wastewater and fertilizers, as the
source of NO3eN.
Phosphorus is monoisotopic, unlike N, but inorganic and organic
PO4 can be traced indirectly using d18O of PO4 (d18OP). Gruau et al.
(2005) doubted the use of d18OP values to identify pollutant sources
because of the restricted isotopic difference between fertilizers
(þ19.6& to þ23.1&) and WWTP effluent (þ17.7& to þ18.1&).
Biogeochemical cycling of P, including processes such as adsorption
that may result in isotopic fractionation, and the rapid equilibration
between d18OP and d18O in ambient water, can result in d18OP values
that reflect equilibrium values, not the isotopic characteristics of
pollutant sources (Blake et al., 2005; Young et al., 2009). Systems
that are not P-limited and exhibit large differences between d18OP
and d18O in water, such as estuaries, create suitable conditions to
trace PO4 (McLaughlin et al., 2006; Kendall et al., 2007). Young et al.
(2009) compiled mean d18OP values for various sources, including
fertilizers (þ20.2&), aerosols (þ20.5&), detergents (þ16.9&),
vegetation (þ16.9&), and WWTPs (þ13.4&). Statistical comparisons (Student’s t-Test) of data from multiple studies indicated
fertilizers had significantly (p < 0.05) higher d18OP values than all
other sources except aerosols; d18OP values for WWTP effluent
were significantly lower than other sources (Young et al., 2009).
80
70
60
Atmospheric
Deposition
(‰)
40
18O
50
30
Trend During
Denitrification
Fertilizers
20
10
Nitrification
in Soils
0
Sewage and
Manure
-10
-20
-10
0
10
15N
20
30
40
(‰)
Fig. 2. Isotopic compositions (d15N and d18O) for nitrateenitrogen derived from
different sources (Mayer et al., 2002).
143
Combining stable isotope analysis with additional data and
techniques provides further options to distinguish sources. Bleifuss
et al. (2000) analyzed d15N values as well as dissolved oxygen and
sodium concentrations from public wells in New York and determined that residential, as opposed to agricultural, NO3eN sources
were primarily responsible for contaminated water. Isotopic
compositions in aquatic organisms have also been used as proxies
to trace N sources (Steffy and Kilham, 2004; Ulseth and Hershey,
2005; Vander Zanden et al., 2005). Nash and Halliwell (2000) discussed several biological markers (e.g., sterols and phospholipids)
that could complement stable isotope techniques identifying PO4
and sediment sources.
Stable isotope analysis can be a useful tool to identify nutrient
sources contributing to water quality impairment (Heaton, 1986;
Peterson and Fry, 1987; Zohar et al., 2010). However, potential
pollutant sources often have overlapping isotopic ranges that can
complicate nutrient analysis (Mayer et al., 2002; Gruau et al., 2005;
Kendall et al., 2007). Fractionation mechanisms that occur during
nutrient cycling create additional complications by changing
typical isotopic characteristics of pollutants (Blake et al., 2005;
Young et al., 2009), though this could be used to identify mitigating
processes (e.g., denitrification). Utilizing multiple isotopes and
tracers increases the likelihood of accurately identifying pollutant
sources (Aravena et al., 1993; Widory et al., 2005; Kendall et al.,
2007; Kaushal et al., 2011). Questions still remain about spatial
and temporal influences on isotopic compositions from different
sources and techniques to distinguish multiple sources from the
same general category (e.g., atmospheric deposition). The emergence of in situ sensors for carbon and nutrients provides new
opportunities to distinguish sources by integrating multiple
instruments operating at fine temporal resolutions that could be
coupled with isotope sampling. Strategic use of in situ sensors can
potentially reveal simultaneous biogeochemical processes
(nutrient uptake, transformation, etc.) occurring during hydrologically variable conditions (Pellerin et al., 2011; Heffernan and
Cohen, 2010). Pairing sensors with techniques such as stable
isotope enrichment could expand currently limited data on the
source and fate of nutrients during storm events in developing
watersheds.
5. Landscape composition and configuration
Nutrient exports reflect the relative proportion and spatial
distribution of watershed land use/land cover (LULC) (Basnyat et al.,
1999; Groffman et al., 2004; Carey et al., 2011b). Anthropogenic
influences in urban and agricultural watersheds produce overall
water quality characteristics that contrast with predominantly
forested watersheds. During baseline conditions over a ten-year
period (1990e1999) in Seattle, Washington, average TN, TP, and
dissolved P concentrations for urban streams were 44%, 95%, and
122% greater, respectively, than forested streams (Brett et al., 2005).
Similar elevated nutrient discharges from agricultural watersheds
derive from variability associated with soils, crops, fertilizer applications, fertilizer storage facilities, and management practices such
as soil tillage methods (Beaulac and Reckhow,1982; Sims et al.,1998;
Nair et al., 2004). The greater distribution of watershed forest cover
relative to urban and/or agricultural land use consequently corresponds to decreased N and P exports (Beaulac and Reckhow, 1982;
Basnyat et al., 1999; Groffman et al., 2004). Wickham and Wade
(2002) used LULC nutrient export coefficients to assess the risk of
nutrient exports from small watersheds in Maryland. For watersheds with less than 70% forest cover, the relative proportion of
urban land use dramatically increased risk estimates for P (Wickham
and Wade, 2002). Wickham et al. (2002) modeled alternative land
use change scenarios in the mid-Atlantic region of the U.S. to identify
144
R.O. Carey et al. / Environmental Pollution 173 (2013) 138e149
areas most vulnerable to increased N and P exports. Areas with
a forested to agricultural ratio of 6:1, and projected urbanization
rates 20%, were vulnerable to increased N exports; at similar
urbanization rates, P vulnerability increased in areas with a 2:1
forested to agricultural ratio (Wickham et al., 2002).
Impervious surfaces in urban watersheds increase runoff
volumes and enhance nutrient transport compared to pervious
ground cover characterized by lower runoff coefficients. Unpaved
parking areas and other locations with compacted soils that hinder
infiltration are considered functionally impervious surfaces (NRC,
2008). The effect of watershed imperviousness on nutrient export
is particularly evident at higher, less frequent discharges (Shields
et al., 2008). Bannerman et al. (1993) identified streets as a major
source of pollutants in Wisconsin stormwater. Elevated runoff
volumes from impervious surfaces, compared to other land uses,
increased pollutant loads. Automobile-related sources (e.g., fluids
from parking lots, service stations, automobile exhaust, etc.)
contribute nutrients to stormwater systems and thus street traffic
volume is another factor affecting nutrient loads. Steuer et al.
(1997) investigated contamination sources in an urban watershed
draining to Lake Superior in Michigan and although nutrient runoff
concentrations from lawns (TN: 9.70 mg L1; TP: 2.33 mg L1) were
five to ten times higher than other areas, concentrations from hightraffic streets (TN: 2.95 mg L1; TP: 0.31 mg L1) were greater than
low-traffic streets (TN: 1.17 mg L1; TP: 0.14 mg L1). The type of
driveway in residential areas also influences nutrient export
because impervious driveways can generate greater runoff volumes
and nutrients loads compared to permeable paver driveways
(Gilbert and Clausen, 2006).
A conservative estimate of impervious cover for the conterminous U.S. in 2000 was 8 million ha and is projected to exceed 11
million ha by 2030 (Theobald et al., 2009). Total impervious area
(TIA) in developing watersheds is used as an indicator of aquatic
health because a threshold of 5e10% TIA can impair water quality
due to urbanization effects (Schiff and Benoit, 2007). Exum et al.
(2005) suggested 5e10% TIA produces modest impacts related to
urbanization that can be addressed through planning and watershed management. Urbanization leads to significant aquatic
degradation at 10e20% TIA, but the likelihood of successful remediation efforts to improve water quality under this scenario is
greater than watersheds exceeding 20% TIA (Exum et al., 2005).
In addition to total watershed imperviousness, another metric
to evaluate pollutant pathways is effective imperviousness, the
directly connected impervious area (DCIA) that is hydrologically
linked to drainage conveyance systems. The difference between TIA
and DCIA is that TIA includes surfaces such as roofs that can drain to
pervious areas (e.g., lawns), but runoff from DCIA will directly enter
drainage systems and/or receiving waters. In Wisconsin,
Bannerman et al. (1993) noted that streets were 100% connected to
the stormwater system while roofs were only 2% connected.
Watershed DCIA percentage consequently has a greater potential
than the equivalent TIA percentage to influence nutrient loads
exported in stormwater runoff (Brabec et al., 2002). Lee and Heaney
(2002) investigated the importance of DCIA percentage in a residential area of Miami, Florida by analyzing runoff-rainfall relationships during a 52-year period. Although DCIA constituted 44%
of the watershed in this study, DCIA generated 72% of the runoff
(Lee and Heaney, 2002). The DCIA percentage in urban watersheds
is a critical factor determining the extent of first-flush runoff effects
and nutrient removal mechanisms required to improve stormwater
quality (Harper and Baker, 2007).
Several researchers have investigated the relationship between
the spatial characteristics of LULC and water quality (Lowrance
et al., 1984; Basnyat et al., 1999; Lee et al., 2009). Carey et al.
(2011b) identified spatially explicit LULC variables, including the
proximity of specific types of urban development to aquatic
systems, which corresponded to increased nutrient exports.
Forested watersheds are particularly effective at reducing nutrient
exports because these areas function as active nutrient transformation zones or sinks (Lowrance et al., 1984; Basnyat et al.,
1999). Properly managed riparian zones in agricultural and urban
watersheds also reduce nutrient transport (Norris, 1993; Groffman
et al., 2002; Correll, 2005). However, hydrological modifications in
urban watersheds alter runoff characteristics, lower water table
depths, and reduce the efficiency of riparian zones (Groffman et al.,
2002). Several studies have investigated alternative methods to
improve water quality in urban watersheds by incorporating urban
forests or forested riparian zones into planning programs. Matteo
et al. (2006) used watershed simulations to evaluate the effect of
applying urban forestry best management practices (BMPs) to
riparian and roadside buffer areas in an urbanizing watershed.
Results from the study indicated that the selection and placement
of tree species affect the ability of increased forest cover to reduce
both sediment and nutrient loading.
In addition to urban forestry BMPs, structural BMPs (e.g., dry
ponds, constructed wetlands, bioretention systems, etc.) are
frequently used in urban watersheds to reduce nutrient exports
(NRC, 2000, 2008; Carey et al., 2012b). Nonstructural BMPs that
emphasize preventative management (e.g., riparian buffers, zoning
restrictions, educational programs, etc.) differ from structural BMPs
that target pollutants in stormwater runoff. Factors such as initial
stormwater concentrations influence variable nutrient removal
efficiencies among common structural BMPs in urban watersheds
(NRC, 2000; Carey et al., 2012b). Groffman and Crawford (2003)
suggested that managing stormwater control structures and
urban riparian areas to maximize soil moisture and organic matter
content would restore NO3eN sinks. Restoration of degraded
riparian zones requires careful attention to soil porosity and
organic carbon levels to promote high microbial activity and low
oxidation/reduction potential in groundwater (Correll, 2005).
The assortment of urban nutrient sources, coupled with riparian
zone degradation, directly affects relative retention in forested,
agricultural, and urban watersheds. In Baltimore, Maryland,
a forested watershed retained 95% of N inputs compared to 77% and
75% in agricultural and suburban watersheds, respectively
(Groffman et al., 2004). The reported N retention in the suburban
watershed may likely be much lower because only atmospheric and
fertilizer inputs were included in nutrient calculations (Groffman
et al., 2004). Castro et al. (2003) reported average atmospheric N
retention for upland forests (92%) and agricultural areas (79%)
along the U.S. Atlantic and Gulf Coast regions. In contrast, urban
watersheds retained only 21e60% of TN inputs (Castro et al., 2003).
Wollheim et al. (2005) also found greater retention in forested
areas (93e97%) compared to urban areas (65e85%) in Massachusetts watersheds.
Despite reduced overall N retention in urban watersheds,
potential N retention in appropriately managed turfgrass can be
significant. Groffman et al. (2009) suggested that similar carbon
cycling rates in turfgrass and forests produce N sinks in soils that
reduce NO3eN and nitrous oxide (N2O) losses. Raciti et al. (2008)
described N cycling processes for lawns that can lead to greater
atmospheric-N retention rates compared to forests. Agricultural
soils, with depleted soil carbon levels relative to turfgrass, were also
characterized by higher levels of soil NO3eN (Groffman et al.,
2009). Understanding the capacity for long-term N retention in
lawns, considering variables such as fertilizer management practices, soil characteristics, and climate, can provide further insight
into urban nutrient cycling (Raciti et al., 2008). For example,
climatic variability (e.g., drought conditions) can negatively affect N
retention rates in agricultural and urban areas much more than
R.O. Carey et al. / Environmental Pollution 173 (2013) 138e149
forested areas (Kaushal et al., 2008). Bijoor et al. (2008) also
demonstrated that warmer temperatures can lead to increased N2O
emissions from lawns.
6. Water quality impacts: harmful algal blooms
In recent decades, there have been increased incidences of
harmful algal blooms (HABs) e also called red or brown tides e that
persist for increasing periods and cover larger areas in aquatic
environments (NRC, 2000; Lapointe and Bedford, 2007; Heisler
et al., 2008). Studies investigating the formation and persistence
of near-shore and offshore HABs often do so within the context of
cumulative pollution derived from nutrient sources in urban
watersheds. Brand and Compton (2007) attributed the increased
prevalence of HABs in Florida’s near-shore waters in the past five
decades to population growth rates and land use disturbances,
Bricker et al. (2008) linked recent blooms in several U.S. estuaries to
population growth in coastal watersheds, and Heisler et al. (2008)
described several examples where improved wastewater treatment
reduced both the extent and occurrence of HABs. Urban populations are projected to increase in the future (United Nations,
2008), which suggests increased delivery of nutrients to aquatic
systems.
If sufficient nutrients are readily available and other growthlimiting conditions are satisfied, both non-toxic and toxic algae
can increase in density to produce HABs. The decomposition of
non-toxic macroalgal blooms leads to depleted dissolved oxygen
levels, the loss of submerged aquatic vegetation, and damaged coral
reefs. Toxic algal species additionally produce toxins that can be
fatal to fish and other aquatic organisms; these toxins also
contribute to neurological and respiratory problems in humans
(Kirkpatrick et al., 2004; Abbott et al., 2009). Toxic cyanobacteria,
the major class of freshwater HABs, produce cyanotoxins (e.g.,
microcystins and cylindrospermopsin) that can present multiple
problems if ingested by humans (Abbott et al., 2009), thus disrupting recreational activities and threatening potable water
resources (Dortch et al., 2008; Bigham et al., 2009; USEPA, 2009).
These toxic cyanobacteria blooms can also be difficult to remove
during water treatment (Westrick et al., 2010). Management
concerns related to HABs therefore include both restoration of
aquatic ecosystem productivity and protection of human health.
Determining specific aspects of HABs has been particularly
important in states with large coastal resources such as Florida,
where all major groups of HABs, including those that cause red
tides and poison fish, have occurred (Steidinger et al., 1999).
Overall, 70 (50 marine and 20 freshwater) harmful algal species
have been identified in Florida (Abbott et al., 2009). Karenia brevis
(also known as Ptychodiscus brevis and Gymnodinium breve) is
a dominant red tide species and its abundance along the southwestern coast of Florida has increased approximately 13e18 fold
from 1954e1963 to 1994e2002 (Brand and Compton, 2007).
Complete life cycle information and exact nutrient pathways for
K. brevis are unknown (Brand and Compton, 2007). Using isotopic
analyses of macroalgae, Barile (2004) investigated potential
nutrient sources contributing to HABs in east-central Florida nearshore waters and reported that ambient concentrations of
wastewater-derived N were sufficient to support K. brevis red tides.
Lapointe and Bedford (2007) also used d15N values to link watershed nutrient sources to rhodophyte blooms in Lee County, Florida.
Isotopic analysis of macroalgae indicated fertilizer-based N was
primarily responsible for HABs in the wet season and wastewater
sources supplied N in the dry season (Lapointe and Bedford, 2007).
Lapointe et al. (2004) suggested both regional agricultural runoff
and local wastewater discharges led to the subsequent development of HABs in the Florida Keys.
145
Nutrient enrichment as a result of human activities contributes
to the increased prevalence of HABs, but the extent of this relationship and the effect of additional factors are unclear. Microcystin
can occur in oligotrophic lakes and may be associated with seasonal
fluctuations in water temperature (Bigham et al., 2009). Toxic
blooms of K. brevis (Florida) and Alexandrium tamarense and
A. catenella (northeastern and northwestern U.S.) can develop miles
away from near-shore pollutant sources (Steidinger et al., 1999). In
addition, a food web link between K. brevis and the cyanobacterium
Trichodesmium in the Gulf of Mexico could stimulate the production
of red tides. Lenes et al. (2001) and Walsh and Steidinger (2001)
suggested that Saharan dust, transported in wind currents from
Africa, provides the iron necessary for N fixation by Trichodesmium,
and dissolved organic N (DON) excreted by the cyanobacterium
serves as the N source for K. brevis. Nitrogen release rates (>50% of
N fixed) from Trichodesmium in the form of DON and NH4eN may
also be sufficient to support other phytoplankton species
(Mulholland et al., 2004). Glibert et al. (2009) demonstrated
through laboratory experiments that K. brevis can satisfy 40% of its
cellular N requirements per hour by directly grazing on another
cyanobacterium, Synechococcus.
The ability of algal species to move toward nutrient sources, and
the effect of climate variations, further complicate the process of
identifying causative factors for HABs. Vertical migration (down to
90 m depths) of K. brevis allows this species to access subsurface
NH4eN and urea (>20 m) transported from Mississippi River plumes
in the northern Gulf of Mexico to the oligotrophic West Florida Shelf
(WFS) (Stumpf et al., 2008). Therefore, local nutrient sources may
not be the primary stimulus for the formation or persistence of high
density HABs. Vargo et al. (2008) suggested a combination of sources
could provide N and P necessary to sustain K. brevis blooms (up to18
months) in the WFS, including zooplankton excretion and fish
remineralization. In 2004, Florida experienced numerous hurricanes
and Hu et al. (2006) suggested that these events led to greater
surface runoff and submarine groundwater discharges that stimulated intense red tides. Contributions from delayed submarine
discharges may provide more nutrients than either riverine or
atmospheric sources (Hu et al., 2006). Warmer temperatures may
also promote the proliferation of toxic cyanobacteria blooms. Davis
et al. (2009) collected samples from northeastern U.S. lakes and
compared growth rates of toxic and non-toxic Microcystis using
different experimental treatments. Elevated P concentrations, along
with increased temperatures, produced the fastest overall growth
rate for the toxic species (Davis et al., 2009).
7. Conclusions
Optimizing watershed management practices to reduce nutrient
losses remains a critical objective to improve urban water quality.
Priority research areas include quantifying nutrient sources and
sinks in urban watersheds, especially understudied aspects such as
pet waste. Determining appropriate management strategies for
urban vegetation such as green roofs and turfgrass will increase
nutrient retention and limit exports. Turfgrass coverage is extensive
in urban watersheds, but long-term nutrient cycling processes in
these landscapes are not completely understood. Evaluating the
fate of atmospheric N, particularly on turfgrass, may enhance
current understanding of the relative significance of atmospheric
deposits on landscape surfaces.
Regulatory controls have reduced direct export of wastewater
into receiving waters, but this export remains an important
component of urban nutrient budgets. Both septic systems and
WWTPs can discharge significant nutrient loads to receiving
waters, depending on the efficiency and extent of wastewater
treatment. Identifying the range of attenuation factors (e.g.,
146
R.O. Carey et al. / Environmental Pollution 173 (2013) 138e149
dilution processes, nutrient transformations, etc.) associated with
septic system discharges is critical to assessing the long-term
capability of watersheds to assimilate nutrients from these sources. Furthermore, there is limited data available on the impact of
long-term reclaimed water use on nutrient cycling in urban
watersheds, especially considering variables such as nutrient
concentrations in water used for irrigation, vegetative nutrient
uptake efficiencies, and soil characteristics.
Continuous in situ sensors of water quality parameters such as
dissolved oxygen and temperature, and more recently, dissolved
organic matter, NO3eN, and PO4, enhance the prospect of investigating multiple ecosystem processes simultaneously and can aid
urban water quality research. Combining in situ sensors that
provide high frequency measurements of water quality and other
biogeochemical parameters with additional methods, such as
isotopic analyses, can reveal both nutrient sources and mechanisms
controlling nutrient cycling in urban watersheds. Improved
assessment of nutrients in urban watersheds will help to advance
our understanding of cultural eutrophication and the development
of algal blooms. The various pathways by which algae can access
nutrients and the processes that can either stimulate bloom
formation or reduce the likelihood of their occurrence in different
regions are unclear. Urban population projections suggest a greater
potential for increased nutrient loads in the future, but adaptive
watershed management requires targeted research on the range of
factors contributing to degraded water quality.
Acknowledgments
This paper was written with financial support from Scotts
Miracle-Gro, Marysville, Ohio, and from the University of Florida,
Institute of Food and Agricultural Science. The University of New
Hampshire Agricultural Experiment Station provided additional
funding. We would also like to thank reviewers of this paper for
their helpful suggestions and comments.
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