During the last 20 years, land use in the Great Plains portion of the St. Vrain Creek basin, in north-central Colorado (Figure 1), has shifted substantially from largely undeveloped native grassland, pasture, and agricultural land to urban and suburban land, particularly in and around the cities of Longmont and Boulder, Colorado. The increase in human population associated with this shift has increased volumes of treated wastewater effluent that is discharged to area streams. Additionally, between 1986 and 1998 impervious surfaces increased 32% in the region, whereas irrigated cropland decreased by 33% and wetlands decreased by 65% (American Forests, 2001). Typically, changes in land cover can lead to increased urban runoff, decreased natural attenuation of discharge and filtering by wetlands, alterations in discharge conditions and water quality (Sprague et al., 2006), and ultimately degraded biological communities (Paul and Meyer, 2001).
During the last two decades new wastewater treatment plant (WWTP) technologies and instream restoration efforts have been implemented by the city of Longmont to improve aquatic habitat and water quality in St. Vrain Creek (Zuellig et al., 2007). The WWTP improvements primarily addressed facility capacity and the reduction of elevated metals, nutrients, and suspended-sediment concentrations. Recent regional and national reconnaissance studies have indicated the presence of previously undocumented contaminants indicative of human sources downstream from WWTPs (Ternes, 2001; Kolpin et al., 2002; Heberer and Adam, 2005; Sprague and Battaglin, 2005). These wastewater-related contaminants include antioxidants, detergents and detergent metabolites, disinfectants, fire retardants, fragrances, insect repellants, pharmaceuticals (prescription and nonprescription drugs), pesticides, plasticizers, polycyclic aromatic hydrocarbons, and steroidal compounds and are referred to hereafter as “emerging contaminants” (ECs). ECs can be released to the aquatic environment through industrial and municipal wastewater discharges, storm drains, agricultural and urban runoff, and individual or multi-facility sewage disposal systems. Until recently, the extent to which these contaminants occurred in the aquatic environment was not well known and the toxicological ramifications with regard to humans or wildlife were largely unknown (Daughton, 2001). However, recent studies have shown that exposure to some ECs, even at very low concentrations, can result in endocrine disruption and histological and immunological alterations in wildlife and humans (Colborn et al., 1993; McLachlan, 2001; Petrovic et al., 2002; Höger, 2003; Bernet et al., 2004; Hoeger et al., 2004; Arslan et al., 2007). Identifying the occurrence, distribution, and fate of ECs in urban streams will aid communities in addressing source-control and reduction efforts to safeguard human and aquatic health.
In this paper we describe the occurrence and transport of selected ECs in St. Vrain Creek through the city of Longmont under two different hydrologic conditions. Stream samples were collected during the two events by using a longitudinal Lagrangian sampling design. Dye tracer studies conducted just prior to each sampling event were used to estimate travel times during each event. Discharge and field measurements were collected and water samples were analyzed by the U.S. Geological Survey (USGS) National Water Quality Laboratory (NWQL) for a series of ECs that are indicative of wastewater by using methods described in Zaugg et al. (2002).
Study Area and Site Selection
St. Vrain Creek flows east from sources along the east side of the Continental Divide and eventually joins the South Platte River, in north-central Colorado (Figure 1). The main mountainous headwater streams, North and South St. Vrain Creeks, are primarily forested until they converge downstream near the town of Lyons to form St. Vrain Creek. Downstream from Lyons, St. Vrain Creek primarily flows through grassland, pastures, and agricultural areas and the city of Longmont on its way to the confluence with the South Platte River.
Seven sites were selected for water-chemistry sampling along a 13.8-kilometer (km) reach of the St. Vrain Creek within the city of Longmont: two sites upstream from the Longmont WWTP outfall (Sites 1 and 2); two sites at the mouths of key tributaries, Left Hand Creek and Boulder Creek (Sites 3 and 6); one site at the Longmont WWTP outfall (Site 4); and two sites on St. Vrain Creek downstream from the WWTP outfall (Sites 5 and 7) (Figure 1). The most upstream site (Site 1) is located at the western edge of Longmont just east (downstream) from Airport Road, and it represents the upstream, primarily nonurban inputs to the creek before it enters the Longmont area. This site is influenced by the small town of Lyons approximately 13.2 km upstream and the surrounding agricultural community. Site 2 is approximately 5.8 km downstream from Site 1 and immediately upstream from the Longmont WWTP outfall; it represents inputs from the urban and agricultural areas before the input of treated wastewater effluent. Left Hand Creek (Site 3) is one of the two largest tributaries entering St. Vrain Creek in the study reach. Left Hand Creek enters St. Vrain Creek just upstream from the Longmont WWTP outfall; it is influenced by the upstream communities of Jamestown and Ward, some historical mining, and suburban development and agricultural activities around Longmont. Site 4 is the WWTP outfall and the flow is composed of treated wastewater effluent from the City of Longmont Water and Wastewater Department. Site 5 is approximately 1 km downstream from the Longmont WWTP outfall and represents the integration of the WWTP effluent and St. Vrain Creek (tributaries do not intervene between the outfall and this site). Boulder Creek (Site 6) is the other major tributary and it enters St. Vrain Creek approximately 7.5 km downstream from the Longmont WWTP outfall and is influenced by upstream operations of Barker Dam, the town of Nederland, historical mining, and WWTP inflows from the urban areas of Boulder, Superior, Louisville, Lafayette, and Erie, and by the surrounding suburban and agricultural land. The WWTP for Boulder is approximately 23 km upstream from Site 6. Site 7 is just downstream from the confluence of Boulder Creek, approximately 7.9 km downstream from the Longmont WWTP and approximately 72.6 km from the confluence with the South Platte River.
A Lagrangian sampling design, which follows the same parcel of water as it moves downstream, was used for each sampling event (Zuellig et al., 2007). Tracer tests with Rhodamine-WT dye were used to determine the time-of-travel between sample-collection sites (Kilpatrick and Wilson, 1989). For this study, time-of-travel was defined as “the amount of elapsed time for the dye peak to travel between two monitoring sites” (Zuellig et al., 2007). Travel-time estimates were made 9 and 15 days prior to the date of sample collection in April 2005 and March 2006, respectively. Minor adjustments were made to the dye travel-time estimates to account for differences in flow conditions during the times of sample collection (Zuellig et al., 2007). In 2005, under higher flow conditions than in 2006, times of travel were estimated as 345 min between Sites 1 and 2, 65 min between Sites 2 and 5, and 235 min between Sites 5 and 7. In 2006, times of travel were estimated as 825 min between Sites 1 and 2, 45 min between Sites 2 and 5, and 290 min between Sites 5 and 7. Stream velocity and travel times can increase or decrease during lower flow conditions depending on stream channel morphology. Uncertainty in the estimates of travel time can be introduced by errors in the measurement of dye concentrations.
Stream samples analyzed for ECs were collected at specific times calculated from the travel-time data using standardized depth- and width-integrating techniques and processed and preserved on-site using methods described in the USGS National Field Manual (variously dated). Samples were analyzed by the USGS NWQL for 61 ECs (Table 1) according to methods described in Zaugg et al. (2002). Field measurements, including dissolved oxygen, pH, specific conductance, water temperature, and discharge, were obtained at the time of sample collection. The complete dataset is presented in Zuellig et al. (2007). Most of the 61 ECs are commercially synthesized compounds or their degradation products, but a few such as phenol, skatol, and the steroids can originate from natural sources.
The analytical results consisted of unqualified concentrations, E coded (or estimated) concentrations, and nondetections [reported as less than the laboratory reporting level (LRL) for the particular compound]. Estimated concentrations include those that are below or above the calibration curve, concentrations for compounds with average recoveries that are less than 60%, or concentrations of compounds routinely detected in laboratory blanks (Furlong et al., 2001). Both unqualified concentrations and E coded concentrations are used in the calculation of summary statistics (Table 1) and contaminant loads (Tables 2 and 3).
|NOD1||Discharge (m3/s)||Total Concentration1 (μg/l)||Total Load1,2 (g/day)||NOD1||Discharge (m3/s)||Total Concentration1 (μg/l)||Total Load1,2 (g/day)|
|Compound Name (common name)||2005 Accumulated Load by Sites (g/day)||2006 Accumulated Load by Sites (g/day)||Log Kow (LogP)|
|Sites 2 + 3 + 4||Site 5||% Difference||Sites 5 + 6||Site 7||% Difference||Sites 2 + 3 + 4||Site 5||% Difference||Sites 5 + 6||Site 7||% Difference|
|Accumulated Discharge (m3/s)||2.06||2.12||4.58||4.39||0.72||0.76||2.05||2.08|
|Triethyl citrate (ethyl citrate)||11.9||14.5||21.8↑||28.8||26.2||−9.03↓||20.4||19.6||−3.92↓||33.6||24.8||−26.2↓||0.33|
|Diethoxynonylphenol (total, NPEO2)||238||294||23.5↑||656||<||-↓||402||445||10.7↑||767||481||−37.3↓||-|
|Acetylhexamethyltetrahydro- naphthalene (AHTN)||11.9||12.1||1.68↑||20.6||17.4||−15.5↓||13.6||15.9||16.9↑||15.9||12.9||−18.9↓||5.70|
|Hexahydrohexamethyl- cyclopentabenzopyran (HHCB)||69.4||78.9||13.7↑||130||106||−18.5↓||77.4||93.4||20.7↑||132||83.6||−36.7↓||5.90|
|Polycyclic aromatic hydrocarbons (PAH)|
Quality Control and Quality Assessment
Quality-control samples were collected as part of this study, including one field blank, one replicate, and one laboratory spike during each sampling event. In 2005, phenol was detected in the field blank and in 2006 methyl salicylate and naphthalene were detected in the field blank. All phenol results for 2005 and naphthalene results for 2006 were considered contaminated. Results for methyl salicylate analyses were qualified as acceptable because all results were nondetectable and the estimated value in the blank (0.0170 μg/l) was much smaller than the LRL (<0.5 μg/l). To facilitate between-year comparisons, no phenol or naphthalene results were further analyzed as part of this study, though the summary data are included in Table 1 for reference.
In 2005, percent difference between environmental and replicate samples ranged from 0 to 57.5% with an overall median percent difference of 9.1 for the 31 detected ECs, excluding phenol (Table 1). In 2006, percent difference between environmental and replicate samples ranged from 6.2 to 30.4% with an overall median percent difference of 15.8 for the three detected ECs, excluding naphthalene (Table 1).
Field and laboratory spike recovery data for the ECs added to environmental sample water and laboratory reagent water at known concentrations are shown in Table 1. Percent recoveries are determined by dividing the measured concentration in the environmental or laboratory sample by the known concentration in the added spike solution. Greater than 60% of most constituents (57 of 61) were recovered for the field or laboratory spikes for one or both sampling events. Only four constituents, β-stigmastanol, d-limonene, isopropylbenzene, and tetrachloroethylene, had environmental and laboratory spike recoveries less than 60% for one or both sampling events, indicating generally poor recovery and increased uncertainty in quantification. Of these four constituents, only β-stigmastanol was detected in an environmental sample. For the 2005 event, field spike recoveries ranged from 44.1 to 171.9% and laboratory recoveries ranged from 19.2 to 102.2% with overall mean and median recoveries of 75.3 and 82.7%, respectively. For the 2006 event, field spike recoveries ranged from −22.5 to 206.1% and the laboratory recoveries ranged from 7.6 to 102.3% with overall mean and median recoveries of 72.2 and 78.3%, respectively. The relatively wide range of recoveries is not atypical for analysis of these types of compounds (Lee et al., 2004).
Differences between paired environmental and replicate samples and field and laboratory spike recoveries could be attributed to analytical variability, contamination in the environmental sample, contamination of the spike solution, or variability in recoveries owing to differences in physical or chemical properties of the ECs (Zaugg and Leiker, 2006).