Persistent Urban Influence on Surface Water Quality via Impacted

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Persistent urban influence on surface water quality via impacted groundwater Rachel S. Gabor, Steven J. Hall, David Eiriksson, yusuf jameel, Mallory Millington, Trinity Stout, Michelle L Barnes, Andrew Gelderloos, Hyrum Tennant, Gabriel J. Bowen, Bethany T Neilson, and Paul D. Brooks Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b00271 • Publication Date (Web): 21 Jul 2017 Downloaded from http://pubs.acs.org on July 23, 2017

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Environmental Science & Technology

Persistent urban influence on surface water quality via impacted groundwater Rachel S. Gabor*1,2, Steven J. Hall3, David Eiriksson1, Yusuf Jameel2, Mallory Millington2, Trinity Stout4, Michelle L. Barnes4, Andrew Gelderloos2, Hyrum Tennant4, Gabriel J. Bowen1,2, Bethany T. Neilson4, Paul D. Brooks1,2 1) Global Change and Sustainability Center, University of Utah, 115 South 1460 East, Salt Lake city, Utah 84112, United States 2) Department of Geology and Geophysics, University of Utah, 115 South 1460 East, Salt Lake city, Utah 84112, United States 3) Department of Ecology, Evolution, and Organismal Biology, Iowa State University, 251 Bessey Hall, Ames, Iowa 50011, United States 4) Civil and Environmental Engineering, Utah Water Research Laboratory, Utah State University, 8200 Old Main Hill, Logan, UT 84322-8200, United States

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Abstract

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We examined a third-order catchment that transitions from an undisturbed mountain

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environment into urban Salt Lake City, Utah. We performed synoptic surveys during a range of

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seasonal baseflow conditions and utilized multiple lines of evidence to identify mechanisms by

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which urbanization impacts water quality. Surface water chemistry did not change appreciably

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until several kilometers into the urban environment, where concentrations of solutes such as

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chloride and nitrate increase quickly in a gaining reach. Groundwater springs discharging in this

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gaining system demonstrate the role of contaminated baseflow from an aquifer in driving stream

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chemistry. Hydrometric and hydrochemical observations were used to estimate that the aquifer

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contains approximately 18% water sourced from the urban area. The carbon and nitrogen

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dynamics indicated the urban aquifer also serves as a biogeochemical reactor. The evidence of

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surface water-groundwater exchange on a spatial scale of kilometers and timescale of months to

Growing urban environments stress hydrologic systems and impact downstream water quality.

ACS Paragon Plus Environment

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years suggests a need to evolve the hydrologic model of anthropogenic impacts to urban water

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quality to include exchange with the subsurface. This has implications on the space and time

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scales of water quality mitigation efforts.

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Introduction

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Urban populations have been steadily climbing since the 1950’s, with 54% of the global

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population and 82% of the North American population living in urban areas by 20141. This

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increasing population and size of urban areas alters ecosystems and stresses natural resources2–4.

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Cities historically have been built along waterways that provide multiple services including

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water source, sanitation, transportation, and recreation. Increasing modification to these natural

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waterways and ecosystems necessitates improved understanding of how urbanization impacts

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water quality to develop effective strategies for attenuating downstream water quality issues.

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Urbanization alters the physical, chemical, and biological structure and function of the aquatic

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ecosystem, often resulting in deleterious downstream impacts. Physically, urbanization alters

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how water reaches and flows within a stream. Headwaters are often buried and directed into

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culverts or pipes5, disrupting linkages between the stream and the catchment and altering the role

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headwaters play in flow regimes and nutrient processing6. Channelization removes meanders and

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can alter exchanges between the surface and groundwater. The addition of impervious surfaces

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can decrease infiltration, channel storm water directly to a stream, and change rates of

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evapotranspiration7. These physical changes result in altered hydrologic connectivity, flow

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variability, and residence times in the surface and surbsurface. Alterations to the connections

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between surface and subsurface water as well as the dynamics of subsurface storage often reduce

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a watershed’s ability to attenuate added nutrients and other urban-derived solutes, resulting in

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decreased overall water quality8–11.

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Changes in urban hydrology combined with inputs from urban environments also alter the

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chemistry and biology of the aquatic system. Salinization of rivers from road salt runoff can limit

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the populations of vegetation and freshwater species, and impair water quality for drinking or

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irrigation12,13. Addition of nutrients alters stream ecosystem function8,14–20 and septic tanks can


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cause streams to have unsafe fecal coliform levels21,22 while other sources of pollution introduce

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toxic chemicals into waters23–25. Stormwater runoff transports nutrients and contaminants from

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urban surfaces and, through drainage infrastructure, delivers them directly to stream channels9,26–

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syndrome”31,32 and while this concept provides a framework for identifying urban impacts to

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aquatic ecosystems, there are still may unknowns in the functioning of urban watersheds.

. This ecological degradation of rivers in urban landscapes has been coined an “urban stream

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The conceptual model driving the management of urban streams tends to view the stream as a

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pipe31,33–36 with minimal or no groundwater exchange, and the assumption that the main transport

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pathway of urbanized water to the stream is through surface runoff during storm events, which is

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routed to streams via gutters, drains, and culverts. Water quality management thus often focuses

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on impervious surfaces and storm runoff as the best target for improving urban water

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quality7,9,31,37. However, globally there is widespread evidence that subsurface water drives

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surface water chemistry, with catchments storing a large amount of water that is quickly released

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during runoff events38–42, and studies have found evidence of urbanization altering subsurface

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water quality12,16,17,23,43–45. If urban catchments behave like other catchments, this widespread

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recharge of urban runoff should come back to the stream during baseflow, which is fed by

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subsurface water. Studies of river restoration have also highlighted the need to integrate whole

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watershed functioning, and not just the surface ecology and topology, in order to see appreciable

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improvements in water quality9,46. Thus, for successful implementation of green water

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infrastructure, river restoration, and water quality management, we need a better understanding

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of the degree of contribution of subsurface water to urban streams. Specifically it is necessary to

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quantify how much of that subsurface water is urban-impacted local recharge vs “cleaner”

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regional water recharged from outside the urban system.

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Given the widespread evidence that longer residence time groundwater forms a major control on

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surface water chemistry, we focused this study around two questions: "What are the relative

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contributions of locally vs regionally recharged groundwater to stream flow in a heavily

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urbanized stream?" and “Do subsurface biogeochemical reactions influence reactive solutes in

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the urban environment?" To address this we sampled seasonal baseflow, an expression of

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subsurface water, of Red Butte Creek in Salt Lake City, Utah along a gradient from forested


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montane to urbanized sites. We also measured urban runoff from several storm events and

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sampled groundwater springs in the urban area. A range of physical and chemical measurements

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allowed us to use multiple lines of evidence to identify the role of the subsurface in impacting

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the urban water quality of the creek. Salt Lake City is a fast growing urban system located in a

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semi-arid enviroment. Red Butte Creek flows through a protected montane forest and foothill

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ecosystem before entering the urban environment. We found that urban-driven changes in water

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quality during baseflow began several kilometers downstream of the start of the urban section, in

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a gaining reach where impacted groundwater springs were prevalent. This impacted groundwater

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indicated that an urban aquifer, which also acts as a biogeochemcal reactor, plays a significant

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role in urban degradation of surface water quality.

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Materials and Methods

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Site Description

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A third order stream, Red Butte Creek originates in the Wasatch Mountains (2300 m elevation)

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in northeastern Salt Lake County, Utah and flows through Red Butte Canyon until transitioning

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to an urban stream and flowing through Salt Lake City (1300 m elevation), eventually

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discharging into the Jordan River and the Great Salt Lake. The watershed has an area of 20.8

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km2 and hydrology typically seen in snowmelt-dominated mountain catchments. Average daily

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discharge (1964-2015) at the canyon outlet (USGS 10172200 Red Butte Creek at Fort Douglas,

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near SLC, UT) is 4.0 cfs and annual peak flows occur in May (12.04 ± 10.47 cfs) with annual

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low flows in September (1.69 ± 0.93 cfs).

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Red Butte Canyon is a designated Research Natural Area by the United States Forest Service,

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with minimal impacts from human activities47 which results in an abrupt transition from a

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protected environment to a heavily urbanized system at the mouth of the canyon. At the canyon

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outlet the creek flows through the University of Utah before entering a residential neighborhood.

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Additional site characteristics can be found in the Supporting Information.

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Sample Collection


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Synoptic surveys were performed in June 2014, September 2014, May 2015, July 2015, January

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2016, and May 2016, covering a range of seasonal baseflow conditions. We measured discharge

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and water chemistry at 500 m intervals beginning approximately 2 km downstream of the

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headwaters (all data is presented with the furthest upstream sampling site set as 0 km) until the

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creek enters subsurface channels in the city (Figure 1). Seasonal snow cover, private property,

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and personnel limitations resulted in some data gaps during each survey. During the May 2016

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survey Salt Lake City Utilities diverted the creek at 9.6 km and the stream channel was dry until

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10.9 km, where groundwater upwelling was observed in the stream bed.

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Discharge was measured using a YSI SonTek Flowtracker Handheld Acoustic Doppler

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Velocimeter following the velocity-area, mid-section method. In situ water quality parameters of

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temperature, specific conductivity, pH, and dissolved oxygen were measured using a YSI 6920

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V2 Sonde. All sensors were calibrated before each field campaign. Discharge was not measured

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during the surveys in 2014.

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Water samples were collected concurrently with discharge measurements. Samples were filtered

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immediately upon collection. Samples for dissolved organic matter (DOM) and total nitrogen

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(TN) analysis were filtered with pre-combusted 0.7 µm Whatman GF/F and collected in an acid-

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washed pre-combusted amber glass bottle. Samples for dissolved inorganic carbon (DIC)

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analysis (concentration and δ13C) were filtered with pre-combusted 0.7 µm Whatman GF/F and

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collected in acid-washed LDPE or HDPE bottles without headspace. Samples for ion analysis

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were filtered with either a pre-combusted 0.7 µm Whatman GF/F (samples from 2014) or a 0.45

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µm Whatman Nylon Filter (samples from 2015 and 2016) into acid-washed LDPE bottles and

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frozen (Cl-, SO42-, PO43-, NO3-, F-) or acidified with nitric acid (Na+, Mg2+, Ca2+, K+, NH4+).

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Water isotope samples were collected unfiltered in a glass vial with no headspace. All samples

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were immediately placed in coolers until transported to lab where they were frozen or

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refrigerated until analysis.

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At six locations (0 km, 5.34 km, 6.53 km, 8.04 km, 8.98 km, 11.4 km) samples were also

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collected on a biweekly or monthly basis beginning in May 2013 and continuing beyond the end


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of this study. These samples were filtered upon collection through a pre-combusted 0.7 µm

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Whatman GF/F, transported on ice, and frozen or refrigerated until analysis48 and were used to

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characterize the mountain recharge chemistry for the mixing model. Discharge is continuously

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monitored at several of these locations using Campbell Scientific CS451 vented pressure

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transducers and site-specific rating curves. Discharge is also measured at four storm culverts,

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located between the 8.04 km and 8.98 km sites, using either acoustic Doppler velocity

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measurements (Teledyne ISCO 2150 area velocity module) or the Manning equation. Water

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depth measurements used for Manning equation derived discharge calculations are made with

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Judd Communications ultrasonic depth sensors.

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Precipitation was collected on the University of Utah campus, in the upper urban section of the

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Red Butte Creek watershed and analyzed for water isotope values. Samples (n = 233) collected

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from January 2013 until May 2016 were used to develop an urban precipitation isotope end-

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member.

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Storm runoff samples in the urban section were collected during multiple storm events between

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June 2015 and November 2016. Samples were collected from commercial and residential roofs,

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streets, parking lots, turf, and storm drains. The storm culverts between 8 and 9 km were also

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sampled at the beginning, middle, and end of the storm hydrograph and between storm events.

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Samples were filtered within 24-48 hours of collection and treated as described above.

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Sample Analysis

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δ18O and δ2H of water were analyzed using a Picarro L2130-i cavity ringdown spectroscopy

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isotope analyzer using isotope ratio infrared spectroscopy49. All isotope values are reported in

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per mil (‰) notation relative to the VSMOW-SLAP scale. δ13C of DIC was measured using the

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GasBench in conjunction with a ThermoFinnigan stable isotope mass spectrometer, expressed

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relative to VPDB50.

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Anions (Cl-, SO42-, PO43-, NO3-, F-) and cations (Na+, Mg2+, Ca2+, K+, NH4+) were measured by

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ion chromatography on a Metrohm Compact IC. Cation samples were diluted 8:1 to remove


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matrix effects that distorted Ca2+ and Mg2+ concentrations. Anion samples with high

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concentrations were diluted to obtain a linear relationship between peak area and concentration.

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Standard curves were calibrated using independent NIST-traceable standards and standards were

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run as unknowns to check analytical precision.

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Dissolved organic carbon (DOC), dissolved inorganic carbon (DIC), and total dissolved nitrogen

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(TDN) analyses were performed on a Shimadzu Total Organic Carbon (TOC-L) analyzer.

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Dissolved organic nitrogen was calculated as DON=TDN-NH4+-NO3--NO2-. Longitudinal plots

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of select solutes form all sampling events can be found in Figures 1-6 in the Supporting

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Information.

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Evaporation, Load, and Mixing Model Analysis

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The Craig-Gordon model was used to calculate percent evaporation at each location in the urban

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section for each survey, following the method outlined in Skrzypek et al, 201551,52. δ18O data was

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used from each sample and local climate data were obtained from a weather station located at the

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8.47 km site. These isotope-estimated evaporative fractions were then used to predict solute

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concentrations to determine whether evapoconcentration could explain urban chemistry. Details

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of this calculation can be found in the Supporting Information.

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For samples with concurrent measurements of discharge and chemistry, incremental loads for

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each location were calculated as:

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=

∗ ∗

(1)

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Where ILD is the incremental load at the downstream site, QD, CD, and DD are the discharge,

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concentration, and location of the downstream site, respectively and QU, CU, and DU are the

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discharge, concentration, and location of site immediately upstream. The incremental load,

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expressed as (mass/time) per kilometer, measures the change in load between sites. Positive

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values indicate a net increase of solutes53,54. Negative values indicate could indicate an uptake of

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nutrients (for non-conservative solutes), loss of flow and mass from the channel, or exchange

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with subsurface water of lower concentration.

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To approximate the effective inflow concentration between each location, Ci54–56 was calculated

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as:

=

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∗ ∗

(2)

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Where QD, CD, QU, and CU are the same as above. Ci was calculated for any pair of points where

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QD > QU, showing evidence of a net inflow. This effective inflow concentration is the average,

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flow-weighted, concentration of any possible inflows within the subreach, including seeps,

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groundwater, culverts, and surface tributaries.

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Two end-members were used in a mixing model to quantify contributions to groundwater in the

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urban system. The water from the canyon, entering the urban aquifer as mountain front and

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mountain block recharge57, was considered the mountain recharge end-member. The urban

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recharge end-member represented water that interacts with the urban environment before

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entering the subsurface. The location at the mouth of the canyon (6.53 km) served as the

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mountain recharge end-member. Monthly values of isotopes and solutes for samples collected

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from 2013-2016 were weighted by the mean monthly fraction of discharge for water years 2014-

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2016 to calculate a discharge-weighted mean. Precipitation samples collected at the University of

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Utah campus in the urban section of the Red Butte Creek Watershed from January 2013-May

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2016 represented the urban recharge end-member. The isotope value of each sample was

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weighted by the total volume collected to develop a volume-weighted mean δ18O value for urban

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precipitation. Details of these calculations can be found in the Supporting Information.

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Based on representing the urban aquifer as a mixture of mountain recharge and urban recharge,

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equation (3) was then used for the end-member analysis: = ∗ + 1 − ∗

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(3)

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Where f is the fraction of urban recharge present in the urban groundwater, CUA is the chemistry

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in the urban aquifer, CUR is the chemistry in the urban recharge, and CMR is the chemistry in the

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mountain recharge.

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Equation (4) was used to calculate f as:

=

(4)


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with δ18O measurements from the urban groundwater springs used as CUA, the volume-weighted

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δ18O from the urban precipitation used as CUR, and the discharge-weighted δ18O from the site at

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the mouth of the canyon (6.53 km) used as CMR.

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Using the f estimated from the isotope mixing analysis, the solute chemistry of the urban

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recharge was predicted as: =

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where CUA is the mean concentration of solutes in the urban groundwater springs, CMR is the

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discharge-weighted solute concentration at 6.53 km, and f is calculated from Equation (4).

(5)

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Results and Discussion

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Hydrochemistry indicates three distinct reaches

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Spatial and temporal trends in discharge (Figure 2a) and conservative tracers (Figures 2b,c)

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identify three reaches with distinct hydrologic regimes. The canyon (0-6 km) is characterized by

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increasing discharge downstream and small seasonal variability (Coefficient of variation (CoV)

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= 0.36). In this reach δ18O values average -16.8‰ at the headwaters, and become steadily more

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enriched to an average of -16.4‰, corresponding with a d-excess that decreases from 9.8‰ to

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9‰ through the canyon. The upper urban reach (6-9 km) has a distinct shift in water isotopes,

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with δ18O values more enriched (-16.2‰ to -15.5‰) than in the canyon, and d-excess dropping

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to between 7‰ and 8‰. The water isotopes also have a different seasonal pattern. In the canyon

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δ18O is higher in during snowmelt than in the summer, while in the upper urban the lowest values

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are in snowmelt and highest in the summer, showing a change in seasonal behavior between the

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reaches. In addition to the shift in water isotopes, there is substantial variability in discharge

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(CoV = 0.57) in this reach. The lower urban reach (9-12 km) has another shift in isotopes, with

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higher δ18O values (-15.8‰ to -15.1‰), and lower d-excess values (6‰ to 7‰) and the same

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seasonal patterns (highest values in summer) as seen in the upper urban. In this section there is

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less variability in discharge (CoV = 0.43) and the range of isotope values across seasons

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decreases through the reach. Chloride (Figure 2d), a conservative tracer and common indicator of


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urban impact12,58,59, stays low in the canyon (5.8 mg/L at the headwaters, 11 mg/L at the canyon

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mouth), has a wider concentration range (12 - 40 mg/L) in the upper urban reach, and increases

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quickly in the lower urban reach (102 +/- 38 mg/L). Nitrate (Figure 2e), a non-conservative

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solute, behaves similarly to chloride with very low values in the canyon (0.01 - 0.05 mg N/L)

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and upper urban (0.03 - 0.1 mg N/L). In the lower urban reach nitrate spikes as high as 3 mg N

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concomitant with the increase in chloride concentrations, then converge to around 2 mg N/L

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furthest downstream.

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The differences in discharge and isotopes between these three reaches indicate varying

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hydrologic processes driving their behavior. In the canyon, the steady increase in discharge is

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indicative of a gaining stream and the discharge and δ18O values suggest consistent groundwater

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inputs from the same aquifer. The decreasing d-excess values downstream indicate evaporation,

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suggesting the downstream increase in Cl- could be due to evapoconcentration. As a whole, the

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canyon appears to be dominated by low-nutrient groundwater, with small seasonal variation. In

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the upper urban reach, the decrease in d-excess values compared to the canyon indicates water

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that has had greater exposure to evaporation, likely due to the reservoir located at the canyon

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mouth. The greater variability in discharge suggests a water table heavily influenced by

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seasonally variable mountain-front recharge. Along with a shift in δ18O this indicates a changing

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mixture of source waters, as well as exchange with the subsurface seen in variably gaining/losing

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subreaches. However, despite the variability in hydrology and source mixing the reach shows

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minimal variability in chemistry, only showing a slight increase in baseflow Cl- concentrations

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during the winter, when there is heavy road salt application, and minimal increase in NO3-. In the

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lower urban reach discharge and water isotopes indicate a return to a hydrology dominated by a

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single groundwater source; however, the increase in concentration of Cl- and NO3- indicates the

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groundwater source in this reach is impacted by the urban system.

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The three hydrologic regimes of the canyon, upper urban, and lower urban reaches also

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demonstrate distinct patterns in hydrochemistry and nutrients (Figure 3, SI Figures 1-6). The

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canyon is characterized by low nutrient concentrations (K+ = 0.7 ± 0.1 mg/L; NO3- =0.04 ± 0.02

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mg N/L; DOC = 1.0 ± 1 0.4 mg C/L), and high, chemostatic, base cation concentrations (Ca2+ =

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84.5 ± 11.9 mg/L, Mg2+ = 24.9 ± 3 mg/L)39,48, consistent with a stream fed by low nutrient


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groundwater. In the upper urban reach the small increase in Cl- coincides with small increases in

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nutrients (NO3-, DOC). Despite the indicators of variable source water, presence of large storm

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culverts, and indications of surface runoff, there is minimal urban impact to water chemistry in

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this reach, which indicates interactions with subsurface water. In the lower urban reach, the

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increase in Cl- corresponds to a proportional increase in K+ and NO3- concentrations. Ca2+

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concentrations also increase in this section, along with K+ and NO3-, while DOC decreases. The

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increased Ca2+, which is likely from interaction with alluvial fill in the subsurface, and low DOC

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concentrations, along with the hydrologic indicators of a gaining stream, all correlate with the

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addition of subsurface water. However, while the groundwater-dominated canyon was fed by

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low nutrient water, this groundwater-dominated urban environment shows significant impact of

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returning urban recharge, through high concentrations of solutes such as Cl-, NO3-, and K+.

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The urban aquifer drives urban impacts to water quality

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The increased concentrations of Cl- and NO3- in the urban reaches are concurrent with water

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isotopes indicating evaporative enrichment, as shown by both increased δ18O and decreased d-

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excess. Based on the Craig-Gordon model, urban samples ranged from 1-3% evaporated during

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spring runoff to as much as 7% evaporated during mid-summer. The increase in Cl-

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concentrations expected due to this degree of evaporation is much less than that observed (S.I.

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Fig 10), indicating that evaporation alone cannot explain increased concentrations in the urban

331

area, implying the addition of new sources of solutes in this region. Additionally, in the lower

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urban reach where these solutes increase significantly, the δ13C of DIC < -10‰ is indicative of

333

water that has recently emerged from the subsurface, yet the d-excess value indicates greater

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evaporation than seen upstream. This suggests an additional water source drives the lower urban

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hydrology and biogeochemistry.

336 337

In the upper urban section, discharge and water isotope data suggests mixing of source waters

338

and subsurface recharge. δ13C of DIC is in approximate equilibrium with the atmosphere and

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incremental loads of Cl- and NO3- both vary (Figure 4), with increased groundwater input during

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snowmelt and decreased groundwater input during the winter. The seasonal variability in the

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incremental loads is thus driven by this variability in discharge and source mixing. Incremental


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loads for both solutes spike in the lower urban section, with an initial increase due to the first

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pulse of water from the subsurface, concomitant with the first increase in measured flow. This

344

sudden increase in incremental load is driven by changes in concentration (Figure 2d, 2e), not

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discharge, and provides evidence of a new water source: urban-recharge returning to the stream.

346 347

Groundwater springs are prevalent on the north side of the stream bank beginning around 10.9

348

km, where concentrations of NO3-, and Cl-, increase, concomitant with a decrease in DOC and

349

depleted δ13C values (Figure 2, 3,4). Additionally, groundwater upwelling was observed

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beginning at 10.9 km in the streambed during the May 2016 sampling event when the city had

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diverted the creek at 9.6 km. In this section of observed groundwater input, water isotopes of the

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creek closely match water isotopes of the groundwater springs (S.I. Figure 7) and water

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chemistry of the creek also appears to be heavily influenced by the chemistry from the springs

354

(Figure 3). This evidence indicates that the urban aquifer is impacted by infiltrated surface runoff

355

and other water sourced from the urban area and provides the majority of the water to the lower

356

urban reach during baseflow.

357 358

Effective inflow concentrations (Ci) were used as further evidence of the role of the urban

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aquifer in driving water chemistry. The Ci values for Cl- and NO3- in the lower urban section are

360

at least an order of magnitude higher than in the upper urban reach (Figure 5). In the upper urban

361

reach Ci was lower during snowmelt, when isotopic evidence suggests groundwater is the

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dominant source to this reach, while Ci was higher in summer and winter, when isotopic

363

evidence suggests local/surface water inputs are larger. Ci falls within the range of values from

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storm culverts before, during, and after storms, except for Cl- concentrations during winter storm

365

runoff, which are higher than estimated inflows. Seeps in the lower urban section have

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concentrations within the range of predicted inflows, further indicating the aquifer releasing

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urban recharge is driving surface water quality at this location. Additionally, seeps at different

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locations but within 1 km of each other display different solute chemistry (Figure 5, open

369

circles), suggesting there may be perched aquifers releasing different water to the stream at

370

different locations. Ci estimates for NO3- represent lower bounds, due its non-conservative

371

behavior, and is on the low end of measured concentrations of storm water input

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We identified two main sources of water for the urban aquifer. The first is the groundwater

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entering the watershed from the canyon, primarily in the form of mountain front and block

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recharge57. The second is water recharged in the urban environment. During storm events in the

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upper urban reach the sum of discharge from the 8.04 km site and the 4 large downstream

377

culverts was higher than the discharge at the 8.98 km site. This indicates that the creek recharges

378

the shallow aquifer within the urban area, where heavier and longer duration storms resulting in

379

larger fractions being lost to recharge (SI Figure 4). The urban recharge is assumed to be largely

380

driven by precipitation, with additional possible from irrigation or leaky pipes. This includes

381

water delivered to the creek through storm culverts or detention basins that also recharge the

382

urban aquifer.

383 384

To estimate the extent to which urban influences impact the aquifer, the two end-member mixing

385

model (Equation 4) was applied with a discharge-weighted average δ18O value of -15.86 ±

386

0.85‰ for the mountain recharge and a volume-weighted average δ18O value of -14.12 ± 4.43‰

387

for the urban recharge. Isotope measurements taken from multiple springs in the lower urban

388

reach fall between these two end-members (S.I. Figure 7). Overall groundwater springs averaged

389

18.2 ± 4.4% urban-infiltrated water. The percent of urban input varied seasonally (S.I. Table 1),

390

and springs furthest upstream had a smaller fraction of urban recharge than downstream springs.

391 392

Sources of urban inputs to the aquifer

393 394

To predict the water chemistry of urban recharge necessary to explain the urban aquifer

395

chemistry, we used discharge-weighted average water chemistry from the mountain recharge and

396

average water chemistry from the urban springs with Equation (5) and f = 18.2%. Concentrations

397

of Cl-, F-, Ca2+, K+, TN, and NO3- were all greater in the urban springs than in the mountain

398

recharge, resulting in even higher predicted concentrations for urban recharge (Table 1). DOC

399

had lower concentrations in the aquifer than in the mountain recharge, resulting in a predicted

400

negative concentration in urban recharge for this non-conservative solute.

401 402

There are many possible sources for the solutes found in the urban aquifer21,23,26,58,60–62. As an

403

initial attempt to assess the sources of these solutes we analyzed the chemistry of samples from


13


Page 14 of 30

404

various urban surfaces during precipitation events. Sources included roofs of commercial and

405

residential buildings, rain gutters, parking lots, street gutters, and turf. Samples were also

406

collected from storm culverts that feed the creek in the upper urban reach during the beginning,

407

middle, and end of precipitation events as well as during dry periods. Other possible sources not

408

included in this analysis include shallow subsurface soils48,63.

409 410

The range of predicted urban recharge solute concentrations for conservative tracers are all

411

plausible based on surface storm runoff measurements (Table 1). In order for the aquifer to have

412

a Cl- concentration of 128 mg/L, urban recharge would have to contribute 646 mg/L.

413

Concentrations that high were found during winter storm runoff (64-3280 mg/L). This suggests

414

that road salt may be a dominant source of these solutes to the subsurface12,13,28 and that it has a

415

deleterious effect on baseflow water quality year round due to long residence times in the

416

aquifer. F- concentrations were not as enriched as Cl-, with an estimated urban recharge value of

417

0.26 mg/L that was within the range found in turf irrigation, pavement runoff, and storm

418

culverts. F- could also enter the subsurface through leaks in culinary water pipes. For these

419

relatively conservative tracers, measured concentrations in storm runoff can explain aquifer

420

chemistry.

421 422

Due to their non-conservative nature, concentrations of nutrients proved less predictable than the

423

more conservative solutes. K+ concentrations in the springs were over twice that found in

424

mountain recharge, leading to predicted urban recharge concentrations of 8.9 mg/L, which was

425

found in runoff from pavement, storm culverts, and winter storms, and other sources such as

426

leachate from fertilized lawns63. TDN concentrations were over an order of magnitude larger in

427

the springs (2.07 mg N/L) than the mountain recharge (0.17 mg N/L) and none of the measured

428

runoff concentrations of TDN came close to the predicted concentration of 10.61 mg N/L for

429

urban recharge. NO3- behaved similarly to TDN and Hall et al 2016 found that δ15N of NO3- was

430

consistent with a human or animal source to the subsurface48, suggesting additional N could be

431

due to leaky sewer pipes or other legacy N inputs64. Shallow urban soils present an additional

432

potential source of nitrogen to the subsurface, which infiltrating water would interact with in

433

areas with permeable surfaces63. The springs had a lower concentration of DOC (0.55 mg C/L)

434

than the mountain recharge (1.63 mg C/L) that resulted in the mixing model predicting a


14

Page 15 of 30


435

negative predicted carbon concentration in urban recharge. However, measurements of DOC in

436

runoff ranged from 0.01 mg C/L (winter storm) to 47 mg C/L (pavement). The observation of

437

nutrient concentrations much higher than conservative mixing would predict suggests the

438

subsurface acts as a biogeochemical reactor where organic carbon and nitrogen are metabolized

439

in respiration and denitrification processes16.

440 441

Implications for urban water quality management

442 443

Our study found that in a small urban creek, urban impacts to water quality occurred during

444

baseflow, specifically in the presence of impacted groundwater springs. We identified an urban

445

aquifer, which is recharged by both “clean” mountain water and impacted urban water, that

446

returns urban-impacted groundwater to the surface at timescales longer than those of individual

447

storm events (Figure 5). This indicates that just as non-urban catchments are found to store and

448

release large amounts of water, resulting in water chemistry that is driven by subsurface inputs39–

449

42

450

locally recharged groundwater vary seasonally as the water table rises and falls and different

451

perched aquifers contribute to the stream. These relative contributions can be identified using

452

multiple isotopic and chemical indicators, as shown above. The importance of the subsurface in

453

driving urban water quality suggests that water management that focuses purely on storm water

454

as a point source, or treats the stream as a pipe where water stays after it enters, misses a

455

substantial contribution of the urban environment to water quality.

urban catchments can demonstrate the same behavior. Relative contributions of regionally vs.

456 457

Additionally, the nutrient concentrations found in the runoff at the urban springs show that the

458

urban aquifer is an area of high biogeochemical activity. The subsurface appears to be an active

459

area of nutrient consumption and transformation, as well as a potential source for nutrients such

460

as NO3-. Water entering the urban system has a mean DOC concentration of 1.63 mg C/L with

461

DON ranging from 0.05 to 0.30 mg N/L, yet while the DOC increases in the upper urban reach,

462

it drops in the lower urban 0.55 mg C/L and there is no measurable DON in this section. This

463

represents a fundamental shift in nutrient dynamics in the stream, as seen in the C:N ratio of the

464

organic matter, that appears to be driven by inputs from the aquifer. Mayer et al (2010) found

465

that in an urban system in Baltimore denitrification of NO3- in urban groundwater was limited by


15


Page 16 of 30

466

DOC availability16. In the carbon-poor subsurface of the Red Butte Creek watershed, time spent

467

in the aquifer causes DOC concentrations to decrease and DIC concentrations to increase (from

468

46.8 mg C/L to 55.4 mg C/L) from the mountain groundwater entering the system. This suggests

469

that some NO3- inputs are processed in the subsurface along with microbial respiration, but there

470

is not enough C to remove the larger NO3- inputs from the urban system via denitrification.

471 472

Overall, these results indicate larger-scale hydrogeologic controls on surface water-groundwater

473

interactions in urban catchments that operate on a scale of kilometers in space and months to

474

decades in time. Substantial work has already expanded the conceptual model of the river as a

475

pipe to include exchange with the hyporheic zone45, which accounts for exchanges with the

476

subsurface over meters and from hours to days. Our evidence of exchange occurring over larger

477

space and time scales necessitates a consideration of subsurface exchange not just with the

478

hyporheic, but with a larger alluvial zone35,38,40,44,45(See Graphical Abstract). This highlights the

479

importance of the vertical dimension, as described in the urban watershed continuum35, in the

480

transport and transformation of solutes in an urban system and the importance of legacy nutrients

481

to water quality64. This also adds a level of complication to the popular storm water management

482

strategies of using bioswales and other low-impact development strategies that aim to reduce

483

storm runoff and instead encourage storm water infiltration to the subsurface34,65,66. If this water

484

infiltrates untreated, bringing with it urban-derived impacts to groundwater quality, and

485

eventually returns to the stream as subsurface exchange, then these strategies only change the

486

time and space scales of urban impacts to water quality, rather than decreasing the impacts. The

487

challenge of urban water quality management and restoration efforts then becomes how to

488

account for the long memory of the system represented by the larger volume and longer

489

residence time of the subsurface water, while also recognizing the subsurface can act as a

490

biogeochemical reactor responding to inputs from the urban system.

491 492 493 494 495 496 497 498

Acknowledgements We thank Brett Boyer, Simone Jackson, Julie Leri, Erik Oerter, Leah Richardson, Rebecca Smith, Carla Valdez, Zinnia Wilson, and Margaret Wolfe for assistance in sample collection and Sagarika Banerjee and Suvankar Chakraborty for analytical assistance. We also thank Tracie Kirkham and the Salt Lake City Department of Public Utilities for information on the city water infrastructure. This research was supported by NSF EPSCoR grant IIA 1208732 awarded to Utah


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Roy, A. H.; Wenger, Æ. S. J.; Fletcher, Æ. T. D.; Walsh, Æ. C. J.; Ladson, A. R.; Shuster, Æ. W. D.; Thurston, Æ. H. W.; Brown, R. R. Impediments and Solutions to Sustainable , Watershed-Scale Urban Stormwater Management : Lessons from Australia and the United States. 2008, 344–359.

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FIGURE 1

Figure 1: Map of sampling locations along Red Butte Creek showing distance downstream from the upper most sample site.


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FIGURE 2

Figure 2: Discharge and water isotopes indicate three distinct hydrologic regimes along Red Butte Creek. The canyon (0-6 km) exhibits minimal variability in discharge and progressive, downstream evaporative enrichment. The upper urban (6-9 km) has high variability in discharge both seasonally and spatially. The lower urban (9-12 km) reach exhibits minimal variability in discharge, a step change increase in evaporation, and large increases in Cl- and NO3- indicative of input of urban-impacted water from the subsurface. .


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FIGURE 3

Figure 3: Boxplots showing solute concentrations within each reach for all sample periods. The canyon (0-6 km) and upper urban (6-9 km) are characterized by low solute concentrations. The increase in base cations indicates input of longer residence time groundwater in the lower urban (9-12 km) reach. This new groundwater input is similar to regional springs and contains nutrient and Cl- concentrations indicative of urban impacts. The chemistry of the springs is also very similar to that of the lower urban reach, suggesting the stream is strongly driven by subsurface inputs. Letters indicate reaches that are significantly different from the immediately upstream reach with a 95% confidence.


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FIGURE 4

Figure 4: In the upper urban (km 6-9) reach, δ13C of DIC is around -8‰, indicating water that has been at the surface long enough to equilibrate with the atmosphere. Higher values in winter indicate less groundwater input than during snowmelt. The drop of δ13C in the lower urban (9-12 km) indicates water that has recently emerged from the subsurface, corresponding with discharge measurements indicating a gaining stream. Incremental loads of Cl- and NO3- show seasonal variability in the upper urban, largely driven by spatial and temporal variability in the location of groundwater recharge. The lower urban shows an initial increase in load with the first pulse of water from the subsurface, located around 11.4 km, and then smaller increases as the stream becomes dominated by subsurface water.


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FIGURE 5

Figure 5: Effective inflow concentrations (Ci) within each subreach. Observed inflow concentrations from storm culverts (taken during both storms and baseflow) are shown in the boxplots between 8 km and 9 km, with the high concentration outliers for Cl- arising from a winter storm event. Chemistry of samples from seeps in the lower urban (9 - 12km) are represented by red open circles. For both NO3- and Cl-, both seep and surface water chemistry in the lower urban reach more closely resembles stormwater runoff from upstream culverts than stream water directly upstream. Additionally, predicted inflow chemistry (Ci) corresponds well with seep chemistry, indicating seeps are likely the dominant inflow source in the lower urban reach.


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FIGURE 6

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Figure 6: Conceptual model of how regional water and urban water mix to drive surface water quality in urban streams with illustrative examples of possible source chemistry (values for stream and groundwater systems are mean, values for urban runoff are median – see Table 1 for variability in measurements). Mountain recharge low in dissolved solutes mixes with local urban recharge high in solutes in an urban aquifer. This impacted water discharges downstream resulting in sustained changes to baseflow chemistry.


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TABLE 1 Predicated concentrations of urban water chemistry based on discharge-weighted mean chemistry of the mounting water and mean chemistry of the aquifer, compared to runoff samples from a range of urban surfaces during multiple storm events. ClF– Ca2+ K+ DOC TDN NO3(mg/L) (mg/L) (mg/L) (mg/L) (mg C/L) (mg N/L) (mg N/L) 12.65 0.10 76.70 1.04 1.63 0.17 0.04 0.76 0.01 14.28 0.11 0.56 0.07 0.05 127.93 0.13 125.02 2.47 0.55 2.07 2.04 12.84 0.05 11.11 0.55 0.35 0.99 1.04 646.19 0.26 342.25 8.90 -4.31 10.61 11.03

Regional mean Water S.D. Urban mean Springs S.D. Predicted Urban Water Range of measurements found in storm runoff samples Turf Runoff Min 0.54 0.02 1.46 Max 58.77 0.88 99.22 Median 35.26 0.46 49.75 Roof Runoff Min 0.11 0.00 0.12 Max 139.72 0.36 102.78 Median 5.07 0.08 3.75 Pavement Min 0.69 0.00 0.63 (streets and Max 172.47 1.92 58.68 parking lots) Median 10.79 0.12 11.81 Storm Min 4.25 0.02 4.21 Culverts Max 396.16 1.41 261.59 Median 25.78 0.21 29.78 Winter Min 63.89 0.01 78.35 Storm Max 3280.25 0.14 87.43 Runoff Median 1573.88 0.04 86.85

0.86 7.42 5.45 0.00 9.95 0.63 0.00 25.00 2.78 0.59 55.47 2.81 4.97 23.46 13.26

2.38 10.54 9.01 0.77 22.92 5.59 1.54 81.13 7.35 1.17 48.27 9.39 0.01 0.96 0.54

0.22 2.37 1.05 0.20 9.84 1.05 0.15 8.46 1.16 0.12 22.16 1.07 0.34 0.72 0.54

0.01 0.21 0.05 0.04 7.58 0.53 0.00 4.60 0.45 0.00 2.54 0.33 0.00 0.18 0.00

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Persistent Urban Influence on Surface Water Quality via Impacted

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