Applied Geochemistry xxx (2010) xxx–xxx

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Applied Geochemistry

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Aquifer Storage Recovery (ASR) of chlorinated municipal drinking waterin a confined aquifer

John A. Izbicki a,*, Christen E. Petersen b, Kenneth J. Glotzbach c, Loren F. Metzger d, Allen H. Christensen a,Gregory A. Smith a, David O’Leary a, Miranda S. Fram d, Trevor Joseph d,e, Heather Shannon b

a US Geological Survey, 4165 Spruance Road, Suite 200, San Diego, CA 92101, USAb Montgomery Watson Harza, Inc., 3321 Power Inn Road, Sacramento, CA 95826-3893, USAc City of Roseville, Environmental Utilities Department, 2005 Hilltop Circle Roseville, CA 95747, USAd US Geological Survey, 6000 J Street, Placer Hall, Sacramento, CA 95819, USAe California Department of Water Resources, 901 P Street, Sacramento, CA 95814, USA

a r t i c l e i n f o a b s t r a c t

Article history:Received 10 December 2009Accepted 28 April 2010Available online xxxx

Editorial handling by R. Fuge

0883-2927/$ - see front matter Published by Elsevierdoi:10.1016/j.apgeochem.2010.04.017

* Corresponding author. Tel.: +1 619 225 6131; faxE-mail address: jaizbick@usgs.gov (J.A. Izbicki).

Please cite this article in press as: Izbicki, J.A., eGeochem. (2010), doi:10.1016/j.apgeochem.201

About 1.02 � 106 m3 of chlorinated municipal drinking water was injected into a confined aquifer, 94–137 m below Roseville, California, between December 2005 and April 2006. The water was stored inthe aquifer for 438 days, and 2.64 � 106 m3 of water were extracted between July 2007 and February2008. On the basis of Cl� data, 35% of the injected water was recovered and 65% of the injected waterand associated disinfection by-products (DBPs) remained in the aquifer at the end of extraction. About46.3 kg of total trihalomethanes (TTHM) entered the aquifer with the injected water and 37.6 kg of TTHMwere extracted. As much as 44 kg of TTHMs remained in the aquifer at the end of extraction because ofincomplete recovery of injected water and formation of THMs within the aquifer by reactions with free-chlorine in the injected water. Well-bore velocity log data collected from the Aquifer Storage Recovery(ASR) well show as much as 60% of the injected water entered the aquifer through a 9 m thick, high-per-meability layer within the confined aquifer near the top of the screened interval. Model simulations ofground-water flow near the ASR well indicate that (1) aquifer heterogeneity allowed injected water tomove rapidly through the aquifer to nearby monitoring wells, (2) aquifer heterogeneity caused injectedwater to move further than expected assuming uniform aquifer properties, and (3) physical clogging ofhigh-permeability layers is the probable cause for the observed change in the distribution of boreholeflow. Aquifer heterogeneity also enhanced mixing of native anoxic ground water with oxic injected water,promoting removal of THMs primarily through sorption. A 3 to 4-fold reduction in TTHM concentrationswas observed in the furthest monitoring well 427 m downgradient from the ASR well, and similar mag-nitude reductions were observed in depth-dependent water samples collected from the upper part of thescreened interval in the ASR well near the end of the extraction phase. Haloacetic acids (HAAs) were com-pletely sorbed or degraded within 10 months of injection.

Published by Elsevier Ltd.

1. Introduction

Although a number of definitions exist, Aquifer Storage Recov-ery (ASR) is commonly defined as the practice of injecting waterinto an aquifer through wells for the purpose of creating a subsur-face water supply that is recovered at a later time to meet seasonal,long-term, emergency, or other demands (Pyne, 1995, 2005; Sedi-ghi et al., 2006). In many projects, the same well is used for bothinjection and recovery (Singer et al., 1993). In 2009, ASR projectswere ongoing at more than 95 sites in the USA and in at least sevenother countries (David Pyne, pers. comm., 2009). Water managersresponsible for the operation of ASR projects need informationon the physical movement of injected water near ASR wells and

Ltd.

: +1 619 225 6101.

t al. Aquifer Storage Recovery0.04.017

on the chemical reactions that occur between injected water andnative ground water to optimize the performance of their wellsand the recovery of injected water.

Water injected into aquifers by wells is often described as abubble within the aquifer near the well. In the most basic concep-tual model, the bubble is geometrically simple and commonly rep-resented as a cylinder occupying the entire thickness of the aquifer(Fram et al., 2003). The radius of the cylinder surrounding the wellincreases as injected water displaces native ground water and de-creases as injected water is recovered from the well (Pyne, 1995,2005). The injected bubble near an ASR well is sometimes viewedas a discrete body with injected water inside the bubble separatefrom native ground water outside the bubble (Vacher et al., 2005).

More complex models of the movement of injected water nearASR wells account for differences in aquifer properties that allowincreased movement of water through higher permeability layers

(ASR) of chlorinated municipal drinking water in a confined aquifer. Appl.

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within the aquifer (Maliva et al., 2006). The resulting models arecommonly stacked cylinders that in confined aquifers resemblean ‘‘inverted tree” (Missimer et al., 2002) and in unconfined aqui-fers may have ‘‘mushroom-cloud” configurations as injected watermoves into formerly unsaturated deposits (Fram et al., 2003). Aswater moves through thin high-conductivity layers within aquifersthe injected bubble is increasingly less of a discrete body and mayultimately approach a complex ‘‘bottle–brush” configuration(Hutchings et al., 2004; Vacher et al., 2005).

Recent work by Pavelic et al. (2006) used flowmeter data fromlong-screened wells that fully-penetrated the injection zone nearan ASR well to extend conceptual models of ground-water flowin a heterogeneous aquifer to a large-scale field problem. Resultsof their work showed that injected water moved preferentiallythrough high-permeability layers within the aquifer. Pavelic et al.(2006) confirmed interpretations of water movement, aquiferproperties, and mixing of injected and native water developedfrom flowmeter data on the basis of temperature data. Pavelicet al. (2006) did not address how the ground-water flow in a het-erogeneous aquifer may have affected mixing and the subsequentchemical reactions between injected water, native water, and aqui-fer materials that may alter the quality of recovered water.

1.1. Disinfection by-products in ASR

The quality of recovered water is of concern in areas where po-table water has been injected into waters having saline or other-wise poor-quality water. In these areas, the ability to recoverinjected water without adverse water-quality effects caused bymixing with the poor-quality native water may be the primarydeterminant of the success or failure of a project (Lowry andAnderson, 2006). Geochemical reactions between injected water,native water, and aquifer materials may further alter the chemistryof the recovered water, often adversely impacting the project (Mir-ecki et al., 1998; Herczeg et al., 2004). In other studies, water-qual-ity issues driven by regulatory requirements may necessitatealmost complete recovery of injected water to prevent disinfectionby-products (DBPs) from leaving the area near the well (Fram et al.,2002, 2003; Nordberg, 2005). If sorption or degradation of DBPs oc-curs within an aquifer during storage, recovery of injected water tomeet regulatory concerns may be less of a constraint on operations.

DBPs include trihalomethanes (THMs), haloacetic acids (HAA),and a number of other compounds (Rook, 1974, 1976; Richardson,2002). Many of these compounds are known or suspected carcino-gens (Morrow and Minear, 1987; Masters, 1998; US EPA, 1998,2006). THMs and HAAs may be present in chlorinated water in-jected into an aquifer (Rook, 1974, 1976; Golfinopolos et al.,1998; Richardson, 2002), and also may form within the aquiferas organic material reacts with free-chlorine residual (primarilyin the form of hypochlorous acid) in the injected water (Thomaset al., 2000; Fram et al., 2002, 2003, 2005). In addition, Br� in nativeor injected water may be oxidized to hypobromous acid by free-chlorine in injected water and the subsequent reaction of hypobro-mous acid with organic material may form bromated THMs andHAAs (Morris, 1978; Rook et al., 1978). Bromated compounds havedifferent physical and chemical properties than chlorinated THMsand HAAs that affect their transport and degradation in groundwater. Although hypobromous acid reacts faster with organic Cto form THMs than hypochlorous acid (Rook et al., 1978; Amyet al., 1985; Symons et al., 1993; Fram et al., 2003) chlorinatedcompounds dominate DBPs formed during chlorination. MaximumContaminant Levels (MCLs) for total THMs (TTHM) are 80 lg/L, andMCLs for total HAA (THAA) are 60 lg/L (US EPA, 1998, 2006).

Abiotic degradation of THMs through hydrolysis is commonlyslow at the pHs encountered in natural systems (Nicholson et al.,2002; Buszka et al., 1994), and sorption and microbially mediated

Please cite this article in press as: Izbicki, J.A., et al. Aquifer Storage RecoveryGeochem. (2010), doi:10.1016/j.apgeochem.2010.04.017

degradation account for almost all of the THM removal in groundwater. In general, bromoform is more strongly sorbed than the lessbromated compounds, which are more strongly sorbed than chlo-roform (Nicholson et al., 2002). Sorption of THMs to organic mate-rials within aquifer deposits has been observed in field andlaboratory experiments (Roberts et al., 1982, 1986; Curtis et al.,1986; Walton et al., 1992; Peng and Dural, 1998). Fram et al.(2003) calculated that sorption of THMs may be measurable foraquifer deposits having an organic C content of 0.05%. Degradationof THMs occurs under anerobic conditions, and bromoform is morereadily degraded than less bromated compounds, which are morereadily degraded than chloroform (Pyne et al., 1996; Pyne, 2005;McQuarrie and Carlson, 2003; Pavelic et al., 2005). In stronglyreducing aquifers complete degradation of THMs has been re-ported (PEER Consultants and CH2M-Hill, Inc., 1997). However,conservative (non-reactive) behavior of chloroform has been re-ported in oxic aquifers (Thomas et al., 2000; Landmeyer et al.,2000; Fram et al., 2002, 2003). Injected water commonly containsO2 and, in the absence of reactions with aquifer materials, can cre-ate an oxic environment near the well—even in reducing environ-ments. As a consequence, degradation may be more likely near themargin of the injection bubble where mixing with native groundwater is greater (Fram et al., 2003).

Although HAAs have been less studied than THMs, availabledata suggest that abiotic degradation of HAAs through hydrolysisor decarboxylation is slow (Nicholson and Ying, 2005). HAAs areless strongly sorbed than THMs (Nicholson et al., 2002) but aremore readily degraded by microorganisms than THMs (Landmeyeret al., 2000; Thomas et al., 2000). Sorption and degradation of HAAsincreases with increased bromination (Nicholson et al., 2002). Ra-pid degradation of HAAs has been reported even in oxic aquiferswhere THMs were not degraded (Singer et al., 1993; Thomaset al., 2000).

1.2. Roseville ASR test

Chlorinated municipal drinking water from surface-watersources was injected into a confined alluvial aquifer underlyingthe City of Roseville in the Sacramento Valley about 100 km NEof San Francisco, California, USA (Fig. 1). ASR has been proposedat this site to ensure a reliable water supply to meet projected in-creased demand without further depletion of the regional groundwater resource (Petersen and Glotzbach, 2005, 2008). In 2004, a96-day pilot test determined the site was suitable for ASR(MWH, Inc., 2004).

The fate of DBPs in the injected water was a concern at theRoseville site because regulatory constraints do not permit in-creases in DBP concentrations in the surrounding ground water.There was concern that DBPs in injected water could move withthe regional ground-water flow during injection and storage be-yond the influence of the ASR well during the recovery phase. IfDBPs remained in the aquifer downgradient from the well, injec-tion at the site could be restricted (CVRWQCB, 2005). If sorptionor degradation of DBPs occurred within the aquifer, movement ofwater beyond the influence of the ASR well would be of less regu-latory concern. Although model simulation results of the pilot ASRtest (Nordberg, 2005) suggested a high recovery of injected waterduring operational periods longer than the 96-day pilot test, meet-ing regulatory constraints required additional understanding of theeffect of aquifer heterogeneity on ground-water flow near the ASRwell and the fate of DBPs in the injected water.

1.3. Purpose and scope

The purpose of this study was to evaluate how aquiferheterogeneity affected (1) the injection, storage, and recovery of

(ASR) of chlorinated municipal drinking water in a confined aquifer. Appl.

Fig. 1. Location of study area.

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municipal drinking water, and (2) the fate and transport of associ-ated DBPs during an 806 day injection, storage, and recovery test(Table 1) in the confined aquifer underlying Roseville, Calif.

The scope of this study included (1) collection of injection andpumping data, water-level, and water-quality data from the ASRwell and three nearby monitoring wells by MWH, Inc. and the Cityof Roseville, and (2) collection of well-bore flow and depth-depen-dent water-quality data from the ASR well and the three monitor-ing wells by the US Geological Survey (USGS). Well-bore flow datafrom the ASR well were interpreted at the well-bore scale with theaid of a numerical ground-water flow model with a particle-tracker.

2. Hydrogeology

The study area is in the North American Subbasin of the Sacra-mento Valley Ground water Basin (California Department of WaterResources, 2003) in the City of Roseville, about 100 km NE of SanFrancisco, California. Alluvial deposits in the subbasin thicken from

Please cite this article in press as: Izbicki, J.A., et al. Aquifer Storage RecoveryGeochem. (2010), doi:10.1016/j.apgeochem.2010.04.017

east to west having a maximum thickness of 650 m near the centerof the valley. Regional ground-water flow is from recharge areasnear the mountain front to a large regional ground water depres-sion created by pumping in excess of recharge west and SW ofthe study area (Petersen and Glotzbach, 2008). In 1996, wateruse by the City of Roseville was about 2.59 � 107 m3 and demandwas expected to increase to 6.78 � 107 m3 annually as the City isfully developed (Petersen and Glotzbach, 2008).

At the study site, the aquifer deposits have been divided into anupper unconfined aquifer and a lower confined aquifer. The upperaquifer extends from the water table to about 94 m bls and con-sists of fine-grained silt and clay deposits. The lower aquifer di-rectly underlies the upper aquifer and extends to a depth of140 m bls (Fig. 2). Alluvial deposits composing the lower aquiferat the site are composed of ‘‘black sands” and gravels believed tobe eroded from volcanic sands of the Mehrten Formation (MWH,Inc., 2004; Nordberg, 2005).

The study site contains an ASR well (well 17F1) screenedthroughout the lower confined aquifer, and three monitoring wells(wells 17F2, 17F3, and 17E1) which are 27.7, 58.8, and 427 m from

(ASR) of chlorinated municipal drinking water in a confined aquifer. Appl.

Table 1Injection, storage, and extraction timeline, Roseville, California.a

Study phase Start date End date

Pilot test (Nordberg, 2005) (includesinjection, storage and extraction)

6/16/2004 9/20/2004

Baseline 5/5/2004 12/14/2005Depth-dependent data collection 11/29/2005 12/07/2005

Injection 12/14/2005 5/5/2006Depth-dependent data collection 12/19/2005 12/23/2005

Storage 5/5/2006 7/18/2007Depth-dependent data collection 12/04/2006 12/08/2006

Extraction 7/18/2007 2/28/2008Depth-dependent data collection 8/05/2007 8/09/2007Depth-dependent data collection 2/27/2008 2/28/2008

Post-test 2/29/2008 3/29/2008

a Depth-dependent work includes velocity log n/extraction well 11N/6E-17F1(DCPW) and monitoring wells 11N/6E-17F2, 17F3, and 17E1 (DCMW-1, 2, and 3,respectively). Velocity logs and depth-dependent samples were not collected fromwell 17F1 during injection and storage phases. No data were collected from mon-itoring wells 17F1, 17F2, and 17E2 during the 2/27–28/2008 sample collection.

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the ASR well (Fig. 2). The monitoring wells were also screenedthroughout the lower aquifer across the injection interval. Geologicdata collected during well drilling indicates that the lower aquiferdips to the SW and is encountered at a depth of 116 m below landsurface near monitoring well 17E1, SW of the ASR well (Fig. 1). Thepotentiometric surface in the lower aquifer underlying Rosevillealso slopes to the SW with a gradient of 0.0012.

The transmissivity of the lower aquifer system at the ASR sitewas estimated to be 2.8 � 103 m2/day from aquifer-tests data(MWH, Inc., 2004). Given a gradient of 0.0012 (MWH, Inc., 2004)and assuming a porosity of 0.2 the rate of ground water movementin the lower aquifer is about 0.3 m/d. Although the lower aquifer isconsidered to be relatively homogeneous on a regional scale, lith-ologic data indicates that the lower aquifer is heterogeneous on thelocal scale. For example, the upper 3 m of the lower aquifer at theASR well (95–98 m bls) were composed of highly-permeable gravel

Fig. 2. Geologic section A-A0 ,

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deposits; whereas, the underlying deposits were finer grained andmore poorly sorted with depth, especially near the bottom of theaquifer (MWH, Inc., 2004). Depth-dependent water-quality data,collected as part of this study, showed that water in the graveldeposits was oxic; whereas, water in deposits below this layerdid not contain O2 and was reducing.

3. Methods

3.1. Field methods

Water-levels were measured continuously, using pressuretransducers, by MWH, Inc. from the ASR well and the three moni-toring wells during the study. Water samples were collected fromthe surface discharge of the ASR well and the three monitoringwells by MWH, Inc. at intervals ranging from weekly (at the begin-ning of the injection, storage, and extraction phases) to monthly(during the later part of each phase when water-quality waschanging less rapidly) (MWH, Inc., 2007). During sample collection,the monitoring wells were pumped at about 24 L (8 gal)/min untilat least three casing volumes were removed and field parametershad stabilized. Samples were chilled and delivered on the day ofcollection to the California Department of Water Resources BryteLaboratory, Sacramento, California, for analysis.

Well-bore flow and depth-dependent water-quality data collec-tion required specialized techniques and access to the wells thatdiffered from procedures used for sample collection from the sur-face discharge of the ASR or monitoring wells. Different procedureswere used for the ASR well and for the monitoring wells because ofaccess constraints.

Access to the ASR well was through a 3.8 cm diameter accesstube that extended from land surface to below the pump intakeand entered the well above the screened interval. Well-bore flowdata were collected under pumping conditions using the tracer-pulse method (Izbicki et al., 1999). When using this method, an

near Roseville, California.

(ASR) of chlorinated municipal drinking water in a confined aquifer. Appl.

Table 2Trihalomethanes (THMs) and haloacetic acids (HAAs) determined as part of thisstudy.

Compound Formula Reporting limits

Bryte Lab (lg/L) USGS Lab (lg/L)

TrihalomethanesChloroform CHCl3 0.5 0.05Bromodichloromethane CHCl2Br 0.5 0.05Dibromochloromethane CHClBr2 0.5 0.05Bromoform CHBr3 0.5 0.05

Haloacetic acidsDibromoacetic acid CHBr2CO2H 1.0 0.16Dichloroacetic acid CHCl2CO2H 1.0 0.42Monobromoacetic acid CH2BrCO2H 1.0 0.25Monochloroacetic acid CH2ClCO2H 1.0 0.45Trichloroacetic acid CCl3CO2H 1.0 0.14

Total THMs and total HAA were calculated as the sum of the individual compounds.For Bryte Lab, THMs determined using US EPA method 502.2; HAAs determineusing US EPA method 552.2. For USGS lab., THMs determined using US EPA method502.2, modified by Crepeau et al. (2004); HAAs determine by Weck laboratoriesusing US EPA method 552.2 modified by Zazzi et al. (2005). Reporting limits inmicrograms per liter, lg/L.

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easily measured tracer (Rhodamine dye) is injected into the well ata known depth, the arrival of the tracer is measured at the surfacedischarge of the well, and the time-of-travel of the tracer is re-corded. The injection is repeated at a deeper depth, the arrival ofthe tracer is measured, and the new travel time is recorded. Thedistance between the injections (L) divided by the difference in tra-vel times (t) is the average velocity (L/t) of water in the well be-tween injection depths. Multiplying this velocity computedbetween two injections by the cross-sectional area of the well(L2) yields the average flowrate (L3/s) in the measured interval.The injections are repeated at different depths throughout the wellto create a profile of well-bore flow (Izbicki et al., 1999).

Depth-dependent water-quality samples from the ASR wellwere collected under pumping conditions at selected depths with-in the well using a small-diameter (2.5-cm) gas-displacementpump (Izbicki, 2004). Sample depths were determined on the basisof the well-bore flow data. The deepest sample collected from thewell is representative of water from the deepest part of the well.The next deepest sample is a mixture of the deepest sample andwater that entered the well from the aquifer between the two sam-ple depths. The composition of water that entered the well fromthe aquifer between the two sample depths was calculated fromwell-bore flow data and depth-dependent water-quality data usinga mass-balance equation as described by Izbicki et al. (1999) andIzbicki (2004). Sample collection was repeated at different depthsdetermined on the basis of the well-bore flow data to obtain a pro-file of water-quality within the well and the aquifer.

Three 10 cm diameter PVC monitoring wells were installed atdistances of 27.7, 58.8, and 427 m from the ASR well (Fig. 2) tomonitor the effects of injection, storage, and extraction. The wellshave long-screen intervals that completely penetrate the confinedaquifer, and are similar in length to the screened interval of theASR well. The design of these monitoring wells is different fromthe short-screen intervals commonly used in ground-water studies(California EPA, 1995). However, the design is similar to the fully-penetrating monitoring wells used to monitor the effects of ASR byPavelic et al. (2005). This design provides the opportunity to devel-op information on the hydraulic properties of aquifer deposits andmovement of water throughout the entire interval screened by thewell.

Access to the monitoring wells was facilitated by removal of thededicated pump column and pump by a commercial pump com-pany. Well-bore flow data were then measured under non-pump-ing conditions using an electromagnetic (EM) flowmeter. Forpumping logs, a temporary pump and pump column were installedwhile the EM flowmeter was in the well below the depth of thepump intake. The well was then pumped at the sample rate ofthe dedicated pumps, about 0.5 L/s. The EM flowmeter has a largedynamic range capable of measuring lower flows present duringunpumped conditions and higher flows present during pumpedconditions (Newhouse et al., 2005). Fluid resistivity and fluid tem-perature logs were collected at the same time as the unpumpedand pumped well-bore flow logs using sensors embedded in theEM flowmeter. After the well-bore flow data were collected, thepump column, pump, and EM flowmeter were removed from thewell. A small-diameter gas-displacement pump (Izbicki, 2004)was then inserted into the well, and the temporary pump andpump column were reinstalled. The well was then pumped at0.5 L/s and samples were collected from within the well at selecteddepths determined on the basis of the EM flow logs. After samplecollection was completed, the gas-displacement pump was re-moved and the dedicated sample pumps were reinstalled.

Depth-dependent water-quality samples from the ASR well andthe monitoring wells were analyzed for field parameters (pH, spe-cific conductance and dissolved O2). Samples were collected fordisinfection by-products (DBPs) including: trihalomethanes

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(THMs), haloacetic acids (HAAs), trihalomethane formation poten-tial (THMFP), and haloacetic acid formation potential (HAAFP).Additional samples were collected for dissolved organic C, opticalproperties (including ultraviolet (UV) absorbance and excitationemission (EEM) spectroscopy), and selected inorganic ions (includ-ing Cl�, F�, Br� and NO�3 ). THMs were collected in four 40-mL glassseptum bottles. Each bottle was completely filled to exclude airand was analyzed separately in the laboratory. If used improperlythe gas-displacement pump can strip volatile compounds from awater sample. In general, agreement between the four replicateTHM samples was good, although occasional low THM concentra-tions were measured. The highest THM concentration in each set ofreplicates was assumed to be the representative THM concentra-tion. Samples for HAAs, THMFP, and HAAFP were collected intwo 250 mL glass bottles. Samples for dissolved organic C (DOC)and optical properties were field filtered and collected in 60 mLglass bottles. All glass bottles used in this study were baked at600 �C prior to use. Samples for inorganic ions were field filteredusing 0.45 mm pore-sized filters and stored in 125 mL plastic bot-tles. Blank samples collected during the study showed no evidenceof contamination from the depth-dependent sample pump andsample collection lines.

3.2. Laboratory methods

Samples collected from the surface discharge of the ASR welland monitoring wells by MWH, Inc. were analyzed for THM,HAA, DOC, Cl� and selected inorganic constituents by the CaliforniaDepartment of Water Resources at the Bryte Laboratory in Sacra-mento. Samples for THMs and HAAs were analyzed using gas chro-matography (GC) with an electron capture detector (ECD) usingEPA method 502.2 (US EPA, 1995a) and method 552.2 (US EPA,1995b), respectively. Total THMs (TTHMs) and total HAAs (THAAs)were calculated as the sum of the individual measured compounds.Laboratory reporting limits for THMs and HAAs are given inTable 2.

Depth-dependent water-quality samples were analyzed forTHMs (Table 2) at the US Geological Survey Laboratory in Sacra-mento, Calif. using a GC with ECD according to a modified versionof US EPA method 502.2 (US EPA, 1995a) described by Crepeauet al. (2004). The modified method provided a lower detection lim-it for measured THMs than the US EPA method. Samples for THMFPwere dosed and incubated using methods described by Crepeau

(ASR) of chlorinated municipal drinking water in a confined aquifer. Appl.

Fig. 3. Volume of water and mass of total trihalomethanes (TTHMs) injected andextracted from well 11N/6E-17F1, Roseville, California, January 2004 to February2008.

6 J.A. Izbicki et al. / Applied Geochemistry xxx (2010) xxx–xxx

et al. (2004). When using this procedure pH is adjusted and buf-fered to 8.3, the sample is dosed with a solution containing free-chlorine, and incubated in the dark for 7 days at 25 �C. Dissolvedorganic C and NH3 concentrations in the sample water are usedto determine the appropriate amount of chlorine to add to main-tain 2–4 mg/L of free-chlorine at the end of the incubation period.If necessary, the samples were diluted to keep the initial DOC inthe range of 0–3 mg/L and the expected THM formation withinspecified ranges. Depth-dependent water-quality samples forHAAs were analyzed by Weck Laboratories, City of Industry, Cali-fornia using GC with ECD using US EPA method 552.2 (US EPA,1995b). Samples for HAAFP were dosed and incubated by the USGeological Survey in Sacramento, in a similar manner to samplesfor THMFP (Zazzi et al., 2005) and analyzed by Weck Laboratories.TTHM, TTHMFP, THAA, and THAAFP were calculated as the sum ofthe individual THMs and HAAs. Laboratory reporting limits forTHMs and HAAs are given in Table 2.

3.3. Estimation of injected water fractions

The fraction of injected water recovered in a water sample wasestimated from Cl� data. Chloride is suitable for this purpose be-cause it is conservative (non-reactive) within the aquifer, and thereis a large difference in the Cl� concentration of native and injectedwater at the site. The average Cl� concentration of the injectedwater (Ci) was 3.7 mg/L, and the average Cl� concentration of thenative ground water (Ca) from the ASR well prior to injection was160 mg/L. Background Cl� concentrations in the three monitoringwells ranged from 150 to 200 mg/L. Assuming only simple mixing,the fraction of injected water in a given sample can be estimatedaccording to the following equations:

CmXm ¼ CiXi þ CaXa ð1Þ

where Cm, Ci, and Ca are the measured Cl� concentration in a watersample, and the average Cl� concentrations in injected water andnative ground water, respectively, and Xm, Xi, and Xa are the frac-tions of extracted water, injected water, and native ground waterin a given sample, respectively.

If the measured fraction (Xm) of any sample is assigned a valueof 1, substituting (1 � Xi) for Xa and rearranging Eq. (1) to solve forthe fraction of injected water in the sample (Xi) produces:

Xi ¼ ðCm � CaÞ=ðCi � CaÞ ð2Þ

The average Cl� concentration of water samples collected dur-ing the baseline sample collection phase was used to representCa for all wells, with the exception of monitoring well 17F3. TheCl� concentration of water from monitoring well 17F3 increasedfrom about 200 mg/L to 300 mg/L during the extraction phase.The change appeared to be abrupt and occurred at the onset ofpumping by the ASR well during the extraction phase. To accountfor that change, the Ca for well 17F3 was assumed to be 300 mg/Lduring the extraction phase.

4. Results

About 1.02 � 106 m3 of chlorinated municipal drinking waterwas injected through the ASR well into the confined aquifer under-lying the City of Roseville between December 14, 2005 and May 5,2006 (Fig. 3). Injected water was stored for 438 days between May6, 2006 and July 17, 2007. About 2.64 � 106 m3 of water was thenextracted from the aquifer through the ASR well between July 18,2007 and February 28, 2008 (Table 1). However, on the basis ofCl� data, the total volume of injected water actually extractedwas only 0.36 � 106 m3, about 35% of the water injected (Fig. 3).The remaining 65% of the injected water was not recovered during

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extraction and presumably moved with the regional ground-waterflow downgradient from the ASR well.

Selected water-quality data from the ASR well and the threenearby monitoring wells are summarized in Table 3. The tableshows the average value of six samples of native water collectedfrom the wells during pre-injection between May 2004 andDecember 2005, the average value of seven samples of water in-jected through the ASR well between December 2005 and May2005, and the highest and lowest values of 35 samples collectedduring the storage and extraction between May 2005 and February2008. Although not differentiated in Table 3, 26 samples were col-lected during storage between May 2006 and July 2007, and ninesamples were collected during extraction between July 2007 andFebruary 2008. Additional water-quality data are available inMWH, Inc. (2004).

Injection rates during the ASR test were as high as 85 L/s. Theinjection rates decreased during the injection period. The ASR wellwas backflushed three times to maintain injection rates (MWH,Inc., 2004). Scanning electron microscopy on particulate materialsampled from the well during backflushing suggested physicalclogging of the well screen, rather than chemical reactions be-tween injected and native water, was the cause of decreasing injec-tion rates (Bassett, 2006).

4.1. TTHMs in the ASR well

Prior to injection, TTHM concentrations in water from the ASRwell averaged 0.8 lg/L (Fig. 4). TTHM concentrations in municipaldrinking water injected into the aquifer ranged from 28 to 53 lg/Lwith an average concentration of 44 lg/L (Fig. 4). During the stor-age phase, TTHM concentrations in water from the ASR well in-creased rapidly to a maximum concentration of 64 lg/L. Thehighest TTHM concentration was 45% higher than the average in-jected TTHM concentration, and was greater than any TTHM con-centration measured in chlorinated municipal drinking waterinjected into the aquifer. These data are consistent with formationof TTHMs within the aquifer after injection. Formation of TTHMsmay occur through the reaction of residual chlorine in the injectedwater, about 0.5 mg/L, with organic C. Dissolved organic C in the

(ASR) of chlorinated municipal drinking water in a confined aquifer. Appl.

Table 3Summary of selected water-quality data from Aquifer Storage Recovery (ASR) well and nearby monitoring wells, Roseville Calif., May 2004 to January 2008.

Constituent ASR well Injectedwateraverage

Monitoring wells

11N/6E-17F1 11N/6E-17F2 11N/6E-17F3 11N/6E-17E1

27.7 meters from ASR well 58.3 meters from ASR well 427 meters from ASR well

Pre-injectionaverage

Rangeduringstorage andextraction

Pre-injectionaverage

Range duringstorage andextraction

Pre-injectionaverage

Range duringstorage andextraction

Pre-injectionaverage

Range duringstorage andextraction

Selected field parameters: pH in standard units, temperature in degrees Celsius, Specific conductance in microSiesmens per centimeter, dissolved oxygen in milligrams per literpH 7.2 6.9 8.7 8.7 7.0 6.8 7.8 7.2 6.8 7.8 7.3 6.7 8.4Temperature 21.5 15.6 20.4 13.8 21.6 13.9 20.5 21.52 13.6 21.4 21.5 21.0 21.5Specific conductance 657 75 662 77 671 69 788 790 69 1115 564 525 728

Major ions: in milligrams per literAlkalinity 62 27 63 25 63 21 60 62 21 57 65 54 62Chloride 162 4 158 4 187 4 189 208 4 303 149 132 153Sulfate 26 8 26 8 29 8 34 33 9 48 24 22 25Calcium 33 10 55 10 38 3 46 43 2 67 28 25 33Sodium 77 4 74 4 82 9 85 86 13 104 74 66 79Potassium 2.4 0.5 2.1 0.6 2.3 0.5 2.4 2.5 0.5 3.2 2.0 1.7 2.1Magnesium 16 2 17 2 18 1 23 20 1 32 14 12 15Silica 67 15 79 8 73 28 82 74 48 84 74 69 88Residue on evaporation

(TDS)455 58 440 34 523 49 539 539 49 706 429 386 474

Nutrients: ammonia and nitrate in milligrams per liter as nitrogen, organic carbon in milligrams per literAmmonia <0.01 <0.01 0.02 <0.01 <0.01 <0.01 0.02 <0.01 <0.01 <0.01 <0.01 <0.01 0.02Nitrate 4.9 0.1 6.0 <0.01 6.4 0.2 7.4 6.5 0.3 7.2 6.2 5.7 6.6Organic carbon 0.5 0.5 9.0 2.0 0.5 0.6 9.4 0.5 0.5 10.4 0.5 0.5 5.0

Selected minor and trace elements: bromide and fluoride in milligrams per liter, iron, manganese, and strontium in micrograms per literBromide 0.35 0.12 0.32 <0.1 0.610 0.02 0.41 0.44 0.05 0.61 0.31 0.25 0.33Fluoride 0.3 0.1 0.7 0.8 0.2 0.1 0.6 0.2 0.1 0.5 0.2 0.1 0.3Iron 39 5 13 <1 5 <1 <1 5 <1 <1 10 <1 <1Manganese <5 <5 <5 <5 <5 <5 <5 <5 <5 <5 <5 <5 <5Strontium 290 40 230 40 360 30 380 0180 10 650 260 180 1000

Selected disinfection by-products: in micrograms per literBromodichloromethane <0.5 0.6 3.5 2 <0.5 0.6 3.6 < 0 <0.5 <0.5 3.6 <0.5 <0.5 <0.5Bromoform <0.5 <1.0 <1.0 <0.5 <0.5 <1.0 1.0 <0.5 <0.5 0.7 <0.5 <0.5 <0.5Chloroform 0.8 5.8 61 42 <0.5 4.5 59 <0.5 <0.5 62 0.3 <0.5 2.4Total trihalomethanes

(THMs)0.8 5.8 63 44 <0.5 0.5 62 <0.5 <0.5 65 0.3 <0.5 2.4

Trichloroacetic acid(TCAA)

<1.0 <1.01 21 19 <1.0 2.5 27 <1.0 <1.0 16 <1.0 <1.0 <1.0

Total haloacetic acids(HAAs)

<1.0 <1.0 26 24 <1.0 <1.0 33 <1.0 <1.0 22 <1.0 <1.0 <1.0

Data analyzed by Bryte Laboratory, California Department of Water Resources, Sacramento, California, from MWH, Inc. (2004). Major ion, nutrient, minor and trace elementdata dissolved.

J.A. Izbicki et al. / Applied Geochemistry xxx (2010) xxx–xxx 7

injected water ranged from 0.9 to 1.8 mg/L, and from <0.2 to1.5 mg/L in native water. Organic C in the aquifer deposits wasnot measured.

After the initial increase, TTHM concentrations gradually de-clined during the remainder of the storage phase to 9.4 lg/L(Fig. 4). Chloride concentrations increased during this period repre-senting an increasing fraction of native water in the ASR well(Fig. 4). The increased fraction of native water accounts for thegradual decrease in TTHM concentrations measured duringstorage.

TTHM concentrations in the ASR well increased at the beginningof the extraction phase reaching a concentration as high as 23 lg/L,about half the TTHM concentration of injected water. Chloride con-centrations decreased as TTHM concentrations increased, consis-tent with an increased fraction of injected water yielded by thewell (Fig. 4). TTHM concentrations in the ASR well at the end ofthe extraction phase were 5.8 lg/L.

4.2. TTHMs in the monitoring wells

The aquifer response to injection, storage and extraction wasalso monitored in monitoring wells 17F2, 17F3, and 17E1 (Fig. 1).As previously discussed, the wells have long-screen intervals that

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completely penetrate the confined aquifer, and are similar inlength to the screened interval of the ASR well.

Water-levels in the three monitoring wells (not shown) in-creased during the injection phase, declined to near pre-injectionlevels and then remained relatively stable during the storagephase. Water-levels declined during extraction from the ASR well.Measured water-levels are discussed by MWH, Inc. (2007) andchanges were consistent with water-level changes expected onthe basis of injection/extraction rates and aquifer properties(MWH, Inc., 2007).

Increases in TTHM concentration associated with the arrival ofinjected water were rapid in monitoring wells 17F2 and 17F3,which are less than 60 m from the ASR well. Within 5 days of thestart of injection TTHM concentrations in wells 17F2 and 17F3were 27 lg/L and 11.7 lg/L, respectively. At that time water fromwells 17F2 and 17F3 was composed of 70% and 30% injected water(estimated from Cl� concentrations), respectively (Fig. 5). The frac-tion of injected water in these wells increased until by early Janu-ary 2006, 27 days after the onset of injection, sampled water fromboth monitoring wells was composed almost entirely of injectedwater (Fig. 5). TTHM concentrations continued to increasethroughout the injection phase, reaching maximum concentrationsgreater than 60 lg/L near the end of the injection phase in May

(ASR) of chlorinated municipal drinking water in a confined aquifer. Appl.

Fig. 4. Chloride, total trihalomethane (TTHM) concentrations, and fraction ofinjected water from ASR well 11N/6E-17F1, Roseville, California, January 2004 toFebruary 2008.

8 J.A. Izbicki et al. / Applied Geochemistry xxx (2010) xxx–xxx

2006, 142 days after the start of injection. Maximum TTHM con-centrations measured in wells 17F2 and 17F3 were greater thanTTHM concentrations in injected water (Fig. 5) consistent with for-mation of THMs within the aquifer after injection.

Fig. 5. Chloride and total trihalomethane (TTHM) concentrations, and fraction of injectCalifornia, January 2004 to July 2008.

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The fraction of injected water in monitoring wells 17F2 and17F3 decreased during storage and extraction in a manner similarto the ASR well (Fig. 4). By the end of the extraction phase, TTHMconcentrations in water from wells 17F2 and 17F3 were 4.5 and0.8 lg/L, respectively (Fig. 5).

Small decreases in Cl� concentrations associated with the arri-val of injected water were apparent in water from 17E1 farthestfrom the ASR well by the end of the injection phase in April2006. However, it was not until depth-dependent samples (dis-cussed later in this paper) were collected in December 2006,359 days after the onset of injection, that low TTHM concentra-tions (between 0.28 and 0.43 lg/L) were first detected in waterfrom this well. After May 2007, TTHM concentrations greater than1 lg/L were consistently present in water from well 17E1 (Fig. 4).TTHM concentrations continued to increase in water from well17E1 until the end of extraction when concentrations reached amaximum of 2.7 lg/L.

4.3. Well-bore flow and depth-dependent water-quality data

Well-bore flow and depth-dependent water-quality data werecollected from the ASR well and the three monitoring wells toprovide a ‘‘snap-shot” of the movement and vertical distributionof injected water within the aquifer (1) prior to injection, (2)during injection, (3) during storage, and (4) during extraction(Table 1).

4.3.1. ASR wellWell-bore flow data collected from the ASR well in December

2005, show that prior to injection almost 60% of the well yieldwas from gravels encountered near the top of the well screen

ed water in water from monitoring wells 11N/6E-17F2, 17F3, and 17E1, Roseville,

(ASR) of chlorinated municipal drinking water in a confined aquifer. Appl.

Fig. 6. Well-bore flow and depth-dependent water-quality data collected under pumped conditions from ASR well 11N/6E-17F1 (DCPW), Roseville, California, December2005 to February 2008.

J.A. Izbicki et al. / Applied Geochemistry xxx (2010) xxx–xxx 9

between 95 and 104 m bls (Fig. 6). Less than 8% of the well yieldwas from the aquifer below 125 m. These deeper deposits havehigher natural gamma counts and lower electrical resistivity valuesthan overlying deposits (Fig. 6), consistent with an increase in claycontent and a decrease in permeability. Depth-dependent water-quality data show that in December 2005, prior to injection, lowconcentrations of TTHMs, about 3 lg/L, were present within theASR well in the low-yielding intervals near the bottom of the well.These low concentrations may have been residual THMs from thepilot test in 2004 (Fig. 6).

It was not possible to collect flow logs from the ASR well duringthe injection or storage phases. However, two well-bore flow logs

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collected during the extraction phase (August 7, 2007 and February8, 2008) showed that the distribution of flow within the well chan-ged during injection. The well-bore flow log collected in August 7,2007, about 19 days after the onset of the extraction phase,showed that only 35% of the flow to the well was from the upper9 m of the well screen compared to 60% in December 2005 priorto injection (Fig. 6). Decreased yield from the upper part of the wellscreen was offset by increased yield immediately below this inter-val (103–113 m bls) and by increased yield from deposits 117 to126 m bls (Fig. 6). Model simulations of well-bore flow presentedlater in this paper suggest that these changes resulted from phys-ical clogging of the well screen.

(ASR) of chlorinated municipal drinking water in a confined aquifer. Appl.

10 J.A. Izbicki et al. / Applied Geochemistry xxx (2010) xxx–xxx

TTHM concentrations in water sampled with depth in the ASRwell shortly after the onset of the extraction phase on August 7,2007 ranged between 18 and 21 lg/L and were relatively uniformwith depth (fig. 6). Chloride data collected with the TTHM datasuggest that the water extracted by the ASR well at this timewas not solely injected water, but rather was an uniform mixtureof as much as 70% native water and as little as 30% injected water.

TTHM data collected from the discharge of the ASR well nearthe end of the extraction phase on February 8, 2008 show thatthe TTHM concentrations in the surface discharge decreased to

Fig. 7. Comparison of well-bore flow, fluid resistivity, Cl� and total trihalomethane data cfrom ASR well 11N/6E-17F1, Roseville, California, December 2005.

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6.3 lg/L. However, the well-bore flow and depth-dependent datashow that the injected and native water were no longer uniformlydistributed throughout the well (Fig. 6). TTHM concentrationsentering the well from the aquifer were as high as 16 lg/L in waterfrom the lower-yielding, finer-grained deposits encountered at125 m bls within the well (Fig. 6). These high TTHM concentrationswere associated with low Cl� concentrations suggesting a highfraction of injected water present at depth. These results suggestthat it was difficult to remove injected water from these lower-per-meability aquifer deposits.

ollected from monitoring wells 11N/6E-17F2 and 17F3, prior to and during injection

(ASR) of chlorinated municipal drinking water in a confined aquifer. Appl.

J.A. Izbicki et al. / Applied Geochemistry xxx (2010) xxx–xxx 11

At most sample depths, TTHM, and Cl� concentrations of waterentering the ASR well near the end of the extraction phase(February 2008) varied inversely (Fig. 6). In general, low Cl�

concentrations, indicative of injected water, had high TTHMconcentrations. However, TTHM concentrations in water yieldedfrom the upper 9 m of the well screen were about 10 lg/L, despitelow Cl� concentrations. On the basis of Cl� data, calculated TTHMconcentrations at that depth would be about 40 lg/L (Fig. 6). Thelow TTHM concentrations in the upper 9 m of the ASR well screennear the end of the extraction phase may be the result of sorptionor partial degradation of THMs within the aquifer.

Fig. 8. Comparison of well-bore flow, fluid resistivity, Cl� and total trihalomethane datduring injection, and during extraction from ASR well 11N/6E-17E1, Roseville, California

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4.3.2. Monitoring wellsWell-bore flow and depth-dependent water-quality data were

collected from monitoring wells 17F2, 17F3, and 17E1 in Novem-ber 2005 prior to injection, in December 2005 about 6 days afterthe onset of the injection phase, in December 2006 during the stor-age phase, and in August 2008 during the extraction phase (Ta-ble 1). Monitoring wells 17F2, 17F3, and 17E1 are about 28 m,58 m and 427 m from the ASR well, respectively.

Prior to injection in November 2005, the distribution of flowinto the three monitoring wells differed and was controlled bythe distribution of hydraulic properties near the well. About 60%

a collected from monitoring wells 11N/6E-17F2, 17F3, and 17E1 prior to injection,, December 2005 to August 2007.

(ASR) of chlorinated municipal drinking water in a confined aquifer. Appl.

Fig. 9. Total trihalomethane (TTHM) and total trihalomethane production ininjected water during incubation, Roseville, California, December 2005.

12 J.A. Izbicki et al. / Applied Geochemistry xxx (2010) xxx–xxx

of the total yield from well 17F2 was from a thin interval between110 and 114 m bls, about 20% of the yield was from about 98 m bls,and about 10% was from about 132 m bls (Fig. 7). In contrast, about90% of the yield from well 17F3 was uniformly distributed between94 and 115 m bls, with only 10% of the yield was from the wellscreen below 115 m bls (Fig. 7). The yield from well 17E1 wasevenly distributed throughout the well screen (Fig. 7). High-yield-ing layers within wells 17F2 and 17F3 were not at the same depthas those in the ASR well (17F1), suggesting that the aquifer is notcomposed of flat-lying layers and that the hydraulic connectionswithin the aquifer may be highly complex.

The distribution of flow into the three monitoring wells chan-ged in response to injection beginning in December 2005. In well17F2, the flow from the thin layer 98 m bls increased from 20%to 40% of the total yield, the yield from about 132 m bls increasedfrom 10% to 30%, and the yield from 110 to 114 m bls decreased to20% of the total yield (Fig. 7). Comparison of the fluid resistivitylogs (Fig. 7) and the temperature logs (not shown) collected priorto and during the injection phase indicate that the injected water(having a higher resistivity than native water) moved into moni-toring well 17F2 through the thin layers identified from the well-bore flow logs (Fig. 7). In well 17F3, the greatest change in wellyield occurred in a thin layer between 113 and 115 m bls. InDecember 2005, after the start of injection, this layer contributedmore than 30% of the total well yield; whereas, prior to the injec-tion phase this layer contributed less than 10% of the well yield(Fig. 7). Similar to monitoring well 17F2, comparison of the resis-tivity and temperature logs collected prior to and during the injec-tion phase indicate that most of the injected water is moving inthin layers (Fig. 7).

TTHM and Cl� data (coupled with fluid resistivity data) suggestalmost complete displacement of native water by injected waterwithin well 17F2 in the high-yielding intervals between 108 and114 m bls and between 130 and 132 m bls after 6 days of injection(Fig. 7). Similar changes in TTHM and Cl� concentrations weremeasured in well 17F3 near 105 m below land surface, althoughthe data show less injected water present in the well at the timeof sample collection (Fig. 7). Injected water completely displacednative water throughout wells 17F2 and 17F3 by February 2006,49 days after the onset of injection (Fig. 5).

Changes in the distribution of well-bore flow in well 17E1 weresmall between well-bore flow data collected prior to and duringthe injection phase (Fig. 7). In addition, there were no apparentchanges in fluid resistivity (Fig. 7); consequently, depth-dependentTTHM and Cl� data were not collected from well 17E1 during theinjection phase.

Well-bore flow data collected during the storage phase (notshown in this paper) about 7 months after injection stopped weresimilar to the pre-injection flow distribution in all three monitor-ing wells. Depth-dependent water-quality data from wells 17F2and 17F3 showed the displacement of injected water by nativeground water flowing with the regional gradient. Displacementof injected water was greatest in the thin, high-yielding intervalsidentified previously. Increases in TTHM concentrations and de-creases in Cl� concentrations measured in well 17E1 farthest fromthe ASR well were consistent with the arrival of a small amount ofinjected water.

Well-bore flow changed in all three monitoring wells as a resultof extraction by the ASR well (Fig. 8). These changes were larger inmagnitude, opposite in direction than changes that occurred dur-ing injection, and occurred within the thin high-yielding layersidentified previously. The changes in the distribution of flow werelarger in magnitude because water was extracted at a greater ratethan it was injected. Pumping from the ASR well induced down-ward fluid movement in wells 17F2 and 17F3 under unpumpedconditions and downward flow occurred within parts of these

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wells even while they were being pumped for logging and samplecollection.

5. THM formation, speciation and degradation

5.1. THM formation

THM formation within aquifers after injection of chlorinatedwater was observed in this study and increased TTHM concentra-tions by 45% compared to TTHM concentrations in injected water.THM formation has been observed in other studies (Miller et al.,1993; Thomas et al., 2000; Fram et al., 2002, 2003) and is limitedprimarily by the amount of free-chlorine in the injected water(Fram et al., 2003).

During a 5-day laboratory incubation period, the TTHM concen-tration in a sample of injected water from the Roseville site, havinga free-chlorine residual of 0.2 mg/L, increased 30% from 50 lg/L to65 lg/L (Fig. 9). During incubation, the free-chlorine declined from0.2 mg/L to less than the detection limit of 0.05 mg/L, after whichformation of THMs ceased. The average chlorine residual in waterinjected into well 17F1 was 0.5 mg/L, suggesting that the 45% in-crease in TTHM concentration estimated from field data at the siteis reasonable.

In contrast, the measured total trihalomethane formation po-tential (TTHMFP) of the injected water was 122 lg/L. As describedin Section 3, TTHMFP data are for samples incubated with suffi-cient free-chlorine to maintain a chlorine residual of 2–4 mg/L atthe end of a 7-day incubation period (Crepeau et al., 2004). Thesehigh formation potentials suggest that the DOC concentration ofthe injected water (0.9–1.8 mg/L) is sufficient to explain the ob-served increase in TTHM concentrations at the site and that if high-er chlorine residuals were maintained in the injected water, evenmore THMs would form during storage within the aquifer.

Given the measured 45% increase in TTHM concentration fromreactions with free-chlorine, the TTHM concentration in extractedwater can be predicted from the measured Cl� concentrations andthe calculated fraction of injected water (as described in Section 3).Calculated TTHM concentrations agree well with TTHM data mea-sured during the storage phase, and only slightly underestimateTTHM concentrations measured during the extraction phase—con-sistent with the expected error associated with changing Cl� con-centrations (Fig. 4).

As stated previously, a mass balance of Cl� indicates that 35% ofthe injected water was recovered and 65% of the injected water re-mained in the aquifer at the end of extraction. Therefore, about30 kg of THMs were not recovered during extraction. Given insituproduction of THMs in the presence of a free-chlorine residual ofabout 45%, TTHMs remaining in the aquifer at the end of extractionwould actually be about 44 kg.

(ASR) of chlorinated municipal drinking water in a confined aquifer. Appl.

J.A. Izbicki et al. / Applied Geochemistry xxx (2010) xxx–xxx 13

5.2. TTHM speciation and degradation

TTHMs in water injected into the ASR well were about 94% chlo-roform and 6% bromodichloromethane by weight. The more bro-mated compounds, bromoform and dibromochloromethane, werenot detected in injected water. The fraction of chloroform and bro-modichloromethane composing the TTHMs remained relativelyconstant during injection (Fig. 10).

During storage and extraction, the fraction of chloroform inwater from the ASR well gradually decreased to about 90% andthe fraction of bromodichloromethane increased. Similar patternswere observed in water from monitoring wells 17F2 and 17F3. Thisbehavior is the opposite of what would be expected given ourunderstanding of the chemical behavior of chloroform and bromi-nated THMs with respect to sorption and degradation (McQuarrieand Carlson, 2003; Pavelic et al., 2005). It is possible that thesechanges may have resulted from preferential formation of bro-mated THMs within the aquifer shortly after injection and therecovery of that initially injected water during the later stages ofextraction.

TTHMFP data for native water from the ASR well and monitor-ing wells show preferential formation of the more highly bromatedcompounds bromoform and dibromochloromethane compared tobromodichloromethane and chloroform. These compounds wouldform in the aquifer during the initial stages of injection if injectedwater mixed with native water before residual free-chlorine wasconsumed (Fram et al., 2003). Depth-dependent samples frommonitoring wells 17F2 and 17F3 collected after the onset of injec-tion during December 2005 show bromoform and dibromochlo-romethane present at concentrations as high as 2.7 and 3 lg/L,respectively. In both monitoring wells, bromoform and dibromo-chloromethane concentrations were higher within intervals whereinjected water moved rapidly from the ASR well through thin high-permeability layers. On the basis of these data, the ratio of chloro-form to bromodichloromethane would shift with time, and withpumping from the ASR well, as water that entered the aquifer dur-ing the initial stages of injection was pumped from the aquifer dur-ing the later stages of extraction.

Fig. 10. Ratio of chloroform to total trihalomethane and ratio of bromodichlorom-ethane to total trihalomethane in water from ASR well 11N/6E-17F1, Roseville,California, November 2005 to February 2008.

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In a homogeneous isotropic aquifer, native water near the injec-tion well would be rapidly displaced by injected water and prefer-ential formation of bromated THMs would occur for only a shortperiod at the beginning of injection. In the more heterogeneousRoseville aquifer system, thin high-permeability layers allow rapidmovement of water away from the ASR well and facilitate mixingof injected water containing free-chlorine and native water foran extended period of time—thereby increasing the formation ofbromated THMs.

The good agreement between measured THMs and THM con-centrations predicted on the basis of Cl� data (Fig. 4) suggest rela-tively little removal of THMs through sorption or degradationduring the early part of the test and that dilution of injected waterwith native water was the primary mechanism of TTHM attenua-tion. However, data collected during later parts of the test frommonitoring well 17E1 (427 m downgradient from the ASR well)indicate that TTHMs and chloroform were removed by sorptionor degradation. Near the end of the extraction phase, as much as10% of the water from well 17E1 was injected water (Fig. 5).Assuming conservative behavior, the expected TTHM concentra-tion in water from this well should have been about 6.5 lg/L. In-stead measured TTHM concentrations in well 17E1 neverexceeded 2.4 lg/L. In addition TTHMs were composed entirely ofchloroform and bromated THMs were absent. These data are con-sistent with preferential removal of bromated THMs through sorp-tion or degradation and even some removal of chloroform.Nicholson et al. (2002) reported that chloroform degrades onlyslowly and under strongly reducing (methanogenic) conditions.Although reducing conditions are present in the aquifer at the sitethe aquifer is not strongly reducing, suggesting that sorption maybe the primary mechanism for chloroform removal in aquifersunderlying the study area.

Similar declines in TTHM concentrations were observed indepth-dependent samples collected from the ASR well in February2006 near the end of extraction. At that time, low Cl� concentra-tions near the top of the well screen (Fig. 6) were consistent witha high fraction of injected water and an expected TTHM concentra-tion of about 37 lg/L. In contrast, the TTHM concentration of waterentering the well from the aquifer through this interval was about10.5 mg/L (Fig. 6). Water recovered at the end of the extractionphase is water that was injected in the early part of the injectionphase. This water has had the longest residence time in the aquifer,was nearest the margin of the injected ‘‘bubble”, and as a conse-quence had the greatest opportunity to interact with native groundwater. The decrease in TTHM concentrations measured from thisinterval of the ASR well is similar in magnitude to the decreasein TTHM concentrations observed in data from the farthest moni-toring well 17E1.

6. THAA formation, speciation, and degradation

Total haloacetic acids (THAAs) concentration in water injectedinto the ASR well averaged about 24 lg/L; 71% of THAAs were tri-chloroacetic acid (TCAA) with smaller amounts of dichloroaceticacid (DCAA). THAA concentrations were slightly more than halfof the DBP concentration in injection water, and as much as 65%of the THM concentration observed in monitoring wells 17F1 and17F2 shortly after injection. In this study, THAA composed a higherfraction of DBPs and persisted longer than THAAs reported in pre-vious studies (Pyne, 2005, 2006). However, THAA concentrationsdecreased more rapidly with time than the THM concentrations.Trichloroacetic acid concentrations decreased to less than thedetection limit of 1 lg/L after about 5–6 months of storage withinthe aquifer leaving dichloroacetic acid as the predominant HAA(Fig. 11). Dichloroacetic acid concentrations subsequently

(ASR) of chlorinated municipal drinking water in a confined aquifer. Appl.

Fig. 11. Total haloacetic acids (THAAs), trichloroacetic acid (TCAA), and dichloro-acetic acid (DCAA) in water from ASR well 11N/6E-17F1, Roseville, California,December 2005 to February 2008.

14 J.A. Izbicki et al. / Applied Geochemistry xxx (2010) xxx–xxx

decreased to less than the detection limit after about 10 months ofstorage (Fig. 11). HAAs were not detected during this study inwater from the farthest monitoring well 17E1, consistent with re-moval and limited transport of HAAs (Pyne et al., 1996; Pyne,2005).

In contrast to field data, total halocetic acid formation potential(THAAFP) data collected as part of this study suggest that fewerhaloacetic acids would form compared to trihalomethanes as a re-sult of the reaction of free-chlorine with native and injected water.On average the THAAFP was only about 20% of the TTHMFP in na-tive and injected water, compared to the 50–60% recorded in thefield data. In addition, THAAFP data showed preferential formationof dibromoacetic acid, to the complete exclusion of the chlorinatedHAAs, while the chlorinated HAAs actually dominated in the fielddata. It is unclear why THAAFP analysis had poor predictive powerrelative to field data.

Fig. 12. Radial movement of water from ASR well 11N/6E-17F1, Roseville,California, assuming a uniform and non-uniform distributions of injected waterinto cylindrical volume of aquifer 50 m thick.

7. Model simulation of well-bore flow

The physical movement of water from the well controls themixing and chemical interaction between injected water and na-tive water. Consequently, understanding the movement of wateris important in understanding the movement and fate of THMsin the aquifer. Specific issues include: (1) movement of injectedwater from the well, (2) the shape of the injected bubble aroundthe ASR well, (3) the extent of interaction between injected and na-tive water, and (4) the long-term efficiency and operation of theASR well. To help address these issues well-bore flow was evalu-ated using simple piston-flow calculations and a two-dimensionalradial ground-water flow model developed using MODFLOW(Harbaugh and McDonald, 1996). Well-bore flow data collectedduring pumping from the ASR well at a rate of 214 L/s were usedfor the piston-flow calculations and model calibration. Injectionwas assumed to be the mirror image of extraction and water move-ment from the ASR well during injection was simulated usingMODPATH at the injection rate of 85 L/s (Pollock, 1994).

Please cite this article in press as: Izbicki, J.A., et al. Aquifer Storage RecoveryGeochem. (2010), doi:10.1016/j.apgeochem.2010.04.017

7.1. Piston-flow calculations

Well-bore flow data indicate that about 60% of the yield to thewell was from a high-permeability layer in the upper 9 m of thescreened aquifer (94 and 103 m bls). Assuming injection is the mir-ror image of extraction and representative porosity values, thensimple piston-flow calculations indicate that injected water wouldmove about 1.5 times farther from the ASR well in the high-perme-ability layer than predicted assuming uniform aquifer properties(Fig. 12). However, even assuming water is moving through theupper 9 m of the aquifer deposits between the ASR well and mon-itoring well 17F3 an unrealistically low porosity is required forwater to move the 58 m between these wells in 5 days (Fig. 12)as observed in field data (Fig. 5). Piston-flow calculations are con-sistent with the results of EM flow logging in the monitoring wells(Fig. 7) that suggest injected water is moving from the ASR wellthrough the aquifer in very thin, highly permeable interconnectedunits. These calculations suggest that although the aquifer depositsappear to be relatively homogeneous geologically, the hydraulicproperties are highly heterogeneous.

7.2. Radial-flow model

To address the effect of aquifer heterogeneity on the movementof water from the ASR well, the computer program MODFLOW(Harbaugh and McDonald, 1996) was used to develop a simplifiedaxially symmetric, radial model of ground-water flow near the ASRwell. The computer program MODPATH (Pollock, 1994) was thenused to simulate the movement of water particles within the mod-el domain to assess the effect of aquifer heterogeneity on themovement of injected water from the ASR well. The two modelshave recently been combined into a single compute program Ana-lyzeHole (Halford, 2009).

(ASR) of chlorinated municipal drinking water in a confined aquifer. Appl.

Fig. 13. Model grid, lithology, and well-screen intervals used to simulate flow to ASR well 11N/6E-17F1, Roseville, California.

Table 4Details of radial ground-water flow model construction.

Data Value

Spatial discretizationGrid dimensions 150 m thick by

60,960 m wideNumber of layers 1Number of rows 100Thickness of rows 1.5 mNumber of columns 60Size of columns VariableColumn 1 (well) 1Column 2 (casing and screen) 0.05Column 3 (gravel pack/seal) 0.5Column 4–60 Multiplier 1.187Lateral boundary condition No flowBottom boundary condition No flowUpper boundary condition Water table(Initial water-level 32 m below land surface.

Pumping water-level of 49.1 m below landsurface estimated from Theis equationfor a pumping rate of 85 L/s)

Hydraulic propertiesPorosity, unitless 0.20Specific storage, m�1 4.5 � 10�7

Vertical anisotropy, unitless 0.1Hydraulic conductivity, in m/dWell casing 0Clay 0.08Sandy clay 0.19Clayey sand 1.1Sand 7.6Gravely sand 60–76Gravel (high-permeability unit) 300Gravel pack 500

Temporal discretizationStress periods 1Length of stress period, in days 180Time steps 42Time step multiplier 1.25Initial time step, in seconds 1Pumping rate, L/s (calibration to flow logs) 214Pumping rate, L/s (particle tracking) 85

J.A. Izbicki et al. / Applied Geochemistry xxx (2010) xxx–xxx 15

7.2.1. Model developmentThe two-dimensional, radial model consisted of one layer, dis-

cretized into 60 rows and 79 columns, representing a cylinder ofaquifer material having a radius of 60,960 m and a thickness of120 m (Fig. 13). The large radius of the model domain ensuredmodel boundary conditions did not influence model output. In thisconfiguration the columns are horizontal and the rows are vertical.The model was finely discretized near the well with columns thatrepresent the well-bore (column 1), the casing and screen (column2), and the gravel pack (column 3) (Table 4). The model has thecapability of simulating changes in well efficiency by changingthe properties of the well screen. Column widths increased withdistance from the well (Table 4).

Hydraulic properties representing aquifer materials (modelrows 4–60) were assigned from lithologic, geophysical, and aqui-fer-test data. The model specific storage of 4.5 � 10�7 m, deter-mined from aquifer-test data (MWH, Inc., 2004), and the verticalanisotropy of 0.1 were held constant during calibration. The aqui-fer materials were simplified into six lithologic classifications: clay,sandy clay, clayey sand, sand, gravely sand and gravel (Fig. 14). Themodel was calibrated to a simulated pumping rate of 214 L/s byadjusting, within reasonable ranges, hydraulic conductivity valuesassociated with the lithology assigned to individual model rows.Calibration continued until there was a reasonable match betweenthe simulated and measured distribution of flow within the ASRwell and between the simulated and measured drawdown of18.6 m. During calibration, the summed hydraulic conductivitiesof the aquifer materials was maintained within ±10% of the aquifertransmissivity of 2.8 � 103 m2/d estimated from aquifer-test data(MWH, Inc., 2004). Additional calibration was done to match thesimulated movement of particles using MODPATH (Pollock, 1994).

The computer program MODPATH (Pollock, 1994) requires esti-mates of aquifer porosity (effective) to simulate the movement ofwater particles within the model domain. MODPATH calculationswere done at the injection rate of 85 L/s rather than the pumpingrate of 214 L/s. As demonstrated with the piston-flow calculations,the effective porosity values need to be between 0.15 and 0.28 inorder to match the arrival times of injected water with data

Please cite this article in press as: Izbicki, J.A., et al. Aquifer Storage Recovery (ASR) of chlorinated municipal drinking water in a confined aquifer. Appl.Geochem. (2010), doi:10.1016/j.apgeochem.2010.04.017

Fig. 14. Measured and simulated well-bore velocity logs from ASR well 11N/6E-17F1, Roseville, California, December 2005 to February 2008.

16 J.A. Izbicki et al. / Applied Geochemistry xxx (2010) xxx–xxx

measured at monitoring wells 17F3 and 17E1. On the basis of thesedata a porosity of 0.2 was used for model calculations presented inthis paper. The particles were initially distributed along the wellscreen and their position within the model domain during simu-lated injection from well 17F1 was calculated for each model timestep. The particles are able to move into model rows having differ-ent hydraulic properties during the simulation.

In order for the particle movement to match the arrival of in-jected water in the 3 monitoring wells it was necessary to furtheradjust the hydraulic conductivity values in the high-permeabilityunit near the top of the screened interval of the ASR well. This unitwas identified on the basis of flow-log data (Fig. 6). Lithologic datashow this 9 m thick layer contains a 3 m thick gravel unit. Thehydraulic conductivity of the gravel unit was increased from76 m/d to 300 m/d to simulate the arrival of particles at monitoringwell 17F3, 58.8 m from the ASR well, within 5 days of the onset ofinjection. Although high, this value is within the published rangefor alluvial gravel deposits (Freeze and Cherry, 1979). Use of thishigh hydraulic conductivity value simulated arrival of particles atmonitoring well 17E1, 427 m from the ASR well, in about 140 daysafter injection. Small changes in Cl� concentration, indicative of thepresence of injected water, first occurred at that time (Fig. 6). Aspreviously discussed, increases in TTHM concentrations did not oc-cur in water from well 17E1until at least December 2006, almostone year after the onset of injection, indicative of loss of TTHMsthrough sorption or degradation.

The final simulated model transmissivity was 2760 m2/d. Thisvalue is about 1% lower than the transmissivity determined fromaquifer-test data. Model calibrated hydraulic conductivities rangedfrom 0.08 m/d for the overlying fine-grained deposits to 300 m/dfor the high-permeability layer near the top of the well screen(Table 4).

7.2.2. Model resultsInjection from the ASR well was simulated at the rate of 208 L/s

(3300 gal/min), the injection rate of the ASR well during collectionof well-bore flow logs. This rate is greater than the pumping rate ofthe well during extraction (158 L/s or 2500 gal/min). The pumpingrate is greater because water was pumped to waste during flowlogging and there was no backpressure on the pump resulting from

Please cite this article in press as: Izbicki, J.A., et al. Aquifer Storage RecoveryGeochem. (2010), doi:10.1016/j.apgeochem.2010.04.017

delivery into the water distribution system. The model simulateddrawdown of 17.7 m (58 ft) closely approximated the measureddrawdown of 18.6 m (61 ft) in the ASR well during the December2005 well-bore flow test. Simulated well-bore flow closelymatched the measured well-bore flow data collected from theASR well in December 2005 prior to the onset of injection, withthe exception of about 98 and 108 m below land surface(Fig. 14). Measured data suggest slightly more aquifer heterogene-ity than the model results actually simulate. Simulated data alsomatch the arrival of injected water at the monitoring wells.

For purposes of this simulation, the well screen was assumed tobe 100% efficient. To determine the effect of physical clogging (wellefficiency) during injection on drawdown and simulated well-boreflow, aquifer properties were held constant and well efficiency wasadjusted by reducing the hydraulic conductivity of the well screen.The well was assumed to be 100% efficient prior to the injectionphase. Well-bore flow data collected prior (December 2005) toand during the injection phase (August 2007) indicate that theyield from the high-permeability layer in the upper 9 m of the wellscreen (94–103 m bls) was reduced from 60% to 35% of the totalwell yield. The decrease in simulated yield between 94 and103 m was offset by increases in yield immediately below thisinterval and from deeper within the well (Fig. 14). In addition,the efficiency of the well screen was reduced by 50%. Physical clog-ging of the well screen and gravel pack during the injection phaseis the suspected cause of the change in well-bore flow distributionand specific capacity.

The simulated change in well efficiency decreased the simu-lated yield from this interval to about 35% of the total well yieldand increased simulated drawdown from 18.6 m to 20.7 m. Thesimulated well-bore flow log and drawdown compare favorablyto the well-bore flow and drawdown data collected in August2007 near the beginning of the extraction phase, after injectionand physical clogging of the well occurred (Fig. 14). The modelsimulations indicate that decreased well efficiency, resulting fromthe injection, is the likely cause for the observed change in well-bore flow. The high-permeability layer is most affected by thephysical clogging because most of the injected water originally flo-wed through this layer.

To simulate the measured well-bore flow data collected duringthe extraction phase (February 2008), the hydraulic conductivity ofthe well screen opposite the high-permeability layer was furtherreduced, representing a well efficiency of 25% for the high-perme-ability layer. The reduction in well efficiency resulted in additionalchanges in the simulated well-bore flow log and an increase insimulated drawdown. The simulated well-bore flow log and draw-down compare favorably with data collected in February 2008 nearthe end of extraction (Fig. 14), and are consistent with continuedclogging of the well by particulate material emplaced in the gravelpack and nearby aquifer deposits during injection being pulledback toward the well by pumping.

MODPATH was used to simulate the movement of the injectedwater away from the ASR well through the heterogeneous aquifersystem. Particles were tracked from the well screens with MOD-PATH to simulate the movement of the injected water from thewell while injecting at the rate of 85 L/s for 140 days. Particleswere seeded in model cells representing the screen in column 3and were distributed proportionately to the simulated flow intoeach cell; therefore, each particle simulated the same amount ofinjected flow from the well. For the purposes of this simulationthe well was assumed to be 100% efficient and the effect of physicalclogging of the well screen was not considered. As a consequence,the model simulations are most representative of the aquifer re-sponse to injection at the beginning of the injection phase. Simu-lated pressure increases from injection were distributedthroughout the aquifer faster than the particles and the results

(ASR) of chlorinated municipal drinking water in a confined aquifer. Appl.

Fig. 15. Simulated pressure response and particle movement during injection fromASR well 11N/6E-17F1, Roseville, California.

Fig. 16. Sensitivity of radial ground-water flow and particle-tracking model resultsto changes in the hydraulic conductivity of the high-permeability layer between 94and 103 m below land surface and to changes in (effective) porosity, Roseville,California.

J.A. Izbicki et al. / Applied Geochemistry xxx (2010) xxx–xxx 17

are consistent with measured increases in water-levels in the mon-itoring wells (Fig. 15). The simulated movement of water from the

Please cite this article in press as: Izbicki, J.A., et al. Aquifer Storage RecoveryGeochem. (2010), doi:10.1016/j.apgeochem.2010.04.017

well was greater through the thin, high-permeability coarse-grained deposits, and the particles reach monitoring well 17F3within the time observed in the field and in about the proportionsexpected on the basis of the fraction of injected water estimatedfrom Cl� data. Model results show the importance of aquifer heter-ogeneity on the movement of injected water, its future recovery,and the recovery of associated THMs.

7.3. Model sensitivity

A sensitivity analysis was done to determine the sensitivity ofmodel results to hydraulic conductivity values in the thin high-permeability layer and porosity values used in the model. The sen-sitivity of model results to changes in the hydraulic conductivity ofthe thin high-permeability layer (94–103 m bls) was tested by firstincreasing the calibrated hydraulic conductivity of the model rowsrepresenting those deposits from 300 m/d to 450 m/d and then bydecreasing the value to 150 m/d. In both scenarios the model wasrerun without adjusting other properties and the overall transmis-sivity changes in relation to the changes in hydraulic conductivity.Increases in hydraulic conductivity produced small changes in thesimulated well-bore flow logs but did not provide the best fit to theobserved data. Decreases in hydraulic conductivity redistributedflow into the well and the simulated flow log no longer closelymatched the observed data. Increases and decreases in hydraulicconductivity of the high-permeability layer altered the simulatedresponse in accordance with the change in the simulated transmis-sivity. Increases in hydraulic conductivity within the high-perme-ability layer increased the simulated movement of water fromwell 17F1 during injection and decreases in hydraulic conductivitydecreased the simulated movement of water from the well(Fig. 16).

The sensitivity of model results to changes in porosity wastested by changing the porosity of the entire model between 0.1and 0.25. Simulated well-bore flow logs and model drawdownwere insensitive to porosity. However, as expected the resultsshow that the simulated particles moved further from the well asporosity decreased. The best match between simulated particlemovement and arrival of water at the observation wells was forthe model simulated porosity of 0.2.

(ASR) of chlorinated municipal drinking water in a confined aquifer. Appl.

18 J.A. Izbicki et al. / Applied Geochemistry xxx (2010) xxx–xxx

7.4. Model utility and limitations

The two-dimensional radial-flow model developed as part ofthis study to interpret well-bore flow and depth-dependentwater-quality data from the ASR well is a simplified representationof the flow system around the well. The simulation of flat-lyingareally extensive aquifer materials is not intended to accuratelyrepresent regional changes in subsurface geology, including theareal extent of these materials and the hydraulic connection be-tween those materials. As previously discussed, comparison ofwell-bore flow logs from the ASR well and the nearby monitoringwells suggest that the hydraulic connections between permeableunits within the aquifer are not flat-lying. Similarly, the flat watertable does not represent regional ground-water flow or interac-tions between pumping wells. However, this simplified radial-flowmodel does provide a qualitative understanding of the effect ofvariations in hydraulic properties on the movement of water in-jected into the ASR well. The model represents an improvementover other simple models used to analyze the effect of aquifer het-erogeneity on ASR by Maliva et al. (2006) in that it allows verticalmovement of water between units having different hydraulicproperties.

The model calibration to well-bore flow data and simulated re-sponse was adequate to meet the needs of this project and closelyapproximated changes in well-bore flow resulting from injectionand the arrival of injected water at nearby monitoring wells. Amore thorough calibration to available water-level data from near-by monitoring wells would yield more refined estimates of aquiferhydraulic conductivity, effective porosity, and storativity.

8. Summary

About 1.02 � 106 m3 of chlorinated municipal drinking waterwas injected into a confined aquifer, 94–137 m below Roseville,California, between December 2005 and April 2006. The waterwas stored in the aquifer for 438 days, after which 2.64 � 106 m3

of water were extracted between July 2007 and February 2008. ACl� mass balance indicates that 35% of the injected water wasrecovered and 65% of the injected water remained in the aquiferat the end of extraction.

Formation of THMs in the presence of free-chlorine in the in-jected water was observed during laboratory experiments andwithin the aquifer. THM formation in the aquifer increased TTHMconcentrations in injected water by 45%. As much as 44 kg ofTTHMs may have remained in the aquifer at the end of extraction.The formation of THMs was limited by the concentration of free-chlorine residual in the injected water. Minimizing the free-chlo-rine in the injected water may minimize TTHM formation afterinjection. TTHMFP data suggest that increasing the free-chlorineresidual in the injected water would increase the formation ofTTHM after injection.

Well-bore flow data collected from the ASR well show that asmuch as 60% of the injected water entered the aquifer through a9 m thick, high-permeability layer near the top of the screenedinterval of the confined aquifer. Aquifer heterogeneity caused in-jected water to move rapidly through a 3 m thick gravel unit with-in this high-permeability layer to nearby monitoring wells, and tomove further than expected assuming uniform aquifer properties.

Aquifer heterogeneity also enhanced mixing of native anoxicground water with oxic injected water, promoting removal ofTHMs near the margin of the injection front. Three to 4-fold re-moval of THMs, primarily through sorption, was observed in thefarthest monitoring well 427 m downgradient, and in depth-dependent water samples collected from the injection/extractionwell near the end of the extraction phase of the study. Highly bro-

Please cite this article in press as: Izbicki, J.A., et al. Aquifer Storage RecoveryGeochem. (2010), doi:10.1016/j.apgeochem.2010.04.017

minated THMs formed near the injection well as injected watercontaining free-chlorine mixed with native water were sorbed ordegraded. Haloacetic acids (HAAs) were completely sorbed or de-graded during the experiment. The more bromated haloacetic acids(HAAs) more rapidly sorbed or degraded than the less brominatedHAAs.

In homogeneous isotropic aquifers, native water near the injec-tion well is rapidly displaced by injected water. Mixing of injectedwater with native water is limited by the simple geometry of theinjected ‘‘bubble” near the ASR well. Assuming a low natural gradi-ent and slow movement of injected water away from the well dur-ing storage, extraction would be the mirror image of injection. Inthe more heterogeneous aquifer underlying Roseville, thin high-permeability layers allow rapid movement of the injected wateraway from the ASR well.

The physical movement of water in the aquifer through thinhigh-permeability layers enhances mixing of native and injectedwater. As a consequence, reactions that occur near the margin ofthe injection bubble in the presence of free Cl� during the earlypart of the injection, such as preferential formation of bromatedTHMs, are enhanced by increased mixing with native water. Simi-larly, THM removal by sorption or degradation within the aquifer isincreased by the mixing of oxygenated injection water with reduc-ing native water in the thin highly permeable layers within theaquifer. As a result, injected water that has moved farthest fromthe well has been subjected to the greatest removal of THMs. Inthe Roseville area, this may be important for management of waterquality because most of the injected water and associated THMswere not recovered during the extraction phase.

Physical clogging of the well during injection resulted in in-creased drawdown during extraction and changes in the distribu-tion of flow in the well. The extent of clogging during the testwas not sufficient to limit the use of the well for ASR, althoughbackflushing was required to remove particulates entrained inthe gravel pack and formation near the well. Physical clogging ofhigh-yielding intervals in the ASR well may have limited the recov-ery of injected water.

Acknowledgements

Injection, storage, and extraction of water, including samplecollection from the ASR well and the monitoring wells, were doneby the City of Roseville and MWH, Inc., with assistance from theCalifornia Department of Water Resources. Well-bore flow logand depth-dependent data collection were cooperatively fundedby the City of Roseville and the US Geological Survey with logisticalsupport by MWH, Inc.

The authors acknowledge assistance from the staff of the City ofRoseville, Environmental Utilities Department and staff of MWH,Inc. for access to wells and assistance with data collection. Theauthors also acknowledge Mark Nordberg and Chris Bonds of theCalifornia Department of Water Resources for their assistance withdata collection and interpretation, and Kathryn Crepeau and Wil-liam Kerlin for laboratory analysis of disinfection by-productsmeasured as part of this study.

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