Cation exchange in a glacial till drumlin at a road salt storage facility

Journal of Contaminant Hydrology 106 (2009) 118–130

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Journal of Contaminant Hydrology j o u r n a l h o m e p a g e : w w w. e l s ev i e r. c o m / l o c a t e / j c o n h yd

Cation exchange in a glacial till drumlin at a road salt storage facility David W. Ostendorf a,⁎, Baoshan Xing b, Niki Kallergis a a b

Civil and Environmental Engineering Department, University of Massachusetts, Amherst, MA, 01003, United States Plant, Soil, and Insect Sciences Department, University of Massachusetts, Amherst, MA, 01003, United States

a r t i c l e

i n f o

Article history: Received 14 February 2008 Revised 8 December 2008 Accepted 2 February 2009 Available online 13 February 2009 Keywords: Road salt storage Cation exchange Glacial till Drumlins Groundwater pollution

a b s t r a c t We use laboratory and field data to calibrate existing geochemical and transport models of cation exchange induced by contamination of an unconfined aquifer at a road salt storage facility built upon a glacial till drumlin in eastern Massachusetts. A Gaines and Thomas selectivity coefficient K models the equilibrium sodium and divalent cation distribution in the groundwater and solid matrix, while an existing method of characteristics model describes the advective transport of total dissolved cations and sorbed sodium. Laboratory isotherms of split spoon soil samples from the drumlin calibrate K with an average value of 0.0048 (L/g)1/2 for a measured cation exchange capacity of 0.057 meq/g dry soil. Ten years of monitoring well data document groundwater flow and the advection of conservative chloride due to outdoor storage and handling of road salt at the site. The monitoring well cation data and retarded transport model offer an independent K calibration of 0.0040 to 0.0047 (L/g)1/2: the consistency of the field and laboratory selectivity coefficient calibrations endorse this application of the Gaines and Thomas and method of characteristics models. The advancing deicing agent plume releases divalent cations from the till into the groundwater, so that monitoring well samples do not reflect the chemical composition of the road salt. In this regard, dissolved divalent cation milliequivalent concentrations are as high as 80% of the total dissolved cationic concentrations in the salt contaminated monitoring well samples, far greater than their 2.5% level in the road salt stored at the site. Cation exchange can thus obscure attempts to hindcast stored road salt sodium water table concentration from monitoring well sample stoichiometry, or to predict sodium impacts on groundwater or receiving stream quality downgradient of the well. © 2009 Elsevier B.V. All rights reserved.

1. Introduction 1.1. Cation exchange and road salt contamination of the subsurface Subsurface cation exchange has been documented in natural settings such as seawater intrusion (Appelo, 1994) as well as anthropogenic applications like coal gasification (Humenick and Mattox, 1978), wastewater infiltration (DeSimone et al., 1997), and brine disposal (Hanor, 2007). Cation exchange chemistry and transport equations must be specified in order to uniquely determine the concentrations of sorbed and dissolved partitions in the subsurface environment. A

⁎ Corresponding author. Fax: +1 413 545-2202. E-mail address: [email protected] (D.W. Ostendorf). 0169-7722/$ – see front matter © 2009 Elsevier B.V. All rights reserved. doi:10.1016/j.jconhyd.2009.02.002

number of controlled field experiments accordingly calibrated a coupled cation transport model with hydraulic and concentration monitoring well data and laboratory isotherms. These experiments included the advection, dispersion, and exchange of injected cations in a carbonate sand unsaturated zone in Texas (Grove and Wood, 1979), a leaky confined aquifer in California (Valocchi et al., 1981), and sandy aquifers in Ontario (Dance and Reardon, 1983) and Denmark (Bjerg and Christensen,1993). Timms and Hendry (2007) studied the geochemistry and diffusion of cations in ambient groundwater in a clay rich aquitard in Saskatechewan. Road salt has also been identified as an agent of cation exchange. Shanley (1994) inferred cation exchange from observed chloride, sodium, and divalent cation concentrations in a Massachusetts stream receiving baseflow impacted by highway runoff. Labadia and Buttle (1996) offered a qualitative

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account of sodium and calcium exchange in soil samples from the unsaturated zone adjacent to deiced highways in southern Ontario. Pugh et al. (1996) used cation exchange to qualitatively interpret cation ratios in near surface groundwater wells adjacent to a deiced highway in Maine. Bench scale experiments were also conducted as part of this qualitative study of exchange chemistry. Norrstrom and Bergstedt (2001) used column studies to attribute mobilization of divalent cations observed in Swedish roadside soils to salt applications; Lofgren (2001) confirmed sodium and divalent cation exchange induced by road salt in soil samples collected at varying distances from Swedish highways. Rhodes et al. (2001)

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suggested that spatial changes of riverine water quality samples throughout a small watershed in western Massachusetts may reflect cation exchange induced by highway deicing agents. Rosfjord et al. (2007) studied the temporal variation of lacustrine water quality samples, and concluded that road salt induced cation exchange increases calcium and magnesium concentrations in the northeastern United States. These soil, water, and groundwater chemistry studies of road salt induced cation exchange were not accompanied by hydraulic data and transport models, as was done in the controlled field experiments by Grove and Wood (1979) and their successors however, and remained qualitative as a consequence.

Fig. 1. Prior outdoor deicing agent storage pile and loading area at a storage facility in eastern Massachusetts. Deicing agents were stored outdoors at the center of the water table atop a glacial till drumlin from the 1960s to the mid 1980s, then stored indoors but loaded and delivered outdoors until 2001 (Ostendorf et al., 2006). Water table contours are shown; they lie 1 to 3 m below the ground surface.

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1.2. Scope of our investigation and relation to previous work We use laboratory isotherms and a spatially resolved data set of sodium and grouped divalent cation concentrations in monitoring well samples with long periods of record to calibrate a geochemical groundwater transport model of road salt induced cation exchange downgradient of an outdoor storage site. We incorporate laboratory cation exchange data with Gaines and Thomas (1953) calibration into the advective transport model of Charbeneau (1981), and calibrate the analysis with ten years of groundwater data from a well characterized salt storage facility constructed over a glacial till drumlin (Ostendorf et al., 2004). The continuous, high sodium concentrations at the water table permit the use of geochemical and transport models simpler than those used to describe the controlled field experiments cited above. In the former regard, exchange between a single monovalent and a single divalent cation is governed by one selectivity coefficient and a closed form equation, rather than the more elaborate codes needed for geochemical equilibrium among multiple constituents (Parkhurst and Appelo, 1999). Sodium dominates all other monovalent cations at our site, and the calcium and magnesium concentrations are grouped in the field and laboratory experiments in order to support this simple application of a single selectivity coefficient. In the latter regard, the use of a continuous water table contaminant source with gradual changes in dissolved concentration permits us to ignore dispersion and focus attention on geochemistry rather than transport. Dispersion is locally important near strong concentration gradients established by injection, and the second order derivatives complicated the finite element (Valocchi et al., 1981), finite difference (Grove and Wood, 1979; Bjerg et al., 1993), and mixing cell (Dance and Reardon, 1983) transport models needed to resolve cation spikes in the groundwater. The advective transport model of Charbeneau (1981) is a far simpler first order analysis of observer travel times along streamlines from the water table to the monitoring wellscreens, readily solved by finite difference approximation. Our choice of total dissolved cation and sorbed sodium concentrations as unknowns further simplifies matters by eliminating reactions from the transport equations—the former are conservative, so that an observer traveling with water molecules along the aquifer streamline will see the same total dissolved concentration, although the composition of the solution will vary. Indeed, the total dissolved cation transport model must match its chloride counterpart to preserve charge neutrality, and the latter has already been published at the site (Ostendorf et al., 2006). The governing transport equation for sorbed sodium exhibits retardation, so that a slower observer along the same streamline will always see the same sorbed sodium concentration, with a retardation factor dependent on the cation exchange reaction. As a practical matter, our period of record is longer than the controlled field experiments cited above, and accordingly better elucidates evolving geochemistry in each well. This investigation quantifies the impact of a widely used salt storage practice on groundwater, corroborates the qualitative conclusions reached by the prior investigators of road salt induced cation exchange, and joins the work of Bjerg et al. (1993) and Timms and Hendry (2007) as another accurate

application of laboratory calibrated Gaines and Thomas (1953) equilibrium to a controlled field experiment. 2. Site description and laboratory isotherm calibration 2.1. Site description Our controlled field experiment is a 20 year period of outdoor road salt storage at a Massachusetts Highway Department facility in eastern Massachusetts (Fig. 1). The deicing agent storage facility was constructed in the mid 1960s, and salt (NaCl) and premix (Na0.884Ca0.116Cl1.12) were stored outdoors in a pile near monitoring well cluster T through the mid 1980s. The deicing agents were stored indoors but delivered and loaded outdoors on the Prior Loading Area shown in Fig. 1 from the mid 1980s through 2001. Delivery, storage, and loading have been performed indoors since the completion of a new structure in 2001. The facility lies on top of a glacial drumlin, and the water table forms 1 to 3 m below the ground surface. Some of the precipitation through the outdoor storage and handling piles infiltrated through the weathered pavement, delivering saline recharge to the water table. MassHighway records suggest that 93.2% by mass of the deicing agents was salt, so that the assumed composition of the saline recharge from the piles is given by a dissolved sodium fraction κ of 0.975 as cited in Table 1, which summarizes site characteristics. The drumlin is 1200 m long and 800 m wide. It consists of 1–5 m of topsoil and fill, underlain by 5–10 m of weathered brown till that in turn overlies a 20–25 m of gray till. Ostendorf et al. (2004, 2006) calibrated an axisymmetric hydraulic model of the northern half of the drumlin with hydraulic head and isotopic data from a series of monitoring well clusters constructed at the site. These data yield the average porosity value n of 0.23 cited in Table 1, which agrees with the average moisture content of 11% found gravimetrically in saturated split spoon samples. The calibrated brown and gray till permeabilities are 1.1 × 10− 13 and 1.4 × 10− 15 m2 respectively, and flow is slower and more vertical in the latter formation. The present analysis focuses on reactions not transport, so we adopt a constant average linear velocity v along the streamline from the water table to each wellscreen in order to calibrate the advective transport theory for cation exchange. Grain size distributions of split spoon samples taken during monitoring well construction suggest that the two till units are comprised of the same soil, with a typical composition of 17% gravel, 39% sand, 33% silt, and 11% clay.

Table 1 Glacial till drumlin parameter values at road salt storage facility. Parameter

Symbol

Units

Value

Bulk density Porosity Sodium fraction Sorbed cation sites Ambient dissolved cations Ambient sorbed sodium Selectivity coefficient

ρB n κ Q CO qNaO K

g/L none none meq/g dry soil meq/L meq/g (L/meq)1/2

2040 a 0.23 a 0.975 b 0.057 c 5c 0.0015 c 0.0048 c

a b c

Ostendorf et al. (2006). Highway deicing agent storage facility records. Analyses of soil and soil moisture from split spoon samples.

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2.2. Soil chemistry and laboratory isotherm experiments The circles in Fig. 2 display the vertical profile of sorbed cation exchange sites Q (expressed in milliequivalents per dry soil mass), which vary from 0.0484 to 0.0585 meq/g dry soil. These values were determined by applying the Buchner Funnel Procedure of the Ammonium Acetate Method at pH of 7.0 (Rhoades, 1982) to split spoon samples from boreholes D2 and C2, which were drilled during June 2003. Shallow samples were corrected for soluble salt contamination in accordance with Richards (1954), and extracts were analyzed by atomic adsorption spectrometry in the Soil Testing Laboratory of the Department of Plant, Soil, and Insect Sciences at the University of Massachusetts Amherst. We equate the average of 0.0570 meq/g dry soil to Q for model calibration, as cited in Table 1. This average is about 0.5 times the clay content of 11%, somewhat lower than the empirical estimate of 0.7 put forward by Appelo and Postma (2005). The figure also shows observed sorbed sodium qNa and divalent q2 cations from the boreholes, and the relative abundance of sorbed sodium above the 10 m depth is striking and suggests the displacement of sorbed divalent cations. Soil samples from the four deepest borehole intervals yield an initial or ambient sorbed sodium concentration qNaO of 0.0015 meq/g, while groundwater from the deepest monitoring wells provide an ambient total dissolved cation concentration CO of 5 meq/L. In the latter regard, all dissolved cation concentrations are expressed in milliequivalents per volume of solution. We conducted a series of batch equilibrium experiments on six of the deeper D2 split spoon samples to determine the selectivity coefficient K for sodium and grouped divalent cation exchange in the Environmental Engineering Program Laboratory of the University of Massachusetts Amherst. Soil moisture extract analyses confirmed that the road salt had not yet infiltrated into these 15–30 m deep samples at the time of drilling, so that uncontaminated soil was used for the batch equilibrium tests. Ten g subsamples were oven dried overnight at 105 deg C, pulverized with a mortar and pestle, passed through a 2 mm sieve, and introduced into a 50 mL polypropylene centrifuge tube. The tube contained 40 mL of the

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target NaCl solution, prepared with certified NaCl (Fisher Scientific; Springfield, NJ) and deionized water. Saline solutions of nominal 4, 20, 40, 100, and 200 meq/L concentration were used to establish the laboratory isotherms. The soil suspension was shaken for 6 h on a Series 51704 horizontal reciprocating shaker (Cole Palmer; Vernon Hills, IL) and spun for 15 min in an IEC Centra CL2 benchtop centrifuge (Cole Palmer; Vernon Hills, IL) at 4500 rpm. The supernatant was filtered with a 20 mL plastic syringe and 0.45 μm filter (Millipore Corp.; Bedford, MA), then acidified to 1% with nitric acid. The acidified samples were analyzed for metals using an Optima 5300 DV ICP (Perkin-Elmer; Shelton, CT) with optical emission spectrometer and an autosampler. 2.3. Groundwater quality methods Monitoring wells were sampled monthly or quarterly in order to calibrate the transport model and offer an independent field calibration of the selectivity coefficient. Ostendorf et al. (2006) describe the monitoring wells and field sampling protocol in the context of conservative chloride transport. Streamline distances xSCREEN and unretarded travel times tSCREEN from the water table to the monitoring well screen were calibrated in this analysis, based on an axisymmetric model by Ostendorf et al. (2004). Briefly, the monitoring wells were of 2 in. diameter, PVC construction, with 5 ft screens set in uniform sand packs at various depths in a series of clusters. We analyzed the groundwater samples for major cations by ICP, generating sodium cNa and divalent (calcium and magnesium) c2 concentrations from June 1997 through December 2007. Fig. 1 displays the plan location of four monitoring clusters (E, H, J and T) used for field calibration; each cluster had a well in the brown till, a well near the top of the gray till, and a well near the bottom of the gray till. Table 2 lists the mid screen depths for each of the 12 monitoring wells, along with streamline distances and unretarded travel times from the water table: the average linear velocity adopted for the present analysis is simply the ratio of xSCREEN to tSCREEN. The deeper wells have generally slower v values due to longer path lengths in the impermeable gray till. 3. Theory 3.1. Gaines and Thomas (1953) geochemistry and assumed water table conditions We consider the advective transport of dissolved divalent cations and sodium through soil with exchange sites completely occupied by sorbed divalent and sodium cations c2 + cNa = C

ð1aÞ

q2 + qNa = Q

ð1bÞ

rffiffiffiffiffi c2 q2

ð1cÞ

K= Fig. 2. Cation exchange sites (circles), sorbed sodium (squares), and sorbed divalent cations (diamonds) observed in split spoon samples from monitoring well clusters C2 and D2 (2004): an average value calibrates Q = 0.057 meq/g, the average of the four lowest values calibrates qNaO = 0.0015 meq/g.

qNa cNa

with total dissolved cation concentration C. A selectivity coefficient characterizes equilibrium cation exchange (Gaines and Thomas, 1953). The equilibrium chemistry permits us to

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Table 2 Field calibrated water table parameter and selectivity coefficient values. Parameter (Units)

EC, ED, EB

JA, JD, JB

HC, HD, HB

TA, TC, TD

Midscreen depth (m) xSCREEN (m) a tSCREEN(yr) a v (m/yr) a Onset (year) CM (meq/L) tM (year) tB (year) b (meq/L-yr) K (L/meq)1/2 δ (%)

4.58, 7.32, 23.6 38.8, 42.5, 58.3 11.8, 16.5, 41.6 3.30, 2.57, 1.40 7/65 33 10/88 10/90 1.75 0.00396 19

5.78, 3.18, 22.2 37.1, 33.7, 53.0 13.9, 8.60, 40.8 2.66, 3.92, 1.30 2/64 29 4/84 9/90 2.63 0.00433 27

5.68, 2.52, 24.8 29.7, 27.1, 48.3 13.6, 8.94, 46.9 2.17, 3.03, 1.03 3/62 179 11/82 3/83 9.07 0.00423 29

26.4, 7.22, 3.14 25.4, 6.04, 9.10 41.6, 6.04, 11.2 0.61, 1.00, 0.81 4/65 366 4/88 7/88 19.5 0.00467 30

a

Calibrated by Ostendorf et al. (2006).

derive an explicit, nonlinear expression for the sodium exchange isotherm in terms of C and Q (Charbeneau, 1981)

cNa

2 3 sffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi  2 q2Na 4CK Q 4−1 + 1 + = −1 5 qNa qNa 2K 2 ðQ − qNa Þ

ð2Þ

The total exchange sites do not vary with time or location in the aquifer, while C changes due to a varying, but known water table concentration CS induced by highway deicing agent storage practices. In this regard, we postulate a linear increase of water table concentration from the onset of contamination at time t equals zero to a maximum level CM at time tM, due to the historical practice of outdoor salt handling and storage, followed by a linear decrease pursuant to implementation of remedial measures at time tB

Eqs. (1a), (1b) and (1c)–(4a) and (4b) anticipate our use of C and qNa as dependent variables: the three equilibrium Eqs. (1a), (1b) and (1c) reduce five unknowns to two when Q is constant. Two transport equations complete the analysis; chemical equilibrium alone does not specify individual cation partitioning in an aquifer. One cannot infer quantitative behavior of cations without knowing site hydraulics as well as boundary and initial conditions. 3.2. Dissolved cation transport model A steady average linear velocity advects the total dissolved cations in the assumed absence of dispersion, subject to known conditions at the water table and uniform initial (ambient) concentration

CS = CO + mt

ð0 b t b tM Þ

ð3aÞ

AC AC +v =0 At Ax

CS = CM

ðtM b t b tB Þ

ð3bÞ

C = CS

ðx = 0Þ

ð5bÞ

CS = CM − bðt − tB Þ

ðtB b t b tD Þ

ð3cÞ

C = CO

ðt = 0Þ

ð5cÞ

CS = CO

ðtD b t Þ

ð3dÞ

with downgradient distance x along a streamline from the water table to a monitoring wellscreen (Fig. 3). If we interpret the coefficient of the spatial derivative in Eq. (5a) as the speed of an observer, then the left hand side of the equation marks the rate of change of the dependent variable seen by the observer. The right hand side of Eq. (5a) implies that this is zero; thus, Eq. (5a) suggests that the total dissolved cation concentration is conserved in a frame of reference moving with the water molecules

with water table contamination and remediation rates m and b, and time tD to reestablish the ambient total dissolved cation concentration CO at the water table. We also assume a constant fraction of sodium in the total dissolved cation concentrations at the water table after the onset of contamination, reflecting the mix of salt and premix that generate the deicing agent infiltration cNaS = κCS

ð0bt Þ

sffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi# CS ðKκ Þ 4Q ð1 − κ Þ −1 + 1 + = 2ð1 − κ Þ CS ðKκ Þ2 2

qNaS

ð5aÞ

dC = 0

ð6aÞ

dx =v dt

ð6bÞ

ð4aÞ

"

ð4bÞ

with dissolved sodium concentration cNaS at the water table. Eq. (4b) constrains the sorbed sodium concentration qNaS at the water table, through the equilibrium chemistry of Eqs. (1a), (1b) and (1c).

The lower box in Fig. 3 refers to an observer moving at speed v along a streamline, a graphical representation of Eq. (6b). This observer retains the water table concentration CS in force at her departure time by virtue of Eq. (6a). The integral of Eq. (6b) sets her position and travel time along the streamline. This travel time coincides with the groundwater

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Fig. 3. Definition sketch for total dissolved cation and sorbed sodium transport models. Boxes refer to observers traveling along streamline between the water table and the monitoring well screen. Lower box travels at average linear velocity, and observer always sees the same total dissolved cation concentration—the water table concentration CS at her departure time. Upper box travels at a retarded velocity, and observer always sees the same sorbed sodium concentration—the water table qNaS at his departure time. He will meet a younger total dissolved cation observer at the wellscreen, and their water table concentrations will specify dissolved and sorbed partitioning at their arrival time.

travel time in the absence of reactions, and may be estimated by hydraulic models (Fetter, 2001) or isotopic dating methods (Solomon et al., 1992) as a consequence. Solution electroneutrality equates C to the conservative chloride concentration in the absence of other anions, so Eqs. (6a) and (6b) corresponds exactly to an anion advection model (Ostendorf et al., 2006). With C solved, we have reduced our problem to one unknown qNa and require a second transport equation to complete the model.

A comparison of Eqs. (5a) and (8a) indicates that the latter tracks sorbed sodium in the slower reference frame of an observer who travels along the same streamlines as his dissolved counterpart. The upper box in Fig. 3 refers to this observer of sorbed sodium, identified by his retarded speed v/RD. He always sees the same sorbed sodium concentration, by virtue of Eq. (8a) dqNa = 0

ð10aÞ

dx v = dt RD ðC; qNa Þ

ð10bÞ

3.3. Sorbed sodium transport model The sodium transport governed by a differential balance of storage, sorption, and advection (Charbeneau, 1981) AcNa ρ Aq Ac + B Na + v Na = 0 At n At Ax

ð7Þ

with bulk dry density ρB of the aquifer. Since cNa(qNa,C) by virtue of Eq. (2), we use the chain rule on Eq. (7), invoke Eq. (5a), and find AqNa v AqNa + =0 RD Ax At

RD = 1 +

ð8aÞ

ρ  B  AcNa n Aq

ð8bÞ

Na

with retardation factor RD requiring the differentiation of Eq. (2) rffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi  ffi3 2 Q 1 − 1 + 4CK −1 7 q q 6 Na Na 2Q − qNa 6 C 7   rffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffiffi =  ffi 4q + 5 2 Na Q 2K 2 qQ − 1 ðQ − qNa Þ 1 + 4CK − 1 Na q q

qNa = qNaO

ðt = 0Þ

ð10cÞ

qNa = qNaS

ðx = 0Þ

ð10dÞ

with sorbed sodium concentration at the water table specified by Eq. (4b). This latter is carried down the streamline towards the monitoring wellscreen by the sorbed sodium observer, at the value in force at his time of departure from the water table. Eq. (10b) yields his retarded travel time to the monitoring well, which is longer than his dissolved counterpart. The retardation factor in Eq. (10b) rests on the invariant sorbed sodium value in his frame, but reflects a varying dissolved concentration, as he is overtaken by total dissolved cation observers carrying different C values. The integration of Eq. (10b) is accordingly numerical, even for a constant average linear velocity.

2

AcNa AqNa

Na

Na

ð9Þ

3.4. Calculation and behavior of solution We use monitoring well observations of total dissolved cations to infer prior total dissolved cation concentrations at

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Fig. 4. Water table concentrations for monitoring well cluster J. Diamonds (JB), squares (JA), and circles (JD) represent measured total dissolved cations, with tSCREEN subtracted from the sample time. Curves represent calibrated CS and qNaS values used by transport models to predict concentrations—observers in Fig. 3 carry these concentrations down the streamline to the monitoring wellscreen.

the water table. For example, the symbols in Fig. 4 represent total dissolved cation concentrations at the water table backcalculated along streamlines from the J cluster wellscreens. Graphically speaking, we follow the total dissolved cation observer in Fig. 3 from her appearance at the wellscreen at the time of sampling backwards to her departure time from the water table. She maintains her observed C value, which becomes CS for her departure time. Since we assume a constant average linear velocity, this latter is obtained by subtracting her tSCREEN value in Table 2 from the sampling time. The deepest well (JB) tracks the oldest observers—who left the water table before it became contaminated (diamonds in Fig. 4). The shallowest well (JD) experiences remediation, and the youngest observers suggest that the reduction of water table salinity began in 1990, with corroboration from their sisters (squares) from monitoring well JA. These backtracked CS values calibrate the water table source parameters (onset of contamination, tM, CM, tB, and b)

of Table 2, represented by the solid line in Fig. 4. Eqs. (4a) and (4b) then completes the calculation of water table concentrations by imposing either ambient or road salt stoichiometry κ for a specified selectivity coefficient. This in turn sets the qNaS calculation in Fig. 4, which is assigned to the sorbed sodium observer (the upper box in Fig. 3) moving down the streamline. Eq. (10b) approximates his location xI at time tI by forward finite difference integration from the water table, using a time step Δt of 0.083 years xI = xI−1 +

vΔt RD ðCI − 1 ;qNaS Þ

  x CI − 1 = CS tI − 1 − I − 1 v

ð11aÞ

ð11bÞ

In this regard, Eqs. (8b) and (9) are used to compute a new retardation factor for each time step, reflecting the arrival of

Fig. 5. Predicted monitoring well concentrations for monitoring well JD. The total dissolved cation concentration is the same as shown in Fig. 4, shifted forward by 8.6 year unretarded travel time to the monitoring well. The first observer of salt contaminated water table groundwater arrived in July 1972, stimulating a decrease of cNa/C. This decrease persisted until arrival of the first observer of sorbed sodium at the water table in November 1998.

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Fig. 6. Laboratory isotherms for batch experiments conducted with uncontaminated split spoon soil samples at various depths from borehole D2. Gaines and Thomas (1953) selectivity coefficients range from 0.00226 to 0.00956 (L/meq)1/2, with an average value K = 0.0048 (L/meq)1/2 (Table 1).

the total dissolved cation concentration CI-1 of an unretarded observer (with her value from the water table) at his prior location on the streamline. The assumption of a constant v simplifies this last computation, as seen in the argument on the right hand side of Eq. (11b). The arrival of the sorbed

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sodium observer at the monitoring well, with the water table value of qNaS set by his departure time, combines with the current total dissolved concentration value in the well to distribute all the cations between the dissolved and sorbed partitions in accordance with Eqs. (1a), (1b) and (1c). Fig. 5 displays the total dissolved cation concentration and dissolved sodium fraction predicted to occur in monitoring well JD. The first follows directly and simply from Fig. 4, obtained by shifting the water table departure times forward by the 8.60 year unretarded travel time to the wellscreen. The pattern of road salt contamination and subsequent remediation is clearly evidenced by the C prediction—and one can assess historical impacts of salt storage practices on groundwater quality from total dissolved cation (or chloride) data bases (Ostendorf et al., 2006). Cation exchange alters the salt stoichiometry between the water table and a monitoring well however, and obscures the interpretation of cation distribution of groundwater samples. In this regard, Eqs. (11a) and (11b) and the calibrated Gaines and Thomas (1953) geochemistry of Eqs. (1a), (1b) and (1c) and Tables 1 and 2 complicate the predicted cNa/C ratio, and four regions can be seen in Fig. 4. The ratio in the monitoring well remained at its ambient value of 0.48 (set by CO and qNaO) until the arrival of the first contaminated total dissolved cation observer in July 1972, who experienced the onset of water table contamination 8.6 years earlier. The cNa/C ratio fell to its

Fig. 7. C (dashed lines and circles) and cNa/C (solid lines and squares) at monitoring well cluster H; symbols are data, lines are calibrations.

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minimum value of 0.24 in November 1998, when the first observer to experience sorbed sodium contamination at the water table arrived at the wellscreen. The sodium fraction rose thereafter to a new equilibrium value that reflected a redistribution of cations at the restored CO value in the aquifer. None of the regions resemble the road salt value of 0.975 for cNa/C in force at the water table. We compare the predicted cNa/C ratio with the observed cNa/C ratio to generate the error δ of the field calibration

KAP values cited by Appelo and Postma (2005). The latter are based on sorbed fractions and eq/L dissolved concentrations, so that 

KAP

 1=2 ð1000c2 Þ =  1 = 2 q2 ð1000cNa Þ Q qNa Q

KAP = 133K ðmeasured − predictedÞ δ= predicted

ðQ = 0:057meq = gÞ

ð13aÞ

ð13bÞ

ð12Þ

A nested Fibonacci search calibrates K for each cluster by minimizing the root mean square of δ.

Our observed average K corresponds to a KAP of 0.63 (L/g)1/2 by virtue of Eq. (13b). This value is at the upper end of the range 0.3 b KAP b 0.6 (L/g)1/2 cited by Appelo and Postma (2005).

4. Results and discussion

4.2. Results of field calibration

4.1. Results of laboratory calibration

Figs. 7–10 display the observed (symbols) and calibrated (lines) C and cNA/C values for the four monitoring well clusters. The latter reflect water table parameters and selectivity coefficients cited in Table 2. A single K and source parameter set for each cluster recovers various regions of the cNa/C curve displayed in Fig. 5, with a calibration error that ranges from 19 to 30%. Generally speaking, the deepest well in each of the four

Fig. 6 displays the Gaines and Thomas (1953) selectivity coefficients implied by the laboratory isotherm experiments. The K values range from 0.0026 to 0.0096 (L/meq)1/2, with an average value of 0.0048 (L/meq)1/2 (Table 2). The laboratory calibrated K may be compared with the selectivity coefficient

Fig. 8. C (dashed lines and circles) and cNa/C (solid lines and squares) at monitoring well cluster J; symbols are data, lines are calibrations.

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Fig. 9. C (dashed lines and circles) and cNa/C (solid lines and squares) at monitoring well cluster E; symbols are data, lines are calibrations.

clusters reflects ambient groundwater, the older of the horizontal regions. The CO, Q, and qNaO values of Table 1, which depend on split spoon sample analyses from borehole D2, represent these deep groundwater quality data (EB, HB, JB, TA) reasonably well, when cluster specific K values are applied. These latter range from 0.0040 to 0.0047 (L/meq)1/2 with an average field value of 0.0043 (L/meq)1/2 in good agreement with the average laboratory value of 0.0048 (L/meq)1/2. The field K values also calibrate the shallow data at each cluster, and these reflect arrival of the dissolved contamination and various stages of sorbed sodium appearance at the monitoring wells. Well EC and ED data (Fig. 9) exhibit the low cNa/C ratio; sorbed sodium frames have not yet arrived from the water table and the strong sodium solution has stripped divalent cations from the glacial till. Wells JA and JD, as shown in Fig. 8, document the arrival of sorbed sodium frames, with subsequent recovery and rise of dissolved sodium ratios. Clusters E and J are further from the water table than H and T, and the shallow wells at these latter clusters sample younger groundwater (more recently recharged to the water table) as a consequence. Thus, the data from HC, HD, TC, and TD exhibit higher cNa/C ratios. The calibrated water table source parameters in Table 2 are generally consistent with historical storage practices at the field site. Dissolved contamination arrived at the water table upgradient of the four clusters between 1962 and 1965, then grew to a maximum value (tM) by the 1980s. Dissolved

concentrations began to fall at the water table by the early 1990s (the tB values), consistent with cessation of outdoor storage at this time. Clusters nearer the center of the outdoor storage area exhibit the highest CM values, so that the parameter is at a high 366 meq/L value at the proximate T cluster. The calibrated maximum dissolved source concentration falls through a 179 meq/L value at cluster H, to 33 and 29 meq/L levels at the more distant E and J. 5. Discussion Fig. 11 summarizes the geochemical analysis of groundwater from monitoring well cluster H as a Piper diagram (Appelo and Postma, 2005). Major cation and anion meq/L fractions are displayed for deep (HB) and shallow (HD) wells. Ambient groundwater comprises the former data set, distinguished by high bicarbonate and moderately high calcium fractions, though the concentrations are low (small symbols). The shallow well features contaminated groundwater (larger symbols) with higher chloride and sodium fractions. The latter are not as large as the former however, due to the dissolution of sorbed calcium into the advancing deicing agent plume. The contamination appears as an upward and rightward shift of symbols of increased size on the cation triangle in the Piper diagram—attributes which would be qualitatively construed as evidence of seawater intrusion in a coastal setting (Appelo and Postma, 2005).

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Fig. 10. C (dashed lines and circles) and cNa/C (solid lines and squares) at monitoring well cluster T; symbols are data, lines are calibrations.

Fig. 7 suggests that the mass transport analysis adds quantitative spatial and temporal insight to this qualitative geochemical appreciation of deicing agent impact on groundwater. Indeed, two transport equations are needed to complete the specification of individual cation partitioning, and a network of monitoring wells with sufficiently long periods of record strengthens the assessment, particularly when accompanied by laboratory isotherms. No single monitoring well at our salt storage site exhibited all four regions of cNa/C behavior—Fig. 5 suggests that a 50 year period of record would have been needed. Ten years of sampling in a cluster of monitoring wells did identify the four regions however. Mass balances that rest on the resulting vertical profile of individual dissolved cations in a cluster must be analyzed cautiously, however, particularly if they are spatially averaged (e.g., DeSimone et al., 1997). The speciated dissolved cation profiles reflect retardation factors that not only vary with depth, but also with location along the streamlines from the wellscreens back to the water table (and forward into the aquifer). The impact of highway deicing agents on the cation quality on baseflow is complicated due to temporal as well as spatial variability. The rise of calcium and magnesium concentrations in surface water, attributed by Shanley (1994), Rhodes et al. (2001), and Rosjford et al. (2007) to road salt, would have been exhibited during the early region of the Fig. 5 prediction if our aquifer intersected a streambed or lake bottom near the JD wellscreen. The divalent dissolved cation fraction would

rise after the arrival of dissolved cation observers but prior to the arrival of sorbed sodium observers at the surface water site. This behavior reflects stripping of divalent cations into the groundwater. As time passes however, the sorbed sodium frames would have arrived at the streambed, and dissolved sodium concentrations would rise and eventually exceed their divalent counterparts. Fig. 5 suggests that the dissolved sodium fraction in monitoring well JD never attained the 0.975 value associated with road salt stoichiometry throughout the period of record—entrained divalent cations change the groundwater composition in variable fashion and complicate any attempt to impute road salt impacts on surface water quality if groundwater lies between the source and receiving stream. A geochemical groundwater transport model is required to quantify the fate of individual cations, with a calibrated water table source and known values for Q, K, and v. 6. Conclusions Chemical equilibrium and coupled transport equations specify the partitioning of dissolved and sorbed sodium and divalent cations in an aquifer subject to contamination from 20 years of outdoor storage of road salt at a highway deicing agent facility in eastern Massachusetts. Laboratory isotherms of split spoon samples from a glacial till drumlin under the facility calibrate a Gaines and Thomas (1953) selectivity

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Fig. 11. Piper diagram of observed groundwater quality data (meq/L fractions) at monitoring well cluster H. Ambient groundwater in the deep well (HB) is of relatively low concentration (smaller symbols), with high bicarbonate fraction and moderately high divalent cation concentration. The shallow well (HD) features contaminated groundwater (larger symbols), with high chloride and sodium fractions, although the latter is not as prevalent as the former due to cation exchange. “Explanation” relates symbol size to meq/L of total dissolved cations.

coefficient of 0.0048 (L/g)1/2. The advective transport model of Charbeneau (1981), when applied to the site hydraulics of Ostendorf et al. (2004, 2006) and 10 years of groundwater quality data, yield a confirming field calibrated K of 0.0043 (L/g)1/2. The excellent agreement of the field and laboratory K values, and reasonable accuracy of the calibrations, endorse the Gaines and Thomas (1953) and Charbeneau (1981) models and elucidate the geochemistry of deicing agent contamination: advancing deicing agents strip divalent cations from the soil into the groundwater. Total dissolved cation concentration is conservative, and monitoring well data reflects the response of the water table to the contamination and subsequent remedial measures. Individual cation concentrations observed in the monitoring well groundwater samples do not resemble road salt stoichiometry due to cation exchange—the dissolved sodium fraction falls to values as low as 20% in the wellscreen, far less than the 97.5% value of road salt. Cation exchange thus complicates any quantitative attempts to assess sodium impacts of highway deicing agent practices on surface or groundwater quality. Nomenclature b water table remediation rate (meq/L3-T) C total dissolved cation concentration (meq/L3) CI total dissolved concentration along streamline at Ith time step CM maximum total dissolved cation concentration at water table (meq/L3)

CO CS cNa cNaS c2 K KAP m n Q qNa qNaO qNaS q2 RD t tB tD tI tM tSCREEN

ambient total dissolved cation concentration (meq/L3) water table dissolved cation concentration (meq/L3) dissolved sodium concentration (meq/L3) dissolved sodium concentration at the water table (meq/L3) dissolved divalent cation concentration (meq/L3) selectivity coefficient of Gaines and Thomas (1953) (L3/meq)1/2 selectivity coefficient of Appelo and Postma (2005) (L3/M dry soil)1/2 water table contamination rate (meq/L3-T) aquifer porosity sorbed cation exchange sites (meq/M dry soil) sorbed sodium concentration (meq/M dry soil) ambient sorbed sodium concentration (meq/M dry soil) water table sorbed sodium concentration (meq/M dry soil) sorbed divalent cation concentration (meq/M dry soil) retardation factor time from onset of contamination at the water table (T) onset of decreasing water table contamination (T) time of reestablishment of ambient total dissolved cation concentration at water table (T) Ith time step along streamline (T) time of maximum total dissolved cation concentration at water table (T) unretarded travel time from water table to wellscreen along streamline (T)

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v x xSCREEN Δt δ κ ρB

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average linear velocity (L/T) distance along streamline from the water table (L) distance from water table to wellscreen along streamline (L) time step (T) error milliequivalent based sodium fraction of total dissolved cations at the water table bulk density of aquifer (M dry soil/L3)

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