Mass-balance analysis of reactive transport and cation exchange in a plume of wastewater-contaminated groundwater

ELSEVIER

Journal of Hydrology

203 ( 1997) 22X-249

Mass-balance analysis of reactive transport and cation exchange in a plume of wastewater-contaminated groundwater Leslie A. DeSimone”.*, “Water Resources Division, hCenter for Marine

US Geological

Brian L. Howesh, Paul M. Barlowa Survey, 28 Lord Road, Suite 280, Murlborough

Science and Technology,

Received 29 January

Universi~

of Massachusetts,

1997; revised 30 July 1997: accepted

New Bedford.

28 August

MA, VI 752. USA MA. 02744.

USA

1997

Abstract Mass-balance calculations were used to quantify reactive transport processes and cation exchange in a plume of groundwater contaminated with septage-effluent wastewater on Cape Cod, Massachusetts. Of the chloride mass recharged to the aquifer in effluent, as much as 72% was accounted for using spatial moment analysis and finite-element integration of groundwater concentrations, which were sampled at 569 wells and supplemented by borehole electromagnetic-induction logging. Comparison of chloride transport and mass balances with transport and mass balances of other species indicated that reactive processes substantially altered concentrations of all major chemical constituents. Calcium in effluent was exchanged for magnesium on aquifer sediments. Potassium also was attenuated, possibly through exchange with magnesium, sodium, and/or hydrogen ions. Sufficient hydrogen ions were generated by microbial nitrification in the unsaturated zone to consume effluent alkalinity and lower the effluent pH from 7.2 to 5.0 in the recharged groundwater; the resultant acid conditions may have facilitated anion adsorption and silicate-mineral dissolution. Retardation factors (R) calculated from breakthrough curves indicated that calcium (R = 1.4-2.2) and boron (R = I .3-2. I) were similarly retarded, whereas potassium experienced greater retardation (R = 1.8-5.2). Retardation of calcium, boron, and potassium was greater in the unsaturated zone than in the saturated zone; this may have resulted from spatial heterogeneity in exchange properties and preferential saturated-zone flow through coarse-grained sediments not present in the unsaturated zone. Although concentrations may stabilize and chemical reactions reach equilibrium at fixed points along paths in the plume, the mass-balance analysis illustrated that steady-state conditions will not be established throughout the aquifer and the cumulative mass of reacted constituents in the plume will increase until the plume reaches its discharge area. The analysis also indicates that retrospective study of dissolved concentrations in an established plume after many years of transport may not identify reactive transport and attenuation of plume constituents, if precise data on source concentrations (or masses) and the spatial distribution of solutes during plume development are not available. Finally, transport of the effluent-contaminated groundwater also altered the geochemistry of the aquifer, for example, through cation exchange, such that the introduction of clean, uncontaminated water into the aquifer will not immediately restore pre-plume conditions. 0 1997 Elsevier Science B.V. Keywords:

Solute transport; Cation exchange;

Groundwater;

Wastewater;

1. Introduction Most

studies

* Corresponding

of groundwater

contamination

are

author. e-mail: [email protected].

0022-1694/97/$17.00 0 1997 Elsevier Science B.V. All rights reserved PfI SOO22-1694(97)00101-7

Artificial recharge

retrospective. With little quantitative information available about the contamination source, physical and reactive transport processes often can only be inferred from the present-day distribution of contaminants in the aquifer. For this reason, small- or

L.A. DeSimone et al./Joumal

large-scale tracer tests that allow for precise source characterization and control are excellent methods to study physical and reactive solute transport in groundwater under natural conditions (Mackay et al., 1986; LeBlanc et al., 1991). However, the discharge history and chemical composition of contamination sources generally are more variable and complex than under tracer-test conditions. Therefore, quantitative studies of contaminant transport at the larger temporal and spatial scales of actual groundwater plumes also are valuable for confirming transport theories. However, to provide information comparable to experimental tracer tests, contaminant-transport studies of actual plumes require the same quantitative information about aquifer and source characteristics and detailed sampling. In the present study, contaminant transport was investigated during development of a wastewater plume that originated from surface discharge of effluent from a new septage-only treatment facility (Fig. 1). Because the facility was newly constructed, it was possible to measure effluent composition and volumes from the initiation of discharge through the 3 years of investigation. Aquifer and groundwater sampling were begun before the initial discharge and the evolving groundwater plume was monitored with an expanding well network. Thus, this study combined the quantitative aspects of a tracer test with the complexities of an actual contaminant source at large spatial and temporal scales. The quantitative source data supported a mass-balance analysis of reactive transport in the septage-effluent plume. Mass balance is an ideal tool for quantifying reactive processes under field conditions because it integrates spatial variability in solute concentrations and aquifer properties. Mass-balance analysis was applied to spatially variable groundwater concentrations measured five times during the study. Chloride, a major, conservative constituent of the septage effluent, was used as a tracer of effluent in groundwater. This paper reports on results of the mass-balance analysis of reactive tranport of major inorganic constituents of the effluent in the aquifer. Cation exchange, an important reaction altering major-ion chemistry during artificial wastewater recharge, was found to be a principal process in this system. Processes affecting nitrogen transformations and transport in the septage-effluent plume are described elsewhere (DeSimone et al., 1996; DeSimone and Howes, 1996).

of Hydrology 203 (I 997) 228-249

2. Description

229

of the study site

2.1. Septage-treatment

facility

The wastewater facility, in Orleans, Cape Cod MA (Fig. 1) treats septage, which is the semi-solid residue from on-site disposal systems, that is pumped from commercial and residential septic systems in surrounding towns. Independent treatment of septage has been recently implemented in the United States in unsewered, densely populated communities such as Cape Cod to avoid groundwater contamination from disposal of untreated septage and to comply with State regulations (US Environmental Protection Agency, 1984). Because of the high solids content of septage and the treatment processes (e.g. chemical conditioning) thereby required, the effluent discharged from septage-treatment facilities is generally more concentrated than secondary sewage effluent (Cantor and Knox, 1985). At the Orleans facility, septage is conditioned with FeC13 and Ca(OH)* (to promote coagulation) and is separated into solid and liquid components with a filter press. The solid component is disposed of off-site. The liquid component receives addition of HCl (for pH control) and phosphoric acid (a nutrient), biological treatment to reduce oxygen demand, and ultraviolet disinfection. The treated effluent is discharged one to two times per day to eight rapid infiltration beds, 10 x 10 m each, which are used on an irregularly rotating basis. Discharge of the effluent averaging 72 m3 per day began in February 1990 (Fig. 2a); effluent chemical characteristics are given in Table 1. 2.2. Aquifer hydrogeology The shallow aquifer at the site is composed of stratified glacial drift of Wisconsin age. Aquifer sediments generally occur in the following stratigraphic sequence: an upper fine-grained unit, 12- 15 m thick, of fine to very fine or fine to medium sand, with varying amounts of silt, mostly above the water table; and intermediate coarse-grained unit, 1S- 18 m thick, of medium to very coarse sand, locally with gravel; a lower fine-grained unit, lo- 18 m thick, of interbedded fine to very fine and medium to fine sand with silt; and a deep coarse-grained unit of medium to coarse sand of undetermined thickness. Sediments

L.A. DeSimone et ul./Journal of H,ydrology 203 (lYY7) 228-249

230

r

&

Namakakst

’ ,

./

’

tment-

.

d

4

I

41046’40” 0 -3.2

150 M

I . . . . . ..

\

WATER-TABLE CONTOUR-Shows altitude of the water table above sea level, December 1992. Contour interval 0.4 m. Dashed where inferred.

Fig. I. Observation-well network and local groundwater flow system in the study area, Orleans. Cape Cod, Massachusetts. Arrows show the general direction of groundwatzr flow: shaded square shows area of inset; solid circles show well-cluster sites, numbers show mass-balance sites, and letters indicate the line of section (A-A’).

in the intermediate coarse-grained unit, with hydraulic conductivities (K) determined from grain-size analyses of 40-60 m per day, were more than lo-fold more permeable than sediments in the upper or lower fine-grained units, with K of 2-4 m per day (DeSimone et al., 1996). However, the aquifer was

heterogeneous, and the three to four lithologic units were not continuous throughout the study area. The hydraulic conductivity of the codrse-grained sediments is similar to that of outwash sediments elsewhere on Cape Cod (Masterson and Barlow, 1994: Millham and Howes, 1995). Porosity averaged 0.34

L.A. DeSimone et al./Joumal Table 1 Chemical constituents

(mg I-‘) and properties Effluent

G? PH (-a’+ Mg” Nat K’ Alkd .%I;~ clSiOq TDSe B (lg I-‘) PONf DON NH;-N NO;-N NOT-N PO;‘-P P, dissolved P. total POC’ DGC

3330 N.D.’ 7.2 440 0.3 100 30 134 19 9.50 1.1 1730 200 3.1 4.0 27 0.76 9.4 0.02 0.04 0.44a 9 14

231

of effluent and groundwatera Groundwater Pre-discharge

SPCh

ofHydrology 203 (1997) 228-249

122 6.8 6.3 3.4 2.2 15 1.2 7.3 13 21 13 77 200 N.D. 0.05 0.02 N.D. 0.07 0.02 0.02 N.D. N.D. 0.4

Water table

3010 N.D. 5.0 330 44 100 16 1.5 1.1 870 17 1420 140 N.D. 1.0 1.0 0.04 35 0.02 0.02 N.D. N.D. 1.8

Plume Center

Leading edge

3020 0.05 5.4 350 28 110 22 < 1.2 2.2 860 14 1430 200 N.D. 0.5 6.4 0.08 32 0.02 0.02 N.D. N.D. 1.9

1500 3.4 5.6 88 59 52 4.5 12 0.9 390 12 631 30 N.D. 0.06 0.02 0.02 I2 0.03 0.02 N.D. N.D. 0.5

“Median values: effluent, March 1990-December 1992 (n = 45); water-table groundwater, January 1991-December 1992 (n = 35): predischarge groundwater, September 1989 (n = 29); plume-center groundwater, December 1992. where Cl- > 800 mg I-’ Oz = I I ): plumeleading-edge groundwater, where Cl- = 200-500 mg I-’ (n = 1 I). “Specific conductance, ys cm-’ at 25°C. ‘N.D., not determined. dAlkalinity, mg I-’ as HCO;. ‘Total dissolved solids, sum of constituents. ‘Particulate organic nitrogen (PGN) and carbon (PGC), estimated from dissolved organic nitrogen (DON) and carbon (DGC) using the empirically derived ratios (N = 15) of total to dissolved organic nitrogen (1.8)and carbon (I .7). ‘Median value, March-December 1992 (n = 15).

(n = 230) in sediment-core samples, but this value was affected by compaction (Wolf et al., 1991; DeSimone et al., 1996). Therefore, a value of 0.39 was used for field porosity in the present study, based on in situ studies of outwash sediments near Falmouth, Cape Cod (Garabedian et al., 1991). Groundwater flow is westward and northwestward across the study area from inland recharge areas of Cape Cod toward regional discharge areas in the adjacent coastal marsh and Cape CodBay (Fig. 1). Beneath the infiltration beds, hydraulic gradients were downward, whereas flow was nearly horizontal downgradient from the beds.

3. Methods Aquifer sediment samples were collected with a wireline-piston corer (Zapico et al., 1987) or splitspoon sampler. Effluent was sampled at 2- to 4week intervals from February 1990 to December 1992 either directly from the discharge or from an 8-h composite sample. Groundwater samples were collected from PVC wells (5 cm, i.d. 0.3 and 1.5 m screen length) with a stainless steel submersible pump (Keck Instruments) or Teflon bailer after removal of at least three well volumes. One to six wells, screened at

232

1990

1991

1992

Fig. 2. (a) Mean daily discharge of effluent. and (b) chloride concentrations in effluent (solid circles) and groundwater at the water table beneath the infiltration beds (open circles, site 7, well screen 3.1-6.2

discrete depths from the water table (1-19 m below land surface) to as much as 43 m below land surface, were sampled at each site. Water samples for analysis of chemical constituents were immediately filtered and preserved (DeSimone et al., 1996). Concentrations of chemical constituents in effluent and groundwater were determined by the US Geological Survey National Water-Quality Laboratory (Arvada, CO) using standard techniques (Wershaw et al., 1987; Fishman and Friedman, 1989). Specific conductance and pH were measured during sample collection with temperature-compensated electrodes (Hach 44600 and Beckman Phi 11 with Orion electrode, respectively). Dissolved oxygen concentrations 2 1 mg 1-l were measured with a temperatureand pressurecompensated electrode, placed in the well between the screen and the submersible pump to ensure flow past the membrane; concentrations < 1 mg I-’ were measured calorimetrically using sealed ampules (Chemetrics) on water collecEed with the submersible pump. Cation-exchange capacity (CEC) and exchangeable cation concentrations in sediments were measured

m above sea level). Location of groundwater site shown in Fig.

I.

by the Soils Testing Laboratory, University of Massachusetts, Waltham MA, using an acidified sodium-acetate extraction method (Shoemaker et al., 1961). The combined electrical conductivity of the aquifer sediments and groundwater was measured at 20-cm intervals through PVC wells with a downhole probe (Geonics EM39 electromagnetic-induction logger). The electromagnetic-induction (EM) logger was used without prior wellbore evacuation because the logger is relatively insensitive to borehole fluids (McNeill, 1986).

4. Results 4.1. Ambient geochemistry

sediment and groundwater

Aquifer sediments consist primarily of quartz and feldspar (62 t 14% and 18 +- 5%, respectively, of sand- and silt-sized fractions, mean relative weight percent -+ 1 SD, n = 19), with plagioclase more abundant

233

L.A. DeSimone et al./Journal of H.vdrology 203 (I 997) 228-249 Table 2 Grain size, cation-exchange Sediment type”

C-VC C-M M-F F-M F-VF

sand sand sand sand, silt sand. silt,

n

4 6 I 2 8

capacity

(CEC) and exchangeable

d tn (mm)b

1.00 + 0.44 + 0.20 0.22 t 0.1 1 t

SP (phi units)’

0.12 0.11 0.01 0.02d

1.55 1.25 0.6 2.01 1.1 I

2 0.26 5 0.21 + 0.08 2 0.53d

cation concentrations Silt and clay

(mean and standard deviation)

Cation exchange

characteristics

of aquifer sediments

(milliequivalents

per 100 g)

(SE)

2.0 3.9 6.5 20.5 27.4

-c 0.6 ? 2.7 z 2.0 2 I 1.7d

CEC

Cal+

0.5 0.5 0.9 I.4 2.0

0.3 0.3 0.7 0.7 0.9

i- 0.3 t 0.1 i- 0.1 2 0.8

Mg” + 0.2 It 0.04 2 0.1 i- 0.4

0.2 0.2 0.2 0.8 0.8

K’ rt: 0.2 t- 0.04 2 0.0 rf: 0.4

0.04 0.03 0.05 0.07 0.08

2 0.02 + 0.01 z 0.01 -+ 0.02

clay “VC. very coarse; C, coarse; M, medium: F, fine; VF, very fine. hdsO,median grain size, grain diameter such that 50% of the grains by weight is titter. ‘s,,, graphic standard deviation, in phi units [(-In (diameter in millimeters)/ln(2)] from Folk (1974) % = 5.

than potassium feldspar (mean plagioclase:potassium feldspar ratio = 1S:l). Accessory minerals include amphibole (probably hornblende), siderite, and glauconite. Calcite occurred in trace amounts (~1%) in about one half of the samples analyzed. The claysized fraction consists primarily of illite/mica (49 + 8%), kaolinite (26 t 7%), mixed-layer illite-smectite (12 -f 6%), and chlorite (I 1 + 6%) > CEC ranged from 0.5 + 0.3 milliequivalents per 100 g for very coarse to coarse sand with gravel (n = 8) to 2.1 ?I 0.8 milliequivalents per 100 g for fine to very fine sand with silt and clay (n = 7). Calcium and magnesium made up about 90% of the exchangeable cations in uncontaminated sedients and were present in subequal concentrations. Sorbed potassium also was present, but in concentrations about an order of magnitude lower than calcium or magnesium (Table 2). The extraction method (Shoemaker et al., 1961) could not measure exchangeable sodium; however, recoverable sodium (using HCl digestion, Fishman and Friedman, 1989) in five samples was a small fraction (O. l-2.5%) of the total recoverable cations, as expected from the low ionic strength of uncontaminated groundwater (McBride, 1994). Ambient (pre-discharge) groundwater in the study area, sampled prior to effluent discharge (September 1989, n = 29), was slightly acidic and low in total dissolved solids (Table 1). Sodium, chloride, and sulfate were the dominant ions and likely resulted from sea spray in atmospheric wetfall and dryfall and/or road-salt application. Dissolved oxygen concentrations ranged from near saturation (10-l 1 mg 1-l) in

groundwater from shallow wells and wells screened in coarse-grained material to nearly zero (O-l mg I-‘) in groundwater from several deep wells (>23 m below the water table). The low dissolved oxygen concentrations probably result from a long groundwater residence time (LeBlanc, 1984) because reduced material that would rapidly deplete dissolved oxygen (e.g. organic matter or pyrite) generally was not found in study-area sediments. Low dissolved oxygen concentrations (CO. l-3.3 mg 1-l) were coincident with significant amounts of dissolved reduced iron (1200-9000 pg 1-l) and manganese (64270 pg l-‘, deep wells at sites 9, 10, 12, 13, and 20, Fig. 1). These concentrations likely resulted from reductive dissolution of iron and manganese oxide or hydroxide weathering-product minerals (e.g. amorphous iron hydroxide) and sand-grain coatings; reddish-orange staining was common in coarsegrained sediments that were immediately above or below fine-grained sediments. 4.2. Septage-efluent

plume in groundwater

After 34 months of discharge, the septage-effluent plume in groundwater (defined by elevated chloride concentrations) extended about 60- 140 m westward and northwestward from the downgradient (northwestern) boundary of the infiltration beds and about 40-50 m upgradient (east and south) of the beds, where ‘upgradient’ and ‘downgradient’ are defined relative to the regional flow directions (Fig. 3a). Its irregular shape in plan view resulted from the

234

LA.

DeSimorw

et d.Nournul

of Hwhdo~y

203

(I YY7J

22%2dY

L.A. DeSimone et al./Joumal

235

of Hydrology 203 (1997) 228-249

600,

AA

A

A PA A

+

Y

A

100

2 3

CHANGE IN SPECIFIC CONDUCTANCE tiS Fig. 4. Aquifer electrical conductivity measured conductance of groundwater from adjacent wells.

cm” at 25 OC)

by the electromagnetic-induction

irregular geometry of the fine- and coarse-grained sediments near the infiltration beds. Mean linear velocity of groundwater in the plume, determined from chloride breakthrough curves, was 0.13-0.14 m per day, corresponding to an apparent integrated hydraulic conductivity of about 10 m per day, (DeSimone et al., 1996). In addition to chloride, major chemical constituents of the septage-effluent plume were calcium, magnesium, sodium, potassium (Fig. 3b-f), bicarbonate (measured as alkalinity), silica and nitrate (Table 1). The large differences between ionic concentrations in effluent (and thus in effluent-contaminated groundwater) and concentrations in ambient groundwater (Table 1) made it possible to use aquifer electrical conductivity as a tracer of plume movement. Aquifer electrical conductivity depends primarily on dissolved ionic concentrations and to a lesser extent on porosity and sediment characteristics (McNeill, 1980). It was correlated with dissolved ion concentrations, measured as specific conductance of groundwater (R* = 0.87, where electrical conductivity and specific conductance are changes from background values, Fig. 4). Similarly, aquifer electrical conductivity was correlated with chloride (as change from background, R* = 0.86, not

logger

(difference

from background)

and specific

shown), the dominant dissolved ion in effluent; chloride was used as a tracer of plume movement for the mass-balance analysis. The vertical distribution of aquifer electrical conductivity indicated that the septage-effluent plume moved primarily through the coarse-grained sediments (Fig. 5). Time-series electrical conductivity logs also showed that the plume moved downward in the aquifer in response to head gradients, lithologic heterogeneity, and possibly density differences (-0.001 g cm-‘, DeSimone et al., 1996) between effluent-contaminated groundwater and ambient groundwater. At some locations, the lower boundary of the plume coincided with the lithologic boundary between coarse- and fine-grained units (Fig. 5); at other locations, the lower boundary of the plume was not defined. The plume ranged from 4 m thick at site 7 to about 20 m thick at site 11. 4.3. Mass-balance 4.3.1. Recharged traveltime

calculations masses and unsaturated-zone

The mass of each major constituent of the effluent discharged from the treatment facility and recharged

236

WEST

A

-

GENERAL DIRECTION OF GROUND-WATER

METERS

t-

251

INFILTRATION

.,VERTICAL EXAGGERATION X 2.5

&g

FLOW

BEDS

site 4

0

EXPLANATION

I

1

I

I

I

r

SO m

I

ELECTRICAL CONMJCTIVITY LOG, DECEMBER 19924cale

VERTICAL EXTENT OF THE PLUME--Based on specific conductance of groundwater and electrical conductivity logs, December 1992

is 0 to 600 fi cm-’ at 25 OC =!l

Fig. 5. Aquifer lithology and vertical distribution of specific conductance and electrical conductivity logs along section A-A’, December 1992. Location of section shown in Fig. I. See text for definition of hydrogeologic units. Vertical lines show depth of well casing; (m), screened interval of well.

to the aquifer was computed from measured volumes of discharge and solute concentrations (sampled at 2to 4-week intervals) after smoothing out daily variation in concentration with a 2-month moving average. The moving average was assumed to yield a representative solute concentration for effluent discharge during the days surrounding the sampling date. Daily discharge volumes were multiplied by the appropriate average solute concentations to obtain the daily mass loading of each effluent constituent. The cumulative mass of solute recharged to the groundwater at any sampling date was calculated as the cumulative discharged mass (sum of daily mass inputs from the initiation of discharge to the sampling date) minus the solute mass in transit in the unsaturated zone; the latter mass was estimated to equal the mass discharged in effluent during the 3 months prior to the groundwater sampling date (see below). In this approach, unsaturated-zone losses of chloride, the conservative tracer of effluent, were assumed to be

negligible; net losses of other constituents in the unsaturated zone are evaluated separately by comparing concentrations in effluent with concentrations in groundwater at water table beneath the infiltration beds. Macrodispersion effects, potentially altering the temporal distribution of discharged chloride masses during transport through the unsaturated zone, also are assumed to be negligible. Macrodispersion effects would have been most significant for those constituents, such as calcium and chloride, for which daily discharged masses varied non-randomly with time. However, errors associated with these effects are likely to have been small relative to other errors associated with the mass-balance calculations (see below). The time required for effluent to travel from the land surface to the water table (12 m) was estimated by correlating chloride concentrations in effluent with chloride concentrations in groundwater beneath the infiltration beds. Chloride concentrations in effluent changed with time as a result of variation in the facility

237

L.A. DeSimone et al.Nournal of Hydrology 203 (1997) 228-249

0

.

I

.

,

.

,

.

,

.

,

.

. WATER TABLE . ..

15

:

CONCENTRATlON x DEPTH INTERVAL

DEPTH INTERVAL

CHLORK’IE ML-1)

(m*mgL-l)

(m)

.-..,

;\ 20

2” 840

r 25

x

6.4

=

5,380

i

5,180

. . . . ..-..-......-...-..--..-..--..-......-...-...... 810

30

x

6.4

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . ..-..--.-. 35

._ 0

200 ELECTRICAL

400

600

CONDUCTMTY

&S cm-l at 25 Oc)

VERTICALLY INTEGRATED CONCENTRATlON FOR THE SITE:

12,700m*mgLS1

Fig. 6. Vertical integration of concentration data for mass-balance calculations (site 11, December 1992). Dashed lines define zones of equal concentration of septage-effluent-plume constituents, as inferred from the electrical conductivity log. Concentrations from well screened at 42.1-42.4 m below land surface were not included because the electrical conductivity log indicates that the well is screened below the plume at this site.

operation and treatment procedures. This variation in input concentrations also created a temporal signal in chloride concentrations in the groundwater at the water table at site 7 (Fig. 2b). For a conservative species such as chloride, the offset in time between changes in concentrations in effluent and changes in concentrations in groundwater provides an estimate of the traveltime from land surface to the water table at this site. The temporal offset that most closely related chloride concentrations in groundwater to concentrations in effluent was 3 months. This estimate represents the best correlation (R* = 0.63) between groundwater concentrations (y,) and separate data sets of temporally offset effluent concentrations (x~), (x,_,), (x~-~), .‘., (x,_,), where t is the order of the sample in time and n is the maximum number of offsets tested (1, 2, 3, 4, 5 and 6 months); concentrations were transformed into equally spaced, monthly data with linear interpolation and correlated with an autoregressive model (SAS Institute Inc., 1988). Ground-

water samples collected prior to January 1991 were not used in the correlations, because they showed a steady increase in concentration with time representing transient conditions that resulted from the initial arrival of effluent at the well (Fig. 2b). A 3-month estimate for the unsaturated-zone traveltime agrees well with time-series measurements of specific conductance in groundwater at site 7 (not shown), which first increased above background values about 3 months (May 1990) after the start of effluent discharge (February 1990). The estimated traveltime may be a slight overestimate, because site 7 is at the margin of the infiltration beds. For example, travel times calculated from the mean daily effluent loading rate, the area of the eight infiltration bends, the depth to the water table, and a specific yield of 0.15-0.30 (Morris and Johnson, 1967) were only 22-43 days. However, this volumetric loading approach assumes only steady, vertical flow. Traveltimes based on steady-state flow may considerably overestimate

LA.

238

DeSimone

et d/Journal

of Hydrology

203 (1997)

7O’JOO’20”

70000’30”

/

I

22X--249

I

60

O-

ME:TERS

I EXPLANATION TREATMENT-FACILTTY BUILDING INFILTRATION

IDENTIFIER

BEDS

Fig. 7. Triangular

OF TRIANGULAR

WELL-CLUSTER site idenifier elements

flow velocities that result from the transient wetting and drying cycles, especially if vertical heterogeneity is present (Russo et al., 1989). Therefore, the 3month estimate of the unsaturated-zone travel time empirically derived from concentrations at site 7 was used for mass-balance calculations in the present study. The paired comparison of chloride concentrations in groundwater at the water table at site 7 with concentrations in effluent (monthly data, temporally offset by 3 months) also indicated that the concentrations in groundwater were significantly lower than concen-

used for mass-balance

SITE-Number

ELEMENT is

calculations

trations in effluent @ = 0.0019, Wilcoxon signed-rank test). The median difference between paired values, I20 mg I-’ (14% of effluent concentrations, 95% contidence interval 6- 19%) likely represents dilution of effluent by atmospheric precipitation (3-4% of effluent recharge in 199 1 and 1992, DeSimone et al., 1996) and by mixing of effluent with ambient groundwater that was recharged to the aquifer upgradient of the infiltration beds. Dilution at site 7 appeared to decrease with time (p = 0.035, Wilcoxon signed-rank test on paired differences vs time) as the septage-effluent plume expanded in area in all directions.

of Hydrology 203 (1997) 228-249

L.A. DeSimone et al./Journal Table 3 Chloride mass balance for the septage-effluent

Number of sites Total Contaminated Number of wells Total Contaminated Sampling density” Total Contaminated Effluent-input mass (MT? Solute mass in groundwater plume (MTh) Solute mass/input mass (5%) “[no. samples]/[plume h metric ton.

plume at five groundwater

sampling

dates

March 1991

June 1991

September

14

I3

7

7

33 I1 I.17 0.39 24 I2 50

volume (cubic dekameters

239

199 1

June 1992

December

19 IO

22 14

24 16

33 12

49 IX

65 29

69 34

I.19 0.43 28 13 46

0.97 0.36 31 I9 61

I .04 0.46 47 34 72

0.90 0.44 66 47 71

1992

or 1000 m’)].

4.3.2. Solute masses in groundwater Solute masses in the septage-effluent plume were calculated from synoptic samplings of concentrations in groundwater using spatial moment analysis and finite-element integration. Concentrations were first vertically integrated at each well-cluster site and then horizontally integrated with a triangular finiteelement analysis in a manner similar to that of Garabedian et al. (1991). For the vertical integration, the concentration measured at each well was multiplied by a depth interval within which concentrations were assumed uniform; these intervals were inferred from the electrical-conductivity log of the plume at the site. A vertically integrated concentration (c;) for each site was obtained by assuming the products of concentration multiplied by depth interval from all wells at the site, such that the entire thickness of the plume at the site was assigned a concentration. This procedure is illustrated in Fig. 6. Linear interpolation was used to determine the solute mass in each of 22-35 triangular elements that were defined by the well-cluster sites at the vertices (Fig. 7). The solute mass in each triangular element is equal to (Garabedian et al., 1991)

M,=n$c, +c*+c3) where: Mj = mass of solute for triangular element i; n = porosity, assumed to equal a uniform value of 0.39; A = area of the triangular element i; c I, cz, c3 = vertically integrated concentration at well sites 1, 2

and 3. A can be calculated

using

A=~((x24’1-~3~2)-(~,~.?-~34’,)+(~,y?-~?1’1))

(2)

where (x,,y ,), (~2.~2) and (.~a,ys) are the coordinates of the three well sites comprising the vertices of each triangular element. The solute mass in the entire plume was calculated as the sum of the masses in each triangular element. Integrations for synoptic samplings in March, June and September 1991 were based on sites l-22 and elements l-30; integrations for samplings in June and December 1992 were based on sites 1-25 and elements 1-35 (Fig. 7). Additional sites and elements were included in the 1992 integrations because the plume expanded and new wells were installed. Finally, the solute masses calculated by the triangular integration were corrected for ambient groundwater incorporated into the plume through dispersion. Because of dispersion, the septage-effluent plume (as defined by spatial integration of high electricalconductivity zones) contained more water than was recharged to the aquifer as effluent through the sampling time. The difference between the total plume water volume, calculated with the triangular integration, and the cumulative volume of effluent recharge was assumed to equal the volume of ambient groundwater incorporated into the plume. Small solute masses corresponding to the background concentrations were associated with this ambient water volume. These masses were estimated by multiplying median background concentrations (Table 1) by the

240

9

750

p

500

z 8

250

(a)

8 YJ

160

9

120

6 i= ii

80

8 8

40

’

’

”

I

I

I

”

”

”

I,

I,

I

”

”

I,

I

’

”

I

I

0 I

I

I

I

Na+

1990

1991

1992

Fig. 8. Concentrations of (a) calcium (Ca”), (b) magnesium (Mg’+), (c) sodium (Na’), (d) potassium (K’), and (e) boron (B), in effluent (solid circles) and water-table groundwater beneath the infiltration beds (site 7, well screen 3.1-6.2 m above sea level, open circles). Location of groundwater site shown in Fig. I.

ambient-water volumes and were subtracted from the total solute masses in the plume. Ambient-water masses, as percentages of the total mass estimates, were 0.8-3% for chloride, 0.4-1.4% for calcium, 0.2-2.2% for magnesium, 4.7-17% for sodium and 2.4- 14% for potassium. Thus, imprecise estimates of ambient-water volumes or concentrations were unlikely to introduce large errors into mass estimates for chloride, calcium, or magnesium, although they may have been a significant source of error for sodium or potassium. This method was used rather than uniformly subtracting background concentrations from

all concentrations in groundwater to avoid applying the correction where mixing of ambient and effluentcontaminated groundwater was minimal (for example, in the plume center). 4.3.3. Chloride mass balance The estimated solute mass of chloride, the conservative tracer of plume movement, ranged from 46 to 72% of the effluent-input masses, with the highest percentages obtained in the later samplings (Table 3). The increased recovery of the solute mass in groundwater as a percentage of effluent-input mass

241

1990

1991

1992

Fig. 8. Continued.

resulted for improvements in the well network, but not simply from an increase in sampling-point density. Overall sampling-point density, measured either by the number of sites or samples, total or contaminated, per unit volume of the plume, was relatively constant with time (Table 3). The percentage of effluent-input mass accounted for in groundwater apparently increased with time because high-concentration areas of the plume were better defined in later synoptic samplings. Better definition was obtained through installation and sampling of new wells (e.g. site 2, July I99 1, Fig. 7) and expansion of the high-concentration areas due to groundwater flow. This is suggested by the distribution of mass among individual triangular elements. In the early sampling rounds. large fractions (27-35%) of the total recovered mass resided within two elements (elements 5 and 6). In later sampling rounds, the total mass was more uniformly distributed. Several factors probably contributed to the < 100% recovery of the effluent-input mass of chloride. Although the vertical extent of the plume was well

defined by electrical conductivity logs, the vertically integrated concentration at a site could be underestimated if wells were not screened in the peakconcentration zones. At some sites. the plume’s lower boundary was not clearly defined, and chloride mass may have been unaccounted for at the bottom of the plume. In addition, the horizontal definition of sampling points was constrained in some areas by obstructions to drilling such as buildings and underground structures, pipes, and utilities. For example, the high-concentration area between sites 13, 14 and 16 was beneath the main building, and probably was underestimated (Fig. 7). Also, linear interpolation may have underestimated the plume thickness and thus the chloride mass between some sites (for example, between site 2, 17 m thick, and site 7, 4 m thick, Fig. 5). Finally, the chloride mass may have been underestimated because the plume extended beyond the well network (Freyberg, 1986), such as southward past site 3 and eastward past site 4 (Fig. 7); thus, some volume of the plume was outside of the integrated area. This effect is demonstrated by the inclusion of

L.A. DeSimone

242

et al./Journal

site 22, which was assigned a vertically integrated concentration of zero based on the specific conductance of groundwater from two wells at the site. Inclusion of site 22 added 5.8 and 6.6% to the chloridemass estimates for June and December 1992. The addition of hypothetical sites with zero vertically integrated concentrations surrounding the well network would have further improved the solute-mass estimates (Garabedian et al., 1991) but the samplingpoint density of the present study did not justify such a procedure. 4.4. Reactive

solute transport and cation exchange

Chemical alteration during unsaturated-zone transport was evident in concentrations of nearly all the measured effluent constituents (Table 1). For calcium (median decrease 28% of effluent concentrations, 95% confidence interval 17-37%), potassium (median decrease 45%, 95% confidence interval 39-55%), and boron (median decrease 32%, 95% confidence interval 25-42%), significant decreases were found that exceeded dilution at the water table beneath the infiltration beds (site 7); these estimates of alteration were made by comparing solute concentrations in groundwater from the water table at site 7 with effluent concentrations, after correcting for the 3-month unsaturated-zone traveltime as described previously for chloride. Decreases in total dissolved nitrogen greater than dilution were not observed (median decrease 19%, 95% confidence interval 3.6-39%).

of Hwirolog~

203

(I 997)

22% 249

However, the distribution of nitrogen species changed substantially; reduced forms, which were dominant in effluent, were transformed to oxidized species. mostly nitrate, which composed about 90% of nitrogen reaching the water table (Table 1). Alkalinity, sulfate, and phosphorus in the effluent were almost entirely depleted after transport through the unsaturated zone. pH decreased by about 2 standard units to 5.0 in water-table groundwater. Magnesium and silica - 1 mg I-’ in effluent, increased over concentrations, one and two orders of magnitude, respectively. Changes in sodium concentrations were less than decreases in chloride, but could not be distinguished from dilution (median decrease 9%, 95% confidence interval I- 16%). Time-series measurements indicated that most unsaturated-zone changes were transient. After 34 concentrations of calcium. months of discharge, potassium, and boron at the water table beneath the infiltration beds nearly equaled their initial concentrations in the effluent (Fig. 8a. d and e). Magnesium concentrations at the water table beneath the infiltration beds peaked at 120 mg 1-l in December 1990 and subsequently decreased to about 30 mg 1-l by the end of the study (Fig. 8b). Transport of the peak magnesium concentrations through the aquifer resulted in a zone of high magnesium concentration (> 100 mg I ‘) encircling the center of the groundwater plume (Fig. 3~). High sulfate concentrations at several sites in the plume center (sites 7 and 8, 4 and 11 mg I-‘, respectively) indicated that unsaturated-zone attenuation of

Table 4 Retardation factors at selected sites in the septage-effluent plume Site

Screened interval

d (m)”

Retardation Factor

(m above sea level)

Unsaturated-zone 7

CaL’

Na’

K-

R

2.22

N.D.h

5.16

2.08

I .57 I .4h I.38 I .42 I .46

0.84

I .4x I .36

0.93

I .I2 I .X0 I .76’

I .08

N.D.

0.90

I .76

I .63 I .48

transport +3.1 to 6.2

0

Saturated- and unsaturated-zone transport 6

+I .7 to +3.2

46

8

-0.94

25

11

-1.8

II

-14.6

to -I 2 to -8.4 to -14.9

Mean “Distance of the three-dimensional ‘N.D.,

56 58

flow path from the water table beneath the infiltration beds to the well.

not determined.

‘Retardation

0.73

factor determined by extrapolating potassium concentrations over time.

1.34

L.A. DeSimone et al./Joumal of Hydrology 203 (1997) 228-249

243

I”““““’

(a)

0.8

0.6

U

UJJASQNQJFUAMJJASQNQ 1992

1991 Fig. 9. Breakthrough curves of chloride (Cl-), Fig. I), (a) 8.1-8.4

m and (b) 14.6-14.9

calcium (Ca”),

sodium (Nat).

potassium (K’),

and boron (B) in groundwater. (site

I I. see

m below sea level.

sulfate also was diminishing by December 1992. In contrast, unsaturated-zone changes in alkalinity and silica showed no time trends. Retardation of individual constituents in the unsaturated and saturated zones of the aquifer was estimated by comparing groundwater concentrations to effluent-input concentrations (breakthrough curves, e.g. Fig. 9). Retardation factors (Table 4) were expressed as the time required for the concentration of a constituent to reach one-half the median effluent concentration divided by the corresponding time for chloride. Retardation factors from breakthrough curves at the water table beneath the infiltration

beds (not shown) were 2.2 for calcium, 5.2 for potasGum, and 2.1 for boron. These values represent unsaturated-zone attenuation only and are consistently higher than the total attenuation in unsaturated and saturated zones represented by retardation factors determined at downgradient wells. Retardation factors from four downgradient wells were similar, averaging 0.9 for sodium, 1.5 for calcium and boron, and 1.8 for potassium (Table 4). Attenuation during unsaturated and combined unsaturated and saturated tmasport indicate three levels of retardation-minimal (sodium), intermediate (calcium and boron), and high (potassium).

Table 5 Mass balances for calcium. magnesium, sodium. potassium, and boron for the septage-effluent plume at hve groundwater sampling dates Constituent Effluent-input Ca”

March 1991

June 1991

September 1991

June 1992

December

1992

mass (MT“) I?

16

IS

22

30

Mg?’

0.15

0.16

0.26

0.17

O.IX

Na’

I .7

2

2.3

1.7

6.6

K+

0.5

0.6

0.7

1I.3

B

0.0055

0.006

0.0065

0.01

I

I .4 0.0I5

Solute mass in the ground-water plume (MT”) Ca’+

3.2

3.1

5.8

M$

I .7

2.7

3.9

Na’

I .6 I .4

II

I5

I .4

2. I

4.2

5.9

K-

0.1 I

0.12

0.21

0.45

0.75

B

0.001 I

0.001 I

0.002

0.005

0.0076

4.5

Solute mass/input mass (%;) Ca’+

27

22

39

Mg”

1100

1100

1700

SO 2300

SO 2500

Na’

82

70

91

88

86

K+

22

20

30

3.5

39

B

20

I8

31

45

51

Solute mass/input mass, normalized to chloride recovery (%) Ca’+

54

48

64

69

70

2200

2400

2800

3200

3500

Nat

160

I50

IS0

120

120

K+

44

43

49

49

55

B

40

40

51

63

71

Mg”’

* Metric ton.

Effluent-input masses and solute masses in the septage-effluent plume of calcium, magnesium, potassium, and boron were determined from the effluent sampling and the five synoptic groundwater surveys (Table 5). The solute masses were compared to the chloride mass balance to quantify reactive transport of each constituent on a mass basis. Solute masses were normalized to the chloride balance by dividing the estimated solute mass, as percentage of input mass, by the estimated percentage of chloride recovered in the same sampling. In this procedure, the factors limiting chloride recovery are assumed to similarly affect recovery of all plume constituents. However, normalizing to chloride may introduce some error if a constituent were distributed differently from chloride, especially where the plume was not well defined. This error would have been negligible for calcium, sodium, and potassium, which were well correlated with chloride (R* = 0.91-0.95 for calcium; R* = 0.87-0.95 for sodium; and R* = 0.71-0.76 for potassium) but may have been more significant for boron

(R2 = 0.35-0.84), and magnesium (R* = 0.35-0.88). Solute masses of calcium (22-50%) and boron (18-51%) in the plume were similar fractions of effluent-input masses in each sampling round, as expected from the calcium and boron retardation factors (Table 4). Masses of potassium in the plume were smaller fractions (20-39%) of input masses, consistent with greater potassium retardation. Magnesium masses were more than lOO-fold greater than input masses. When normalized to the chloride mass balance and expressed in kiloequivalents (keq), decreases in calcium (140-340 keq) were 84-l 14% of increases in magnesium mass (122-360 keq), suggesting that calcium-magnesium exchange was likely responsible for these changes (Fig. 10). Sodium masses in the plume, when normalized to the chloride mass balance, were slightly larger than input masses, which agrees with the observed sodium retardation factor of < 1 (Table 4). Mass decreases in potassium (3.6-16 keq) represented large fractions of potassium inputs, but were small relative to changes in other

L.A. DeSimone et al./Journal of Hydrology

203 (1997)

GAIN

+346CAT +85 W r

+239CAT +I45

CAT

+29H+

1 -145

CAT

+158

CAT

+ 37.5 -

_,;n r

”

+53.4

245

1;1+ u

m

400 CAT +92

ii+

Ii+

h

CAT “&I\

-186 CAT -30 ALK

I -301 CAT -64 ALK

LOSS t

MARCH 1991

228-249

JUNE 1991

SEPTEMBER 1991

JUNE 1992

DECEMBER 1992

EXPLANATION CALCIUM MAGNESIUM SODIUM POTASSIUM

0

AMMONIUM

’

ii?kt$i?;:

?? ALKALINITY

FRoM

Fig. 10. Mass losses and gains from solution in the plume of calcium, magnesium, sodium, potassium, ammonium. hydrogen ions from nitrification, and alkalinity. Numbers are the sum of cations (calcium, magnesium, sodium, potassium, ammonium) lossed or gained (CAT); hydrogen ions from nitrification (H’); and alkalinity lost (ALK), in thousands of equivalents (keq). Calcium, magnesium, sodium, potassium. and alkalinity masses are equal to the difference between effluent-input masses and solute masses in the septage-effluent plume, normalized to the chloride mass balance. Ammonium masses (sorbed ammonium) and hydrogen-ion masses are from nitrification from DeSimone et al. (1996)

ions. Potassium ions may have displaced sorbed sodium ions, but decreases in potassium were only 16-39% of increases in sodium (21-42 keq), suggesting that other factors were involved (Fig. 10); potassium also may have exchanged with magnesium on sediments. Ammonium sorption also represented a significant loss (11 and 27 keq) of dissolved cations from the saturated zone in June and December 1992. Sorbed ammonium masses in the saturated zone were separately determined from dissolved ammonium concentrations, sediment extractions, and laboratory batch tests (DeSimone et al., 1996). Ammonium sorption in early sampling rounds was relatively small (0.7-3.1 keq), because it occurred primarily in the anoxic central volume of the plume, which was small prior to June 1992. Oxidation of ammonium to nitrate through microbial nitrification, although significant in the unsaturated zone, was unlikely to have

had a major role in the saturated zone, where ammonium concentrations generally were much lower and oxygen in the plume center was depleted (DeSimone et al., 1996). Hydrogen ions produced in the unsaturated zone by nitrification of effluent ammonium (two hydrogen ions per ammonium ion oxidized) could have exchanged with sorbed cations (Fig. 10). Because little of the nitrate produced was assimilated or denitrified by soil microbes (DeSimone et al., 1996), nitrification resulted in net acid production. Nitrification of 23 mg N I-’ ammonium (median difference between paired ammonium concentrations in effluent and water-table groundwater, January 1991 to December 1992) would thus produce 3.3 meq I-’ of Effluent alkalinity, 2.2 meq 1-l hydrogen ion. (134 mg 1-l as HCO;), was insufficient to consume all of this quantity of hydrogen ion; moreover, the

246

LA. DeSimone CI ul.NourrmlofHydrology 20.1(lYY7) 22X-24Y

measured pH decrease of 2 standard units corresponds to an increase of only 0.01 meq H’ 1-l. Little carbonate-buffering capacity was available in the aquifer, as evidenced by the pH of uncontaminated groundwater and the negligible calcite content of sediments. Sorption of phosphate, produced by mineralization of effluent organic phosphorus, may have consumed a small amount (-0.05 meq 1-l or less) of the hydrogen ion produced by nitrification (Walter et al., 1996). Hydrogen ions also may have been consumed by base cation exchange, but the large influx of base cations, calcium, sodium, and potassium, may have limited the effectiveness of this mechanism.

5. Discussion Infiltration of septage effluent into a sandy glacial aquifer resulted in a wastewater plume in groundwater that was highly reactive with aquifer sediments. Although the CEC of aquifer sediments was low, as is typical of sandy sediments (Dance and Reardon, 1983; Ceazan et al., 1989) cation exchange appeared to have a large effect on solute transport and the chemical composition of the plume. Calcium-magnesium exchange was evident from time-series measurements in effluent and water-table groundwater beneath the infiltration beds, from retardation of calcium relative to conservative transport of chloride, and from mass balances of calcium and magnesium. As the calcium-rich, magnesium-poor effluent percolated through the unsaturated-zone and aquifer sediments, calcium ions displaced sorbed magnesium ions, which were present in sediments before effluent discharge at about equimolar concentrations with calcium (Table 2). The exchange process appeared most active as effluent initially contacted unsaturated-zone sediments, when contrast between dissolved and sorbed calcium-to-magnesium ratios was greatest. Retardation factors from downgradient wells and sodium mass balances suggested a small net increase in sodium ion during transport of effluentcontaminated groundwater. Sodium may have been gained from sediments through exchange with potassium, hydrogen, and(or) ammonium ion, but exchangeable sodium represented only a small fraction (<2.5%) of the pre-discharge exchangeable cations. The measured increase in sodium also may have

resulted from dissolution of plagioclase feldspar (e.g. Lee, 1991). Dissolution of feldspar and other aluminosilicate minerals would consume hydrogen ions (McBride, 1994). and along with phosphate sorption and base cation exchange may account for the hydrogen ions produced by nitrification that were not accounted for by changes in effluent alkalinity. Dissolution of aluminosilicate minerals also might explain the measured increase in silica, although desorption of monosilicic acid is an alternative explanation (Wood and Signor, 1975). Attenuation of boron and sulfate suggest that anion adsorption could have occurred. At low pH, variablecharge minerals such as iron oxides and hydroxides can develop anion adsorption capacity (McBride, 1994). lron oxides and/or hydroxides were present in the aquifer sediments, as indicated by visible reddish-orange staining on mineral surfaces and high dissolved iron concentrations in oxygendepleted, uncontaminated groundwater. Sulfate retention through adsorption to iron and aluminum oxides is a widely recognized effect of acidification in some soils (Reuss and Johnson, 1986; Mulder and Cresser. 1994). Increased sulfate adsorption has been associated with increased nitrification in a forest soil following clear cuts (Nodvin et al.. 1986) and with artiticial wastewater recharge by sorption to ironhydroxide coated sediments at low pH (Wood, 1978). The measured pH of water-table groundwater was below the isoelectric points of common iron oxide and hydroxide minerals (Drever, 1988) and may have facilitated sulfate adsorption. However. bacterial uptake is an alternative explanation for the sulfate attenuation. Sulfate reduction is not likely to have caused sulfate attenuation because of the high concentrations of nitrate and oxygen in the unsaturated zone, which typically are preferentially reduced before sulfate. Boron (as borate) can adsorb to clay minerals, iron oxides and hydi,oxides, and organic matter, although the reaction is most active in alkaline conditoins (Keren and Bingham, 1985; McBride. 1994). Both attenuation and conservative transport of boron have been reported during recharge and groundwater transport of secondary-level wastewater (Bouwer et al.. 1980; Idelovitch and Michail, 1984; LeBlanc. 1984). In the present study, boron was clearly retarded, despite the acidic conditions. Retardation of boron

L.A. DeSimone et al./Journal of Hydrology 203 (1997) 228-249

and calcium were readily detected in the present study because of the detailed source data and observations of the initial transport of effluent-contaminated groundwater through aquifer sediments. Finally, phosphorus also was removed in the unsaturated zone in the present study. Phosphorus is more reactive with soil minerals than sulfate through precipitation and ion-specific adsorption (Hem, 1985; McBride, 1994). Phosphate is adsorbed by iron and manganese oxy-hydroxides, which are insoluble in oxic conditions and were present in aquifer sediments at the study site. If the unsaturated zone became anoxic such that iron minerals were dissolved, the phosphorus removed from effluent by phosphate sorption could become mobile; however, this is unlikely under the current conditions of effluent loading. Phosphorus removal during unsaturated-zone transport also was found to occur in septic-system effluent (Robertson et al., 1991; Weiskel and Howes, 1992). Reactions involving calcium, magnesium, potassium, and boron during unsaturated-zone transport were approaching steady-state after 34 months of effluent discharge, or 11 pore volumes of the unsaturated-zone sediments. The transient nature of ion exchange and other reactions during artificial recharge has been reported in several studies (Wood and Signor. 1975; Idelovitch and Michail, 1984; van Beek and van Puffelen, 1987). As an increasing number of pore volumes percolate through the sediments, sorbed concentrations gradually equilibrate with input dissolved concentrations. Thus, the chemical composition at any fixed point along the transport path, such as at the water table beneath the infiltration beds, gradually approaches that of the infiltrating recharge water. However, because of physical transport, a plume of waste water-contaminated groundwater continually advances through sediments with which it is in disequilibrium until it reaches a discharge area. After the plume has reached the discharge area, a steady-state, or equilibrium, distribution of sorbed and dissolved concentrations will be approached at all points along the flow paths as long as input concentrations remain constant. Because the wastewater plume studied here had not reached a discharge area during the study period, the cumulative solute mass removed from groundwater in the plume by sorption increased with time as the plume advanced through uncontaminated aquifer sediments

247

(Fig. 10). After the plume reaches its discharge area and equilibrium conditions are established throughout, the distribution of reactive and conservative constituents in groundwater should no longer reflect the effects of reactive transport and attenuation, in the same way that concentrations at the water table beneath the infiltration beds resembled effluent concentrations after 34 months of discharge. Thus, retrospective identification of the reactive transport and attenuation would not be possible based on dissolved concentrations alone. In the present study, attenuation of calcium is not readily apparent in the general distributions of dissolved calcium and chloride in the aquifer after less than 3 years of transport (Fig. 3a and b), although calcium clearly was retarded relative to chloride. Although mass losses and gains increased in absolute magnitude with time, comparison of solute masses with the chloride mass balance also indicated that the solute masses in the plume represented increasingly larger percentages of the total input mass as the plume evolved. After normalizing to the chloride mass balance, the mass accounted for in the plume increased from 54 to 70% of the input mass for calcium, from 44 to 55% for potassium, and from 40 to 7 1% for boron. Sodium, which exceeded the input mass after correction to chloride, decreased from 160 to 120% of the input mass. There are two possible explanations for these trends. In contrast to the pulsed injection commonly used in tracer tests, the total plume mass from a continuous contamination source continually increases. Because only a fraction of each incremental input mass is removed by reactive processes, the mass of the reacted pool increases more slowly than does the total recharged mass. In the present study, temporal trends in exchange reactions apparently also resulted from the heterogeneity of aquifer sediments. Effluent discharged at land surface subsequently percolated through sediments of the upper fine-grained unit. These sediments, consisting of fine to medium sand with silt, extend from land surface to just below the water table. Fine-grained sediments in the study unit had higher CECs (by a factor of 3-4) and higher concentrations of exchangeable cations than the medium to coarse sand and gravel (Table 2), such as make up the intermediate coarse-grained unit and through which effluentcontaminated groundwater preferentially flowed in

248

[..A. DeSimone

rt ul.Nournal

the saturated zone. Thus, mass losses or gains due to exchange reactions per unit transport distance were greater in the unsaturated zone than in the saturated zone. The mass-balance estimates of losses or gains integrated the cumulative effects of all the retardation experienced in the unsaturated and saturated zones through the time of the estimate. As equilibrium is reached with unsaturated-zone sediments and the plume expands with time, the unsaturated-zone residence times and reactions became less significant fractions of the total losses and gains from sorption. Finally, as reaction with aquifer sediments altered the chemistry of the effluent and of groundwater in the effluent plume, transport of the effluent-contaminated groundwater altered the geochemistry of the aquifer. For example, the pool of sorbed magnesium, which occupies nearly half of cation exchange sites in uncontaminated sediments, was stripped from aquifer sediments in the unsaturated zone and should be expected to be similarly removed from sediments in the saturated-zone path of the plume. Dissolved iron and manganese concentratons were not measured in the septage-effluent plume, but it is likely that iron oxy-hydroxide coatings will dissolve in the anoxic center of the plume, reducing the capacity of these sediments to remove phosphate or other anions. Thus, should the discharge of effluent cease, the introduction of clean, uncontaminated water into the aquifer will not immediately restore pre-plume geochemical conditions.

of‘Hydrology203 (1097) 22%249

work was funded by the USGS, the Massachusetts of Environmental Protection, and Department National Science Foundation grant no. X7 17701. WHOI Contribution no. 897 1.

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We thank D. Pelton, D. White, D. Schlezinger, N. Millham, S. Brown-Ledger, and D. Geohringer. Woods Hole Oceanographic Institution (WHOI); W. MacDonald and J. Burgess, Orleans, Brewster and Eastham Groundwater Protection Districts; and the Town of Orleans. R.W. Lee and G.E. Granato, US Geological Survey (USGS), provided technical reviews of the manuscript. Mineralogical analyses were provided by L. Poppe and A. Moffett, USGS, Woods Hole, MA. Cation exchange analyses were provided by S. Bodine, Soil Testing Service, University of Massachusetts, Waltham, MA. Use of firm or trade names is for identification purposes only and does not constitute endorsement by the USGS. This

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