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Journal of Environmental Radioactivity 101 (2010) 33–44

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Journal of Environmental Radioactivity

journal homepage: www.elsevier .com/locate/ jenvrad

Environmental fate of Ra in cation-exchange regeneration brine waste disposedto septic tanks, New Jersey Coastal Plain, USA: migration to the water table

Zoltan Szabo a,*, Eric Jacobsen a, Thomas F. Kraemer b, Bahman Parsa c

a U.S. Geological Survey, 810 Bear Tavern Rd., Site 206, W. Trenton, NJ 08628, USAb U.S. Geological Survey, 12201 Sunrise Valley Rd., Reston, VA 20192, USAc N.J. Department of Health and Senior Services, CN-308, Trenton, NJ 08625, USA

a r t i c l e i n f o

Article history:Received 20 August 2008Received in revised form30 July 2009Accepted 16 August 2009Available online 11 November 2009

Keywords:226Ra228RaAlpha-particle radioactivityGround waterCation exchangeAciditySeptic systemNitrificationChlorideNew Jersey coastal plain

* Corresponding author. Tel.: þ1 609 771 3929; faxE-mail address: zszabo@usgs.gov (Z. Szabo).

0265-931X/$ – see front matter Published by Elsevierdoi:10.1016/j.jenvrad.2009.08.007

a b s t r a c t

Fate of radium (Ra) in liquid regeneration brine wastes from water softeners disposed to septic tanks inthe New Jersey Coastal Plain was studied. Before treatment, combined Ra (226Ra plus 228Ra) concen-trations (maximum, 1.54 Bq L�1) exceeded the 0.185 Bq L�1 Maximum Contaminant Level in 4 of10 studied domestic-well waters (median pH, 4.90). At the water table downgradient from leachfields,combined Ra concentrations were low (commonly �0.019 Bq L�1) when pH was >5.3, indicatingsequestration; when pH was �5.3 (acidic), concentrations were elevated (maximum, 0.985 Bq L�1 –greater than concentrations in corresponding discharged septic-tank effluents (maximum, 0.243 BqL�1)), indicating Ra mobilization from leachfield sediments. Confidence in quantification of Ra massbalance was reduced by study design limitations, including synoptic sampling of effluents and groundwaters, and large uncertainties associated with analytical methods. The trend of Ra mobilization in acidicenvironments does match observations from regional water-quality assessments.

Published by Elsevier Ltd.

1. Introduction

The U.S. Environmental Protection Agency (USEPA) finalizeda Maximum Contaminant Level (MCL) of 0.555 Bq L�1 (15 pCi L�1)for gross alpha-particle activity and 0.185 Bq L�1 (5 pCi L�1) forcombined Ra (sum of 226Ra and 228Ra) for drinking water (USEPA,2000). Radium is a known human carcinogen that poses healthrisks when ingested (Mays et al., 1985; USEPA, 1999). Concentra-tions of Ra isotopes in water samples from the aquifers of the NewJersey Coastal Plain have been identified as frequently exceedingthe combined Ra MCL. In a series of studies conducted by the U.S.Geological Survey (USGS) in cooperation with the New JerseyDepartment of Environmental Protection (NJDEP), concentrationsof Ra isotopes in water samples from the two major unconfinedaquifer systems in the New Jersey Coastal Plain, the Kirkwood-Cohansey in southeastern New Jersey and the Potomac-Raritan-Magothy immediately adjoining the Fall Line (see Fig. 1 for general

: þ1 609 771 3915.

Ltd.

location), were determined (Focazio et al., 2001; Szabo et al., 2005).Combined Ra activities were greater than the MCL of 0.185 Bq L�1 inabout 31 percent of samples.

Water softeners (cation-exchange units) are widely availableand are convenient whole-house Ra treatment units for domestic-well water, because the units efficiently remove Ra (about90% removals; Lucas, 1987; Szabo et al., 2008) along with constit-uents that cause water hardness, such as calcium (Ca) andmagnesium (Mg), as well as iron (Fe). The findings regardingwidespread elevated Ra concentrations in ground water used assources of drinking waters in the Atlantic Coastal Plain (Boltonet al., 2000), including New Jersey (Kozinski et al., 1995; Szabo et al.,2005), as well as occasionally elsewhere in the United States(Focazio et al., 2001), have led to the increased use of water soft-eners for purposes of Ra removal, and not just water softening(NJDEP, 2004). Proper maintenance includes regular regenerationof the capacity of the cation-exchange media with sodium (Na) orpotassium (K) chloride (Cl) brine solution (Bowie, 1995). The waste(regeneration) brines bearing large concentrations of Ra(maximum, 81.2 Bq L�1; Szabo et al., 2008) commonly are flushedto septic systems. It is assumed that there they are diluted and

Z. Szabo et al. / Journal of Environmental Radioactivity 101 (2010) 33–4434

dispersed to the environment. The environmental fate of thisRa-bearing brine waste after disposal to septic tanks has not beenstudied, warranting this investigation.

The authors hypothesized that there is the likelihood ofcontinued Ra mobility, or an increase in Ra mobilization, in water–effluent mixtures in the leachfield and at the water table for theunconfined aquifers receiving the cation-exchange regenerationbrine waste through discharge of effluent from the septic systems.Kraemer and Reid (1984) and Miller and Sutcliffe (1985) showedthat Ra mobilization was occurring in aquifers in the Gulf Coastalregion of the USA subject to high or strongly increasing salinity,respectively. A proposed mechanism for Ra mobilization withincreased salinity is the increase in competition for exchange siteswith the associated divalent cations (Nathwani and Phillips, 1979),though Szabo et al. (1997) emphasized that increasing amounts ofhydrogen ion can also be a source of increasing competition. Boltonet al. (2000) noted that in the aquifers of the Maryland CoastalPlain, concentrations of Ra were typically highest where salinitywas highest and pH was lowest.

Fig. 1. Sampling site locations for Ra in ground water, and in septic-tank effluent, Kirkwoo2003–04.

Rapid development of densely populated residential areasreliant on septic systems for waste disposal, especially in areas withprivate wells that produce water that is treated with softeners, hasbeen shown to be associated with increased salinity (Cl concen-tration) and nitrogen (N) and phosphorus (P) concentrations in thewaters of the underlying sandy aquifers. These changes in waterquality have been studied in the Atlantic Coastal Plain in New Jersey(Bunnell et al., 1999; Barringer et al., 2006), in Massachusetts (Yates,1985; DeSimmone and Howes, 1998), and in the glaciated upperMidwestern United States and Canada (Wilhelm et al., 1996;Robertson et al., 1998; Thomas, 2000). This study characterizes Raconcentration and fate in liquid waste media (septic effluent)associated with the cation-exchange treatment of drinking waterfor Ra removal in the (rarely studied) private well setting in theCoastal Plain of New Jersey (Fig. 1) after episodic disposal of thewater-softener regeneration brine wastes into septic systems anddischarge of the effluent to the environment. The brines weredescribed by Szabo et al. (2008). Sampling at the water tabledowngradient from the leachfields described in this article was

d-Cohansey and Potomac-Raritan-Magothy aquifer systems, New Jersey Coastal Plain,

Z. Szabo et al. / Journal of Environmental Radioactivity 101 (2010) 33–44 35

designed to determine the Ra concentrations there, and to assess ifthe Ra remained mobile or was sequestered in the vadose-zone soiland sediment after septic-tank effluent discharged to the leach-field. An assumption of the study was that synoptic sampling ofseptic-tank effluent and downgradient ground water was suffi-ciently representative to allow for general characterization andapproximate mass-balance calculation.

2. Study area, hydrology, and previous assessments

2.1. Sampling location and hydrology

Residential sites with softened domestic-well water wereconsidered, where cation-exchange treatment units were used andepisodic (weekly) regeneration was practiced with brine wastedisposal to septic tanks. Sites without cation-exchange treatmentsystems were not included as part of the sampling effort, becausethe objective of this narrowly focused study was to characterize thefate of Ra in water-treatment waste disposed to the environment.Sampling was conducted at 10 sites in the Coastal Plain of southernNew Jersey (Fig. 1) and consisted of water or effluent collected fromthe private (domestic) drinking-water wells, septic tanks, anddowngradient drive-point monitoring wells (Fig. 2). The sampleddrinking-water wells were in the two major Coastal Plain aquifersystems of southern New Jersey. The unconsolidated iron-rich,occasionally feldspathic, Cretaceous-age delta sand deposits formthe Potomac-Raritan-Magothy aquifer system along the Fall Line(unconformity) that establishes the northwestern boundary of theCoastal Plain (Fig. 1). Farther to the south, sampled wells were inthe areally extensive quartzose Miocene and Tertiary age seawarddipping beach and river sand deposits that form the Kirkwood-Cohansey aquifer system (Fig. 1).

Pumping fromwell

Septicsystem

NOT TO SCALE

EXPLANATION

Water table

Ground-water flow

1. Initial radium source from aquifer: well water for drinking, into the home --> 2. Human ingestion exposure endpoint: treated drinking water after flow through treatment unit--> 3. Radium-brine waste: regeneration brine waste discharge from treatment unit--> 4. Septic system: 4a. Septic system effluent (liquid), and/or 4b. Septic sludge (solid)--> 5. Environmental dispersion: Shallow ground water (at the water table) from shallow observation well downgradient of shallow soil in leachfield receiving briny effluent waste.

2

1

3

4ab

5

Treatmentsystem waste

Water tablesampling point

Media 1, 4a, and 5detailed this article

Fig. 2. Sampling scheme for Ra and other constituents in ground water used fordrinking water collected from domestic wells, in septic-tank effluent after receivingregeneration brine waste from water treatment (softening) in the home, and in waterat the water table after dispersal of the septic-tank effluents to the environment.

Detailed descriptions of the geology, hydrology, and chemistryof these major unconfined Coastal Plain aquifer systems studiedwere provided by Barksdale et al. (1958), Rhodehamel (1973),Yuretich et al. (1981), Barton et al. (1987), Zapecza (1989), Kozinskiet al. (1995), Szabo et al. (1996), Kauffman et al. (2001), and Caullerand Carleton (2006). Waters are dilute and acidic. The median pHsin these quartzose aquifer systems is in the range of 4.7–5.2 (Bartonet al., 1987; Kozinski et al., 1995; Szabo et al., 1997, 2005). The‘‘pooled’’ median for these studies is 4.85, and 4.9 is used herein asthe typical median for general descriptive purposes for the aquifersas a whole. Hardness is typically low (median, 17 mg L�1; inter-quartile range, 8–42 mg L�1; Kozinski et al., 1995). The backgroundCl concentration in water in the Kirkwood-Cohansey aquifer systemin forested (undeveloped) areas is about 5 mg L�1 (Yuretich et al.,1981), though Szabo et al. (1997) determined a median Clconcentration at the water table of 13 mg L�1 in a samplingprogram focused in rural (mixed agricultural/residential) areas.Barton et al. (1987) reported a median Cl concentration for waterfrom the Potomac-Raritan-Magothy aquifer system from undevel-oped (background) areas of 14 mg L�1. The predominant sources ofCl to the ground water are surficial recharge that bears Cl from seaspray in precipitation, and from non-point-source leachates fromroad salt and agricultural fertilizers (Yuretich et al., 1981; Bartonet al., 1987; Szabo et al., 1997; Kauffman et al., 2001). Concentra-tions of Fe can be elevated, up to about 5 mg L�1 in sandy zones, buthigher where the aquifer is clayey (Kozinski et al., 1995; Szabo et al.,1997). Before the elevated ground-water Ra concentrations weredocumented, the water softener (cation-exchange) treatmentsystems were commonly used to remove the Fe.

Typical soil is sand to loamy sand with porosity of 35–45 percentand moisture content of 5–25 percent (Baehr et al., 2003). Rainfall isabout 1.05 m yr�1, with abundant but variable recharge (typicallyfrom about 0.29 to 0.46 m yr�1 for the Kirkwood-Cohansey aquifersystem; Szabo et al., 1996; Baehr et al., 2003), though recharge isless where soil is clayey or where extensive compaction fromresidential development has taken place. Horizontal hydraulicconductivities are typically 50–70 m day�1 (Rhodehamel, 1973;Cauller and Carleton, 2006), though the full extent of the range maybe much larger; ground-water flow velocity is typically about 0.1 mday�1. Distribution of transient tracers in the Kirkwood-Cohanseyaquifer system indicates dispersion is minimal (Szabo et al., 1996).Travel time from the septic leachfield to the water table ispresumed to be equal to or less than 3–6 months on the basis of thesoil properties, the high recharge rates, the high vertical conduc-tivities, the distribution of transient tracers at the water table(Szabo et al., 1996), and also on the basis of the documentedappearance of the large concentration of nitrate (>5 mg L�1) thatwas detected at the water table 3–6 months after land applicationof about a 1 cm thick blanket of nitrogen-rich treated sludge to soilstudy plots overlying the aquifer system (Jacobsen, 2000).

The sands comprising the Potomac-Raritan-Magothy aquifersystem are slightly coarser grained than those that comprise theKirkwood-Cohansey aquifer system, and historical aquifer test dataindicate slightly higher typical horizontal conductivities (50–80 mday�1) (Barksdale et al., 1958). Soils are coarse and well drained,and recharge rates are assumed by Barton et al. (1987) to be aboutthe same as for the Kirkwood-Cohansey aquifer system, on thebasis of the historical data available at that time.

2.2. Previous assessments in the study area

The median concentration of combined Ra in the ground waterfrom the aquifer systems in previous investigations was about0.165 Bq L�1 (Kozinski et al., 1995; Szabo et al., 2005). Concentra-tions of Ra are substantially greater in those samples where pH is

Z. Szabo et al. / Journal of Environmental Radioactivity 101 (2010) 33–4436

less than 5.2 than in those samples where pH is greater than5.2 (Kozinski et al., 1995). In a detailed study, Szabo et al. (1997)collected water samples from a variety of depth intervals from theaquifers, including at the water table, and determined that themedian concentration of Ra at the water table for both 226Ra and228Ra was 0.055 Bq L�1 (1.5 pCi L�1), and was 0.110 Bq L�1 forcombined Ra. The inter-quartile range for combined Ra concen-trations at the water table was small (0.060–0.115 Bq L�1). Thecombined Ra concentrations increased with depth in the aquifer,and were greatest at medium depths (15–25 m; maximumconcentration, 1.12 Bq L�1 was at 16.2 m) used most frequently forwithdrawal by the domestic wells in the area (Szabo et al., 1997).

An initial phase of study of treatment of acidic water bearing Radetermined the fate of Ra within the water softening treatmentsystem and septic tank (Szabo et al., 2008). Concentrations ofcombined Ra were determined in the (raw) ground water, treateddrinking water, regeneration brine waste, septic-tank effluent, andsettled sludge in the septic tank. The water softeners (cation-exchange treatment systems), when maintained, effectivelyremoved Ra from water, typically resulting in about a 10-foldreduction (median concentrations for 226Ra and 228Ra in treatedwater were <0.0037 and 0.011 Bq L�1, respectively) relative to theuntreated ground water. The concentration of Ra in the backwashbrine waste typically was enriched about 30-fold relative to theconcentration in untreated ground water (maximum combined Raconcentration, 81.2 Bq L�1 (2200 pCi L�1)). The median concen-tration ratios of combined Ra in regeneration brine wastes relativeto the septic-tank effluents was 35, and estimated mass ratios near1 indicated return of much of the Ra in the liquid brine to the septictank (Szabo et al., 2008). Radium in the liquid effluent is lost to thesludge layer by the settling of Ra-sorbing particles (Szabo et al.,2008), but the Ra-accumulation rate in the sludge, while notable,was small enough (about 5% of the mass) that the most substantialfraction of the Ra introduced to the septic tank in regularly recur-ring pulses of the regeneration brine waste migrated to the leach-field. The consistent episodic (weekly) discharge of the Ra-bearingregeneration brine waste to the septic tank and the substantial Ramass transfer to the leachfield allows, in this second phase of study,for reasonable approximation of relative mass balance for Ra in thedischarging liquid waste as it percolated through the leachfieldsediments to the water table over a period of weeks to months.

3. Sampling and analytical methods

3.1. Drive-point emplacement and sampling procedures

A series of temporary drive points at the 10 Coastal Plain siteswere emplaced to the water table 3–10 m downgradient from theseptic leachfield in about a 30-degree arc in the likely direction ofground-water flow using direct push technology (Geoprobe)without use of drilling fluids. The final drive point for samplingpurposes was installed immediately adjacent to the drive pointwhere the highest specific conductance was determined indicatingproximity to the center of the effluent ‘‘plume’’. The median depthto the water table was 6.1 m (range, 3.7–14.3 m). These water tabledepths are representative for the region (Szabo et al., 1996; Jacob-sen, 2000; Baehr et al., 2003; Cauller and Carleton, 2006). Screensemplaced for sampling were 1.25 cm inner diameter 10-slot size,high-density polyethylene, 30 or 60 cm in length, and were pre-cleaned in overnight soak in 5% HCl.

Purging, for a period of about 60 min, was used to develop thedrive points sufficiently to lower the water turbidity to levels on theorder of 5–10 NTU or less, the critical threshold for the collection ofrepresentative trace-element samples in most sandy oxic aquifers,including those in the New Jersey Coastal Plain, even when using

filtration and ‘‘low-flow’’ or ‘‘near-low-flow’’ purging techniques(Gibs et al., 2000). Turbidity was monitored every 3–5 min by use ofa Hach Ratio/2100P Portable Turbidimeter. Pumping of water wasaccomplished with either a portable peristaltic or mechanicalbladder pump (Geoprobe MB470) with steady flow rates forsampling (0.3–0.5 L min�1). Water level was measured two to fourtimes with an accuracy of �0.3 cm by use of a weighted steel tapeboth before emplacement of the portable pumping equipment andin the first 10 min after initiation of purging. Private supply wellswere purged for a period of 30–60 min until physical (temperature,specific conductance) and chemical properties (pH, dissolvedoxygen concentrations) were stabilized, and water turbidity waslowered to levels on the order of 0.5 NTU, in order to insurecollection of representative samples from the aquifers (Gibs et al.,2000; U.S. Geological Survey, 2004). The 10 domestic drinking-water wells sampled were between 16.5 and 33.5 m (54–110 ft)deep and were screened 9.8–28.7 m (32–94 ft) below the watertable (Deluca et al., 2006).

Low-level trace-element sampling techniques recommended byHorowitz et al. (1994), Ivahnenko et al. (1996), and Olson andDeWild (1999), including acid (5% HCl) purges and soaks of allequipment contacting the samples to preserve sample integrity,were incorporated into the sampling program, including that forthe waste media (septic-tank liquids). All samples were pumpedthrough pre-cleaned Teflon-lined polyethylene tubing. Samples forradionuclide and trace-element analyses were filtered (0.45-mmpolysulfone capsule filter) and preserved with ultrapure nitric acid(Ivahnenko et al., 1996; Szabo et al., 2008). The septic-tank liquidswere pumped directly from 4 to 5 points in the septic tank (Miller,1996) using a high-volume peristaltic pump, and were filteredthrough a 100-mm mesh polyethylene or nylon bag filter to removesolid particles, discharged directly into a pre-cleaned polyethylenebucket (compositing and homogenization), then pumped througha 0.45-mm filter into sample bottles. Filtration was reasonablebecause the Ra reaching the leachfield and potentially migrating tothe water table was most likely in the dissolved form, and the Ra inparticulate form (adsorbed?) settled with particulates to the sludgelayer in the bottom of the septic tank (Szabo et al., 2008). Replicateswere collected sequentially into bottles from pumped ground waterand sequentially from the compositing buckets for septic-tankliquids. Replicate samples were collected for one or more matrixesreported here at 6 of 10 sites. Equipment blanks were collectedbefore sampling was initiated.

3.2. Analytical methods

Concentrations of Ra radionuclides were determined afterchemical separation from the water sample as a co-precipitate withbarium sulfate (Table 1). Concentrations of 226Ra were determinedeither by the planchet counting method or the 222Rn de-emanationmethod (Krieger and Whittaker, 1980; USEPA, 1997); the latter isthe more sensitive technique with lower detection levels (Table 1).Concentrations of 228Ra were determined by the beta counting ofthe ingrown 228Ac progeny (Parsa and Hoffman, 1992; USEPA,1997). The sample specific minimum detectable concentration(SSMDC) of 228Ra ranged from 0.0111 to 0.037 Bq L�1; the mode was0.015 Bq L�1, and this value was reported for statistical populationconcentration distributions. Gross alpha-particle and beta-particleactivity was analyzed within 48 h (maximum, 72 h) after samplecollection, as required by the State of New Jersey (New JerseyAdministrative Code, 2002) on the basis of the studies of Parsa(1998), and was determined using planchet counting with 230Thand 137Cs as the standards, respectively. The co-precipitationtechnique was used for sample preparation of the most brinysamples from the septic-tank liquids and water from the water

Table 1Description of methods for analysis of ground water and septic-tank effluent con-taining cation-exchange regeneration brine waste, New Jersey Coastal Plain, 2003–04. [1 Bq L�1 ¼ 27.027 pCi L�1].

Constituent Laboratoryreporting orminimumdetection level

Method Citation

226Ra 0.0037Bq L�1

222Rn de-emanation(EPA 903.1)

USEPA (1997)

226Ra Variable,0.011–0.019Bq L�1

Planchet count afterbarium sulfateco-precipitation(EPA 903.0)

Krieger andWhittaker (1980)

228Ra Variable,0.011–0.037Bq L�1

Beta counting of Ac-228ingrowth after bariumor lead sulfateco-precipitation(EPA 904.0)

USEPA (1997);Parsa and Hoffman(1992)

Gross alpha(48 to 72 hand 30 dayholding time)

Variable,0.037–0.111Bq L�1

Low-backgroundproportional count(EPA 900.0) afterevaporation for lowdissolved solidsground water or drinkingwater samples or afterco-precipitation for highdissolved solids brineor septic effluentsamples

Parsa (1998); USPA(1997); Krieger andWhittaker (1980)

Gross beta(48 to 72 hand 30 dayholding time)

Variable,0.037–0.148Bq L�1

Low-backgroundproportional countafter evaporation(EPA 900.0)

Krieger andWhittaker (1980)

pH 0.1 standardunits

Field. Electrode. U.S. GeologicalSurvey (2004)

Specificconductance

1 microsiemenper centimeter(uS cm�1)

Field. Electrode. U.S. GeologicalSurvey (2004)

Turbidity 0.1 NTU Field. Nephelometry Gibs et al. (2000)Major cations

and ironVariable byanalyte;0.001–0.02mg L�1

Inductively coupledplasma – opticalspectroscopy (ICP-OES)

Harris et al. (1997)

Uranium, selecttrace elements

Variableby analyte;typically0.0002 mg L�1

Inductively coupledplasma – Massspectrometry (ICP-MS)

Faires (1993)

Nitrate plusnitrite;Nitrite;Ammonia;Chloride

0.1; 0.01;0.02, allas mg L�1 N;and 0.1mg L�1,respectively

Ion Chromatography/Colorimetry

Fishman andFriedman (1989)

Table 2Mean relative analytical precision estimate for analyses of radionuclides or radio-activity for ground water used for drinking water and for shallow ground watermixed with septic-tank effluent, for homeowner study, New Jersey Coastal Plain,2003–04; mean relative analytical precision estimate for samples from sites wherereplicates were collected; and mean relative percent difference for replicatesamples. [Samples at or below the MDC were excluded.].

Quality-Assurance 226Ra (%) 228Ra (%) Alpha Beta activity

Z. Szabo et al. / Journal of Environmental Radioactivity 101 (2010) 33–44 37

table; evaporation was used for the remaining samples. Grossalpha-particle activity of samples of water from the drinking-waterwells at 7 of the sites were re-counted after 30 days; the noteddecline was consistent with the presence of 224Ra (Szabo et al.,2005, 2008), but because of the short half life, this isotope was notconsidered for mass balance computations.

Measure activity(48 to 72-hmeasurement)(%)

(48 to 72-hmeasurement) (%)

Mean relative analyticalprecision estimate,all sites

25 30 33 16

Mean relative analyticalprecision estimate,sites with replicatesamples

21 27 34 17

Mean relative percentdifference, replicates

12 4.8 18 8.4

3.3. Quality assurance (analytical precision, replicates, andequipment blanks)

The laboratory analytical precision (pe, propagated precision (oruncertainty) estimate, of which the counting error is typically thelargest component) was about �20% to �35% for concentrations of226Ra, 228Ra, and gross alpha-particle activity (means, �25, 30, and33%, respectively) for the ground-water samples collected from themoderate-depth drinking-water wells and from drive points at thewater table (Table 2). Especially the SSMDC of 228Ra (0.011–0.037

Bq L�1), which was higher than that of 226Ra, and the relativelylarge pe’s associated with the results at or near the SSMDC, limitedconfidence in individual results for that isotope. The precision inrelative terms was lowest (greatest uncertainty) when the Raconcentrations and radioactivity were lowest. These lowestconcentrations typically were from the ground-water samplescollected at the water table that had high dissolved solidsconcentrations. Excluding samples with concentrations at theSSMDC, the pe was greater than �50% for concentrations of 226Ra,228Ra, and gross beta-particle activity for 1 sample among the 10sites, but for 5 samples for gross alpha-particle activity. Sequentialreplicate samples indicated results were reproducible within theerror bounds of the analytical techniques. All replicate results werewithin the bounds of the pe for the concentrations of 228Ra andgross beta-particle activity. One set of replicate results forconcentrations of 226Ra and gross alpha-particle activity wereoutside the range of overlap of the respective precision estimates inthis study. For the one sample where concentrations of 226Ra didnot overlap within the analytical error, the difference between thereplicates exceeded the pe by 0.004 Bq L�1. The mean relativepercent difference (RPD) among all the replicates (for 226Ra, �12%)was less than the mean pe for all the samples for which replicateshad been collected and for all the samples from the 10 sites asa whole (Table 2). For gross measures of radioactivity generalreproducibility for replicates was about �20%, but were withinlarger measurement error bounds (up to �35%) than for the Raisotopes. Radium concentrations and gross measures of radioac-tivity generally were reproducible. Radium concentrations could beresolved among samples of the various media at individual sitesdespite moderately large relative measurement error bounds; onlyat 2 sites did large relative measurement error pose a problem indistinguishing results. (The results for reproducibility are consis-tent with those from the larger data set for the various water andwaste media compiled by Szabo et al., 2008). Neither radionuclideof Ra was detected in the equipment-blank samples indicating thatrandom sample-handling bias in the field or the laboratory was notan additional source of uncertainty.

3.4. Statistical and mass-balance methods

To characterize the distribution of Ra in each of the sampledmedia, the quartiles (median, and the 25th and 75th percentile) ofthe Ra-concentration data were determined, as were the minimaand maxima. Raw ‘‘uncensored’’ concentration values were used for

Z. Szabo et al. / Journal of Environmental Radioactivity 101 (2010) 33–4438

Ra concentrations and measured radioactivity in statistical evalu-ations even if less than the individual SSMDC, although for thepresentation of the concentration distributions, the ‘‘censored’’(<MDC) values (or their mode if computable) are reported. Corre-lations of combined Ra concentrations with ancillary water-qualityconstituents or characteristics were evaluated by the non-para-metric (rank) Spearman correlation test. Statistical significance wasdetermined at the one-tailed 0.1 confidence level on the basis of thesmall sample population. Relative amounts of Ra mobilization by,or sequestration from, the discharge of saline waters from septictanks to the shallow water table in these quartz sand aquifersystems were evaluated by mass balance approximation using Cl asa conservative tracer, with equations provided in the appropriatesection, and limitations discussed below.

3.5. Limitations

The mass-balance estimates were limited by large uncertainties.1. Relative analytical precision for Ra was typically on the order of�20% to �25%, and were especially large for both Ra isotopes forthe lowest concentrations. The analytical uncertainty issue isespecially important for 228Ra, because of the higher reporting level(lower detection capability), and poorer relative precision, whencompared to those for 226Ra. The Ra replicate results indicatedgeneral (mean) reproducibility of about �12%, but within largemeasurement error bounds (mean pe up to � 25%). Because grossmeasures of radioactivity had larger measurement error (mean peup to� 35%) than for Ra isotopes, these measures were not used forprocess oriented analysis, such as mass/activity balance. Relativeprecision estimates of �12%, �20%, to � 33% can result in as muchas a 1.3-fold, 1.5-fold, to a 2-fold uncertainty in the mass-balancecalculation. 2. Despite recurring episodic discharge at a frequencyless than the time needed for the effluent to migrate to the watertable, or perhaps even to the leachfield, and despite the efforts tocomposite samples to account for spatial variability (Miller, 1996),the study design was synoptic sampling, in which neither spatialnor temporal variability of the waste materials at individual sitescould be addressed. The synoptic study design was intended tocharacterize the range of variability that is possible on the regionalscale. The spatial and temporal variability of the waste materials atany given site is likely substantial, and could reasonably beassumed to exceed analytical uncertainty for individual results,perhaps doubling the uncertainty. Characterization of variability atsingle sites is beyond the scope of this study.3. Background Raconcentrations at the water table were not defined at each site(logistical constraints). The median value of Ra concentrations atthe water table for the region as determined by Szabo et al. (1997)was used in the mass-balance calculation. The inter-quartile rangeabout the median was small, but still showed variation by abouta factor of 2. While background Ra is only one component in thecalculation of relative Ra mass balance, natural variability mayaccount for an estimated 2-fold factor of variation. Substantial localvariation in radionuclide leaching to the water table is possible asreported by Tricca et al. (2001) in a glacially derived sand aquifer onnearby Long Island. 4. Flow volumes were only estimates in mostcases. Flow meter installations are needed.

Coupling the uncertainties resulting from generalized back-ground Ra concentration characterization, the sample heteroge-neity (especially for the waste liquids), the imprecise analyticalmeasurement, and the imprecision of the estimated flows, it ispossible that the relative uncertainty in the mass-balance calcula-tion for most samples might be equal to or exceed from 4-fold theresult to perhaps 10-fold the result (order of magnitudeapproximation).

4. Results

4.1. Radium concentrations and radioactivity in ground water usedfor drinking water

The combined Ra concentrations in 8 of the 10 filtered untreatedwater samples from the 10 domestic wells sampled for this studywere greater than 0.110 Bq L�1 and in 4 (40%) were greater than0.185 Bq L�1 (the USEPA 5 pCi L�1 MCL) (Fig. 3). (Analytical resultsfor all the samples are presented by Deluca et al., 2006). The medianconcentration of combined Ra in the 10 domestic (drinking watersource) wells was 0.165 Bq L�1, nearly identical to that found for theaquifer systems in previous investigations (Kozinski et al., 1995;Szabo et al., 2005). The maximum concentrations of 226Ra and 228Rawere 0.806 and 0.729 Bq L�1 (21.8 and 19.7 pCi L�1) respectively,with combined Ra of 1.54 Bq L�1 (41.5 pCi L�1), and the maximumgross alpha-particle activity 72 h after sample collection was 4.22Bq L�1 (114 pCi L�1; well 011406). The median pH of the ground-water samples was 4.90, typical of the aquifers as a whole (forwhich the ‘‘pooled’’ median pH value from previous studies was4.9). The median concentration of Cl was 18.4 mg L�1. The medianCl concentration is consistent with the previously determinedmedian values for unconfined aquifers in the New Jersey CoastalPlain (between 5 and 14 mg L�1; Yuretich et al., 1981; Barton et al.,1987; Kozinski et al., 1995; Szabo et al., 1997), with these previousstudies containing sizable sample populations from undeveloped(forested) areas. Three of 10 domestic-well water samples had Clconcentrations between 5 and 14 mg L�1 in this study. Themaximum concentration of Cl in this study was 37.6 mg L�1.

4.2. Radium concentrations and radioactivity in septic-tank liquids

Concentrations of combined Ra in the septic-tank liquids at the10 sampled sites ranged from 0.026 to 0.243 Bq L�1 (Fig. 3) andwere typically about three-fourths (median, 75 percent) that of theuntreated ground water at the site. The combined Ra concentra-tions in 7 of the 10 septic-system effluents sampled were greaterthan 0.110 Bq L�1 and in 3 (30%) were greater than 0.185 Bq L�1

(concentrations, 0.228, 0.237, and 0.243 Bq L�1, respectively)(Fig. 3). The 226Ra was higher in the septic system liquid than theinput untreated ground water at only 1 site, but the 228Ra washighest in the septic system liquids at 3 sites. The source of the Ra inthe septic tanks was the regular discharge of the regeneration brinewaste, which was typically enriched 30-fold in combined Ra rela-tive to the native (untreated) ground water (Szabo et al., 2008); thedesign of each septic tank affected the mixing and residence time ofthe brine waste within the tank.

Gross alpha-particle radioactivity distribution patterns followedthe same trend in the septic-tank liquids as the 226Ra, though nosample exceeded the MCL, perhaps because of decay of 224Raduring residence on the ion-exchange resin and in the septic tank.The gross beta-particle activity in septic-tank liquids ranged from0.108 to 2.78 Bq L�1 and was greater than that of the medium-depthground water (used for drinking) at all the sites. The primary reasonfor the elevated beta-particle activity was the high concentrationsof K in the septic-tank liquids.

All liquid samples from the septic tanks were considerablydifferent in composition than the native ground water. Notablewere the high concentrations of Cl (maximum, 6330 mg L�1;median, 457 mg L�1), Na (maximum, 3990 mg L�1; median, 289 mgL�1), K (maximum, 31.2 mg L�1; median, 12.5 mg L�1), andammonium (as N, maximum, 123 mg L�1; median, 32 mg L�1)(Deluca et al., 2006; Barringer et al., 2006; Szabo et al., 2008). ThepH of the septic effluent at all 10 sites, ranging from 6.23 to 7.43with a median value of 6.83, was at least 1–2 full pH units greater

Ground water (well)

Septic effluent

Water table near leach field

0.01

0.1

1

10

SITES SAMPLED

1109

16/1

1121

5

0114

06/0

1143

6

0114

07/0

1143

7

3306

82/3

3096

8

1515

22/1

5153

6

1515

20/1

5153

7

2314

07/2

3141

8

2314

10/2

3141

9

2313

63/2

3141

7*0.001

MCL

Kirkwood-Cohanseyaquifer system

Potomac-Raritan-Magothy aquifer system

3309

58/3

3099

2*

44.

7.0

5.1

4.4

6.2

5.9

4.96.1 5.2

4.9 6.8

5.9

5.3

7.4

5.3

4.97.0

4.8

4.4

7.0

6.0

4.1

4.9

5.2

7.0

5.5

5.3

6.8

5.5

4.4, pH in standard units

226 R

AD

IUM

PLU

S 2

28 R

AD

IUM

, BE

CQ

UE

RE

LS P

ER

LIT

ER

6.5

Fig. 3. Concentrations of 226Ra plus 228Ra, by media, including ground water used for drinking water, septic-tank effluent, and septic effluent and ground-water-recharge mixture atthe water table, at 10 sites where septic effluent waste dispersal occurs to the environment, Kirkwood-Cohansey and Potomac-Raritan-Magothy aquifer systems, New Jersey CoastalPlain, 2003–04. [The upper horizontal dashed line indicates 0.185 Bq L�1, the USEPA 5 pCi L�1 MCL. The site numbers identify the drinking-water well first then the correspondingobservation well that penetrated to the water table downgradient from the septic leachfield. Sites marked with an * had substantial relative analytical error. All site locations areshown on Fig. 1.].

Z. Szabo et al. / Journal of Environmental Radioactivity 101 (2010) 33–44 39

than water from the associated drinking-water wells sampled atthe sites (median pH, 4.90).

4.3. Radium concentrations and radioactivity in ground water atthe water table downgradient from the septic leachfield

Combined Ra concentrations in water samples from the drivepoints downgradient from septic tanks, presumably representingnative water and septic-tank effluent mixtures, were generally lessthan those in the liquid in the septic tanks (Fig. 3). At half (5 of 10) ofthe sites, combined Ra concentrations were �0.019 Bq L�1 (Table 3;Fig. 3), and of these samples, all but one had pH �5.5. Of the 3 sitesposing exceptions to the concentration trend, that is, combined Raconcentrations at the water table were greater than or equal tothose in the input septic-tank effluent, all were found where the pHat the water table was �5.3 (Figs. 3 and 4). At 3 of the sites,

Table 3Summary of distributional statistics for selected radionuclide and water-quality constituetreatment (septic tank) leachfields, New Jersey Coastal Plain, 2003–04. [1 Bq L�1 ¼ 27.027alpha-particle activity greater than 0.555 Bq L�1 (USEPA 15 pCi L�1 MCL), and gross beta-pbold type. The concentration of 228Ra could not be determined in one sample; combined

Constituent Units Number Minimum

All Samples

pH Standard 10 4.80Specific Conductance mS cm�1 10 188Chloride mg L�1 10 17.7Nitrate (as N) mg L�1 10 3.95Ammonia (as N) mg L�1 10 0.03Sodium mg L�1 10 3.95Calcium mg L�1 10 0.867Potassium mg L�1 10 1.90226Ra Bq L�1 10 <0.0037228Ra Bq L�1 9 <0.015226Ra plus 228Ra Bq L�1 10 <0.019Alpha activity (48 to 72 h) Bq L�1 10 <0.037Beta activity (48 to 72 h) Bq L�1 10 0.117

combined Ra concentrations were greater than 0.110 Bq L�1 (0.118,0.139, and 0.985 Bq L�1 – this maximum concentration consider-ably exceeded the 0.185 Bq L�1 MCL). The 2 highest gross alpha-particle activities (0.518 and 2.66 Bq L�1) were from sites withamong the lowest of pH’s (�5.14) and among the highest combinedRa concentrations.

The maximum combined Ra concentration at the water table of0.985 Bq L�1 was from drive point 151536 where pH at the watertable was the minimum, 4.80 (Figs. 3 and 4). The Cl and nitrateconcentrations at the water table at this site were about 50-foldgreater than the concentrations at moderate depths in the aquifer(drinking-water well 151522) there. Combined Ra concentration atthe water table from drive point 011437 (adjoining well site 011407)was about the same as in the effluent in the septic system; pH at thewater table there was 5.18 (Fig. 3). The combined Ra concentration inthe sample from the water table at drive point 330992 (pH, 5.31) was

nts in water samples collected at the water table 3–10 m downgradient from on-sitepCi L�1. Concentrations of Ra greater than 0.185 Bq L�1 (USEPA 5 pCi L�1 MCL) grossarticle activity greater than 1.85 Bq L�1 (USEPA 50 pCi L�1 MCL screen) are shown inRa was set equal to the concentration of 226Ra for statistical computation.].

1st Quartile Median 3rd Quartile Maximum

5.15 5.40 5.77 6.04316 588 1370 178039.6 113 366 4115.68 9.76 13.7 25.1<0.04 0.05 0.11 5.423.2 99.0 200 2839.39 14.1 29.3 55.63.19 4.36 9.50 11.6<0.0037 0.0165 0.046 0.336<0.015 <0.015 0.096 0.650<0.019 <0.019 0.128 0.9850.055 0.099 0.464 2.660.211 0.370 0.768 2.28

Ground water (well)water table (near leach field)

MCL Regional background at the water table

0

0. 2

0.4

0.6

0.8

1.0

1.2

1.4

4 5 4.5 5.5 6 76.5PH, S TANDARD UNITS

1.6

226

RAD

IUIM

P

LUS

228

RAD

IUM

, BE

CQ

UER

ELS

PER

LIT

ER

Fig. 4. Concentrations of 226Ra plus 228Ra in ground water at depth and at the watertable where septic effluent and ground-water-recharge mix, as a function of pH,Kirkwood-Cohansey and Potomac-Raritan-Magothy aquifer systems, New JerseyCoastal Plain, 2003–04. [The upper horizontal dashed line indicates 0.185 Bq L�1, theUSEPA 5 pCi L�1 MCL. The lower horizontal dashed line indicates 0.110 Bq L�1, previ-ously determined by Szabo et al. (1997) as the median concentration of combined Ra atthe water table in the Kirkwood-Cohansey aquifer system in and near the sameagricultural/residential areas studied in this investigation. Ground-water samples fromdepth at 5 additional sites where drive points to the water table could not be installedare included for purpose of comparison; details for those 5 sites are given by Szaboet al. (2008).].

Z. Szabo et al. / Journal of Environmental Radioactivity 101 (2010) 33–4440

greater than the concentration in water from the deeper drinking-water well at the site (well 330958). The measured Ra concentra-tions at the water table and in the deeper ground water at this lastsite were the smallest or among the smallest of any of the samplesfrom the 10 studied sites, and could not be resolved among eachother with confidence when considering the magnitude of themeasurement error (the precision estimates were relatively largewhen compared to the concentrations). Combined Ra was present inwater from the water table from drive point 111215 (adjoining wellsite 110916) at moderately elevated concentrations (0.118 Bq L�1)and pH was low (5.14), but the concentration was less than that inthe septic-tank effluent at the site.

The pH (median, 5.40; range, 4.80–6.04) at the water table wasgreater than that for the moderate-depth ground water withdrawn(for drinking) at the same site (median pH, 4.90; range, 4.15–5.35),though at 2 sites (well/drive point 330958/330992 and 151522/151536), the pH’s for water from the two depths were about equal(Fig. 3). The median difference among results at the water table andthe (drinking) water sample from depth was þ0.54 standard pHunits and ranged to þ1.67 pH units. The water samples from thewater table also had high concentrations of Cl (median, 113 mg L�1;maximum, 411 mg L�1), with the exception of 2 sites where the Clconcentrations were 17.7 and 18.3 mg L�1, respectively. At all thesites, the Cl concentration was higher at the water table than in thedeeper ground water produced by the drinking-water wells.

The gross beta-particle activity in water samples from the watertable ranged from 0.185 to 2.28 Bq L�1.(Table 3). Samples from thewater table also had moderately high concentrations of K (median,4.36 mg L�1; maximum, 11.6 mg L�1), with the exception of 2 siteswhere the K concentrations were<3 mg L�1. At all but 2 sites, the Kconcentration was higher at the water table than in the deeperground water from the drinking-water wells. At the water table,then, where septic-tank effluent is present, 40K may contributelargely to the measured gross beta-particle activity, and thus, evenif elevated, this measure of radioactivity is not necessarily a reliableindicator for elevated levels of Ra.

5. Discussion

5.1. Relative mass-balance estimates and fate of Ra in septic effluentreleased to the environment

Mass-balance estimates were a tool to further evaluate thepossible fate of the discharged Ra. Dilution of the effluent afterdischarge to the environment occurs by mixing with rechargeoriginating from natural sources (precipitation and/or existing soilmoisture). To initially evaluate dilution, the concentrations of Raand Cl in ground-water samples from the water table relative to theliquids in the adjoining septic tanks were compared. The inter-quartile range of the ratios of the Ra concentrations in groundwater from the water table downgradient of the septic leachfieldrelative to the Ra in the septic effluent itself was from 0.03 to 1.00for combined Ra, and 0.22 to 1.33 for 226Ra, with median values of0.21 and 0.75, respectively (Table 4). The median ratio for Clconcentrations among the two media was 0.345, with inter-quartilerange from 0.125 to 0.65 (Table 4). Presuming conservativebehavior for Cl, dilution was notable. The ratio of concentrations atthe water table relative to that in the septic effluent (source) washigher for combined Ra than Cl at 6 of 10 sites (Table 4; 7 of 10 ifonly 226Ra is considered for drive point 330968 for which 228Raconcentration could not be determined).

Relative mass loss of Ra from (or gain to) the septic-tank effluentduring flow through the leachate field and underlying vadose zoneafter outflow from the septic tank was estimated using theconservative tracer (chloride) method (Kroeger et al., 2006).Because the soils in the southern New Jersey Coastal Plain are acidic(pH typically �5.0), anion adsorption from percolating wateroccurs. Chemical extractions of the local soils indicated that thesulfate anion was adsorbed preferentially to Cl and nitrate, however(Reilly and Baehr, 2006).

The dilution of the effluent by uncontaminated recharge usingthe concentration of Cl as the conservative tracer was determinedas the following:

f ¼�Cl;wt � Cl;bk

���Cl;septic � Cl;bk

�(1)

where Cl represents Cl concentrations in the various media indi-cated by subscripts (wt, measured at the water table; bk, back-ground, represented by measured concentration of medium-depthground water from the drinking-water wells; septic, measured inseptic effluent), and f is the fraction of the water-table samplecomposed of septic-tank liquid outflow. The median Cl concentra-tion among all the ground-water samples collected in this studyfrom the (moderate depth) wells used for drinking-water supplywas 18 mg L�1. The background Cl concentration in water from theaquifer systems in forested (undeveloped) areas is about 5–14 mgL�1 (Yuretich et al., 1981; Barton et al., 1987; Szabo et al., 1997). Asthere is no in-situ source of Cl in the aquifer, the measured Clconcentrations from samples from the drinking-water wells wereconsidered reasonable approximations of the background Clconcentrations of water recharged at each site.

The Ra concentration at the water table after dilution was thendetermined as

Ra;cs ¼��

Ra;septic � f�þ�Ra;bk � ð1� f Þ

��(2)

where Ra,cs is the expected Ra concentration at the water table ifeffluent bearing Ra had mixed conservatively with ‘‘background’’recharge water bearing Ra concentration of Ra,bk with no additionalloss (change in mass), and (1 � f) is the fraction of the samplecomposed of recharge water with Cl concentration similar to that ofbackground soil and ground water. Because there is clearly a source

Table 4Concentration ratio of samples of ground water from the water table 3–10 m downgradient from the leachfield and the source effluent (liquid component of septage from theseptic tank), and the relative mass loss or gain of Ra and radioactivity of the leachfield–effluent-and-soil–water mix at the water table relative to the source effluent (liquidseptage) after passing through the unsaturated zone and mixing with soil–water recharge and ground water, New Jersey Coastal Plain, 2003–04. [A negative sign indicatesrelative mass gain whereas positive sign indicates relative mass loss. The mass balance and mixing equations are detailed in the text. Site numbers are those of the observationwells and are consistent with Fig. 3. Sites where pH at the water table is less than or about equal to 5.3 are shown in bold type. Results from sites marked with an * hadsubstantial relative analytical error. Original units for measurement were Bq L�1, except mg L�1 for chloride, and standard units for pH, as reported in Table 3.].

Site number pH Beta radioactivity, gross Alpha radioactivity, gross 226Ra 228Ra 226Ra plus 228Ra Chloride 226Ra Plus 228Ra 226Ra 228Ra

Concentration Ratio, Water Table at Leachfield: Effluent (Liquid Septage) Relative Mass Loss or Gain, Water table111215 5.14 0.5 1.37 0.89 0.36 0.50 0.27 �0.026 0.004 �0.05011436 5.91 0.17 0.12 0.03 0.02 0.02 0.11 0.95 0.94 0.97011437 5.18 1.63 1.76 0.99 1.01 1.00 0.53 �0.11 0.27 �0.31330968 5.86 – – 0.3 – – 0.02 – 0.56 –330992* 5.31 3.11 1.55 3.00 1.90 2.45 0.91 �0.80 �1.20 �0.40151536 4.80 4.88 9.38 21.94 14.95 16.77 1.02 �15.77 �20.94 �13.95151537 6.04 1.08 – 0.03 0.04 0.03 0.69 0.96 0.97 0.96231418 4.88 0.04 0.01 0.19 0.03 0.07 0.01 0.89 0.85 0.93231419 5.50 0.77 0.23 1.44 0.02 0.03 0.42 0.95 0.93 0.97231417* 5.48 0.25 0.23 0.60 0.13 0.21 0.17 0.94 0.94 0.94

Per-centiles Summary of Population Distribution Statisticsa

25th 5.15 0.25 0.2025 0.22 0.03 0.03 0.125 �0.11 0.00a �0.3150th 5.40 0.77 0.8 0.75 0.13 0.21 0.345 0.89 0.85a 0.9375th 5.77 1.63 1.6025 1.33 1.01 1.00 0.65 0.95 0.94a 0.95

a Number of results are not equal for each constituent; only the result from the relative mass loss for 226Ra for site 330968 was excluded from calculation of populationdistribution statistics (for ease of comparison).

Z. Szabo et al. / Journal of Environmental Radioactivity 101 (2010) 33–44 41

of Ra within the aquifer, the Ra concentrations from the on-sitedrinking-water wells sampled in this study were not, therefore,considered representative of background Ra concentration of waterat the water table or of recharge (soil pore water) in the vadosezone. The assumption was made for the mass-balance calculationthat actual background concentrations of combined Ra in rechargewater at each site (Ra,bk) were reasonably approximated by theregional water-table Ra concentration median value of 0.055 Bq L�1

for each Ra radionuclide that was determined during the previousregional study by Szabo et al. (1997).

The fraction (F) of relative Ra mass loss or gain after mixing wasthen calculated by:

F;Ra mass loss ¼�Ra;cs � Ra;wt

��Ra;cs (3)

where Ra,cs is defined as before (computed concentration ofconservative mixture), and Ra,wt is the actual Ra concentrationmeasured at the water table. The fraction of Ra mass loss (or gain) isshown in Table 4. Positive numbers indicate Ra sequestration or(and) combined Ra concentration substantially less than 0.110 BqL�1 in local recharge. Negative numbers indicate Ra gain, eitherfrom leaching from the sediment or (and) contribution from Ra inpore water (percolating recharge) with combined Ra concentrationgreater than 0.110 Bq L�1 (the occurrence of which is possible atindividual locales in the New Jersey Coastal Plain; Szabo et al.,1997). The median value of the fraction, F, indicated that the relativemass loss of combined Ra was 89% (Table 4), with maximum, 96%.The median relative mass losses were 85 and 93% for 226Ra and228Ra, respectively (9 sites, 228Ra concentration was not determinedat the water table at site 330958, thus for sake of comparability,results from this site are excluded from the population distributiondetermination for 226Ra mass balance). Since both 226Ra and 228Raconcentrations were near, at, or below detection at the water tableat 5 of the sites, it is not surprising that substantial mass lossbeyond that from dilution is indicated as typical. The lowest quar-tile of calculated results for the fraction, F, indicates negative values(relative mass gain), consistent with the fact that at 3 sitescombined Ra concentrations at the water table were greater than orequal to those in the input Ra-bearing septic-tank effluent.

Differing fates for the masses of Ra during transport to the watertable by the effluents were related to the pH. At half the sites, the

pH at the water table of the effluent–water mixture was mildlyacidic (pH �5.3, with maximum, 6.04), only slightly less than thoseof the septic-tank effluent, and combined Ra concentration for theeffluent–water mixture was low or not detected. The pH (median,5.40) at the water table was higher than typical for the deeperground water from the aquifers (median, 4.90) as the result of thepresence of (mixing with) near-neutral septic effluent (median pH,6.83) that contained notable carbonate alkalinity and hardness. Theplausible explanation is that Ra from the discharging septic effluentwas removed (sequestered) during transit through the soil andvadose zone in the leachfield where the water was near neutral oronly mildly acidic (Table 4). Sorption (or ion-exchange) mecha-nisms were effective in the near-neutral pH environments forcationic constituents, such as Ra.

At half the sites, the septic effluent evolved from a neutralsolution to a moderately acidic (pH �5.3) solution (effluent–watermixture) at the water table. Net Ra mass addition (enrichment) inthe septic effluent at the water table is evident at three of thesesites (first quartile; Table 4) despite the noted dilution. At one site,relative mass balance (subtraction/addition) is near zero (nochange). At drive-point site 151536 with the maximum Raconcentration at the water table (greater than the MCL) and thelowest pH (Fig. 4), the relative mass of Ra entering the effluent–water mixture in the vadose zone was such as to increase Ra massin solution by nearly an order of magnitude relative to that in theinitially discharged septic effluent (Table 4; Fig. 3). The maximumcombined concentrations of Ra from soils and shallow sedimentsamples adjoining or near the studied sites was 40 Bq kg�1 (Szaboet al., 2008) indicating the presence of low but measurable sourceof Ra.

5.2. Regional comparison and statistical analysis for assessingsignificance of the mass-balance estimate

The trend of the results of the mass-balance calculations pre-sented, despite the substantial limitations (described earlier as 4-fold to 10-fold uncertainty), is in agreement with the documentedtendencies of Ra concentration distribution of the unconfinedAtlantic Coastal Plain aquifer systems as a whole – that is, Ramobility is greatest at low pH, and the concentrations of Ra are

Z. Szabo et al. / Journal of Environmental Radioactivity 101 (2010) 33–4442

substantially greater in those samples where pH is less than 5.2than in those samples where pH is greater than 5.2, as observed inregional studies (Kozinski et al., 1995; Szabo et al., 1997, 2005). Thetrend in mass loss or gain relative to pH was the same for both Raisotopes despite minor differences in precision and detectioncapability (Table 4), and was consistent among the sampled sitesdespite minor local differences among geological and hydrologicalcharacteristics.

The critical inverse relation of combined Ra concentrations withpH is also indicated by the statistically significant Spearmancorrelation coefficients, r, of�0.52 for samples from the water tableand �0.77 for samples from the deeper ground water (drinking-water wells). In contrast, the positive relation of combined Raconcentrations with dissolved solids concentration as representedby the specific conductance is weak and is not statistically signifi-cant (r, þ0.14 and þ0.18, respectively, for samples from the watertable and from the deeper ground water from the drinking-waterwells). Nathwani and Phillips (1979), Kraemer and Reid (1984), andMiller and Sutcliffe (1985) all showed Ra mobilization in soils oraquifers subject to increasing dissolved salinity and presumed theincrease in competition for exchange sites, perhaps with divalentcations, was the mechanism responsible. The hydrogen ion,however, can compete successfully for exchange and sorption sitesrelative to divalent cations (Apello, 1994), especially in dilutesolutions. Competition with free hydrogen ion (acidic water)appears the primary mechanism controlling Ra desorption at thewater table and at depth in these unconfined quartzose CoastalPlain aquifer systems, whereas absence of such competition (innear-neutral or neutral waters) can result in sequestration.

The mass-balance estimates presented here have limitationsand are considered a preliminary tool in understanding the natureand magnitude of the potential problems related to Ra-bearingwaste disposal and in guiding further research efforts. The trends inrelative mass loss or gain for Ra with respect to pH establisha plausible mechanism explaining Ra mobility. Because quantifi-cation was somewhat imprecise, capability to understand therelative importance of other factors that might affect the fate of Rawere limited. A designated research site, where some criticalexplanatory variables might be controlled for a monitoring periodof perhaps several seasons, might improve confidence (decreaseuncertainty) in the mass-balance estimates. Because of the highdegree of Ra mobility, site 151522 (with drive point 151536) mightbe an example of one end member to prioritize for additionalcharacterizations. Conversely, a site similar to 151520 (with drivepoint 151537), where Ra appears to be strongly sequestered, mightbe the opposing end member.

5.3. Implications of study results for protection of ground-waterquality

The protection of the quality of the ground-water resource usedfor drinking water is a critical issue when considering the effects ofthe septic effluent discharge (USEPA, 2003). Concerns are height-ened when Ra-bearing septic-system effluent is discharged to theground because of the health risk posed by ingestion of Ra-richwater (Mays et al., 1985; USEPA, 1999, 2000). To the knowledge ofthe investigators, this study is the first to characterize the fate of Rain the liquid brine waste media derived from the cation-exchangetreatment of Ra-bearing drinking water and disposed to the envi-ronment in the Atlantic Coastal Plain of New Jersey, where thehighly permeable and coarse surficial deposits result in aquiferswith high sensitivity to surficial contamination. At half the studiedsites, Ra concentration (�0.019 Bq L�1) was barely detectable at thewater table, and dilution and sequestration from effluent wasindicated. At one of 10 (10%) water-table drive points (151536),

however, the combined Ra concentration of 0.985 Bq L�1 (Fig. 3) atthe water table was greater than the MCL, and was greater than thecombined Ra concentration in ground water at depth used fordrinking-water supply (from adjacent well 151522). Thus dispersalof Ra wastes to the environment and enhanced Ra mobilization atthe most acidic sites (pH �5.3) poses some risk of future water-quality degradation, and thus may pose additional future exposurerisk to those private-well owners in the region who are unaware ofthe problem. On the other hand, ingestion exposure through thedrinking-water pathway is being avoided now by homeownerswith private wells using well-maintained cation-exchange treat-ment units (Szabo et al., 2008). Use of properly maintained watersofteners in order to avoid Ra exposure is one of several optionsdetailed by New Jersey State officials to private-well owners in theAtlantic Coastal Plain of New Jersey where radiation (gross alpha-particle activity) in the water is elevated (NJDEP, 2004).

Percolation of septic-system effluents (or leachates from sludgethat was land applied; Jacobsen, 2000) typically affects each of fourenvironmental geochemical factors that most control leaching of Rafrom natural soils and vadose-zone sediments: pH, oxidation–reduction potential, total dissolved solids content, and the overallcomposition of the soil solution (Champlin and Eichholz, 1976;Nathwani and Phillips, 1979; Landa et al., 1991). The pH is the mostcritical variable for Ra mobility in southern New Jersey unconfinedCoastal Plain aquifers. Septic-system effluents generate acidity byoxidation reactions of commonly occurring waste compounds, suchas ammonia. The concentrations of nitrate at the water tabledowngradient from the septic leachfields at the 10 sites sampled inthis study ranged from 3.95 to 25.1 mg L�1 (Table 3) indicating thepotential for acidification as the effluent–water mixture (bearingammonia) was oxidized. The nitrification reaction produces freehydrogen ion as shown:

NH3 þ 2O2/NO�3 þ H2Oþ Hþ (4)

The conversion of the requisite 0.6–3.6 mmol L�1 (millimolesper liter) of ammonia (as nitrogen) from the percolating effluent toform the nitrate at the observed concentration levels at the watertable could release enough free hydrogen ion (2 mmol of freehydrogen ion per mmol of ammonia consumed) to neutralize 73–440 mg L�1 of bicarbonate alkalinity present in the original septiceffluent. Half the samples from the water table had bicarbonatealkalinity concentrations �10 mg L�1 despite the discharge ofbicarbonate-bearing septic effluent. Oxidation of ammonia in theseptic effluent is one reason (mixing; and hydrolysis of dissolvediron or other cations are others) that pH (and bicarbonate alka-linity) of the effluent and soil–water mixture decreased relative tothe septic effluent. Where pH is lowest (�5.3), the aquifer isperhaps locally most quartzose and has the least buffering capacity.Using mass-balance calculations, Wilhelm et al. (1996) showed thatoxidation of ammonium alone was sufficient to consume most orall of the alkalinity in effluents from the septic systems they studiedoverlying a carbonate-mineral-free glacial sand aquifer – pHdecreased at or near the water table there as well. The fate ofnitrogen in fertilizer might be considered a reasonable, though notidentical, analogue for oxidation of ammonia from septic-tankeffluents. The pH in water from southern New Jersey unconfinedquartzose aquifers in areas of fertilized agricultural land use waslowest (typically <4.8) where nitrate concentrations were highest(5–23 mg L�1; Szabo et al., 1997), and was about the same or lowerthan those observed in this study downgradient from the septicsystems.

In areas where the Ra appears to be currently sequestered, andthereby disposal of regeneration brine waste does not appear topose immediate water-quality concerns, only long-term monitoring

Z. Szabo et al. / Journal of Environmental Radioactivity 101 (2010) 33–44 43

can insure that ground-water contamination with Ra does not occurin the future. Regional-scale monitoring of drinking-water qualitymight be completed within the framework of a private-well testingprogram, such as that instituted in New Jersey (New JerseyAdministrative Code, 2002), and could be coupled perhaps withmore detailed sampling of observation wells in the areas most at risk(pH <5.3), using recommended monitoring guidelines (Interna-tional Steering Committee of Radionuclide Standards ISCORS, 2004).The sequestration capability of the vadose-zone sediments mighteventually be exhausted or geochemical conditions might change(pH might decrease further, for example, or dissolved solids mightincrease sharply), thereby liberating the Ra, indicating the need formonitoring.

6. Conclusions

The fate of Ra in the septic-tank effluent–pore water mixture atthe water table downgradient of the leachfields in unconfinedaquifers in the New Jersey Coastal Plain was studied. Combined Raconcentrations were commonly low at the water table, with half (5of 10) sites having concentrations �0.019 Bq L�1. Combined Raconcentrations in the discharged septic-tank effluents ranged from0.026 to 0.243 Bq L�1, with the Ra in the effluents ultimatelyderived from discharge of cation-exchange regeneration brinewaste to the septic tanks. The effluents were about neutral in pH(median, 6.83). Transit through the soil to the water table resultedin varied degrees of acidification of the septic effluent plumes bymixing with acidic recharge, and pH may have been furtherreduced by various oxidation reactions characteristic of septiceffluent, such as nitrification. The median pH at the water table ofthe effluent–water mixtures was 5.40 (range, 4.80–6.04). Radiummass reaching the water table was strongly dependent upon pH.Combined Ra concentrations at the water table were less than inthe discharging septic-tank effluents when pH at the water tablewas >5.3, and the decrease in mass was more than could beaccounted for by simple dilution indicating Ra sequestration ontoleachfield sediments. In moderately acidic conditions (pH �5.3) atthe water table, Ra was readily detectable, and at 3 sites, combinedRa concentrations were greater than in the discharging septic-tankeffluents despite dilution. Maximum combined Ra concentration atthe water table was 0.985 Bq L�1 where pH was lowest (4.80).Radium was contributed to the effluent–water mixture in acidicenvironments, likely from vadose-zone sediments, which containlow but detectable (up to 0.40 Bq kg�1) concentrations of Ra. Thus,while in the majority of cases the water at the water table was notdirectly contaminated by Ra from the discharging septic-tankeffluents, Ra in the discharging effluent did remain mobile andadditional amounts of Ra were mobilized from the aquifer matrix inacidic environments (pH �5.3). Acidic environments can beattained in the quartzose (carbonate-mineral-free) sands of theAtlantic Coastal Plain unconfined aquifer systems.

Acknowledgements

This project was completed cooperatively by the United StatesGeological Survey (USGS) and the New Jersey Department Envi-ronmental Protection (NJDEP) Division of Science, Research andTechnology. Additional analytical support was provided by theUSGS Toxics Hydrology Program and the USGS National WaterQuality Assessment (NAWQA) Program. We thank NJDEPcolleagues Dr. R. Lee Lippincott and Patricia Gardner for assistancein devising the sampling program and administering it. Assistancein finding sampling locations was provided by numerous localassociations including Delaware Valley Regional PlanningCommission, Salem County Watershed Taskforce, and the

Middlesex County Department of Health. We thank USGScolleagues Timothy Reilly and Larry Feinson for operation of theGeoprobe equipment, Elizabeth Keller for equipment cleaning, andWilliam Ellis for assistance with graphics. We acknowledge helpfulsuggestions to improve this manuscript by USGS colleagues RalphSeiler, Brian Katz, and Edward Landa, and by Joseph Drago ofKennedy Jenks Associates. The use of trade names is for identifi-cation purposes only and does not imply endorsement by theUnited States government or the State of New Jersey.

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