Toxic Substances Hydrology Program

Prepared in cooperation with the

Air Force Center for Engineering and the Environment

Transport of Nitrogen in a Treated-Wastewater Plume to 

Coastal Discharge Areas, Ashumet Valley, 

Cape Cod, Massachusetts

Scientific Investigations Report 2013–5061

U.S. Department of the Interior

U.S. Geological Survey
Cape Cod Bay

Buzzards Bay

Coonamessett Pond Ashumet Pond Johns Pond

Little Pond

Great Pond

Green Pond

Bournes Pond

Nantucket Sound

Cover photo. View northward from above Nantucket Sound of the narrow coastal saltwater embayments, referred to as coastal ponds, 

in southern Falmouth, Cape Cod, Massachusetts, and several freshwater ponds 2 to 4 miles inland from the coast. The Massachusetts 

Military Reservation is the tan area in the top right of the photo. Photography by Joseph R. Melanson, Aerials Only (http://www.skypic.com), 

used with permission. 
Transport of Nitrogen in a TreatedWastewater Plume to Coastal Discharge 

Areas, Ashumet Valley, Cape Cod, 

Massachusetts

By Jeffrey R. Barbaro, Donald A. Walter, and Denis R. LeBlanc

Toxic Substances Hydrology Program

Prepared in cooperation with the

Air Force Center for Engineering and the Environment

Scientific Investigations Report 2013–5061

U.S. Department of the Interior

U.S. Geological Survey
U.S. Department of the Interior

SALLY JEWELL, Secretary

U.S. Geological Survey

Suzette M. Kimball, Acting Director

U.S. Geological Survey, Reston, Virginia: 2013

For more information on the USGS—the Federal source for science about the Earth, its natural and living 

resources, natural hazards, and the environment, visit http://www.usgs.gov or call 1–888–ASK–USGS.

For an overview of USGS information products, including maps, imagery, and publications, 

visit http://www.usgs.gov/pubprod

To order this and other USGS information products, visit http://store.usgs.gov

Any use of trade, product, or firm names is for descriptive purposes only and does not imply endorsement by the 

U.S. Government.

Although this report is in the public domain, permission must be secured from the individual copyright owners to 

reproduce any copyrighted materials contained within this report.

Suggested citation:

Barbaro, J.R., Walter, D.A., and LeBlanc, D.R., 2013, Transport of nitrogen in a treated-wastewater plume to coastal 

discharge areas, Ashumet Valley, Cape Cod, Massachusetts: U.S. Geological Survey Scientific Investigations Report 

2013–5061, 37 p., http://pubs.usgs.gov/sir/2013/5061/.

ISBN 978-1-4113-3591-2
iii

Contents

Abstract...........................................................................................................................................................1

Introduction.....................................................................................................................................................1

Purpose and Scope ..............................................................................................................................3

Hydrogeologic Setting .........................................................................................................................3

Description of the Treated-Wastewater Plume........................................................................................4

Data Collection and Analysis..............................................................................................................4

Development of Gridded Concentration Data..................................................................................4

Plume Mass .........................................................................................................................................13

Plume Mass Balance .........................................................................................................................14

Reactive Transport of Nitrogen in the Treated-Wastewater Plume...........................................16

Nitrate .........................................................................................................................................16

Ammonium ..................................................................................................................................17

Simulation of Nitrogen Transport..............................................................................................................20

Regional Model Design.............................................................................................................20

Mass-Loaded Particle Tracking ..............................................................................................22

Nitrogen Loads to Receptors............................................................................................................24

Nitrate Loads ..............................................................................................................................24

Ammonium Loads.......................................................................................................................26

Total Nitrogen Loads .................................................................................................................26

Comparison of Nitrogen Loads From the Treated-Wastewater Plume to Nitrogen

Loads From Nonpoint Sources..............................................................................................27

Limitations of Modeling Analysis ..............................................................................................................30

Summary........................................................................................................................................................33

References Cited..........................................................................................................................................34

Figures

1. Map showing the study area, including the locations of the Massachusetts Military

Reservation, the treated-wastewater infiltration beds, Ashumet Pond, the Ashumet

Valley, and rivers and coastal embayments, western Cape Cod, Massachusetts............2

2. Map showing locations of 2007 sampling sites used in the plume-mapping analysis

and sections E–E’, F–F’, and Z–Z’, Ashumet Valley, Cape Cod, Massachusetts ................5

3. Graphs showing vertical nitrate profiles at well cluster F254 for A, sample

concentrations and B, interpolated concentrations, Ashumet Valley, Cape Cod,

Massachusetts..............................................................................................................................6

4. Graphs showing vertical profiles of dissolved oxygen, nitrate, ammonium, and

chloride concentrations in 2007 at sampling sites A, F262, B, F350, C, F271, D, D215,

and E, D222, Ashumet Valley, Cape Cod, Massachusetts......................................................7

5. Map and sections showing distribution of nitrate in 2007. A, Plan view at 16.8

meters below sea level; B, longitudinal section Z–Z’; and C, transverse section

E–E’, Ashumet Valley, Cape Cod, Massachusetts...................................................................9

6. Map and sections showing distribution of ammonium in 2007. A, Plan view at 7.6

meters below sea level; B, longitudinal section Z–Z’; and C, transverse section F–F’,

Ashumet Valley, Cape Cod, Massachusetts...........................................................................10
iv

7. Map and sections showing distribution of chloride in 2007. A, Plan view at 16.8

meters below sea level; B, longitudinal section Z–Z’; and C, transverse section

E–E’, Ashumet Valley, Cape Cod, Massachusetts.................................................................11

8. Map and sections showing distribution of boron in 2007. A, Plan view at 16.8 meters

below sea level; B, longitudinal section Z–Z’; and C, transverse section E–E’,

Ashumet Valley, Cape Cod, Massachusetts...........................................................................12

9. Graphs showing concentrations of A, nitrate compared to specific conductance,

and B, nitrate and ammonium compared to distance on section Z–Z’ in 1994 and

2007, Ashumet Valley, Cape Cod, Massachusetts.................................................................18

10. Graphs showing A, ammonium concentration in 1994 compared to ammonium

concentration in 2007, and B, difference in dissolved oxygen concentration

between 1994 and 2007 in relation to distance from the infiltration beds for the set

of groundwater samples (number of samples (n) = 356) analyzed for ammonium in

1994 and 2007, Ashumet Valley, Cape Cod, Massachusetts................................................20

11. Maps showing distribution of ammonium in A, 1994 and B, 2007 at altitudes of –22.9,

–16.8, –7.6, and +1.5 meters, Ashumet Valley, Cape Cod, Massachusetts........................21

12. Map showing rivers and embayments (major receptors) that receive most of the

simulated discharge of nitrogen in the treated-wastewater plume, Ashumet Valley,

Cape Cod, Massachusetts ........................................................................................................23

13. Map showing three-dimensional steady-state particle tracks originating from the

2007 nitrate plume, Ashumet Valley, Cape Cod, Massachusetts........................................25

14. Graphs showing simulated discharge of nitrate in the 2007 treated-wastewater

plume to major receptors, Ashumet Valley, Cape Cod, Massachusetts. A,

Cumulative nitrate-nitrogen mass, and B, nitrate-nitrogen loads ......................................27

15. Map showing three-dimensional steady-state particle tracks originating from the

2007 ammonium plume, Ashumet Valley, Cape Cod, Massachusetts................................28

16. Graphs showing simulated discharge of ammonium in the 2007 treated-wastewater

plume to major receptors, Ashumet Valley, Cape Cod, Massachusetts. A,

Cumulative ammonium-N mass, and B, ammonium-nitrogen loads ..................................29

17. Graph showing simulated discharge of total nitrogen in the 2007

treated-wastewater plume to major receptors, Ashumet Valley, Cape Cod,

Massachusetts............................................................................................................................30

18. Map showing generalized land use near the treated-wastewater plume, Ashumet

Valley, Cape Cod, Massachusetts............................................................................................31

Tables

1. Mass of nitrate, ammonium, chloride, and boron in the treated-wastewater plume

in 1994 and 2007, Ashumet Valley, Cape Cod, Massachusetts............................................14

2. Mean concentration and total mass of selected constituents in treated wastewater

discharged from1936 to 1995 from the Massachusetts Military Reservation

treatment plant, Ashumet Valley, Cape Cod, Massachusetts .............................................15

3. Median simulated steady-state traveltimes to major receptors for particles

originating from the 2007 nitrate and ammonium plumes, Ashumet Valley, Cape Cod,

Massachusetts............................................................................................................................26

4. Steady-state nonpoint-source and peak treated-wastewater plume nitrogen loads

to major receptors, Ashumet Valley, Cape Cod, Massachusetts .......................................32
v

Conversion Factors and Datum

SI to Inch/Pound

Multiply By To obtain

Length

centimeter (cm) 0.3937 inch (in.)

meter (m) 3.281 foot (ft)

kilometer (km) 0.6214 mile (mi)

Area

hectare (ha) 0.4047 acre

square kilometer (km2) 247.1 acre

square meter (m2) 10.76 square foot (ft2) 

square kilometer (km2) 0.3861 square mile (mi2)

Volume

liter (L) 0.2642 gallon (gal)

cubic meter (m3) 0.0002642 million gallons (Mgal)

cubic centimeter (cm3) 0.06102 cubic inch (in3)

liter (L) 61.02 cubic inch (in3)

cubic meter (m3) 35.31 cubic foot (ft3)

Flow rate

meter per day (m/d) 3.281 foot per day (ft/d)

cubic meter per second (m3/s) 35.31 cubic foot per second (ft3/s)

liter per second (L/s) 15.85 gallon per minute (gal/min)

cubic meter per day (m3/d) 264.2 gallon per day (gal/d)

Mass

gram (g) 0.03527 ounce, avoirdupois (oz)

kilogram (kg) 2.205 pound avoirdupois (lb)

Hydraulic conductivity

meter per day (m/d) 3.281 foot per day (ft/d)

Temperature in degrees Celsius (°C) may be converted to degrees Fahrenheit (°F) as follows:

°F = (1.8 × °C) + 32

Temperature in degrees Fahrenheit (°F) may be converted to degrees Celsius (°C) as follows:

°C = (°F – 32) / 1.8

Vertical coordinate information is referenced to the National Vertical Geodetic Datum of 1929 

(NVGD 29).

Horizontal coordinate information is referenced to the North American Datum of 1927 (NAD 27).

Altitude, as used in this report, refers to distance above the vertical datum.

Specific conductance is given in microsiemens per centimeter at 25 degrees Celsius (μS/cm at 

25°C).

Concentrations of chemical constituents in water are given either in milligrams per liter (mg/L) 

or micrograms per liter (μg/L).
THIS PAGE INTENTIONALLY LEFT BLANK
Transport of Nitrogen in a Treated-Wastewater Plume 

to Coastal Discharge Areas, Ashumet Valley, Cape Cod, 

Massachusetts

By Jeffrey R. Barbaro, Donald A. Walter, and Denis R. LeBlanc

Abstract

Land disposal of treated wastewater from a treatment plant on the Massachusetts Military Reservation in 

operation from 1936 to 1995 has created a plume of contaminated groundwater that is migrating toward coastal 

discharge areas in the town of Falmouth, Massachusetts. To 

develop a better understanding of the potential impact of the 

treated-wastewater plume on coastal discharge areas, the 

U.S. Geological Survey, in cooperation with the Air Force 

Center for Engineering and the Environment, evaluated the 

fate of nitrogen (N) in the plume. Groundwater samples from 

two large sampling events in 1994 and 2007 were used to map 

the size and location of the plume, calculate the masses of 

nitrate-N and ammonium-N, evaluate changes in mass since 

cessation of disposal in 1995, and create a gridded dataset 

suitable for use in nitrogen-transport simulations. In 2007, 

the treated-wastewater plume was about 1,200 meters (m) 

wide, 30 m thick, and 7,700 m long and contained approximately 87,000 kilograms (kg) nitrate-N and 31,600 kg total 

ammonium-N. An analysis of previous studies and data from 

the 1994 and 2007 sampling events suggests that most of 

the biologically reactive nitrogen in the plume in 2007 will 

be transported to coastal discharge areas as either nitrate or 

ammonium with relatively little transformation to an environmentally nonreactive end product such as nitrogen gas.

Nitrogen-transport simulations were conducted with a 

previously calibrated regional three-dimensional MODFLOW 

groundwater flow model. Mass-loaded particle tracking 

was used to simulate the advective transport of nitrogen to 

discharge areas (or receptors) along the coast. In the simulations, nonreactive transport (no mass loss in the aquifer) 

was assumed, providing an upper-end estimate of nitrogen 

loads to receptors. Simulations indicate that approximately 

95 percent of the nitrate-N and 99 percent of the ammoniumN in the wastewater plume will eventually discharge to the 

Coonamessett River, Backus River, Green Pond, and Bournes 

River. Approximately 76 percent of the total nitrate-N mass 

in the plume will discharge to these receptors within 100 

years of 2007; 90 and 94 percent will discharge within 200 

and 500 years, respectively. Nitrate loads will peak within 

about 50 years at all of the major receptors. The highest peak 

loads will occur at the Coonamessett River (450 kg per year 

(kg/yr) nitrate-N) and the Backus River (350 kg/yr nitrate-N). 

Because of adsorption, travel times are longer for ammonium than for nitrate; approximately 5 percent of the total 

ammonium-N mass in the plume will discharge to receptors 

within 100 years; 46 and 81 percent will discharge within 200 

and 500 years, respectively. The simulations indicate that the 

Coonamessett River will receive the largest cumulative nitrogen mass and the highest rate of discharge (load). Ongoing 

discharge to Ashumet Pond is relatively minor because most 

of the wastewater plume mass has already migrated downgradient from the pond.

To evaluate the contribution of the nitrogen loads from 

the treated-wastewater plume to total nitrogen loads to the 

discharge areas, the simulated treated-wastewater plume loads 

were compared to steady-state nonpoint-source loads calculated by the Massachusetts Estuaries Project for 2005. Simulation results indicate that the total nitrogen loads from the 

treated-wastewater plume are much lower than corresponding 

steady-state nonpoint-source loads from the watersheds; peak 

plume loads are equal to 11 percent or less of the nonpointsource loads.

Introduction

Land disposal of treated wastewater from a treatment 

plant on the Massachusetts Military Reservation (MMR) has 

created a plume of contaminated groundwater in the unconfined aquifer of Cape Cod, Massachusetts. Treated effluent 

containing phosphorous, nitrogen (N), detergents, metals, 

and other wastewater-derived constituents was discharged 

from 1936 to 1995 to rapid-infiltration beds near the southern 

boundary of the MMR (fig. 1). The disposal created a plume 

of wastewater-contaminated groundwater that extends south 

toward Ashumet Pond and the coastal embayments in the town 

of Falmouth, Massachusetts. In 2007, the treated-wastewater 

plume was about 1,200 meters (m) wide, 30 m thick, and 

7,700 m long and was generally aligned with the Ashumet Valley, a local topographic feature that extends from the western 
2 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

Base from U.S. Geological Survey digital data, 1:25,000, 1995

Massachusetts State Plane Coordinate System

North American Datum of 1927 0 0.5 1 MILE

0 0.5 1 1.5 2 KILOMETERS

Massachusetts

Military Reservation

FALMOUTH

MASHPEE

BOURNE

SANDWICH

Coonamessett

River

Backus

River

Childs

River

Bournes

River

Quashnet

River

Snake

Pond

Coonamessett

Pond

Johns

Pond

Ashumet

Pond

ATLANTIC OCEAN

Waquoit

Bay

Bournes

Pond

Eel

Pond

Great

Pond

Green

Pond

Mashpee

Pond

Wakeby

Pond

Mashpee

River

70°36' 70°34' 70°32' 70°30' 70°28'

41°40'

41°38'

41°36'

41°34'

Town Boundary

Open water

Cranberry bog

Massachusetts Military Reservation (MMR) boundary

Generalized boundary and orientation of the 

 Ashumet Valley (a relict glacial-outwash landform) 

EXPLANATION

Infiltration beds

MASSACHUSETTS

Study area

Cape 

Cod

ASHUMET VALLEY

ATLANTIC OCEAN

Figure 1. The study area, including the locations of the Massachusetts Military Reservation, the treated-wastewater infiltration beds, 

Ashumet Pond, the Ashumet Valley, and rivers and coastal embayments, western Cape Cod, Massachusetts.
Introduction 3

side of Ashumet Pond southward to Green Pond. Potential 

discharge areas for wastewater-contaminated groundwater 

include the freshwater rivers and ponds and associated saltwater coastal embayments along the coast in Falmouth (fig. 1), 

collectively referred to as coastal discharge areas. Although 

parts of the plume have been investigated extensively by the 

U.S. Geological Survey (USGS) and other researchers since 

the late 1970s, the three-dimensional (3D) distribution and 

transport of nitrogen at the plume scale (that is, the entire 

plume) have received relatively little study.

Nitrate and ammonium are the predominant nitrogen species in the plume, and unlike other nutrients such as phosphorous, they are relatively mobile in groundwater and have the 

potential to reach coastal discharge areas. Both nitrogen species also are reactive in the plume. Under hypoxic to anoxic 

conditions and in the presence of a suitable organic substrate, 

nitrate undergoes denitrification, the microbially mediated 

reduction of nitrate to nitrogen gas, which is not biologically 

reactive. Before cessation of wastewater disposal, denitrification was found to be an important process in a reactive zone 

near the infiltration beds (Smith and Duff, 1988; Smith and 

others, 1991). In contrast, ammonium undergoes nitrification, 

the microbially mediated oxidation of ammonium to nitrate 

that occurs in the presence of dissolved oxygen. Nitrification 

was found to occur mainly in the mixing zone at the upper 

boundary of the wastewater plume where both dissolved 

oxygen and ammonium are present (Böhlke and others, 2006; 

Smith and others, 2006). If post-cessation biogeochemical conditions are favorable for these reactions at the plume 

scale, then substantial loss of biologically reactive nitrogen 

may occur in the aquifer before the plume reaches coastal 

discharge areas.

The fate of nitrogen in the wastewater plume is of 

concern because the embayments in Falmouth have received 

increasing nitrogen loads in response to rapid population 

growth and development over the past several decades (Bowen 

and Valiela, 2001; Howes and others, 2005). Excessive nutrient enrichment (eutrophication) adversely affects water quality 

and aquatic ecosystems (Bowen and others, 2007). Elevated 

nitrogen concentrations are the result of increased wastewater 

disposal, atmospheric nitrogen deposition, and fertilizer application, as land use in coastal watersheds changes from forest 

and open, undeveloped land to residential and commercial 

development. Discharge from septic systems often is the largest source of nitrogen in these watersheds (Howes and others, 

2005; Cole and others, 2006; Kroeger and others, 2006a). The 

quantity and form of nitrogen reaching individual embayments depends on the density and spatial pattern of residential 

development in the embayment’s watershed (Kroeger and 

others, 2006a). In coming years, the MMR treated-wastewater 

plume will contribute to the nitrogen loading to these embayments, but the magnitude, spatial distribution, and timing of its 

discharge are not well understood.

To develop a better understanding of the the ptotential 

impact of the treated-wastewater plume on coastal discharge 

areas, the USGS, in cooperation with the Air Force Center for 

Engineering and the Environment, investigated the transport of 

nitrogen in the plume. The objectives of the study were to calculate the mass of nitrogen in the wastewater plume in 2007, 

evaluate the occurrence and extent of reactions that could 

reduce the mass of nitrogen in the aquifer before discharge, 

simulate the spatial and temporal discharge of nitrogen to the 

rivers and coastal embayments in Falmouth, and compare 

treated-wastewater plume loads to steady-state nonpointsource loads. The first two objectives were met by analyzing 

large groundwater datasets collected in 1994 and 2007 and 

reviewing results from previous studies on the reactivity of 

nitrogen in the plume. Nitrogen discharge was evaluated 

by simulating the movement of the plume with an existing 

groundwater flow model.

Purpose and Scope

The purpose of this report is to describe an investigation 

of the fate of nitrogen in the MMR treated-wastewater plume 

on Cape Cod, Massachusetts. Groundwater samples from two 

sample-collection events, referred to as the 1994 and 2007 

snapshots, were used to determine the size and location of the 

wastewater plume, calculate the masses of reactive nitrogen 

species and conservative constituents (chloride and boron) in 

the plume, and evaluate the reactivity of nitrogen species at 

the plume scale.

This report also describes nitrogen-transport simulations conducted with a previously calibrated 3D MODFLOW 

groundwater flow model and advective particle tracking. The 

report briefly describes the development of the groundwater 

flow model and details the results of nonreactive transport 

simulations conducted to estimate wastewater-plume nitrogen 

loads to coastal discharge areas. The report also compares 

the simulated wastewater-plume loads to published nonpointsource loads to these coastal discharge areas.

Hydrogeologic Setting

The unconfined aquifer of western Cape Cod consists of 

glacial drift deposited during the late Wisconsinan stage of the 

Pleistocene Epoch (Masterson and others, 1997). The study 

area is on an outwash plain deposited as a progradational 

deltaic sequence. Outwash-plain sediments fine downward and 

consist of stratified medium-to-coarse-grained glaciofluvial 

sand and gravel underlain by fine-to-medium-grained glaciolacustrine sand and silt. Ice-block collapse sedimentary structures and kettle ponds are common features in the outwash 

plain. The stratified drift is underlain in most places by a thin 

layer of basal till. The glacial deposits are underlain by crystalline bedrock, which is impermeable relative to the drift and 

forms the base of the unconfined aquifer. Regional groundwater flow on western Cape Cod has a radial pattern that extends 

outward from a circular groundwater mound near Snake Pond 

toward the surrounding coastlines (Walter and Masterson, 

2003). Areal recharge from precipitation is the major external 
4 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

source of water to the aquifer; the recharge rate on Cape Cod 

and southeastern Massachusetts is approximately 0.68 meters 

per year (m/yr; Masterson and others, 1998; Walter and 

Masterson, 2003; Masterson and others, 2009).

The aquifer is approximately 80 m thick in the study area. 

The altitude of the water table decreases from about 15 m 

above sea level at the infiltration beds to sea level at the coast, 

a distance of about 10 kilometers (km). The unsaturated zone 

is about 6 m thick beneath the infiltration beds. Groundwater 

flow is predominantly horizontal except near discharge boundaries such as Ashumet Pond and the coast where vertical flow 

is evident (McCobb and others, 2003). Horizontal groundwater flow velocities are in the range of 0.2 to 0.7 meters per day 

(m/d) (LeBlanc, 1984a).

Ashumet Pond, an 88-hectare, flow-through kettle pond, 

substantially affects the local groundwater-flow field and the 

wastewater plume originating from the infiltration beds, which 

are about 1 km to the northwest of the pond (fig. 1). Kettle 

ponds receive water by groundwater inflow and direct precipitation, and lose water by pond water outflow and evaporation. Because precipitation exceeds evaporation on Cape Cod, 

kettle ponds in this region are areas of net recharge (Walter 

and others, 2002). The flow-through nature of kettle ponds 

creates a spatial pattern in which groundwater flow converges 

toward ponds in upgradient areas and diverges away from 

them in downgradient areas. The extent of the influence on the 

groundwater-flow field is determined by the size and bathymetry of the pond, thickness of the aquifer, hydraulic conductivity of pond-bottom sediments, and temporal variability in 

precipitation and aquifer recharge (Walter and others, 2002). 

Under variable recharge conditions, simulated and measured 

horizontal groundwater-flow directions were found to vary by 

20 to 45 percent near Ashumet Pond (Walter and Masterson, 

2003; McCobb and others, 1999) as groundwater-flow patterns 

and the position of the hinge line (the line that separates the 

area of the pond receiving water from the area losing water) 

shift with fluctuating groundwater levels. Away from the pond, 

transient changes in the magnitudes of gradients and directions 

of groundwater flow are small (Walter and Masterson, 2003).

Both the spatial and temporal variability of the flow field 

near Ashumet Pond have implications for the fate of the wastewater plume. Reactive wastewater constituents such as nitrate 

and ammonium that discharge into the pond are likely attenuated by trophic processes in the pond. In contrast, conservative 

plume constituents such as chloride and boron are diluted and 

then discharged back into the aquifer in the pond water outflow. The discharged pond water, which may extend nearly to 

the bedrock, lies to the east of the treated-wastewater plume. 

In this study, the discharged pond water is not considered to 

be part of the wastewater plume. The temporal variability of 

the flow field near the pond (and the infiltration beds) may 

also affect the wastewater plume; flow simulations indicate 

that transient fluctuations in advective flow directions near 

the infiltration beds produce a wider wastewater plume than 

otherwise would develop in a steady-state flow field (Walter 

and Masterson, 2003).

Description of the Treated-Wastewater 

Plume

The size and location of the plume has been determined 

periodically by snapshot sampling, which is defined as a comprehensive sampling of the monitoring network over a short 

time interval of several months. As the most current snapshot 

for the plume, the 2007 snapshot was used to determine the 

size and location of the treated-wastewater plume and the 

total mass of nitrogen in the plume. The 1994 snapshot also 

was evaluated to compare changes in the plume between 1994 

(before cessation of wastewater disposal in 1995) and 2007. 

Nitrogen species (nitrate and ammonium), dissolved oxygen, 

conservative constituents (chloride and boron), and specific 

conductance were evaluated in detail for this study.

Data Collection and Analysis

Groundwater samples were collected by the USGS from 

observation-well clusters, multilevel samplers (MLSs), and 

direct-push borings. At nearly all sites, vertical concentration 

profiles were obtained, ranging from 2 samples in some well 

clusters to about 40 samples in the MLSs. The datasets from 

1994 and 2007 differ in the number and locations of sampled 

sites, but both are adequate for mapping the wastewater plume 

in three dimensions and estimating contaminant mass. For the 

1994 snapshot, approximately 760 samples were collected 

from 107 well-cluster and MLS sites (Savoie and LeBlanc, 

1998). For the 2007 snapshot, approximately 1,160 samples 

were collected from 154 well-cluster, MLS, and direct-push 

sites (Savoie and others, 2012). Descriptions of well installation, sampling techniques, analytical methods, quality-control 

procedures, and the tabulated data are provided in Savoie and 

LeBlanc (1998) and Savoie and others (2012).

Development of Gridded Concentration Data

Concentration data were interpolated onto a 3D grid to 

calculate contaminant mass in the plume in 1994 and 2007 

and create gridded concentration datasets suitable for use in 

nitrogen-transport simulations. Gridded datasets were created 

for nitrate-N, ammonium-N, chloride, boron, and specific 

conductance. To develop the gridded concentration data, the 

1994 and 2007 datasets were screened to remove samples 

from wells or MLSs in densely sampled parts of the plume 

that were deemed to be redundant, long-screened wells that 

were not part of a vertical cluster, and locations that were 

outside (generally above) the interpreted plume boundaries. 

The screened 1994 dataset consists of 534 samples from 70 

sampling sites, and the screened 2007 dataset consists of 1,005 

samples from 122 sampling sites (fig. 2). In 1994, samples 

from the MLSs were not analyzed for chloride, so the screened 

dataset for chloride consists of 236 samples from 62 sampling 

sites. Despite the size of the datasets, well locations were not 
Description of the Treated-Wastewater Plume 5

MILE

KILOMETERS

0 0.5

0 0.5 1 1.5

1

2

EXPLANATION

Multilevel sampler, well cluster, 

 or direct-push site and number

Open water or cranberry bog

Infiltration bed

Section line and identifier

70°34' 70°32'

41°38'

41°36'

D224

D223

D222 D221

D220

D219 D218

D217

D216

D215

S556

S421

S313

F695

F693

F694

F691

F687

F683

F680

F679

F678

F677

F676

F674

F670

F663

F660

F657 F655

F615

F613

F587

F585

F584

F582

F570

F508

F502

F501

F497

F493

F487

F484

F474

F472

F471

F470

F460

F459

F443

F442

F436

F433

F429

F239 F422

F418

F412

F411

F375

F359

F358

F357 F356

F355

F350

F294

F282

F279

F271

F267

F262

F254

F244

F236

F168

S434

Coonamessett

Pond

Johns

Pond

Ashumet

Pond

Coonamessett

River

Backus

River

Bournes

River

S521

S518

S517

S500 S499

S474

S473

S472

S471

S470

S469

S468 S436

S434

S344

S318

S316 S314

S313

F670

F661

F627

F624 F622

F621

F620

F619

F591 F618

F587

F586

F577

F575

F573

F572

F570

F567

F566 F565

F564

F512

F432

F424

F422

F388

F383

F379

F347

F343

F300

F239

F236

F230

MILES

KILOMETERS

0 0.05

0 0.1

0.1

0.2

Z

Z'

E

E'

F

F'

E E'

Base from U.S. Geological Survey digital data, 1:25,000, 1995

Massachusetts State Plane Coordinate System

North American Datum of 1927 

ASHUMET VALLEY

Figure 2. Locations of 2007 sampling sites used in the plume-mapping analysis and sections E–E’, F–F’, and Z–Z’, Ashumet Valley, 

Cape Cod, Massachusetts.
6 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

sufficiently spread out to completely define the size and shape 

of the wastewater plume. Consequently, control points and 

hydrologic judgment were used to fill in data gaps so that the 

locations of the upper, lower, and lateral boundaries of the 

wastewater plume could be delineated, as described below.

To develop the gridded datasets, each vertical concentration profile was first filled out by linearly interpolating concentrations between each sample pair in the profile. Interpolated 

profiles contained values at 0.30-m vertical intervals (fig. 3). 

To develop complete profiles, it was necessary to determine 

the locations of the upper and lower plume boundaries. The 

upper plume boundary typically was identified by the sharp 

dissolved oxygen, chloride, boron, specific conductance, and 

ammonium concentration gradients between the plume and 

the lens of uncontaminated groundwater above the plume that 

is produced by areal recharge. For example, detailed vertical concentration profiles obtained in 2007 indicate that the 

position of the upper boundary declines from an altitude of 

about –5 m at F262 (2,350 m from the infiltration beds) to 

about –14 m at D222 (5,800 m from the infiltration beds) as 

the recharge lens pushed the plume deeper into the aquifer 

with increasing distance from the source area (fig. 4). The high 

chloride concentrations above the top of the plume likely are 

the result of road salting along Mass. Route 151 (LeBlanc, 

1984a). The high nitrate concentrations in the recharge lens 

apparently are from nonpoint sources such as septic-system 

discharge and inorganic fertilizer application.

Because of a paucity of concentration data, the position 

of the lower boundary was more difficult to discern than the 

upper plume boundary; available data from deeper samples 

indicated that the plume was about 33 m thick. Consequently, 

it was assumed that the lower plume boundary was approximately 33 m below the upper plume boundary at all sampling 

sites. LeBlanc (1984a) shows the bottom of the plume coinciding with a lithologic boundary between permeable sand and 

gravel and less permeable fine-grained sediments. The altitude 

of the boundary rises from about –33 to –23 m below sea level 

over a distance extending 4,500 m downgradient from the 

infiltration beds (fig. 17; LeBlanc, 1984a). Beyond 4,500 m, 

lithologic data are not available to determine the altitude, 

thickness, and areal extent of the fine-grained layer. Because 

constituent concentrations appeared to be elevated below the 

altitude of the lithologic boundary in the 0- to 4,500-m interval (fig. 4) and the occurrence and location of the boundary 

farther downgradient is uncertain, the top of the fine-grained 

sediments generally does not coincide with the base of the 

plume in the current study. However, if these sediments are 

areally extensive in the Ashumet Valley, then they likely affect 

groundwater-flow paths and the vertical location of the plume 

downgradient from the infiltration beds.

Between the 1994 and 2007 snapshots, two substantial 

changes to the groundwater-flow system occurred that potentially affected the plume boundaries. Wastewater disposal 

ceased in 1995 and a remedial system designed to control 

chlorinated-solvent migration was installed in 1999 (Air Force 

Center for Engineering and the Environment, 2010). The 

-40

-30

-20

-10

0

10

0.0 0.5 1.0 1.5 2.0 2.5 3.0

Nitrate, in milligrams per liter as nitrogen

Altitude, in meters

A.

-40

-30

-20

-10

0

10

0.0 0.5 1.0 1.5 2.0 2.5 3.0

Top of plume: 2 meters

Background nitrate concentration

Bottom of plume: -35 meters

B.

Figure 3. Vertical nitrate profiles at well cluster F254 for A,

sample concentrations and B, interpolated concentrations, 

Ashumet Valley, Cape Cod, Massachusetts. Location of well 

cluster shown on figure 2.
Description of the Treated-Wastewater Plume 7

0 5 10 15 20 25

-40

-35

-30

-25

-20

-15

-10

-5

0

5

10

0 5 10 15 20 25

-40

-35

-30

-25

-20

-15

-10

-5

0

5

10

-40

-35

-30

-25

-20

-15

-10

-5

0

5

10

0 5 10 15 20 25

0 5 10 15 20 25

-40

-35

-30

-25

-20

-15

-10

-5

0

5

10

0 5 10 15 20 25 -40

-35

-30

-25

-20

-15

-10

-5

0

5

10

Concentration, in milligrams per liter 

Concentration, in milligrams per liter 

Altitude, in meters

Maximum 

chloride= 

76.3 mg/L

Maximum 

chloride= 

36.6 mg/L

Maximum 

chloride=

26.5 mg/L

Dissolved oxygen

Nitrate-nitrogen (N)

Ammonium-nitrogen (N)

Chloride

EXPLANATION

Top of plume

D222, 5,800

meters 

Sampling site and distance

 from infiltration beds

A F262, 2,350 meters B F350, 3,250 meters

C F271, 3,600 meters

E D222, 5,800 meters 

D D215, 4,600 meters 

Figure 4. Vertical profiles of dissolved oxygen, nitrate, ammonium, and chloride concentrations in 2007 at sampling sites A, F262, B, 

F350, C, F271, D, D215, and E, D222, Ashumet Valley, Cape Cod, Massachusetts. Locations of sampling sites shown on figure 2.
8 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

remedial system consisted of three extraction wells and two 

shallow infiltration trenches for treated water and was located 

near section E–E′ (fig. 2). Concentration data indicated that 

the position of the upper plume boundary was not appreciably 

affected by these changes to the flow system. Consequently, 

the same upper and lower boundary altitudes were used to 

delineate the treated-wastewater plume in 1994 and 2007.

At most sites, vertical concentration profiles included 

samples from the recharge lens above the plume; these samples were not used in the analysis. At other sites, concentration 

data and altitudes of the sampling points indicated that the 

uppermost sample in the profile was below the upper plume 

boundary. At nearly all sites, concentration data and altitudes 

indicated that the lowermost sample in the profile was above 

the lower plume boundary. Therefore, to calculate complete 

interpolated vertical concentration profiles, it was necessary 

to insert control points at the upper and lower plume boundaries at nearly all sampling sites in the 1994 and 2007 datasets. 

Control points were assigned concentrations representing 

background groundwater quality (0.35 milligrams per liter 

(mg/L) nitrate-N; 0 mg/L ammonium-N; 10 mg/L chloride; 

6 micrograms per liter (μg/L) boron; and 50 microsiemens 

per centimeter at 25 degrees Celsius (μS/cm) specific conductance. Background concentrations were consistent with values 

used in other studies of the wastewater plume and Cape Cod 

groundwater (Persky, 1986; Ceazan and others, 1989; 

Frimpter, 1990; Parkhurst and others, 2003; Kroeger and others, 2006b; Repert and others, 2006). At a few sites, control 

points also were added within the plume to fill in large vertical 

gaps between sampling sites and the interpreted plume boundaries. For samples with concentrations reported as less than 

the detection limit, the value of the detection limit was used 

for boron and chloride, whereas zero was used for nitrate and 

ammonium because these constituents may be fully consumed 

by reactions within the plume. Because there were relatively 

few of these samples and detection limits are low relative 

to concentrations in the plume (Savoie and LeBlanc, 1998; 

Savoie and others, 2012), the choice of zero or the detection 

limit had a small effect on nitrogen mass and distribution.

Nineteen horizontal layers from altitudes of –41.1 m 

(–135 ft) to +13.7 m (+45 ft) at 3-m (10-ft) vertical intervals 

were created from the interpolated vertical concentration 

profiles. Each layer contained a concentration value for each 

sampling site. For example, for the 2007 dataset, each layer 

contained 122 concentration values, although for layers near 

the top and bottom of the plume, many of the values were at 

background concentrations, indicating that the plume was not 

present at that location. Most of the data in a layer are interpolated values rather than actual sample concentrations because 

few well-screen or sampling-port altitudes coincide with layer 

altitudes. For each layer, the ArcGIS TOPOGRID procedure 

was used to interpolate the concentration data to an orthogonal 

grid. The interpolation procedure requires a bounding polygon 

to enclose the concentration data, which served as the lateral 

plume boundary. The bounding polygon size and shape for 

each layer was based on the concentration data at that altitude, 

and, to a lesser extent, on the simulated steady-state hydraulichead distribution. Because the spatial extents of the mobile 

constituents were similar, a single set of 19 bounding polygons 

was used for the nitrate, chloride, and boron plumes, whereas 

a second set was used for the smaller ammonium plume. 

Bounding polygons were assigned concentrations representing background groundwater quality. Thus, the plumes of the 

various constituents were defined as the volumes of the aquifer 

with concentrations generally above the constituents’ respective background levels. It should be noted that in the lowconcentration parts of the plume, concentrations may be lower 

than background concentrations.

The horizontally interpolated layers were stacked on 

top of each other to form the 3D depiction of the wastewater 

plume. Groundwater Modeling System (GMS) version 7.0 

(Aquaveo, 2012) software was used visualize the 3D contours 

of the plume and create vertical and horizontal cross-sections 

at selected locations (figs. 5–8). On the basis of the interpolated data, the wastewater plume, as defined by nitrate, chloride, and boron concentrations, was approximately 1,200 m 

wide, 30 m thick, and 7,700 m long in 2007 (1994 plume not 

shown). Consistent with a relatively slow migration rate, the 

ammonium plume was substantially smaller than the nitrate 

plume, extending roughly 3,300 m downgradient from the 

infiltration beds in 2007 (fig. 6). The spatial heterogeneity of 

the solutes is best represented near the infiltration beds and 

Ashumet Pond because the sample density is greatest in that 

part of the plume. The wastewater plume is substantially wider 

than the infiltration beds mainly because the steady-state flow 

field diverges downgradient from Ashumet Pond. Transient 

flow near Ashumet Pond may also cause lateral spreading 

closer to the beds.

The position of the upper boundary of the plume is 

relatively well defined, followed by the position of the lateral 

boundaries. The positions of the lower boundary and downgradient edge of the plume are less well defined. The downgradient extent of the plume, in particular, is poorly defined, 

and longitudinal dispersion creates a large mixing zone so that 

the mapped plume length will be dependent on the concentration used to differentiate contaminated from uncontaminated 

water. The farthest downgradient well cluster, F497, is about 

7 km from the infiltration beds and appears to be within the 

plume. The maximum downgradient extent of the wastewater plume in 2007 was assumed to be approximately 0.7 km 

beyond this well cluster (figs. 7 and 8). This plume length (7.7 

km) is roughly consistent with the results of a Darcy’s Law 

calculation: the 71-year (1936–2007) travel distance for the 

mean position of a contaminant front migrating at an average 

linear groundwater velocity of 0.3 m/d is 7.8 km. It should be 

noted, however, that if the average linear groundwater velocity was 0.4 m/d, a reasonable rate for the conductive upper 

part of the aquifer (LeBlanc, 1984a), then the mean position 

of the contaminant front would be more than 10 km from the 

source area.
Description of the Treated-Wastewater Plume 9

EXPLANATION

Open water or cranberry bog

Ground surface

Bedrock surface from Walter and Whealan (2005)

Water-table contour—Altitude of simulated

steady-state water table in feet above NGVD 29

shown in plan view (A) 

Base from U.S. Geological Survey digital data,

Massachusetts State Plane Coordinate System,

North American Datum of 1927,

National Geodetic Vertical Datum of 1929

Section line and identifier

Multilevel sampler, well cluster, or direct-push 

boring site (plan view)

Multilevel sampler or well cluster site and 

number (Sectional view. For clarity, 

selected sites on Z-Z’ and E-E’ were omitted)

F254

Nitrate concentration, in milligrams per liter

as nitrogen

 0.0 to 0.4

> 0.4 to 0.8

> 0.8 to 2.0

> 2.0 to 2.8 

> 2.8 to 4.4

> 4.4 

Approximate water table in 2007 shown in

 sectional view (B and C )

Infiltration bed

20x vertical exaggeration

30

20

10

 0

-10

-20

-30

-40

-50

-60

-70

-80

0 1,000 2,000 3,000 4,000 5,000 6,000 7,000 8,000 

Z Z' S421 F424 F254 F411 F262 F350 F271 F375 F655 F484 F497 Section Section E-E’ F-F’

Distance, in meters

Altitude, in meters 

Infiltration beds Sandwich

Road Carriage Shop

Road Old Barnstable

Road

Route

28

B.

Mass.

Route 151

A.

8

6

38

34

32

10

28

26

24

22

20

18

12

44

14

14

12

16

18

42

40

46

4

48

48

46

50

50

52 52

2

24

26

28

30

20

22

10

8

32

34

36

38

40

42

44

36

30

16

41°38'

41°36'

70°34' 70°32'

MILE

KILOMETERS

0 0.5 1

0 0.5 1 1.5 2

Z

Z'

E

E'

F

F'

F F'

30

20

10

 0

 0 500 1,000 1,500 2,000 

20x vertical exaggeration

Distance, in meters

 -10

 -20

 -30

 -40

 -50

 -60

 -70

 -80

Altitude, in meters

E E' F350 Section Z-Z’ F418 F282 F357 F359 F412

C

Figure 5. Distribution of nitrate in 2007. A, Plan view at 16.8 meters below sea level; B, longitudinal section Z–Z’; and C, transverse 

section E–E’, Ashumet Valley, Cape Cod, Massachusetts. Section locations shown in figure 2.
10 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

EXPLANATION

Ammonium concentration, in milligrams per liter

as nitrogen

Open water or cranberry bog

Ground surface

Bedrock surface from Walter and Whealan (2005)

Z Z' Section line and identifier

Multilevel sampler, well cluster, or direct-push 

boring site (plan view)

Multilevel sampler or well cluster site and 

number (Sectional view. For clarity, 

selected sites on Z-Z’ and E-E’ were omitted)

F254

0.0 to 0.4

> 0.4 to 0.8

> 0.8 to 2.0

> 2.0 to 2.8 

> 2.8 to 4.4

> 4.4 

Water-table contour—Altitude of simulated

steady-state water table in feet above NGVD 29

shown in plan view (A)

Base from U.S. Geological Survey digital data,

Massachusetts State Plane Coordinate System,

North American Datum of 1927, National Geodetic Vertical Datum of 1929

Approximate water table in 2007 shown in

 sectional view (B and C )

Infiltration bed

A.

8

6

38

34

36

32

30

10

28

12

14

16

44

14

16

42

40

26

18

22

24

20

18

46

4

48

52

50

50

2

24

26

28

30

12

10

8

20

22

32

34

36

38

40

42

44

46

48

KILOMETERS

1 MILE

1.5 2

0

0 0.5

0.5

1

Z

Z'

E

E'

F

F'

F279

Section  Z-Z’

30

F267F677F411F585

20

10

 0

 -10

 -20

 -30

 -40

 -50

 -60

 -70

 -80 Altitude, in meters

 0 500 1,000 1,500 

F F’

Distance, in meters

20x vertical exaggeration

C

30

20

10

 0

-10

-20

-30

-40

-50

-60

-70

-80

0 1,000 2,000 3,000 4,000 5,000 6,000 7,000 8,000 

Z

S421

F424F254F411F262F350F271F375F655F484F497

Section E-E’

Altitude, in meters

Carriage Shop

Road Old Barnstable

Road

Route

28

20x Vertical exaggeration

3,000

Z’

F350F271F375F655F484F497

Section F-F’

Distance, in meters

Infiltration beds Sandwich Road

Carriage Shop

Road Old Barnstable

Road

Route

28

B.

Mass.

Route 151

70°34' 70°32'

41°38'

41°36'

Figure 6. Distribution of ammonium in 2007. A, Plan view at 7.6 meters below sea level; B, longitudinal section Z–Z’; and C, transverse 

section F–F’, Ashumet Valley, Cape Cod, Massachusetts. Section locations shown in figure 2.
Description of the Treated-Wastewater Plume 11

32 30

26

18

24

22

20

A

0 0.5 1 1.5 2 KILOMETERS

0 0.5 1 MILE

8

6

38

34

36

32

30

10

28

12

14

16

44

14

16

42

40

26

18

22

24

20

18

46

4

48

52

52

50

50

2

24

26

28

30

12

10

8

20

22

32

34

36

38

40

42

44

46

48

30

20

10

 0

 0 500 1,000 1,500 2,000 

20x vertical exaggeration

Distance, in meters

 -10

 -20

 -30

 -40

 -50

 -60

 -70

 -80 Altitude, in meters

E E' F418 F282 F350 F357 F359 F412

C

20x vertical exaggeration

30

20

10

 0

-10

-20

-30

-40

-50

-60

-70

-80

0 1,000 2,000 3,000 4,000 5,000 6,000 7,000 8,000 

Z Z' S421 F424 F254 F411 F262 F350 F271 F375 F655 F484 F497 Section F-F’ Section E-E’

Distance, in meters

Altitude, in meters

Infiltration

beds Sandwich

Road

Carriage Shop

Road Old Barnstable

Road

Route

28

B

Mass.

Route 151

EXPLANATION

Chloride concentration, in milligrams per liter

 6.0 to 12.0

>12.0 to 18.0

> 18.0 to 30.0

> 30.0 to 42.0

> 42.0 to 78.0

> 78.0

Base from U.S. Geological Survey digital data,

Massachusetts State Plane Coordinate System,

North American Datum of 1927,

National Geodetic Vertical Datum of 1929

Open water or cranberry bog

Ground surface

Bedrock surface from Walter and Whealan (2005)

Section line and identifier

F254

Z

Z'

E

E'

F

F'

70°34' 70°32'

41°38'

41°36'

Z Z'

Multilevel sampler, well cluster, or direct-push 

boring site (plan view)

Multilevel sampler or well cluster site and 

number (Sectional view. For clarity, 

selected sites on Z-Z’ and E-E’ were omitted)

Water-table contour—Altitude of simulated

steady-state water table in feet above NGVD 29

shown in plan view (A)

Approximate water table in 2007 shown in

 sectional view (B and C )

Infiltration bed

Section  Z-Z’

Figure 7. Distribution of chloride in 2007. A, Plan view at 16.8 meters below sea level; B, longitudinal section Z–Z’; and C, transverse 

section E–E’, Ashumet Valley, Cape Cod, Massachusetts. Section locations shown in figure 2.
12 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

! !

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

! ! ! !

!

!

!

! !

!

!

!

!

!

!

!

!

!

!

!

!

! ! !

!

! ! !

!

!

!

!

!

!

!

!

!

!

!

!

!

!

!

! !

!

!

!

!

! ! !! !!

!

!!!

!

! !

!

!

!

!

! !

!

! !

!

!

A

6

28

26

4

2

8 8

10

24

22

20

18

12

14

14

12

16

18

24

26

28

30

20

22

10

16

46

48

50

52

32

34

36

38

40

42

44

38

34

32

44

42

40

48

46

50

52

36

30

30

20

10

 0

 0 500 1,000 1,500 2,000 

20x vertical exaggeration

Distance, in meters

 -10

 -20

 -30

 -40

 -50

 -60

 -70

 -80

Altitude, in meters

F418F282F350F357F359F412

C

E

Section  Z-Z’

E'

20x vertical exaggeration

30

20

10

0

-10

-20

-30

-40

-50

-60

-70

-80

0 1,000 2,000 3,000 4,000 5,000 6,000 7,000 8,000 

Z Z' S421 F424 F254 F411 F262 F350 F271 F375 F655 F484 F497 Section F-F’ Section E-E’

Distance, in meters

Altitude, in meters

Infiltration

beds Sandwich

Road

Mass.

Route 151

Carriage Shop

Road Old Barnstable

Road

Route

28

EXPLANATION

Boron concentration, in micrograms per liter

 6 to 24

> 24 to 42

> 42 to 78

> 78 to 114

> 114 to 186 

> 186 

B

Bedrock surface from Walter and Whealan (2005)

Open water or cranberry bog

Ground surface

Z Z’ Section line and identifier

!F254

Water-table contour—Altitude of simulated

steady-state water table in feet above NGVD 29

shown in plan view (A)

Multilevel sampler, well cluster, or direct-push 

boring site (plan view)

Multilevel sampler or well cluster site and 

number (Sectional view. For clarity, 

selected sites on Z-Z’ and E-E’ were omitted)

Base from U.S. Geological Survey digital data,

Massachusetts State Plane Coordinate System,

North American Datum of 1927,

National Geodetic Vertical Datum of 1929

Approximate water table in 2007 shown in

 sectional view (B and C )

Infiltration bed

Z

Z'

E

E'

F

F'

41°38'

41°36'

70°34' 70°32'

MILE

KILOMETERS

0 0.5 1

0 0.5 1 1.5 2

Figure 8. Distribution of boron in 2007. A, Plan view at 16.8 meters below sea level; B, longitudinal section Z–Z’; and C, transverse 

section E–E’, Ashumet Valley, Cape Cod, Massachusetts. Section locations shown in figure 2.
Description of the Treated-Wastewater Plume 13

Plume Mass

Gridded concentration data were used to calculate the 

total mass of nitrate-N, ammonium-N, boron and chloride in 

the treated-wastewater plume in 1994 and 2007. Masses of 

other nitrogen species, such as nitrite and organic N, were not 

calculated. In 2007, the numbers of samples for nitrite and 

organic N were insufficient to calculate the total mass in the 

plume. In 1994, the numbers of samples were sufficient, but 

concentrations of both constituents in the plume were relatively low. Nitrite-N concentrations were below the detection 

limit in most samples, with a maximum value of 0.87 mg/L 

at F566 (Savoie and LeBlanc, 1998). Organic-N was detected 

mainly near the source area (within 1,500 m of the infiltration 

beds). Concentrations typically were less than about 0.5 mg/L 

(Savoie and LeBlanc, 1998).

Mass was calculated for each horizontally interpolated 

layer. Total mass then was obtained by summing the results for 

the layers. The dissolved-phase mass in a horizontally interpolated layer was calculated from

(1)

where

mz is the constituent mass in layer z

(in kilograms),

	 θ	 is the aquifer porosity (dimensionless),

V is the total aquifer volume containing solute 

mass in layer z (in cubic meters),

A is a unit conversion factor,

Ci is the interpolated constituent concentration 

for grid cell i (in milligrams per liter), and

n is the number of grid cells in layer z.

Use of this equation assumes that porosity does not vary 

spatially in the aquifer, and that grid-cell dimensions and layer 

thickness are uniform. To compute the total aquifer volume 

containing solute mass for layer z, the total plume area in 

the layer was multiplied by 3 m, the distance between the 

horizontally interpolated layers. Interpolated concentrations 

were assumed to be centered in the 3-m vertical interval; for 

example, concentrations in the –44.2-m layer represent the 

–45.7- to –42.7-m interval in the aquifer. The total mass then 

was obtained by summing the masses for the set of horizontal 

layers that compose the wastewater plume. For mass estimation, interpolated concentrations were computed for 122-m 

(400-ft) by 122-m (400-ft) grid cells to be consistent with the 

discretization of the regional model grid. Possibly because 

of discretization effects, plume-mass results were found 

to be dependent on the size of grid cell; mass increased by 

approximately 12 percent as grid-cell length increased from 

3 to 152 m.

The above approach is appropriate for calculating mass 

in the dissolved phase. Mass is calculated from interpolated 

concentrations, which are assumed to be constant in the pore 

m V z A Ci

i

n

= =

θ ∑1

 ,

space of the block of aquifer material surrounding the value. 

For nitrate and conservative constituents (boron, and chloride), adsorption is negligible and dissolved phase mass is 

equivalent to total mass. Ammonium, however, undergoes 

adsorption by cation exchange (Ceazan and others, 1989), 

and a second equation is required to calculate adsorbed mass. 

The adsorbed-phase ammonium mass in a horizontally interpolated layer was calculated from

(2)

where

	 ρb is the dry bulk density of aquifer sediment 

(grams per cubic centimeter of porous 

medium), and

Kd is the adorption isotherm distribution 

coefficient (milliliters per gram of dry 

sediment).

Use of this equation assumes that both Kd

 and ρb do not vary 

spatially and that dissolved- and adsorbed-phase ammonium 

are in equilibrium. Dry bulk density was calculated from 

ρb

 = ρs (1 – θ), where ρs is the aquifer solid density (in grams 

per cubic centimeter solid). Using ρs

 = 2.65 g/cm3 for quartz 

and θ = 0.35, ρb

 = 1.72 g/cm3. The value of the distribution 

coefficient (Kd

 = 0.34 mL/g) was determined from sediment 

batch experiments (Ceazan and others, 1989). The ammonium 

sorption isotherm was linear over an aqueous concentration 

range that is greater than the range observed in the aquifer.

In 2007, the wastewater plume contained an estimated 

118,600 kg of nitrogen, consisting of 87,000 kg of nitrate-N 

and 31,600 kg of total ammonium-N; 2,900 kg of boron; and 

1,513,500 kg of chloride (table 1). To estimate changes in 

plume mass for 1994–2007, mass calculations were done as 

before except that only the set of wells and MLSs common 

to both sampling events were used in the analysis (table 1). 

This reduced the number of sampling sites to 62 for nitrate, 

ammonium, and boron, and 54 for chloride. The 1994 and 

2007 sample datasets developed in this manner were not identical because, although the sampling sites were common to 

both datasets, the numbers of samples composing the vertical 

concentration profiles at some of the sites differed by sampling 

event. Results indicate that the mass of nitrate-N, total ammonium-N, and boron declined by 7.8, 37.7, and 41.3 percent, 

respectively, between 1994 and 2007 (table 1). In contrast, the 

mass of chloride increased by 17.9 percent.

One possible explanation for the increase in the mass of 

chloride is the lateral migration of chloride from upgradient 

sources (such as road salt) into the wastewater plume under 

the new groundwater flow conditions that developed near the 

infiltration beds after cessation of treated-wastewater disposal 

in 1995. The occurrence of high concentrations of chloride 

and low concentrations of other wastewater constituents 

(boron and nitrate) immediately below the infiltration beds is 

consistent with this explanation (figs. 5B, 7B, and 8B). The 

decrease in the mass of boron between 1994 and 2007 is not 

m K VA C , z b d i

i

n

= =

ρ ∑1
14 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

Table 1. Mass of nitrate, ammonium, chloride, and boron in the treated-wastewater plume in 1994 and 2007, Ashumet Valley, Cape 

Cod, Massachusetts.

[kg, kilogram; N, nitrogen; na, not applicable; --, not determined; Common sites are the subset of sampling sites that were sampled for both the 1994 and 

2007 snapshots. Plume masses calculated from the common sites were used to compare changes between 1994 and 2007]

All sites Common sites

Constituent 1994 plume 

mass (kg)

2007 plume mass 

(kg)

1994 plume mass 

(kg)

2007 plume mass 

(kg)

Percent change 

from 1994

Number of sites for nitrate,

 ammonium, and boron

70 122 62 62 na

Number of sites for chloride 62 122 54 54 na

Nitrate-N 90,000 87,000 87,700 80,900 -7.8

Total Ammonium-N 53,400 31,600 48,000 29,900 -37.7

 (Dissolved) 20,000 11,800 18,000 11,200 --

 (Adsorbed) 33,400 19,800 30,000 18,700 --

Boron 4,800 2,900 4,600 2,700 -41.3

Chloride 1,279,300 1,513,500 1,258,300 1,483,700 17.9

well understood. The decreases in the masses of nitrogen 

species are discussed in greater detail in the Reactive Transport of Nitrogen in the Treated-Wastewater Plume section of 

this report.

Plume Mass Balance

The total mass of nitrogen, dissolved organic carbon 

(DOC), chloride, and boron in the treated wastewater discharged to the aquifer during plant operation (table 2) was 

estimated by Repert and others (2006) from effluent samples 

collected between 1974 and 1995 (n = 3–8 samples) and the 

history of effluent discharge rates compiled by Parkhurst and 

others (2003). Total mass was calculated by multiplying the 

mean concentration in the effluent by the estimated total volume of water (45 × 109

 liters) discharged to the aquifer over 

the 60-year period of operation. Calculations of total mass in 

the treated wastewater based on samples collected only during 

the last 20 years of operation are considered uncertain because 

of several factors. Concentrations in effluent were quite variable for most constituents, as indicated by sample standard 

deviations (table 2). The human population at the MMR and 

associated wastewater disposal rates varied substantially over 

60 years of operation. Based on limited records, disposal 

rates were probably highest during 1941 to 1945 and 1950 to 

1970. Effluent composition during these two periods may have 

been different than during the later period when the effluent 

samples were collected. Concentrations of the nitrogen species likely were not constant over time because of changes in 

treatment-plant operation and flow rates. Ammonium levels 

may have been higher during the 1940s to 1960s because 

oxidation of ammonium to nitrate was less likely during this 

period than during the 1970s and later when wastewater was 

recycled through the plant, thus increasing the oxidation of 

ammonium before discharge. In addition, detergent formulations, household and personal-care products, and dietary habits 

changed over time, possibly affecting concentrations of boron 

and chloride. Despite the uncertainly, the estimated total mass 

in the treated effluent was used to conduct a mass balance 

at the plume scale to gain insight into processes affecting 

solute mass.

The total nitrogen mass in the treated effluent was calculated as the sum of nitrate-N, nitrite-N, ammonium-N, and 

organic N. In addition to an estimated 560,000 kg of nitrate-N 

and 210,000 kg of ammonium-N, the wastewater effluent also 

contained an estimated 154,000 kg of organic N and 13,000 kg 

of nitrite-N. Thus, approximately 937,000 kg of total nitrogen was distributed as 59.8 percent nitrate-N, 22.4 percent 

ammonium-N, 16.4 percent organic N, and 1.4 percent nitriteN in the treated-wastewater effluent. Comparison of tables 1 

and 2 indicates that the masses of nitrate-N and ammonium-N 

in the 2007 plume were about 15 percent of the estimated total 

masses in wastewater effluent released to the aquifer. However, biotransformation of the organic N may have produced 

additional ammonium and possibly nitrate in the aquifer after 

recharge. Smith and Duff (1988) concluded that because soil 

water and shallow groundwater had the same concentrations 

of nitrate and ammonium as the effluent, biological processes 

were not having an observable effect on these species in the 

unsaturated zone. This is consistent with the high effluent 

flow rates and relatively thin unsaturated zone (6 m) beneath 

the infiltration beds, which results in a relatively short residence time the unsaturated zone. If assimilation of nitrogen 

in biomass is neglected and all organic N in the effluent 

was transformed to ammonium, then total ammonium mass 
Description of the Treated-Wastewater Plume 15

Table 2. Mean concentration and total mass of selected constituents in treated wastewater discharged from1936 to 1995 

from the Massachusetts Military Reservation treatment plant, Ashumet Valley, Cape Cod, Massachusetts.

[mg/L, milligrams per liter; kg, kilograms; N, nitrogen; --, not available; Mean wastewater concentration based on 3 to 8 samples. Concentration data 

from D.A. Repert, U.S. Geological Survey, written commun., 2010]

Treated wastewater concentration (mg/L)

Constituent Mean Standard 

deviation

Number of 

samples

Total mass in treated 

wastewater, 1936–1995 

(kg)

Nitrate-N 12.2 5.4 8 560,000

Ammonium-N 4.6 3.7 8 210,000

Boron 0.45 0.097 5 20,000

Chloride 39 10.1 4 1,760,000

Nitrite-N 0.30 0.20 7 13,000

Organic-N 3.4 1.9 3 154,000

Dissolved organic carbon 13.3 5.9 5 603,000

Specific conductance 4071 64.7 5 --

1

Specific conductance units in microsiemens per centimeter at 25 degrees Celsius.

released to the aquifer over 60 years of operation would be 

364,000 kg and the mass of ammonium-N in the plume in 

2007 plume could be as low as 9 percent of the total organic N 

and ammonium-N mass that was discharged to the aquifer.

The masses of boron and chloride in the 2007 plume were 

14.5 and 86.0 percent, respectively, of total estimated masses 

in wastewater effluent. One possible explanation for the apparent large loss of boron, which is presumed to be conservative 

in the outwash aquifer (Barber and others, 1988; Repert and 

others, 2006), is that the total estimated mass in the wastewater effluent is uncertain, as discussed above. Alternatively, 

boron may have been attenuated by an unknown reaction, an 

explanation which also is consistent with the apparent boron 

mass loss between 1994 and 2007 (table 1).

A general mass-balance equation for any constituent in 

the wastewater plume expressed in terms of the mass in the 

aquifer is

(3)

where

ma is the total mass in aquifer (adsorbed plus 

dissolved),

me is the mass input to the aquifer from the 

wastewater effluent,

mpr is the mass produced in the aquifer by 

reaction, and

mr is the mass removed from the aquifer by 

reactions, pumping, and discharge.

The mr term expanded to represent losses from physical processes and reactions is

(4)

m m m m , a e = + pr r −

m m r l = +r w m m + +p c m ,

where

mlr is the mass removed from the aquifer by 

reactions,

mw is the mass removed by pumping wells,

mp is the mass removed by discharge to ponds 

(mainly Ashumet Pond), and

mc is the mass removed by discharge to all other 

groundwater discharge areas (rivers and 

coastal embayments).

Losses to Ashumet Pond (mp ) are treated separately because 

a part of the plume has been discharging to Ashumet Pond for 

decades. In contrast, it appears that relatively little mass has 

discharged to downgradient discharge areas (mc ). Based on 

locations of known irrigation and remediation withdrawals, it 

also appears likely that the plume has been largely unaffected 

by major withdrawals (mw ) over the seven-decade residence 

time in the aquifer. To develop simplified mass balance equations for the 2007 plume, mc

 and mw are assumed to be negligible. For conservative species such as chloride and boron, 

reactions can be omitted, and the simplified mass-balance 

equation is

(5)

For reactive constituents such as ammonium and nitrate, the 

simplified mass-balance equation is

(6)

Determining net mass loss in the wastewater plume is complex because nitrate and ammonium are involved in multiple, 

coupled reactions (see Reactive Transport of Nitrogen in the 

Treated-Wastewater Plume section).

m m a e =− mp .

m m a e = + m m pr − − lr mp .
16 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

Although the mass-balance approach provides a useful 

conceptual framework for determining the fate of nitrogen in 

the wastewater plume, most of the terms in the equations are 

either unknown or known with a high degree of uncertainty. 

Plume mass in 2007 is the best-known quantity. Mass input 

to the aquifer from wastewater effluent, mass discharged to 

Ashumet Pond, and mass produced and consumed in the 

aquifer are the least well understood components of the massbalance equation at the spatial (plume) and temporal (71-year) 

scales for this study. The mass flux to Ashumet Pond appears 

to be highly sensitive to the loading history of the disposal 

beds and groundwater-flow system near the pond (LeBlanc, 

1984b; Parkhurst and others, 2003). The flux can be computed 

with a groundwater flow model, but large uncertainties associated with the flow system and the loading of the wastewater 

beds will propagate to the flux estimate. The total mass of 

nitrogen produced or consumed by reactions at the plume 

scale probably cannot be reliably estimated for the entire 

period of operation of the treatment plant. For the coastal 

nitrogen loading analysis, however, a transient analysis of 

plume evolution and mass balance is not necessary. Rather, in 

this study, a steady-state model was used to transport the nitrogen known to be in the aquifer in 2007 to the coastal discharge 

areas without quantification of prior mass loss. Nonreactive 

transport (that is, no mass loss in the aquifer) provides an 

upper-end estimate of the amount of nitrogen discharge from 

the plume to the coast in the years following 2007.

Reactive Transport of Nitrogen in the TreatedWastewater Plume

Previous field and laboratory studies and the 1994 and 

2007 snapshot datasets were evaluated to develop conceptual 

models for reactive transport of nitrogen in the aquifer and 

determine the representativeness of the nonreactive-transport 

simulations used in this study. If the post-cessation reactivity of nitrogen is low, then most of the mass in the aquifer (in 

2007) will reach discharge areas, and nonreactive-transport 

simulations will provide a reasonable estimate of nitrogen 

loading. On the other hand, if mass loss during transport to 

discharge areas is more substantial, then the nonreactivetransport simulations will overestimate nitrogen loading and 

provide a worst case estimate. Comparison of the 1994 and 

2007 datasets also allows for an evaluation of biogeochemical 

changes that have taken place since cessation of treated-wastewater disposal in 1995.

Determining net mass loss in the treated-wastewater 

plume is complex because nitrate and ammonium may be 

involved in multiple, coupled reactions. Nitrate is potentially 

consumed by heterotrophic denitrification, autotrophic denitrification coupled to oxidation of reduced inorganic species 

(Korom, 1992), dissimilatory nitrate reduction to ammonium 

(DNRA), and anammox. Nitrate is potentially produced by 

nitrification. Ammonium is potentially consumed by nitrification and anammox and produced by ammonification of 

organic N. Anammox is a recently identified reaction in which 

ammonium is oxidized to nitrogen gas under anaerobic conditions (Mulder and others, 1995). Ammonium serves as the 

electron donor and nitrate or nitrite as the electron acceptors. 

Anammox was first identified in wastewater systems (Mulder 

and others, 1995; Jetten and others, 2001), and has since been 

observed in estuarine sediments (Rich and others, 2008), soils 

(Jetten, 1999) and groundwater (Clarke and others, 2008; 

Robertson and others, 2012).

Nitrate

Denitrification, the microbially mediated reduction of 

nitrate to nitrogen gas under anoxic conditions, is the most 

important reaction potentially affecting nitrate at the plume 

scale. DNRA has also been observed, but the reaction had 

only a small effect on nitrate transport relative to denitrification (Smith and others, 1991). Autotrophic denitrification 

potentially can occur, but the highly reducing conditions 

needed for the production of electron donors such as hydrogen 

gas, reduced iron, and sulfur do not appear to be prevalent at 

the plume scale. Recent studies have shown that anammox 

also occurs in the plume (R.L. Smith, U.S. Geological Survey, 

written commun., 2012). Thus, the potential exists for simultaneous conversion of nitrate and ammonium to nitrogen gas 

where these species co-occur, but the importance of the reaction at the plume scale is unclear.

The occurrence of denitrification within 300 m of the 

infiltration beds has been demonstrated in the laboratory 

and field (Smith and Duff, 1988; Smith and others, 1991; 

Smith and others, 1996; Smith and others, 2004). Because 

denitrifying bacteria commonly are heterotrophic and couple 

the oxidation of organic carbon to the reduction of nitrate to 

obtain energy for growth, the spatial distribution of the occurrence and rate of denitrification may be determined by the 

amount and quality of the DOC in the plume. Previous studies 

have shown that biodegradation of the labile organic carbon 

in the wastewater effluent occurred predominantly near the 

infiltration beds (within about 500 m), and that the organic 

compounds that serve as the electron donors for denitrification 

are increasingly recalcitrant with distance from the infiltration 

beds (Harvey and others, 1984; Thurman and others, 1986; 

Barber, 1992; Harvey and Barber, 1992). For example, Thurman and others (1986) estimated that only 15 to 30 percent of 

the DOC originally in the effluent was present in downgradient parts of the plume. Moreover, about 50 percent of the 

remaining DOC was composed of nonbiodegradable alkylbenzene sulfonate (ABS) detergent. A comparison of effluent 

DOC concentrations with groundwater DOC concentrations 

near the infiltration beds also provides evidence for rapid 

uptake (biodegradation or sorption) of labile DOC. Before cessation of wastewater disposal, concentrations of DOC rarely 

exceeded about 4 mg/L near the infiltration beds (Harvey and 

Barber, 1992; Savoie and LeBlanc, 1998). In contrast, effluent DOC was about 12 mg/L (and possibly up to 20–25 mg/L 

in the past), suggesting that most of the DOC in the effluent 
Description of the Treated-Wastewater Plume 17

had been biodegraded in the unsaturated zone and groundwater adjacent to the beds, or most of the DOC was adsorbed 

to aquifer solids. In addition, Harvey and others (1984) and 

Harvey and Barber (1992) showed that the abundance of 

free-living bacteria and heterotrophic activity decreased with 

increasing distance from the infiltration beds, and these trends 

were attributed to the increasing recalcitrance of the wastewater-derived DOC in the plume. Based on the composition and 

spatial patterns of DOC and the distribution and abundance of 

bacterial biomass, denitrification may be electron donor (DOC 

and reduced inorganic constituents) limited in most of the 

treated-wastewater plume.

A comparison of longitudinal nitrate distributions within 

the mapped plume along section Z–Z′ before and after cessation of wastewater disposal also supports a conceptual model 

in which the rate of nitrate loss is very low or negligible 

beyond a reactive zone near the infiltration beds (fig. 9). In 

1994, during active wastewater disposal, nitrate concentrations 

were relatively high (ranging from 0 to nearly 30 mg/L–N) 

within 1,000 m of the infiltration beds (fig. 9). Within this 

same interval, specific conductance ranged from about 50 

to 550 μS/cm, and there was a weak linear relation between 

nitrate and specific conductance (R2 = 0.57; p<0.001; fig. 9A). 

In contrast, in the 1,000 to 6,500 m interval, nitrate concentrations in the plume were substantially less variable, ranging 

from about 1 to 4 mg/L–N, with a modest decrease in concentration with distance from the infiltration beds (fig. 9B). 

In 2007, 12 years after cessation of wastewater disposal, 

specific conductance and nitrate concentrations were low in 

the 0 to 1,000-m interval, and nitrate concentrations in this 

interval were similar to concentrations in downgradient parts 

of the plume (fig. 9). These longitudinal patterns indicate that, 

during active wastewater disposal, high nitrate concentrations (and labile DOC) were being injected into the aquifer 

and that nitrate was being reduced within a reactive zone that 

extended roughly 1,000 m downgradient from the infiltration 

beds to a residual concentration of 1 to 4 mg/L–N. Perhaps 

because the DOC was too recalcitrant to support continued 

denitrification, nitrate concentrations declined relatively little 

beyond the reactive zone. The most extensive denitrification in 

the reactive zone was in the core of the plume, leaving a shell 

of relatively elevated nitrate in the mixing zone at the plume 

boundaries (LeBlanc, 1984a; Repert and others, 2006). By 

2007, however, labile DOC and nitrate were no longer being 

discharged to the aquifer, and ongoing losses with distance 

from the source were not evident. Wastewater-derived DOC 

and nitrate had been transported roughly 1,300 m downgradient from the wastewater infiltration beds (12 years of flow at 

a horizontal groundwater flow velocity of 0.3 m/yr), and the 

nitrate in the 0- to 1,000-m interval probably was generated 

from oxidation of adsorbed ammonium and organic N near the 

beds (Repert and others, 2006).

To further evaluate spatial trends in the nitrate concentration data, the conservative-tracer method of Kroeger and 

others (2006b) also was applied to the data along the Z–Z′ 

transect. Mean specific conductance and nitrate concentrations 

in the effluent and background groundwater were used to correct for dilution before calculating percent nitrate loss in the 

plume. Results from this method show that losses of nitrate in 

samples from near the leading edge of the plume ranged from 

60 to 80 percent in comparison to effluent; however, most 

samples from close to the source had percent losses similar to 

samples from farther downgradient, so that there was no clear 

trend of ongoing loss as nitrate migrated downgradient (data 

not shown). Results from this method were consistent with the 

2007 nitrate concentration data on the Z–Z′ transect (fig. 9).

Because by 2007 nitrate concentrations throughout the 

plume were low and ongoing losses with distance from the 

source were not evident, it is reasonable to expect that the 

entire mass of nitrate in the 2007 plume will be transported 

to discharge areas. However, studies to measure denitrifying 

activity in distal parts of the plume have not been conducted; 

a low rate of ongoing denitrification or other nitrate removal 

reaction could have an appreciable effect on plume-scale 

nitrate loading to the coast because traveltimes range from 

decades to hundreds of years. Another possible location 

for mass loss is the groundwater/surface-water interface 

at the discharge location, where lithologic and geochemical changes occur (Böhlke and Denver, 1995; Kroeger and 

Charette, 2008). In this study, losses at the interface were 

not considered.

Ammonium

Nitrification, the microbially mediated oxidation of 

ammonium to nitrate in the presence of dissolved oxygen, 

appears to be the most important reaction affecting ammonium 

at the plume scale. Previous work at the site has demonstrated 

that nitrification occurs in the mixing zone at the upper boundary of the wastewater plume where both dissolved oxygen and 

ammonium are present (Böhlke and others, 2006; Smith and 

others, 2006; Miller and Smith, 2009). Ammonium loss in the 

anoxic core of the ammonium plume has not been observed. 

Because vertical transverse dispersion is a weak mixing 

process (Sudicky, 1986; Garabedian, and others, 1991), the 

rate of nitrification at the upper boundary is limited by the 

vertical transport of dissolved oxygen into the mixing zone, 

and consequently is less than the potential maximum rate for 

the microbial community in the aquifer (Böhlke and others, 

2006; Smith and others, 2006). On the basis of the limited size 

of the nitrification zone and low rate of reaction, relatively 

little ammonium is consumed by nitrification along the upper 

boundary of the plume. Previous work has also shown that 

geochemical changes resulting from cessation of wastewater 

disposal, in particular depletion of immobile electron donors 

and eventual downgradient migration of a dissolved oxygen 

front, may play an important role in determining the fate of 

the ammonium plume (Böhlke and others, 2006). If oxygen 

demand from the mass of immobile electron donors (reduced 

metals, organic carbon, and ammonium) in the aquifer was 

persistent, then the downgradient migration of the dissolved 

oxygen front would be slow relative to groundwater flow. 
18 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

A 

Distance from infiltration beds, in meters

B 2007

Ammonium-nitrogen

Nitrate-nitrogen

0

5

10

15

20

25

30

0 2,000 4,000 6,000 8,000

EXPLANATION

Ammonium-nitrogen

Nitrate-nitrogen

EXPLANATION

1994

Nitrate or ammonium, in milligrams per liter as nitrogen

0

5

10

15

20

25

30

0 2,000 4,000 6,000 8,000

0 200 400 600

 0 to 1,000 meters

> 3,000 to 7,200 meters

> 1,000 to 3,000 meters

0

5

10

15

20

25

30

2007

Distance from 

infiltration beds

EXPLANATION

0

5

10

15

20

25

30

Specific conductance, in microsiemens per centimeter at 25 degrees Celsius

Nitrate, in milligrams per liter as nitrogen

0 200 400 600

 0 to 1,000 meters

> 3,000 to 6,500 meters

> 1,000 to 3,000 meters

1994

Distance from 

infiltration beds

EXPLANATION

Figure 9. Concentrations of A, nitrate compared to specific conductance, and B, nitrate and ammonium compared to distance on 

section Z–Z’ in 1994 and 2007, Ashumet Valley, Cape Cod, Massachusetts. Location of section Z–Z’ shown on figure 2.
Description of the Treated-Wastewater Plume 19

Alternatively, if the oxygen demand was not persistent, then 

the dissolved oxygen front would begin migrating conservatively and eventually reach the back end of the ammonium 

plume. Extensive data collection near the infiltration beds indicates that oxygen demand in the core of the plume substantially slows the migration of the dissolved oxygen front. For 

example, by 2004, 9 years after cessation, the dissolved-oxygen front had not migrated far and remained near the downgradient edge of the disposal beds (Repert and others, 2006); 

consequently, lateral advection of dissolved oxygen into the 

slower moving ammonium plume was not occurring at this 

location. Taken together, previous studies and data collection 

indicate that ammonium loss has been minor in comparison to 

the total mass of ammonium in the plume.

In contrast, estimates of ammonium mass loss and 

changes in the distribution of ammonium and dissolved 

oxygen in the aquifer from evaluation of the 1994 and 2007 

plume snapshot data indicate that reoxygenation of the aquifer 

and ammonium mass loss are more extensive. As discussed 

previously, these plume-scale datasets indicate an apparent 

ammonium-N mass loss of 37.7 percent between 1994 and 

2007, or from 48,000 to 29,900 kg (table 1). This change 

corresponds to a zero-order (constant) rate of 1,400 kg/yr 

of total (adsorbed plus dissolved) ammonium-N. If this rate 

of loss were to continue, then the ammonium plume would 

persist for only 22 years and fail to reach a discharge boundary 

before complete removal from the aquifer. Assuming that such 

a large difference in plume mass is unlikely to be an artifact 

of the plume-mapping methodology or analytical differences 

between sampling events, post-cessation ammonium loss may 

be more extensive than previously recognized from analyses 

of more spatially limited datasets.

A scatterplot of samples from within the mapped plume 

boundaries, analyzed in both 1994 and 2007 (n = 356), demonstrates that the most extensive change in ammonium concentrations occurred in the back of the plume, nearest the infiltration beds (fig. 10A). Ammonium concentrations decreased at 

the back end of the plume (0–800 m), were roughly equivalent 

in the body of the plume (800–2,800 m), and increased at 

the leading edge of the plume (2,800–3,200 m). This spatial 

pattern is consistent with the downgradient movement of 

the plume; however, the measured decrease in plume mass 

between 1994 and 2007 indicates that mass loss has occurred 

in addition to simple displacement. Of the 235 samples in 

the 0- to 800-m interval that had ammonium concentrations 

of 0.1 mg/L–N (10 times the detection limit) or greater in 

1994, 100 percent had lower concentrations in 2007 than in 

1994. In addition, a plot of the difference in dissolved oxygen 

concentration with distance from the infiltration beds (samples 

were projected perpendicularly onto the plume centerline to 

determine distance from the beds) for the same dataset of 

sample pairs shows that dissolved oxygen concentrations in 

the 0- to 800-m interval increased between 1994 and 2007 

(fig. 10B). Dissolved oxygen concentrations in 2007 were 

equal to or greater than 1994 concentrations in 64 percent 

of the sample pairs (n = 269) in this spatial interval, and in 

31 percent of pairs (n = 87) farther downgradient in the plume. 

Increases in dissolved oxygen concentration at the leading edge of the plume (3,300 m from source, fig. 10B) may 

be due in part to reinjection of oxygenated water from the 

chlorinated-solvent remedial system that began operation in 

1999 (Air Force Center for Engineering and the Environment, 

2010). The system discharged treated water to two shallow 

infiltration trenches near the ends of section E–E′. Dissolved 

oxygen concentrations changed least and remained low in the 

core of the plume near the infiltration beds (data not shown). 

Plan-view distributions of the ammonium plume at selected 

depths also show a pattern of change between 1994 and 2007 

that is consistent with the scatter plots (fig. 11). Between 1994 

and 2007, the ammonium plume decreased in size between the 

infiltration beds and Ashumet Pond (fig. 11). Concentrations 

of ammonium in the interior of the plume (for example, in the 

thin, high-concentration layer at –7.6 m) also decreased, but 

the largest change in the plume was at the margins, where the 

plume width decreased at most depths. One possible explanation for these observations is that post-cessation downgradient 

migration of oxygenated groundwater provided appropriate 

conditions for nitrification in the back (excluding the core 

of the plume) and sides of the plume in addition to along the 

upper boundary.

Although the change in plume mass and spatial patterns 

in ammonium and dissolved oxygen concentrations are suggestive of ammonium oxidation during 1994 to 2007, these 

data do not provide direct evidence for widespread nitrification in the wastewater plume. Moreover, the nitrate-N mass 

that would have been produced by nitrification of 18,100 kg 

of ammonium-N was not evident in the difference between 

the 1994 and 2007 nitrate plume mass estimates (table 1) or 

the nitrate concentrations along section Z–Z′ (fig. 9B), which 

likely would have been elevated near the ammonium plume. 

Another possible reaction that may account for some of the 

apparent ammonium mass loss is anammox; in this case, both 

nitrate and ammonium would be consumed, and nitrate losses 

from anammox could partially offset nitrate gains from nitrification. The possible occurrence of multiple nitrogen producing 

and consuming reactions in different parts of the plume and 

at different times following cessation of wastewater disposal 

makes it difficult to determine the net effect of ammonium 

oxidation on the mass of nitrate. In addition, error related to 

plume-mapping procedures could have affected the calculation of nitrate-N and ammonium-N mass in the plume and the 

resulting differences in mass between 1994 and 2007. An additional plume snapshot for nitrogen species, DOC, dissolved 

oxygen, chloride, boron, and pH (nitrification and anammox 

produce acidity) within approximately 10 years from 2007 

would be useful for evaluating changes in nitrogen mass at the 

plume scale and developing a better understanding of the reactions responsible for the observed changes.
20 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

0

2

4

6

8

10

12

0 2 4 6 8 10 12

2007 ammonium,

in milligrams per liter as nitrogen

1994 ammonium, in milligrams per liter as nitrogen

A

 0 to 800 meters (n=269)

> 800 - 2,800 meters (n=40)

> 2,800 - 3,300 meters (n=47)

EXPLANATION

Distance from infiltration beds

-15

-10

-5

0

5

10

15

0 1,000 2,000 3,000

Distance from infiltration beds, in meters

Difference in dissolved oxgyen concentration 

between 2007 and 1994, in milligrams per liter

B

Line of equal 

concentration

Figure 10. A, ammonium concentration in 1994 compared to ammonium concentration in 2007, and B, difference in dissolved oxygen 

concentration between 1994 and 2007 in relation to distance from the infiltration beds for the set of groundwater samples (number of 

samples (n) = 356) analyzed for ammonium in 1994 and 2007, Ashumet Valley, Cape Cod, Massachusetts.

Simulation of Nitrogen Transport

Walter (2013) modified an existing regional model of 

western Cape Cod (Walter and Whealan, 2005) to conduct an 

analysis of nitrogen transport from surficial sources to ecological receptors in coastal areas on western Cape Cod. Because 

the scope of Walter (2013) is similar to the scope of this study 

and considers the same regional area, the Walter (2013) model 

was used to simulate the advective transport of nitrogen in the 

treated-wastewater plume to coastal discharge areas. To simulate the transport of plume nitrogen, additional modifications 

were made to the Walter (2013) model, including linearization 

of the model by converting water-table layers to confined layers, addition of layers to accommodate the gridded plume, and 

alteration of pumping rates and recharge. The “base” version 

of the regional model and the associated analyses are fully 

documented in Walter (2013); a synopsis of the development 

of this model is given below.

Groundwater flow was simulated by using the finitedifference model MODFLOW–2005 (Harbaugh, 2005). The 

USGS particle-tracking software program MODPATH was 

used to simulate the advective transport of nitrogen in the 

aquifer (Pollock, 1994). The USGS software application 

MODTOOLS (Orzol, 1997) was used to convert model output 

to georeferenced data layers.

Regional Model Design

The regional model encompasses the entire western end 

of Cape Cod (Walter, 2013). The model grid has 246 rows 

and 365 columns with a uniform horizontal discretization 

of 122 m (400 ft). The total active modeled area is 637 km2

. 

The modified regional model has 27 layers with thicknesses 

ranging from less than 3 m (10 ft) above sea level in parts of 

the aquifer to more than 30 m (100 ft) in the deepest part of 

the aquifer.

Coastal embayments, open coastal water, and rivers and 

streams are represented as head-dependent flux boundaries 

in the model. The General Head Boundary (GHB) and Drain 

(DRN) packages of MODFLOW–2000 (Harbaugh and others, 

2000; McDonald and Harbaugh, 1988) were used to simulate the open coastal waters, coastal embayments (saltwater 

estuaries), and some of the freshwater wetlands in the model 

domain. The Streamflow Routing Package (STR; Prudic, 

1989) was used to simulate other freshwater wetlands and 

streams; this package allows for an accounting of streamflow 

generated from groundwater discharge and for pumpinginduced streamflow depletion. Freshwater ponds were simulated as regions of high hydraulic conductivity (30,500 m/d) 

within the model and, as a result, groundwater flows through 

ponds under steady-state conditions.
Simulation of Nitrogen Transport 21

-22.9 meters 

-16.8 meters 

-7.6 meters

+1.5 meters

-22.9 meters 

-16.8 meters

-7.6 meters

+1.5 meters

MILE

KILOMETERS

0 0.5

0 0.5 1 1.5

1

2

Base from U.S. Geological Survey digital data,

Massachusetts State Plane Coordinate System,

North American Datum of 1927,

National Geodetic Vertical Datum of 1929

1994

2007

EXPLANATION

Ammonium concentration, 

in milligrams per liter as nitrogen

 0.0 to 0.4

> 0.4 to 0.8

> 0.8 to 2.0

> 2.0 to 2.8 

> 2.8 to 4.4

> 4.4 

Multilevel sampler, well cluster, 

or direct-push boring site

Open water or infiltration bed

A.

B.

Figure 11. Distribution of ammonium in A, 1994 and B, 2007 at altitudes of –22.9, –16.8, –7.6, and +1.5 meters, Ashumet Valley, Cape Cod, 

Massachusetts. 
22 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

Aquifer properties were estimated from lithologic 

logs, aquifer tests, and a depositional model of western 

Cape Cod (Masterson and others, 1997; Byron Stone, 

U.S. Geological Survey, written commun., 2002). Hydraulic 

conductivity varies spatially within layers and with depth 

and ranges from 107 m/d for sand and gravel to 3 m/d for silt 

and clay. Hydraulic conductivity in the Mashpee pitted plain 

outwash sediments generally decreases to the south and with 

depth (Walter and Whealan, 2005).

Recharge, which is the only source of water to the aquifer, was applied to the aquifer at a constant rate of 0.68 m/yr, 

a value derived from a long-term record of precipitation and 

temperature at Hatchville, Mass., and adjusted during model 

calibration (Walter and Whealan, 2005). Net recharge to 

ponds was 0.41 m/yr after accounting for pan evaporation 

(Farnsworth and others, 1982), and recharge to wetlands was 

zero because these areas function as net discharge areas in 

the groundwater-flow system. Recharge to the aquifer was 

adjusted further to account for septic-system return flow in 

unsewered residential areas that receive water from publicsupply systems and for point discharges of wastewater at existing wastewater treatment plants. Withdrawals of water from 

production wells totaling about 17.3 million gallons per day 

(Mgal/d) were simulated by use of the Well (WEL) package 

(McDonald and Harbaugh, 1988). One production well (Fresh 

Pond well) is near the observed path of the plume (fig. 12). 

The steady-state pumping rate of the well, as represented in 

the regional model, is about 0.6 Mgal/d.

The regional steady-state flow model was calibrated by 

adjusting initial estimates of recharge and aquifer properties to 

achieve an acceptable match between observed and simulated 

heads and streamflow. Details of the development, calibration, and limitations of the regional model are provided by 

Walter (2013).

Mass-Loaded Particle Tracking

Particle tracking was used to simulate the advective 

transport of nitrogen in the treated-wastewater plume to discharge areas along the coast. The particle-tracking algorithm 

in MODPATH uses model-calculated cell-by-cell flows to 

track the movement of particles of water through the aquifer. 

Particle tracking represents advective transport only. Other 

processes that can affect solute migration in groundwater—

namely dispersion, adsorption, and reactions—are not explicitly accounted for in a particle-tracking analysis. Neglecting 

these processes affects mass-loading estimates to discharge 

areas in different ways. Dispersion is a mechanical mixing 

process that spreads solute mass in transverse and longitudinal directions relative to the flow direction, thereby lowering concentrations at the plume margins. However, because 

dispersive mixing transverse to the direction of groundwater 

flow is a relatively weak process and longitudinal mixing only 

affects the leading and trailing edges of the plume, dispersion likely has only a minimal effect on the concentration and 

timing of nitrogen reaching a given receptor and, therefore, 

on the nitrogen load calculated by a particle-tracking analysis 

(Walter, 2007). Reversible adsorption, a process that affects 

ammonium, slows the rate of solute migration compared to the 

rate of groundwater flow. In contrast, nonsorbing species, such 

as nitrate, move at the same rate as groundwater. Neglecting 

adsorption in a particle-tracking analysis will affect the timing of mass loading (the solute will arrive at the receptor too 

quickly in the simulations), but total mass will be conserved. 

Biologically mediated mass loss of nitrogen in the aquifer 

potentially is the most consequential process. If nitrogen is 

biodegraded in the aquifer, then mass-loading rates calculated 

by a particle-tracking analysis would overestimate true massloading rates and represent a worst case scenario. Overall, the 

use of particle tracking to simulate nitrogen loading to coastal 

receptors is most appropriate for nonsorbing, nonreactive 

species such as nitrate; however, particle tracking also can be 

used to simulate the migration of a species such as ammonium 

that is affected by reversible adsorption.

The movement and discharge of nitrogen in the treatedwastewater plume were quantified by assigning a particle 

to each model cell that contained a nonzero interpolated 

concentration from the gridded plume. The number of particles released from the nitrate and ammonium plumes were 

6,169 and 695, respectively. Particles from the plumes then 

were transported to discharge areas under steady-state flow 

conditions. Separate simulations were conducted for nitrate 

and ammonium. Once a particle reached a surface-water 

discharge area, it was removed from the model domain. For 

the flow-through kettle ponds, this is equivalent to assuming 

that the nitrogen undergoes rapid uptake in the ponds and, 

consequently, does not reenter the aquifer with recharging 

pond water. The major coastal receptors receiving mass from 

the wastewater plume, represented by groups of MODFLOW 

model cells, are shown in figure 12. An accounting of the 

number and traveltime of particles reaching receptors was 

used to estimate the time history of nitrogen mass loading to 

these receptors. Particles were assigned a mass by multiplying 

the nitrogen concentration by the volume of water in the cell 

of origin. The spatial distribution of discharge within a given 

receptor was not determined; the model produced a curve 

of cumulative nitrate-N or ammonium-N mass in relation to 

time for the entire area of the receptor, irrespective of where 

the individual particles discharged to the river or embayment. 

Curves of cumulative total nitrogen mass in relation to time 

were calculated for the receptors by summing the curves of 

cumulative nitrate-N and ammonium-N. Nitrate-N, ammonium-N and total-N loads in kilograms per year were computed from smoothed (LOWESS, locally weighted scatterplot 

smoothing) curves of cumulative-mass. Limitations of the 

mass-loaded particle-tracking approach are discussed in detail 

in the Limitations of Modeling Analysis section.
Simulation of Nitrogen Transport 23

ATLANTIC OCEAN

Johns

Pond

Ashumet

Pond

Waquoit

Bay Eel

Pond

Coonamessett

Pond

Great

Pond

EXPLANATION

Receptor boundary and name

Open water or cranberry bog

Production well

Infiltration bed

0 0.5 1 MILE

0 0.5 1 1.5 2 KILOMETERS

Bournes River

Bournes Pond

Bournes Pond

Backus River

Green 

Pond

Fresh Pond well

70°36' 70°33' 70°30'

41°38'

41°36'

41°34'

Coonamessett River

FALMOUTH

MASHPEE

BOURNE

SANDWICH

Massachusetts

Military Reservation

Boundary

ASHUMET

VALLEY

Base from U.S. Geological Survey digital data, 1:25,000, 1995;

Massachusetts State Plane Coordinate System,

North American Datum of 1927, National Geodetic Vertical Datum of 1929 

Figure 12. Rivers and embayments (major receptors) that receive most of the simulated discharge of nitrogen in the treatedwastewater plume, Ashumet Valley, Cape Cod, Massachusetts. Major receptors are shown as groups of cells in the MODFLOW model 

grid.
24 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

Nitrogen Loads to Receptors

Particle-tracking simulations indicate that nearly all of 

the nitrogen in the 2007 wastewater plume will be transported 

to the coastal discharge areas near the leading edge of the 

plume: the Coonamessett River, Backus River, Green Pond, 

and Bournes River (fig. 12). The remainder will discharge 

mainly to Bournes Pond and Ashumet Pond, with minor 

quantities transported by deep flowpaths to Waquoit Bay and 

the Atlantic Ocean. The Coonamessett River, Backus River, 

and Bournes River are freshwater streams that derive most of 

their flow from groundwater and discharge to the headwaters 

of Great Pond, Green Pond, and Bournes Pond, respectively 

(fig. 12). Great Pond, Green Pond, Bournes Pond, Eel Pond, 

and Waquoit Bay (fig. 1) are brackish to saltwater coastal 

embayments that are separated from the Atlantic Ocean by 

barrier beaches except at connecting channels. These embayments receive direct groundwater discharge as well as surfacewater inflows from headwater streams. With the exception 

of Green and Bournes Ponds, the simulated mass of nitrogen 

discharging directly to the embayments was negligible. Cumulative mass and loads were computed for nitrate-N, ammonium-N, and total N (sum of nitrate-N and ammonium-N). 

The loads reported here assume that transport of nitrogen to 

the groundwater/surface-water interface results in discharge 

to surface water; biogeochemical processes near the interface 

(in the aquifer or subterranean estuary) that may reduce or 

eliminate nitrate and ammonium loads before discharge are 

not considered.

Nitrate Loads

Because particle tracking was used to simulate transport, nonreactive transport was assumed, and the total mass 

of nitrate-N in the treated-wastewater plume (87,000 kg) in 

2007 eventually reaches discharge areas. A lack of nitrate 

mass loss along flow paths is consistent with the conceptual 

model for nitrate reactivity that was developed from the 1994 

and 2007 plume snapshot data. By 2007, 12 years after cessation, concentrations in virtually the entire plume were near 

residual levels and ongoing denitrification that would reduce 

concentrations further did not appear to be occurring on a 

large spatial scale. However, a low rate of denitrification along 

groundwater flow paths or biotransformation at the groundwater/surface-water interface could reduce nitrate concentrations 

in the 2007 plume. Consequently, the mass-loaded particletracking provides an upper-end estimate (worst case scenario) 

of nitrate loads to the coastal discharge areas.

Groundwater flow generally transports the nitrate plume 

southward toward the Atlantic Ocean. The plan-view distribution of simulated 3D particle tracks originating from the 

cells with nonzero concentrations in the gridded nitrate plume 

is shown in figure 13. For reference, the plume outline at 

–16.8 m, an altitude at which the plume is areally extensive, 

also is shown. Some of the differences between the particle 

tracks and the plume outline are caused by juxtaposing the 3D 

particle tracks for the entire plume against a two-dimensional 

(2D) slice of the plume. However, as discussed in greater 

detail in the Limitations of Modeling Analysis section, most 

of the differences are related to the simulation of the surfacewater features in the model.

Simulations indicate that approximately 95 percent of 

the 2007 nitrate plume discharges to the Coonamessett River, 

Backus River, Green Pond, and Bournes River (hereafter 

referred to as the major receptors of nitrate). Bournes Pond 

and Ashumet Pond receive most of the remaining mass, with 

much smaller amounts discharging to Waquoit Bay and the 

Atlantic Ocean. Traveltimes for individual particles range 

from less than a year to more than 1,000 years, depending on 

the proximity of the receptor to the plume and the flow paths 

taken to reach the receptor. Median simulated traveltimes 

for transport from the wastewater plume to receptors ranges 

from 16 years for the Backus River to 69 years for Green 

Pond (table 3), which is farther from the plume and receives 

more mass from deeper flowpaths than the other receptors. 

Simulations indicate that the Fresh Pond well is in an area that 

receives discharged Ashumet Pond water, and, therefore, is 

likely not a receptor for wastewater nitrogen from the plume.

Simulations indicate that most of the nitrate in the 2007 

plume will discharge to the major receptors within 200 years 

of 2007 (fig. 14). The Coonamessett River receives the 

most nitrate-N mass (about 42 percent of the total nitrate-N 

mass), followed by the Bournes River, Green Pond, and the 

Backus River (fig. 14A). Simulated discharge to the Coonamessett and Backus Rivers, which are relatively close to the 

treated-wastewater plume, will occur mainly within 100 years 

from 2007. In contrast, discharge to the Bournes River and 

Green Pond will be more gradual. Future discharge to Ashumet Pond is minor because most of the nitrate-N mass in the 

treated-wastewater plume is already downgradient from the 

pond. Overall, approximately 76 percent of the total nitrateN mass in the plume will discharge to the major receptors 

within 100 years from 2007, 90 percent within 200 years, and 

94 percent within 500 years. The Coonamessett River will 

receive 47 percent (31,100 kg) of the total nitrate-N discharged to the major receptors within 100 years from 2007.

The timing and magnitude of simulated nitrate-N loads 

will vary among the major receptors (fig. 14B). Simulations 

indicate that loads will peak within about 50 years of 2007 

at all locations. The simulated load to the Backus River will 

peak at about 350 kg/yr nitrate-N relatively quickly (25 years 

from 2007) and then decline rapidly; the last particle will 

reach the Backus River approximately 150 years from 2007. 

The load to the Coonamessett River will peak at about 450 kg/

yr nitrate-N and then decline rapidly over the following 50 

years. In contrast, loads to Green Pond and the Bournes River 

will peak below 200 kg/yr nitrate-N and decline relatively 

slowly thereafter. With the exception of the Backus River, all 

of the major receptors receive water from deep flowpaths and, 
Simulation of Nitrogen Transport 25

50

44

46

48

50

52

52

54

36

32

30

28

22

24

26

20

14

16

10

12

6

4

2

34

36

8

18

34

38

40

42

44

46

48

38

40

42

32

30

28

26

24

22

20

18

8

6

4

2

16

14

12

10

2

54

Base from U.S. Geological Survey digital data,

Massachusetts State Plane Coodinate System,

North American Datum of 1927, National Geodetic Vertical Datum of 1929

0 0.5 1 1.5 2 KILOMETERS

0 0.5 1 MILE

70°36' 70°33' 70°30'

41°38'

41°36'

41°34'

Coonamessett

River

Backus

River

Bournes

River

Coonamessett

Pond

Ashumet

Pond

ATLANTIC OCEAN

Waquoit

Bay

Bournes

Pond

Eel

Pond Green

Pond

Fresh Pond well

EXPLANATION

Particle track

Open water or cranberry bog

Plume outline at -16.8 meters altitude

Water-table contour—Shows altitude of 

simulated steady-state water table,

in feet above NGVD 29

Production well 

Infiltration bed

Figure 13. Three-dimensional steady-state particle tracks originating from the 2007 nitrate plume, Ashumet Valley, 

Cape Cod, Massachusetts.
26 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

Table 3. Median simulated steady-state traveltimes to major 

receptors for particles originating from the 2007 nitrate and 

ammonium plumes, Ashumet Valley, Cape Cod, Massachusetts

[Location of receptors shown on figure 12. yr, years]

Nitrate plume Ammonium plume

Receptor

Median 

particle

traveltime 

(yr)

Number 

of

particles

Median 

particle 

traveltime 

(yr)

Number of

particles

Backus River 16 614 119 80

Bournes River 67 1,619 321 122

Coonamessett 

 River 41 2,520 171 302

Green Pond 69 1,010 452 176

consequently, will continue to receive low simulated nitrateN loads (1–10 kg/yr) from the treated-wastewater plume for 

hundreds of years (data not shown).

Ammonium Loads

As was the case for the nitrate simulation, nonreactive 

transport was assumed and the total mass of ammonium-N 

in the treated-wastewater plume in 2007 (31,600 kg; table 1) 

eventually reaches discharges areas. Because total ammoniumN mass consists of dissolved-phase and adsorbed-phase 

pools and an advective particle-tracking simulation cannot 

transport the adsorbed mass, an equivalent dissolved-phase 

concentration representing the total mass of ammonium-N 

was calculated from CT

 = Cdissolved + (Cdissolved × 1.6708), where 

CT

 is the equivalent dissolved-phase concentration, Cdissolved

is the dissolved-phase concentration , and 1.6708 is the ratio 

of adsorbed-phase mass to dissolved-phase mass per unit 

volume of the aquifer. The simulation assumes that no oxidation reactions occur and the entire plume migrates to coastal 

receptors as ammonium. Although such a simulation is not 

consistent with observed ammonium-N mass loss, it is not 

possible to simulate reactive-ammonium transport realistically with a mass-loaded particle-tracking analysis. Reactions 

could affect the timing and total mass of biologically reactive 

nitrogen reaching discharge areas. The present simulation can 

be viewed as an end-member transport simulation, reflecting 

maximum traveltimes to receptors and conservation of total 

ammonium mass in the 2007 plume.

To simulate ammonium transport, particle traveltimes 

were multiplied by the ammonium retardation factor (R) of 

2.5, resulting in particle migration that is slower than the 

rate of groundwater flow (v; that is, at a velocity of v/R) and, 

therefore, longer traveltimes to reach receptors. The retardation factor was calculated by use of the distribution coefficient 

(0.34 mL/g) determined from sediment batch experiments and 

is consistent with retardation factors determined from tracer 

experiments (Ceazan and others, 1989) and the position of the 

ammonium front in the aquifer (fig. 6).

The plan-view distribution of simulated 3D particle 

tracks originating from the cells with nonzero concentrations 

in the gridded ammonium plume is shown in figure 15. For 

reference, the plume outline at –7.6 m, an altitude at which the 

plume is areally extensive, also is shown. Similar to nitrate, 

some of the differences between the particle tracks and the 

plume outline are caused by juxtaposing the 3D particle tracks 

for the entire plume against a 2D slice of the plume. However, 

as discussed in greater detail in the Limitations of Modeling 

Analysis section, most of the differences are related to the 

simulation of the surface-water features in the model.

Simulations indicate that approximately 99 percent of 

the 2007 ammonium plume will discharge to the Coonamessett River, Backus River, Green Pond, and the Bournes River 

(hereafter referred to as the major receptors of ammonium). 

Ashumet Pond will receive negligible ammonium from the 

2007 plume. Traveltimes for individual particles range from 

about 40 to more than 1,000 years, depending on the proximity of the receptor to the plume and the flow paths taken to 

reach the receptor. Because of sorption and the position of the 

leading edge of the ammonium plume, median traveltimes to 

the major receptors are longer for ammonium than for nitrate, 

ranging from less than 120 years for the Backus River to 

450 years for Green Pond.

Simulations indicate that most of the ammonium 

plume will discharge to receptors within 400 years of 2007 

(fig. 16A). Similar to nitrate, the Coonamessett River will 

receive the most ammonium-N mass (about 45 percent of 

the total ammonium-N mass), followed by Green Pond, the 

Backus River, and the Bournes River. Overall, approximately 

5 percent of the total ammonium-N mass in the plume will 

discharge to receptors within 100 years, 46 percent within 

200 years, and 81 percent within 500 years. The discharge to 

Green Pond accounts for most of the discharge after about 

350 years.

The timing and magnitude of simulated ammonium 

loads varies among receptors (fig. 16B). The distribution of 

ammonium loads among receptors is similar to that of nitrate. 

Ammonium loads to the Backus River will peak at about 50 kg 

ammonium-N/yr relatively quickly (125 years from 2007) and 

then decline over the following 200 years. Loads to the Coonamessett River will peak at nearly 200 kg ammonium-N/yr 

approximately 200 years from 2007 and then decline rapidly 

over the following 100 years. Loads to Green Pond and the 

Bournes River will peak at about 20 kg ammonium-N/yr and 

remain elevated for several hundred years (fig. 16B).

Total Nitrogen Loads

Because total-N loads to most of the major receptors 

(Coonamessett River, Backus River, Green Pond, and Bournes 

River) are dominated by discharge of nitrate-N mass, the totalN load curves (fig. 17) are similar in shape and magnitude to 

nitrate-N load curves (fig. 14). An exception is the total-N load 
Simulation of Nitrogen Transport 27

Cumulative mass, in kilograms of nitrate-nitrogen

A.

0

10,000

20,000

30,000

40,000

0 50 100 150 200

0 50 100 150 200

Load, in kilograms per year of nitrate-nitrogen

B.

Green Pond

Coonamessett River

Backus River

Bournes River

EXPLANATION

0

100

200

300

400

500

Time since 2007, in years

Figure 14. Simulated discharge of nitrate in the 2007 treatedwastewater plume to major receptors, Ashumet Valley, Cape 

Cod, Massachusetts. A, Cumulative nitrate-nitrogen mass, and 

B, nitrate-nitrogen loads.

curve to the Coonamessett River, which is affected by substantial ammonium-N mass. At this location a secondary peak will 

arrive around 175 years from 2007.

The simulated load curves shown in figures 16 and 17 

include the transport of nonreactive ammonium as a reversibly 

adsorbing species. If nitrification were a dominant reaction 

and ammonium was being converted rapidly to nitrate relative 

to residence times in the aquifer, then traveltimes from the 

ammonium plume to receptors would decrease, peak loads 

would be higher, and loads would be spread over shorter 

time intervals compared to migration retarded by adsorption. 

Because geochemical conditions in the plume are considered to be unfavorable for denitrification, nitrogen mass in 

the ammonium plume likely would be conserved, and the 

most significant change would be in the timing and rate of 

total N discharge.

Simulations indicate that the Coonamessett River will 

receive the most cumulative nitrogen mass and the highest 

rates of discharge (loads) from the 2007 treated-wastewater 

plume. The Backus and Bournes Rivers and Green Pond also 

will receive substantial direct discharge of nitrogen, mostly 

as nitrate, from the plume. Surface-water inflows from the 

Backus, Bournes, and Coonamessett Rivers (which contain nitrogen from groundwater discharge) to Green Pond, 

Bournes Pond, and Great Pond, respectively, contribute to the 

total load of nitrogen to these embayments.

Comparison of Nitrogen Loads From the TreatedWastewater Plume to Nitrogen Loads From 

Nonpoint Sources

The coastal rivers and embayments in Falmouth receive 

nitrogen from sources other than the treated-wastewater 

plume, such as atmospheric deposition, fertilizer, and wastewater from septic systems. Septic systems and fertilizer 

applied to residential, commercial, agricultural, and recreational areas are the major sources of anthropogenic nitrogen 

in the coastal watersheds of Falmouth (Bowen and Valiela, 

2001; Howes and others, 2005). At the watershed scale, these 

numerous and widely distributed small point sources are 

typically referred to as nonpoint nitrogen sources. In contrast, 

larger, discrete sources such as the wastewater plume produced by the MMR wastewater treatment plant and landfills 

are considered separately as point sources. Nonpoint-source 

nitrogen is released to the ground surface as fertilizer or to the 

soil zone as wastewater effluent and conveyed to discharge 

areas by groundwater and surface-water flow. Nitrate is often 

the dominant form of nitrogen reaching the water table from 

wastewater releases (for example, DeSimone and Howes, 

1998), but under certain circumstances, such as when flow 

paths are short, a substantial portion of anthropogenic nitrogen reaching discharge areas may be in the form of dissolved 

organic nitrogen (Kroeger and others, 2006a).

In 2005, there was substantial residential development 

near the treated-wastewater plume (fig. 18). The 11 land-use 
28 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

50

50

52

52

44

44

40

42

40

42

34

32

32

30

30

36

36

28

28

22

22

20

20

24

24

26

26

14

16

16

18

18

10

10

12

12

6

4

6

2

2

8

8

2

4

34

14

38

38

46

46

48

48

54

54

Coonamessett

River

Backus

River

Bournes

River

Coonamessett

Pond

Ashumet

Pond

ATLANTIC OCEAN

Waquoit

Bay Bournes

Pond

Eel

Pond Green

Pond

Fresh Pond well

Base from U.S. Geological Survey digital data,

Massachusetts State Plane Coordinate System,

North American Datum of 1927, National Geodetic Vertical Datum of 1929

0 0.5 1 1.5 2 KILOMETERS

0 0.5 1 MILE

70°36' 70°33' 70°30'

41°38'

41°36'

41°34'

EXPLANATION

Particle track

Open water or cranberry bog

Plume outline at -7.6 meters altitude

Water-table contour—Shows altitude of 

simulated steady-state water table,

in feet above NGVD 29 

Production well 

Infiltration bed

Figure 15. Three-dimensional steady-state particle tracks originating from the 2007 ammonium plume, Ashumet Valley, Cape Cod, 

Massachusetts. 
Simulation of Nitrogen Transport 29

0

2,000

4,000

6,000

8,000

10,000

12,000

Cumulative mass, in kilograms of ammonium-nitrogen

0 100 200 300 400 500

A

0

50

100

150

200

0 100 200 300 400 500

Load, in kilograms per year of ammonium-nitrogen 

Time since 2007, in years

B

Green Pond

Coonamessett River

Backus River

Bournes River

EXPLANATION

Figure 16. Simulated discharge of ammonium in the 2007 

treated-wastewater plume to major receptors, Ashumet Valley, 

Cape Cod, Massachusetts. A, Cumulative ammonium-N mass, and 

B, ammonium-nitrogen loads.

categories shown in figure 18 were aggregated from 33 

land-use codes in the 2005 Massachusetts land-use data layer 

(Massachusetts Office of Geographic Information, 2010). The 

aggregated developed land-use categories include high-density 

residential (single-family houses on less than 1⁄4-acre lots and 

multifamily dwellings), medium-density residential (houses 

on 1⁄4- to 1⁄2-acre lots), low-density residential (houses on 1⁄2- to 

1-acre lots), very-low-density residential (rural properties with 

houses on greater than 1-acre lots), commercial-industrialtransportation (shopping centers, stores, medical offices, 

schools, churches, colleges, hospitals, municipal properties, 

industrial facilities, and major roads), agricultural (cropland, 

pasture, orchards, nurseries, and cranberry bogs), and open 

urban (cemeteries and participation recreational areas such 

as golf courses and athletic fields). Residential development 

is particularly dense near the coast. Because these areas are 

unsewered, wastewater is disposed to the ground through 

onsite septic systems.

To evaluate the contribution of the nitrogen loads from 

the treated-wastewater plume to the total nitrogen loads to the 

discharge areas, the simulated wastewater-plume loads were 

compared to nonpoint-source (watershed) loads calculated 

by the Massachusetts Estuaries Project (MEP) (Howes and 

others, 2005). As part of a study to determine nitrogen load 

thresholds for use in evaluating nitrogen management alternatives in the watersheds, the MEP calculated nonpoint-source 

nitrogen loads for the watersheds to the embayments in the 

town of Falmouth by using a linked watershed-embayment 

model. The linked model uses a nitrogen loading submodel 

to calculate nonpoint-source loads based on predetermined 

nitrogen loading rates for specific land-use types. The MEP 

study determined that most of the load (66 to 86 percent) 

occurred in parts of the watershed with relatively little recent 

land-use change and groundwater traveltimes of less than 

10 years. Consequently, it was assumed that nitrogen entering a watershed was approximately in balance with nitrogen 

discharging to embayments and temporally varying loads were 

not calculated.

Results of the MEP study (Howes and others, 2005; 

table IV–4) indicate that the 2005 nonpoint-source nitrogen loads were 16,866 kg/yr (46 kg/d) to the Coonamessett River; 3,210 kg/yr (9 kg/d) to the Backus River; 

7,356 kg/yr (20 kg/d) to Green Pond; and 2,583 kg/yr (7 kg/d) 

to the Bournes River (table 4). Total watershed loads are 

composed of contributions from natural background sources, 

lawn fertilizers, direct runoff from impervious surfaces, and 

wastewater from septic systems. Of these nonpoint sources, 

wastewater from septic systems was determined to be the 

largest contributor, accounting for up to 75 percent of the 

total nonpoint-source load. The MEP nonpoint-source loads 

discussed herein exclude the nitrogen load from the MMR 

treated-wastewater plume, which is estimated to be about 

1,030 kg/yr (Howes and others, 2005). The MEP wastewaterplume nitrogen load (1,030 kg/yr) is in agreement with the 

results of this study; simulations indicate that the sum of the 
30 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

0 100 200 300

Time since 2007, in years

0

100

200

300

400

500

600

Load, in kilograms per year of total nitrogen

Green Pond

EXPLANATION

Coonamessett River

Backus River

Bournes River

Figure 17. Simulated discharge of total nitrogen in the 2007 

treated-wastewater plume to major receptors, Ashumet Valley, 

Cape Cod, Massachusetts.

peak total-N loads to the major receptors will be about 1,250 

kg/yr over the next 100 years (fig. 17).

MEP steady-state nonpoint-source loads generally are 

much higher than total N wastewater-plume loads for the 

receptors receiving nitrogen from the plume. For the Coonamessett River, the simulated peak total-N wastewater-plume 

load of 550 kg/yr that occurs about 35 years from 2007 

(fig. 17) is 3.3 percent of the 16,866 kg/yr nonpoint-source 

load (table 4). Simulated wastewater-plume nitrogen loads 

before and after the peak are less than about 1 percent of the 

nonpoint-source loads. For the Backus River, the peak load of 

350 kg/yr that occurs about 25 years from 2007 is 10.9 percent 

of the 3,210 kg/yr nonpoint-source load. Simulated treatedwastewater plume nitrogen loads after the peak are less than 

about 2 percent of the nonpoint-source load. For Green Pond, 

the peak load of 150 kg/yr that occurs about 25 years from 

2007 is 2.0 percent of the 7,356 kg/yr nonpoint-source load. 

For the Bournes River, the peak load of 200 kg/yr that occurs 

about 25 years from 2007 is 7.7 percent of the 2,583 kg/yr 

nonpoint-source load. The simulated treated-wastewater 

plume load for the Bournes and Green River decreases gradually over 200 years to approximately 50 kg/yr, which is less 

than 2 percent of the nonpoint-source load to the respective 

receptors. Overall, the simulated total-nitrogen loads from the 

treated-wastewater plume are much lower than the corresponding steady-state nonpoint-source loads from the watershed; peak plume loads are equal to 11 percent or less of the 

nonpoint-source loads.

Limitations of Modeling Analysis

The modeling analysis in this study was done with a 

deterministic groundwater-flow model in which hydraulic conductivities, recharge, and boundary leakances were determined 

by trial-and-error calibration to observations of hydraulic head 

and streamflow. The model is assumed to represent the flow 

system adequately because the spatial distribution of hydraulic conductivity is based on an understanding of the depositional history of the glacial sediments in the region; values of 

hydraulic conductivity, boundary leakance, and recharge are 

reasonable and consistent with previous investigations; and 

simulation results reasonably match observations of heads and 

streamflows. However, the model produces a simulated flow 

field that is nonunique because different model configurations can produce similar head distributions. Flowlines in the 

aquifer also are strongly affected by hydraulic boundaries such 

as ponds, streams, and coastal waters, but because boundary 

geometries are static, simulated particle tracks likely would 

be similar for a range of model configurations. Despite these 

issues, if the model represents the flow system adequately and 

is calibrated to observed conditions, then simulated particle 

tracks are reasonably representative of advective flowpaths in 

the aquifer at a regional scale.

Simulated traveltimes also can be affected by model 

construction and parameterization. If hydraulic conductivity 

and recharge rates are changed in a correlated manner, then 

a similar calibrated hydraulic head distribution is produced, 

but groundwater fluxes and traveltimes change. For example, 

higher hydraulic conductivity and recharge rates that produce 

a reasonable calibrated head distribution would lead to higher 

groundwater fluxes and decreased traveltimes compared to 

the calibrated model used in the study. Although the values of 

these parameters are constrained to an extent by independent 

observations, they are often adjusted during model calibration 

to achieve the best model fit to observed heads and streamflows, and the final combination of values directly affects 

traveltimes in the aquifer. Another factor affecting traveltimes 

is aquifer porosity. A uniform value of 0.35 was used for these 

simulations, but the porosity of glacial sand and gravel can 

range from 0.25 and 0.40. Groundwater velocity is inversely 

proportional to porosity; decreasing porosity increases velocity 

and decreases traveltime.

The analysis also assumes steady-state flow conditions 

based on a constant recharge rate. Although recharge varies 

monthly and yearly, hydraulic gradients in the aquifer generally do not vary over time because of the small time scale of 

recharge variability compared to the much larger time scale of 
Limitations of Modeling Analysis 31

0 0.5 1 1.5 2 KILOMETERS

0 0.5 1 MILE

EXPLANATION

Agriculture

Open undeveloped

Commercial-industrial-transportation

High-density residential

Medium-density residential

Low-density residential

Very-low-density residential

Open urban

Forest

Forested and nonforested wetland

Water

Base from U.S. Geological Survey digital data

Massachusetts State Plane Coordinate System,

North American Datum of 1927

2005 land-use data from Massachusetts

 Office of Geographic Information

Land use

Waquoit

Bay

Bournes

Pond

Eel

Pond

Great

Pond

Green

Pond

Coonamessett

Pond

Johns

Pond

Ashumet

Pond

Infiltration beds

70°36' 70°33' 70°30'

41°38'

41°36'

41°34'

Figure 18. Generalized land use near the treated-wastewater plume, Ashumet Valley, Cape Cod, Massachusetts.
32 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

Table 4. Steady-state nonpoint-source and peak treated-wastewater plume nitrogen loads to major receptors, Ashumet Valley, 

Cape Cod, Massachusetts

[kg/yr, kilograms per year; Steady-state nonpoint-source nitrogen loads from Howes and others (2005). Location of receptors shown on figure 12]

Receptor

Steady-state nonpoint-source 

nitrogen load

(kg/yr)

Peak nitrogen treated-wastewater 

plume load

(kg/yr)

Ratio of peak wastewater plume 

load to nonpoint-source load 

(percent)

Backus River 3,210 350 10.9

Bournes River 2,583 200 7.7

Coonamessett River 16,866 550 3.3

Green Pond 7,356 150 2.0

advective transport. Walter and Masterson (2003) compared 

particle tracks on northwestern Cape Cod by using steadystate and transient flow simulations and found that the particle 

tracks were nearly identical in most parts of the aquifer 

(transient conditions were more influential near water-table 

divides and hydrologic boundaries). This result indicates that 

it is reasonable to use a steady-state model to simulate transport of the treated-wastewater plume even though recharge is 

variable. Variable groundwater withdrawals for irrigation and 

water supply also may affect hydraulic gradients. In particular, 

operation of wells for cranberry agriculture (bog irrigation 

and flooding) may affect local gradients near certain discharge 

areas of the treated-wastewater plume; these wells are not represented in the model. Long-term changes in recharge rates or 

sea level that may occur in response to possible future changes 

to the climate in the Northeast also are not represented in the 

analysis; therefore, load curves represent a projection of current conditions into the future.

Another limitation is that the regional model lacks 

detailed hydraulic data for the surface-water features (streams, 

ponds, and coastal embayments) that compose the discharge 

areas for the treated-wastewater plume. For streams these data 

include streamflow, streambed altitude, and hydraulic characteristics of streambed sediments. As shown in figures 13 and 

15, the large deviations of particle pathlines from the plumes 

appear largely to be related to the simulated surface-water features in the model. For example, the few particle pathlines that 

reach Waquoit Bay result from particles flowing under Ashumet Pond; a higher conductance for the pond bottom sediment 

might have resulted in more capture of deeper flowpaths in 

this area. Similarly, the particle pathlines that deviate from the 

western edge of the plume are affected by the Coonamessett 

River. This effect may indicate that flows in this river are overrepresented in the model. Even if the match between simulated 

and observed streamflow at the mouth of the Coonamesett 

River is reasonable, the model could be simulating too much 

discharge in the northern portion of the river, and, therefore, 

drawing particle pathlines too far in a westerly direction 

across the mapped plume boundary. Streamflow data for the 

northern part of the river were limited. The distributions of 

particles discharging to the streams and the downgradient 

headwaters of the adjoining embayments also are based on 

current knowledge of the distribution of flows between these 

surface-water bodies. Any changes in the simulated distribution of flows would affect the distribution of particles and the 

timing and magnitude of simulated loads to these receiving 

waters that are shown in figures 14, 16, and 17 and table 3. In 

addition, manipulation of flows in the cranberry bogs adds to 

the complexity of understanding average flow conditions in 

these streams.

The regional model also tends to pull particles deeper 

than the lower plume boundary (data not shown), which 

increases traveltimes in the aquifer and affects the distribution of the particles that discharge to receptors. Adjustment of 

model parameters such as hydraulic conductivity to minimize 

downward flow and keep particles within the gridded plume 

would improve the agreement between simulated and observed 

transport. In general, explicit consideration of the wastewaterplume location during calibration likely would lead to better 

agreement between simulated and observed nitrogen transport 

and reduce particle-path excursions across mapped plume 

boundaries. A different modeling approach also could provide 

better agreement between simulated and observed nitrogen 

transport in the aquifer. For example, solute-transport models, 

which typically account for advection, dispersion, adsorption, and biological reactions, could improve the simulation 

of nitrogen transport in the aquifer and loading to coastal 

discharge areas.

The regional model (Walter, 2013) originally was 

designed and used for advective transport of nitrate on western 

Cape Cod, including in this study area, so it was considered to 

be appropriate for the simulation of nitrogen transport in the 

treated-wastewater plume. However, given the modeling (and 

plume mapping) limitations discussed above, there is considerable uncertainty associated with the results presented in this 

study. The simulated nitrogen loading curves should be viewed 

as only one plausible set of outcomes based on particle-tracking simulations with the current regional model for western 

Cape Cod. Future modeling analyses would benefit from additional streamflow calibration data to better represent flows in 

the surface-water features near the treated-wastewater plume; 

inclusion of the location of the plume during calibration of the 

flow model; and use of a solute-transport model to simulate 

nitrogen transport in the aquifer.
Summary 33

Summary

Land disposal of treated wastewater from a treatment 

plant in operation from 1936 to 1995 on the Massachusetts 

Military Reservation (MMR) created a plume of contaminated 

groundwater that is migrating toward the coastal embayments 

in the town of Falmouth, Massachusetts. Although the plume 

has been investigated extensively by the U.S. Geological Survey (USGS) and other researchers since the late 1970s, the 3D 

distribution and transport of nitrogen at the plume scale (that 

is, the entire plume) has received relatively little study. The 

fate of nitrogen (N) in the treated-wastewater plume is of concern because land-use changes in coastal watersheds have led 

to nitrogen enrichment in the coastal embayments. To develop 

a better understanding of the potential effect of the treatedwastewater plume on coastal discharge areas, the USGS, in 

cooperation with the Air Force Center for Engineering and the 

Environment, evaluated the fate of nitrogen in the plume. The 

objectives of the study were to calculate the mass of nitrogen in the treated-wastewater plume as of 2007, evaluate the 

occurrence and extent of reactions that could remove nitrogen mass in the aquifer before discharge, simulate the spatial 

and temporal discharge of nitrogen to the rivers and coastal 

embayments in Falmouth, and compare treated-wastewater 

plume loads to steady-state nonpoint-source loads.

Groundwater samples from two large sampling events 

in 1994 and 2007 were used to map the size and location of 

the wastewater plume, calculate the mass of nitrogen species 

(nitrate and ammonium are the predominant nitrogen species 

in the plume), evaluate changes in plume mass since cessation 

of wastewater disposal in 1995, and create a gridded dataset 

suitable for use in nitrogen-transport simulations. In 2007, the 

treated-wastewater plume was about 1,200 meters (m) wide, 

30 m thick, and 7,700 m long and contained approximately 

87,000 kilograms (kg) nitrate-N and 31,600 kg total ammonium-N. Consistent with a relatively slow rate of migration 

caused by adsorption, the ammonium plume was substantially 

smaller than the nitrate plume, extending roughly 3,300 m 

downgradient from the infiltration beds in 2007. Between 

1994 and 2007, the mass of nitrate-N in the plume decreased 

by 7.8 percent, whereas the mass of ammonium-N decreased 

by 37.7 percent. Spatial patterns of nitrate concentrations and 

geochemical conditions in the wastewater plume indicate that 

the post-cessation rate of nitrate removal by denitrification is 

low or negligible. In contrast to nitrate, the ammonium plume 

appears to be undergoing biotransformation in the aquifer. The 

reactions responsible for ammonium mass loss are unclear. 

However, based on changes in spatial patterns in concentrations of ammonium and dissolved oxygen, nitrification is most 

likely responsible for the apparent loss. Nitrification converts 

ammonium to nitrate, which is more mobile and persistent in 

the aquifer. Consequently, nearly all of the biologically reactive nitrogen in the 2007 plume may be transported to coastal 

discharge areas as either nitrate or ammonium with relatively 

little loss to an environmentally nonreactive end product such 

as nitrogen gas. If a reaction such as anammox (the anaerobic 

oxidation of ammonium), which consumes both ammonium 

and nitrate, is widespread in the plume, then nitrogen-transport 

simulations that assume nonreactive (no mass loss in the aquifer) transport of nitrogen in the 2007 plume provide an upperend estimate of loading to coastal discharge areas. Additional 

data collection would be useful for evaluating nitrogen mass 

and reactivity at the plume scale.

Nitrogen transport was simulated with a previously 

calibrated regional 3D MODFLOW groundwater-flow model. 

Mass-loaded particle tracking was used to simulate the advective transport of nitrogen in the 2007 treated-wastewater 

plume to groundwater discharge areas (or receptors) along the 

coast. The movement and discharge of nitrogen were quantified by assigning particles to the model cells encompassing 

the wastewater plume and then transporting them to discharge 

areas under steady-state flow conditions. An accounting of 

the number and traveltime of particles reaching receptors was 

used to estimate the time history of nitrogen mass loading to 

these receptors. Particles were converted to mass by multiplying the nitrogen concentration assigned to the cell of particle 

origin by the volume of water in the cell. Overall, the use of 

particle tracking to simulate mass loading to coastal discharge 

areas is most appropriate for nonsorbing, nonreactive nitrogen species such as nitrate. To simulate ammonium transport, 

particle traveltimes were multiplied by the ammonium retardation factor of 2.5. The simulation, therefore, assumes that no 

oxidation reactions occur and that the entire ammonium plume 

migrates to coastal receptors as ammonium. Although such a 

simulation is not consistent with the observed ammonium-N 

mass loss, it is not possible to simulate reactive ammonium 

transport realistically with a mass-loaded particle-tracking 

analysis. The present simulation can be viewed as an endmember-transport simulation, reflecting maximum traveltimes 

to receptors and conservation of the total ammonium mass in 

the 2007 plume.

Simulation results indicate that about 95 percent of the 

nitrate-N and 99 percent of the ammonium-N in the treatedwastewater plume will discharge to the coastal receptors 

nearest to the downgradient end of the plume: the Coonamessett River, Backus River, Green Pond, and Bournes River. 

The Coonamessett River will receive the most nitrate-N 

mass (about 42 percent of the total nitrate-N mass) followed 

by the Bournes River, Green Pond, and the Backus River. 

Approximately 76 percent of the total nitrate-N mass in the 

plume will discharge to these receptors within 100 years of 

2007, 90 percent within 200 years, and 94 percent within 

500 years. Nitrate loads will peak within about 50 years (of 

2007) at all major receptors. The highest peak loads will occur 

at the Coonamessett River (450 kilograms per year (kg/yr) 

nitrate-N) and Backus River (350 kg/yr nitrate-N). As was 

the case for nitrate, the Coonamessett River will receive the 

most ammonium-N mass (about 45 percent of the total mass) 

followed by Green Pond, the Backus River, and the Bournes 

River. Because of adsorption, traveltimes for ammonium are 

longer than for nitrate; as a result, approximately 5 percent 

of the total ammonium-N mass in the plume will discharge 
34 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

to receptors within 100 years of 2007, 46 percent within 200 

years, and 81 percent within 500 years. The discharge to 

Green Pond accounts for most of the ammonium discharge 

after about 350 years. Simulations indicate that the Coonamessett River will receive the most cumulative nitrogen mass and 

the highest rate of discharge (load) from the 2007 treatedwastewater plume. The Backus and Bournes Rivers and 

Green Pond also will receive substantial discharge of nitrogen 

(mostly as nitrate) from the plume. Ongoing discharge to 

Ashumet Pond is minor because most of the nitrogen mass in 

the plume already is downgradient from the pond. The loads 

reported in this report assume that transport of nitrogen to the 

groundwater/surface-water interface results in discharge to 

surface water; biogeochemical processes near the interface (in 

the aquifer or subterranean estuary) that may reduce or eliminate loads before discharge are not considered.

To evaluate the contribution of the nitrogen loads from 

the treated-wastewater plume to total nitrogen loads to the 

discharge areas, the simulated treated-wastewater plume 

loads were compared to steady-state nonpoint-source loads 

calculated by the Massachusetts Estuaries Project for 2005. 

Wastewater nitrogen from residential septic systems is the 

major source of anthropogenic nitrogen in the groundwater 

watersheds of the coastal embayment systems. Nonpointsource nitrogen loads in 2005 were 16,866 kg/yr to the Coonamessett River, 3,210 kg/yr to the Backus River, 7,356 kg/yr 

to Green Pond, and 2,583 kg/yr to the Bournes River. Simulated peak total-N wastewater-plume loads were 550 kg/yr 

to the Coonamessett River, 350 kg/yr to the Backus River, 

150 kg/yr to Green Pond, and 200 kg/yr to the Bournes River. 

Before and after peak loads, simulated treated-wastewater 

plume loads are about 1 to 2 percent of nonpoint-source loads. 

Simulation results indicate that the total-N loads from the 

treated-wastewater plume are much lower than corresponding 

steady-state nonpoint-source loads from the watersheds; peak 

plume loads are equal to 11 percent or less of the nonpointsource loads.

References Cited

Air Force Center for Engineering and the Environment, 

2010, Groundwater plume maps and information booklet: 

Air Force Center for Engineering and the Environment, 

Massachusetts Military Reservation, Cape Cod, Mass., 

April 2010, 31 p.

Aquaveo, 2012, GMS software documentation: Groundwater 

Modeling System version 7.0, accessed September 2012, at 

http://www.aquaveo.com/.

Barber, L.B., II, 1992, A hierarchical analytical approach to 

evaluating the occurrence, distribution, and biogeochemical fate of organic contaminants in sewage-contaminated 

groundwater, in Lesage, S., and Jackson, R.E., eds., 

Groundwater quality and analysis at hazardous waste sites, 

New York, Marcel Dekker, p. 73–120.

Barber, L.B., II, Thurman, E.M., Schroeder, M.P., and 

LeBlanc, D.R., 1988, Long-term fate of organic micropollutants in sewage-contaminated groundwater: Environmental Science & Technology, v. 22, no. 2, p. 205–211.

Böhlke, J.K., and Denver, J.M., 1995, Combined use of 

groundwater dating, chemical, and isotopic analyses to 

resolve the history and fate of nitrate contamination in two 

agricultural watersheds, Atlantic Coastal Plain, Maryland: 

Water Resources Research, v. 31, no. 9, p. 2319–2339.

Böhlke, J.K., Smith, R.L., and Miller, D.N., 2006, Ammonium transport and reaction in contaminated groundwater—Application of isotope tracers and isotope fractionation factors: Water Resources Research, v. 42, W05411, 

doi:10.1029/2005WR004349.

Bowen, J.L., Kroeger, K.D., Tomasky, G., Pabich, W.J., Cole, 

M.L., Carmichael, R.H., and Valiela, I., 2007, A review of 

land-sea coupling by groundwater discharge of nitrogen to 

New England estuaries—Mechanisms and effects: Applied 

Geochemistry, v. 22, p. 175–191.

Bowen, J.L., and Valiela, I., 2001, The ecological effects of 

urbanization of coastal watersheds— Historical increases in 

nitrogen loads and eutrophication of Waquoit Bay estuaries: Canadian Journal of Fish and Aquatic Sciences, v. 58, 

p. 1489–1500.

Ceazan, M.L., Thurman, E.M., and Smith, R.L., 1989, Retardation of ammonium and potassium through a contaminated 

sand and gravel aquifer—The role of cation exchange: 

Environmental Science &Technology, v. 23, no. 11, 

p. 1402–1408.

Clarke, I., Timlin, R., Bourbonnais, A., Jones, K., LaFleur, D., 

and Wickens, K., 2008, Origin and fate of industrial ammonium in anoxic ground water–15N evidence for anaerobic 

oxidation (anammox): Ground Water Monitoring and 

Remediation, v. 28, no. 3, p. 73–82.

Cole, M.L., Kroeger, K.D., McClelland, J.W., and Valiela, I., 

2006, Effects of watershed land use on nitrogen concentrations and δ15 nitrogen in groundwater: Biogeochemistry, 

v. 77, p. 199–215.

DeSimone, L.A., and Howes, B.L., 1998, Nitrogen transport 

and transformation in a shallow aquifer receiving wastewater discharge—A mass balance approach: Water Resources 

Research, v. 34, no. 2, p. 271–285.
References Cited 35

Farnsworth, R.K., Thompson, E.S., and Peck, E.L., 1982, 

Evaporation atlas for the contiguous 48 States: National 

Oceanic and Atmospheric Administration Technical Report 

NWS 33, 26 p.

Frimpter, M.H., 1990, A mass-balance nitrate model for predicting the effects of land use on ground-water quality: U.S. 

Geological Survey Open-File Report 88–493, 17 p.

Garabedian, S.P., LeBlanc, D.R., Gelhar, L.W., and Celia, 

M.A., 1991, Large-scale natural gradient tracer test in 

sand and gravel, Cape Cod, Massachusetts—2. Analysis of 

spatial moments for a nonreactive tracer: Water Resources 

Research, v. 27, no. 5, p. 911–924.

Harbaugh, A.W., 2005, MODFLOW–2005, the U.S. Geological Survey modular ground-water model—The groundwater flow process: U.S. Geological Survey Techniques and 

Methods, book 6, chap. A16, variously paged.

Harbaugh, A.W., Banta, E.R., Hill, M.C., and McDonald, 

M.G., 2000, MODFLOW–2000, The U.S. Geological Survey modular ground-water model—User guide to modularization concepts and the ground-water flow process: U.S. 

Geological Survey Open-File Report 00–92, 121 p.

Harvey, R.W., and Barber, L.B., II, 1992, Associations of freeliving bacteria and dissolved organic compounds in a plume 

of contaminated groundwater: Journal of Contaminant 

Hydrology, no. 9, p. 91–103.

Harvey, R.W., Smith, R.L., and George, L., 1984, Effect 

of organic contamination upon microbial distributions 

and heterotrophic uptake in a Cape Cod, Mass., aquifer: 

Applied and Environmental Microbiology, v. 48, no. 2, 

p. 1197–1202.

Howes, B., Ramsey, J.S., Kelley, S.W., Samimy, R., Schlezinger, D., and Eichner, E., 2005, Linked watershed-embayment model to determine critical nitrogen loading thresholds for Great/Perch Pond, Green Pond and Bournes Pond, 

Falmouth, Massachusetts: Boston, Massachusetts Department of Environmental Protection, Massachusetts Estuaries 

Project, 205 p.

Jetten, M.S.M., 1999, New pathways for ammonia conversion in soil and aquatic ecosystems: Plant and Soil, v. 230, 

p. 9–19.

Jetten, M.S.M., Wagner, M., Fuerst, J., van Loosdrecht, 

M., Kuenen, G., and Stous, M., 2001, Microbiology and 

application of the anaerobic ammonium oxidation (anammox) process: Current Opinions in Biotechnology, v. 12, 

p. 283–288.

Korom, S.F., 1992, Natural denitrification in the saturated 

zone—A review: Water Resources Research, v. 28, no. 6, 

p. 1657–1668.

Kroeger, K.D., and Charette, M.A., 2008, Nitrogen biogeochemistry of submarine groundwater discharge: Limnology 

and Oceanography, v. 53, no. 3, p. 1025–1039.

Kroeger, K.D., Cole, M.L., and Valiela, I., 2006a, Groundwater-transported dissolved organic nitrogen exports from 

coastal watersheds: Limnology and Oceanography, v. 51, 

no. 5, p. 2248–2261.

Kroeger, K.D., Cole, M.L., York, J.K., and Valiela, I., 2006b, 

Nitrogen loads to estuaries from waste water plumes—

Modeling and isotopic approaches: Ground Water, v. 44, 

no. 2, p. 188–200.

LeBlanc, D.R., 1984a, Sewage plume in a sand and gravel 

aquifer, Cape Cod, Massachusetts: U.S. Geological Survey 

Water-Supply Paper 2218, 28 p.

LeBlanc, D.R., 1984b, Digital modeling of solute transport in 

a plume of sewage-contaminated ground water, in LeBlanc, D.R., ed., Movement and fate of solutes in a plume of 

sewage-contaminated ground water, Cape Cod, Massachusetts—U.S. Geological Survey Toxic Waste Ground-Water 

Contamination Program—Papers presented at the toxic 

waste technical meeting, Tucson, Arizona, March 20–22, 

1984: U.S. Geological Survey Open-File Report 84–475, 

p. 11–46.

Massachusetts Office of Geographic Information, 2010, Datalayers and metadata: Massachusetts Office of Geographic 

Information, accessed 2010 at http://www.state.ma.us./mgis/

massgis.htm.

Masterson, J.P., Carlson, C.S., and Walter, D.A., 2009, 

Hydrogeology and simulation of groundwater flow in the 

Plymouth-Carver-Kingston-Duxbury aquifer system, southeastern Massachusetts: U.S. Geological Survey Scientific 

Investigations Report 2009–5063, 110 p.

Masterson, J.P., Stone, B.D., Walter, D.A., and Savoie, J.G., 

1997, Hydrogeologic framework of western Cape Cod, 

Massachusetts: U.S. Geological Survey Hydrologic Investigations Atlas HA–741, 1 pl., scale 1:50,000.

Masterson, J.P., Walter, D.A., and LeBlanc, D.R., 1998, Delineation of contributing areas to selected public-supply wells, 

western Cape Cod, Massachusetts: U.S. Geological Survey 

Water-Resources Investigations Report 98–4237, 45 p.

McCobb, T.D., LeBlanc, D.R., and Hess, K.M., 1999, Determination of temporal and spatial variability of hydraulic 

gradients in an unconfined aquifer using three-point triangulation, Cape Cod, Massachusetts, in Morganwalp, D.W., and 

Buxton, H.T., eds., U.S. Geological Survey Toxic Substances Hydrology Program—Proceedings of the technical 

meeting, Charleston, South Carolina, March 8–12, 1999—

volume 3—Subsurface contamination from point sources: 

U.S. Geological Survey Water-Resources Investigations 

Report 99–4018C, p. 349–360.
36 Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas, Ashumet Valley, Massachusetts

McCobb, T.D., LeBlanc, D.R., Walter, D.A., Hess, K.M., 

Kent, D.B., and Smith, R.L., 2003, Phosphorus in a groundwater contaminant plume discharging to Ashumet Pond, 

Cape Cod, Massachusetts, 1999: U.S. Geological Survey 

Water-Resources Investigations Report 02–4306, 69 p.

McDonald, M.G., and Harbaugh, A.W., 1988, A modular 

three-dimensional finite-difference ground-water-flow 

model: U.S. Geological Survey Techniques of WaterResources Investigations, book 6, chap. A1, 586 p.

Miller, D.N., and Smith, R.L., 2009, Microbial characterization of nitrification in a shallow, nitrogen-contaminated 

aquifer, Cape Cod, Massachusetts and detection of a novel 

cluster associated with nitrifying Betaproteobacteria: Journal of Contaminant Hydrology, v. 103, p.182–193.

Mulder, A., van de Graaf, A.A., Robertson, L.A., and Kuenen, 

J.G., 1995, Anaerobic ammonium oxidation discovered in a 

denitrifying fluidized bed reactor: FEMS Microbiology and 

Ecology, v. 16, p. 177–184.

Orzol, L.L., 1997, User’s guide for MODTOOLS—Computer programs for translating data of MODFLOW and 

MODPATH into geographic information systems files: U.S. 

Geological Survey Open-File Report 97–240, 86 p.

Parkhurst, D.L., Stollenwerk, K.G., and Colman, J.A., 2003, 

Reactive-transport simulation of phosphorus in the sewage plume at the Massachusetts Military Reservation, 

Cape Cod, Massachusetts: U.S. Geological Survey WaterResources Investigations Report 03–4017, 33 p.

Persky, J.H., 1986, The relation of ground-water quality to 

housing density, Cape Cod, Massachusetts: U.S. Geological 

Survey Water-Resources Investigations Report 86–4093, 

28 p.

Pollock, D.W., 1994, Users’ guide for MODPATH/MODPATH–PLOT, version 3—A particle tracking post-processing package for MODFLOW, the U.S. Geological Survey 

finite-difference ground-water flow model: U.S. Geological 

Survey Open-File Report 94–464, 234 p.

Prudic, D.E., 1989, Documentation of a computer program to 

simulate stream-aquifer relations using a modular, finite-difference, ground-water flow model: U.S. Geological Survey 

Open-File Report 88–729, 113 p.

Repert, D.A., Barber, L.B., Hess, K.M., Keefe, S.H., Kent, 

D.B., LeBlanc, D.R., and Smith, R.L., 2006, Long-term 

natural attenuation of carbon and nitrogen within a groundwater plume after removal of the treated wastewater 

source: Environmental Science & Technology, v. 40, no. 4, 

p. 1154–1162.

Rich, J.J., Dale, O.R., Song, B., and Ward, B.B., 2008, 

Anaerobic ammonium oxidation (anammox) in Chesapeake 

Bay sediments: Microbial Ecology, v. 55, p. 311–320.

Robertson, W.D., Moore, T.A., Spoelstra, J., Li, L., Elgood, 

R.J., Clark, I.D., Schiff, S.L., Aravena, R., and Neufeld, 

J.D., 2012, Natural attenuation of septic system nitrogen by 

anammox: Ground Water, v. 50, no. 4, p. 541–553.

Savoie, J.G., and LeBlanc, D.R., eds., 1998, Water-quality 

data and methods of analysis for samples collected near 

a plume of sewage-contaminated ground water, Ashumet 

Valley, Cape Cod, Massachusetts, 1993–94: U.S. Geological Survey Water-Resources Investigations Report 97–4269, 

208 p.

Savoie, J.G., LeBlanc, D.R., Fairchild, G.M., Smith, R.L., 

Kent, D.B., Barber, L.B., Repert, D.A., Hart, C.P., Keefe, 

S.H., and Parsons, L.A., 2012, Groundwater-quality data for 

a treated-wastewater plume near the Massachusetts Military 

Reservation, Ashumet Valley, Cape Cod, Massachusetts, 

2006–08: U.S. Geological Survey Data Series 648, 11 p., 

CD–ROM.

Smith, R.L., Baumgartner, L.K., Miller, D.N., Repert, D.A., 

and Böhlke, J.K., 2006, Assessment of nitrification potential in ground water using short term, single-well injection 

experiments: Microbial Ecology, v. 51, p. 22–35.

Smith, R.L., Böhlke, J.K., Garabedian, S.P., Révész, 

K.M., and Yoshinari, T., 2004, Assessing denitrification in groundwater using natural gradient tracer tests 

with 15N—In situ measurement of a sequential multistep 

reaction: Water Resources Research, v. 40, W07101, 

doi:1029/2003WR002919.

Smith, R.L., and Duff, J.H., 1988, Denitrification in a sand and 

gravel aquifer: Applied and Environmental Microbiology, 

v. 52, no. 5, p. 1071–1078.

Smith, R.L., Garabedian, S.P., and Brooks, M.H., 1996, Comparison of denitrification activity measurements in groundwater using cores and natural-gradient tracer tests: Environmental Science & Technology, v. 30, no. 12, p. 3448–3456.

Smith, R.L., Howes, B.L., and Duff, J.H., 1991, Denitrification in nitrate-contaminated groundwater—Occurrence in 

steep vertical geochemical gradients: Geochimica et Cosmochimica Acta, v. 55, p. 1815–1825.

Sudicky, E.A., 1986, A natural gradient experiment on solute 

transport in a sand aquifer—Spatial variability of hydraulic 

conductivity and its role in the dispersion process: Water 

Resources Research, v. 22, no. 13, p. 2069–2082.

Thurman, E.M., Barber, L.B., II, and LeBlanc, D., 1986, 

Movement and fate of detergents in groundwater—A 

field study: Journal of Contaminant Hydrology, no. 1, 

p. 143–161.
References Cited 37

Walter, D.A., 2007, Use of numerical models to simulate 

transport of sewage-derived nitrate in a coastal aquifer, Cen-

tral and Western Cape Cod, Massachusetts: U.S. Geological 

Survey Scientific Investigations Report 2007–5259, 41 p.

Walter, D.A., 2013, The effect of wastewater management 

actions on the hydrologic system and nitrogen loading rates 

to well and ecological receptors, Popponesset Bay water-

shed, Cape Cod, Massachusetts: U.S. Geological Survey 

Scientific Investigations Report 2013–5060, In press

Walter, D.A., and Masterson, J.P., 2003, Simulation of advec-

tive flow under steady-state and transient recharge condi-

tions, Camp Edwards, Massachusetts Military Reservation, 

Cape Cod, Massachusetts: U.S. Geological Survey WaterResources Investigations Report 03–4053, 51 p.

Walter, D.A., Masterson, J.P., and LeBlanc, D.R., 2002, Simu-

lated pond-aquifer interactions under natural and stressed 

conditions near Snake Pond, Cape Cod, Massachusetts: U.S. 

Geological Survey Water-Resources Investigations Report 

99–4174, 35 p.

Walter, D.A., and Whealan, A.T., 2005, Simulated water 

sources and effects of pumping on surface and ground 

water, Sagamore and Monomoy flow lenses, Cape Cod, 

Massachusetts: U.S. Geological Survey Water-Resources 

Investigations Report 2004–5181, 85 p.
THIS PAGE INTENTIONALLY LEFT BLANK
Prepared by the Pembroke Publishing Service Center.

For more information concerning this report, contact:

Director

U.S. Geological Survey

New England Water Science Center

Massachusetts-Rhode Island Office

10 Bearfoot Road

Northborough, MA 01532

dc_ma@usgs.gov

or visit our Web site at:

http://ma.water.usgs.gov
Barbaro and others—Transport of Nitrogen in a Treated-Wastewater Plume to Coastal Discharge Areas—Scientific Investigations Report 2013–5061

Printed on recycled paper