Knee and Paytan Ra - University of California, Santa Cruz

Transcription

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Knee KL and Paytan A (2011) Submarine Groundwater Discharge: A Source of
Nutrients, Metals, and Pollutants to the Coastal Ocean. In: Wolanski E and
McLusky DS (eds.) Treatise on Estuarine and Coastal Science, Vol 4, pp. 205–233.
Waltham: Academic Press.
© 2011 Elsevier Inc. All rights reserved.
Author's personal copy
4.08 Submarine Groundwater Discharge: A Source of Nutrients, Metals,
and Pollutants to the Coastal Ocean
KL Knee, Smithsonian Environmental Research Center, Edgewater, MD, USA
A Paytan, University of California, Santa Cruz, CA, USA
© 2011 Elsevier Inc. All rights reserved.
4.08.1
4.08.1.1
4.08.1.2
4.08.1.2.1
4.08.1.2.2
4.08.1.3
4.08.2
4.08.2.1
4.08.2.2
4.08.2.3
4.08.2.4
4.08.3
4.08.3.1
4.08.3.2
4.08.3.3
4.08.3.4
4.08.4
4.08.4.1
4.08.4.2
4.08.4.3
4.08.4.4
4.08.4.5
4.08.5
4.08.5.1
4.08.5.2
4.08.5.3
4.08.5.4
4.08.5.5
4.08.5.6
4.08.5.7
4.08.6
4.08.6.1
4.08.6.2
4.08.6.3
4.08.6.4
4.08.7
4.08.7.1
4.08.7.2
4.08.7.3
4.08.7.4
References
Introduction
Drivers of SGD
Measurement of SGD
Direct measurement of SGD
Indirect measurement of SGD
Summary
Flux of Water into the Coastal Ocean from SGD
Fresh versus Saline Groundwater Discharge
Temporal Variability in SGD
Spatial Variability in SGD
Comparison of SGD Fluxes to Input from Other Water Sources
Nutrient Loads from SGD
Nutrient Concentrations in Coastal Groundwater
SGD-Related Nutrient Fluxes into Coastal Waters
Ecological Importance of Groundwater-Borne Nutrients
Summary and Conclusions
Metal Loads from SGD
Metal Concentrations in Coastal Groundwater
Metal Species and Transformations in the Subterranean Estuary
Fluxes of Metals into the Coastal Ocean from SGD
Effects of SGD-Derived Metals in Coastal Waters
SGD-Related Fluxes of Other Constituents
Dissolved Organic Carbon
FIB and Other Microbes
Caffeine, Pharmaceuticals, and Personal Care Products
Pesticides
Volatile Organic Compounds
Methane
SGD and Climate Change
Groundwater Temperature
SGD Quantity
SGD Chemistry and Quality
Effect of SGD on Climate Change
Summary
SGD at the Local, Regional, and Global Scales
SGD in a Changing World
Current and Future Research Directions
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Abstract
This chapter provides a review of the current state of knowledge on submarine groundwater discharge (SGD) and the
associated fluxes of nutrients, trace metals, microbes, pharmaceuticals, and other terrestrial constituents to coastal waters. We
review methods of estimating SGD, present flux estimates from different locations worldwide, and discuss how various
hydrogeologic features such as topography, aquifer substrate, climate, waves, and tides affect SGD. We discuss the range of
nutrient and metal concentrations observed in groundwater and their relationship to land use, and explore the chemical
changes that nutrients and metals undergo during their seaward journey through the aquifer. Climate change is likely to affect
both the quantity and the quality of SGD, and we investigate these effects, which are only beginning to be studied. The
Treatise on Estuarine and Coastal Science, 2011, Vol.4, 205-233, DOI: 10.1016/B978-0-12-374711-2.00410-1
205
206
Submarine Groundwater Discharge: A Source of Nutrients, Metals, and Pollutants to the Coastal Ocean
chapter concludes with a discussion of active areas of SGD research, including expanding the geographic scope of SGD
studies; characterizing and reducing the uncertainty associated with SGD measurements; understanding the behavior of
nutrients, metals, and other pollutants in the subterranean estuary; and modeling SGD on a global scale.
4.08.1 Introduction
(Moore, 1999). However, unlike river flow, SGD is often invi
sible, occurring as diffuse seepage through porous rocks or
sediments rather than in discrete springs or discharge conduits.
Submarine groundwater discharge (SGD) has been defined as
“direct groundwater outflow across the ocean–land interface
into the ocean” (Church, 1996), including both fresh ground
water flow originating from inland recharge areas and seawater
that circulates through the coastal aquifer on tidal to seasonal
timescales (Michael et al., 2005). It also includes discharge
from both confined and unconfined aquifers, occurring at the
shoreline or offshore.
The existence of offshore freshwater springs has been recog
nized since the time of the Roman Empire (Kohout, 1966;
Taniguchi et al., 2002; Burnett et al., 2006). However, only
since the 1960s have scientists had the tools to study SGD
quantitatively. Over the past two decades in particular, SGD
studies have been conducted at many sites worldwide
(Taniguchi et al., 2002; Figure 1).
SGD is important in the context of estuarine and coastal
science because it connects land and sea. Moore (1999) coined
the term ‘subterranean estuary’ to describe that part of the
coastal aquifer where freshwater and saltwater mix and where
SGD occurs (Figure 2). Like a river, SGD delivers freshwater,
nutrients, metals, and other constituents to coastal waters, and,
like river estuaries, subterranean estuaries are dynamic mixing
zones where biological and chemical transformations occur
4.08.1.1
Drivers of SGD
Precipitation, water pumping, aquifer characteristics, beach
morphology, the presence and level of development of stream
systems, waves, and tides all control the quantity and location
of SGD. Compaction and dewatering of sediments may also
contribute to SGD (Burnett et al., 2003; Hays and Ullman,
2007); however, these effects have not been quantified.
Rainfall recharges the aquifer, maintaining the hydraulic gradi
ent necessary to drive seaward groundwater flow. Aquifer
characteristics that influence groundwater flow in general also
influence SGD. These include the porosity, permeability, and
hydraulic conductivity of the aquifer substrate; the homogene
ity or heterogeneity of the aquifer substrate; the presence of
fractures or other preferred flow paths such as faults or lava
tubes; and the hydraulic head contours within the aquifer.
Lakes, rivers, and streams can serve as either sources or sinks
for groundwater flow; however, the relative importance of SGD
is expected to be greater in areas where stream systems are
poorly developed or absent.
12
17
42
2,16,41
39
45
17 4,6,7,8,9,30
13,28
14
43
18,29
34
25
21
33
43
36
1,15,31
5
23
26,27
3
23
38
51
23
23 23
37
40
22
23
23
10,35
11
23
49
52
24
50
53
48
54
47
46
19,20
32
44
Figure 1 Map of sites worldwide where submarine groundwater discharge (SGD) has been estimated or measured. Numbers 1–45 refer to Table 1 of
Taniguchi et al. (2002). Numbers 46–54 (in red) indicate new sites where SGD fluxes have been estimated since 2002; details of these studies may be
found in Table 1. Many additional SGD studies have been conducted at or near sites 1–45 since 2002; however, these are not indicated on the map. Base
map reproduced from Taniguchi, M., Burnett, W.C., Cable, J.E., Turner, J.V., 2002. Investigation of submarine groundwater discharge. Hydrological
Processes 16, 2115–2129.
Land surface
207
Submarine groundwater discharge = meteoric water (groundwater of recent atmospheric origin)
+ connate water (water incorporated into rock pores when the rocks form)
+ recirculated seawater
Water table
Seepage at
shoreface
Fresh pore water
one
ng z
ixi
rg e
ter m
twa
ischa
/sal
ter d
r
a
e
w
t
a
h
Fres
shw
Fre
Seawater
Brackish-pore-water
recirculation cell
ter
e wa
e por
Salin
Submarine spring
Confining unit
Seaflo
or
ter
Fresh pore water
arge
disch
water
h
s
e
r
F
a
ltw
r/sa ne
e
t
a
zo
shw ng
Fre mixi
Saline pore water
Figure 2 Schematic drawing of submarine groundwater discharge (SGD) including fresh and saline components. Reproduced with permission from
Swarzenski and Kindinger (2003).
Waves and tides affect both fresh and saline SGD. Waves
promote the recirculation of seawater through the coastal aqui
fer, resulting in brackish and saline SGD (Li et al., 1999; Horn,
2002). They deposit seawater on the beach surface, and this
water then flows down through the aquifer to the shoreline.
High tides can have a similar effect, saturating portions of the
aquifer that then drain at ebb and low tides. Tidal variations in
sea-surface height on semidiurnal, diurnal, and fortnightly
timescales also affect SGD (Lee, 1977; Taniguchi, 2002; De
Sieyes et al., 2008). Generally, SGD flow is greatest at low
tide, when the hydraulic gradient between the inland water
table and sea-surface height is steepest. It is also higher at spring
tide compared to neap tide because the greater tidal pumping
oscillations at spring tide enhance the flow of recirculated
seawater.
4.08.1.2
Measurement of SGD
While it is fairly straightforward to gauge the flow of a river,
measuring SGD fluxes can be considerably more complicated.
Various methodologies have been used to estimate SGD,
including seepage meters, natural and artificial groundwater
tracers, water balances, and hydrogeological modeling.
4.08.1.2.1
Direct measurement of SGD
Seepage meters are devices installed in the seafloor that
measure groundwater seepage directly. They may also col
lect groundwater for later chemical analysis. Lee (1977)
designed a simple, manual seepage meter consisting of a
plastic bag attached to the top or bottom of a 55-gallon
(208-l) drum (Figure 3). The drum is pressed into perme
able bottom sediments, and water seeping up through the
sediments collects in the bag. Shaw and Prepas (1989)
improved upon this design by recognizing that sampling
artifacts can be greatly reduced by prefilling the plastic bag
with a known volume (typically, 1 l) of water. More sophis
ticated, automated designs that measure flow through heat
transport, including the constant-heat and heat-pulse types,
have been developed by Taniguchi and collaborators
(Taniguchi and Fukuo, 1993; Taniguchi and Iwakawa,
2001; Taniguchi et al., 2003). Dye-diffusion (Sholkovitz
et al., 2003) and ultrasonic (Paulsen et al., 2001) seepage
meters have also been developed.
Seepage meters have the advantage of allowing direct mea
surement of discharging groundwater. In addition, manual
seepage meters are relatively inexpensive and simple to con
struct (Cable et al., 1997). The disadvantage of seepage meters
is that, while they provide an accurate measure of seepage
occurring at the exact installation site, a large number are
required to estimate discharge over a larger area, such as a bay
or beach, because SGD can be highly heterogeneous on small
spatial scales (Shaw and Prepas, 1990a, 1990b). In addition,
the meters themselves can alter pressure gradients and flow
patterns, and they can be difficult or impossible to install in
turbulent environments such as wave-dominated coasts
(Burnett et al., 2006).
208
(a) Lee-type manual seepage meter
(b) Ultrasonic seepage meter. (Left) Schematic drawing of
entire seepage meter. (Below) Detail showing interior of
flow tube.
Data
logger
Plastic bag
Flow tube
Flow
outlet
Flow
inlet
Rubber-band wrap
Amber
latex
tube
Polyethylene tube
Rubber stopper
Seafloor
Piezoelectric transducers
End section of steel drum
Stainless-steel funnel
(d) Dye-diffusion seepage meter
(c) Taniguchi-type heat pulse seepage meter
Top view
e
lic tub
e
m pip line
minu
Alum Nichrome
Waste
Water
Dye
1
Acry
Water
TH1
TH2
A
C
2
TH3
T/S/D
SGD
00
be
PVC tu
SGD
B
4 5
6
+ −
+ −
T/S/D
TH : Thermistor
Recorder
and
timer
00
TH0
3
Side view
Surface
Dye
5 6
65 cm
Water Bottom
00
T/S/D
T/S/D
Iron
funnel
Seepage water
B
A
Water
00
Heat
insulator
2
1
SGD
3
D ye
injection
C
SGD
NAS
4
Waste
+ −
+ −
Battery
Pump
battery
Sediment
SGD
Figure 3 Schematic diagrams of the four main types of seepage meters. Based on figures from (a) Lee, D.R., 1977. A device for measuring seepage flux in
lakes and estuaries. Limnology and Oceanography 22, 140–147; (b) Paulsen, R.J., Smith, C.F., O’Rourke, D., Wong, T., 2001. Development and evaluation of
an ultrasonic ground water seepage meter. Ground Water 39, 904–911; (c) Taniguchi, M., Fukuo, Y., 1993. Continuous measurements of ground-water
seepage using an automatic seepage meter. Ground Water 31, 675–679; (d) Sholkovitz, E., Herbold, C., Charette, M., 2003. An automated dye-dilution based
seepage meter for the time-series measurement of submarine groundwater discharge. Limnology and Oceanography: Methods 1, 16–28.
A theoretically similar but larger-scale method of directly
measuring SGD was developed by Hays and Ullman (2007).
Taking advantage of the natural, partially enclosed morphology
of a beach at Cape Henlopen, Delaware, they created a fully
enclosed receiving pond using a movable Plexiglas weir. They
installed the weir at low tide when the pond area had drained of
seawater, and then measured the rate at which groundwater
impounded behind the weir drained out through a gutter.
Another direct method of estimating SGD is by using multilevel
piezometer nests that measure the hydraulic head at various
depths in the aquifer, either on land or beneath the seafloor
(Freeze and Cherry, 1979; Taniguchi et al., 2003). The hydraulic
gradient in the aquifer can be calculated from hydraulic head
measurements, and, if the hydraulic conductivity of the aquifer
is known, the discharge can be calculated using Darcy’s law:
q ¼ −K
dh
dl
where q is the volume of groundwater discharge per unit area
per unit time, K the hydraulic conductivity, and dh/dl the
hydraulic gradient. The main difficulty with this method is
obtaining representative values of K, which can vary by orders
of magnitude within an aquifer and vary depending on the
direction of groundwater flow (Taniguchi et al., 2003).
4.08.1.2.2
Indirect measurement of SGD
Tracers. Constituents with high concentrations in groundwater
and lower or negligible concentrations in surface water have
been used as tracers of SGD. Natural tracers, including radium,
radon, salinity, silica, barium, and methane, are elements or
compounds that are naturally present in groundwater. Artificial
tracers, including SF6, 131I, 32P, and fluorescein dye (Corbett
et al., 2000; Cable and Martin, 2008), can be added to the
aquifer at a specific location and their movement is then mon
itored. Natural tracers have been much more widely used in SGD
studies than artificial tracers (Burnett et al., 2006). Tracers pro
vide an integrated estimate of SGD over a larger spatial scale.
They can also be used in settings that would be unsuitable for
seepage meters, and, in some cases, tracer characteristics such as
Ra isotope ratios can be used to distinguish different ground
water endmembers (Moore, 2006; Young et al., 2008).
The element radium (Ra) has four naturally occurring iso
topes (223Ra, 224Ra, 226Ra, and 228Ra), which are products of
the 238U, 235U, and 232Th decay series, and are thus present in
232-Thorium decay series
238-Uranium decay series
U
Th
234U
235U
(4.47E9 a)
(2.45E5 a)
(7E8 a)
234Pa
(6.69 h)
234Th
(24.1 d)
231Pa
(3.3E4 a)
230Th
(7.5E4 a)
Rn
Po
(1.6E3 a)
227Ac
(6.15 h)
(21.8 a)
224Ra
223Ra
(3.66 d)
(11.4 d)
Radionuclide
(half-life)
220Rn
(3.823 d)
(55.6 s)
Alpha
decay
210Po
214Po
214Po
216Po
212Po
(3.04 m)
(1.6E–4 s)
(138.4d)
(0.15 s)
(3E–7 s)
(19.7 m)
214Pb
(26.9 m)
(18.7 d)
220Ac
226Ra
214Bi
227Th
231Th
(1.06 d)
(1.91 a)
228Ra
(5.75 a)
226Ra
Bi
Pb
220Th
232Th
(1.4E10 a)
Ac
Ra
235-Uranium decay series
230U
Pa
209
Beta
decay
219Rn
(3.96 s)
215Po
(1.78 ms)
212Bi
210Bi
(5.01 d)
211Bi
(1.01 h)
210Pb
206Pb
212Pb
200Pb
(22.6 y)
(stable)
(10.6 h)
(stable)
Tl
203
Tl
(3.05 m)
(2.14 m)
207Pb
(stable)
(36.1 m)
207Tl
(4.77 m)
211Pb
Figure 4 The uranium and thorium decay series, which produce the natural groundwater tracers Ra and Rn. Reproduced with permission from Swarzenski,
P.W., 2007. U/Th series radionuclides as coastal groundwater tracers. Chemical Reviews 107, 663–674. Copyright 2007 American Chemical Society.
most rocks and sediments on Earth (Figure 4). These isotopes
have half-lives ranging from 3.7 days to over 1600 years, mak
ing them useful for quantifying coastal mixing processes on a
variety of timescales. Activities of all four isotopes tend to be
highest in the brackish mixing zone that characterizes many
subterranean estuaries and elevated in coastal compared to
offshore waters; they also decrease with distance offshore
(Moore and Arnold, 1996; Moore, 2003; Hwang et al., 2005a;
Paytan et al., 2006).
The main advantage of Ra over other natural tracers is that it
can be used to calculate coastal mixing rates and residence
times; the main disadvantage is that the desorption of Ra
from rock and particle surfaces depends on salinity (Webster
et al., 1995). In freshwater, Ra is completely particle bound,
and its desorption increases with salinity, with the exact rela
tion depending on particle size (Webster et al., 1995) and pH
(Gonneea et al., 2008). Thus, fresh SGD may not have high Ra
activity (Burnett et al., 2006). Radon (222Rn) and methane
(CH4) are usually present in both fresh and saline groundwaters, and recent advances in 222Rn (Burnett et al., 2001;
Dulaiova et al., 2005) and CH4 (Kim and Hwang, 2002) detec
tion allow continuous, in situ measurement and mapping of
concentrations over large areas. However, because they are
gases, exchange between the dissolved and atmospheric phases
must be accounted for.
Natural tracers are usually incorporated into mass-balance
or ‘box’ models to estimate SGD fluxes (Burnett et al., 2006;
Figure 5); thus, it is essential to quantify any non-SGD sources
of the tracer, such as rivers or diffusive flux from sediments, to
the system. It is also important that the other model inputs,
namely, the volume of the box and the residence time of water
within it, be well constrained. Finally, the spatial heterogeneity
of groundwater chemistry can make it difficult to characterize
the discharging groundwater (Moore et al., 2006; Swearman
et al., 2006).
Water-balance approach. The water balance approach relies
on quantifying inputs (precipitation) and outputs (evapotran
spiration, surface runoff, and groundwater discharge) of
freshwater into a system and setting them equal to each other.
If precipitation, evapotranspiration, and surface runoff are
known, fresh groundwater discharge can be calculated
(Taniguchi et al., 2003). This method is only suitable for calcu
lating fresh SGD flux and does not address the recirculated
seawater component, which can be an important source of
nutrients and other land-derived chemicals (Li et al., 1999;
Kim et al., 2003; Shellenbarger et al., 2006). In addition,
water balances are generally applicable on a basin or larger
scale, and they may not work well in situations where
watershed boundaries do not correspond with groundwater
recharge zones. Finally, evapotranspiration, surface runoff,
and precipitation must all be well constrained, which is gen
erally not the case except in very well-studied areas (Taniguchi
et al., 2003).
Hydrogeologic modeling. Hydrogeologic modeling using
numerical simulations can be used both to estimate the net
freshwater discharge from an aquifer to the ocean (Oki, 1999;
Kaleris et al., 2002; Langevin, 2003) and to investigate specific
processes in the subterranean estuary, such as wave- and tidedriven flows (Turner et al., 1997; Horn, 2002; Robinson et al.,
2007). Numerical SGD simulations are typically performed
210
Groundwater inputs
Characteristics:
Coastal ocean control volume (Box)
- Groundwater radium
activity (Ragw)
Characteristics:
- Submarine ground
water discharge flux
(Fsgd)
Opean ocean
Characteristic:
- Volume (Vbox)
- Radium activity (Raco)
Other inputs (river,
diffusive sediment
flux, etc;)
Offshore radium
activity (Raos)
- Residence time (Tr)
Rax = Raco − Raos
Fsgd =
Rax
Vbox
×
Ragw
Tr
Figure 5 Schematic diagram of the mass-balance, or box model, method of calculating submarine groundwater discharge fluxes using natural
groundwater tracers.
using MODFLOW, a modular, finite-difference groundwater
flow model that was developed in the 1980s by the United
States Geological Survey and is available for free. Some studies,
especially those focusing on the groundwater–seawater inter
face, use SEAWAT, which, unlike MODFLOW, allows modeling
of variable-density, transient flow of the type that would be
expected within the subterranean estuary.
The main challenge of hydrogeologic modeling is obtaining
representative values for the input parameters, which include
hydraulic heads, hydraulic conductivities, and boundary con
ditions. All these parameters can vary considerably in space and
time, and it can be difficult to obtain sufficient field data to
characterize the variability (Burnett et al., 2006). Like the waterbalance approach, hydrogeologic modeling is best suited to
regional or basin-scale studies and is not well suited to captur
ing fine-scale spatial variability in SGD.
Intercomparison experiments. Because several very different
methods have been used to measure SGD fluxes, it is important
to understand how the results of these methods compare. To
address this issue, the International Atomic Energy Agency
(IAEA) and the United Nations Educational, Scientific and
Cultural Organization (UNESCO) sponsored SGD intercom
parison experiments at five sites: Cockburn Sound, Australia;
Donnalucata, Sicily; Shelter Island, New York; Ubatuba, Brazil;
and the island of Mauritius, off the southwest coast of Africa.
These five sites represent a range of hydrogeological settings,
including a sandy coastal plain, a carbonate aquifer with abun
dant springs, a fractured crystalline rock aquifer, and a volcanic
island, and they were all sites where SGD had been documen
ted previously (Burnett et al., 2006). All five studies included
radon and seepage meters, as well as different combinations of
other methods. Four included radium isotopes, two included
geophysical studies to map the flow field in the coastal aquifer,
one included artificial tracers, one included numerical model
ing, and one included a water-balance approach. An additional
intercomparison experiment was carried out in Apalachee Bay,
Florida, and it included seepage meters (Taniguchi et al., 2003),
continuous radon monitoring (Lambert and Burnett, 2003),
radium isotopes (Moore, 2003), and numerical modeling
(Smith and Zawadzki, 2003). The results of these six intercom
parison studies are summarized in Table 1.
Good agreement between methods was observed at some
locations, while at others there were considerable discrepan
cies. At Cockburn Sound, good agreement between manual
seepage meters, radon, and radium isotopes was observed. At
Donnalucata, seepage meters, radium, and radon yielded dif
ferent results (Table 1). Seepage meters yielded the lowest
fluxes, and those estimated from radon were somewhat
higher, although of a similar magnitude. Radium estimates
were the highest of all, which the authors attributed to the fact
that radon and seepage meter sampling was restricted to a
small protected basin, while radium sampling took place in
a larger area up to several kilometers from shore. The basin
may have had fewer springs relative to the larger area and/or
may have been less influenced by wave-driven recirculated
seawater flow.
In Shelter Island, SGD estimates from dye-dilution seepage
meters agreed well with those calculated using radon and
radium. However, a high degree of variability was seen between
seepage meters of different types (manual, continuous heat,
heat pulse, dye-dilution, and ultrasonic). This was attributed
to the positioning of individual seepage meters within the
study area; specifically, those recording higher flow rates were
located closer to a pier with pilings piercing an aquitard
(Burnett et al., 2006). In Ubatuba, all methods produced a
similar range of results, but this range was very large because
the fractured rock aquifer produced spatially heterogeneous
flow. A similar situation was observed in Mauritius, where
radon, seepage meter, and water-balance estimates all fell
within the same wide range (Table 1). Discharge from this
volcanic aquifer was characterized by many springs, which
served as point sources of SGD. Finally, in Apalachee Bay,
discrepancies between results from seepage meters, chemical
tracers, and modeling led the authors to conclude that pressuredriven seawater circulation through the coastal aquifer was an
important contributor to SGD.
Table 1
Results of intercomparison experiments, where different methods were used to quantity submarine groundwater discharge at the same location
Estimated groundwater discharge
Seepage meters
Natural tracers
Artificial
tracers
Water
balance
Numerical
modeling
Site
Units
References
Dates
Manual
Heat
pulse
Continuous
heat
Dye-diffusion Ultrasonic
Radium Radon
Cockburn Sound,
Western Australia
Donnalucata, Sicily
(Basin)
Donnalucata, Sicily
(Shoreline flux)
Shelter Island, New
York, USA
Ubatuba, Sao Paulo
State, Brazil
m3 m−1
day−1
m3 day−1
Burnett et al. (2006),
Loveless (2007)
Burnett et al. (2006),
Moore (2006),
Taniguchi et al.
(2006)
Burnett et al. (2006)
25 November to 6
December 2000
18–24 March 2002
2.5–3.7
-
-
-
-
3.2
2.0–2.7
-
-
2.5–4.8
300–
1000
10–30
-
-
-
-
-
-
-
-
-
-
-
-
1000
1200–
7400
30–200
-
-
-
11.5
0.4–0.8
2.5
3.4
17.5
16–26
8–20
-
0.5
10
Mauritius Islands
(Spring)
Mauritius Islands
(South Beach)
Mauritius Islands
(Klondike Hotel)
Apalachee Bay,
Florida, USA
m3 m−2
day−1
m3 m−1
day−1
cm day−1
cm day−1
m3 min−1
18–24 May 2002
Burnett et al. (2006), 16–22 November 2003 5–270
Moore and de
Oliveira (2008),
Stieglitz et al. (2008)
Burnett et al. (2006)
19–26 March 2005
1–490
-
0–360
2–109
-
-
1–29
28–184
-
-
-
-
-
-
-
65–140
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-
2.5–22
-
-
-
-
-
13–23
-
-
-
-
-
-
5–28
-
-
14–24
-
-
-
1.6–2.5
-
Similar to manual -
1.7–2.5
-
-
7.5–86
Lambert and Burnett
(2003), Moore
(2003), Smith and
Zawadzki (2003),
Taniguchi et al.
(2003)
13–18 August 2000
Similar to manual 1.5
212
4.08.1.3
Summary
SGD studies have been completed in numerous locations
around the world, although some areas, including Southeast
Asia, Africa, and South America, have not yet been well studied.
At each of these locations, evidence of some type of SGD, be it
freshwater or recirculated seawater, has been observed. This
indicates that SGD is a pervasive phenomenon on a global
scale and that it is likely to affect water chemistry and quality
in coastal areas, particularly those where other hydrologic con
nections between land and sea, such as rivers, are less important.
Relatively few studies of SGD at high latitudes have been
completed (Taniguchi et al., 2002), perhaps because aquifers in
such locations are frozen for much of the year and access is
difficult. The few studies that have been done (Hay, 1984;
Deming et al., 1992; Piekarek-Jankowska, 1996) report low
SGD compared to temperate or tropical locations, with the
exception of Vanek and Lee (1991) who measured a flux of
75 m yr−1 in Laholm Bay, Sweden. However, the magnitude of
SGD in polar regions could increase if these areas thaw due to
global warming.
The most common methods used to measure or estimate
SGD fluxes are seepage meters, radon, radium isotopes, water
balances, and numerical modeling. Because SGD can be highly
heterogeneous on a small spatial scale, methods that have
different scales of applicability can produce different results.
For example, a seepage meter provides a point measurement of
SGD, whereas a water balance provides a basin-level estimate.
While estimates of SGD made using different methods some
times agree well, in other cases large discrepancies exist. These
discrepancies may shed light on gaps in our knowledge and
lead to a more sophisticated understanding of the system.
The greatest challenges in SGD research today are the
following:
1. reducing the uncertainty in SGD estimates so that estimates
from different locations and times can be compared to each
other and to other inputs, such as those from rivers;
2. understanding how and why SGD measurement methods
differ in order to choose the best method for a particular
situation; and
3. understanding the importance of SGD on a worldwide scale,
including spatial variability.
the coastal aquifer are expected to occur at virtually any coastal
setting. However, both the magnitude and the chemical com
position of SGD can vary greatly in space and time. SGD fluxes
reported from different locations around the world range from
less than one to over 400 cubic meters per square meter of
seafloor per year (m3 m−2 yr−1 or m yr−1), with typical values in
the range of 40 m yr−1 (Taniguchi et al., 2002).
4.08.2.1
Fresh versus Saline Groundwater Discharge
Burnett et al. (2003) identified three main components of
SGD: meteoric or freshwater (Figure 6), recirculated seawater,
and connate groundwater, which has a salinity greater than that
of seawater because of the dissolution of salts within the aqui
fer. Coastal SGD studies typically focus on fresh SGD and/or
saline flow resulting from seawater recirculation; connate water
discharge at most coastal settings is much less common.
Although fresh and saline SGD often occur concurrently,
they can be distinguished from each other using chemical
tracers, and the specific contribution of each SGD type to the
total flux can be calculated. Fresh SGD can be identified by its
low salinity signature; by definition, any depression in salinity
below the offshore seawater value indicates the presence of
freshwater at the coast. Radon (Rn) is commonly used as a
tracer of total SGD, because groundwater of any salinity will
attain an elevated Rn activity through contact with the aquifer
substrate (Dulaiova et al., 2008). In contrast, high radium (Ra)
activities are observed in brackish or saline, but not fresh,
groundwater, because of the salinity-dependent desorption
behavior of Ra (Webster et al., 1995). Ra is thus a good tracer
of SGD in coastal areas where fresh groundwater mixes with
higher-salinity seawater in the coastal aquifer prior to discharge
or where SGD is predominantly saline (Moore, 1999). The
relative contributions of various endmembers, including fresh
groundwater, saline groundwater, and unaltered seawater, to a
given sample can be presented visually using mixing diagrams
(Figure 7).
The relative magnitudes of fresh and saline SGD can vary
considerably from site to site, with fresh SGD typically consti
tuting up to 30% of the total flux. Li et al. (1999) showed that
Recently, a number of coordinated research efforts including the
International Atomic Energy Agency/United Nations
Educational, Scientific and Cultural Organization (IAEA/
UNESCO) and GEOTRACES (SCOR Working Group, 2007)
intercomparison experiments, and special issues of journals
such as Marine Chemistry (vol. 109) and Estuarine, Coastal and
Shelf Science (vol. 76) have sought to identify and address these
challenges, and we can expect more such efforts in the future as
the field progresses and matures.
4.08.2 Flux of Water into the Coastal Ocean from SGD
SGD is a pervasive phenomenon that occurs to a measurable
extent at every location where it has been studied. This is not
surprising, given that the seaward flow of fresh groundwater
and/or tide- and wave-driven recirculation of seawater through
Figure 6 A submarine freshwater spring discharging off the Florida
coast is visible at the sea surface. However, most submarine groundwater
discharge is diffuse and invisible. Used by permission of the St. Johns
River Water Management District.
213
5000
Oct 2001 lagoon
July 2002 lagoon
Mar 2003 lagoon
All middle springs
All wells
226Ra
(dpm per 100 l)
4000
3000
2000
Drowned lagoon well
and harbor well
Seawater
1000
0
10
20
30
40
Salinity
Figure 7 A three-endmember mixing diagram from Celestún Lagoon, Mexico. The three endmembers are indicated with red squares. Reproduced with
permission from Young, M.B., Gonneea, M.E., Fong, D.A., Moore, W.S., Herrera-Silveira, J., Paytan, A., 2008. Characterizing sources of groundwater to a
tropical coastal lagoon in a karstic area using radium isotopes and water chemistry. Marine Chemistry 109, 377–394. Copyright (2008), with permission
from Elsevier.
at some sites saline SGD resulting from seawater recirculation
could comprise up to 96% of the total SGD flux, with only 4%
coming from fresh SGD. Similar results were observed by
Shellenbarger et al. (2006) in the Gulf of Aqaba, Israel. Over
a 2-year monitoring period in Kahana Bay, Oàhu, Hawaiì,
about 16% of total SGD was fresh (Garrison et al., 2003), and
at some sites on the Kona Coast of Hawaiì Island, fresh
groundwater made up over 40% of the nearshore water volume
(Knee et al., 2010). Freshwater constituted up to 29% of total
SGD input to Osaka Bay, Japan (Taniguchi et al., 2005), and
10% of total SGD input to Chesapeake Bay (Hussain et al.,
1999).
4.08.2.2
Temporal Variability in SGD
At many sites, temporal variation in SGD fluxes at the semidiurnal, diurnal, fortnightly, and seasonal timescales has been
observed. Semidiurnal and diurnal variations in SGD result
from changes in sea-surface height between high and low
tides. When the sea-surface height is lower, at low tide, the
hydraulic gradient between land and sea is steeper, increasing
seaward groundwater flow (Taniguchi, 2002), and most studies
where SGD has been measured at both high and low tides have
recorded this effect (e.g., Robinson et al., 1998; Garrison et al.,
2003; Lambert and Burnett, 2003; Shellenbarger et al., 2006).
For example, at Shelter Island, New York, SGD flux was
approximately 30 times greater at low tide than at high tide
(Sholkovitz et al., 2003). A lesser, but still significant, degree of
variability was observed in Waquoit Bay, Massachusetts, where
SGD fluxes into nearshore waters were about 4 times greater at
low tide than at high tide (Michael et al., 2003).
Spring–neap tidal variability on a fortnightly timescale also
affects SGD. A spring tide occurs when the tidal range, or
difference between daily high and low tide levels, is greatest,
whereas a neap tide occurs when the tidal range is least. While
the mean sea-surface height does not differ between spring and
neap tides, the greater range at spring tides causes heightened
tidal pumping though the subterranean estuary. A number of
studies (e.g., Kim and Hwang, 2002; Taniguchi, 2002; Boehm
et al., 2004; Rapaglia, 2005; de Sieyes et al., 2008) have demon
strated that total SGD is significantly higher during spring tides.
For example, at one site in Venice Lagoon, the SGD flux during
spring tides was 60% greater than during neap tides (Rapaglia,
2005), and in Tolo Harbor, Hong Kong, SGD flux during spring
tides was over 100% greater than during neap tides (Tse and
Jiao, 2008).
In addition, SGD fluxes vary seasonally due to differences in
water table elevation resulting from variation in precipitation
and evapotranspiration. SGD, particularly the fresh compo
nent, is often greatest during or after the rainy season, when
fresh groundwater is recharged (e.g., Lewis, 1987; Tobias et al.,
2001; Young et al., 2008). Lewis (1987) found that SGD in
Barbados was approximately twice as great during the rainy
season, and in Ubatuba, Brazil, SGD was 21–67% higher in
the summer than during other times of year (Oliveira et al.,
2006).
Sometimes, however, seasonal SGD variations are out of
phase with precipitation, most likely as a result of the time it
takes for water to flow from the recharge location to the
214
coastline. Moreover, seasonal variations have also been
observed for saline SGD; this, however, cannot be explained
as a direct effect of precipitation, which would only affect fresh
groundwater. Michael et al. (2005) proposed the idea that
seasonal variation in precipitation and evapotranspiration
causes the freshwater–saltwater boundary to shift back and
forth with the seasons, drawing saltwater into the aquifer at
some times of year and discharging it at others. This would
explain lags between changes in precipitation and SGD fluxes,
seasonal variation in the saline SGD, and the fact that saline
SGD fluxes at times are too large to be explained by tides and
waves alone (Michael et al., 2005).
Not only does the quantity of SGD vary over tidal and
seasonal timescales, but the chemical makeup of the dischar
ging water does as well. In Celestún Lagoon, Mexico, the
relative contributions of fresh and brackish SGD to the total
flux varied both over the daily tidal cycle and between
seasons (Young et al., 2008). As a result, SGD-related fluxes
of chemicals also showed seasonal variations. De Sieyes
et al. (2008) found that total SGD flux at Stinson Beach,
California, was lower at neap tide than at spring tide; how
ever, fresh SGD flux was almost 10 times higher at neap tide.
Fresh groundwater at Stinson Beach had much higher con
centrations of silicate, dissolved inorganic nitrogen (DIN),
and soluble reactive phosphate (SRP) than did saline
groundwater; hence, during neap tides greater quantities of
these nutrients were delivered to the surf zone. The greater
fresh groundwater flux caused surf zone concentrations of
silicate, DIN, and SRP to be 14%, 35%, and 27% higher,
respectively, at neap tide than at spring tide. Campbell and
Bate (1998) also reported greater fluxes of fresh SGD and
nutrients during neap tides in South Africa.
considered, SGD is greatest near the shoreline at the land–sea
interface and decreases with distance offshore.
Studies where multiple seepage meters have been installed
at one site have shown that SGD can be highly variable over
small spatial scales (meters to tens of meters), and, in many
cases, there is no obvious explanation for this variation. For
example, Michael et al. (2003) found that SGD fluxes mea
sured by seepage meters installed <2 m apart differed by more
than a factor of 7, and that the standard deviation within a
‘cluster’ of seepage meters located within a 2 m by 2 m area was
50–100% as great as for a much larger 60 m transect. SGD can
also vary considerably between sites within the same region.
For example, total SGD fluxes measured along the coastline of
the approximately 1800 km2 Jeju Island in S. Korea ranged
from 50 to 300 m yr−1, and the proportions of fresh and saline
groundwater also differed greatly between sites (Kim et al.,
2003). Street et al. (2008) also documented a great deal of
variability in fresh and saline SGD along the leeward coasts of
three Hawaiian Islands.
Several specific site features are known to affect SGD and
result in a heterogeneous flux distribution. Dulaiova et al.
(2006a) found that SGD was enhanced by approximately an
order of magnitude near the mouth of the Chao Phraya River in
Thailand compared to other sites nearby. Coastal ponds or
unconsolidated bluffs can create a steep hydraulic gradient
nearshore, driving SGD (Bokuniewicz et al., 2003). Natural
embayments can focus groundwater flow (Bokuniewicz et al.,
2003), and the excavation of artificial harbors can cut through
previously confined aquifer layers or preferential flow paths,
increasing SGD (Gallagher, 1980; Knee et al., 2010). In addi
tion, other human constructions, such as dock pilings, can
puncture confining layers and cause increased local SGD flux
(Burnett et al., 2006).
4.08.2.3
4.08.2.4 Comparison of SGD Fluxes to Input from Other
Water Sources
Spatial Variability in SGD
Numerical modeling of SGD from a homogeneous aquifer has
predicted that fresh SGD should decrease exponentially with
distance offshore (McBride and Pfannkuch, 1975; Fukuo and
Kaihotsu, 1988; Taniguchi et al., 2003), and most field data
show the expected pattern (e.g., Attanayake and Waller, 1988;
Bokuniewicz et al., 2004; Paytan et al., 2004). However, at
some sites SGD does not decrease as expected with distance
from shore (e.g., Connor and Belanger, 1981; Cable et al.,
1997; Bokuniewicz et al., 2004). SGD fluxes may be greater at
some distance offshore than at the shoreline for several reasons.
Aquifer heterogeneity can generate preferential flow paths that
happen to discharge offshore (Bokuniewicz et al., 2004). In
karst systems, the flow of groundwater through underground
channels can cause a positive feedback loop where the ground
water itself alters the aquifer geochemistry and promotes
channel development and discharge at channel openings
located at varying distances offshore (Evans and Lizarralde,
2003). Even in a homogeneous aquifer, the interplay of forcing
from fresh groundwater recharge and seawater recirculation
through the coastal aquifer can lead to the development of a
freshwater discharge ‘tube’, which causes a flux distribution
that is not a simple function of distance offshore. If the wave
amplitude is large, this tube can discharge a considerable dis
tance offshore (Robinson et al., 2007). However, in general,
and especially when scales of kilometers rather than meters are
Many authors have attempted to put SGD fluxes in context by
comparing them to fluxes from rivers and streams, and some
attempts to estimate SGD fluxes on a global scale have also
been made. In the Chesapeake Bay region, total SGD into two
estuaries made up 0.6–10% of river fluxes into the Bay
(Charette and Buesseler, 2004). Garrison et al. (2003) esti
mated that the total SGD flux into Kahana Bay, Oàhu,
Hawaiì, was 9.0 106 l day−1, similar to the discharge of the
nearby Kahana River; approximately 16% of the SGD flux was
freshwater. In contrast, on the north shore of Kauaì, total SGD
represented only 2–10% of river flow (Knee et al., 2008). Most
estimates of global fresh SGD flux range from 0.3% to 16% of
fresh surface water inputs (Burnett et al., 2003). The magnitude
of SGD fluxes observed in regional studies is often higher than
estimates of freshwater recharge by precipitation in the respec
tive watersheds, likely due to the inclusion of recirculated
seawater in total SGD flux estimates (Slomp and Van
Cappellen, 2004). Another explanation is that SGD studies
are often conducted in locations that are known or suspected
to have significant SGD fluxes; thus, they may neglect sites in
the same region where discharge is low or negligible.
Although both fresh and total SGD fluxes are typically lower
in magnitude than river fluxes at locations where rivers are
present, SGD occurs more continuously in space (i.e., all
215
along the coastline), and integration over long coastlines trans
lates into considerable fluxes. In certain settings, such as karst
terrains (Moore and Shaw, 1998; Taniguchi et al., 2003; Young
et al., 2008) and volcanic islands (Lewis, 1987; Kim et al.,
2003; Street et al., 2008; Knee et al., 2010), SGD can be the
most important, or even the only, hydrologic connection
between land and sea. In addition, groundwater nutrient con
centrations can be much higher than those in rivers and
streams, so that even a relatively small flux can have a large
impact on nutrient subsidies (Valiela et al., 1990; Uchiyama
et al., 2000; Hwang et al., 2005b). Even in locations where the
salinity of SGD is not noticeably lower than that of the receiv
ing seawater (e.g., Boehm et al., 2006; Shellenbarger et al.,
2006), saline groundwater can provide important nutrient sub
sidies because of pollution or nutrient leaching from the
aquifer substrate.
4.08.3 Nutrient Loads from SGD
SGD has long been suspected of providing nutrient subsidies to
the coastal ocean; however, until recently, scientists have lacked
the means to measure these subsidies. Groundwater usually
has higher nutrient concentrations than the receiving seawater,
and sometimes these concentrations are also significantly
higher than those in rivers or streams. For example, at Hasaki
Beach in Japan, nitrate concentrations were about six times
higher in shallow groundwater than in water from the nearby
Tone River, and nitrite, phosphate, and silica concentrations
were also higher in groundwater than in the river (Uchiyama
et al., 2000). In Chesapeake Bay, groundwater nutrient concen
trations were up to two orders of magnitude higher than those
found in estuarine surface waters (Charette and Buesseler,
2004). Because groundwater nutrient concentrations can be
so high compared to other potential sources, even a relatively
small volumetric flux could provide significant nutrient sub
sidies. Excess nutrient additions to coastal waters can lead to
unwanted outcomes, such as algal blooms (Paerl, 1997;
Anderson et al., 2002), red tides (Hodgkiss and Ho, 1997;
Townsend et al., 2001; see Figure 8), and damage to coral
reefs (Hallock and Schlager, 1986; Parsons et al., 2008; see
Figure 9). Thus, in order to prevent or mitigate these outcomes,
it is necessary to determine the sources of nutrients to coastal
areas and to quantify contributions from each source.
Quantifying SGD-related nutrient fluxes would be an essential
component of this analysis in many areas.
4.08.3.1
Nutrient Concentrations in Coastal Groundwater
Nitrogen (N) and phosphorus (P) are generally considered the
primary limiting nutrients in coastal and estuarine systems,
although there is some debate about which of the two is
the ultimate limiting nutrient, whether nutrient limitation can
change seasonally, and whether some systems may be
co-limited by both N and P (Howarth, 1988). Silicon limita
tion can also occur (Rocha et al., 2002; Gilbert et al., 2006),
especially for diatoms (Paasche, 1973; Gobler et al., 2006).
Thus, studies of nutrient fluxes into coastal waters typically
focus on various inorganic nitrogen species (nitrate, nitrite,
and ammonium) or total DIN, as well as phosphate or dis
solved inorganic phosphorus and silicate.
Figure 8 A red tide. Nutrient additions to coastal waters from SGD have
been associated with red tides.
Nitrate (NO3−) is the most common groundwater pollutant
worldwide (Spalding and Exner, 1993) and it is also the form
of inorganic nitrogen that is most prevalent under freshwater,
oxic conditions. It can enter groundwater from several sources,
both natural and anthropogenic. The main natural source of
nitrate is nitrogen fixation, in which relatively inert nitrogen gas
from the air (N2) is converted biologically to ammonia (NH3)
and organic nitrogen, which, in turn, are converted to nitrate by
nitrifying bacteria in the soil. Some species of plants, micro
organisms, and cyanobacteria are capable of nitrogen fixation.
The majority of nitrogen-fixing plants are in the legume family
(Fabaceae), and their ability to fix nitrogen comes from a
symbiotic association with rhizobia, bacteria that live in their
root nodules. This nitrogen then enters the soil and undergoes
various transformations there. In pristine environments where
natural soil nitrogen is the only source of nitrate to ground
water, groundwater nitrate concentrations are typically less
than 2 mg l−1 (Mueller et al., 1996; Burkart et al., 2002); how
ever, in certain areas, such as the Sahel region of Africa,
naturally occurring groundwater nitrate concentrations can be
much higher (Edmunds and Gaye, 1997).
In comparison to pristine settings, areas affected by human
activities usually have much higher groundwater nitrate concen
trations. These concentrations can originate from point sources,
such as sewage leaks, chemical facilities, or animal feedlots
(Gardner and Vogel, 2005; Wakida and Lerner, 2005).
However, they more commonly originate from nonpoint
sources and local or regional land uses. Numerous studies
(e.g., Carpenter et al., 1998; Nolan, 2001; Choi et al., 2007)
have linked high population densities and urban development
to elevated groundwater nitrate concentrations. For example, the
average groundwater nitrate concentration in the Indonesian
megacity of Jakarta was almost 15 mg l−1, and in densely
216
(a)
(b)
Figure 9 A healthy coral reef (a) and one that has been overgrown by
macroalgae (b). Macroalgal overgrowth of coral reefs has been associated
with nutrient inputs from submarine groundwater discharge. The data
used in this effort were acquired as part of the activities of NASA’s Science
Mission Directorate, and are archived and distributed by the Goddard
Earth Sciences (GES) Data and Information Services Center (DISC).
populated residential areas in Korea, groundwater nitrate con
centrations reached over 200 mg l−1 (Choi et al., 2007). For
reference, drinking water in the United States is considered safe
if its nitrate concentration is <10 mg l−1. Agricultural land use,
including crop and livestock production, is also associated
with high groundwater nitrate concentrations. In Korea,
cropping–livestock areas had groundwater nitrate concentrations
as high as 160 mg l−1 (Choi et al., 2007), and some agricultural
areas in Connecticut had groundwater nitrate concentrations as
high as 265 mg l−1 (Hamilton and Helsel, 1995).
Nitrite and ammonium concentrations in coastal, fresh,
surficial aquifers are typically much lower than those of
nitrate (Slomp and Van Cappellen, 2004). For example, in
groundwater from a Japanese coastal aquifer, nitrate con
centrations reached over 20 mg l−1, while nitrite
concentrations were generally <1 mg l−1 with a maximum
of 5, and ammonium concentrations never exceeded 0.5 mg l−1
(Uchiyama et al., 2000). However, in some anaerobic or sub
oxic aquifers, ammonium can be the dominant form of
inorganic nitrogen (Joye et al., 2006), especially if ammo
nium-rich wastewater is released directly into the anoxic
zone or if the decomposition of nitrogen-containing organic
matter occurs under anoxic conditions (Slomp and Van
Cappellen, 2004). Saline groundwater sometimes also has
high ammonium concentrations (Talbot et al., 2003;
Santoro et al., 2006).
Dissolved inorganic phosphorus, typically in the form of
phosphate (PO43−), can enter groundwater from dissolution of
the minerals making up the aquifer substrate, decomposition
of organic matter in soils, or anthropogenic activity. Similar to
nitrate, phosphate concentrations in groundwater are often
considerably higher than in surface waters. For example, in
Tomales Bay, California, groundwater phosphate concentra
tions reached 5.3 mg l−1, while the highest surface water
concentrations were <0.1 mg l−1 (Oberdorfer et al., 1990).
However, at other locations, such as Hanalei Bay, Kauaì
(Knee et al., 2008), groundwater and surface water phosphate
concentrations were similar.
Phosphate in groundwater often forms insoluble, inorganic
compounds that sorb to rock and particle surfaces (Wong et al.,
1998). Thus, a larger proportion of this nutrient is retained by
the aquifer substrate, and less is dissolved in groundwater and
carried away as SGD. However, under certain circumstances,
such as in sandy soils with low phosphate sorption capacity or
when very high phosphate loads are added, phosphate can be
transported by groundwater flow (Peaslee and Phillips, 1981;
Wong et al., 1998). Two main sources of high anthropogenic
phosphate loads are wastewater or sewage plumes and inor
ganic fertilizers. Near an old septic system in Ontario, Canada,
phosphate concentrations were up to 200 times greater than in
uncontaminated groundwater (Harman et al., 1996), and
groundwater phosphate concentrations in Pennsylvania crop
land were 5–7 times higher than in nearby forested land
(Pionke and Urban, 1985).
Silicon dissolved in groundwater exists as Si(OH)4, com
monly referred to as silica. The major source of silica to
groundwater is the rocks, sediments, and soils making up the
aquifer substrate. Silica concentrations in groundwater are less
variable than those of nitrate, phosphate, or other dissolved
constituents, with typical values of about 17 mg l−1 (Davis,
1964), but varying somewhat based on rock type. Silica does
not sorb to aquifer surfaces or react with other groundwater
constituents (Haines and Lloyd, 1985). In contrast to nitrate
and phosphate, the concentrations of which are influenced
both by anthropogenic sources and by groundwater chemistry,
groundwater silica concentrations are mainly determined by
the mineral characteristics of the aquifer substrate (Davis,
1964) and how long the water has been in contact with that
substrate (Haines and Lloyd, 1985).
4.08.3.2
SGD-Related Nutrient Fluxes into Coastal Waters
The simplest approach to estimating SGD-related nutrient fluxes
to coastal waters is to estimate the SGD flux using one or more of
the methods described in Section 4.08.2, determine the nutrient
concentrations of the discharging groundwater endmember, and
multiply the two together to calculate a nutrient flux. This
approach has been used in numerous studies (e.g., Valiela and
Costa, 1988; Krest et al., 2000; Kelly and Moran, 2002; Paytan
et al., 2004) over the past two decades.
However, the assumption that nutrients behave conserva
tively in the subsurface, which is inherent in this approach, is
not always valid. Over the course of decades-long flow paths
from inland recharge areas to the sea, nitrate removal from
groundwater via microbial denitrification can occur, as seen
in a study of the coastal plains in the southeastern United States
(Tesoriero et al., 2005). Patterns of inorganic nitrogen species
distribution in groundwater collected along a salinity gradient
at Waquoit Bay, Massachusetts, also suggested nitrate removal
via denitrification (Talbot et al., 2003). In addition, a recent
study by Kroeger and Charette (2008) revealed significant
attenuation of both nitrate and ammonium in fresh ground
water, with both species being reduced to approximately 10%
of their original concentrations through various microbial pro
cesses in the aquifer.
Groundwater phosphate concentrations can also be reduced
by biological uptake, sorption, and precipitation occurring
along the flow path (Corbett et al., 2002). Silica is generally
conservative in groundwater (Davis, 1964; Haines and Lloyd,
1985; Corbett et al., 2002), although its concentration
increases with the length of time groundwater has spent in
contact with aquifer materials up to an equilibrium concentra
tion, and it can also be decreased through dilution with lowersilica water. The nonconservative behavior of nitrogen and
phosphorus in some coastal aquifers indicates the importance
of characterizing the chemical transformations these nutrients
undergo as groundwater flows seaward in order to more accu
rately estimate SGD-related nutrient fluxes. Alternatively,
nutrient concentrations can be measured directly at the dis
charge point, after biogeochemical transformations within the
aquifer have already occurred.
Both SGD rates and groundwater nutrient concentrations
vary widely, leading to an even greater degree of variability in
their product, the SGD-derived nutrient flux. In their review
paper, Slomp and Van Cappellen (2004) documented
SGD-related nitrogen and phosphorus fluxes ranging over
several orders of magnitude. Nitrogen fluxes ranged from
160 μmol m−2 day−1 at Kahana Bay, Hawaiì (Garrison
et al., 2003) to 72 000 μmol m−2 day−1 at Nauset Marsh
estuary, Massachusetts (Portnoy et al., 1998). Likewise,
phosphorus fluxes ranged from <1 μmol m−2 day−1 at one
site in the Florida Keys (Corbett et al., 1999) to 900 μmol
m−2 day−1 at North Inlet, South Carolina (Krest et al., 2000).
The relative importance of groundwater-borne nutrient
loads, compared with other sources, is also highly variable. At
some sites with little or no surface flow, SGD is the only path
way by which terrestrial nutrients are transported to the sea.
These sites include volcanic islands such as Hawaiì (Kay et al.,
1977; Oki, 1999) and Jeju Island, Korea (Kim et al., 2003),
karst environments such as the Yucatan Peninsula of Mexico
(Bokuniewicz et al., 2003; Young et al., 2008), Florida (Brooks,
1961; Simmons, 1992; Corbett et al., 1999), and Jamaica
(D’Elia et al., 1981), and arid locations such as Eilat, Israel
(Shellenbarger et al., 2006). In some locations where rivers
are present, groundwater nutrient loads are comparable to, or
greater than, river-borne loads (e.g., Crotwell and Moore, 2003;
Garrison et al., 2003; Burnett et al., 2007); however, in other
locations, SGD is likely to be a negligible contributor to the
nutrient budget (Burnett et al., 2001).
4.08.3.3 Ecological Importance of Groundwater-Borne
Nutrients
Nutrients delivered to estuaries, bays, and coastal waters by
SGD can have important ecological effects. One such effect is
red tides (Figure 8), also known as harmful algal blooms
217
(HABs). Hu et al. (2006) observed that nutrient loads from
rivers were insufficient to support a HAB in coastal waters off
west-central Florida, and suggested that SGD was the most
likely source of the additional nutrients. SGD was also identi
fied as the most likely source of nutrients supporting recurrent
HABs in Yeoja Bay (Hwang et al., 2005a) and Masan Bay (Lee
et al., 2009) in South Korea. In the case of Masan Bay, the silica
concentrations in SGD, which were 2–3 times higher than in
surface waters, were instrumental in promoting the diatom
bloom (Lee et al., 2009). Brown tides, or blooms of the uni
cellular microalgae Aureococcus anophagefferens, have also been
associated with SGD and have caused considerable harm to the
scallop industry in waters off Long Island, New York
(Bokuniewicz et al., 2003).
Another potential negative consequence of SGD-related
nutrient subsidies is eutrophication. Eutrophication occurs
when high nutrient concentrations allow phytoplankton to
bloom. Excessive phytoplankton growth blocks light from
reaching deeper parts of the water column, killing aquatic
plants and algae below. When the aquatic plants, algae, and
the phytoplankton themselves die, their decomposition
depletes the water of oxygen, which can lead to hypoxic or
even anoxic conditions dangerous for fish and other aquatic
life. In Bangdu Bay on the volcanic Jeju Island, SGD provided
the majority of nutrients and was determined to be the cause of
benthic eutrophication (Hwang et al., 2005b). Discharge of
high-nutrient groundwater was also implicated in eutrophica
tion in the Florida Keys (Lapointe and Clark, 1992) and in Tolo
Harbor, Hong Kong (Tse and Jiao, 2008). A feature shared by
most settings where SGD-related eutrophication has occurred is
their enclosed or semi-enclosed nature, which restricts circula
tion and exchange with lower-nutrient offshore seawater.
Finally, SGD-borne nutrients have been associated with
damage to coral reefs and ‘phase shifts’ from coral-dominated
to macroalgae-dominated reef areas (Figure 9). On Reunion
Island in the Indian Ocean, high-nutrient SGD was associated
with coral death and macroalgal blooms on fringing reefs
(Naim, 1994). Similar results have been observed in Florida
(Lapointe and Matzie, 1996), Hawaiì (Parsons et al., 2008),
and Brazil (Costa et al., 2000).
4.08.3.4
Several decades of scientific investigation have shown that SGD
is an important nutrient source to estuarine and coastal areas,
and that in many locations it is impossible to balance coastal
nutrient budgets without accounting for SGD. However, a step
that has not yet been taken is the detailed estimation or mod
eling of SGD-related nutrient inputs to coastal waters at a
global scale. A more global view of SGD is important both for
understanding global nutrient budgets and for assessing how
specific sites fit into more general trends. In addition, such
models can be used to predict which areas might be susceptible
to future climatic or other environmental change impacts.
Recently, researchers in the Land–Ocean Interactions in the
Coastal Zone (LOICZ) project have begun working on a typo
logical approach to SGD. This approach involves the
identification of factors that influence SGD quantity, such as
topography, precipitation, aquifer thickness, hydraulic gradi
ent, anisotropy, and the fractal dimension of the coastline
(Bokuniewicz, 2001). Then, sections of the world’s coastline
218
are classified in terms of these characteristics so that zones of
high and low SGD can be predicted (Bokuniewicz et al., 2003).
Global modeling of nutrient transport by groundwater into the
coastal ocean is also underway (Dürr et al., 2008) although, at
the time of writing, the results of this work have not yet been
published.
Other active areas of study in this field include estimating
SGD-related nutrient fluxes at previously unstudied locations
(e.g., Lee et al., 2009; Leote et al., 2008; Tse and Jiao, 2008),
investigating biological processes that affect nutrient concen
trations in groundwater (Kroeger and Charette, 2008), and
investigating how natural processes such as tides (de Sieyes
et al., 2008) and storm events (Smith et al., 2008) can affect
the quantity and quality of SGD. Links between SGD-related
nutrient fluxes and land use are also being studied (e.g., Knee
et al., 2010). As land uses associated with groundwater nutrient
contamination continue to proliferate, the accurate estimation
of SGD-related nutrient fluxes is likely to become even more
important.
4.08.4 Metal Loads from SGD
In addition to providing macronutrients, such as nitrate, phos
phate, and silica, to coastal waters, SGD can also be a source of
dissolved metals. These metals may serve as micronutrients
necessary for coastal ecosystems to function; however, they
may also be pollutants that interfere with ecosystem functions
or pose a threat to human health. Due to its prolonged contact
with the metal-containing aquifer substrate, groundwater
would be expected to have high dissolved metal concentrations
relative to surface waters, at least for the more soluble metals or
in areas where the subterranean estuary is characterized by
conditions favorable for metal dissolution. As it has been
shown throughout the world that SGD is a prevalent phenom
enon in coastal areas, it is important to quantify SGD-related
metal fluxes into coastal waters in order to understand local,
regional, and global metal budgets. Understanding how SGD
may carry metals of anthropogenic origin into coastal waters is
also important for coastal water-quality management.
4.08.4.1
Metal Concentrations in Coastal Groundwater
Metal concentrations in uncontaminated aquifers can be very
low. For example, Sañudo-Wilhelmy et al. (2002) measured
the concentrations of Al, Ag, Cd, Cu, Mn, Pb, and Zn in a
residential area of Long Island, New York, believed to have
uncontaminated groundwater and found that they were as
low as concentrations in the open ocean. However, in other
cases, the composition of the aquifer substrate can lead to
naturally high groundwater metal concentrations. In West
Bengal and Bangladesh, groundwater in the alluvial Ganges
aquifer has natural levels of arsenic high enough to endanger
the health of humans who drink it. This problem results from
arsenic-rich iron oxyhydroxide minerals that undergo reductive
dissolution within the aquifer (Nickson et al., 1998). Similar,
although less well-known, cases of natural groundwater arsenic
contamination have been described in Taiwan’s Chianan
coastal plain (Chen and Liu, 2007), and in Vietnam (Berg
et al., 2001). Groundwater iron concentrations tend to be
high in the same areas (Berg et al., 2001). Iron and manganese
concentrations were found to be thousands of times higher in
Brazilian groundwater than in offshore seawater, a condition
that resulted from the aquifer’s geologic composition and sub
oxic status (Windom et al., 2006). Concentrations of Ba are
also often high in the subterranean estuary because of this
element’s salinity-dependent desorption (Shaw et al., 1998).
Finally, concentrations of total and/or monomethyl mercury
were elevated in coastal groundwater in central California due
to high mercury levels in the aquifer rocks and hydrothermal
activity (Black et al., 2009). These examples illustrate that local
geology is an important control on SGD-related metal inputs to
the coastal ocean.
Although high groundwater metal concentrations can occur
naturally, they may also result from human activity. In the MidLevels area of Hong Kong, one of the most developed coastal
areas in the world, mean groundwater concentrations of Mn, V,
Co, and Mo were 30, 10, 17, and 13 times, respectively, greater
than in undeveloped areas immediately upslope. However, Zn,
Cr, Cu, Cd, Pb, and Fe concentrations were not significantly
different in developed areas. This was attributed to attenuation
of these elements in the vadose zone, as other evidence indi
cated the presence of sewage, which contains all of them at high
concentrations (Leung and Jiao, 2006). Elevated concentra
tions of Cu, Cr, Ni, Pb, Cd, and Zn in Sicilian springs were
attributed to industrial activity (Povinec et al., 2006).
Surprisingly, metal concentrations observed in landfill leachate
have generally been relatively low (Christensen et al., 2001;
Øygard et al., 2004), although still higher than pristine ground
water or open ocean seawater. However, these concentrations
are variable and can be high in some places (Jensen and
Christensen, 1999), especially when mining waste is involved
(Martin and Pedersen, 2002). In such cases, leachate could
contaminate groundwater with dissolved metals.
High concentrations of Cu, Zn, Cd, Pb, and Ni have been
observed in stormwater (Mikkelsen et al., 1997), with stormwater runoff from developed areas having much higher trace
metal concentrations than that from natural areas (Yoon and
Stein, 2008). This stormwater could infiltrate into the aquifer
and raise groundwater metal concentrations; however, natural
soils and stormwater infiltration systems often remove the
metals from solution before they reach the water table
(Mikkelsen et al., 1997). Karst systems are more sensitive to
the infiltration of high-metal stormwater than most other types
of aquifers, and high metal concentrations have been observed
at karst springs following storm events (Vesper and White,
2003).
Land reclamation, the process of dumping fill materials on
top of marine sediments in order to create more coastal land for
human use, has also been associated with elevated ground
water metal concentrations. In the coastal Shenzhen area of
China, concentrations of V, Cr, Mn, Co, Ni, and Cd were higher
in groundwater under reclaimed land than in groundwater
under natural coastal land located nearby (Chen and Jiao,
2007). Coastal land reclamation is a common practice in
Europe (Healy and Hickey, 2002), the United States
(Kennish, 2001), and Asia (Chen and Jiao, 2007; Guo and
Jiao, 2008). To cite an extreme example, 45% of the population
of South Korea lives on reclaimed land. As populations grow
and sea levels rise in the future, the amount of reclaimed land,
as well as its effect on coastal metal budgets, is likely to
increase.
4.08.4.2 Metal Species and Transformations in the
Subterranean Estuary
In the subterranean estuary, metals can occur in various forms.
They may be dissolved in groundwater, suspended as colloids,
or bound to particle surfaces. In addition, they can form myriad
chemical compounds depending on the groundwater chemis
try and redox state. The form in which a metal occurs can affect
its mobility, reactivity, and toxicity. Moreover, as metals are
carried seaward from inland aquifer areas through the subter
ranean estuary, they often undergo chemical transformations,
and the nature of these transformations affects the eventual
impact of the groundwater on the coastal waters into which it
discharges.
A number of studies have indicated that metals do not
always move conservatively through the subterranean estuary;
rather, they may be either released into or removed from solu
tion along the seaward flow path (Figure 10). In a study of
groundwater metal concentrations along a salinity gradient in
West Neck Bay, Long Island (New York, USA), dissolved Co
and Ni displayed nonconservative inputs mid-gradient, while
Cu, Pb, Ag, Al, and Mn had variable concentrations that were
unrelated to salinity (Beck et al., 2007). Ba and Sr showed
nonconservative inputs along a groundwater salinity gradient
in Waquoit Bay (Massachusetts, USA), with mid-transect pro
duction possibly related to the redox chemistry of Mn at
brackish salinities. In contrast, U was nonconservatively
removed over the same gradient (Charette and Sholkovitz,
2006).
One of the main controls on metal behavior in the subsur
face is the phase in which the metal occurs: solid, dissolved, or
colloid. In the solid phase, metals are immobile and attached
to or incorporated into the aquifer substrate. In the dissolved
phase, metals are the most mobile, and the colloid phase is
intermediate.
A substantial proportion of the total groundwater metal
pool occurs as colloids in which the metal is evenly distributed
Metal concentration
Fresh groundwater endmember
Seawater endmember
Conservative species
Nonconservative addition
Nonconservative removal
Conservative mixing line
Salinity
Figure 10 Idealized schematic drawing of conservative and nonconser
vative behavior of a dissolved metal in the subterranean estuary.
219
throughout the groundwater as a suspension, rather than a
solution (see Chapter 4.06 for further details). This proportion
varies depending on the metal under consideration. For exam
ple, in unpolluted Long Island (New York, USA) groundwater,
approximately 85% of Al occurred in the colloidal phase, while
only 37% of Mn did (Sañudo-Wilhelmy et al., 2002). The
partitioning of metals between dissolved and colloidal phases
may be related to their affinity for humic substances in the soil.
Humic substances are organic compounds that form from
decaying plant and animal matter, or humus. Those metals
with a greater affinity for humic substances are more likely to
form colloids (Sañudo-Wilhelmy et al., 2002).
When metals are present as colloids rather than in the
dissolved phase, they are less mobile and less likely to emerge
at the coast via SGD. In a study conducted in Germany
(Baumann et al., 2006), landfill-derived metals in groundwater
were mainly present in the colloidal, rather than in the dis
solved phase, and this behavior inhibited their transport away
from the landfill. Thus, although concentrations of Fe, Mn, Cd,
Co, Cu, Ni, Pb, and Zn were higher down-gradient of the
landfill compared to up-gradient, this increase was much less
than expected, and down-gradient concentrations were much
lower than groundwater metal concentrations within the land
fill. The formation of colloids prevented much of the metal
load from being carried away from the landfill via groundwater
flow.
Besides colloid formation, four main chemical processes
control metal concentrations in groundwater: redox-controlled
solubility, adsorption onto Fe and Mn oxides or organic matter,
release from oxides undergoing reductive dissolution, and des
orption from sediments via ion-exchange reactions (Charette
and Sholkovitz, 2006). Both the redox potential (EH) and pH
of groundwater can affect the solubility of metals, although the
actual effect differs from one metal to the next. For example, in
a study of rice paddies (Reddy and Patrick, 1977), Cd became
more soluble with increasing EH, while the opposite was true of
Pb. In sediments and soils, trace metals frequently adsorb to Fe
and Mn oxides. Both the abundance of these oxides and the
propensity of trace metals to adsorb on them are affected by
redox conditions. For example, Co, Pb, and Zn were found to
be more soluble at low EH because this condition was less
favorable for adsorption onto Fe and Mn oxides (Reddy and
Patrick, 1977; Iu et al., 1982). Finally, desorption from sedi
ments via ion-exchange reactions can be particularly important
in the subterranean estuary because of the steep gradients in
ionic strength that characterize the aquifer zone where fresh
and saline waters mix. These reactions produce nonconserva
tive inputs of Ra, Ba, and Sr at intermediate groundwater
salinities (Charette and Sholkovitz, 2006). The oxidation of
reduced metal species, usually sulfides, in anaerobic or
tailing-contaminated sediments can also result in trace metal
release.
Plants and microbes can affect groundwater metal concen
trations. Most terrestrial plants have the ability to take up
metals used as micronutrients, including Fe, Mn, Zn, Cu, Mg.
Mo, and Ni. Some can also take up and accumulate in their
body tissue significant quantities of metals that have no known
biological function, including Ba, Cd, Cr, Pb, Co, Ag, Se, and
Hg (Salt et al., 1995). Because plants are only capable of taking
up dissolved nutrients, the effect of plants would be modulated
by the solubility of the metal in question under particular
220
conditions. Many types of microbes also reduce metals as part
of their metabolic pathways or for detoxification (Lovely,
1993). Fe, Mn, U, Se, Cr, Hg, Tc, V, Mo, Cu, Au, and Ag can
all be reduced through microbial action. Although reducing
these metals does not physically remove them from the aquifer,
it can change their solubility or reactivity. For example, the
microbial reduction of highly soluble U(VI) to highly insoluble
U(IV) has the result of removing U from water and adding it to
sediment (Lovely, 1993).
Finally, anthropogenic activity can affect the partitioning
and oxidation states of metals in the subsurface. For example,
Audry et al. (2010) observed that trace metals, including Zn,
Cd, Pb, and Cu, displayed distinct chemical partitioning upand downstream of a mining and smelting site. Neumann et al.
(2009) found that infiltration ponds in Bangladesh contribu
ted to dangerously high dissolved arsenic concentrations in
groundwater by providing the dissolved organic carbon
(DOC) that fuels the microbially mediated geochemical trans
formations that release arsenic from sediments. In addition,
anthropogenic contamination of the subsurface with chemical
pollution can permanently alter aquifer properties, including
the ability to attenuate metals (Yaron et al., 2010).
4.08.4.3
Fluxes of Metals into the Coastal Ocean from SGD
Only recently has SGD been recognized as a potentially impor
tant pathway for metals to enter the coastal ocean. Depending
on the magnitude of other inputs to the coastal metal budget,
SGD may be an important or even dominant component. For
example, during periods of high discharge, SGD accounted for
up to 58% of the total flux of Cu into Flanders Bay (Long
Island, New York, USA; Montluçon and Sañudo-Wilhelmy,
2001). However, in the Elizabeth River estuary (Virginia,
USA), SGD accounted for only 3% of the total Cu flux because
of large Cu inputs from paints associated with the nearby naval
shipyard (Charette and Buesseler, 2004). SGD delivers a similar
quantity of Sr to the Bengal Basin as the Ganges and
Brahmaputra Rivers (Basu et al., 2001), and, in the southeast
ern United States, SGD accounts for about 75% of Ba inputs to
the coastal ocean. SGD has also been shown to be a major
source of Fe, a key micronutrient for marine productivity, to the
South Atlantic Ocean (Windom et al., 2006).
SGD may contribute to the impairment of coastal water
bodies by toxic metals such as Hg. Estimated SGD-related Hg
fluxes at Waquoit Bay (Massachusetts, USA) were an order of
magnitude higher than the atmospheric deposition rate of Hg
in the same region (Bone et al., 2007). Similar results were
obtained at two sites on the central California coast, where total
Hg fluxes from SGD were considerably higher than atmo
spheric deposition. In the same study, fluxes of monomethyl
Hg, the most toxic form of this element, from SGD were com
parable to typical diffusive fluxes from estuarine bottom
sediments (Black et al., 2009). In contrast, the subterranean
estuary appears to act as a natural barrier for As, trapping it in
sediments and preventing its flux into coastal surface waters
(Bone et al., 2006).
Estimating metal fluxes into coastal waters can be a chal
lenge because groundwater metal concentrations can be highly
variable, and metals can display nonconservative behavior over
short spatial scales. For example, calculated fluxes of Co into
West Neck Bay (Long Island, New York, USA) ranged from
3.4 102 to 8.2 103 μmol day−1, depending on sampling
location and season. Even the choice of different wells less
than 30 m apart on the same beach yielded order-of-magnitude
variations in flux estimates for Mn, Cu, Co, and Ag (Beck et al.,
2007). Despite the difficulty involved, estimating SGD-related
metal fluxes is important because, as illustrated above, it may
be impossible to make sense of coastal metal budgets without
them.
4.08.4.4
Effects of SGD-Derived Metals in Coastal Waters
Metals contributed to coastal waters by SGD can provide
essential micronutrients, have toxic effects, and affect the inter
pretation of paleoceanographic records. Fe, Mn, and Zn can
limit phytoplankton growth; interestingly, higher concentra
tions of these metals are necessary to support the growth of
phytoplankton living in nearshore waters compared to openocean species (Brand et al., 1983). In cyanobacteria, Fe is
essential for photosynthesis, and both Fe and Mo are necessary
for nitrogen fixation and nitrate reduction (Reuter and
Petersen, 1987). Recent research has suggested that micronu
trients, including B, Co, Cu, Fe, and Mo, may increase algal or
cyanobacterial growth when concentrations of macronutrients
(i.e., nitrogen and phosphorus) are already high and the system
is already characterized as eutrophic (Zhang, 2000; Downs
et al., 2008). As eutrophic conditions in coastal waters are
expected to become more prevalent in the future (Rabalais
et al., 2009), the role of metal micronutrients, including those
associated with SGD, may become more important.
In addition to supporting the growth of both desirable and
undesirable primary producers, some metals can also be toxic
to plants, animals, and humans. For example, Cu can interfere
with nitrogen fixation in cyanobacteria (Reuter and Petersen,
1987) and it has toxic effects on picoeukaryotes and
Synechococcus, one of the most common marine phytoplankton
species (Paytan et al., 2009). Cd (Miao and Wang, 2006), as
well as high concentrations of Zn, Co, and Mn (Rosko and
Rachlin, 1975), can depress phytoplankton growth. Some
SGD-borne metals, notably Hg, bioaccumulate at higher
trophic levels, harming top predators and humans even when
concentrations in the water itself are relatively low. For exam
ple, mussels growing in high-SGD areas of the Seine Bay
(France) had higher Hg levels than those growing in low-SGD
areas, posing a potential health threat to human consumers
(Laurier et al., 2007).
Finally, past SGD-related inputs of metals such as Sr and Ba
can add a twist to the paleoceanographic record. For example,
the 87Sr/86Sr isotopic ratio in marine sediments and sedimen
tary rocks has been used to reconstruct events, such as tectonic
collisions and uplift of mountains, which led to changes in Sr
sources and fluxes over the past 40 million years. Groundwater,
which today has Sr concentrations 10 times higher than river
water, could have a significant effect on the Sr mass balance and
must be taken into consideration (Basu et al., 2001). Ba is used
as a tracer of paleoproductivity and paleocirculation; therefore,
an accurate understanding of the marine Ba cycle is necessary in
order to correctly interpret the marine barite (BaSO4) record. As
SGD-related Ba fluxes to the coastal ocean are often high in the
present day (Shaw et al., 1998), it is likely that they were also
significant in the past, making them an important component
of the global Ba cycle preserved in the geologic record.
4.08.4.5
Metal concentrations in groundwater are determined by the
aquifer substrate, anthropogenic land use, pH, oxidation–
reduction potential, salinity, and the abundance of humic sub
stances in the soil or sediment. In some pristine locations,
groundwater metal concentrations are very low, but in other
locations they may be comparable to or higher than surface
water concentrations. This may be due to the aquifer geology or
to human inputs from land uses such as waste disposal. Metals
often behave nonconservatively as they move seaward through
the aquifer, which can make it difficult to characterize the
concentration in the discharging groundwater endmember.
Nonetheless, in those locations where they have been esti
mated, fluxes of metals including Ba, Fe, Sr, Hg, and Cu were
often large relative to fluxes from other sources, such as rivers
and atmospheric deposition.
Estimating SGD-related fluxes of more metals, at more loca
tions, is one of the active areas in SGD research today. One key
component of this research is understanding how different
metals behave in the dynamic environment of the subterranean
estuary, which is characterized by steep gradients in salinity,
pH, and redox state. This is particularly important for metals
that tend not to behave conservatively in the subsurface.
Research completed so far indicates that SGD is often an
important term in coastal metal budgets; thus, these efforts
should lead to a better understanding of how metals enter
and move through coastal ecosystems. This information, in
turn, will allow for more effective pollution prevention and
management.
4.08.5 SGD-Related Fluxes of Other Constituents
Fluxes of nutrients and, to a lesser extent, metals have been the
focus of most SGD studies conducted thus far. However, SGD
may also supply other constituents to coastal waters. These
include dissolved organic carbon (DOC), fecal indicator bac
teria (FIB), caffeine, pharmaceuticals, methane, and volatile
organic compounds (VOCs). Microorganisms in groundwater
and coastal waters consume DOC (Baker et al., 2000), so its
presence can affect microbial abundance. FIB, caffeine, and
some pharmaceuticals are of concern mainly because they can
indicate the presence of sewage or wastewater, while VOCs
and some other pharmaceuticals may have negative effects of
their own once released into environmental waters.
4.08.5.1
Dissolved Organic Carbon
One key difference between surface and subterranean estuaries
is that the former are exposed to sunlight, while the latter are
almost completely shielded from it. Because of this, photosyn
thetic organisms such as plants, algae, and phytoplankton are
less important in a subterranean estuary’s biogeochemistry,
and microbes, many of which do not require light, are more
important. DOC, which originates from the decomposition of
organic matter, provides food for groundwater microbes and
thus fuels much of the subterranean estuary’s biological activ
ity. While it can be added to groundwater by human activities
such as waste disposal, groundwater DOC usually originates
from natural sources.
221
In addition to sustaining groundwater microbes, DOC also
affects the behavior of metals in the subsurface. In general,
DOC either mobilizes metals and increases their concentra
tions in groundwater, or has no clear effect. For example, Cd,
Ni, and Zn in landfill leachate can form complexes with DOC,
leading to modest increases in these metals’ mobility
(Christensen et al., 1996). Another study, in an industrial area
of Germany, found that DOC concentration was correlated
with concentrations of Hg, Cr, Cu, and As, indicating that it
helped mobilize these metals; however, in the same study,
DOC had no effect on Cd or Zn (Kalbitz and Wennrich, 1998).
Once it enters coastal surface waters, DOC can continue to
influence microbial activity and metal mobility. In many cases,
SGD is a significant source of DOC to coastal waters. For
example, in Celestún Lagoon on Mexico’s Yucatán Peninsula,
SGD accounted for 16% of the total annual carbon export from
the lagoon, leading the authors to conclude that changes in
groundwater DOC concentration could affect biogeochemical
cycling not only in the lagoon itself, but also in nearby coastal
waters receiving the lagoon’s outflow (Young et al., 2005). In
Thailand, DOC fluxes from SGD were 20–30% of those from
the Chao Phraya River, the largest source of freshwater to the
Gulf of Thailand (Burnett et al., 2007). In North Inlet, South
Carolina (USA), SGD constituted a significant fraction of the
total DOC export, with the majority of the flux coming from
saline SGD (Goñi and Gardner, 2003).
4.08.5.2
FIB and Other Microbes
FIB are bacteria that are used to indicate the presence of human
waste in environmental waters. They include Enterococcus, fecal
coliform, and Escherichia coli, which is one particular species of
fecal coliform bacteria. Although they do not generally pose a
serious threat to human health, some epidemiological studies
(e.g., Cabelli et al., 1982; Seyfried et al., 1985; Corbett et al.,
1993) have shown a link between high FIB concentrations in
bathing waters and increased risk of water-borne illness in
bathers. This link presumably exists because both FIB and
human pathogens could originate from the same source –
human sewage. Based on these studies, federal and state reg
ulatory agencies use FIB to assess whether water bodies are safe
for swimming. However, it has been noted that this relation
between FIB and human sewage-borne pathogens is not always
seen (Lipp et al., 2001; Thompson et al., 2003; Colford et al.,
2007).
SGD may become a pathway for FIB to enter coastal waters
if it is contaminated by sewage through activities such as waste
treatment and disposal. In Key Largo, Florida (USA), FIB were
observed in the onshore and nearshore aquifer as well as
coastal ocean waters. This contamination was attributed to
the use of septic tanks and other onsite waste disposal systems,
which can leak into the highly porous limestone or karst aqui
fer (Paul et al., 1995). Karst aquifers may be particularly
susceptible to fecal pollution because of the strong connectivity
between surface and groundwater. For example, a study of a
French karst aquifer (Mahler et al., 2000) showed high FIB
concentrations, often associated with suspended sediment
within the aquifer. These concentrations were highest after
heavy rains, when FIB and sediments were flushed in.
However, contamination of groundwater by FIB is not lim
ited to highly porous karst systems. In Huntington Beach,
222
California, FIB concentrations in the surf zone were correlated
with radium, a natural groundwater tracer, suggesting that
these bacteria were entering the surf zone via SGD, and/or
that one of the components in SGD was increasing the FIB
survival rate (Boehm et al., 2004; Paytan et al., 2004).
Another line of evidence connecting SGD to FIB concentrations
in the coastal ocean is the finding that, just like SGD,
Enterococcus concentrations at 60 Southern California beaches
were correlated with the spring–neap tidal cycle (Boehm and
Weisberg, 2005). The highest Enterococcus concentrations at
most beaches occurred at spring tides, when the tidal range
was greatest. A different result was observed at Stinson Beach,
California, where the highest FIB inputs from SGD occurred at
neap tide. This was because, although the total SGD flux was
greatest at spring tide, the fresh SGD flux was greatest at neap
tide, and fresh SGD was much more enriched in FIB than saline
SGD (de Sieyes et al., 2008). Although SGD is an important
transport mechanism for FIB, in some cases its role is not
evident. For example, on the north shore of Kauaì (Hawaiì,
USA), river and stream fluxes, rather than SGD, dominated FIB
inputs to coastal waters (Knee et al., 2008).
4.08.5.3 Caffeine, Pharmaceuticals, and Personal Care
Products
Caffeine, pharmaceuticals, and chemicals found in personal
care products have recently received attention as possible tra
cers of wastewater contamination in groundwater, freshwater,
and marine systems. These chemicals are usually present at
high concentrations in wastewater and absent in pristine or
natural waters. In addition to indicating wastewater inputs,
some of these chemicals may have negative environmental
impacts of their own, either alone or in combination.
Caffeine has been investigated as a possible wastewater
tracer because of its nearly ubiquitous consumption by
humans. It enters the wastewater stream mainly through dis
posal of caffeine-containing beverages and products, with
human waste as a secondary source due to incomplete meta
bolism. In Swiss rivers and lakes, caffeine concentration was
correlated with human population density, indicating that it is
a good quantitative marker of wastewater discharge in that
system (Buerge et al., 2006). Caffeine was also detected in
Boston Harbor (Massachusetts, USA; Siegener and Chen,
2002) and Singapore coastal water (Wu et al., 2008), where
its distribution indicated a land-derived, anthropogenic source.
However, caffeine may not be an ideal wastewater tracer
because it can be degraded by light and microbial action
(Bradley et al., 2007) and, in some places, it has natural in
addition to anthropogenic sources (Peeler et al., 2006).
To the best of our knowledge, no studies calculating caffeine
fluxes to coastal waters from SGD have been published.
However, given the high caffeine concentrations in contami
nated groundwater and the recognition that discharge of
sewage-contaminated groundwater does occur, it is likely that
SGD provides caffeine to some coastal waters. The levels mea
sured in environmental waters are typically on the order of
parts per billion (Siegener and Chen, 2002; Buerge et al.,
2006; Peeler et al., 2006), much lower than those observed to
cause harmful effects on humans or wildlife, including highly
sensitive amphibians (Richards and Cole, 2006). Nonetheless,
in some cases, caffeine may be useful as an indicator of waste
water contamination in SGD.
Other pharmaceuticals and personal care products have also
been suggested as wastewater markers in coastal environments.
A Japanese study (Nakada et al., 2008) found that concentra
tions of crotamiton (a topical drug for the treatment of scabies),
carbamazepine (an anticonvulsant and mood-stabilizing drug),
and mefenamic acid (an anti-inflammatory drug) were corre
lated with population density, and that these compounds
behaved conservatively in a surface estuary with a salinity
range of 0–30. Crotamiton, carbamazepine, and several other
pharmaceutical compounds were also detected in groundwater
in the same study, suggesting that if SGD were occurring in this
area, it would be a source of pharmaceuticals to coastal waters.
In Cape Cod, groundwater-fed coastal ponds in areas with
higher residential density had higher concentrations of steroidal
hormones, pharmaceuticals, and other organic wastewater com
pounds than those in areas of lower residential density (Standley
et al., 2008). High concentrations of estrogens, both natural and
artificial, occur in sewage, and they may leach into groundwater
from septic systems and be transported to the coastal ocean via
SGD (Atkinson et al., 2003). Recently, acesulfame, an artificial
sweetener, has been suggested as an ideal wastewater tracer in
groundwater because it is widely consumed and present at high
concentrations in the wastewater stream; additionally, it does
not appear to be subject to physical, chemical, or biological
degradation over relevant environmental timescales (Buerge
et al., 2009).
The potential addition of pharmaceutical compounds, par
ticularly estrogenic hormones, to coastal waters by SGD is
worrisome because these compounds may harm aquatic organ
isms. In a Cape Cod study (Standley et al., 2008), estrogenic
compounds were present in coastal ponds at concentrations
high enough to induce physiologic changes in fish. Some phar
maceuticals can affect wildlife at trace concentrations of only a
few parts per trillion (Snyder et al., 2003), and, because of
widespread human use, trace concentrations of many pharma
ceuticals are ubiquitous in the wastewater stream. Because the
levels are so low, they are difficult to detect and remove with
current wastewater treatment technologies (Snyder et al.,
2003). Both the effect of trace concentrations of single pharma
ceuticals on humans and the effects of combinations of
pharmaceuticals on humans and wildlife are poorly under
stood at present. Although SGD is likely to contribute
pharmaceuticals to coastal waters, at least in some locations,
further research is necessary to understand the nature and
extent of this effect.
4.08.5.4
Pesticides
Pesticides are chemicals used to control or eliminate weeds,
plant diseases, or unwanted insects in agricultural, residential,
or recreational areas. When applied to fields or lawns, they can
leach into groundwater and be transported to estuaries and
coastal waters via SGD. A 10-year-long, nationwide study by
the United States Geological Survey (Gilliom, 2007) detected
pesticides in over 50% of shallow wells in urban and agricul
tural areas, and many studies (e.g., do Nascimento et al., 2004;
Almasri, 2008; Anderson et al., 2010) have reported elevated
pesticide concentrations in coastal groundwater. Although
SGD-related transport of pesticides into coastal waters is clearly
possible, few studies have reported pesticide concentrations in
discharging groundwater or SGD-related pesticide fluxes.
One study (Gallagher et al., 1996) reported alachlor,
atrazine, cyanazine, and metolachlor concentrations of
0.05–0.5 μg l−1 in groundwater discharging into the southern
Chesapeake Bay.
4.08.5.5
Volatile Organic Compounds
Volatile organic compounds (VOCs) are defined as organic
chemical compounds that, due to their high vapor pressures,
vaporize to a significant extent under normal conditions. They
are used in a wide range of industrial, chemical, commercial,
and household applications (Fusillo et al., 1985), and are
frequently found in municipal and industrial wastes (Edil,
2003; Fan et al., 2009). Urban air may be an additional source
of VOCs to shallow groundwater because it can contain signif
icant concentrations of these compounds, many of which
display strong partitioning from air into water (Pankow et al.,
1998). Several VOCs are known carcinogens (Belpomme et al.,
2007; Claxton and Woodall, 2007) and may pose a threat to
the health of humans, wildlife, and aquatic or coastal
ecosystems.
Due to direct and indirect anthropogenic influences, VOCs
are common in groundwater in industrialized areas. A study of
almost 3000 wells in the United States, conducted as part of the
United States Geological Survey’s National Water-Quality
Assessment Program between 1985 and 1995, showed that
14% of all well samples contained measurable levels of at least
one VOC; this rose to 47% in urban areas (Squillace et al., 1999).
VOCs were also prevalent in contaminated groundwater in other
industrialized areas, including Taiwan (Fan et al., 2009),
Northern Ireland (Phillips et al., 2007), and Sicily (Pecoraino
et al., 2008). In Sicily, VOC concentrations were highest in
groundwater near industrial and intense agricultural areas
(Pecoraino et al., 2008). The lack of light in aquifers prevents
VOCs from degrading by photolysis (Rathbun, 2000), which
may allow them to persist longer than in surface waters.
Although VOCs are common in groundwater, several pro
cesses may remove them before the groundwater discharges into
the ocean as SGD. These include sorption, dilution, and volati
lization, which change the form of VOCs but do not destroy
them, as well as biodegradation, which breaks down the mole
cules (Lorah and Olsen, 1999). Biodegradation of various VOCs
has been documented in sand (Semprini et al., 1995), unconso
lidated sediment (McNab and Narasimhan, 1994), and
fractured bedrock aquifers (Lenczewski et al., 2003), as well as
in tidal marshes (Lorah and Olsen, 1999). Wetlands may be
particularly effective at breaking down VOCs in groundwater
because of their naturally high levels of DOC, low redox state,
and anoxic conditions (Lorah and Olsen, 1999).
To the best of our knowledge, no study specifically quanti
fying SGD-related inputs of VOCs to coastal waters has been
conducted. However, discharge of VOC-containing ground
water into rivers has been documented (Conant et al., 2004;
Ellis and Rivett, 2007), and VOCs have been detected in coastal
aquifers (Fusillo et al., 1985; Graber et al., 2008). Thus, it is
likely that discharge of VOC-contaminated groundwater into
coastal waters occurs as well, particularly in places where VOC
sources are located close to the shore and conditions are not
favorable for rapid microbial degradation. Because of the
223
health threats posed by VOCs, inputs of these compounds to
coastal waters from SGD should be investigated further.
4.08.5.6
Methane
Methane (CH4) is a potent greenhouse gas that can be pro
duced by some microbes through the anaerobic process of
methanogenesis. Observed methane concentrations in ground
water vary widely, from 40 to 6300 μmol l−1 (Barker and Fritz,
1981; Grossman et al., 1989; Watanabe et al., 1994), based on
aquifer conditions. Aquifers with high organic matter content
and low dissolved oxygen tend to have high methane concen
trations, whereas those with less organic matter and more
oxygen are lower in methane. Because its concentrations in
groundwater are frequently high, some workers have used it
as an SGD tracer (e.g., Cable et al., 1996; Kim and Hwang,
2002; Santos et al., 2009).
In some areas, SGD is associated with large fluxes of
methane into the coastal ocean, while at others such fluxes
are very small. In the northeastern Gulf of Mexico, the methane
flux from SGD into nearshore waters was 21 nmol m−2 min−1,
accounting for 83–99% of all methane inputs into the water
column (Bugna et al., 1996). More recent studies at the same
location (Santos et al., 2009) and in the South Sea of Korea
(Kim and Hwang, 2002) showed that variations in methane
concentration tracked SGD variations over timescales of hours
and days. However, in West Neck Bay, New York (Dulaiova
et al., 2006b), Toyama Bay, Japan (Kameyama et al., 2005),
and the Yucatan Peninsula of Mexico (Young, 2005), ground
water methane concentrations were lower than those of coastal
waters. In these areas, SGD did not contribute significantly to
the coastal ocean methane budget.
4.08.5.7
SGD represents an important hydrologic connection between
land and sea. Thus, any constituent that can enter groundwater
through natural processes or human activity may eventually
find its way into coastal waters via SGD. These constituents
include not only nutrients and trace metals, but also microbes,
DOC, caffeine and other pharmaceuticals, and VOCs. Some are
relatively harmless, whereas others indicate the presence of
human sewage or pose a threat to human and ecosystem health
in their own right.
Some of the pollutants discussed in this section have only
recently been recognized and studied. Future research in this
field will likely focus on quantifying different sources of each
pollutant to water bodies and understanding how each source
responds to various driving factors, such as seasonal cycles,
waste disposal practices, land use, and land cover. In order to
gain a complete understanding of pollution inputs to water
bodies, the role of SGD in supplying these pollutants must also
be understood.
4.08.6 SGD and Climate Change
Climate change is one of the most pressing issues in environ
mental science today. According to the Intergovernmental
Panel on Climate Change (IPCC), over the next century the
Earth’s average temperature is expected to rise by 1–4 °C,
224
global precipitation is expected to increase by up to 6%, and
average global sea level is expected to rise by 0.1–0.8 m. Clearly,
local, regional, and global hydrologic cycles will be affected by
climate change, and both the quantity and quality of SGD will
likely change as a result.
4.08.6.1
Groundwater Temperature
Although temperature is not always considered as an aspect of
groundwater quality, it is important for several reasons. First, in
some locations, temperature can be used as a groundwater
tracer. SGD at high latitudes tends to be warmer than the
ocean into which it discharges (Roxburgh, 1985; Banks et al.,
1996; Miller and Ullman, 2004), whereas at low latitudes it
tends to be cooler (Shaban et al., 2005; Duarte et al., 2006;
Johnson et al., 2008). Such temperature anomalies, which may
also vary seasonally (Miller and Ullman, 2004), can be used to
identify and map SGD plumes and quantify the magnitude of
discharge (Johnson et al., 2008). Second, some chemical and
particularly microbial processes that occur in groundwater are
temperature dependent, so temperature can affect biogeochem
ical cycling and the concentrations of better-recognized
groundwater-quality constituents, such as nutrients, metals,
and microbes (Taniguchi et al., 2007). Finally, if SGD has a
different temperature than the receiving water and the dis
charge rate is high, the change in temperature may have
ecological effects in coastal systems. Fish and other aquatic
organisms can be highly sensitive to temperature, as their
body temperature changes with that of the water in which
they live (Gooseff et al., 2005). In addition, changes in tem
perature affect the water’s dissolved oxygen capacity.
Both climate change and urbanization can affect the tem
perature of groundwater and, ultimately, SGD. In a study of
four Asian megacities, Taniguchi et al. (2007) found that surfi
cial groundwater was 2–3 °C warmer than would be expected
based on the natural geothermal gradient. These temperature
anomalies corresponded to changes in air temperature in the
same cities over the past century, and the authors attributed
them to warming caused by a combination of global climate
change and the urban heat island effect (Taniguchi et al., 2007).
The effect of global warming on groundwater temperatures has
not been studied extensively, but it is reasonable to hypothe
size that as the temperature of the air increases, that of the
groundwater will as well, at least in shallow, unconfined
aquifers.
Increased groundwater temperatures resulting from global
warming may have several effects. They may obscure actual
temporal trends in SGD if temperature is used as a groundwater
tracer. If, for example, the temperature of a cold SGD plume
increases, the plume may appear to have decreased in size over
time when in fact it is the same size, but warmer. Because water
temperature is related to its buoyancy, changes in SGD tem
perature may affect the way SGD plumes either stratify from, or
mix with, coastal waters. Higher water temperatures can pro
mote the survival and growth of FIB (He et al., 2007), and
increase their propensity to attach to sand (Castro and
Tufenkji, 2007). The metabolic rates of bacteria involved in
mercury methylation or metal reduction may also increase,
elevating the fluxes of toxic compounds and metals to the
coastal ocean. In addition, groundwater discharge in streams
creates pockets of cooler water called ‘thermal refugia’ that
harbor fish adapted to colder temperatures. The overall effect
of the heterogeneity in temperature is to increase species diver
sity (Chu et al., 2008). The same may occur in coastal
environments (Dale and Miller, 2007), especially in the
absence of strong mixing from waves and other processes.
With increased groundwater temperatures, the occurrence of
such refugia may diminish.
4.08.6.2
SGD Quantity
Several factors related to climate change may affect future volu
metric fluxes of both fresh and saline SGD. Changes in
precipitation and evapotranspiration will alter local and regio
nal water budgets and fresh groundwater recharge, which is the
precursor of fresh SGD. Changes in the timing and frequency of
precipitation – whether rain falls in a few large storms or many
smaller ones, and whether it falls during hot or cold times of
year – may also affect the partitioning of rainfall between run
off, evaporation, and groundwater recharge. In areas where
some precipitation occurs in the form of snow, the balance
between rain and snow may affect groundwater recharge.
Climate change may also affect some of the driving factors for
saline SGD, including waves, storm surges, and sea-surface
height.
Global average precipitation is projected to increase slightly
over the next century as a result of climate change (Cubasch
et al., 2001), but these changes depend greatly on the region.
High-latitude areas in the Northern Hemisphere may see their
annual precipitation increase by over 20%, while Central
America, Australia, the Mediterranean region, and southern
Africa will likely experience decreases in precipitation (Giorgi
et al., 2001). Precipitation drives fresh groundwater recharge,
which occurs through three main mechanisms:
1. direct recharge, or the infiltration of water from the surface,
through the vadose zone, to the water table;
2. localized recharge, or the infiltration of water through pre
ferential flow paths such as joints, depressions, or rivulets;
and
3. indirect recharge, or the infiltration of water through the
beds of losing streams (Figure 11).
Thus, although increased precipitation would generally be
associated with increased groundwater recharge and discharge,
this effect could be modulated by changes in temperature,
which affects evaporation, or streamflow (de Vries and
Simmers, 2002).
Certain areas are expected to experience particularly severe
decreases in groundwater recharge. Northeastern Brazil, south
western Africa, and areas along the southern rim of the
Mediterranean Sea may have 70% less groundwater recharge a
century from now than they do today. In other places, such as
the Sahel, the Near East, northern China, the western United
States, and Siberia, groundwater recharge is projected to
increase by up to 30% over the same time period (Döll and
Flörke, 2005). In some high-latitude areas, such as the Yukon
River Basin (Walvoord and Striegl, 2007), ground that was
previously frozen solid is predicted to thaw for longer parts of
the year, allowing more and deeper groundwater flow to occur.
In this area, groundwater is higher in dissolved inorganic car
bon and nitrogen, whereas surface runoff is higher in dissolved
E = evapotranspiration
E
E
Rivers
Joints,
depressions,
rivulets
Infiltration
Infiltration
E
Indirect
recharge
Precipitation
reaching
surface
Runoff
and
interflow
E
Infiltration
E
E
Localized
recharge
225
Direct
recharge
Figure 11 Mechanisms of fresh groundwater recharge, which, at steady state, equals fresh groundwater discharge at the shoreline. Reproduced from
de Vries, J.J., Simmers, I., 2002. Groundwater recharge: an overview of processes and challenges. Hydrogeology Journal 10, 5–17.
inorganic carbon and nitrogen. Thus, if groundwater makes up
more of the total freshwater flux to the Bering Sea in the future,
the predominant forms of carbon and nitrogen entering the sea
may also shift, perhaps with ecological consequences.
A few studies have attempted to model the effects of climate
change on groundwater discharge in certain areas. In a study
focusing on Taiwan (Lin et al., 2007), groundwater discharge
decreased as a result of climate change between 2000 and 2020
regardless of what land-use changes were included in the model.
Another study (Tague et al., 2008) considered how climate
change would affect groundwater discharge into streams in the
Oregon Cascades, and found that streams fed by deep ground
water flow were more resilient than those fed by shallow
subsurface flow. These studies underscore the importance of
accounting for local and regional changes in temperature and
precipitation when projecting future SGD fluxes.
Aquifer characteristics such as hydraulic conductivity and
homo- or heterogeneity modulate the effect of climate change
on SGD quantity. For example, a modeling study focusing on
two regions in Denmark with very different aquifer geology
indicated that they would experience different climatechange-related impacts (van Roosmalen et al., 2007). One
region, Jyland, had sandy soil and large interconnected
aquifers, while the other, Sjaelland, had an impermeable top
soil layer and many clay layers impeding flow. The climate
changes predicted for these regions, most importantly
increased annual rainfall and decreased summer rainfall,
affected the high-permeability Jyland aquifer more than the
low-permeability Sjaelland aquifer. Groundwater recharge, as
well as groundwater discharge into streams, increased markedly
in Jyland, while in Sjaelland there was little effect. It is likely
that similar patterns would be observed in terms of ground
water discharge into the sea.
4.08.6.3
SGD Chemistry and Quality
Climate change may affect not only the volume of groundwater
discharged, but also its chemistry and quality. Different aspects
of climate change, namely temperature and precipitation,
may have opposite effects on groundwater quality. In addition,
the concentrations of different types of contaminants may be
affected in different ways. In one modeling study considering the
effects of projected climate change scenarios on a karst ground
water system in Switzerland (Butscher and Huggenberger,
2009), increased summer heat waves were expected to increase
groundwater contaminant concentrations because more water
would evaporate before infiltrating into the ground, leading to
reduced dilution. In contrast, the projected increase in severe
rain events led to improved water quality because of increased
dilution. The concentrations of short-lived pollutants, such as
microbes, were predicted to decrease under climate change con
ditions of increased temperature and rainfall. Conversely, the
response of long-lived pollutants such as pesticides depended on
the balance between summer heat waves, which would increase
groundwater pollutant concentrations, and heavy rain events,
which would decrease them.
Changes in rainfall intensity and timing can also affect the
leaching of nutrients from soils and fertilizers into groundwater.
In a laboratory column experiment, Sugita and Nakane (2007)
found that heavy rainfall events caused more leaching than lighter
ones, even when the total amount of rainfall was held constant.
The effect of rainfall intensity was modulated by aquifer properties
such as layering and the presence of macropores, or preferred flow
paths. In the homogeneous columns, nitrate was transported to
the bottom of the column under all simulated rainfall conditions,
but the leaching was greatest at the most intense rainfall level. In
the layered columns, microbial processes removed nitrate before
it reached the bottom of the column except at the most intense
rainfall level. Macropores increased the transport of nitrate, but
only in the heavy rain condition. These results indicate that the
timing and intensity of rainfall events may be more important in
determining groundwater quality than the total amount of rain,
and that different systems may be affected differently depending
on aquifer properties.
Finally, the sea-level rise associated with global climate
change may also affect SGD. Global sea level is predicted to
rise between 0.11 and 0.77 m by the year 2100 as a result of the
thermal expansion of the world’s oceans, as well as the melting
226
of alpine glaciers and of the Greenland and Antarctic ice sheets
(Church et al., 2001). Rising sea levels will change the location
of the land–sea and freshwater–saltwater interfaces, and parts
of the coastal aquifer that were previously fresh and brackish
are likely to become more saline. The exposure of these pre
viously fresh parts of the aquifer to water of high ionic strength
is expected to cause adsorbed ions to desorb and SGD-related
nutrient and metal fluxes to increase (Moore, 1999). Low-lying
coastal areas are likely to experience greater changes because a
given sea-level change would correspond to a greater degree of
seawater incursion. Likewise, nutrients and metals would des
orb and enter groundwater in aquifers characterized by iron
and manganese oxyhydroxides. Such aquifers accumulate high
levels of adsorbed nutrients and metals when the water in them
is fresh, and these constituents can desorb and be flushed away
if salinity increases (Moore, 1999).
4.08.6.4
Effect of SGD on Climate Change
Although climate change is more likely to affect SGD than vice
versa, some evidence suggests that SGD-borne nutrients may
influence the Earth’s climate by altering primary productivity in
the ocean. Phytoplankton use carbon dioxide, the most abun
dant greenhouse gas, for photosynthesis, thus removing it from
the atmosphere. If the phytoplankton sink to the ocean bot
tom, the carbon may be sequestered, or removed from the
rapidly cycling part of the global carbon cycle, for thousands
of years. The total amount of carbon dioxide in the atmosphere
would then decrease, slowing down global warming. One pro
posed solution to the climate problem that has arisen from this
line of reasoning is to increase primary productivity in certain
parts of the ocean so that more carbon can be taken up. Yet it
must be stressed that the ecological consequences of such
manipulations have not been thoroughly studied.
The regions of the ocean targeted for this productivity
enhancement are those that have high nutrient concentrations,
but low levels of chlorophyll, an indicator of primary produc
tivity (Figure 11). Phytoplankton growth in these regions is
limited by iron (Behrenfeld and Kolber, 1999). Iron is in short
supply because its concentrations in upwelled water are low
compared to those of other nutrients, and river iron inputs are
usually in the colloidal or particulate forms, which are harder
for phytoplankton to use. Thus, the main source of iron to most
open ocean areas is thought to be atmospheric dust deposition
(Duce and Tindale, 1991; Jickells and Spokes, 2001).
Recent research indicates that SGD may also play an impor
tant role in supplying bioavailable iron to the ocean and
regulating primary productivity. Windom et al. (2006) esti
mated that SGD from a 240-km-long stretch of the Brazilian
coastline was equivalent to approximately 10% of atmospheric
Fe inputs to the entire South Atlantic Ocean. Additionally, the
iron added to coastal waters from SGD persisted long enough
to be transported a significant distance offshore, influencing
open ocean iron budgets and regional primary productivity. As
both the magnitude and the iron concentration of SGD fluxes
into the Southern Ocean may vary in response to the Earth’s
temperature and rainfall, dynamic feedback loops between
SGD and climate may exist. As high-latitude areas continue to
become warm, SGD in these areas is likely to increase, and so
may SGD-related iron inputs. This may be especially important
because most high-nutrient, low-chlorophyll areas are located
at high latitudes.
4.08.7 Summary and Conclusions
4.08.7.1
Summary
SGD is a pervasive phenomenon that affects coastal water chem
istry and quality at many locations worldwide. SGD may be
fresh, brackish, or saline. Fresh SGD results from precipitation
in recharge areas, whereas saline SGD can result from the recir
culation of seawater through the aquifer on daily, fortnightly, or
seasonal timescales, or from the discharge of deeper, saline
groundwater that has been separated from seawater for a long
time. SGD fluxes are much more difficult to quantify than river
discharges because they are often diffuse, highly variable in space
and time, and difficult or impossible to observe directly.
Over the past several decades, a number of methods have
been developed to quantify SGD. These include seepage meters
of various types, the use of natural and artificial groundwater
tracers, water balances, and hydrogeologic modeling. SGD fluxes
ranging from <1 to over 400 m3 of water per square meter of
seafloor per year (m3 m−2 yr−1) have been reported in the litera
ture, with typical values of about 40 m3 m−2 yr−1. Usually, the
saline component of SGD constitutes the majority of its volume,
with the fresh component accounting for up to 30%.
SGD is often an important term in the chemical budgets of
coastal waters. Nitrogen and phosphorus concentrations can be
orders of magnitude higher in groundwater than in surface
waters, so even groundwater discharges that are relatively
small in volume can have a large ecological impact. In some
settings, such as volcanic islands, karst regions, or arid deserts,
SGD is the major or only pathway by which terrestrial nutrients
are transported to the sea. In other locations where rivers are
abundant, SGD-related nutrient additions may be comparable
to, or greater than, those from rivers.
In addition to nutrients, SGD can add numerous other
constituents to coastal waters. These include dissolved metals,
DOC, FIB and other microbes, caffeine, pharmaceuticals, per
sonal care products, VOCs, and methane. Estimating the SGDrelated fluxes of these constituents into coastal waters can be
difficult because their behavior in the subterranean estuary is
often nonconservative and poorly understood. As our under
standing of contaminant behavior within coastal aquifers
improves, so also will our ability to quantify SGD-related fluxes
of these contaminants.
4.08.7.2
SGD at the Local, Regional, and Global Scales
Scale is an important issue when measuring SGD and consider
ing its importance to coastal ocean water quality. In terms of
measurement, seepage meters are usually the best tool for
measuring SGD at the local scale, such as a single beach or
individual locations within the beach. They provide precise
point measurements of SGD, but due to fine-scale spatial het
erogeneity, a large number must be used to obtain a seepage
rate representative of a larger area. The use of SGD tracers,
including temperature, salinity, radium, and radon, is most
appropriate at a regional scale, such as a bay, lagoon, or a
segment of coastline on the scale of kilometers to hundreds
of kilometers. Water budgets and hydrogeologic modeling are
most appropriate at the regional or larger scale. The data that
are used in these methods, such as precipitation, evaporation,
and hydraulic conductivity, are usually only available at a fairly
large scale, so small-scale variations cannot be incorporated.
Intercomparison studies such as those reported by Burnett et al.
(2006) have shown general agreement between these different
methods, although sometimes discrepancies were observed.
The importance of SGD to the budgets of various constitu
ents in the coastal ocean is also dependent on scale. Some
features of the coastline, such as bays and river mouths, can
focus SGD, while others, such as promontories, can decrease it.
Thus, SGD might be very important to the nutrient budget of a
single bay, but less important to the larger stretch of coastline
containing it. It is apparent that at some sites, SGD represents a
major source of nutrients and pollutants, but its effect on water
chemistry and quality at the regional, subcontinental, conti
nental, and global scales is less well understood. SGD is
believed to account for 0.3–16% of the total flux of freshwater
into the ocean (Burnett et al., 2003). Because this range is wide,
groundwater chemistry is highly variable, and many constitu
ents do not behave conservatively in the subterranean estuary,
it is difficult to say with certainty how important SGD is to
global oceanic budgets of nutrients and metals. However, the
research that has been done already suggests that SGD’s con
tribution to these budgets is significant, and should be
investigated further.
4.08.7.3
SGD in a Changing World
One of the challenges in SGD research is that, just as scientists
are beginning to understand how SGD works in various sys
tems, those systems are rapidly changing as a result of human
activities. In particular, climate change and coastal land devel
opment are likely to impact SGD quantity and quality in the
future. Most global climate models predict a warmer, rainier
world over the next century, although changes in temperature
and precipitation will vary greatly with location. Precipitation
increases would be expected to lead to increases in fresh SGD,
while increased temperatures could increase evaporation, and
thus decrease fresh SGD. Changes in fresh SGD quantity will
depend on the interplay between these two effects, and are
likely to vary greatly from place to place.
The effects of urban development on groundwater recharge
and discharge have not been well studied (Lerner, 2002), but
urban development would be expected to decrease fresh SGD
because the impermeable surfaces that characterize urban areas
prevent water from infiltrating into the coastal aquifer. In addi
tion, the large human populations in urbanized areas have
correspondingly large water needs, which in some areas are
met by pumping groundwater. Pumping groundwater would
remove it from the aquifer before it reaches the coastline, thus
reducing the amount of fresh SGD. Alternatively, many cities
import significant quantities of water, which may escape
through leaky pipes and recharge the aquifer. Other types of
land development, such as irrigated agriculture, can increase
SGD by providing artificial groundwater recharge if the water is
imported rather than being pumped locally.
Global climate change and land use can also affect SGD
quality. Both global warming and urbanization can increase
the temperature of groundwater (Taniguchi et al., 2007).
Increased precipitation resulting from global climate change
227
may improve groundwater quality by diluting pollutants.
Other, indirect effects of global climate change on groundwater
quality are more difficult to predict. For example, temperature
influences aquifer microbes, which, in turn, affect concentra
tions of nutrients, metals, and other constituents. Most types of
human land alteration have been tied to groundwater quality
impairments such as high nutrient and metal concentrations,
high levels of FIB, and the presence of pharmaceuticals and
personal care products. Natural biogeochemical processes can
attenuate some of this pollution. However, in general, the
potential for SGD to carry harmful substances into coastal
waters would be expected to increase with human alteration
of the coastal landscape.
4.08.7.4
Current and Future Research Directions
It is now well recognized that SGD occurs in most coastal loca
tions and that it is an important source of nutrients, metals, and
pollutants to the coastal ocean. However, recognizing that SGD
may be important is only the first step toward understanding its
role in varied coastal settings. Although SGD has been thor
oughly studied at some locations, including Waquoit Bay,
Massachussetts (e.g., Abraham et al., 2003; Michael et al.,
2005; Mulligan and Charette, 2006), the South Atlantic Bight
(e.g., Moore, 1996; Crotwell and Moore, 2003), and the Gulf
Coast of Florida (e.g., Cable et al., 1996; Moore, 2003; Santos
et al., 2009), other parts of the world have received very little
attention. One current research focus is simply to estimate SGD
at a wider range of locations, such as the southwest coast of India
(Babu et al., 2009), the Yellow River Delta of China (Peterson
et al., 2008), and Hong Kong (Tse and Jiao, 2008). Conducting
SGD studies in a wider range of locations will not only improve
understanding of those locations, but will also aid in estimating
SGD fluxes at the global level.
A second active research area is comparing and refining the
methods used to measure SGD. For example, a recent, largescale intercomparison experiment (Burnett et al., 2006) com
pared SGD flux rates calculated using seepage meters, natural
tracers, and water balances at four sites with a wide range of
hydrogeological characteristics. A number of recent papers have
refined the use of natural tracers, for example, by investigating
in detail their behavior in groundwater (Gonneea et al., 2008),
presenting a more accurate way of calculating the errors asso
ciated with their activities (García-Solsona et al., 2008),
improving preparation methods for Ra standards (Dimova
et al., 2008), and preparing a set of international calibration
Ra standards. Advancements in direct SGD measurement have
also been made, including the development of a seepage meter
suitable for use in flowing water (Rosenberry, 2008), and an
investigation of how waves, which are common in coastal areas
where SGD occurs, affect seepage meter functioning (Smith
et al., 2008).
A third active research area addresses the behavior of nutri
ents, metals, and other constituents in the coastal aquifer.
Groundwater chemistry and quality can be highly heteroge
neous, so the earlier approach of assuming that the average
groundwater concentration of a certain constituent is represen
tative of the concentration at the discharge point may not
always be valid. Recent studies have investigated how nitrogen
is transported, transformed, and attenuated along its path to
the sea (Kroeger and Charette, 2008; Santos et al., 2008), as
228
well as the role of microbes in these processes (Santoro et al.,
2008). The transformations that metals undergo in coastal
aquifers (Charette and Sholkovitz, 2006; Rouxel et al., 2008;
Black et al., 2009) and the behavior of microbes in beach sands
and groundwater (Ishii et al., 2007; Yamahara et al., 2007,
2009) are also being investigated. Such investigations are cru
cial for understanding how changes in inland groundwater
quality may affect the coastal ocean via SGD.
Researchers are also attempting to develop a quantitative
understanding of the factors that contribute to variation in
SGD fluxes. While it has long been recognized that tides affect
SGD fluxes, recent studies (e.g., de Sieyes et al., 2008; Li et al.,
2009) have delved deeper into these effects, for example by
considering variations in tidal range in addition to simply tidal
height. A model of how seasonal differences in precipitation
can affect recharge and discharge of both fresh and saline
groundwater over the course of the year (Michael et al., 2005)
has been presented. Additionally, hydrogeologic modeling of
coastal aquifers has been expanded to include characteristics
representative of actual aquifers, including a heterogeneous
composition (Li et al., 2009) and the interaction between
fresh SGD and tidal pumping (Robinson et al., 2007).
Efforts are also focused on estimating SGD fluxes of water,
nutrients, and metals at a global scale. What percentage of
nitrogen, phosphorus, or iron in the Earth’s oceans originates
from SGD? Answering this question is important not only in
order to understand biogeochemical cycling today, but also in
order to correctly interpret the paleoceanographic record,
which is used to infer the characteristics of the Earth’s climate
in the geologic past. As groundwater often has different ele
mental concentrations and isotopic compositions than river
water, assuming that all terrestrial inputs were derived from
rivers can be problematic if SGD inputs were actually signifi
cant. Finally, work is in progress to determine the potential
effects of SGD on coastal ecosystems (Slomp and Van
Cappellen, 2004; Dale and Miller, 2008; Taniguchi et al.,
2008; Herrera-Silveira and Morales-Ojeda, 2009).
SGD is an important component of the hydrologic cycle
that affects coastal water quality in myriad ways. Although the
potential importance of SGD has been recognized and a num
ber of effective methods have been developed to measure it, the
study of this phenomenon is only beginning to mature. Many
questions remain unanswered; the answers to these questions
will affect coastal environmental management as well as our
understanding of global climate and biogeochemistry. Thus,
SGD research is likely to remain an active and exciting field for
decades to come.
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Knee and Paytan Ra - University of California, Santa Cruz

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