Marine Environmental Research 52 (2001) 51±79
www.elsevier.com/locate/marenvrev
Historical contamination of PAHs, PCBs,
DDTs, and heavy metals in Mississippi River
Delta, Galveston Bay and Tampa Bay sediment
cores
P.H. Santschi a,*, B.J. Presley b, T.L. Wade c,
B. Garcia-Romero c, M. Baskaran a,1
a
Texas A&M University, Laboratory of Oceanographic and Environmental Research, Department of
Oceanography, Galveston, TX 77551, USA
b
Texas A&M University, Department of Oceanography, College Station, TX 77845, USA
c
Texas A&M University, Geochemical and Environmental Research Group, College Station, TX 77845, USA
Received 15 January 2000; received in revised form 15 June 2000; accepted 15 August 2000
Abstract
Pro®les of trace contaminant concentrations in sediment columns can be a natural archive
from which pollutant inputs into coastal areas can be reconstructed. Reconstruction of historical inputs of anthropogenic chemicals is important for improving management strategies
and evaluating the success of recent pollution controls measures. Here we report a reconstruction of historical contamination into three coastal sites along the US Gulf Coast: Mississippi River Delta, Galveston Bay and Tampa Bay. Within the watersheds of these areas are
extensive agricultural lands as well as more than 50% of the chemical and re®nery capacity of
the USA. Despite this pollution potential, relatively low concentrations of trace metals and
trace organic contaminants were found in one core from each of the three sites. Concentrations and ¯uxes of most trace metals found in surface sediments at these three sites, when
normalized to Al, are typical for uncontaminated Gulf Coast sediments. Hydrophobic trace
organic contaminants that are anthropogenic (polycyclic aromatic hydrocarbons, DDTs, and
polychlorinated biphenyls) are found in sediments from all locations. The presence in surface
sediments from the Mississippi River Delta of low level trace contaminants such as DDTs,
which were banned in the early 1970's, indicate that they are still washed out from cultivated
* Corresponding author. Tel.: +1-409-740-4476; fax: +1-409-740-4786.
E-mail address: [email protected] (P.H. Santschi).
1
Present address: Wayne State University, Department of Geology, Detroit, MI 48202, USA.
0141-1136/01/$ - see front matter # 2001 Elsevier Science Ltd. All rights reserved.
PII: S0141-1136(00)00260-9
52
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
soils. It appears that the DDTs pro®le in that sediment core was produced by a combination
of erosion processes of riverine and other sedimentary deposits during ¯oods. Most of the
pollutant pro®les indicate that present-day conditions have improved from the more contaminated conditions in the 1950±1970's, before the advent of the Clean Water Act. # 2001
Elsevier Science Ltd. All rights reserved.
Keywords: PAHs; PCBs; DDTs; Metals; Sediments;
Bay; Tampa Bay
210
Pb;
239,240
Pu; Mississippi River Delta; Galveston
1. Introduction
The pro®les of trace contaminants in undisturbed sediment cores can be utilized to
estimate the extent and history of pollution in estuarine and coastal areas (e.g.
Santschi et al., 1999; Valette-Silver, 1993). Most contaminants of concern are particlereactive, with a tendency to sorb to suspended particles and are eventually deposited
as sediments. Particle-reactive contaminants are eciently removed to the underlying sediments, even on time scales smaller than the hydraulic residence time of
most estuaries, and the resulting deposition may be preserved in the sediment column. This allows for the reconstruction of historical pollutant inputs into coastal
areas. Reconstruction of historical inputs of anthropogenic chemicals is important
for improving management strategies and evaluating the success of recent pollution
control measures. Because trace elements can have both anthropogenic and natural
sources, the anthropogenic component must be identi®ed in order to recognize the
extent of pollution. One way to approach this is to normalize the data to ®ne
grained sediment content, as metals and organic contaminants are often concentrated in ®ne particles due to their greater surface areas and amounts of organic
carbon, clay, iron or aluminum (Blomqvist, Larsson, & Borg, 1992; Hanson, Evans,
Colby, & Zdanowicz, 1993; Mayer & Fink, 1980; Summers, Wade, Engle, &
Malaeb, 1996; Windom et al., 1989), thus facilitating comparisons. Another way is
to interpret the pro®le based on local sediment criteria.
Quantitative reconstruction of contaminant inputs requires a precise chronology. In estuarine and coastal areas, this chronology can be complicated by postdepositional mixing of sediments through physical, chemical and/or biological
mechanisms, which can potentially alter the original imprint in the sediments. In
addition, there are multiple sources of sediments resulting from riverine input,
shoreline erosion and dredging activities. These can provide di€erent types of
material with di€erent composition and grain size and may have experienced di€erent intermediate storage times before reaching a particular site. Additional geological processes, which can alter the fabric of sediments, include slumps, turbidity ¯ows
and diagenesis. None of the common radionuclides alone can take all of these
complications into account. However, use of two or more nuclides with di€erent
input functions has been successfully used in sorting out postdepositional processes,
which can alter the original imprint (e.g. Santschi, Li, Bell, Trier, & Kawtaluk, 1980;
Santschi et al., 1999). A common approach has been to use 210Pb, which has a constant atmospheric source, and one or several of the bomb fallout nuclides 239,240Pu
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
53
or 137Cs, which were introduced into the environment mainly in the late 1950s and
early 1960s, with a peak in 1963 (e.g. Santschi & Honeyman, 1989).
In areas where the e€ects of physical and/or biological mixing are small compared
to the sedimentation rate, the peak fallout retained in the sedimentary record corresponds to the year 1963. Concentrations of these nuclides can, however, remain
elevated to the present time due to drainage basin inputs (Ravichandran, Baskaran,
Santschi, & Bianchi, 1995a, b; Santschi, Nixon, Pilson, & Hunt, 1984; Santschi et
al., 1980, 1999).
A recent summary by Valette-Silver (1993) on the historical trends of contamination of estuarine and coastal sediments shows a recent decline of many organic and
inorganic pollutants concentrations, indicating that pollution control measures have
been e€ective. The objective of this study was to evaluate dated sediment cores from
Galveston Bay, the Mississippi River Delta, and Tampa Bay, to establish pollution
trends in these important Gulf Coast areas.
2. Sampling, materials and methods
2.1. Sampling sites
2.1.1. Mississippi Delta
The Mississippi River, as the dominant US river, drains more than 40% of the
conterminous US states, i.e. 2.9106 km2, including numerous cities and scores of
industrial facilities. The river is estimated to carry 60% of the dissolved salts and
66% of the suspended solids transported to the ocean from the US (Curtis, Curlberson & Chase, 1973; Leifeste, 1974). In its lower reaches the river winds through
extensive farmlands, marshlands and population centers before entering the Gulf
through the famous birdfoot delta. An Army Corps of Engineers diversion dam,
which sends about one third of the Mississippi River water down the Atchafalaya
River, is located just above Baton Rouge. The stretch of the river between Baton
Rouge and New Orleans is lined with dozens of chemical and petrochemical plants.
It has been well documented (Shokes, 1976) that sedimentation rates decrease
rapidly with distance from the mouth of the Mississippi River (e.g. the mouth of
Southwest Pass). Trefry, Metz, Trocine, and Nelson (1985) documented a decrease
in stable Pb concentrations, since its concentration peaked in the early 1970s, in a
core collected in 1982 about 20 km southwest of the mouth of Southwest Pass of the
Mississippi River, and dated with 210Pb. The 1970±1982 decrease was attributed to a
decrease in the use of leaded gasoline.
2.1.2. Galveston Bay
Galveston Bay, with a surface area of 1600 km2, is the second largest estuary in
Texas. The bay water is, however, shallow, averaging only about 2 m in depth, and
its exchange with the Gulf of Mexico is restricted because of the barrier islands of
Bolivar Peninsula and Galveston Island. Its average water residence time is about 40
days (Solis & Powell, 1999). The Dallas-Fort Worth Metroplex with a population of
54
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
about 4 million is located 640 km up the Trinity River from Galveston Bay. The
Houston metropolitan area with a population of about 3 million is immediately
adjacent to upper Galveston Bay and drains directly into it through the San Jacinto
River and the Houston Ship channel. The Houston±Texas City±Galveston area is
highly industrialized, especially by the petroleum, petrochemical, and chemical
industries. It has been estimated that 30±50% of the US chemical production and oil
re®neries is situated around Galveston Bay. Furthermore, Houston is the third largest seaport in the USA in terms of total shipping tonnage. Galveston Bay receives
more than half of the total permitted wastewater discharges for the State of Texas
and a total of about 5 km3 per year of wastewater input. The Houston Ship Channel
is often cited as one of the most polluted water bodies in the United States (EPA,
1980). Massive ®sh kills and other visible signs of pollution were common in the
1960s and 1970s. Because of these concerns, serious e€orts to clean up the ship
channel began in the 1970s. By 1976, the EPA was able to report that several Texas
waterways were getting cleaner and the Houston Ship Channel was singled out as
showing ``the most notable improvement'' (EPA, 1980). There is general agreement
that industrial discharges of pollutants to the ship channel and the rest of Galveston
Bay have declined in the past 15 years or so, yet the population in the drainage basin
has continued to increase and massive amounts of oil and chemicals are still shipped
across the Bay daily. Concentrations of nutrients (e.g. Santschi, 1995), trace metals
(e.g. Benoit et al., 1994; Stordal, Gill, Wen, & Santschi, 1996; Wen, Santschi, &
Tang, 1997; Wen, Stordal, Gill, & Santschi, 1996; Wen, Santschi, Paternostro, &
Gill, 1999), and trace organics (Gorham-Test, Wade, Engle, Summers, & Hornig,
1999) in open waters of Galveston Bay, as well as in surface sediments (e.g. Morse,
Presley, Taylor, Benoit, & Santschi, 1993; Warnken, Gill, Grin, & Santschi, 1999;
Wen, Shiller, Santschi, & Gill, 1999) do not indicate a greatly polluted system.
2.1.3. Tampa Bay
The Tampa Bay estuary, occupying 967 km2, is a shallow (average depth of 3.5 m
at mid-tide level), Y-shaped embayment located at the northern periphery of south
Florida's subtropical environment (Lewis & Estevez, 1988; Lewis, Clark, Fehring,
Greening, Johansson, & Paul, 1998) and is Florida's second largest estuarine system
(NOAA, 1990). The Tampa Bay watershed is the smallest (6739 km2) of the three
estuaries, resulting in a relatively long water residence time of 150 days (Solis &
Powell, 1999). This estuary's physical structure, such as historical circulation and
transport processes, has been considerably modi®ed within the past century, mainly
due to dredging of ship channels, construction of major causeways, and usage of the
estuarine sediments as shoreline land®lls. According to Lewis et al. (1988), approximately 10% (i.e. 101 km2) of the total open water area have shallow (<2 m) shelves
vegetated with seagrasses. The watershed supports a population of 2 million (as of
1995) with the cities of Tampa, St. Petersburg, Clearwater, Bradenton and surrounding suburban communities. Pollution from municipal discharges with nonexistent or inadequate treatment, industrial wastes from phosphate mines, citrus
canneries, and other sources were widespread in the 1950s (summarized in Lewis et
al.,). The City of Tampa installed an advanced wastewater treatment system for
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
55
nutrient removal in 1979, and the other cities followed after that. The estuary has
been a major seaport for over 100 years, and is rated the tenth largest in the United
States in overall tonnage (52 million tons/year).
2.2. Coring operation
One core from each sampling site was collected. The Mississippi Delta sediment
core was collected in 1993 approximately 24 km west of the mouth of the Southwest Pass of the Mississippi River (28 55.48270 N and 89 40.63520 W, Fig. 1) at a
water depth of approximately 60 m. This site was chosen so the changes in the
concentrations of trace metals could be compared with data, which were obtained
about 20 years ago at a nearby location (Presley, Trefry & Shokes, 1980). The
sediment cores were collected from Galveston Bay (28.630 N, 94.820 W, Fig. 1) in
1995 in lower Trinity Bay and in 1993 from Tampa Bay (27.740 N, 82.620 W, Fig. 1).
Cores were taken with a custom-built 2-m long gravity corer in the Mississippi
River Delta, and at the other sites, by a SCUBA diver using push corers with 6-cm
diameter plastic tubes. No compaction of sediment was noted in the core tubes
during coring. Total core lengths for each core is given in Table 1. While the cores
from Galveston Bay and Mississippi River Delta were taken in open waters using a
boat, the one from Tampa Bay was taken close to the shore line, at a site which
appeared physically undisturbed by dredging activities or other human activity
(Fig. 1). A 51-cm long core was taken by driving a plastic core tube into the ®rm
sediment by hand and digging it out in order to retain the largely coarse grained
material. The sediment varied in appearance and texture with depth as well as in
chemistry, and therefore, had not been homogenized physically. Cores were stored
refrigerated (2 C) until sectioning within less than one month, extruded in the lab
and cut into 1-cm sections, and the water content of each section was determined
after drying at 90 C.
2.3. Radiochemistry
The porosity (é) of each sample was calculated from the water content by
assuming a solid density of 2.50 g cm 3 (Santschi et al., 1999). For 210Pb analysis
by 210Po, the dried core sections were pulverized using an agate mortar and pestle.
Two gram aliquots were taken into solution by repeated digestion in conc. HF,
HNO3, HCl acid sequentially. A known amount of 209Po spike was added at the
beginning of each digestion to assess 210Po recovery. Polonium was electroplated
onto silver planchets (Ravichandran et al., 1995a, b; Santschi et al., 1980, 1984,
1999), and then assayed for 210Pb by using an alpha spectrometer with a surface
barrier alpha detector coupled S-100 Canberra Alpha Spectrometer. For 239,240Pu,
about 10 g of dried, powdered sediment was leached with hot 6 M HCl three times.
The leachates were combined and then processed for Pu after the addition of
242
Pu as a yield monitor. The separation and puri®cation of Pu was done by standard ion exchange techniques (Santschi et al., 1980, 1984, 1999; Ravichandran et al.,
1995a, b).
56
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
Fig. 1. Location of the Mississippi River Delta, Galveston Bay, and Tampa Bay sampling sites, in relation to the Gulf of Mexico.
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
57
Table 1
Inventories of excess 210Pb and 239,240Pu in sediment cores
Station
Core length (cm)
Galveston Bay
Mississippi Delta
Tampa Bay
44
83
51
210
Pbxs inventory (dpm cm 2)
245
1054
±
239,240
Pu inventory (dpm cm 2)
0.190.02
0.670.05
0.110.02
About 10±15 g of dried powdered sediment sample was placed in a gamma
counting vial and speci®c concentrations of 226Ra were determined using the 351
keV 214Pb gamma line for 226Ra. Excess-210Pb (210Pbxs) was calculated as the di€erence between total and 226Ra supported activities. Supported 210Pb activities were
calculated from constant 210Pb activities at depth and/or 226Ra activities. Even
though gamma counting allows for the determination of 137Cs (661.6 keV), the
pro®les we measured were deemed unreliable due to low count rates, and are not
presented. Either National Institute of Standards and Technology (NIST) radioactive standard solutions or solid material, or radioactive standards that are calibrated with NIST standards, were used for calibrating the counting equipment.
Precision in the concentration of 210Pb, 239,240Pu and 226Ra was better than 5%.
Since radioactive spikes are traceable to those supplied by NIST and counting
equipment was well-calibrated, accuracy was similar to precision.
The sediment inventory, I (dpm cm 2) of 210Pb or 239,240Pu in a sediment core was
calculated by summing each of their speci®c activities Ai (dpm g 1) down to a depth
where 210Pbxs or 239,240Pu concentrations were detectable. The inventory of 210Pbxs
or 239,240Pu, in dpm cm 2, in a sediment core was calculated using the equation:
X
Ai mi
I …dpm cm 2 † ˆ
where mi is the mass-depth increment corresponding to the depth interval (g cm 2)
and Ai is the activity concentration.
2.4. Elemental analysis
Elemental analyses were performed by atomic absorption spectroscopy (AAS),
instrumental neutron activation analysis (INAA) and/or inductively coupled plasma
spectrometry (ICP) depending on the metal and the concentration (Presley, Wade,
Santschi, & Baskaran, 1998). The most sensitive method for each metal was always
used when concentrations were low to insure accurate and precise values for all
metals in all samples. This requirement meant that many of the analyses were by
graphite furnace AAS.
Sample preparation started with freeze-drying a representative sediment aliquot
and grinding it to a ®ne powder. No further treatment was needed for INAA; thus
this technique provided a check on the sample dissolution techniques that were
applied for AAS analysis. For INAA, 0.5 g aliquots of the powdered samples were
weighed directly into plastic vials and heat sealed. They were then irradiated for 8 h
58
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
in the 1 megawatt TRIGA reactor at Texas A&M University. After a 7-day cooling
period to allow Na, Cl and other interfering isotopes to decay to low levels, the
samples were counted using a hyper pure germanium detector coupled to a Nuclear
Data Corp. model 9900 multichannel analyzer integrated with a Digital VAX II/
GPX graphics workstation. Concentrations were obtained by comparing counts for
each sample with those for sediment and rock reference materials of accurately
known elemental composition. Details of this method are given in James, Boothe
and Presley (1998), including information on counting geometry, reference materials,
spikes, blanks and other aspects of Quality Assurance/Quality Control (QA/QC).
The National Status and Trends Program methods (Lauenstein & Cantillo, 1993)
were used in the AAS analysis. Brie¯y, 200-mg aliquots of the powdered sediment
samples were weighed into te¯on ``bombs'' and completely dissolved in a mixture of
nitric, hydro¯uoric and boric acids by prolonged exposure of the closed bombs to a
temperature of 130 C. Various dilutions were made on the clear digests to bring
them into the working range of the AAS. A Perkin-Elmer Corp model 306 ¯ame
AAS or a Spectro ICP was used for Fe, Mn and Zn analysis essentially following the
manufacturers instructions. Other elements were determined using a Perkin-Elmer
3030Z equipped with an HGA-600 graphite furnace and an auto sampler. Details of
furnace programs, matrix modi®ers, blanks, spikes, reference materials and other
QA/QC information can be found in the reference given above. Matrix spike
recovery for all elements was almost always >90%, as were recoveries of certi®ed
values on reference materials from the National Research Council of Canada. Several elements were determined by both AAS and INAA and di€erences between
results from the two methods were generally less than 10%.
2.5. Organic trace contaminant analysis
Samples were stored frozen ( 20 C) until analysis. Sediments for organic analyses were extracted within the prescribed holding times using the methods described
in Wade et al. (1988). Approximately 10 g of freeze-dried sediment were soxhletextracted with methylene chloride. The solvent was concentrated to approximately
20 ml in a ¯at-bottomed ¯ask equipped with a three-ball Snyder column condenser.
The extract was then transferred to Kuderna-Danish tubes, which were heated in a
water bath (60 C) to concentrate the extract to a ®nal volume of 2 ml. During concentration of the solvent, dichloromethane was exchanged for hexane.
The extracts were fractionated by alumina:silica (80±100 mesh) open column
chromatography. Silica gel was activated at 170 C for 12 h and partially deactivated
with 3% (v/w) distilled water. Twenty grams of silica gel was slurry packed in
dichloromethane over 10 g of alumina. Alumina was activated at 400 C for 4 h and
partially deactivated with 1% distilled water (v/w). The dichloromethane was
replaced with pentane by elution, and the extract was applied to the top of the column. The extract was sequentially eluted from the column with 50 ml of pentane
(aliphatic fraction) and 200 ml of 1:1 pentane-dichloromethane (aromatic-pesticide
fraction). The fractions were then concentrated to 1 ml using Kuderma-Danish
tubes heated in a water bath at 60 C.
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
59
Quality assurance for each set of 20 samples included a procedural blank and a
matrix spike, which were carried through the entire analytical scheme and an
appropriate standard reference material (i.e. SRM 1941a). All internal standards (surrogates) were added to the samples prior to extraction and were used for
quantitation.
Aromatic hydrocarbons were separated and quanti®ed by gas chromatographymass spectrometry (GC-MS) (HP5890-GC and HP5970-MSD). The samples were
injected in the splitless mode onto a 0.25 mm30 m (0.32 mm ®lm thickness) DB-5
fused silica capillary column (J&W Scienti®c Inc., or equivalent) at an initial temperature of 60 C and temperature programmed at 12 C/min to 300 C and held at he
®nal temperature for 6 min. The mass spectral data were acquired using selected ions
for each of the polycyclic aromatic hydrocarbon (PAH) analytes. The pesticides and
polychlorinated biphenyls (PCBs) were separated by gas chromatography in the
splitless mode using an electron capture detector. A 30 m0.32 mm I.D. fused silica
column with DB-5 bonded phase (J&W Scienti®c or equivalent) was used. The
chromatographic conditions for the pesticide-PCB analysis were 100 C for 1
min, then 5 C/min to 140 C, hold for 1 min, then 2.5 C /min to 250 C, hold for
1 min, and then 10 C/min to 300 C, and a ®nal hold of 5 min.
Holding times were less than speci®ed by EPA. Precision and accuracy of the
organic chemicals were established by analytes of standard reference materials from
NIST. The sum of 24 major PAH compounds was used to calculate total PAHs, and
18 major congeners for PCBs.
3. Results and discussion
Downcore variations of water and Al content can be taken as proxies for grain
size variations in sediment cores. Water, Al, and Fe, and often, also organic carbon
content covary in sediment cores (Hansen et al., 1993; Horowitz & Elrick, 1987;
Presley, Tayor, & Boothe, 1992; Ravichandran et al., 1995a), allowing one to use
one in place of the other. Fig. 2 shows co-varying water and Al content pro®les for
the three di€erent sediment cores, suggesting that grain size variations in Mississippi
River Delta and Galveston Bay cores are relatively small, while those at the Tampa
Bay site are substantial. Before 1970, it appears that the Tampa Bay site was poor in
Al, i.e. sandy, while after that time, sediment became more Al rich, i.e. more ®ne
grained.
3.1. Geochronology
Sedimentation rates can be calculated from time markers, or from pro®les of
chemicals with a known decay rate. Methods applied for age dating of sediments are
as follows:
1. From bomb fallout nuclide (137Cs, 239,240Pu) peak concentration in 1963: This
method can be applied to sediment cores, even under conditions of sediment
60
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
Fig. 2. Pro®les of water and Al content for the three di€erent sampling sites.
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
61
focusing. Sedimentation rates were calculated assuming a fallout maximum in
1963.
2. From excess- 210Pb (210Pbxs=[210Pb]total [226Ra]) pro®les: In the simplest case,
sedimentation rates can be calculated under steady state conditions from Eq.
(1a±c).
‰210 Pbxs …z†Š ˆ ‰210 Pbxs …0†Š exp… z†
…1a†
ˆ …l=S†
…1b†
ˆ …l=Sa †
…1c†
where [210Pbxs(z)] and [210Pbxs(0)] represent the excess 210Pb concentration at depth z
(in cm or g cm 2) and at the sediment surface, respectively; l, decay constant of
210
Pb (0.031 yr 1); S, linear sedimentation rate, in cm/yr; Sa, sediment accumulation
rate, in g cm 2 yr 1.
This approach assumes that: (1) particle reworking rates are negligible over the
depth interval of the 210Pb pro®le; and (2) initial surface sediment concentrations of
210
Pb had been constant over time. If these assumptions are ful®lled, the Constant
Initial Concentration model can be applied to 210Pbxs data where sedimentation
rates and 210Pb ¯uxes vary little with time. This model is valid even under conditions
where sediment focusing is occurring, provided that conditions (i.e., initial concentrations) had been constant over time as well. Even though numerical models
utilizing bomb fallout input functions are available for including particle reworking
into sedimentation models (e.g. Fuller, van Geen, Baskavan, & Anima, 1999; Olsen,
Simpson, Peng, Bopp, & Trier, 1981; Santschi et al., 1980, 1999), they have not been
used here due to the more complicated nature of the sedimentation processes in
estuaries, e.g. drainage basin inputs, which are an unknown part of the input function. Instead, we relied on the apparent agreement between 210Pbxs and 239,240Pu
based geochronologies.
The depth distribution of excess 210Pb (210Pbxs) and 239,240Pu activity concentration in sediment cores collected from all three estuaries are shown in Figs. 3 and 4,
respectively. The inventories of these nuclides for the three estuaries are given in
Table 1. The radionuclide pro®les and inventories of individual estuarine systems
are discussed below.
3.1.1. Mississippi Delta
The 239,240Pu pro®le (Fig. 4) clearly shows a sharp peak corresponding to the 1963
peak fallout. The retention of the atmospheric fallout history in this core is an indication of negligible biological and/or physical mixing. 239,240Pu becomes nondetectable below a depth of 40 cm. The linear sedimentation rate based on the
239,2340
Pu peak (corresponding to 1963), 0.7 cm yr 1, is within the error of
210
the Pbxs-based linear sedimentation rate of 0.62 cm yr 1. The mass accumulation
rates, which take compaction of sediments into account, obtained by both methods
are in good agreement as well (239,240Pu-peak accumulation rate=0.39 g cm 2
62
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
Fig. 3.
210
Pbxs as a function of mass depth, for two di€erent sampling sites.
year 1; 210Pbxs-based mass accumulation rate=0.4 g cm 2 year 1, Figs. 3 and 4).
The close agreement between the 239,240Pu-peak and 210Pbxs-based methods con®rms
that there was very little mixing of sediment layers in this core. This observation is
consistent with earlier studies that bioturbation is minimal in areas where sediment
accumulation is fast, i.e., above 0.5 cm year 1 (Fuller et al., 1999; Olsen et al., 1981).
The calculated inventories of 210Pbxs and 239,240Pu are 105 and 0.67 dpm cm 2,
respectively (Table 1). These values are considerably higher (235 and 218% higher
for 210Pb and Pu, respectively) than would be expected from the direct atmospheric
fallout (45 dpm cm 2 for 210Pb and 0.3 dpm cm 2 for 239,240Pu, based on 1.5 m
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
Fig. 4.
239,240
Pu activity concentration pro®les for the three di€erent sampling sites.
63
64
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
year 1 rainfall, and Houston area estimates (Baskaran, Coleman, & Santschi, 1993;
Ravichandran et al., 1995a, b; Santschi et al., 1999). These higher values indicate
some focusing e€ect due to riverine material sedimenting out in the delta region, as
well as from boundary scavenging e€ects (Oktay, Santschi, Moran, & Sharma,
2000), thus enhancing the marine derived inventory from direct fallout. Provided
that this condition has been constant over time, sedimentation rates calculated from
210
Pbxs would not be a€ected. The recent (i.e. post-1963) sediment accumulation rate
in this core of 0.6±0.7 cm year 1, derived from Pu and Pb isotopic data, is within the
error of 0.7 cm year 1, reported for another core taken in parallel with this one, by
Oktay et al.
3.1.2. Galveston Bay
The 239,240Pu pro®le from Galveston Bay core (Fig. 4) shows a broad peak from
10±14 cm depth (5.5 g cm 2 cumulative mass depth), which is attributed to the
year 1963. Dividing the average peak depth of 12 cm by the time di€erence to 1963
(i.e. 32 years) would result in a sediment accumulation rate of 0.17 g cm 2 yr 1 (or
0.38 cm/yr). The sediment inventory in this core of 239,240Pu (0.19 dpm cm 2, Table
1) is comparable to the expected atmospheric fallout value of 0.20 dpm cm 2
(Santschi et al., 1999; Ravichandran et al., 1995a, b).
Since Pu shows low concentrations much deeper than expected by assuming the
same sedimentation rate to 1953, when the bomb fallout began (Fig. 4), the geochronological interpretation of this core is not as straightforward. This deeper
penetration of Pu down to 40 cm depth in this core from Galveston Bay could be
attributed to a non-steady state sedimentation regime, where sediment material has
been deposited at higher rates before the year 1963 (see below). Alternatively, it
could be taken as an indication of a lowering of sedimentation rates after the 1969,
when 90% of the sediment delivered to Galveston Bay started to be retained behind
the Livingston Lake dam (e.g. Van Metre & Callender, 1996).
The 210Pbxs pro®le in the upper 20 cm of sediments from Galveston Bay (Fig. 3) is
consistent with a sediment accumulation rate of 0.16 g cm 2 year 1. Below 20 cm
depth, the 210Pbxs decreases at uncertain and variable rates (not shown). Small
210
Pbxs activities appear until the core bottom, suggesting a non-steady state system.
The higher apparent sedimentation rates derived from 210Pbxs in that lower section,
taken together with the near-constant 239,240Pu activities in that layer, could have
been caused by either material eroded from shore lines in the late 1950s and subsequently deposited in the sampling area, or by the higher sediment delivery before the
construction of the Lake Livingston dam. The total inventory of 210Pbxs (above 20
cm) in that core of 24 dpm cm 2 is somewhat lower than what would be expected
from direct atmospheric fallout of 33 dpm cm 2 (based on 3 years average fallout
value at Galveston of 1.03 dpm cm 2 year 1). If the low values below 20 cm would
be included, the inventory in that core would be close to that expected from fallout.
3.1.3. Tampa Bay, Florida
No signi®cant 210Pbxs was measurable in this sediment core. The only measured
isotope that can provide geochronological information is thus 239,240Pu. However,
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
65
the highest Pu concentration was observed in the upper 1-cm layer. Because Al and Fe
concentrations are low, tightly correlated (R=0.99), and vary by an order of magnitude in this core, Pu was normalized to Al (Fig. 4). This normalization identi®es the
depositional event corresponding to the 1963 peak fallout at about 35 cm. Thus, the
average sediment accumulation and sedimentation rate would be about 1.4 g cm 2
year 1 and 1.1 cm year 1, respectively. Despite this high sediment accumulation rate,
the inventory of 239,240Pu is low, i.e. only 0.11 dpm cm 2, which is considerably lower
than the inventory expected from direct atmospheric fallout of 0.20 dpm cm 2
(assumed to be similar to that of the Houston area). Even though the geochronology
of this core is thus not as well constrained as the others, pesticide distributions in this
core (described below) con®rm the relatively high sedimentation rates.
3.2. Trace contaminants
The concentrations of trace metals can vary not only due to contamination but also
due to variations in grain size, mineralogy, and concentrations of organic carbon, Al,
Fe and Mn. A commonly used method is to normalize the metal concentrations to
Al, which corrects for variations of grain size and mineralogy (Alexander, Smith,
Calder, Schropp, & Windom, 1993; Bruland, Bertine, Koide, & Golberg, 1974;
Goldberg, Grin, Hodge, Koide, & Windom, 1979; Schropp, Lewis, Windom,
Ryan, Calder, & Burney, 1990; Summers et al., 1996; Trefry et al., 1985; Windom et
al., 1989). Iron (Sinex & Wright, 1988; Trefry & Presley, 1976) has also been used
even though it can be susceptible to anthropogenic changes (in the form of Fe sul®des
and Fe oxides) and post-depositional redox processes. The metal to Al ratios are
relatively constant in soil and crustal material and are less likely to be a€ected by
human activities (Alexander et al., 1993; Schropp et al., 1990). This would only be
true for elements whose carrier phases are closely associated with clays. Normalization to Al might therefore not always be appropriate for Hg, which shows strong
association with sul®des in sediments (Coakley & Poulton, 1993; Schropp et al., 1990,
Summers et al., 1996). In our study, concentrations of metals were normalized to Al when
Al showed substantial downcore variations. This was the case for the Tampa Bay core,
where Al concentrations were very low, and varying by an order of magnitude downcore.
The historical trends of selected trace metal and trace organic contaminant concentrations are plotted in Figs. 5±8 for all three estuarine systems, using the sediment geochronology described above.
3.2.1. Mississippi River Delta
As shown in Fig. 5, most individual PAHs increased in concentrations starting in
the early 1940s, peaking in the 1970s. The increasing concentrations of PAHs
observed are likely the result of increasing anthropogenic activities in the drainage
basin of the Mississippi River. Sporadic input events are identi®ed as isolated concentration peaks (much larger than the analytical error) in one sediment section;
whereas episodic input events, which lasted longer, are seen where at least two
adjacent core sections show increasing or decreasing concentrations reaching a
maximum or a minimum concentration. Concentrations of PAHs quadrupled from
66
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
Fig. 5.
P
P
P
PAHs, PCBs and DDT concentration pro®les for the three di€erent sampling sites.
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
Fig. 6. Comparison between (a)
sissippi River site.
67
P
P
DDTs and Chlordanes and (b) DDTs and Hg pro®les at the Mis-
1940 to 1960, peaked in the early 1970s, and show only a moderate decrease thereafter (Fig. 6). The fastest rate of increase occurred during the 1950s. The PAH distribution does not show the transition from coal to petroleum fuel which is
characterized by a large PAH concentration peak in the 1940±1950 period (Bates,
Hamilton, & Cline, 1984; Zhang, Christensen, & Yan, 1993). Distribution of individual PAH compounds indicate varying proportions of petroleum and combustion
sources. Ranges of individual PAH concentrations, and their respective concentrations, are further discussed in Presley etPal. (1998).
The concentrations of total DDTs ( DDT), calculated as the sum of o,p0 and
p,p0 -DDT isomers and respective metabolites DDE and DDD, were very low
throughout the core (Fig. 5). However, a sharp concentration increase is evident in
the early 1950s, following the increasing production and use of DDT, increasing
steadily towards the present after the
P 1960s. The highest concentration is found at
the sediment surface. The fact that DDT concentrations did not decrease after the
ban in 1972 likely indicates that their concentrations are controlled by both agricultural runo€ (e.g. Barber & Writer, 1998), as they are still present in cultivated
soils, and by a combination of processes controlling erosion (i.e. resuspension) of
riverine and other sedimentary deposits P
(Barber & Writer, 1998).
A similar distribution is observed for Chlordanes (Fig. 6a). These ``anomalous''
chemicals, DDTs and Chlordanes, now have mainly di€use sources in the watershed, and experienced broad applications to agricultural soils as well as local (point)
discharges in the 1960s and 1970s. Point sources of these chemicals, however, have
decreased dramatically in recent years (Barber & Writer, 1998; Bergqvist, Strandberg, Ekeluna, Rappe, & Granmo, 1998). The increasing concentrations to the present of these chemicals can be attributed to drainage basin and riverine deposit
erosion, which has been observed to increase in recent years for many herbicides,
especially during the 1993 ¯ood (Clark, Goolsby, & Battaglin, 1999), which occurred before the core was taken. Flooding events have increased in frequency and
intensity since the 1960s (USGS data). The data from this core are not sucient to
document the occurrence of speci®c ¯ood events.
Concentrations of o,p0 -isomers of DDT were approximately 10 times lower than
the p,p0 -isomers, which may be a result of their lower concentration proportions in
68
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
Fig. 7. Pb, Ag, Cd concentration pro®les for the three di€erent sampling sites.
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
69
Fig. 8. Cu, Zn, Ba concentration pro®les for the three di€erent sampling sites.
70
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
commercial DDT technical compositions, higher water solubility and higher rates of
microbial degradation. Concentrations of p,p0 DDT ranged from ND to a maximum
of 0.34 ppb in 1993, those of p,p0 DDE from ND to 0.69 ppb in 1993, those of
p,p0 DDD from ND to 0.44 ppb in 1982. It is certainly noteworthy that bed sediments of the upper Mississippi River showed enhanced p,p0 DDE and p,p0 DDD
concentrations after the 1993 ¯ood (Barber & Writer, 1998), with values of 0.5±1
ppb, similar to values in the 1993
P layer of our core.
The vertical distribution of PCBs, calculated as the sum of all PCB congeners
(Fig. 5), showed a steady increase after its production increased in the early 1950s,
peaking in the early 1970s, following the decline in its use after regulations were
imposed. PCBs with four chlorine atoms had consistently the highest percentages
(25%) of the PCBs. Individual congeners had di€erent vertical distributions which
were not related to their molecular weight, concentration ranges or percentages in
commercial arochlors. Approximately 60% of the individual congener concentrations showed no maximum
in the early 1970s, and only around 20% had similar
P
distributions to the PCBs. The fact that the vertical distributions and concentration ratios of individual PCB congeners are dissimilar to commercial PCB aroclor
compositions (WHO, 1993) suggests that alterations due to various physical and
biochemical fractionation processes must haveP
occurred.
The di€erence between the distribution of PCBs, which decreased drastically
after their ban in the early 1970s, and that of DDTs, which, even though banned
in 1972, continue
to reside in agricultural soils, is most dramatic. It is noteP
worthy that PCBs inputs into the Mississippi River drainage basin were often
through point sources (e.g.
P Steingraeber, Schwartz, Wiener, & Lebo, 1994; Barber
& Writer, 1998), while DDT inputs are through di€use sources such as runo€ of
DDTs contaminated sediments in the upper watershed during ¯oods (Barber &
Writer, 1998). PCB emissions resulting from insulator material point sources were
likely removed, while DDTs in agricultural soils and stream and river deposits
continue to pose a source to the coastal environment during high run-o€ conditions (e.g. Bergqvist et al., 1998; Cooper, 1991; Cooper, Dendy, McHenry,
& Ritchie, 1987). Similarly to DDTs, Chlordane, another pesticide (Fig. 6a)
showed a strong increase in the 1980s up to the present time. Sources of DDTs in
the lower Mississippi River watershed (e.g. Cooper, 1991; Cooper et al., 1987;
Reich, Perkins, & Cutter, 1986) resulting from heavy prior applications of DDTs
in cotton ®elds and other agricultural settings likely contribute to the present-day
load as well.
A distribution similar to the DDTs and Chlordanes is also seen for one trace
metal, Hg (Fig. 6b), which shows an increase towards the sediment surface. The
reason could be that an important fraction of Hg is presently introduced with pesticides. These constituents are present in agricultural soils and streambeds, which are
subject to ¯ood erosion. For example, Balogh, Engstrom, Almendinger, Meyer and
Johnson (1999) state that human activities resulting in landscape disturbance, such
as agriculture, logging and urbanization enhance the delivery of Hg to surface
waters. These di€use Hg sources now exceed point source discharges, and constitute
the major inputs to the upper Mississippi River.
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
71
Trace metal concentrations (Figs. 7 and 8) in this core are similar to those of
average crustal abundances and average uncontaminated Gulf Coastal ®ne-grained
sediment. Thus, there is no indication of massive metal contamination. There is,
however, clear evidence of human in¯uence in the metal pro®les. Such in¯uence had
been noted in previous work on cores collected within a few miles of the location of
this one (e.g. Trefry et al., 1985). Several trace metal concentrations followed similar
changes (Figs. 7 and 8). Pb distributions (Fig. 7), which showed a peak in the early
1970s, are consistent with Trefry's suggestions that Pb follows its atmospheric
deposition pattern from gasoline exhausts. Not only are the peak shapes we found
similar to Trefry's, also the concentration levels are similar, though not identical.
Concentrations in the late 1890s and early 1900s are about 25 ppm. Concentrations
start to increase in the early 1950s, peaking in the early 1970s, at about 37 ppm. The
decline in recent years follows the decreased use of leaded gasoline in the USA, as
suggested by Trefry et al. (1985) Other metals, such as Ag, Cd (Fig. 7), and Zn (Fig.
8), but not Cu (Fig. 8), show increases and decreases similar to those of Pb. The
reason why Cu is behaving di€erently from Pb is that Cu was not as elevated over
ambient concentrations in the 1970s as was Pb. Thus, even though both elements
show peaks during that time, and improvements towards the sediment surface, they
di€er in their shape before the 1970s, possibly due to di€erent input functions Ba
shows a steady increase from the 1960s to the mid-1980s. Ba concentrations are near
the expected concentrations for uncontaminated Gulf Coast sediments in the core
bottom (about 450 ppm), begin to increase in the 1940s, coincident with the beginning of oil well drilling mud use in this area. The concentrations increase gradually
until 1980, when a sharp increase doubles the concentration. It then triples in the
late 1980s. Values in the upper 4 cm of the core decrease to about double background concentrations, probably due to fewer drill mud discharges in this area in
recent times. Concentrations of other metals, such as Cr, Sn, Sb, Ni, Hg, etc., show
variations similar to some of the metals discussed here, and are further discussed in
Presley et al. (1998). No evidence exists for signi®cant diagenetic remobilization of
Fe and Mn.
Van Metre, Wilson, Callender, and Fuller (1998) recently noticed a similar rate of
decrease of bomb fallout nuclides, selected trace metal and trace organic pollutants
in sediment cores from reservoir lakes in riverine systems of the USA, with highest
inputs in the 1960s and 1970s, and with di€erences in half-times explainable by land
use practises in the drainage system. Urban watersheds had half-times of the order
of 10 years, while agricultural soils resulted in slower rates with half-times of the
order of 15 years or longer. Pu concentrations in our core decrease after 1963 with a
half-time of about 15 years, PCBs decrease after the early 1970s with a 10 year halftime, Cd with a 17 year half-time, and Pb with a 47 year half-time (Fig. 9). While
10±15 years for Pu, Cd and PCBs appear to be characteristic for urban point and
di€use sources, 50 year half-times for Pb indicate continuing but decreasing sources in the Mississippi River watershed. For example, power plant emissions from the
combustion of fossil fuels, municipal incineration of waste products and discharge
of waste water, even though decreasing in magnitude, are still major sources for Pb
(summarized in Moore, 1991).
72
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
P
Fig. 9. Rate of decrease for 239,240Pu, Cd, Pb, and PCBs after their subsurface maxima in the 1960s
P
(Pu) and 1970s (Cd, Pb, PCBs) in the core from the Mississippi River Delta.
3.2.2. Galveston Bay
Even though the core was collected in the vicinity of the highly industrialized
western shore of Galveston Bay, it showed low concentrations of all measured
organic contaminants (Fig. 5). The highest concentrations of PAHs, DDTs and
PCBs occurred in the late 1960's, when the Houston Ship channel was most polluted
(EPA, 1980), and before pollution clean-up measures were initiated for this area.
Although not as dramatic as for the trace organic contaminants, trace metal concentrations (Figs. 7 and 8) were generally also highest in the 1960s and 1970s (e.g.
Al, Pb, and especially Ag and Cd). However, slightly elevated concentrations were
observed for Zn and Cu (Fig. 8) since the early 1900s. Barium (Fig. 8) showed larger
concentration changes with depth than any other metal. The bottom half of the
core, below about 19 cm, had about 300 ppm Ba, a typical value for uncontaminated
Gulf Coast estuarine sediment. From 19 to 14 cm, Ba concentrations increase sharply, reaching 1800 ppm. The source of the added Ba is almost certainly oil well
drilling mud. There is an approximately 5-cm thick layer of this Ba-rich sediment
and then a decrease to the top of the core. Barium then, like most metals, seems to
indicate improved conditions in Galveston Bay in recent times. Similarly decreasing
Ba concentrations were also observed in sediment cores from Lake Livingston, a
reservoir lake of the main freshwater source of Galveston Bay, the Trinity River.
3.2.3. Tampa Bay
The sediment appeared coarse-grained, carbonate-rich, varied in appearance, texture and chemistry with depth. The pro®les of percent water and Al (Fig. 2) testify
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
73
to this layered structure, showing a lower water content until the 1970s, and an
increase thereafter, accompanied with an increase in Al content and likely, grain
size (not determined). Therefore, we conclude that the pro®les had not been seriously compromised. Even though this sediment core was coarse-grained and had
the lowest content of Al (and thus, of clay minerals) and of natural and fallout
radioisotopes, and was low in trace metals, the organic contaminant concentrations
were higher than those of the Mississippi Delta and Galveston Bay cores. All
organic contaminants (except DDTs) were in higher concentrations in surface
sediments, but were detected at all depths. The concentrations of PAHs (Fig. 5)
were exceptionally high in this core, with a peak in the 1980s. Carr et al. (1996)
reported sediment toxicity to sea urchins in fertilization tests, and attributed this to
high concentrations of PAHs, PCBs, DDTs, Pb and Zn in these sediments. Peaks of
DDTs occurred in the mid 1960s, and those of PCBs increased steadily to the present (Fig. 5). The exceptionally high DDTs spikes deposited in the 1960s, during the
times of highest production and agricultural applications, are likely local in nature,
as the core was collected closest to shore. Conversely, we can take the fact that we,
as expected, found DDTs to peak in the 1960s as a con®rmation of the sediment
dating technique. OC was not determined for this core. Normalization to organic
carbon (OC) concentrations would not necessarily help in the interpretation. In
most cases, OC concentrations in sediments are closely related to grain size, and
thus, to Al (Horowitz & Elrick, 1987).
Because of the large variations of the Al content in this sediment core, and its
clay-poor nature, trace metal concentrations shown were normalized to Al (Figs. 7
and 8). Since Al concentrations highly correlate with those of Fe (R value of 0.99),
this normalization to the clay carrier phase also normalizes it to the iron oxide/sul®de carrier phase. The normalized depth distribution of the di€erent trace metals
resembles those in the Mississippi River Delta. Consistent Al-normalized concentration maxima are observed in the late 1960s to early 1970s for Ag, Pb, Zn and
Cu, and for Cd in the late 1950s. The maximum Ba/Al ratio occurred near the
sediment surface (Fig. 8). The absolute concentrations of some of the metals are
higher than would be expected for coarse-grained sediment, and therefore, indicate
e€ects of pollution.
3.3. Degree of contamination of surface sediments at all three sites, from
comparisons to natural levels
Concentrations of trace metals, predicted for uncontaminated sites in the Gulf of
Mexico from their Al content and the algorithms given by Summers et al. (1996), are
given in Table 2. When compared to the actual surface concentrations in the upper 2
cm, given in Table 3, there are only a few elements or compounds that are slightly
outside the upper limit (95% con®dence limit) predicted for uncontaminated sites:
Cu (by 30%), Hg (by 17%) for the Mississippi River site, and Cd (by a factor of
6), Cu (by a factor of 4), Hg (by a factor of 4) and Pb (by a factor of 4) for the
Tampa Bay site. Trace organic contaminants showed much larger variations from
site to site, with highest concentrations in the Tampa Bay core as well. For
74
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
Table 2
Concentrations for uncontaminated sediments in Galveston Bay, the Mississippi River Delta region, and
Tampa Bay, predicted from their Al content (Summers et al., 1996)a
Element
Galveston Bay
Al (%)
Pb (ppm)
Hg (ppm)
Cd (ppm)
Ag (ppm)
Cu (ppm)
Zn (ppm)
Cr (ppm)
a
Mississippi River
7.89
23 (30)
0.07 (0.13)
0.26 (0.50)
0.17 (0.27)
17 (20)
101 (170)
69 (84)
Tampa Bay
6.94
21 (28)
0.06 (0.12)
0.22 (0.44)
0.15 (0.25)
15 (18)
89 (155)
62 (77)
0.64
4 (7)
0.02 (0.06)
0.04 (0.16)
0.05 (0.15)
2 (5)
11 40)
10 (25)
Values given in brackets are for the upper limit, 95% con®dence limit.
Table 3
Present-day trace contaminant ¯uxes to surface sediments of Galveston Bay, the Mississippi River Delta
region, and Tampa Bay
Species
Galveston Bay
(Sa=0.17 g cm2 year 1)
Mississippi River
(Sa=0.39 g cm 2 year 1)
Tampa Bay
(Sa=1.385 g cm2 year 1)
Concentration Total ¯ux
Concentration Total ¯ux
Concentration Total ¯ux
(g cm2 year 1) (g/g)
(g cm2 year 1)
(g g)
(g cm2 year 1) (g g 1)
Pb
Ba
Hg
Cd
Ag
Sb
Cu
Mn
Sn
Zn
Cr
Fe
P
PAHs
P
PCBs
P
DDTs
26.61
798
0.08
0.157
0.15
0.59
13.94
792
1.99
107
68.32
3.77
0.32
0.0068
0.0003
4.5237
135.66
0.0136
0.02669
0.0255
0.1003
2.3698
134.64
0.3383
18.19
11.614
0.6409
0.054
0.0012
0.000054
27.41
871
0.139
0.175
0.16
1.38
21.31
1044
1.94
144
71.62
4.15
0.66
0.007
0.0016
10.416
330.98
0.05282
0.0665
0.0608
0.5244
8.0978
396.72
0.7372
54.72
27.216
1.577
0.25
0.0027
0.00061
15.7
32
0.06
0.256
0.12
0.08
9.1
9.6
0.49
30
20
3172
6.26
0.028
0.011
21.7
44.3
0.083
0.355
0.166
0.111
12.6
13.3
0.68
41.6
28
4393
8.67
0.039
0.015
example, concentrations of PAHs, PCBs and DDTs in surface sediments of Tampa
Bay were higher by an order of magnitude than at the two other sites. High pesticide
and heavy metal concentrations in the Tampa Bay core might be due to the much
closer proximity to the shoreline than for the other two sites. The higher concentrations in the Tampa Bay core are fortuitous. For example, if a sediment core
for the Houston ship channel were analyzed for the same compounds, concentrations similar or higher than those in the Tampa Bay would be found. Due to the
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
75
variable extent contaminated riverine particles in these three estuaries have had the
chance to be diluted by ambient particles, or a€ected by local inputs, comparisons
between the three di€erent sites are dicult.
3.4. Recent trace metal ¯uxes
Present-day ¯uxes (F) of total metals (in g cm 2 yr 1) were estimated in Table 3
for all three sites. Fluxes, as well as concentrations, of most trace metals are within
the range reported previously for Gulf of Mexico sediments which are relatively
uncontaminated (e.g. Beck et al., 1990; Ravichandran et al., 1995a, b; Santschi et al.,
1999, as summarized in Wen, Shiller et al., 1999). PAHs, PCBs and DDTs ¯uxes at
the Tampa Bay site are, however, higher than ¯uxes at the other two sites by orders
of magnitude. Because it is not possible to rule out pesticide and PAH contamination from local sources close to shore at that site, we cannot necessarily conclude
that all of Tampa Bay is as polluted by these compounds as much as the coring site.
More studies are needed to show how representative these cores are for the three
estuaries.
4. Summary and conclusions
Despite some uncertainties caused by the complex history of sediment transport
in coastal areas, radionuclide geochronology allowed a historical reconstruction of
the evolution of trace contaminant inputs into three di€erent coastal areas: Mississippi River Delta, Galveston Bay and Tampa Bay. These areas receive contaminants from a drainage basin, which contains more than 50% of the chemical
and re®nery capacity of the USA. Generally, low concentrations of trace metals
and trace organic contaminants
P were encountered in surface sediments at these
three sites. Exceptions include PAHs in Tampa Bay, where the PAHs are attributed to local sources due to the proximity to shore. Many of the pollutant pro®les
indicate that the present-day situation is continuing to improve from the more
contaminated conditions in the 1950±1970s. With a few exceptions, concentrations
and ¯uxes of most trace metals found in surface sediments at these three sites, when
normalized to Al, are typical for uncontaminated Gulf Coast sediments (Summers
et al., 1996; Wen, Shiller et al., 1999). Even though caution is advised when only
one core from each site is available for interpretation, it is clear that the management strategies to ban the use of organochlorine contaminants have lead to declining concentrations. Exceptions to this trend include trace contaminants such as
DDTs, which were banned in the early 1970s, but are still present in cultivated
soils, riverine deposits, and in Mississippi River Delta sediments, and which appear
to be controlled by a combination of erosion processes of riverine and other sedimentary deposits during ¯oods. Their continued presence many years after their
ban, albeit at low concentrations, indicates that environmental processes will continue to provide these trace contaminants to the marine environment for the foreseeable future.
76
P.H. Santschi et al. / Marine Environmental Research 52 (2001) 51±79
Acknowledgments
We thank the four anonymous reviewers of this manuscript for their helpful
comments. This work was funded by the National Oceanic and Atmospheric
Administration (Grant # 003581).
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