1
Bioaccumulation Factor for Polychlorinated Biphenyls in Fish in
the Houston Ship Channel
Civil and Environmental Engineering Research Experience for Undergraduates
University of Houston
Prepared by:
Sean Carbonaro
Dr. Hanadi Rifai
July 30, 2009
2
3
Abstract
Polychlorinated biphenyl (PCB) concentrations in water, sediment and fish
samples were studied in the Houston Ship Channel and Galveston Bay to determine a
bioaccumulation factor (BAF) for common fish species. The objective of the study was to
test the validity of the use of a simple BAF parameter in water quality regulation. Water
PCB concentration was found using a high-volume sampling technique. Fish were
filleted for tissue analysis. A general decline in PCB levels from 2002-2003 to 2008
across all media was found in the studied area. In the 2008 study, 96% (n=25) of the
stations sampled exceeded the PCB standard for fish tissue and/or water. A strong linear
correlation was found between total water concentration and lipid-normalized tissue
concentration of PCBs for Atlantic Croaker (R2=0.696, p<0.0005), while there was no
correlation for Hardhead Catfish (R2=0.022, p=0.563). In calculating a BAF (in L/kg
lipid) for both species, a lower mean and smaller range was found for Atlantic Croaker
(2.41·106, 1.19·106-4.12·106) than for Hardhead Catfish (7.02·106, 8.50·105-5.82·107) in
the 2008 data set. Similar results were found for Hardhead Catfish in the 2002-2003
study (5.65·106, 5.37·105-3.04·107). No correlation (R2=0.0029, p=0.862) was found
between the BAF of Atlantic Croaker and Hardhead Catfish at common sampling sites in
the 2008 study. A linear deterministic model was not found to be suitable to determine a
BAF that applied to both species.
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Table of Contents
1. INTRODUCTION .............................................................................................................................. 5
2. METHODS ......................................................................................................................................... 9
2.1 STUDY AREA ........................................................................................................................................ 9 2.2 SAMPLE COLLECTION ......................................................................................................................... 10 2.3 ANALYTICAL METHODS ..................................................................................................................... 12 2.4 DATA PROCESSING METHODS ............................................................................................................ 13
3. RESULTS AND DISCUSSION ........................................................................................................ 14
3.1 TEMPORAL VARIATION OF PCB CONCENTRATIONS IN WATER, SEDIMENT AND FISH TISSUE ................ 14 3.2 VARIABILITY OF BAF WITH FISH TISSUE CONCENTRATION AND WATER CONCENTRATION ................. 16
3.2.1 Lipid-normalized fish tissue concentration ................................................................................ 16 3.2.2 Water concentration ................................................................................................................... 18
3.3 CALCULATION OF BAF FOR HARDHEAD CATFISH AND ATLANTIC CROAKER ..................................... 20 3.4 COMPARISON OF SPECIES-SPECIFIC BAF............................................................................................. 24
4. CONCLUSIONS ............................................................................................................................... 27
ACKNOWLEDGEMENTS .................................................................................................................. 27
REFERENCES ..................................................................................................................................... 29
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1. Introduction
Polychlorinated biphenyls (PCBs) are a suite of chemicals that have been used for
a variety of industrial purposes, namely as dielectric fluids in transformers and capacitors,
and in hydraulic systems (Bodin et al., 2008). Large-scale industrial production began in
1929 with the creation of Arochlors, a family of PCB congeners with different levels of
chlorination (Lipnick et al., 2001). In 1976, the ban of PCB manufacturing, processing,
distribution and use began in the United States, followed by Japan, Canada and western
European countries (Lipnick et al., 2001). The prohibition of PCBs occurred as a result of
the discovery of the toxicological effects of PCBs exposure. The effects of PCBs were
found as early as 1937 when occupational exposure caused acute toxic health effects and
researchers reported that “these experiments leave no doubt as to the possibility of
systemic effects from … chlorinated diphenyl” (Drinker et al., 1937). Chronic toxicity
has yet to be proved, but toxicological data from animal studies tend to show that it is
indeed toxic and carcinogenic (Robertson and Hansen, 2001). It is important to note that
each congener has different toxicity, with more chlorinated congeners generally being
more toxic. The consequences of PCB exposure are observed mainly in the thyroid gland,
liver, immune system and reproductive system of humans (Sharma et al., 2009). The
majority of the attention to PCB toxicity occurred as a result of an incident in western
Japan in 1968. Cooking oil was contaminated with PCBs and 1291 patients reported toxic
effects including various somatic complaints, low birth weights, chloracne and
pigmentation (Robertson and Hansen, 2001). However, some additional studies have
suggested the observed effects have occurred as a result of co-contamination with more
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toxic compounds, such as polychlorinated dibenzofurans (PCDFs), which complicates
toxicological analysis (Robertson and Hansen, 2001).
The current standards for fish and water for PCBs in the state of Texas have been
set by the Texas Department of State Health Services (TDSHS) and Texas Commission
on Environmental Quality (TCEQ), respectively. The fish tissue standard is 0.047 mg/kg,
or 47 ng/g, measured on a wet weight basis (TDSHS, 2008a). The surface water quality
standard for areas with sustainable saltwater fisheries is 0.885 ng/L (TCEQ, 2008). The
surface water quality standard for freshwater areas designated or used for public drinking
water supply and recreational fishing is 1.3 ng/L (TCEQ, 2008). These standards are
currently below those set by the United States Environmental Protection Agency
(USEPA).
The ubiquitous nature of PCBs was reported by Jensen in 1966 with a study
indicating significant concentrations of PCBs in the tissue of eagles, herring, and other
environmental species in Swedish waters (Jensen et al., 1969). PCBs are persistent in
nature due to the high physical and chemical stabilities that made them attractive to
industry (Lin et al., 2006). This persistence along with their hydrophobicity and
resistance to degradation lends way to PCBs’ ability to bioaccumulate (Bodin et al.,
2008). The criteria for bioaccumulation is generally when the Bioaccumulation Factor
(BAF) exceeds 5000 L/kg on a wet weight basis, with the BAF being the ratio of
concentration in fish tissue to concentration of the water (Harrad, 2001). The BAF is used
by the USEPA in calculating a Water Quality Target (WQTarget) for a particular body of
water. Toxicological effects of PCBs generally occur from human consumption of
aquatic organisms, however the concentration of PCBs in the tissue of aquatic organisms
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cannot be easily monitored and regulated. The purpose of the BAF is to translate the
desired concentration in aquatic organism tissue into a water concentration, WQTarget, to
assist in the regulation process. The accuracy of the number used for BAF is important as
it relates to permitting and clean-up efforts.
The ability of a chemical to bioaccumulate has been related to several potential
factors. Substances will partition differently between fish tissue and water depending
upon their chemical properties, the most relevant of which is the octanol-water partition
coefficient (Kow). There are 209 different congeners of PCBs, over which there is a Kow
range of 105, with the more chlorinated congeners being more hydrophobic. This range of
chemical properties among PCB congeners themselves leads to the problem of
calculating a singular BAF for such a wide range of chemicals. Another important
consideration for calculating the BAF is the bioavailability of the PCBs. In the freely
dissolved phase, the contaminants are able to move freely across biological membranes
of aquatic organisms, while those in the suspended phase cannot (Sethajintanin and
Anderson, 2006). The study by Sethajintanin and Anderson, 2006 found bioavailability
varies greatly among PCB congeners, with some being consistently more bioavailable
than others.
In the calculation and understanding of the BAF, it has been accepted that lipid
normalization of the chemical concentration is essential (USEPA, 1995). United States
Environmental Protection Agency (USEPA) guidelines suggest that lipid content is
determined from the same sample that is analyzed for chemical residue (USEPA, 1995).
Bioconcentration tends to be proportional to the lipid content of the organism, and the use
of lipid normalization allows for the calculation of a BCF or BAF independent of the
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lipid content of the organism (Harrad, 2001). Based on the value of normalizing fish
tissue concentration with lipid content and using bioavailable, or dissolved phase, water
concentration, the equation that will be preferred in the calculation of BAF in this paper
will be as follows:
BAF =
Tissue PCB concentrationLipid content
Dissolved phase PCB concentration
In this study, it will be important to consider the vast differences between
congeners. Some congeners can potentially be neglected as their concentrations are
consistently very low in comparison to others due to the lack of industrial production and
other factors. Other congeners can also possible be neglected due to their low octanol-
water coefficient (Kow) since they likely do not partition into fish tissue as well as other
more chlorinated congeners. Several studies have been done to attempt to use log Kow as
a predictor of BAF, typically in the form of log BAF. It has been suggested that if
equilibrium partitioning between the water and fish tissue were to occur, log Kow would
be equal to log BAF (Borga et al., 2005). However Borga et al. 2005 found that the BAF
was at least 10 times higher than that predicted by Kow. A special case of
bioaccumulation is the bioconcentration factor (BCF), which like BAF is the ratio of
concentration in fish tissue to concentration of the water, but considers only abiotic
exposure and must be found in laboratory experiments. The relationship between log
BCF and log Kow has been found to be stronger and more reflective of the equilibrium
partitioning assumption than the relationship between log BAF and log Kow (Arnot and
Gobas, 2006). However, the BCF is not a realistic predictor of the true concentration in
aquatic organisms because it does not account for factors including diet and sediment
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interactions. The BAF has been found to be up to 10000 times greater than the BCF for
high Kow chemicals (Harrad, 2001). There has also been a parabolic correlation found
between log BAF and log Kow. A species-specific parabolic model more sufficiently
explained the relationship of log BAF to log Kow than a linear model in 4 out of 6 species
in a heavy polluted reservoir (Wu et al., 2008). No consensus has been found concerning
the relationship between BAF and Kow as many other factors contribute to the BAF
including the organism’s size, diet, behavior, and several other potential factors.
2. Methods
2.1 Study Area
The Houston Ship Channel (HSC), Upper and Lower Galveston Bay, and Trinity
Bay have received several advisories about fish consumption in recent years (TDSHS,
2008a) (TDSHS, 2008b). TDSHS has advised against the consumption of more than one
eight-ounce meal per month of all catfish species and speckled sea trout due to the
presence of polychlorinated dibenzo-p-dioxins and polychlorinated dibenzofurans
(PCDDs/PCDFs or ‘dioxins’) and polychlorinated biphenyls (PCBs). Galveston Bay is
the seventh largest estuary (600 square miles or 384,000 acres, 232 miles of shortline) in
the United States. Nearly half of all United States petrochemical production occurs in the
greater Houston area. A map summarizing the sample locations for the 2008 study is
included in Figure 1.
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Figure 1. Map of Houston Ship Channel, Galveston Bay, contributing bayous and
sample sites for 2008 study.
2.2 Sample Collection
Sampling was performed at the selected sites for two periods of time, from
summer 2002 to spring 2003, then again during 2008. In the summer 2002 to spring 2003
study, a sum total of 53, 98 and 84 samples were collected for water, sediment and fish,
respectively. In the 2008 study, a sum total of 45, 95 and 50 samples were collected for
water, sediment and fish, respectively. The selected sites were spaced throughout the
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HSC, adjoining bays, and Houston area bayous. At each site, samples were taken from
water, sediment and fish tissue to measure various parameters.
Water column samples were taken using a high-volume sampling technique. The
sampling unit that was used is designed to collect both particle-bound and dissolved
phase PCBs. This is achieved at low detection limits by the use of 1 µm Glass Fiber
Filters (GFFs), which collect the particle-bound PCBs, and XAD-2 resins, which collect
the dissolved phase PCBs. The sampling unit was operated continuously from a boat until
the desired volume of water was filtered. This desired volume of water can range from
100 L to 1000 L to detect ultra trace (well below 0.1 ppb, or 0.1 µg/L) concentrations. A
representative sample of the entire water column was ensured by changing the depth of
the sampling unit inlet every 30 minutes. Some in-stream sampling was performed by
operation of the sampling unit from the shore of the stream. These sampling points taken
from land are generally assumed to have no concentration deviation with depth. After the
desired volume is achieved, the GFFs and XAD-2 resins are packaged separately for
analysis. The Total Suspended Solids (TSS) samples were collected from a depth of 1 ft
using pre-cleaned glass bottles. The Dissolved Organic Carbon (DOC) and Total Organic
Carbon (TOC) samples were also collected from a depth of 1 ft using pre-cleaned
borosilicate glass bottles.
Sediment samples were taken using a stainless steel Ponar, Eckman, or Peterson
dredge. Prior to sample collection, the dredge, stainless steel spoon, and stainless steel
bucket were rinsed with de-ionized (DI) water, then ambient water. A minimum of three
sediment grab samples were collected from only the top 5 cm of sediment. A
representative sample of the water cross section was ensured by collecting samples in
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equal amounts from river left, right and center. The samples were then homogenized,
deposited into a labeled, pre-cleaned amber glass jar with a Teflon seal. Separate samples
were prepared to measure for Total Organic Carbon (TOC), solids content, and Total
Petroleum Hydrocarbons (TPH). The dredge was not used in the same area for all grab
samples, but rather various locations around the sampling boat or across the stream.
Fish tissue was collected from several species, with preference in the following
order; Trout/Croakers: Speckled Seatrout (Cynoscion nebulosus), Sand Seatrout
(Cynoscion arenarious), and Atlantic Croaker (Micropogonias undulates); Catfish:
Hardhead Catfish (Arius felis), Blue Catfish (Ictalurus furcatus), and Channel Catfish
(Ictalurus punctatus). Speckled Seatrout, Sand Seatrout and Atlantic Croaker were not
sampled in the 2002-2003 study. A target length of 30 cm, or 12 in, and larger was used
for collection. The samples were all individually measured for length and weight. A
composite of fish tissue from a minimum of three fish from the same species was
obtained by filleting fish samples using a stainless steel knife, packing in aluminum foil
and plastic bags, then freezing the composite. The composite was then shipped to a
commercial lab for analysis of PCB concentration and lipid content.
2.3 Analytical Methods
PCB congeners in water, sediment, and fish were quantified by high-resolution
gas chromatography/high resolution mass spectrometry (HRGC/HRMS) using USEPA
method 1668A, which quantifies all 209 congeners (USEPA, 1999). Lipid content was
determined by a commercial laboratory’s in-house method. Sediment TOC was
determined via USEPA Method 415.2. All of these analyses were completed by a
commercial laboratory.
a The 43 congeners, along with co-eluters, are PCB-8, 18/30, 20/28, 37, 44/47/65, 49/69, 52, 60, 61/70/74/76, 66, 77, 81, 82, 83/99, 86/87/97/109/119/125, 90/101/113, 105, 114, 118, 123, 126, 128/166, 129/138/163, 135/151, 153/168, 156/157, 158, 167, 169, 170, 177, 179, 180/193, 183, 187, 189 b The NOAA 18 congeners, along with co-eluters, are PCB-8, 18/30, 20/28, 44/47/65, 52, 66, 77, 90/101/113, 105, 118, 126, 128/166, 129/138/163, 153, 169, 170, 180/193, 187
2.4 Data Processing Methods
Data analysis was performed using databases containing the results from two
sampling sets from 2002 to 2003 and 2008. Two congener-specific grouping methods
were used in analyzing the data in addition to the total concentration method. One of the
methods used is a grouping of 43 specific congenersa (McFarland and Clark, 1989). The
other used was the NOAA 18b (NOAA, 1989). These congener groupings were found by
their respective sources to be of particular concern due to their ability to occur frequently
in environmental samples, accumulate in animal tissue, and display toxic effects.
Other considerations were taken in the use of the database. Duplicate samples
were taken at several of the sites used for this study. For those samples, the average of the
results were taken and used as a singular data point for that particular station. The
concentrations of the individual PCB congeners were reported separately and summed to
produce the results for all congeners, McFarland and Clarke 43a, and NOAA 18b. On
several samples the results of some congeners fell below the Method Detection Limit
(MDL). For these individual congener concentrations, the result was assumed to be one
half of the MDL. This method did not significantly change the results for the congener
concentration sums and was found to be the best option over assuming the non-detectable
samples to be zero or at the MDL.
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3. Results and Discussion
The fish tissue and surface water quality standards for PCBs were regularly
exceeded in the study area. In the 2008 study, 23 out of 25 stations sampled for water
concentration exceeded the TCEQ surface water quality standard of 0.885 ng/L. The
average total water concentration, using all congeners, was 2.22 ng/L. In the 2008 study,
22 out of the 25 stations sampled for catfish and/or sportfish exceeded the standard of 47
ng/g wet weight. The average fish tissue concentration for all fish species sampled, using
all congeners, was 174 ng/g wet weight. In the 2002 study, 27 out of the 32 stations
sampled for water concentration exceeded the water quality standard of 0.885 ng/L. The
average total water concentration, using all congeners, was 1.96 ng/L. In the 2002 study,
30 out of the 32 stations sampled for catfish exceeded the standard of 47 ng/g wet weight.
The average tissue concentration for all fish species sampled in the 2002 study, using all
congeners, was 129 ng/g wet weight.
3.1 Temporal variation of PCB concentrations in water, sediment and fish tissue
The studies in 2002-2003 and 2008 had 18 common stations that were sampled
for water, sediment and catfish tissue concentrations. Atlantic Croaker and Speckled
Seatrout were not sampled in the 2002-2003 study. A decline in the concentration of
PCBs across all media over time can be seen in Table 1. However, the concentrations of
the congener groupings in all water phases considered to be of most importance show less
of a reduction than when all congeners are included. This trend is not seen in sediment or
fish tissue. This could indicate that either the specific congeners identified by McFarland
and Clarke and NOAA are indeed more persistent, but only in the water phase, than other
15
congeners or that the recent inputs of PCBs into the HSC have a congener profile more
representative of the specific congener groupings than the entire congener range. On a
lipid basis, the PCB concentration within the catfish tissue was relatively unchanged.
However, on a wet weight basis, the concentration was significantly reduced. This could
be indicative of a difference in selection criteria between the two studies or a possible
trend in the lipid content of the fillets of catfish in the HSC.
Table 1. Comparison of 2002-2003 and 2008 Data for various media
2002-2003 2008
Water (Total) (ng/L)
All congeners 2.58 1.98 23.1
M & C 43 1.10 0.91 17.3
NOAA 18 0.77 0.68 12.0
Water (Dissolved) (ng/L)
All congeners 1.68 1.55 8.0
M & C 43 0.74 0.69 6.4
NOAA 18 0.52 0.53 -1.2
Water (Suspended) (ng/L)
All congeners 0.89 0.43 51.6
M & C 43 0.36 0.22 39.5
NOAA 18 0.25 0.15 39.9
Sediment (ng/g)
All congeners 228 26 88.8
M & C 43 117 14 88.3
NOAA 18 78 9 87.8
Catfish tissue (ng/g ww)
All congeners 154 112 27.6
M & C 43 113 84 24.9
NOAA 18 86 63 26.7
Catfish tissue (ng/g lipid)
All congeners 9033 9023 0.1
M & C 43 6604 6871 -4
NOAA 18 5061 5135 -1.5
Year
Percent reduction (%)
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3.2 Variability of BAF with fish tissue concentration and water concentration
The lipid-normalized BAF was plotted against the lipid-normalized fish tissue
PCB concentration, total PCB water concentration, and dissolved phase PCB water
concentration. The regressions were performed on individual species of fish from the
study, the largest two groups being Atlantic Croaker and Hardhead Catfish.
3.2.1 Lipid-normalized fish tissue concentration
The regressions of lipid-normalized fish tissue concentration against lipid-
normalized BAF for Hardhead Catfish and Atlantic Croaker are shown in Figure 2. A
significant linear correlation (R2 = 0.8027, p<0.0005, 2 outliers removed) was found
between the lipid-normalized Hardhead Catfish tissue concentration and lipid-normalized
BAF for Hardhead Catfish. Another regression of lipid-normalized tissue concentration
against lipid-normalized BAF was also performed using the 2002 Hardhead Catfish data.
A significant linear correlation (R2=0.522, p<0.0005, 31 samples, 3 outliers removed)
was found using this Hardhead Catfish data as well. The range of water and fish tissue
concentration was comparable, as can be seen in Table 1. The relationship found from
this plot indicates that the BAF may not be independent of fish tissue concentration as
has been previously suggested. It is possible that in the particular water and fish tissue
concentration range in this study the PCB intake rate of the Hardhead Catfish far exceeds
that of a potential maximum depuration rate. However, for the Atlantic Croaker, this
relationship does not seem to be very strong (R2=0.3908, p<0.05). The relationship
becomes weaker when considering the M & C 43 and NOAA 18 congener groupings, and
BAFs calculated from dissolved water concentrations. This finding suggests that the
17
Atlantic Croaker is more capable of equilibrating its intake and depuration rates at the
water and tissue concentrations found in this study than the Hardhead Catfish.
Figure 2. Relationship of lipid-normalized BAF to tissue concentration for
Hardhead Catfish (blue) and Atlantic Croaker (red). Lipid-normalized BAF
y = 210.37x + 2E+06
R² = 0.3908
p=0.0168
1.E+06
2.E+06
3.E+06
4.E+06
5.E+06
0 2 4 6 8 10 12
Cro
ak
er
BA
F (L
/kg
lip
id)
Tissue concentration (µg/g lipid)
y = 532.99x + 44127
R² = 0.8027
p=7.54E-09
5.00E+05
5.50E+06
1.05E+07
1.55E+07
0 5 10 15 20 25 30
Ca
tfis
h B
AF
(L/k
g li
pid
)
Tissue concentration (µg/g lipid)
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calculated using total water concentration and all PCB congeners. (2 outliers
removed for Hardhead Catfish regression)
3.2.2 Water concentration
The plot of water concentration against lipid-normalized BAF for Hardhead
Catfish and Atlantic Croaker is shown in Figure 3. No significant relationship was found
between the lipid-normalized BAF and dissolved water concentration for Hardhead
Catfish (R2=0.0963-0.1405, p>0.05) or Atlantic Croaker (R2=0.0064-0.0143, p>0.5). The
Hardhead Catfish data has a less random distribution than the Atlantic Croaker data, as
shown by the p-values and R2 values of the regressions. The results of the regressions
show that the BAF of the Hardhead Catfish seems to be more correlated to the water
concentration than the BAF of the Atlantic Croaker, which shows almost absolutely no
correlation with the water concentration, both total and dissolved phase. The calculation
of a BAF does not necessarily require equilibrium conditions, but the consistency of the
results would be improved if equilibrium had been reached. The fish specimens used in
this study do not remain in the spot they happened to be caught in, but are instead a
composite of the nearby waters. The specimens are exposed to a small, but still variable,
range of PCB concentrations in the HSC and its tributaries. The fact that the Atlantic
Croaker BAF is more random with respect to water concentration may be an indication of
how quickly it reaches equilibrium compared to the Hardhead Catfish. The intake and
depuration rates of the Atlantic Croaker may be faster, leading to a more accurate
representation of current PCB concentrations, rather than lifetime exposure another
species with slower intake and depuration rates may display.
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Figure 3. Relationship of lipid-normalized BAF with water concentration for Hardhead Catfish (blue) and Atlantic Croaker
(red). Lipid-normalized BAF calculated using respective water phase and all PCB congeners.
y = -1E+06x + 7E+06
R² = 0.0963
p=0.092
1.0E+06
6.0E+06
1.1E+07
1.6E+07
0 1 2 3 4
Ca
tfis
h B
AF
(L/k
g)
Total water concentration (ng/L)
y = -3E+06x + 1E+07
R² = 0.1405
p=0.061
5.00E+05
5.50E+06
1.05E+07
1.55E+07
2.05E+07
0 1 2 3
Ca
tfis
h B
AF
(L/k
g)
Dissolved water concentration (ng/L)
y = -165542x + 3E+06
R² = 0.0064
p=0.786
1.E+06
3.E+06
5.E+06
7.E+06
0 1 2 3C
roa
ke
r B
AF
(L/k
g)
Dissolved water concentration (ng/L)
y = 127837x + 2E+06
R² = 0.0143
p=0.683
1.E+06
2.E+06
3.E+06
4.E+06
0 1 2 3 4
Cro
ak
er
BA
F (L
/kg
)
Total water concentration (ng/L)
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3.3 Calculation of BAF for Hardhead Catfish and Atlantic Croaker
The bioaccumulation factor (BAF) can be calculated by two possible
deterministic methods, performing a linear regression on a plot of fish tissue
concentration against water concentration or dividing fish tissue concentration by water
concentration and analyzing the distribution. This study used both methods, the
regression method is shown in Table 2, a box and whisker plot of division method is
shown in Figure 4. The regression method showed that the use of congener groupings
McFarland and Clark 43 and NOAA 18 had no particular advantage over the use of all
congeners in the calculation of the BAF. The BAF for the congener groupings calculated
by the regression method slightly increased over that for all congeners. This would
suggest that the congeners selected for these groupings are indeed more bioaccumulative
than the suite of all 209 PCB congeners. The relationship of Atlantic Croaker lipid-
normalized tissue concentration and both total and dissolved phase concentration had a
strong linear correlation by the regression (R2=0.656-0.707, p<0.0005). The Hardhead
Catfish data for neither 2008 nor 2002-2003 data indicated the same relationship (p>0.05
for all regressions). The suspended phase water concentration surprisingly had a stronger
correlation with lipid-normalized fish tissue concentration than both total and dissolved
phase water concentration for Hardhead Catfish. The reason for this may be explained by
the typical diets of the two species. The mature Atlantic Croaker generally eats crabs,
shrimp and other fish like eels and minnows while the Hardhead Catfish has been found
to eat virtually anything, including algae, zooplankton, smaller Hardhead Catfish, mud,
and sand (Horst and Lane, 2006). This indicates that the Hardhead Catfish may actually
be more likely to consume and absorb PCBs that have adsorbed onto suspended particles.
21
The BAF for the Atlantic Croaker was approximately double that of the Hardhead Catfish
for most regressions. However, by the box plot distribution method, the BAF of the
Hardhead Catfish was approximately triple that of the Atlantic Croaker. This is likely due
to the poor fit of the Hardhead Catfish data in the regression.
Several physical parameters, length, weight, and lipid content, of the species were
recorded during processing and analysis. These results are summarized in Table 3. On
average, the Atlantic Croaker had over four times the lipid content of the Hardhead
Catfish. On average, the Hardhead catfish was larger than the Atlantic Croaker, 40% by
length, 130% by weight.
The Hardhead Catfish was found to have a better correlation (R2=0.42, p<0.0005,
all congeners) between total water concentration and tissue concentration when tissue
concentration was not lipid-normalized. The Atlantic Croaker was found to have a similar
correlation (R2=0.43, p<0.005, all congeners). Considering the results of Table 2, we can
see that the goodness of fit for Hardhead Catfish worsened after lipid-normalization while
that for the Atlantic Croaker improved.
A t-test on the lipid-normalized BAF values for Hardhead Catfish in the 2002-
2003 and 2008 study reported a value of 0.472, showing no significant statistical
difference between the mean of the two data sets. Given the number of samples (31 in
2002-2003, 19 in 2008), these results, along with visual inspection of Figure 4, show that
the bioaccumulation behavior of the Hardhead Catfish was mostly unchanged from 2002-
2003 to 2008. Histograms of the two data sets also appear to be very similar, showing
little significant change in the BAF distribution from 2002-2003 to 2008.
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The values of the BAF found in this study agreed well with those found in related
literature. The range of lipid-normalized BAF calculated from dissolved water
concentrations was 8.95·105 to 5.82·107 and 1.33·106 to 6.09·106 for Hardhead Catfish
and Atlantic Croaker, respectively. On a logarithmic scale, this range is 5.95 to 7.76 and
6.12 to 6.78 for Hardhead Catfish and Atlantic Croaker, respectively. The EPA has
reported a log BAF, calculated on the same basis, range from 5.52 to 9.26 for Great
Lakes Trout and salmonids (USEPA, 1995). Another study on Great Lakes Trout
reported a log BAF range from 5.5 to 8.5, calculated on a dissolved water and wet weight
basis (Streets et al., 2006). A study on a Chinese e-waste site with elevated PCB water
concentration (mean concentration of 204 ng/L) reported a log BAF range of 1.2 to 8.4
(Wu et al., 2008). The Wu et al. 2008 study calculated BAF for several species including
carp, snails and water snakes and by individual PCB congeners. All of the BAFs
calculated for this study fall within the range of those found in the literature that
calculated BAFs for various species in various bodies of water.
Another study that created a small, artificial ecosystem contaminated with PCBs
within a glass and acrylic board cube (5.4 m3) found similar results of different BAFs for
different species (Lin et al., 2006). The Lin et al. 2006 study found that the five species
(three specimens of each) accumulated differing amounts of PCBs through the run of the
experiment. While the study performed by Lin et al. likely does not well represent a
natural environment, it does show that when different species are exposed to the same
conditions, they bioaccumulate different amounts of PCBs.
23
Table 2. Lipid-normalized BAF for Hardhead Catfish (2002-2003 and 2008 studies)
and Atlantic Croaker (only 2008 study) calculated by regression method with
corresponding goodness of fit and p-value.
Table 3. Physical parameters of fish samples from 2008 study.
Water Species R² p-value BAF (L/kg lipid) R² p-value BAF (L/kg lipid)
Total (All cong) Catfish 0.022 0.563 1.28E+06 0.102 0.17 1.02E+06
Total (M & C 43) Catfish 0.046 0.407 2.78E+06 0.114 0.068 2.38E+06
Total (NOAA 18) Catfish 0.025 0.542 2.32E+06 0.106 0.073 2.53E+06
Dissolved (All cong) Catfish 0.007 0.752 1.28E+06 0.042 0.329 1.03E+06
Dissolved (M & C 43) Catfish 0.034 0.475 3.48E+06 0.069 0.177 2.39E+06
Dissolved (NOAA 18) Catfish 0.013 0.661 2.53E+06 0.065 0.184 2.53E+06
Suspended (All cong) Catfish 0.064 0.325 6.09E+06 0.145 <0.05 2.48E+06
Suspended (M & C 43) Catfish 0.058 0.349 9.92E+06 0.233 <0.05 1.23E+07
Suspended (NOAA 18) Catfish 0.051 0.383 1.01E+07 0.220 <0.05 1.37E+07
Total (All cong) Croaker 0.696 <0.0005 2.65E+06
Total (M & C 43) Croaker 0.701 <0.0005 3.46E+06
Total (NOAA 18) Croaker 0.696 <0.0005 3.38E+06
Dissolved (All cong) Croaker 0.656 <0.0005 3.45E+06
Dissolved (M & C 43) Croaker 0.707 <0.0005 4.90E+06
Dissolved (NOAA 18) Croaker 0.669 <0.0005 4.45E+06
Suspended (All cong) Croaker 0.394 0.016 5.38E+06
Suspended (M & C 43) Croaker 0.454 0.008 7.81E+06
Suspended (NOAA 18) Croaker 0.451 0.008 8.09E+06
2002-20032008
24
Figure 4. Box plot of lipid-normalized BAF by species and time on a logarithmic
scale. BAF calculated by indicated water phase. All PCB congeners used. Outliers
are marked by diamonds. Extreme outliers marked by circles.
3.4 Comparison of species-specific BAF
The relationship of Hardhead Catfish BAF to Atlantic Croaker BAF is shown in
Figure 5. The lipid-normalized BAF was used for this regression, however an equivalent
representation would be the lipid-normalized tissue concentration of each species, since
the BAF was calculated simply by dividing the lipid-normalized tissue concentration by
the respective water phase concentration. The 2008 study had 13 common stations where
both species were collected and analyzed. There was definitely no correlation between
Hardhead Catfish BAF and Atlantic Croaker BAF. These common stations, as shown in
Figure 1, were generally part of Trinity Bay or a nearby section of the HSC.
The reason for the difference in the BAF in the two species likely occurs due to
the variation in habitat and life cycle. Atlantic Croaker is usually found in estuaries or
offshore water while the Hardhead Catfish can be found both in nearshore waters and
5.E+05
5.E+06
5.E+07
2002 Catfish
Total
2008 Catfish
Total
2008 Croaker
Total
2002 Catfish
Dissolved
2008 Catfish
Dissolved
2008 Croaker
Dissolved
BA
F (L
/kg
lip
id)
25
occasionally in freshwater, as seen by our finding of Hardhead Catfish in non-tidal
bayous (Horst and Lane, 2006). The concentrations found at stations in Trinity Bay were
generally lower than those found further up the HSC, meaning the Atlantic Croaker may
not be exposed to as high of PCB concentrations. The main reason for the deviation likely
occurs as a result of the life cycle of the species. The Atlantic Croaker generally does not
survive past 4 or 5 years, while the average life span for the Hardhead Catfish is 23 years
(Horst and Lane, 2006). For this reason, the Hardhead Catfish is likely more
representative of historical concentrations of PCBs than the Atlantic Croaker, which
reflects better on the current concentrations.
26
Figure 5. Relationship of lipid-normalized BAF calculated for Hardhead Catfish
and Atlantic Croaker from 13 common sites in 2008 study. Top plot used BAF
calculated from total water concentration, bottom plot used BAF calculated from
dissolved water concentration. All PCB congeners used.
y = 0.0056x + 2E+06
R² = 0.0029
p=0.862
1.E+06
2.E+06
3.E+06
4.E+06
1.0E+06 1.1E+07 2.1E+07 3.1E+07
Cro
ak
er
BA
F (L
/kg
lip
id)
Catfish BAF (L/kg lipid)
y = 0.0266x + 3E+06
R² = 0.0835
p=0.338
1.E+06
3.E+06
5.E+06
1.0E+06 2.1E+07 4.1E+07 6.1E+07
Cro
ak
er
BA
F (L
/kg
lip
id)
Catfish BAF (L/kg lipid)
27
4. Conclusions
Regressions of water and fish tissue data found that congener groupings
established by NOAA and McFarland and Clarke may be more bioaccumulative in
comparison to the suite of 209 PCB congeners. The bioaccumulation factor (BAF) for
PCBs was found to behave differently for the Hardhead Catfish and Atlantic Croaker.
The BAF distribution for Hardhead Catfish did not significantly change from 2002-2003
to 2008. The lack of correlation between the BAF or lipid-normalized tissue
concentration for each species at common sites suggests that the mechanism of
bioaccumulation differs by species. Due to differences between species, including diet,
intake, depuration and metabolism rates, length, weight and lipid content, a simple linear
deterministic model cannot predict a BAF for all fish species. The BAF calculated for a
species may also vary by region, such as a specific river or bay, due to potential changes
in the species’ overall behavior. On this basis, it would be most appropriate to determine
a region and species-specific BAF for bioaccumulative chemicals, including PCBs.
Acknowledgements
The research study described herein was sponsored by the National Science
Foundation under the Award No. EEC-0649163. The opinions expressed in this study are
those of the authors and do not necessarily reflect the views of the sponsor.
I am glad to have been under the guidance of Dr. Hanadi Rifai and Nathan
Howell. It has also been a pleasure to work with this group both in the office and the
field: Yaa Amoah, Anu Desai, Matt Feaga, Lisa Grecho, Emil Helfer, Becky Jimenez,
28
Divagar Lakshmanan, Jenni McFarland, Maria Modelska, Norma Moreno, Stephen Ray,
Scott Rauschhuber, Zack Van Brunt and Sharon Wells.
29
References
Arnot, J.A., Gobas, F., 2006. A review of bioconcentration factor (BCF) and bioaccumulation factor (BAF) assessments for organic chemicals in aquatic organisms. Environ. Rev. 14, 257-297. Bodin, N., Le Loc'h, F., Caisey, X., Le Guellec, A.M., Abarnou, A., Loizeau, V., Latrouite, D., 2008. Congener-specific accumulation and trophic transfer of polychlorinated biphenyls in spider crab food webs revealed by stable isotope analysis. Environmental Pollution 151, 252-261. Borga, K., Fisk, A.T., Hargrave, B., Hoekstra, P.F., Swackhamer, D., Muir, D.C.G., 2005. Bioaccumulation factors for PCBs revisited. Environmental Science and Technology 39, 4523. Drinker, C.K., Warren, M.F., Bennett, G.A., 1937. The Problem of Possible Systemic Effects from Certain Chlorinated Hydrocarbons. The Journal of Industrial Hygiene and Toxicology 19. Harrad, S., 2001. Persistent Organic Pollutants: Environmental Behavior and Pathways of Human Exposure. Kluwer Academic Publishers, Norwell, Massachusetts. Horst, J., Lane, M., 2006. Angler's Guide to Fishes of the Gulf of Mexico. Pelican Publishing Company, Gretna. Jensen, S., Johnels, A.G., Olsson, M., Otterlin.G, 1969. DDT AND PCB IN MARINE ANIMALS FROM SWEDISH WATERS. Nature 224, 247-&. Lin, Y.-J., Liu, H.-C., Hseu, Z.-Y., Wu, W.-J., 2006. Study of transportation and distribution of PCBs using an ecologically simulated growth chamber. Chemosphere 64, 565-573. Lipnick, R.L., Hermens, J.L.M., Jones, K.C., Muir, D.C.G., 2001. Persistent, Bioaccumulative and Toxic Chemicals I: Fate and Exposure. American Chemical Society. McFarland, V.A., Clark, J.U., 1989. Environmental Occurrence, Abundance, and Potential Toxicity of Polychlorinated Biphenyl Congeners: Considerations for a Congener-Specific Analysis. Environmental Health Perspectives 81, 225-239. NOAA, 1989. A Summary of Data on Tissue Contamination from the First Three Years (1986-1988) of the Mussel Watch Project. National Oceanic and Atmospheric Administration, Rockville, Maryland. Robertson, L.W., Hansen, L.G., 2001. PCBs: Recent Advances in Environmental Toxicology and Health Effects. The University Press of Kentucky, Lexington, Kentucky. Sethajintanin, D., Anderson, K.A., 2006. Temporal bioavailability of organochlorine pesticides and PCBs. Environmental Science & Technology 40, 3689-3695. Sharma, C.M., Rosseland, B.O., Almvik, M., Eklo, O.M., 2009. Bioaccumulation of organochlorine pollutants in the fish community in Lake Årungen, Norway. Environmental Pollution 157, 2452-2458. Streets, S.S., Henderson, S.A., Stoner, A.D., Carlson, D.L., Simcik, M.F., Swackhamer, D.L., 2006. Partitioning and Bioaccumulation of PBDEs and PCBs in Lake Michigan†Environmental Science & Technology 40, 7263-7269. TCEQ, 2008. Revisions to §307 - Texas Surface Water Quality Standards. Texas Commission on Environmental Quality, Austin, p. 144.
30
TDSHS, 2008a. Characterization of Potential Adverse Health Effects Associated with Consuming Fish or Blue Crab from Trinity Bay and Upper Galveston Bay: Chambers, Galveston, and Harris Counties, Texas. Department of State Health Services; Division of Regulatory Services; Policy, Standards and Quality Assurance Unit; Seafood and Aquatic Life Group, Austin, TX, p. 55. TDSHS, 2008b. Texas Department Of State Health Services Fish and Shellfish Consumption Advisory ADV-35. In: Services, T.D.o.S.H. (Ed.), Austin, TX. USEPA, 1995. Great Lakes Water Quality Initiative Technical Support Document for the Procedure to Determine Bioaccumulation Factors. United States Environmental Protection Agency. USEPA, 1999. Method 1668, Revision A: Chlorinated Biphenyl Congeners in Water, Soil, Sediment, and Tissue by HRGC/HRMS. USEPA National Office, Washington, D.C., p. 133. Wu, J.-P., Luo, X.-J., Zhang, Y., Luo, Y., Chen, S.-J., Mai, B.-X., Yang, Z.-Y., 2008. Bioaccumulation of polybrominated diphenyl ethers (PBDEs) and polychlorinated biphenyls (PCBs) in wild aquatic species from an electronic waste (e-waste) recycling site in South China. Environment International 34, 1109-1113.