UNITED STATES ENVIRONMENTAL PROTECTION AGENCY

WASHINGTON, DC 20460

			OFFICE OF  PREVENTION, PESTICIDES,  AND TOXIC SUBSTANCES

 									September 11, 2008

	MEMORANDUM

Subject: 	Environmental Fate and Transport Assessment of  Creosote for
the Reregistration Eligibility Decision (RED) Process.

From:	A. Najm Shamim, PhD, Chemist

	Regulatory Management Branch II

	Antimicrobials Division (7510P)

To:	Jackie McFarlane, CRM for Creosote RED

	Regulatory Management Branch I

	Antimicrobials Division (7510P)

			And

	 Timothy McMahon, PhD, Senior Toxicologist 

and  Risk Assessor  for Creosote RED

	Risk Assessment and Science Support Branch

	Antimicrobials Division (7510P)

Thru:	Mark Hartman, Chief

	Regulatory Management Branch II

	Antimicrobials Division (7510P)

DP Barcode :  		D246550

CREOSOTE - ENVIRONMENTAL FATE AND TRANSPORT RISK ASSESSMENT

I.  Executive Summary:

Creosote is an oil-based wood preservative which is primarily used for
preserving wood used in railroad ties and utility poles.  Coal tar
creosote or the P1/P13 and P2 fractions of  coal tar creosote are
obtained from the carbonization of bituminous coal and are mixtures of
organic molecules (about 200-250  identifiable substances) consisting of
simple polyaromatic hydrocarbons (PAH), multi-aromatic fused rings,
cyclic nitrogen-containing heteronuclear compounds, and phenolic
substances.  It is estimated that about 85 percent of the creosote
mixture consists of PAHs and the remainder consists of cyclic
heteronuclear nitrogen and oxygen containing molecules.  Since the PAHs
constitute the majority of the percent mass of the P1/P13 and P2
fractions of creosote, they weigh heavily in the environmental fate
assessment.  Therefore, the Agency has carried out an open literature
search for environmental fate  and transport studies on the creosote
mixture and has decided to make the assessment based on the PAH
constituents only.

Coal tar creosote has been used as a wood preservative pesticide for
over 125 years.  According to the American Wood Preservers’
Association (AWPA,1997) statistics, an estimated 728 million cubic feet
of wood was treated with preservatives, and creosote and its mixtures
represent 13.3 percent of the treatments.  Nearly 100 percent of
railroad crossties, switch ties and bridge timbers are pressure-treated
with creosote and 14.6 percent of utility poles are treated with
creosote (Webb, N.D.).  

Process wastewater, dumpsite leachate, storage tank leaks, and spills
are the major creosote sources to the environment (Merril and Wade,
1985). In addition, leachates from pressure treated wood can migrate out
into soils, and water.  The environmental fate and transport of creosote
as addressed here focuses primarily on the likely exposure of PAHs into
three environmental compartments: 1) leaching of creosote mixture into
surface and ground waters from the railroad ties and utility poles; 2)
migration into soil/sediments from the railroad ties and utility poles,
and 3) bioaccumulation into the aqueous and benthic organisms. 

The major uses of creosote since 1988 have been railroad ties,
crossbars, decks on marinas and utility poles.  Polyaromatic
hydrocarbons (PAHs) constitute the highest percent 

(85%) of coal tar creosote while the phenolic substances are about 10
percent, and N– and S- containing substances represent the remainder
of the mixture.  Most of the PAHs are non-volatile, therefore; creosote
normally does not contaminate the air.  The major route of exposure from
creosote is through water and soil, and from these environmental
compartments into the aquatic and benthic organisms (bioaccumulation). 

Abiotic Degradation

The PAHs are fused aromatic polycyclic rings which have no hydrolyzable
hydrogens and the solubility of these compounds are very low in water. 
Environmentally, hydrolysis does not appear to be an important pathway
for dissipation of the composite mixture of PAHs in water; however, some
molecules like benzo[k]fluoranthene and benz[a]pyrene could persist in
water.

Very few studies are reported in the open literature on field
volatilities for PAHs present in creosote.  Gevao et al. (1998) showed
that acenaphthene, fluorene, phenanthrene, anthracene, and fluoranthene
volatilized at a faster rate at 30(C than at 4(C.  The study also showed
that 85 percent of these components remained in the wood after seven
weeks.  The rate of volatilization was slow for acenaphthene (half-life,
one year) and fluoranthene (half-life, one year).  In most cases, the
initial rates of dissipation are caused by partitioning between the wood
and air and biodegradation in the presence of microbial populations. 
Therefore, exposure to air does not appear to be an important factor in
fate assessment for most PAHs.

Photooxidation is a common phenomenon, and therefore an important
degradation pathway for the creosote PAHs.  The photolytic half-lives of
the PAHs in aqueous medium are dependent on the season, geographical
location, surface water measurements, and complexities of the parent
molecules (two fused rings vs. five fused rings).  In most cases,
half-lives under the conditions mentioned do not appear very long. 
Because of this, and the fact that most of the PAHs are not readily
soluble (except for a few low molecular weight ones), the PAHs may not
be a problem in surface and groundwater runoffs.  However, it should be
noted that the photooxidized products of PAHs are stable; therefore, may
persist in air/water and soils and become an environmental concern as
these photooxidized products are also bioaccumulative.

Mobility

Once introduced to an aquatic environment, creosote components are
subjected to several fractionation processes.  Many PAHs adsorb to
sediments and may persist for long periods of time.  Creosote
contaminated sediments usually contain relatively higher levels of
hydrophobic PAHs than whole creosote (Bieri et al., 1986).  Eventually,
sediment adsorbed PAHs may dissolve or become re-suspended in the water
by tides, storms, bioturbation, shipping, or dredging.  As a result,
local biota may be exposed to low level PAHs over the long term (Fowler
et al., 1993).  Therefore, the adsorption/desorption processes in water
involved with creosote-derived PAHs are a significant consideration in
fate assessments of creosote contamination.  

Additionally, colloidal matter present in a cresote-contaminated
environment has been found to affect the desorption rates of specific
PAHs.  One study found that PAHs partitioned to course (of sizes >100
nm) colloid fractions and were linearly correlated with the PAH
octanol-water partition coefficient, indicating the partitioning was
hydrophobic (Villholth, 1999).

The PAHs from creosote-treated utility poles and/or railway ties tend to
leach out initially in the first seven days and remain in the sediment
surrounding the poles or railroad ties not migrating far from the wood. 
Most of the PAHs, however, tend to stay inside the wood (~85%).  One
study showed that background levels for PAHs leached from wood were
attained within three months and may have been due to photolysis or
biodegradation of the PAHs.  A detailed study of 200 U.S. estuaries
showed that PAHs that leached from the treated wood of decks and
bulkheads, 175 had muddy sediments.  Higher amounts of PAHs leached into
such soil types.  No systematic work has been carried out on all the
PAHs and the representative soil types to show which one would have a
higher tendency of retention for the PAHs. 

Migration studies of PAHs into groundwater have shown that migration of
some of the PAHs does take place.  Vertical or lateral migration of the
PAHs from the utility poles indicated that at ground level the migration
was not significant beyond 150 meters from the site of contamination
(base of utility poles).  The vertical or downward migration of the PAHs
was even more limited and the existence of the PAHs were not found below
a 12 meter depth.  

Biodegradation

Most of the PAHs have a tendency to biodegrade under aerobic conditions.
 It has been reported that over eighty percent of biodegradation occurs
in the first month after the wood preservative application, with the
exception of benz[a]pyrene and benzo[k]fluoranthene, which have shown
resistance to biodegradation.  A number of aerobic soil metabolism
studies on PAHs conducted at various contaminated sites as well as in
simulated microcosms reported that low molecular weight PAHs generally
metabolized in aerobic conditions and the greater the oxic environment,
the higher the  biodegradation level. 

Aerobic degradation of PAHs associated with soil and groundwater often
leads to a rapid depletion of dissolved oxygen which eventually
decreases the redox potential.  This decrease in the redox potential can
result in favorable environments for denitrifying, sulfate-reducing, or
methanogenic microbes.  Therefore, anaerobic transformations may be a
significant factor in oxygen-depleted habitats (Karthikeyan and
Bhandari, 2001).  Under these conditions, anoxic or anaerobic
degradation stimulated by denitrifying or sulfate-reducing bacteria can
become an important pathway for the cleanup of contaminated sites.

Bioaccumulation in Fish

The major components of the PAHs in creosote have shown the ability to
form neutral to oxidized transformation products under aerobic
soil/aquatic conditions.  For example, fluorene forms hydroxy fluorene
and acenaphthene converts into diacetic acid acenaphthene.  These
oxidized species are stable and bioaccumulative.  Numerous studies have
shown that photooxidized transformation products of these PAHs are
bioaccumulative and result in adverse effects on the aqueous biota as
well as on the organisms in the soils and benthic sediments.

In aquatic habitats, fish, shellfish, and crustaceans readily
bioaccumulate PAHs from the environment and store these at high levels
in the tissues.  Seven PAHs: naphthalene, anthracene, phenanthrene,
pyrene, 9-methyl anthracene, benz[a]anthracene and perylene were shown
to bioaccumulate in Daphnia pulex.  PAHs like naphthalene,
biphenyl/acenaphthylene, fluorene, phenanthrene/anthracene/chrysene, and
benzopyrene were found to bioaccumulate in clams (Rangia cuneate).  The
most dramatic increases were in cases such as 
benz[a]anthracene/chrysene which reported a bioaccumulation of 41 ppb
(week zero) and increased to 190 ppb (week 4).  Similarly, benzopyrene
bioaccumulated from 8 ppb (week zero) to 600 ppb (week 4).  Depuration
was within two weeks.  This study was conducted after the creosote spill
into the Bayou Bonfuca at the American Creosote Works Plant Site at
Slidell, Louisiana.A study on benthic invertebrates showed a
bioaccumulation concentration ranging from 0.10 to  11.00 ppm.  A
bioconcentration study on zebra mussels in the Great Lakes found that
pre-spawning species (high lipid) bioaccumulated benzo[a] pyrene at a
faster rate than the post- spawning (low lipid) species.

Bioaccumulation data on marine mammals are not readily available, and
only one study on whales and seals has been reported.  That study
indicated a bioaccumulation of 0.10 and 1.21 ppm in the muscles of these
mammals, respectively.

Some of the PAHs, particularly those that have a high molecular mass
(higher number of the fused aromatic rings) have a higher tendency to
adsorb to soil organic carbon.  Such adsorption coefficients (KOC) have
been reported in literature.  Some PAHs with a high Koc can bind
strongly with the organic carbon of the soils/sediments and may not be
bioavailable to the aquatic organisms. However, if the octanol/ water
coefficients (Kow) is high, and if the PAHs are desorbed from the
soils/sediments to which they are bound, some of these PAHs can become
bioaccumulative to the benthic organisms. 

It has been suggested that based on theoretical calculations and
modeling, the half-lives of the PAHs obtained from coal tar creosote can
be estimated in air, water, soils and sediments.   From these
calculations and modeling, one can arbitrarily divide the PAHs into 3
distinct groups: PAHs with two fused rings, PAHs with three fused rings
and PAHs that have 4 to 5 fused rings. The half-lives in the
environmental compartments (air, water, and soils) for PAHs are as
follows: the half-lives of two fused rings PAHs < three fused rings <
4/5 fused rings.  The Kow values lie between 3 and 4 for PAHs with two
fused rings, between 4 and 5 for PAHs with 3 fused rings, and at 6 or
above for the third group of the PAHs.  In general, the half-lives in
air and water environmental compartments are much lower than in
soils/sediments because the soil adsorption coefficients are higher. 
The longer the half-life, the greater the persistence of the PAHs in
soils.  Some of the 4/5 fused ring PAHs are more persistent in soils and
sediments and since their Kow values are higher, they can bioaccumulate
but some of them adsorb onto soils and they may not be bioavailable for
benthic organisms.

The third group of PAHs show a higher degree of bioaccumulation,
persistence in soils and water, resistance to biodegradation and
photooxidation.  Additionally the components of these group have less of
a  tendency to leach from the wood structure and have high sorption
constants to soils.  On the other hand, these higher members of the PAHs
(4/5 fused ring compounds) are not readily soluble in water and their
percent on a mass basis in the creosote mixture is very low compared to
the first group (2 fused rings) of the PAHs, and may not be available
for biomagnification and migration into surface and ground waters.

A recent two year mesocosm study (K. Brooks, 2004) of  leaching and
migration of  PAHs from creosote-treated railway ties into  ballast,
wetland sediments, ground water, storm water and soil showed that
ballast and sediment core samples found  the presence of PAHs at an
initial level of 1.207 µg/g at 0 cm to 0.482 µg/g  at 60 cm depth. One
out 16 storm water sample, collected after 18 months, showed the
presence of two PAHs: benzo(a) anthracene (0.00019 mg/L) and
phenanthrene (0.00066 mg/L)

B.  APPENDIX OF THE ENVIRONMENTAL FATE AND TRANSPSORT ASSESSMENT For
CREOSOTE

A. Chemical Profile

Common Name(s):	Creosote, Creosote Oil, Dead Oil, Brick Oil, Coal Tar
Oil, Creosote P1, Heavy Oil, Liquid Pitch Oil, Wash Oil, Creosotum,
Cresylic Creosote, Naphthalene Oil, Tar Oil, AWPA #1, and Preserv-o-sote

Chemical Name:	Coal Tar Creosote

Trade Name:		Sakresote 100

Formulations:		Distillate mixture obtained from bituminous coal;
oil-based

Physical/Chemical Properties:

Molecular formula:	It is a multi-component mixture. Not applicable

Molecular weight:	It is a multi-component mixture. Not applicable

Physical State:	Translucent brown to black; yellowish to dark
green-brown; oily liquid

Melting Point:	Not available

Boiling Point:	194-400(C. Not possible to define a boiling point of  a
multi-component mixture

Viscosity:	14.60 mm/s (P1/P13); 15.5 mm/s (P2)

Vapor Pressure:	11.1 mm Hg at 24.4(C (P1/P13); 8.6 mm Hg at 24.4 to
24.5(C (P2)

Dissociation Constant: 3.247 (pKa, P1/P13); 3.311 (pKa, P2)

Solubility: (Water)	313 µg/mL (P1/P13); 306 µg/mL (P2); temperature
not specified

Henry’s Law Constant: Not available. Not possible to determine for a
multi-component  mixture

Octanol/Water Partition Coefficient (log Kow): 1.0 

 Studies have not been submitted to the Agency and specific guideline
requirements have not been fulfilled for creosote.  Therefore, the
Agency has carried out an open literature search for environmental fate
studies on creosote mixtures and has decided to make the assessment
based on the PAH constituents only.  The literature search on
environmental fate studies provided data on the volatility, photolysis
in water, aerobic/anaerobic metabolism, leaching and
adsorption/desorption, bioaccumulation in aquatic and benthic organisms,
and migration from poles into soils.  The following are summaries of
studies obtained from the literature search.

II a.

Volatility

Lindhardt, B. and T.H. Christensen, 1996

A laboratory volatilization study was conducted on the non-steady-state
fluxes of aromatic hydrocarbons from coal tar contaminated soil.  The
contaminated soil samples, obtained from Holte, Denmark, were placed
below a 5 cm deep layer of uncontaminated soil and monitored for 53
days.  The contaminated soil contained 50 to 840 µg/cm3 of 11 selected
aromatic hydrocarbons.  In analyses where the microbial activity was
inhibited, the fluxes stabilized on a semi-steady state level for
monocyclic aromatic hydrocarbons, naphthalene and 1-methylnaphthalene
after 10 to 20 days.  Acenaphthene and fluorene fluxes were measurable
only in experiments using a soil cover with a low organic content.  When
the soil cover was adapted to degrade naphthalene, the fluxes of
naphthalene and 1-methylnaphthalene were approaching the detection limit
at 5 to 8 days.  

Gevao, B. and K.C. Jones, 1998 

A volatilization study for five PAHs (acenaphthene, fluorene,
phenanthrene, anthracene and fluoranthene) from treated (painted) wood
was conducted in the United Kingdom.  This laboratory study was
performed using glass chambers equipped with an air inlet/outlet in 4(C
and 30(C environments.  Wood samples painted with 110 grams of creosote
were placed in the glass chambers and the air traps were changed at each
sampling interval.  Samples were stored for 2 to 4 weeks at -17(C prior
to extraction.  The rate of desorption was analyzed using first order
kinetics for all five PAHs and was found to be higher at 30o C than at
4oC.  The mean PAH values ranged from 2.57 ± 1.52 mg/m2 treated
wood/day and 29.5 ± 6.1 mg/m2 treated wood/day at 4(C and 30(C,
respectively.  From the desorption rates, the half-life at 4oC ranged
from 0.70 year to 31 years for fluoranthene and acenaphthene,
respectively.  When the temperature was raised to 30(C, the half-life
ranged from 0.3 year to 1 year for fluroanthene and acenaphthene,
respectively.  Following a long-term study at 4(C, it was observed that
the volatilization rate was constant for about seven weeks after which
it was estimated that >85 percent of the PAHs remained in the wood.  The
authors noted that initial desorption rates were caused by partitioning
between the wood and air, and by the rates of compound diffusion from
the interstices of the wood.

II b. Aqueous Photolysis

Kirso, U., 1991

A photolysis study in natural sunlight in aqueous (~5 x 10-8 M) PAH,
aza-PAH and benzene media was performed with selected PAHs.  Table 1
summarizes the photooxidation half-lives of these commercially available
highly pure PAHs.  The study also collected data on photolysis by
natural sunlight of benzo[a]pyrene under open-sea conditions at northern
and southern latitudes.  In this European study, first-order
photooxidation rates were measured and first-order rate constants were
found.  The Agency has calculated the half-lives using the first-order
rate constants from this study.  PAH appear to photolyse through
oxidation processes. The results of the calculations are presented in
Table 1.

Table 1.  Photochemical Oxidation and Half Lives of Selected PAHs

Compound	

Half-life in Water (minutes)	

Half-life in Benzene (minutes)



Fluorene	

119.97	

2.05



Anthracene	

8.06	

*



Phenanthrene	

64.95	

107.05



Triphenylene	

93.93	

136.80



Pyrene	

19.99	

97.47



Chrysene	

25.88	

100.54



Benz[a]anthracene	

20.92	

41.72



Benz[b]anthracene	

3.56	

0.68



Dibenz[a,c]anthracene	

25.57	

55.06



Dibenz[a,h]anthracene	

22.94	

55.06



Dibenz[a,j]anthracene	

16.58	

41.04



Dibenzo[a]pyrene	

15.5	

34.20



Dibenzo[e]pyrene	

22.94	

117.99



Perylene	

374.94	

51.98



Coronene	

312.01	

63.95



Benzo[k]fluoranthene	

110.98	

171.00



Benzo[b]fluoranthene	

312.01	

190.15

Notes:	1.  Not all the PAHs used in the study are present in the P1/P13
and P2 fractions of the creosote.

2.  PAHs with high molar masses have longer  photooxidation half-lives
than smaller PAHs with small molar masses.

3.  * =  Could not be determined under the experimental conditions.



Table 2.  Rate Constants and Half-lives of Photooxidation of
Benzo[a]pyrene by Sunlight in Sea Water

Region	

Water Temperature 

(oC)	

Initial Concentration

(10-8 M)	

Rate Constant

(k x 10-4) 1/s	

Half-life

(minutes)



Bering Sea	

14	

1.47	

1.69	

68.3



Bering Sea	

16	

4.20	

1.60	

72.2



Tropical Pacific	

26	

6.59	

2.84	

40.6



Tropical Pacific	

27	

2.06	

1.60	

72.2



Lagune Caroline Atoll	

27	

1.90	

2.99	

38.61



Lagune Caroline Atoll	

27	

0.70	

4.20	

27.50



Baltic Sea	

6	

15.6	

0.70	

146.2

Notes:	1.  ‘K’ is a first-order rate constant.

2.  The Rate constants of photooxidation of benzo[a]pyrene (and half
lives) are higher in tropical longitudes than in northern and temperate
climatical zones.

Experimental and calculated data on the photolysis of most of the PAHs
present in the P1/P13 and P2 fractions of creosote can be found in the
international scientific literature.  These data are summarized in the
Table 3 below.

Table 3.  Photolysis of Selected PAHs Found in P1/P13 and P2 Fractions
of Coal Tar Creosote

Compound	

Description	

Half lives



Naphthalene	

Direct sunlight, 40(N, midday, midsummer (calculated)	

71 Hours (Harris, 1982)

	

Distilled water at 25 o C	

25 Hours (Fukuda, 1988)



Quinoline	

Sunlight at 40(N, aqueous hydrolysis, pH 6.9 	

3851 Hours, summer

535 Hours, winter (Mill et al., 1981)



1-Methyl naphthalene	

Summer sunlight in surface water	

180 Days (Miller, 1985)





2-Methyl naphthalene	

Summer sunlight in surface water	

410 Days (Mill et al., 1981)

	

Distilled water	

16.4 Hours (Fukuda, 1988)



Acenaphthene	

Determined by rotary photoreactor technique on different atmospheric
particulate substrates:

-- silica gel

-- alumina

-- fly ash	

2.0 Hours (Behymer & Hites, 1985)

2.2 Hours (Ibid.)

44 Hours (Ibid.)

	

Distilled water(irradiated light at wavelength 290 nm)	

3 Hours (Fukuda, 1988)



Fluorene	

Determined by rotary photoreactor technique on different atmospheric
particulate substrates:

-- silica gel

-- alumina

-- fly ash	

110 Hours (Behymer & Hites, 1985)

62 Hours (Ibid.)

37 Hours (Ibid.)



Anthracene	

Midsummer sunlight:

-- deep, slow, somewhat turbid water

-- deep, slow and muddy water

-- deep, slow and clear water

-- Shallow, fast and clear water

-- shallow, very fast and clear water	

173.2 hours (Southworth, 1977)

693 hours (Ibid.)

38.5 hours (Ibid.)

8.1 hours (Ibid.)

2.91 hours

	

At 35( on latitude

-- summer

-- winter	

1.6 hours (Ibid.)

4.8 hours (Ibid.)

	

At 35( N, winter

-- in water	

4.62 hours (Callahan, 1979)

	

Different atmospheric particulate substrate:

-- silica gel

-- alumina

-- fly ash

-- distilled water	

2.9 hours (Beymer & Hites,1985)

0.50 hours (Ibid.)

48 hours (Ibid.)

1.0 hour (Fukuda, 1988)



Carbazole	

40(N , midday, sunlight in late Jan. and river (calc.)	

6.0 hours (Smith, 1978)

	

In eutrophic pond and eutrophic lake	

15.0 hours (Ibid.)

	

In oligotrophic lake	

3.0 hours (Ibid.)

	

Aqueous medium	

1.0 hours (Ibid.)



Fluoranthene	

Atmospheric/aqueous photolysis, based on the measured sunlight
photolysis rate constant in water

-- summer sunlight, surface water	

21 hours (Howard, 1991)

160 days (Mabey, 1982)

	

Different atmospheric particulate substrates:

-- silica gel

-- alumina

-- fly ash	

74 hours (Behymer & Hites, 1985)

23 hours (Ibid.)

44 hours (Ibid.)



Chrysene	

Different atmospheric particulate substrates:

-- silica gel

-- alumina

-- fly ash	

100 hours (Behymer & Hites, 1985)

78 hours (Ibid.)

38 hours (Ibid.)



Acenaphthylene	

Different atmospheric particulate substrates:

-- silica gel

-- alumina

-- fly ash	

0.7 hours (Behymer & Hites, 1985)

2.2 hours (Ibid.)

44 hours (Ibid.)



Benz[a]anthracene	

Aquatic	

10-50 hours (Callahan, 1979)

	

Stream	

20 hours (Smith, 1978)

	

Eutrophic pond or lake 	

50 hours (Ibid.)

	

Oligotrophic lake 	

10 hours (Ibid.)

	

Aquatics	

0.58 hours (EPA Report 600/7-78-074)

	

Early March	

0.2 days (Zepp, 1980)

	

-- 1% acetonitrile in filter-sterilized natural water at 313 nm wave
length	

5 hours

	

Different atmospheric particulate substrates:

-- silica gel

-- alumina

-- fly ash	

4.0 hours (Behymer & Hites)

2.0 hours (Ibid.)

38 hours (Ibid.)



IIc. Aerobic/anaerobic Metabolism

Mueller, J.G. et al., 1993

This laboratory study was designed to evaluate the ability of a
sequential inoculation process using selected microorganisms to enhance
the bioremediation technologies for the treatment of groundwater
contaminated with creosote and PCP.  The contaminated groundwater was
obtained from a monitoring well at the American Creosote Works site in
Pensacola, Florida.  Both 1.2 L (bench scale) and 454 L (pilot scale)
bioreactors were utilized for the analysis.  The bench scale study
showed that after 32 days of continuous-flow operation, the majority of
the monitored creosote components were degraded.  Overall, for groups 1,
2, and 3 PAHs, the biodegradation values were determined to be 98.0,
96.2, and 89.4 percent, respectively.  The amount of group 2 and 3 PAHs
found in the bioreactor residues were 3.5 and 9.2 percent, respectively.
 In the pilot scale study, the system was effective in treating the
contaminated water in a two-step process.  Since the pattern of
degradation favored the low molecular weight components, an additional
inoculum of microorganisms selected for their ability to degrade this
components were added.  As a result, >98 percent of all the monitored
creosote components were removed.  A mass balance distribution analysis
showed that of the various routes of removal (adsorption,
volatilization, and biodegradation), biodegradation was the primary
mechanism for the removal of the creosote components.

National Oceanic and Atmospheric Administration, 1988

The National Oceanic and Atmospheric Administration (NOAA) conducted a
detailed study on the analysis of PCBs, PAHs, and 12 trace metals
present in surface water sediments of 200 estuaries.  One hundred
seventy-five estuaries were contaminated with PAHs.  These estuaries
have muddy rather than sandy sediments with the exception of two
estuaries in Long Island, where the sediments were predominantly sandy. 
The results of the study indicated that PAH contamination was higher in
the muddy sediment. 

Bouwer, E.J., W. Zhang, L.P. Wilson, and N.D. Durant, 1996

A four-week aerobic study was conducted on the sediment and groundwater
of an abandoned manufactured gas plant (MGP) site which was contaminated
with creosote PAHs .  The experiment was conducted to assess the ability
of the bacteria in sediment to mineralize

14C-labeled benzene, naphthalene, and phenanthrene under simulated field
conditions.  The results of this study showed that the bacteria present
in the aquifer sediments were able to degrade low molecular weight PAHs.
 The study also determined that the higher the concentration of oxygen,
the greater the biodegradation of these molecules.  Mineralization for
anthracene and phenanthrene ranged from 4 to 23 percent, ranged from 4
to 42 percent for benzene, and from 8 and 55 percent for naphthalene.

Chapman, P.J., M. Shelton, M. Grifoll, and S. Selifonov, 1995

In this study, bacterial cultures were obtained from washing
creosote-contaminated soils at lumber treatment facilities.  These
cultures were to be utilized as an enrichment for biodegradation
studies.  Purified PAHs were used as the sole carbon sources (0.1%) and
were inoculated with the bacterial cultures.  Increases in turbidity and
color changes were monitored to determine growth and biological
activity, respectively.  Low molecular weight PAHs were easily degraded
by the bacterial enrichments.  Approximately 72 percent of the measured
PAH were degraded, which accounted for 52.5 percent of the weight of the
initial PAHs.  These aerobic degradation processes of creosote PAHs have
shown that biodegradation in the presence of certain bacteria is
accompanied with the formation of neutral and acidic oxidized products. 


Schocken, M.J. and D.T. Gibson, 1984

The metabolism of acenaphthene and acenaphthylene by two strains of
bacteria was examined in this study.  Biejerinckia species and a mutant
strain Biejerinckia species B8/36 bacteria were found to oxidize the two
PAHs.  The study was carried out with large-scale incubations at 30(C.
Acenaphthene oxidized into 1-acenaphthenone, 1,2-acenaphthenediol,
acenapthenequinone and 1-acenaphthenol; acenaphthylene oxidized into
acenapthenequinone.  The results indicated that even though these PAHs
were both oxidized to acenapthenequinone, the pathways to form this
product are quite different.  

Godsy, E.M., D.F. Goerlitz, and D. Grbic-Galic, 1992

(Anaerobic Study) The US Geological Survey selected an abandoned wood
preserving site once used for creosote and pentachlorphenol pressure
treatments to analyze and identify the contaminants in groundwater from
plant waste.  This waste had been discharged into the unlined surface
impoundments that were in direct hydraulic contact with groundwater. The
groundwater was determined to be anaerobic and showed the presence of
methane and hydrogen sulfide (indicating the presence of methanogenic
and sulfuryl microbes).  To conduct the study, holes were bored to a
depth of 6.1 meters at various sites on a downward slope from the
contamination source ponds.  Table 4 lists the concentrations of PAHs at
the various sites.  The results indicated that lateral migration of most
of the PAHs became undetectable 150 meters from the source of
contamination.  However, the study authors did not indicate the age of
the wood and how long the preserving plant had been abandoned before
testing began.

Table 4.  Amounts of PAHs in the Water Samples From 6.1 Meter Wellsa

Compoundb	

Site 1c

(uncontam.)	

Site 3 

(6.0 m)	

Site 39

(53 m)	

Site 40

(99 m)	

Site 4 

(122 m)	

Site 37

(150 m)



Indene	

ND	

1.25	

0.24	

ND	

ND	

ND



Naphthalene	

ND	

9.38	

3.39	

2.89	

0.93	

1.54



1-Methylnaphthalene	

ND	

0.41	

0.32	

0.25	

0.06	

0.11



2-Methylnaphthalene	

ND	

0.99	

0.32	

0.54	

0.10	

0.10



Acenaphthene	

ND	

0.52	

0.29	

0.33	

0.05	

ND



Indole	

ND	

ND	

ND	

ND	

ND	

ND



Quinoline	

ND	

11.2	

0.01	

ND	

ND	

ND



Benzothiophene	

ND	

0.83	

0.31	

0.22	

0.16	

0.16



Dibenzofuran	

ND	

0.30	

0.04	

0.16	

ND	

ND

ND = Not detected

a - Water samples collected from wells on these sites.

b - Analyzed and quantified  through GC/MS methods.

c - Uncontaminated site

Goerlitz, D.F. et al., 1985

A similar study was conducted by the US Geological Survey at the same
wood preserving plant that was used in the study of Godsy et al. (1992).
 Results showed that on Site 3 at a 30-meter well, naphthalene (15.60
ppm) was detected 6 meters from the source of contamination and the
level of naphthalene progressively decreased with an increase in well
depth.  At 24 meters, only 0.60 ppm of naphthalene was detected.  PAHs
were not detected at a depth greater than 12 meters.

Genther, B.R., et al., 1997

A year long study on anaerobic soil biodegradation was conducted at the
American Creosote Works Superfund site located in Pensacola, Florida
using soil samples collected from the creek bed sediment.  Samples were
collected at a depth of 5 to 8 cm and 12 meters, which was below the
main discharge pond.  These soil samples were selected because they
represented bacterial populations exposed to minimum and maximum
concentrations of the PAHs in situ.  The PAHs used were from an
artificial mixture simulating the PAH components of creosote.  Various
batches of the artificial mixture were made and consisted of the
following components: naphthalene (36 mg), 1-methylnaphthalene (10.8
mg), 2-methylnaphthalene (10.8 mg), 2,6-dimethylnaphthalene (10.8 mg),
biphenyl (5.4 mg), acenaphthene (5.4 mg), fluorene (10.8 mg),
phenanthrene (18.0 mg), anthracene (18.0 mg), 2-methylanthracene (9.0
mg), anthraquinone (3.6 mg), fluoranthene (9.0 mg), pyrene (3.6 mg),
chrysene (3.6 mg), 2,3-benzo[b]fluorene (3.6 mg), and benzo[a]pyrene
(3.6 mg).  Various batches of this artificial PAH mixture were
inoculated with the contaminated soil samples under methanogenic,
sulfidogenic, and nitrate-reducing conditions. 

The majority of the PAHs did not degrade in the soil samples collected
from the creek bed (5-8 cm depth).  Loss of some bicyclic and tricyclics
were observed; however, 4- and 5-membered PAHs did not degrade under
methanogenic, sulfidogenic, or nitrate-reducing conditions.  By 16
weeks, under methanogenic conditions, the maximum loss of naphthalene
was 47 percent (similar to naphthalene loss under an abiotic control (40
percent)).  Among the tricyclics, anthraquionone was the only substance
where any loss under methanogenic conditions was  reported. 
Anthraquinone loss reached 48 percent and about 65 percent by weeks 28
and 52, respectively.  Two other tricyclics, anthracene and
acenaphthene, degraded 22 percent under the same conditions. Under
nitrate-reducing conditions, only degradation for 2-methylanthracene was
observed and the concentration of this chemical was below the detection
limit between weeks 8 to 28.  Under sulfidogenic conditions, only
anthraquinone degraded 22 percent by week 8.  Under these conditions, no
other PAHs showed any biodegradation up to week 52.  For the 12-meter
soil samples, no appreciable biodegradation processes were observed for
most of the PAHs under all three anaerobic conditions.

III Miscilinous Studies

Anaerobic soil metabolism studies for individual PAHs have been
conducted over a period of time using the three anaerobic conditions. 
For nitrate reducing conditions, the studies conducted were Bouwer and
McCarty, 1983; Al-Bashir, et al., 1990; Ehrlich et al., 1982a; Flyvberg
et al. 1993; Hambrick et al., 1980; Kuhn et al., 1988; and Mihelcic and
Luthy, 1988a, 1988b.  The study was conducted for sulfidogenic
conditions by Flyvjberg et al., 1993.  The studies  for methanogenic
conditions include Ehrlich et al. 1982a; and Godsy et al., 1992.  Other
in situ studies for biodegradation of PAHs at creosote contaminated
sites are: Ehrlich et al. 1982a, 1982b; Godsy et al. 1992; Goerlitz et
al., 1985; and Mattraw and Frank, 1986.

Bauer, J.E. and D.G. Capone, 1985

Microorganisms present in the intertidal sediments were investigated for
the degradation of anthracene and naphthalene.  No mineralization was
observed under anaerobic conditions.  However, mineralization did show
dependence on the amounts of the polyaromatics present, oxygen level,
and pre-exposure time.  Maximum mineralization of these two PAHs
occurred after one or two weeks of pre-exposure.  A similar study
conducted by the same authors (in 1988) showed that the pre-exposure of
anthracene and naphthalene under aerobic conditions (marine sediments
collected from 0 to 1 cm depth) to benzene and other PAHs accelerated
their mineralization.

Mueller, J.G. et al., 1991

Groundwater samples collected from a depth of 7 meters from the American
Creosote Works superfund site in Pensacola, Florida were examined for
biodegradation (aerobic aquatic metabolism) of PAHs from coal-tar
creosote.  Creosote-contaminated soil from the site was collected and
used to prepare the microbial inoculum.  Groundwater samples and the
microbial inoculum were incubated for 14 days at 30(C.  The groundwater
samples contained the PAHs present in coal-tar derived creosote and in
the phenolic components (which constituted 5% of the creosote sample). 
In addition, PCP was also present.  Table 5 summarizes the results of
the 14-day incubation experiment.  The analysis showed that bicyclic
PAHs and phenolics are metabolized readily by the microorganisms in the
aquatic soil while the PAHs with a higher number of fused rings
biodegrade more slowly.  By Day 8 most of the biodegradation was
complete.  A few nitrogenous heterocyclics like quinoline, isoquinoline,
and acridine readily metabolized while quinaldine, carbazole, and
pentachlorphenol showed resistance to the metabolic process under the
conditions of the experiment.

Table 5.  Aquatic Soil Metabolism (Biodegradation) Results for
Creosote/PCP Contaminants  In Groundwater Samples From Pensacola,
Florida Superfund Site

Compound	

Concentration (µg/mL) After Incubation 

	

Initial Conc.	

1 Day	

3 Days	

5 Days	

8 Days	

14 Days	

Sterile Control



Naphthalene	

28.7	

17.2	

0.1	

U	

0.1	

U	

25.6



2-Methylnaphthalene	

4.7	

3.0	

U	

U	

0.1	

U	

4.5



1-Methylnaphthalene	

9.5	

5.7	

2.1	

1.5	

U	

U	

8.2



Biphenyl	

3.0	

1.7	

1.2	

U	

U	

U	

2.6



2,6-Dimethylnaphthalene	

2.4	

1.4	

1.2	

1.2	

1.0	

0.3	

2.1



2,3-Dimethylnaphthalene	

1.3	

0.8	

0.5	

0.8	

0.7	

0.2	

1.0



Acenaphthalene	

0.6	

0.3	

0.4	

0.6	

0.6	

0.2	

0.4



Acenaphthene	

13.6	

9.0	

8.3	

9.6	

9.7	

1.8	

11.9



Fluorene	

11.6	

7.8	

8.0	

5.2	

1.8	

0.1	

9.9



Phenanthrene	

32.8	

23.5	

23.1	

15.4	

0.3	

U	

27.7



Anthracene	

4.7	

3.2	

3.0	

2.7	

2.2	

0.5	

3.9



2-Methylanthracene	

5.2	

3.7	

3.7	

4.0	

4.2	

1.5	

4.4



Anthraquinone	

3.3	

2.7	

1.9	

U	

U	

U	

2.9



Fluoranthene	

16.2	

11.5	

11.5	

13.3	

13.5	

7.6	

14.4



Pyrene	

10.4	

7.8	

7.3	

8.2	

8.3	

4.7	

9.8



Benzo[b]fluorene	

2.5	

1.7	

1.7	

1.8	

2.0	

1.2	

2.0



Chrysene	

2.7	

1.8	

1.8	

2.0	

2.1	

1.2	

2.4



Benzo[a]pyrene	

2.1	

0.5	

U	

U	

U	

0.9	

2.0



Benz[a]anthracene	

2.9	

2.0	

2.0	

2.0	

2.2	

1.3	

2.7



Benzo[b]fluoranthene/benzo[k]-

Fluoranthene	

2.9	

2.8	

2.0	

2.1	

2.1	

1.7	

2.8



Indeno[1,2,3-c,d]pyrene	

1.9	

1.3	

1.4	

1.4	

1.2	

0.9	

1.8



2,6-Xylenol	

1.1	

0.6	

0.2	

0.1	

0.1	

U	

0.8



o-Cresol	

4.2	

2.7	

0.3	

0.2	

0.2	

U	

4.9



2,5-Xylenol	

0.1	

U	

U	

U	

U	

U	

0.1



2,4-Xylenol	

0.2	

U	

U	

U	

U	

U	

0.2



p-Cresol	

2.0	

0.1	

U	

U	

U	

U	

2.3



m-Cresol	

2.5	

1.9	

U	

U	

U	

U	

2.3



2,3-Xylenol	

0.2	

0.1	

U	

U	

U	

U	

0.1



3,5-Xylenol	

1.3	

0.5	

0.2	

0.1	

U	

U	

1.1



3,4-Xylenol/2,3,5-trimethylphenol	

0.4	

0.1	

0.1	

0.1	

U	

U	

0.3



PCP	

0.1	

0.3	

0.1	

0.1	

0.1	

0.1	

0.1



2-Picoline	

0.3	

0.2	

0.2	

0.2	

0.1	

U	

U



3-Picoline/4-picoline	

U	

U	

U	

U	

U	

U	

U



Lutidine	

0.9	

0.7	

0.6	

0.5	

0.5	

0.6	

0.8



Thianaphthene	

20.3	

12.5	

2.6	

1.2	

0.5	

0.3	

23.4



Quinoline	

4.3	

1.7	

0.5	

0.2	

0.2	

0.2	

3.6



Isoquinoline	

1.5	

0.9	

0.3	

0.3	

0.1	

U	

1.4



Quinaldine	

3.4	

3.2	

2.8	

2.4	

0.6	

0.3	

4.9



Lepidine	

0.7	

0.6	

0.4	

0.3	

0.3	

0.2	

0.7



Dibenzofuran	

5.5	

5.9	

5.8	

3.4	

1.0	

0.7	

6.1



Dibenzothiophene	

3.8	

2.8	

3.1	

2.0	

1.3	

1.2	

3.1



Acridine	

22.5	

18.2	

14.1	

1.9	

2.0	

2.0	

26.2



Carbazole	

2.9	

2.1	

1.2	

0.9	

0.8	

1.0	

3.0

Note:	Reported data are averages of duplicate samples.

U - Below the detection limit.

Hurst, C.J. et al., 1996

An aerobic and anaerobic soil metabolism study was conducted at the
Champion International superfund site in Libby, Montana.  The
contaminated soil samples were spiked with for analysis.  Biodegradation
for 14C-pyrene and seventeen PAHs (including pyrene) was followed by 0
percent, 2 percent, and 21 percent oxygen as soil gas.  The PAHs chosen
for the study were: naphthalene, acenaphthylene, acenaphthene, fluorene,
anthracene, phenanthrene, fluoranthene, pyrene, benzo[a]anthracene,
chrysene, benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[a] pyrene,
dibenzo[a,h] anthracene, benzo[g,h,i]perylene, and indeno(1,2,3-cd). 
After 70 days, 45 to 55 percent of the 14C-pyrene was mineralized in 2
percent  and 21 percent oxygen.  At 0 percent oxygen, no statistically
significant mineralization was observed.  For eight of the
non-radiolabeled PAHs, biodegradation in an oxygen atmosphere ranged
from 6.2 percent (naphthalene) to 57 percent (pyrene).  The remaining
PAHs were below the limit of detection.

IV Leaching and Adsorption/Desorption

Villholth, K.G., 1999

The objective of this study was to determine the amount of colloids in
the groundwater of two creosote-contaminated aquifers in Denmark and to
determine the in situ distribution of the PAHs between the dissolved and
colloidal phases in the water.  The colloids identified at the sites
were clay minerals, iron-oxides, iron-sulfides, and quartz particles
containing significant amounts of organic carbon.  The results of a
two-step fractionation procedure, showed that the PAHs partitioned to
the course (>100 nm) colloid fraction (log Koc) and was linearly
correlated with the PAH octanol-water partition coefficient (log Kow),
indicating the partitioning was hydrophobic.  This suggested a potential
for colloid-facilitated transport of PAHs.  The lack of PAH partitioning
to colloids <100 nm, indicates a weaker binding to the smaller, more
hydrophilic colloids in the groundwater.

Rutherford, P.M., et al., 1997

This study was designed to determine if a 10-week slurry phase
bioremediation treatment altered the desorptive properties of
two-creosote-contaminated soils.  Soil samples were collected from an
inactive wood preserving facility in Edmonton, Alberta (EDM site) and
Prince Albert, Saskatchewan (PAA site).  For bioremediation, soil
samples were combined with aqueous nutrient media under aerobic
conditions in a bioreactor.  Desorption of 14C-naphthalene from the
contaminated soils was measured before and after bioremediation in a
sequential batch experiment.  The contributions of the contaminant
organic phase and the soil organic matter to desorption were determined
by experiments on soils with and without nonaqueous phase liquid
contaminants.  Results showed that total extractable organics were 43 to
31 percent lower in the bioremediated soils for both sites.  This
reduction in total organic carbon content lowered the sorption capacity
of the soils.  The desorptive partition coefficient (Kd) for the
nonbioremediated EDM soil was significantly (p = 0.022) greater than the
nonbioremediated PAA soil.  However, after bioremediation, no
significant difference was found between the two soils (p = 0.11). 
Although the Kd decreased due to bioremediation, the carbon-based
partition coefficients on the nonaqueous phase liquid did not change
significantly once the changes in the overall composition of the soil
had been accounted for. 

Priddle, M.W. and K.T.B. MacQuarrie, 1994

A laboratory study was conducted to evaluate the dissolution of
industrial creosote in water using a generator column to determine the
impacts on groundwater quality and the kinetics of the dissolution
process.  The laboratory results were also compared to an equilibrium
model and a quasi-kinetic model.  For the laboratory study, the
generator column was packed with a 10 percent creosote effluent obtained
from Carbochem Inc. (Hamilton, Ontario, Canada).  The study focused on
10 specific PAHs including naphthalene, phenanthrene and benzo[a]pyrene.
 A mass-transfer rate test was conducted to evaluate the rate the
creosote components reached equilibrium concentrations.  Additionally,
two long-term dissolution tests were conducted by passing water through
the column for a designated period of time.  The mean contact time in
the column was 0.56 and 1.02 hours during the two testing periods.  The
results of the mass transfer test indicated that the creosote components
reached equilibrium with the aqueous phase in about 60 hours.  The two
long-term tests found that the concentrations of the targeted PAHs
detected in the effluent all decreased steadily throughout the
monitoring period.  The initial concentrations of the components were
approximately 40 percent of the calculated effective solubilities.  The
higher molecular weight compounds were not detected which was expected
due to the low effective solubilities (<0.002 mg/L-1).  Overall, the
ratios of these concentrations were in proportion to their effective
solubilities which were calculated using Raoult’s law and the creosote
composition data.

Padma, T.V. et al., 1999

A study was conducted to monitor the effects that various processes
(tides, storms, bioturbation, shipping, and dredging) may have in the
dissolution and resuspension of sediment-associated PAHs.  These
environmental processes were mimicked by creating a water-soluble
fraction from the creosote-contaminated sediment and artificial
estuarine water.  Creosote-contaminated sediment samples were collected
near Atlantic Wood Industries on the Southern Branch of the Elizabeth
River in Virginia.  The results showed that the creosote-contaminated
sediment source  contained more intermediate weight (three aromatic
rings) and high molecular weight (more than three aromatic rings) PAHs,
in contrast to the water-soluble fraction, which contained high levels
of low molecular weight (less than three aromatic rings) and
heterocyclic compounds.  These differences were the result of
fractionation and degradation of creosote in the water soluble fraction.

EPRI, 1992

EPA’s toxicity characteristic (TC) rule declares three phenolic
isomers (o-, m-, and p-cresol) are highly toxic, and therefore,
regulated substances.  The Electric Power Research Institute (EPRI) used
the TCLP technique, a leaching method (EPRI, 1992), to determine the
concentrations of the three phenolic isomers that leached from treated
wood poles and crossarms when nearing disposal.  The EPA regulation
requires that the leachates should not contain >200 mg/L; otherwise, the
treated wood is classified as hazardous and can not be disposed of as
solid waste into landfills.  Fifty four samples from seventeen poles and
six crossarms, (ages 7 to 53 years) were chosen from the Northeast,
Mid-Atlantic, Southeast, Midwest, North Central and Western regions of
the United States.  Southern Pine, Douglas Fir, Western Red Cedar, and
Cedar woods were chosen for the study.  Total cresol concentration (all
three isomers) present in the leachates ranged from below the detection
limit (0.01 mg/mL) to 14.95 mg/mL, and the mean concentration was 1.63
mg/mL.  This was below the Agency’s toxic characteristic regulatory
level of 200 mg/mL. 

Gile, J.D. et al., 1982

Seven blocks of Ponderosa pine blocks (3.3 x 2.6 x14 cm)
pressure-treated with radiolabeled phenanthrene, acenaphthene, and
bis(tri-n-butyl oxide) were tested in a terrestrial microcosm chamber
(TMC-II).  This microcosm contained Williamette Valley topsoil,
ryegrass, invertebrates, and a gravid gray-tailed vole.  The
impregnation mixture contained dieldrin as a reference compound.  The
study was conducted for 2.5 months at which time it was found that 95
percent of the pesticides remained in the wood and most of the materials
that leached remained in the upper soil layer immediately surrounding
the pine blocks.  In the plants of the microcosm, 

0.7 ppm of dieldrin was detected while phenanthrene was detected at a
8.8 ppm.  Bioaccumulation in the invertebrates was variable and the
concentration of phenanthrene in the vole body was 7.2 ppm, while
acenaphthene was detected at 37.0 ppm.

Wendt, P.H., 1996

In a six-week two-phase field study, private, residential docks located
on ten tidal creeks at South Carolina’s Charleston Harbor Estuary
treated recently with CCA, but originally treated with creosote were
studied.  Samples were collected from sediments and oysters (Crassostrea
virginica) <1 meter, and >10 meters from the docks.  Reference samples
were also collected.  Mean concentrations of the 12 PAHs monitored from
the sediment were 978.3 µg/kg (dry wt.), 690.0 µg/kg (dry wt.), and
1183.8 µg/kg (dry wt.) for the <1 m, >10 m, and the reference samples,
respectively.  Mean PAH concentrations from the oysters were 3547.3
µg/kg (dry wt.), 2057.6 µg/kg (dry wt.), and 2173.1 µg/kg (dry wt.)
for the <1 m, >10 m, and reference samples, respectively.  The study
author reported that these concentrations were generally within the
range of values found at nearby marinas.  Most concentrations of PAHs in
whole sediments were generally below Long et al.’s (1995) “ER-L”
(Effects Range-Low) levels, suggesting that these values reported were
insufficient to cause any adverse biological/toxic effects. 

Bestari K.T. Jim et al., 1998a

 water were 7.3 μg/L (0.06 ppm  application) and 5,803.2 μg/L (109 ppm
application).  By Day 84, total PAHs remaining in the water were 0.80
μg/L and 13.9 μg/L from the 0.06 ppm and 109 ppm applications,
respectively.  In sediments, total PAHs ranged from 0.91 μg/g to 63.9
μg/g at Day 28, then declining thereafter.  A similar trend of
declining concentrations was shown by the PAHs on the PVC strips.  A
mass balance calculation reported a loss of 88.3 percent of PAHs from
the  microcosms after one month.  Based on the total PAH concentrations,
the half-life of most PAHs in water was reported to be approximately one
week.

Bestari, K.T. Jim et al., 1998b

Marine-grade Douglas fir pilings (15 to 20 cm diameter and 1.2 m length)
were pressure-treated with creosote using the same concentrations (0.06
ppm to 109 ppm) as in Bestari K.T. Jim et al. (1998a) in a similar
simulated microcosm.  This microcosm contained sediment, rooted and
floating macrophytes, and fish and invertebrate communities which
consisted of phytoplankton, zooplankton, and benthos.  The total organic
content was 5.1 percent in water that had been  circulated in the
microcosm through a holding tank for four weeks to maximize the chemical
and biological compositions.  The pressure-treated pilings were
suspended vertically in each microcosm in such a way that they were just
above the water surface and not touching the sediment.

 7 after the treatment (7.3 μg/L to 97.2 μg/L) and then declined to
concentrations close to the background (0.80 to 6.7 μg/L) by the end of
the study (Day 84).  Total PAHs from the leachates did bind to the PVC
liner, concentrations on Day 31 ranged from 0.3 to 2.4 μg/cm2, and
ranged from 0.2 to 2.2 μg/cm2 58 days after the treatment.  The rate of
loss of creosote from the pilings was 50 μg/cm2/day.  The study
suggested that the rapid loss of creosote was primarily due to
degradation processes like photolysis and microbial decomposition, and
partial adsorption to PVC liners.  The PAHs identified in this and
previous experiments were: naphthalene, acenaphthene, fluorene,
phenanthrene, anthracene, fluoranthene, pyrene, benz[a]anthracene,
chrysene, benzo[b]pyrene, benzo[k]fluoranthene, benzo[a]pyrene,
benzo[ghi]perylene, indeno[1,2,3-cd]pyrene and dibenz[a,h]anthracene. 
These are fifteen of the sixteen PAHs that EPA has recognized as EPA
Priority Pollutants.

Wan, M.T., 1994

In this study, runoff water from pressure-treated utility and telephone
wood poles was collected from fourteen utility and six railway ditches. 
The utility and telephone poles were initially pressure-treated with
pentachlorophenol and later in situ treated with a mixture of creosote
and chlorophenols.  The treatment was typically at the base up to 0.5 m
above and below soil level, and these applications were either used as
wrappings or painted with a creosote/chlorophenol mixture.  The runoff
water was analyzed for the presence of 15 PAHs.  These right of way
(ROW) ditches which were sampled flow into salmon streams in the Lower
Mainland and Vancouver Island of British Columbia, Canada.  Ditches of
parklands, farmlands, and railway ROWs were also sampled to establish
background and reference PAH concentrations.  PAHs were not detected in
the parkland ditches; however, they were found in farmlands and in
utility and railway ROW ditches.  The data from this study showed that
the maximum concentrations of the PAHs are in the wood and the PAHs that
leach out of the wood were mostly present around the base of the poles. 
Tables 6 through 8 summarize the presence/absence of the PAHs in various
environmental compartments.

Table 6.  PAHs in Treated Wood

Compound	

Poles

 (mg/kg)	

Railway Ties (mg/kg)



Acenaphthene	

3,566	

1,238



Acenaphthylene	

244	

77



Anthracene	

17,686	

5,100



Benz[a]anthracene	

1,938	

465



Benz[b+k]fluoranthene	

811	

254



Benz[a]pyrene	

1,116	

461



Benz[ghi]perylene	

69	

37



Chrysene	

1,093	

383



Dibenz[a,h]anthracene	

7	

16



Fluoranthene	

7,832	

1,880



Fluorene	

5,313	

1,141



Indeno[1,2,3-cd]pyrene	

36	

40



Naphthalene	

1,514	

342



Phenanthrene	

15,378	

3,173



Pyrene	

5,289	

1,487



Total PAHs	

61,182	

16,094

Note:  These PAHs were extracted from a sample of wood chips/scraps
which were collected and made into a composite mixture of 25 poles or 25
railway ties.

Table 7.  PAHs in Ditch Water of Parklands, Farmlands and Railway
Rights-of-Way  in the Lower Mainland of British Columbia

Compound	

Parkland Ditches (μg/L)	

Farm Ditches (μg/L)	

Railway Ditches (w/poles, μg/L)	

Railway Ditches (w/o poles, μg/L)



Acenaphthene	

ND	

1.04	

206	

0.57



Acenaphthylene	

ND	

1.72	

5.5	

1



Anthracene	

ND	

0.16	

81	

01.3



Benz[a]anthracene	

ND	

0.10	

195	

0.12



Benz[b+k]fluoranth-ene	

ND	

0.19	

144	

0.19



Bnez[a]pyrene	

ND	

0.10	

43	

0.10



Benz[g,h,i]perylene	

ND	

0.19	

12.2	

0.14



Chrysene	

ND	

0.10	

228	

0.17



Dibenz[a,h]anthrace-ne	

ND	

0.26	

4.1	

0.10



Fluoranthene	

ND	

0.30	

2035	

0.26



Fluorene	

ND	

0.30	

116	

0.22



Indeno[1,2,3-cd]pyrene	

ND	

0.16	

17.6	

0.15



Naphthalene	

ND	

0.35	

8.5	

0.19



Phenanthrene	

ND	

0.40	

1027	

0.44



Pyrene	

ND	

0.19	

1233	

0.19



Total PAHs	

ND	

5.56	

5,356.3	

3.97

Note:  In all cases the number of sample sites was never more than two. 

ND = Not detected.

Table 8.    PAH Concentrations in Utility ROW Sediments in the Lower
Mainland of British Columbia

Compound	

Concentration (µg/L)

	

Base of Pole	

Ditches 4 m

Upstream of Pole	

Ditches Adjacent to Pole (0.1-0.3 m)	

Ditches 4 m Downstream of Pole



Acenaphthene	

221	

0.11	

1.03	

ND



Acenaphthylene	

36	

ND	

1.37	

ND



Anthracene	

706	

0.16	

2.12	

0.09



Benz[a]anthracene	

93	

0.05	

0.46	

0.25



Benz[b+k] fluoranthene	

71	

ND	

0.65	

0.73



Benz[a]pyrene	

67	

0.11	

0.41	

0.64



Benz[g,h,i]perylene	

73	

ND	

0.43	

0.20



Chrysene	

92	

0.11	

0.41	

0.17



Dibenz[a,h]anthracene	

91	

ND	

0.20	

0.27



Fluoranthene	

211	

0.19	

1.24	

0.24



Fluorene	

469	

0.10	

1.58	

ND



Indeno[1,2,3-cd]pyrene	

75	

0.25	

0.71	

0.34



Naphthalene	

34	

0.12	

0.08	

ND



Phenanthrene	

666	

0.07	

3.29	

0.12



Pyrene	

171	

0.06	

1	

0.22



Total PAHs	

3076	

1.33	

15	

3.27

Note:  Sampling size varied from 5 to a maximum of 8 samples.

ND = Not detected

Middaugh, D.P. et al., 1991

An Agency sponsored study was conducted on the leaching of creosote
components from an abandoned American Creosote Works Site, a freshwater
stream that flows into the Florida Pensacola Bay.  This site also
utilized pentachlorophenol (PCP) and chromated copper arsenate (CCA). 
Adjacent to the site, a well was dug to a depth of 21 meters and the
ground water was analyzed.  The PAH concentrations in groundwater were
reported as follows: phenanthrene (32.8 mg/kg), naphthalene (28.8mg/kg),
fluoranthene (16.1 mg/kg), acenaphthene (13.6 mg/kg), fluorene(11.6
mg/kg), pyrene (10.4 mg/kg), 1-methyl naphthalene(9.5 mg/kg), and
2-methylanthracene (5.2 mg/kg).  The concentrations of other PAHs
detected were either less than 5 percent of the ones noted above or were
very low.

V Bioaccumulation

Stegman and Teal, 1973 and Varsani, et al., 1978

Studies on bioaccumulation showed that in an aquatic habitat, organisms
such as fish, shellfish, and crustaceans readily accumulate PAHs from
the environment and store them at a high level in their tissues.

Southworth, G.R., etal., 1978

A study was designed to investigate the bioaccumulation of seven
selected PAHs in Daphnia pulex.  The PAHs selected were: naphthalene,
anthracene, phenanthrene, pyrene, 9-methylanthracene, benz[a]anthracene,
and perylene.  Bioaccumulation corresponded to the  octanol/water
partition coefficients (KOW).  The benz[a]anthracene bioaccumulation
factor was 10,000 fold higher and about 100 fold higher for naphthalene.
 

A bioaccumulation study on clams (Rangia cuneate) was also conducted
after a creosote spill into Bayou Bonfuca at the American Creosote Works
Plant site at Slidell, Louisiana.  Results showed that the levels of
PAHs increased gradually for two weeks and increased dramatically by
week four at the monitoring station closest to the spill site.  The PAH
results are shown in Table 9.  

Another station (control station) in the same study showed evidence of
depuration after two weeks and equilibration after four weeks.  The
concentration of PAHs in water was very low at the site close to the
spill (13 ppb in two weeks and 26 ppb after four weeks).  Among the
PAHs, benzopyrenes were detected at a very high level of 600 ppb at the
station closest to the spill site. Other studies (Politzer, 1985; Neff,
1976) also showed that depuration of PAHs in bivalves vary from a couple
of weeks to a few months.  These studies support the possibility of
bioavailability of PAHs and contamination of the food chain.

Table 9.  Amounts of PAHs Detected in Clams at the Closest Station to
the Spill Site

PAH Component	

Week 0 - Pre-exposure (ppb)	

Week 2 

(ppb)	

Week 4

(ppb)



Naphthalene	

43	

60	

120



Biphenyl/Acenapthylene	

17	

13	

42



Fluorene	

7	

5	

11



Phenanthrene/ Anthracene	

34	

10	

28



Fluoranthene/Pyrene	

120	

88	

130



Benz[a]anthracene/Chrysene	

41	

81	

190



Benzopyrenes	

87	

132	

600



Spacie, A., et al., 1983

Bluegill sunfish were used to investigate the uptake (bioaccumulation),
biotransformation, and depuration rates of anthracene and
benzo[a]pyrene.  The uptake half-life for anthracene was 0.019 hours and
did not appear to be affected by the exposure concentration and humics. 
The uptake half-life of benzo[a]pyrene was also similar, but was
affected by the presence of humics.  Biotransformation for anthracene
was determined to be 0.22 nmol/hr while for benzo[a]pyrene  varied from
0.044 nmol/hr to 0.088 nmol/hr, between 1 and 2 hours of exposure.  The
depuration half-life for anthracene was 17 hours and 67 hours for
benzo[a]pyrene.  Due to the  biotransformation, the bioconcentrations
for both anthracene and benzo[a]pyrene were lower than predicted from
the Kow.

Tay, K.L., et al., 1992

A detailed bioassessment study was conducted in Canada on the Halifax
Harbor Sediment. Bioaccumulation of PAHs and other organic and inorganic
contaminants were investigated on the bivalve mollusk (Macoma balthica).
 The exposure of the species to the contaminated sediment was 30 days. 
The study authors suggested that the duration of 30 days may have been
too short for uptake of the PAHs into the Macoma system.  Organic
content in the Tuft’s Cove sediment was high and that may have
prevented the bioaccumulation process.  Table 10 summarizes the results.

Table 10.  Bioaccumulation of PAHs in Macoma balthica

Station	

PAHs in Tissues

(mg/kg wet wt.)



Walton (Control sediment)	

0.11



Drakes Gut 1 (Reference sediment)	

0.13



Tuft’s Cove (Contaminated sediment)	

ND



Original tissues (not exposed to test sediment)	

0.01



J.F. Elder and P. Dresler, 1988

A study on bioaccumulation was conducted on mollusk species at the
Pensacola Bay, 500 m from the creosote wood-preserving facility in
Pensacola, Florida.  Only four PAH compounds (phenanthrene,
fluoranthene, pyrene, and naphthalene) were found in the mollusk species
studied.  Depuration rates were not reported.  The same study also
estimated the PAH presence in water and the surface layer of estuarine
sediments.  The study, conducted on seven sites, found that the
sediments of the drainage streams were heavily contaminated with PAHs. 
The analysis of sediments at the sites showed very little contamination
with PAHs except at one site (Site 4).  This site was closest to the
wood-preserving facility.  The same study also reported that the
bioaccumulation of fluoranthene, pyrene and phenanthrene in both species
of mollusks was ten times higher at the test site than at the control
site.  Table 11 summarizes the results.

Most of the PAHs were insoluble in water and the solubilities were
inversely related to the molecular weights of the polyaromatics. 
Bioaccumulation of a pesticide depended on many external factors such as
resistance to biodegradation, chemical degradation, photolysis, tendency
for migration, and bioavailability.  The data on bioaccumulation of PAHs
are not extensive; however, a few generalities emerged from the existing
data.  First, molecules with lower molar masses had a tendency to
bioaccumulate more than high molar mass substances.  Second,
bioaccumulation was also dependent on the concentrations of a substance.
 For example, between naphthalene (two fused-ring compound) and
phenanthrene (three fused-ring compound), it is the later which was
found to bioaccumulate more than naphthalene because in the original
mixture of PAHs, phenanthrene was 20 percent of the mixture while
naphthalene was 3 percent.

Table 11.  Concentrations of PAHs in the Sediment and Discharge Stream
at Pensacola Bay

PAH	

Stream Site 3 (g/kg)	

Stream Site 2 (g/kg)	

Pensacola Bay

Site 4 (g/kg)	

Pensacola Bay/ Other Sites (g/kg)



Naphthalene	

300	

200	

ND	

ND



Phenanthrene	

ND	

12000	

ND	

ND



Fluoranthene	

62000	

17000	

190	

ND



Pyrene	

32000	

11000	

160	

ND



Benzoanthracene	

15000	

5000	

75	

ND



Chrysene	

10000	

7000	

100	

ND



Acenaphthene	

19000	

5000	

ND	

ND



Fluorene	

32000	

3000	

ND	

ND



Anthracene	

140000	

3000	

ND	

ND

ND - Not detected.  The limit of detection was 40 g/kg.

Bruner, K.A. et al., 1994

To measure PAH bioconcentration factors, a study was conducted on
pre-spawning (high lipid) and post-spawning (low lipid) mussel
populations of zebra mussel in the Great Lakes.  Pre-spawning mussels
had greater bioconcentration factors and a faster rate of accumulation
for benzo[a]pyrene than post-spawning mussels.  Lipid content, however,
did not influence bioconcentration factors or the rates of accumulation.
 Rates of depuration were not influenced by either factor (high or low
lipid contents).

Hellou, J.G., et al., 1990

A study was conducted on the bioaccumulation of PAHs in marine mammals. 
Four species of seals and six species of whales from the waters around
Newfoundland and Labrador were utilized as test subjects.  Accumulated
values, when expressed in terms of chrysene-equivalents were 0.10 to
1.21 ppm in the muscles of these mammals on a dry weight basis.

Howard, P.H. et al., 1991

It was suggested, based on theoretical calculations and modeling, that
the half-lives of the PAHs can be estimated in air, water, soil and
sediments.  This data are presented in Table 12.  From the table, one
can arbitrarily divide the PAHs into three groups: PAHs with two
aromatic fused rings, with three aromatic fused rings, and 4-5 aromatic
fused rings.  The study authors came to the following conclusions: 1)
The half-lives of the PAHs in these environmental compartments increased
as the complexity of the molecules increased.  Generally, half-lives of
2 aromatic fused rings < 3 aromatic fused rings < 4/5 aromatic fused
rings; 2) Kow values also show a similar trend: three sets of Kows were
observed.  The Kow values ranged from 3 to 4, from 4 to 5, and 6 and
above.  As noted for the half-lives, the Log Kow increased as the
complexity of the molecule increased;  3) In general, the half-lives in
air and water environmental compartments were lower than in
soils/sediments, since the adsorption constant in these two compartments
were larger than in air and water media.  Those molecules with a longer
half-life also exhibited persistence in that environmental compartment. 
PAHs were more persistence in soils/sediments than in other
environmental compartments; and 4) The 4/5 aromatic-fused ring molecules
were persistent and because they also had high Kows, they were also
bioaccumulative in the organisms present in the soil/sediments. 

Table 12.  Estimated/Modeled PAH Half-lives in Air, Water, Soil and
Sediment

Compound	

Air/

Class	

Half- live	

Water/Class	

Half-life	

Soil/Class	

Half-life	

Sediment/Class	

Half-life	

Kow



Indan	

2	

1 day	

4	

1 wk.	

6	

2 mos.	

7	

8 mos.	

3.33



Naphthalene	

2	

1 day	

4	

1 wk.	

6	

2 mos.	

7	

8 mos.	

3.37



1-methyl naphthalene	

2	

1 day	

4	

1 wk.	

6	

2 mos.	

7	

8 mos.	

3.87



Acenaphthalene	

3	

2 days	

5	

3 wks.	

7	

8 mos.	

8	

2 yrs.	

3.92



Fluorene	

3	

2 days	

5	

3 wks.	

7	

8 mos.	

8	

2 yrs.	

4.12



Phenathrene	

3	

2 days	

5	

3 wks.	

7	

8 mos.	

8	

2 yrs.	

4.57



Anthracene	

3	

2 days	

5	

3 wks.	

7	

8 mos.	

8	

2 yrs.	

4.54



Pyrene	

4	

1 wk.	

6	

2 mos.	

8	

2 yrs.	

9	

~ 6 yrs.	

5.18



Fluoranthene	

4	

1 wk.	

6	

2 mos.	

8	

2 yrs.	

9	

~ 6 yrs.	

5.22



Chrysene	

4	

1 wk.	

6	

2 mos.	

8	

2 yrs.	

9	

~ 6 yrs.	

1.65



Benz[a]anthracene	

4	

1 wk.	

6	

2 mos.	

8	

2 yrs.	

9	

~ 6 yrs.	

5.91



Benzo[k]flouranthene	

4	

1 wk.	

6	

2 mos.	

8	

2 yrs.	

9	

~ 6 yrs.	

6.00



Benzo[a]pyrene	

4	

1 wk.	

6	

2 mos.	

8	

2 yrs.	

9	

~ 6 yrs.	

6.04



VI Migration of PAHs From Poles to Soils

Mississippi State University, 1981

Fifty-six soil samples were collected radially around fourteen
creosote-treated poles and analyzed for migration of creosote components
into the surrounding soils.  The concentrations of naphthalene,
2-methylnaphthalene, 1-methylnapthalene, and biphenyl varied from 25 to
50 ηg/g. Acenaphthylene, acenaphthene, dibenzofuran, fluorene,
phenanthrene, and anthracene were present at lower concentrations than
in the original mixture of creosote.  The author suggested that
vaporization, water solubility, and biological degradation might be
contributing factors for reduction in the concentrations of these
polycyclic components of creosote.  Fluoranthene, pyrene, carbazole,
1,2-benzanthracene, and chrysene were persistent in the soil around the
poles and the concentration of all the components decreased as the
distance from the poles increased. 

McGroddy, S.E. and J.W. Farrington, no date available

In a study on sediment porewater partitioning of PAHs in Boston Harbor,
it was shown that the PAH concentrations measured in sediments and
porewaters from three cores were notably lower than the amounts of the
PAHs predicted by two- and three-phase equilibrium partitioning models. 
The study author suggested lower amounts might have been available for
partitioning in sediment porewaters.

K. Brooks’  Study, 2004 :

 “Polycyclic Aromatic Hydrocarbon Migration From Creosote-Treated
Railway Ties into Ballast and Adjacent Wetlands.

*A simulated mesocosm on a wetland was created in Will County, Illinois.


*Newly Treated, untreated and weathered-treated railroad ties were
placed in this simulated wetland.

*The study was conducted for two years.

*Samples were  analyzed  from ballast ( surface on which ties are
placed),  wetland sediments, groundwater, stormwater, and soil cores.

*The detection and analysis of the PAHs was done:  by examining  direct
contamination from surface stormwater, PAHs infiltration into shallow
ground water, and lateral distance from the railroad ties (adjacent
wetland soils). 

*PAHs were detected up to 60 cm  vertically down  into the ballast.

*Mesocosm soil type was: Romeo Silty clay loam.

*Baseline PAH concentrations were determined using EPA’s Method 8310.
The samples were extracted by Soxhlet extraction method.

*p-Terphenyl was used as a surrogate for PAH determinations.

*Baseline PAH levels in the surface soils on this simulated wetland was
measured at 0.00 to 0.17 µg/g.

* Retentions for new  railway ties varied from 28 to 57.28 kg/m3 and for
the old railway ties it varied from 39.36 to 55.52 kg/m3 Each number
represents an average of 12 borings.

*13 stormwater samples which included newly treated, untreated and
weathered ties were analyzed using EPA’s Method 2003a( GC/MS
technique).

*229 samples collected form the surface soils of the wetland at a
distance of 0, 0.25, 0.50 and 0.75 meters were collected and analyzed
using EPA’s Method 2003b(HPLC technique)

*174 samples of ballast (rock) from around (east/west) the railway ties
at the distance of 5, 20 and 30 cm were collected and analyzed using
modified EPA method 8310 (HPLC technique).

* 36 core samples of  ballast rock and  11 sediments) collected at a
distance 10 cm form each other (up to 60 cm)  were collected and
analyzed  using EPA’s method 8310.

*LOD of the analytical methods used  was set between 0.20 to 0.100
µg/g.

* First sample analysis was done 10 days after the construction of the
simulated wetland.

Thereafter the sample analysis and data collections were done at 3, 6,
9, 12 , 15 and 18 months. 

*The  measured concentrations of  PAHs form new, weathered  generally
declined progressively from  5 cm to 30 cm distances for all samples
taken after 10 days, 3, 6, 9, 12, 15, and 18 months.

* Similarly, core samples at 10-cm depth increment  up to 60 cm down,
showed that the concentration decreased from 1.207 µg/g  at 0 cm to
0.482 µg/g at 60 cm depth.

* From stormwater samples collected after 18 months, two PAHs were
detected from newly treated railway ties; Benzo(a)anthracene (0.00019
mg/L) and phenanthrene (0.00066 mg/L). This was observed in only one out
of 16 samples that were collected.

*Fluoranthene, phenanthrene and pyrene were detected from weathered
railway ties.

*Because in this particular simulated site the ballast did not have
organic carbon and therefore the likelihood of microbes in such an
environment were not likely to survive,. no biodegradation study was
conducted.

Conclusions:  This is a well designed study conducted for about two
years. Sample size is also large. However, this mesocosm could not yield
a biodegradation of PAH data.

It should be noted that this study is not a real life wetland and no
rails run on these railroad ties. Thus we do not know how the
concentration of  PAH leaching out, and migration would be impacted .

We can ask for a mesocosm study.

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 PAGE   

 PAGE   1 

 PAGE  6 

 PAGE  25 

Table 3.  Photolysis of Selected PAHs Found in P1/P13 and P2 Fractions
of Coal Tar Creosote (Cont’d)

Table 5.  Aquatic Soil Metabolism (Biodegradation) Results for
Creosote/PCP Contaminants  In Groundwater Samples From Pensacola,
Florida Superfund Site (Cont’d)

