1

 OCS Study MMS 2004-001

Final Report

Phenanthrene Adsorption and Desorption by Melanoidins

and Marine Sediment Humic Acids

by

John A. Terschak

Susan M. Henrichs*

Principal Investigator

David G. Shaw

Institute of Marine Science

University of Alaska Fairbanks

Fairbanks, AK 99775-7220

January 2004

*Corresponding author

e-mail: henrichs@ims.uaf.edu
2

Contact information

e-mail: cmi@sfos.uaf.edu

phone: 907.474.7707

fax: 907.474.7204

postal: Coastal Marine Institute

School of Fisheries and Ocean Sciences

University of Alaska Fairbanks

Fairbanks, AK 99775-7220
iii

Table of Contents

List of Tables ............................................................................................................................ v

List of Figures ......................................................................................................................... vi

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

Chapter 1. Introduction ............................................................................................................ 1

Background ........................................................................................................................ 1

Site Descriptions ................................................................................................................ 5

Beaufort Sea ................................................................................................................ 5

Lower Cook Inlet ......................................................................................................... 7

Port Valdez .................................................................................................................. 9

Summary .......................................................................................................................... 10

Acronyms and Abbreviations ........................................................................................... 11

Chapter 2. Effects of Humic Acid Properties on Phenanthrene Adsorption

J.A. Terschak, S.M. Henrichs and D.G. Shaw .............................................................. 13

Abstract ............................................................................................................................ 13

Introduction ...................................................................................................................... 13

Experimental Section ....................................................................................................... 14

Sampling sites and sample collection ......................................................................... 14

Sediment organic matter and humic acid characterization .......................................... 15

Sediment organic carbon and nitrogen determination by combustion

and isotope ratio mass spectrometry ........................................................................ 15

Sediment hydrocarbon analyses by extraction and gas chromatography ..................... 16

Humic acid extraction ................................................................................................ 16

Synthetic humic acid preparation ............................................................................... 17

UV/Visible spectroscopy ........................................................................................... 17

FTIR spectra acquisition ............................................................................................ 17

CPMAS/13CNMR spectra acquisition ........................................................................ 17

Adsorption experiments ............................................................................................. 17

Statistical treatments .................................................................................................. 18

Results ............................................................................................................................. 19

Humic acid properties ................................................................................................ 19

Adsorption measurements .......................................................................................... 21

Discussion ........................................................................................................................ 25
iv

Chapter 3. Phenanthrene Adsorption to Mineral-Bound Humic Acid: Kinetics

and Influence of Previous Phenanthrene Adsorption

J.A. Terschak and S.M. Henrichs ............................................................................... 29

Abstract ............................................................................................................................ 29

Introduction ...................................................................................................................... 29

Experimental Section ....................................................................................................... 31

Substrate preparation ................................................................................................. 31

Humic acid sources .................................................................................................... 31

Substrate coating ........................................................................................................ 32

Total organic carbon measurements ........................................................................... 32

Adsorption experiments ............................................................................................. 32

Results ............................................................................................................................. 34

Discussion ........................................................................................................................ 40

Chapter 4. Observed Desorption Kinetics of Phenanthrene from Mineral-Bound

Humic Acids: Consequences of Conformational Changes

J.A. Terschak and S.M. Henrichs ............................................................................... 43

Abstract ............................................................................................................................ 43

Introduction ...................................................................................................................... 43

Experimental Section ....................................................................................................... 44

Substrate preparation ................................................................................................. 44

Humic acid sources .................................................................................................... 45

Substrate coating ........................................................................................................ 45

Total organic carbon measurements ........................................................................... 45

Desorption experiments ............................................................................................. 45

Results ............................................................................................................................. 46

Discussion ........................................................................................................................ 51

Chapter 5. Conclusion ............................................................................................................ 55

Acknowledgments .................................................................................................................. 56

Study Products ........................................................................................................................ 57

References .............................................................................................................................. 57
v

List of Tables

Table 1. Sample sites – Beaufort Sea region ............................................................................................ 6

Table 2. Sample sites – Lower Cook Inlet region .................................................................................... 8

Table 3. Sample sites – Port Valdez region ........................................................................................... 10

Table 4. Site locations of sediments discussed in this study ................................................................. 15

Table 5. The measured properties of sediment organic matter, natural humic acids,

and synthetic humic acids evaluated in this study and the methods by which

they were obtained .................................................................................................................... 16

Table 6. N/C ratios and carbon and nitrogen stable isotope values versus PDB

and air, respectively .................................................................................................................. 26

Table 7. Site locations of sediments discussed in this study ................................................................. 31

Table 8a. Initial rates of phenanthrene adsorption at various concentrations based on

1 hour initial time increments ................................................................................................... 37

Table 8b. Initial rates of phenanthrene adsorption at various concentrations based on

1 day initial time increments .................................................................................................... 37

Table 9. Initial rates of phenanthrene desorption after 1 day of adsorption at various

concentrations based on 1 day time increments ...................................................................... 50
vi

List of Figures

Figure 1. Sample regions within Alaska .................................................................................................... 5

Figure 2. Sample sites within the Beaufort Sea region ............................................................................. 6

Figure 3. Cook Inlet .................................................................................................................................... 7

Figure 4. Sample sites within the Lower Cook Inlet region ..................................................................... 8

Figure 5. Sample sites within the Port Valdez region ............................................................................... 9

Figure 6. Cross-polarized magic angle spinning/13C nuclear resonance spectroscopy .......................... 20

Figure 7. Increasing nitrogen to carbon ratios of humic acid were associated with an

increase in the fraction of carbon atoms associated with oxygen and nitrogen

bonds in natural and synthetic humic acids ............................................................................. 20

Figure 8. Fourier transform infrared spectroscopy .................................................................................. 21

Figure 9a. Partition coefficients (KOC) and percent aromaticity of humic acids were

not significantly correlated ....................................................................................................... 22

Figure 9b. Partition coefficients (KOC) and percent aliphaticity of humic acids had

no significant correlation .......................................................................................................... 23

Figure 9c. Partition coefficients (KOC) increased with the fraction of nonpolar carbons ........................ 23

Figure 10a. Partition coefficients (KOC) decreased with increasing nitrogen to carbon

(N/C) ratios of humic acid ........................................................................................................ 24

Figure 10b. Partition coefficients (KOC) decreased as the number of carbons in O and N

bonds (fraction carboxyl and amide C) increased ................................................................... 24

Figure 11. Montmorillonite particle size distribution ............................................................................... 34

Figure 12. Linear adsorption of phenanthrene to pristine sediments

(without Port Valdez samples) ................................................................................................. 35

Figure 13. Phenanthrene adsorption to pristine sediments and to PAH-loaded sediments ..................... 36

Figure 14. Initial rate of phenanthrene adsorption as a function of initial phenanthrene

concentration to pristine sediment (without Port Valdez samples) ........................................ 36

Figure 15. Comparison of initial rates of phenanthrene adsorption to pristine sediments

and to PAH-loaded sediments .................................................................................................. 38

Figure 16a. Partition coefficients for phenanthrene adsorption to JB2 ...................................................... 39

Figure 16b. Partition coefficients for phenanthrene adsorption to 1:10–48 ............................................... 39

Figure 17a. Desorption plot for a synthetic humic acid (10:1–48) ............................................................. 47

Figure 17b. Desorption plot for a natural humic acid (TB2) ...................................................................... 48

Figure 18a. Extent of desorption for a synthetic humic acid (10:1–48) ..................................................... 48

Figure 18b. Extent of desorption for a natural humic acid (TB2) .............................................................. 49

Figure 19. Initial desorption rates after 1 day of desorption time ............................................................. 50
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Abstract

Sediments are major reservoirs of persistent petroleum contamination in marine environments. Petroleum

hydrocarbons associate with the sediment organic matter, of which humic acids are an important

constituent. This study examined the role that humic acids and their structures play in the kinetics and

mechanisms of polycyclic aromatic hydrocarbon (PAH) interactions with sediments. Natural humic acids,

with a wide range of properties, were isolated from Alaska coastal marine sediments. Melanoidins were

synthesized and used as humic acid analogs. The humic acids were characterized by elemental and

isotopic analyses, Fourier transform infrared spectroscopy, and cross-polarized magic angle spinning 13C

nuclear magnetic resonance spectroscopy. The humic acids were coated onto a standard montmorillonite

clay, and the adsorption and desorption of phenanthrene was measured using a radiotracer.

Adsorption required about one week to reach steady state, indicative of slow diffusion of PAH within the

humic acid. The composition of the humic acids had a greater effect on phenanthrene adsorption than

their concentrations on the clay. Organic carbon normalized adsorption partition coefficients were closely

correlated with the sum of amide and carboxylic carbons, a measure of the polarity of the humic acids, but

were independent of initial phenanthrene concentration, indicating that the binding sites were unlimited

and uniform in strength. This explains the fact that the initial adsorbed concentration of phenanthrene had

no effect on subsequent phenanthrene adsorption.

Desorption of phenanthrene was not related to any of the humic acid structural characteristics measured.

The initial desorption rate was linearly related to the initial adsorbed concentration, as expected for a

diffusive process, and was negatively correlated with the carbon content of the humic acid coated clay.

Under most conditions, desorption was complete after one to seven days; there was little evidence for

irreversible adsorption.

Because of the substantial variability of adsorption and desorption behavior with organic matter

characteristics, interactions of aromatic hydrocarbons with marine sediments cannot be predicted based

on total organic matter concentration alone. Information on aspects of organic matter composition is

needed in order to make accurate predictions.

Chapter 1. Introduction

Background

Hydrocarbon contamination of the coastal marine environment results from both acute and chronic

anthropogenic sources. Catastrophic accidents such as the Exxon Valdez oil spill are widely publicized,

while smaller spills, permitted discharges, and non-point sources, which in sum contribute a larger

quantity of petroleum to the environment, receive much less public scrutiny. The fate and effects of

petroleum are a complex function of both the chemical characteristics of the petroleum and the physical

and biological characteristics of the marine environment that is contaminated. Although polycyclic

aromatic hydrocarbons (PAH) (see p. 10 for a list of abbreviations and acronyms) make up only about

10% of petroleum, they are of special concern because, in addition to being resistant to biological and

chemical breakdown under many conditions, they are toxic, mutagenic, and carcinogenic [Black et al.

1983; White 1986; Pahlman and Pelkonen 1987]. In addition to petroleum, combustion is an important

and widespread source of PAH to the environment [Wakeham and Farrington 1980; Gschwend and Hites

1981].

PAH have very low solubilities in seawater [Shaw et al. 1989] and therefore tend to concentrate in

sediments rather than remaining in the water column. The adsorption of PAH onto particulate matter
2

is an important process affecting their movement and fate in the environment. For example, the observed

trapping of PAH and other hydrophobic pollutants in estuarine and coastal marine environments has been

commonly attributed to adsorption to sediments. Bates et al. [1987] and Murphy et al. [1988] examined

suspended sediment concentrations of PAH in a fjord-like estuary near Seattle, Washington. Bates et al.

[1987] concluded that vertical rather than horizontal transport dominated, and Murphy et al. [1998] found

that greater than 90% of sediment-bound PAH remained within the estuary. Although sedimentary

contamination decreases markedly with time after a spill, due to biodegradation and other processes,

aromatic hydrocarbons can persist for at least 20 years in contaminated sediments [Teal et al. 1992]. A

likely factor in this persistence is a decrease in biodegradability associated with adsorption [Guerin and

Boyd 1992; Luthy et al. 1997]. The effect of desorption on bioavailability has been noted in laboratory

studies, in that freshly added hydrophobic organic compounds (HOC) are metabolized, while they persist

in “aged” environmental samples, even though they contain the appropriate degrading microorganisms

[Braddock and Richter 1998].

Biological uptake of petroleum occurs mainly from solution, leading to biodegradation, bioaccumulation,

or toxicity [Landrum et al. 1985; Efroymson and Alexander 1995]. There is evidence that petroleum

associated with sediments can be a major source of polycyclic aromatic hydrocarbons and other

hydrophobic organic compound pollutants for marine and lacustrine organisms [reviewed in McElroy et

al. 1989], and exposure of benthic invertebrates can result in subsequent transfer to higher trophic levels

[Long et al. 1998; Thompson et al. 1999; Valette-Silver 1999; Krantzberg et al. 2000]. Harmful effects

of contaminated sediments have been found for polychaetes [Knauss and Hamdy 1991; Weston 1990;

Weston and Mayer 1998], bacteria [Moll and Mansfield 1991], amphipods [Daan et al. 1994], echinoids

[Daan et al. 1994], and bivalves [Knauss and Hamdy 1991; Daan et al. 1994], among other groups of

organisms. Strong binding of pollutants to sediments can result in lower bioavailability, and thus,

lessened environmental toxicity [Weissenfels et al. 1992]. On the other hand, weak and reversible binding

can result in greater dispersal and availability of pollutants, and therefore higher effective toxicity.

Many studies have found non-equilibrium (or rate-limited) adsorption of organic chemicals, including

aromatic hydrocarbons and hydroxy- and carboxy-substituted aromatic compounds, by sediments and

soils [Brusseau et al. 1991; Hatzinger and Alexander 1995]. Specifically, desorption is often slower than

adsorption and is biphasic, with the rate decreasing over time [Brusseau et al. 1991; Isaacson and Frink

1984]. Henrichs et al. [1997] reported that Lower Cook Inlet sediments strongly adsorbed phenanthrene,

naphthalene, and benzene, and that this adsorption was not completely reversible using their experimental

conditions. Their desorption experiments were carried out for a maximum of three days and used

water:particle concentration ratios equal to those in the adsorption experiments. They recognized that

adsorption could be reversible under more favorable conditions. For example, Means [1995] found that

pyrene adsorption by sediments was reversible, but he exhaustively eluted sediments with seawater,

which would allow even slow desorption of strongly adsorbed hydrocarbons to proceed to completion.

Observations of slow or apparently irreversible adsorption are most consistent with slow diffusion of the

hydrocarbons into a three-dimensional organic structure; reversal of this process would also be slow

[Brusseau et al. 1991; Isaacson and Frink 1984]. Although adsorption of aromatic hydrocarbons at several

types of sites (with varying partition coefficients) on the particle surface did not explain the observations

as well, this remains a possible interpretation [Henrichs et al. 1997].

Another observation which is not consistent with a simple, reversible adsorption process is that partition

coefficients (Kp) sometimes decrease with decreasing particle concentrations, an effect that could

be due to binding of hydrocarbons by dissolved or colloidal organic matter [Voice and Weber 1985].

Complexation of nonpolar organic compounds by dissolved or colloidal organic matter in soil solutions

or sediment pore waters increases their mobility [Brownawell and Farrington 1986; Chin and Gschwend

1992; Hebert et al. 1993]. Bacterial extracellular polymers are one category of soluble compounds that

have been shown to enhance pore water transport of phenanthrene [Dohse and Lion 1994].
3

Within sediments, PAH associate with organic matter and organic coatings of mineral particles of

low polarity [Luthy et al. 1997]. The adsorption of hydrophobic compounds by sediments and soils is

positively correlated with their organic matter content [Means et al. 1980; Schwarzenbach and Westall

1981; Shaw and Terschak 1998]. The extent of the association is known to vary depending on the

character of the organic matter present [Means et al. 1980; Garbarini and Lion 1986]. Adsorption

of aromatic hydrocarbons is positively correlated with the nonpolar-to-polar functional group ratio

[Garbarini and Lion 1986; Rutherford et al. 1992] and the proportion of the organic structure that is

composed of aromatic rings (aromaticity) [Gauthier et al. 1987]. This specific organic matter–PAH

association appears to result from weak dipole interactions between aromatic rings of the PAH and

similar rings which are part of many organic compounds, particularly those derived from lignins and

related materials of woody terrigenous plants [Chin et al. 1994]. Terrigenous organic matter has, on

average, greater aromaticity and lower oxygen content (polarity) than marine sediment organic matter,

which would tend to increase the affinity for aromatic hydrocarbons. Adsorption of phenanthrene on

natural soils is also related to the extent of diagenetic alteration of soil organic matter [Young and Weber

1995].

Humic acid is the fraction of naturally occurring dissolved, sedimentary, or soil organic material that is

soluble in aqueous base but insoluble in aqueous acid [Parsons 1988]. Humic acid associates strongly

with other organic molecules, including organic pollutants [Piccolo 1988; Karickhoff and Morris 1985;

Karickhoff et al. 1979]. Despite their similarity in solubility behavior, soil and marine sediment humic

acids differ in structure and reactivity [Sastre et al. 1994], reflecting the structural characteristics of

the original detrital organic material as well as the extent and pathways of biological and chemical

transformation. Terrestrial humic material includes structural components derived from lignin, a

component of all vascular plants, but not algae or other nonvascular plants. The sources of marine humic

acid are generally understood, although many of the specific transformation processes leading to these

materials are not. Organic material is released to the marine environment mainly by plankton, and by

macroalgae in some coastal areas. Once in the environment, most of these materials are consumed as food

by detritivores and bacteria [Fenchel and Blackburn 1979]. However, it appears that a fraction undergoes

chemical condensation reactions [Tissot and Welte 1978] in which the reactive portions of the molecules

combine, usually without enzymatic control. This gradually leads to large organic molecules with few

reactive sites and low biodegradability. An alternate hypothesis on the origin of humic substances is that

they represent a largely unmodified residue of the decomposition process, composed of certain natural

products of organisms that are inherently resistant to decay [DeLeeuw and Largeau 1993].

Lignin degradation products can react with proteins and amino acids to form humic substances [Flaig

1964], and evidence suggests that natural humic substances can also be produced by condensation

reactions between reducing sugars and free amino acids [Maillard 1913; Enders and Theis 1938; Hodge

1953; Abelson and Hare 1971]. The evidence includes chemical similarities between natural humic

substances and sugar–amino acid condensation products (melanoidins) produced in the laboratory

[Hoering 1973; Ertel and Hedges 1983]. Also, in marine environments, carbohydrates, because of their

much greater abundance, are more likely precursors of humic substances than lignin [Nissenbaum and

Kaplan 1972; Hedges and Parker 1976].

Ishiwatari [1985] found that artificial kerogen, a highly refractory product of further condensation and

other transformations of humic substances, formed rapidly in a reaction between glucose and casein.

Yamamoto and Ishiwatari [1989] hypothesized that proteins are the more probable reactants with

carbohydrates than free amino acids in the natural environment. Their kinetic study indicated that kerogen

could form in 11 years at 20°C and 140 years at 5 °C, not unrealistic for the natural environment. Other

investigations show substantial similarities between humic substances synthesized from a variety of

starting materials and naturally occurring humic substances [Haider et al. 1975; Martin and Haider 1980;

Gilam and Wilson 1985; Taguchi and Sampei 1986]. Using cross-polarized magic angle spinning 13C
4

nuclear magnetic resonance (CPMAS/13CNMR), Ikan et al. [1990] found that melanoidins synthesized

with higher initial proportions of sugar than amino acids yielded products with considerably greater

aromaticity compared with those prepared from higher proportions of amino acids.

Although there have been a number of investigations of the role of sediment organic matter in aromatic

hydrocarbon adsorption, there is reason to believe that these are not entirely applicable to Alaska

sediments. The terrestrial vegetation that supplies a substantial fraction of the organic matter to coastal

sediments is substantially different from that in temperate climates, especially on the North Slope of

Alaska. Decomposition processes, which affect the properties of sediment organic matter, are influenced

by lower temperatures and greater seasonality of organic matter inputs. The physically-weathered

lithogenic minerals of high-latitude sediments probably have different adsorptive properties from those

derived from chemical weathering regimes, potentially leading to differences in the type and quantity of

natural organic matter adsorption to the sediment surface, which is thought to be an important mechanism

of sediment organic matter preservation [Hedges and Keil 1995].

The following hypotheses were tested:

1. Aromatic hydrocarbon adsorption that was apparently irreversible in the experiments of

Henrichs et al. [1997] and Braddock and Richter [1998] is reversible with longer desorption

times. The kinetics of desorption largely controls the persistence of petroleum-derived

aromatic hydrocarbons in sediments.

2. Interactions of aromatic hydrocarbons with sediment organic matter are responsible for

adsorption that appears to be irreversible, at least under some conditions.

• Diffusion into the three-dimensional structure of organic matter slows desorption.

• Binding to a variety of adsorption sites with different partition coefficients causes

differences in the apparent, experimental partition coefficient under conditions of

adsorption and desorption.

3. Variations in organic matter properties influence the rate and extent of adsorption and

desorption processes. In particular, there is a direct correlation between the aromatic

character of sediment humic acid and the adsorption partition coefficient of phenanthrene.

• Model organic substances similar to those occurring in sediments, but with well-controlled

composition and properties, will show a systematic relationship between adsorption and

desorption properties and organic structure.

• The aspects of organic structure identified as important to adsorption properties using

model compounds will also show a systematic relationship to adsorption and desorption

of aromatic hydrocarbons by natural sediment organic matter.

4. Previous phenanthrene adsorption of sediments affects subsequent adsorption.

Both natural humic acids representing a wide range of organic matter properties and synthetic humic

acids (melanoidins) were used in the experiments. Melanoidins were used because of the ability to

control various aspects of their structural components during synthesis, such as the degree of aromatic

constituents, and their initial lack of PAH.

Included within this report are three manuscripts that address the major hypotheses of this investigation.

The first (Chapter 2), “Effects of Humic Acid Properties on Phenanthrene Adsorption”, looks closely at

PAH partition coefficients and how they are related to the molecular structures of organic coatings. The

second manuscript (Chapter 3), “Phenanthrene Adsorption to Mineral-Bound Humic Acid: Kinetics and

Influence of Previous Phenanthrene Adsorption”, reports on investigations of the processes of PAH
5

adsorption—more specifically, the rates and the effect initial sediment phenanthrene concentration has on

subsequent PAH adsorption. The final paper (Chapter 4), “Observed Desorption Kinetics of Phenanthrene

from Mineral-Bound Humic Acids: Consequences of Conformational Changes”, examines biphasic

patterns seen in the rates of PAH desorption and offers explanations for incomplete desorption.

Site Descriptions

Sediment samples were taken from three distinct regions within Alaska (Figure 1). These regions

included portions of the Beaufort Sea, Lower Cook Inlet, and Port Valdez. It was expected that

characteristics of the sediments and the humic acids associated with those sediments would vary due

to differences in physical and chemical weathering processes, terrestrial run-off, marine influences,

and anthropogenic inputs within the selected areas.

400 km

400 mi

170 160 150 140

60

60

70

180

70

Bering Sea Gulf of Alaska

Chukchi

Sea

Beaufort Sea

Cook Inlet

Shelikof Strait

Arctic Circle 

Figure 1. Sample regions within Alaska. The northernmost box is the Beaufort Sea

region, the southernmost box is the Lower Cook Inlet region, and the

easternmost box is the Port Valdez region.

Beaufort Sea

Offshore exploratory drilling in the Beaufort Sea began in 1973, with 15 exploratory wells drilled by the

end of 1981 [Norton and Weller 1984]. A total of 57 wells had been drilled in the Alaskan Beaufort Sea

by 1986 [Boehm et al. 1987]. The sample locations (Figure 2, Table 1) were originally selected for

a trace metal contaminant study under the Outer Continental Shelf Environmental Assessment Program

(OCSEAP) [Crecelius et al. 1991]. Naidu et al. [2001] also used these sites in a study of trace metal and

hydrocarbon contamination.
6

6D

6B

5A

WPB

2F

6G

7E

River

Colville

River

Harrison Bay

Camden Bay

Sagavanirktok

Beaufort Sea

152 N 150 N 148 N 146 N 144 N

71 N

70 N

Barter I.

Kuparuk

River

Canning

River

Figure 2. Sample sites within the Beaufort Sea region.

Table 1. Sample sites – Beaufort Sea region. Site denoted with an asterisk (*)

indicates sample used in the adsorption and desorption experiments.

Station Latitude (N) Longitude (W) Depth (m)

2F 70° 10.3′ 146° 02.0′ 1.9

5A* 70° 29.7′ 148° 46.0′ 11.4

6B 70° 33.4′ 150° 24.6′ 5.5

6D 70° 44.9′ 150° 28.5′ 18.4

6G 70° 31.3′ 149° 53.9′ 2.1

7E 70° 43.6′ 152° 04.4′ 3.3

WPB 70° 20.6′ 148° 23.2′ 2.5

The bordering terrestrial coastal plains consist of nearshore marine, fluvial, and aeolian deposits of midto late-Quaternary age [Black 1964]. The uppermost soils of the seasonal thaw layer, ranging from

0.5 to 2.5 m in thickness, are made up of silts and fine sands and often have high organic matter content,

including both finely divided material and larger plant fragments [Everett and Parkinson 1977]. Tundra

vegetation is low in species diversity. The lichens Collema fuscovirens (present nomenclature; C.

tunaeforme in the referenced report), Evernia perfragilis, and Fulgensia bracteata, and the vascular plants

Carex atrofusca, C. membranacea and C. misandra are common [Webber and Walker 1975]. Average air

temperatures at Barter Island range from 4.4°C in July to –28.6 °C in February [Brown et al. 1975]. The

ground is frozen to depths greater than 625 m [AEIDC 1975] and only surface thawing (15 cm to 1 m)

occurs in the summer. Nearshore bottom water averages –1.3°C and the sediment temperatures in waters

shallower than 2 m range between –2.2 °C and –4.6 °C [Reimnitz and Barnes 1974].

The input of sediments is highly seasonal. In spring the large quantity of water tied up in snow is released

over a short period. Approximately 80% of the total discharge into the Beaufort Sea occurs in June via the
7

Colville, Kuparuk, Sagavanirktok, and Canning rivers [Sharma 1979]. In summer, coastal bluffs suffer

thermoerosion due to melting permafrost and are a locally important source of terrigenous material for

coastal sediments. Sedimentation on the shelf is controlled by wave- and current-related processes. There

is a general westward movement of sediment resulting from the nearshore current [Sharma 1979] driven

by the prevailing easterly winds. Sea ice covers the shelf from mid-September through early July [Sharma

1979]. Ice-related sedimentary processes occur during the 8–10 months of shorefast ice in winter, but

sediment transport during ice cover is minimal [Sharma 1979]. The thickness of Recent sediments on

the shelf is generally <5 m [Reimnitz and Barnes 1974]. Surface sediments are generally poorly sorted

gravel, moderately sorted silt and sand, and well-sorted clays. Naidu and Mowatt [1974] found that the

fine clay fraction consists mostly of illite (56–69%), with smaller amounts of chlorite (14–30%) and

kaolinite (5–10%), and trace amounts of smectite (<9%).

Lower Cook Inlet

Offshore oil exploration and production in Cook Inlet (Figure 3) have occurred since the mid-1960s.

Since construction of the Trans Alaska Pipeline, tankers have carried Prudhoe Bay crude oil from Port

Valdez to refineries located near Kenai. In Lower Cook Inlet, merchant and fishing vessels are sources

of small but constant inputs of fuel and combustion products. Additional hydrocarbons come from erosion

of coastal and submarine coal outcrops and natural oil seeps [Shaw and Wiggs 1980]. The sampling

locations in Lower Cook Inlet are given in Figure 4 and Table 2. The samples were all collected from

the intertidal zone near mean low-low water.

West Foreland

Kenai/Soldotna

Ninilchik

Anchor Point

Homer

Nanwalek

Tyonek

Nikiski

Anchorage

Seldovia Port

Augustine I. Graham

Kenai Peninsula

Cook Inlet

Gulf of Alaska

154 W 153 W 152 W 151 W 150 W 149 W

62 N

61 N

60 N

59 N

Kalgin I.

North Foreland

East Foreland

Figure 3. Cook Inlet.
8

59 36′N

Bishop Beach

Homer Mud Flat

Homer Boat Harbor

Seldovia Beach

Jakolof-1

Jakolof-3

Jakolof-5

Jakolof-4

Jakolof-2

Tutka Bay-1

Tutka Bay-2

59 30′N

59 24′N

151 45′W 151 30′W 151 15′W

Figure 4. Sample sites within the Lower Cook Inlet region.

Table 2. Sample sites – Lower Cook Inlet region. Sites denoted with an asterisk

(*) indicate samples used in the adsorption and desorption experiments.

Station Latitude (N) Longitude (W)

Bishop Beach 59° 38.3′ 151° 33.3′

Homer Boat Harbor* (HBH) 59° 36.3′ 151° 25.0′

Homer Mud Flat 59° 38.3′ 151° 29.3′

Jakolof-1 59° 27.1′ 151° 29.6′

Jakolof-2* (JB2) 59° 27.1′ 151° 29.3′

Jakolof-3 59° 27.1′ 151° 29.6′

Jakolof-4 59° 27.1′ 151° 29.3′

Jakolof-5 59° 27.1′ 151° 29.7′

Seldovia Beach 59° 23.8′ 151° 41.0′

Tutka Bay-1 59° 26.8′ 151° 20.5′

Tutka Bay-2* (TB2) 59° 24.8′ 151° 17.0′
9

Feely et al. [1980] and Larrance and Chester [1979] studied the distribution and dynamics of Cook Inlet

sediments. Concentrations of suspended matter, largely rock flour input by streams draining surrounding

glaciers, are high in upper Cook Inlet, ranging from about 50 to 100 mg L–1near the Forelands to 1 g L–1

near the mouths of some rivers. These materials tend to have low organic content, 0.5 to 1% by weight.

Suspended sediment concentrations in Lower Cook Inlet are much lower, often near the 1 mg L–1 typical

of the adjacent Gulf of Alaska waters. The organic content of suspended particles is higher than in the

upper inlet, averaging 4% in surface waters, due to seasonally high phytoplankton productivity and to

fluvial inputs of terrigenous organic matter. The change in suspended sediment concentration does not

appear to be largely due to deposition in the central area of the inlet. Feely et al. [1980] theorized that

much of the sediment is either carried out into Shelikof Strait or deposited in small embayments along

the coastline. There are also substantial cross-channel gradients in suspended sediment concentrations,

which are generally lower on the eastern side due to water circulation patterns. The available information

suggests that any sediments contaminated by petroleum in the upper or lower sections of Cook Inlet

might ultimately be deposited in bays of the lower inlet or outside the inlet.

Port Valdez

Port Valdez is the terminus of the Trans Alaska Pipeline, where major oil-related activities have been

ongoing since 1977. In addition to the oil terminal operations, tanker movements, merchant shipping,

and fisheries activities all contribute petroleum to the marine environment [Hameedi 1988]. Sampling

locations for this study in Port Valdez (Figure 5 and Table 3) are a subset of sites originally selected for

initial environmental impact studies and long-term monitoring of petroleum operations within the area

[Feder and Shaw 1996].

50 45

91

37

11

16 80 77

82

25

+

+

+ +

+ +

+

+

++

61 08′N

146 40′W

61 06′N

61 04′N

146 30′W 146 20′W

+ +

51 33

Valdez

Figure 5. Sample sites within the Port Valdez region.

The region is characterized by glaciated, steep, high mountains (150 to 1000 m) which create a fjord that

is 21 km long and 4.5 km wide. Port Valdez is a flat-bottomed, steep-sided, glacially-carved trough, with

two sills near the mouth of the port [Feder et al. 1976]. Surface waters range from less than 2.5°C in the

winter to greater than 11°C in the summer, but temperatures below 75 m remain in the 3 to 6°C range.

Typical precipitation in the area is in on the order of 158 cm yr–1 [Hood et al. 1973], causing surface

salinities to be extremely low in the summer (<1.0‰), but during the winter, freezing air temperatures

and limited freshwater runoff allow salinities to increase to about 32‰ and result in mixing of the entire

water column within Port Valdez.
10

Table 3. Sample sites – Port Valdez region. Sites denoted with an asterisk (*)

indicate samples used in the adsorption and desorption experiments.

Station Latitude (N) Longitude (W) Depth (m)

PV11 61° 06.4′ 146° 20.0′ 198

PV16 61° 05.9′ 146° 21.8′ 239

PV25* 61° 05.5′ 146° 23.3′ 73

PV33 61° 05.4′ 146° 23.1′ 55

PV37 61° 07.7′ 146° 26.9′ 51

PV45 61° 06.4′ 146° 32.3′ 241

PV50 61° 06.4′ 146° 35.7′ 241

PV51 61° 05.4′ 146° 23.5′ 79

PV77 61° 05.8′ 146° 22.8′ 231

PV80 61° 05.7′ 146° 20.7′ 79

PV82* 61° 05.4′ 146° 22.3′ 79

PV91 61° 04.9′ 146° 33.9′ 45

Much of the sediment within tidal flats, outwash deltas, and alluvial fans consists of poorly sorted debris

ranging from gravels to muddy sands, with anoxic conditions existing 3 cm below the surface of the mud

flats. The sedimentary clay minerals are composed mainly of chlorite and illite which have undergone

little in the way of chemical weathering [Feder et al. 1976].

Summary

This report describes an experimental investigation of the adsorption and desorption of PAH from Lower

Cook Inlet, Port Valdez, and inner Beaufort Sea shelf sediments, including an examination of the reasons

that desorption of aromatic hydrocarbons from sediments is often slow and incomplete. The work is

built upon that of Henrichs et al. [1997], who investigated PAH adsorption and desorption by Lower

Cook Inlet sediments, and Braddock and Richter [1998], who studied the biodegradation of PAH in

seawater–sediment systems from Lower Cook Inlet. The goal of the research was to improve the ability

to predict the persistence and fate of these compounds in slightly- to moderately-contaminated sediments.
11

Acronyms and Abbreviations

CHN carbon–hydrogen–nitrogen

CPMAS/13CNMR cross-polarized magic angle spinning 13C nuclear magnetic resonance

DOM dissolved organic matter

dpm disintegrations per minute

E4/E6 ratio of absorbances at 465 and 665 nm

EPA Environmental Protection Agency

FTIR Fourier transform infrared

GC gas chromatography

HBH Homer Boat Harbor

HOC hydrophobic organic compounds

IRMS isotope ratio mass spectrometry

KOC organic carbon normalized partition coefficient

Kp partition coefficient

MCT-B mercury/cadmium/telluride detector at the 400–4000 cm–1 wavenumber range

N/C nitrogen to carbon ratio

NIST National Institute of Standards and Technology

NMR nuclear magnetic resonance

OC organic carbon

PAH polycyclic aromatic hydrocarbons

PCB polychlorinated biphenol

PCP pentachlorphenol

PDB PeeDee Belemnite (reference standard for δ13C values)

TARO total aromatics

TOC total organic carbon

UHP-N2 ultra high purity nitrogen gas
12
13

Chapter 2. Effects of Humic Acid Properties on Phenanthrene Adsorption1

Abstract

This study examined the role humic acid and its structure plays in the adsorption of organic pollutants

in coastal sediments. Humic acids were extracted from intertidal and subtidal Alaska coastal marine

sediments, representing a variety of organic matter sources. Melanoidins, used as model humic acids,

were synthesized via the Maillard reaction between glucose and bovine casein. The natural and synthetic

humic acids were characterized by elemental and isotopic analyses, as well as ultraviolet/visible,

Fourier transform infrared, and cross-polarized magic angle spinning 13C nuclear magnetic resonance

spectroscopies. Adsorption of phenanthrene to the humic substances was investigated using a

radiotracer. Concentrations of humic materials on clay were not related to partition coefficients (K).

The percent nonpolar carbon of humic material, but not aromaticity alone, was weakly correlated to the

organic carbon normalized K (KOC ). However, KOC was more closely negatively correlated with the sum

of amide and carboxylic carbons, a measure of the polarity of the humic acids.

Introduction

Aromatic hydrocarbons make up only about 10% of the complex mixture of compounds in petroleum,

but are a special concern as environmental contaminants, because they are among the most toxic,

mutagenic, and carcinogenic constituents [Black et al. 1983; White 1986; Pahlman and Pelkonen 1987].

They are also the most soluble constituents, and thus the most likely to be transported in solution, to

affect organisms from the dissolved phase, and to be affected by adsorption–desorption reactions in the

marine environment. However, in absolute terms, aromatic hydrocarbons have low to very low aqueous

solubilities and high affinities for surfaces in aquatic ecosystems [Farrington and Westall 1986]. Their

adsorption is usually not rapidly or completely reversible, indicating that this process is probably a factor

in polycyclic aromatic hydrocarbon (PAH) retention by contaminated sediments [Brusseau and Rao 1991;

Hatzinger and Alexander 1995]. Adsorption may decrease the availability of contaminants to microbial

degraders [Weissenfels et al. 1992; Manilal and Alexander 1991; Braddock and Richter 1998]. Sediments

are important reservoirs for aromatic hydrocarbons in the marine environment [Wakeham and Farrington

1980; Gschwend and Hites 1981].

Weissenfels et al. [1992] concluded that biodegradability and biotoxicity of PAH were decreased by

adsorption and migration into an organic matter matrix. It has been suggested by some that the

concentration is less important in determining the pollutant’s biological hazard than is the ratio of PAH to

total organic matter [Hansen et al. 1991]. Proponents of this hypothesis reason that there is a characteristic

partition ratio of PAH in organic matter, and only when this is exceeded do substantial amounts of

pollutant become bioavailable. This hypothesis was the basis for the U.S. Environmental Protection

Agency (EPA) to write “Proposed Sediment Quality Criteria for the Protection of Benthic Organisms”,

in which acceptable concentrations of pollutants in sediments were related to the concentration of organic

materials in those sediments [Hansen et al. 1991].

For uncharged organic pollutants, organic matter is considered the primary sorbent component

[Karickhoff and Morris 1985]. The adsorption partition coefficient, Kp, increases with increasing organic

content of sediments and soils [Means et al. 1980; Schwartzenbach and Westall 1981; Shaw and Terschak

1998]. Organic matter properties also influence adsorption. Adsorption of various aromatic hydrocarbons

has correlated positively with the nonpolar–polar functional group ratio [Rutherford et al. 1992; Garbarini

and Lion 1986] and aromaticity [Gauthier et al. 1987]. The mechanisms for binding of PAH to organic

 

1

Terschak, J.A., S.M. Henrichs and D.G. Shaw. Effects of Humic Acid Properties on Phenanthrene Adsorption. In

preparation for submission to Environmental Science and Technology.
14

matter have not been completely elucidated [Ragle et al. 1997 and references therein]. Interactions

between PAH and humic acids result, at least in part, from weak dipole interactions between aromatic

rings of the PAH and those of humic acids, particularly humic acids derived from lignins and related

materials of woody terrigenous plants [Chin et al. 1997; Chin et al. 1994]. The adsorption of neutral

hydrophobic organic compounds (HOC) to organic matter is generally considered to involve rapid

van der Waals type interactions [Brusseau and Rao 1991].

Organic matter, of which humic acid is a part, is a flexible, cross-linked, branched, amorphous,

polyelectrolytic, polymeric substance within which organic pollutants can diffuse [Brusseau et al.

1991]. Schnitzer and Khan’s [1972] theory is that humic acid is an open structure with hydrophobic

cavities [Schlautman and Morgan 1993 and references therein]. Direct confirmation of the “porous”

nature of organic matter has been reported [Brusseau et al. 1991 and references therein]. While the pores

of sediment mineral particles are fixed and rigid, organic matter pores are dynamic and ephemeral

[Brusseau et al. 1991]. The size, shape, and hydrophobicity of the pores within organic matter are

sensitive to variations in solution chemistry, and, therefore, changes in the ability to bind HOC occur

[Schlautman and Morgan 1993]. The hydrophobicity of PAH suggests that PAH association is governed

by diffusion into nonpolar environments of organic matter [Gauthier et al. 1987; Ragle et al. 1997; Uhle

et al. 1999].

A number of authors have reported evidence for irreversible adsorption of organic compounds to particles

and organic matter [Brusseau and Rao 1991; Hatzinger and Alexander 1995; Karickhoff and Morris 1985;

Brusseau et al. 1991; Carmichael et al. 1997; Kan et al 1994; Di Toro and Horzempa 1982; Pavlostathis

and Mathavan 1992], but differ over the exact causes for the observations. Hypothesized mechanisms

include adsorption to the particle’s surface [Mingelgrin and Gerstl 1983], diffusion into tortuous particle

micropores [Brusseau et al. 1991; Wu and Gschwend 1986; Steinberg et al. 1987], and diffusion into the

organic matrix [Brusseau and Rao 1991; Karickhoff and Morris 1985; Brusseau et al. 1991; Schlautman

and Morgan 1993; Kan et al. 1994]. Intraorganic matter diffusion is thought to be the most likely

mechanism for HOC [Brusseau and Rao 1991; Brusseau et al. 1991]. Adamson [1990] suggests that

irreversible adsorption is probably due to a mechanical or structural rearrangement of the adsorbent, that

is, the matrix from which desorption takes place is different from that during adsorption.

This study was conducted to examine the role that humic acids play in the adsorption of PAH by the

sediments of coastal marine environments. This chapter addresses the effects of humic acid structure and

other characteristics on the extent and rate of phenanthrene adsorption. Both natural humic substances and

an artificial analog, melanoidins [Gagosian and Stuermer 1977; Nissenbaum and Kaplan 1972; Hedges

and Parker 1976], were investigated. These organic substances represent a wide range of properties

potentially affecting associations with PAH.

Experimental Section

Sampling sites and sample collection

Samples were taken from three separate coastal locations in Alaska (Table 4). In Port Valdez, natural

petroleum seeps [Page et al. 1996], oil terminal operations, tanker movements, general shipping, and

fisheries activities all contribute petroleum to the marine environment [Hameedi 1988]. In Lower Cook

Inlet, shipping and fisheries activities contribute petroleum and additional PAH come from natural oil

seeps and erosion of coastal and submarine coal outcrops [Shaw and Wiggs 1980]. Although the Beaufort

Sea is a site of petroleum exploration and extraction, the main contributors of PAH to the marine

environment are non-anthropogenic sources such as natural oil seeps and tundra vegetation [Naidu et al.

2001; Yunker et al. 1991; Steinhauer and Boehm 1992; Yunker and MacDonald 1995].
15

Intertidal sediment samples from Lower Cook Inlet were collected in July 1995, July 1996, and May

1999. Sampling occurred during spring tides below the mean lower-low water level. Subtidal sediment

samples were collected from the Port Valdez region in August 1996 using a Haps corer and from the

Beaufort Sea region during September 1997 using a Kynar-coated van Veen grab sampler. Any overlying

debris was removed and then sediment was collected from the oxic layer (upper 2 cm) with metal

implements that had been heated to redness before use. All samples were stored until needed at –50°C

in pre-combusted glass jars.

Table 4. Site locations of sediments discussed in this study; n/a – not applicable, as Lower

Cook Inlet samples were intertidal and collected during low tide.

Location Station Latitude (N) Longitude (W) Depth (m)

Port Valdez PV25 61° 05.5′ 146° 23.3′ 73

Port Valdez PV82 61° 05.4′ 146° 22.3′ 79

Lower Cook Inlet HBH 59° 36.3′ 151° 25.0′ n/a

Lower Cook Inlet JB2 59° 27.1′ 151° 29.3′ n/a

Lower Cook Inlet TB2 59° 24.8′ 151° 17.0′ n/a

Beaufort Sea 5A 70° 29.7′ 148° 46.0′ 11.4

Beaufort Sea 6G 70° 31.3′ 149° 53.9′ 2.1

Sediment organic matter and humic acid characterization

Table 5 lists the characteristics examined in the sediment organic matter and humic acids used in this

study. The methods associated with each are described below.

Sediment organic carbon and nitrogen determination by combustion and isotope ratio mass

spectrometry

Approximately 2 grams of each thawed sediment were weighed into pre-combusted glass vials and oven

dried for 48 hours. The dry sediment was acidified with 1 N HCl and homogenized. Samples were redried

and ground to a powder. The analyses of Port Valdez and Lower Cook Inlet sediments were performed

by the MSI Analytical Lab (Marine Science Institute, University of California, Santa Barbara), which

uses one of two automated CHN analyzers (CE 240 X and Leeman Lab CE 440). Orchard leaves

(NIST [National Institute of Standards and Technology] standard reference material) and previously

characterized marine sediments were the standards. Total organic carbon (TOC) and nitrogen analyses

of the Beaufort Sea sediments and all humic acids were performed at the University of Alaska Fairbanks

using a Finnigan MAT Delta Plus mass spectrometer with a Conflo II interface and a Carlo Erba

elemental analyzer.
16

Table 5. The measured properties of sediment organic matter, natural humic acids, and synthetic

humic acids evaluated in this study and the methods by which they were obtained.

IRMS – isotope ratio mass spectrometry, GC – gas chromatography, CPMAS/13CNMR

– cross-polarized magic angle spinning 13C nuclear magnetic resonance, N/C – nitrogen

to carbon ratio, E4/E6 – ratio of absorbances at 465 and 665 nm.

Organic Constituent Characteristic Method

Sediment organic matter % C Combustion & IRMS

% N Combustion & IRMS

N/C Combustion & IRMS

Phenanthrene Extraction & GC

Total aromatics Extraction & GC

Humic acid % Extractable humic acid Gravimetric

% C Combustion & IRMS

% N Combustion & IRMS

N/C Combustion & IRMS

δ

13C & δ15N ratios Combustion & IRMS

E4/E6 UV/Visible spectroscopy

% Aromaticity (carboxyl-free,

aromatic carbon) CPMAS/13CNMR

% Aliphatic carbon CPMAS/13CNMR

% Oxygen bound alkyl carbon CPMAS/13CNMR

% Carboxyl + amide carbon CPMAS/13CNMR

% Carbonyl carbon CPMAS/13CNMR

% Non-polar carbon (aromatic +

aliphatic carbon) CPMAS/13CNMR

% Polar carbon (O-alkyl + carboxyl

+ amide + carbonyl carbon) CPMAS/13CNMR

Partition coefficient Radiotracer

Sediment hydrocarbon analyses by extraction and gas chromatography

Hydrocarbon analyses of Port Valdez and Lower Cook Inlet sediment samples were done as described

by Feder and Shaw [1996]. Beaufort Sea samples were analyzed by M.I. Venkatesan’s laboratory at

the University of California Los Angeles following a similar method [Naidu et al. 2001]. The sum

of 18 individual PAH (naphthalene, 2-methylnaphthalene, 1-methylnaphthalene, biphenyl,

2,6-dimethylnaphthalene, acenaphthene, fluorene, phenanthrene, anthracene, 1-methylphenanthrene,

fluoranthene, pyrene, benz[a]anthracene, chrysene, benzo[e]pyrene, benzo[a]pyrene, perylene, and

dibenz[a,h]anthracene) are reported as total aromatics (TARO) in μg PAH kg–1 dry sediment.

Humic acid extraction

Humic acid was obtained using the alkali (0.5 M NaOH) extraction procedure of Anderson and Schoenau

[1983]. After extraction from approximately 30 g of wet sediment, humic acid was precipitated by the

addition of 6 M HCl until a pH of 1.5 was obtained. The precipitated extract was allowed to stand for

15 minutes and was then centrifuged for 30 minutes at 14,000 g. The precipitated humic acid was
17

lyophilized at –85 °C for 24 hours. The resulting dry humic acid was stored at room temperature in a glass

vial under N2.

Synthetic humic acid preparation

Artificial humic acids were prepared in the laboratory by refluxing several ratios of glucose and bovine

casein (1:1, 10:1, and 1:10) in buffered solutions for 24 to 48 hours as described by Yamamoto and

Ishiwatari [1989]. The buffer (50%/30%/20%) was composed of 125 ml 0.1 M KH2PO4, 75 ml 0.1 M

NaOH and 50 ml of glass-distilled water to give a final pH of 7.0. After the prescribed reaction time, the

mixture was allowed to cool with stirring, then transferred to 250-ml Nalgene® centrifuge bottles and the

artificial humic acid was isolated using the extraction procedure described in the previous paragraph.

UV/Visible spectroscopy

The ratio of absorbance at 465 and 665 nm (E4/E6 ratio) of the extracted humic acids was determined

using ultraviolet/visible light spectroscopy. A Milton Roy Spectronic 21D UV/visible spectrometer was

used to measure light absorbance characteristics of the humic acids. Humic acid (2–4 mg) was dissolved

in 10 ml of 0.05 N NaHCO3 [Chen et al. 1977] and light absorbance readings were taken.

FTIR spectra acquisition

Fourier transform infrared (FTIR) spectra were obtained using a Nicolet Magna 560 FT spectrometer

with an infrared source, KBr beam splitter, and an MCT-B detector. Sixty milligrams of a ground mixture

of 0.200 g KBr and 0.005 g humic acid were made into a pellet. Signals were averaged from 200 scans at

a resolution of 0.121 cm–1. Spectra were obtained between 4000 cm–1 and 400 cm–1 and processed against

a background of KBr.

CPMAS/13CNMR spectra acquisition

Cross-polarized magic angle spinning 13C nuclear magnetic resonance (CPMAS/13CNMR) spectroscopic

data were obtained from the NMR Laboratory at Florida State University in Tallahassee using the

standard methods described by Hatcher et al. [1981]. The NMR spectrometer was a Bruker/IBM

WP200SY with a solid state accessory package. A 7-mm standard Doty Scientific multifrequency

CPMAS probe was employed. Sample weights of approximately 225–300 mg were used and spun at

3.5–4.3 kHz. The frequency for 13C on this instrument was 50.325 MHz. Five major regions of chemical

shifts between 0 ppm and 300 ppm were assigned. These corresponded to carbons associated with

aliphatic (0–50 ppm), aromatic (110–160 ppm), oxygen bound alkyl (50–110 ppm), carboxylic and

amide (160–190 ppm), and carbonyl (190–240 ppm) functional groups [Malcolm 1990]. Carboxyl-free

aromaticities were reported by subtracting the integrated area of carboxylic and amide carbon peaks from

those assigned to the aromatic carbons. The nonpolar fraction was calculated by adding the fractions of

aliphatic and aromatic carbons.

Adsorption experiments

Radiolabeled phenanthrene was used to examine adsorption [Henrichs et al. 1997]. All glassware and

Teflon® liners used in this study were treated to remove any hydrocarbon contamination before use.

Glassware was baked in a muffle furnace at 450 °C for at least 8 hours. Teflon® cap liners were soaked

in chromic acid for 20 minutes and then rinsed 3 times with organic-free water. A stock solution

was prepared by dissolving [9-14C] phenanthrene (5–15 mCi mmol–1; Sigma Chemical Co.) and

nonlabeled phenanthrene in acetonitrile. After evaporation of the acetonitrile, artificial seawater was

added to aliquots of the stock solution to produce 50, 300, and 700 μg L–1 phenanthrene solutions.

Previous experiments within this laboratory have shown that for concentrations of phenanthrene ranging
18

from 10 to 750 μg L–1, there was no loss of hydrocarbon due to evaporation or adhesion to the walls of the

container [Henrichs et al. 1997]. In addition, all experiments had control vials containing no sediment.

Artificial seawater was prepared from organic-free water that had been glass distilled over a saturated

solution of potassium permanganate. To each liter of water the following salts were added: 23.260 g

NaCl, 10.636 g MgCl2 •6H2O, 3.918 g anhydrous Na2SO4, 1.102 g CaCl2, 0.664 g KCl, 0.192 g NaHCO3,

0.096 g KBr, and 0.026 g H3BO3 [Lyman and Fleming 1940]. Also, 0.500 g HgCl2 was added as an

antibiologic agent.

A montmorillonite standard (A.P.I. # 26, 49 E 2600 from Clay Spur, Wyoming; Ward’s Natural Science

Establishment, Inc.) was ground for 2 minutes using a shatter box equipped with carbide rings. The

resulting powder was passed through a 270 mesh (53 μm) sieve using a Ro-Tap® apparatus. Organic

material was removed from the clay by oxidizing it with 30% hydrogen peroxide.

Humic acids were coated onto inorganic substrate for use in the adsorption experiments. Approximately

1 g of lyophilized humic acid was dissolved in 100 mL of organic-free water, maintained at a pH of 10

by additions of 0.5 M NaOH, and mixed under an atmosphere of ultra high purity nitrogen gas (UHP-N2)

for 3 to 7 days. The humic acid solution was then added to the washed clay and mixed for 24 hours. The

suspension was transferred to a centrifuge bottle containing the salts resulting from the evaporation of an

equal volume of artificial seawater and mixed for 48 hours, then centrifuged at 14,000 g for 2 hours and

the supernatant discarded. A sample of the humic acid coated clay was dried and submitted for TOC

analysis. The remaining humic acid coated clay was stored in a refrigerator at 5–10°C in a container

flushed with UHP-N2.

Adsorption experiments began by weighing 0.1 g of wet, humic-coated clay into each of a number of

2-dram vials with Teflon® lined caps and adding 5 ml of a radiolabeled phenanthrene/seawater solution.

After mixing for 15 seconds using a vortex mixer, the vials were placed on a table shaker at 150–200 rpm.

After 14 days, the vials were centrifuged at 2400 g for 30 minutes. A 1.0 mL aliquot of each of the

radiolabeled reaction solutions was removed and the 14C activity was measured by liquid scintillation

counting. Partition coefficients, KOC, were reported as the ratio of adsorbed phenanthrene to dissolved

phenanthrene, normalized to organic carbon (Equation 2–1).

K

a

b

c

a

b

d

f OC OC =

− ×

×

100 100

100

where a is the measured dpm of the sample, b is the measured dpm of the control, c is the dry sediment

weight of the sample, d is the solution volume, and fOC is the sediment weight fraction of organic carbon.

Statistical treatments

A Microsoft® Excel 2000 spreadsheet was used to calculate linear correlation coefficients (r

2

) between

pairs of variables. The statistical significance of correlations was determined using a two-tailed t-test

with (n – 2) degrees of freedom at both the 5% and 10% significance levels (p = 0.05 and p = 0.10,

respectively) [Miller and Miller 1988; Mendenhall and Sincich 1996]. All correlations are reported at

the p = 0.05 significance level unless otherwise noted.

(2–1)
19

Results

Humic acid properties

The δ13C (PDB standard) of the extracted sediment humic acids averaged –24.00 ± 0.99‰ in Port Valdez,

–19.82 ± 1.36‰ in Lower Cook Inlet, and –26.01 ± 0.72‰ in the Beaufort Sea. The δ15N (air standard) of

the extracted sediment humic acids averaged 6.46 ± 1.99‰ in Port Valdez, 7.71 ± 1.37‰ in Lower Cook

Inlet, and 5.42 ± 2.21‰ in the Beaufort Sea.

Nitrogen/carbon ratios (N/C) of natural humic acids ranged from 0.026 to 0.14 with a mean of

0.088 ± 0.027; they were correlated with, but consistently lower than, those of the total sediment

organic matter. The synthetic humic acids had greater N/C, ranging from 0.16 to 0.27 with a mean of

0.22 ± 0.041. The N/C of laboratory prepared humic acids increased as the casein/glucose in the reaction

mixture increased.

Spectral results of CPMAS/13CNMR showed that natural humic acids and melanoidins were similar

(Figure 6). As indicated by 13CNMR data, the carboxyl-free aromaticities ranged from 15 to 34% for

natural humic acids and 10 to 18% for the synthetic humic acids. The percentage of carbon atoms

involved in carboxyl and amide bonds ranged from 10 to 14% for natural humic acids and 16 to 22% for

the synthetic humic acids. As expected, increasing N/C ratios corresponded to an increase in the relative

proportions of carbon atoms associated with amide bonds (r

2 = 0.90, t = 10.091) (Figure 7). The trend was

similar for natural and synthetic humic acids, although the natural humic acids were consistently lower in

amide carbon and N/C. Increasing N/C ratios also correlated with a decrease in carboxyl-free aromaticity,

but this trend was caused by the difference between natural and synthetic humic acids and did not exist

within the natural humic acids. Negative correlations existed between aromaticity and the degree of

polarity, as measured by N/C (r 

2

= 0.55, t = 3.649), and fraction carboxyl and amide C (r

2 = 0.44,

t = 2.914), but aromaticity did not correlate significantly with the aliphatic fraction (r

2 = 0.05, t = 1.718)

in the humic acids of this study. E4/E6 ratios showed no relationship to the other measured properties of

the humic acids, including carboxyl-free aromaticity (r

2 = 0.10, t = 1.120).

Fourier transform infrared spectra of the natural and synthetic humic acids were similar to each other

[cf. Ertel and Hedges 1983], with some minor differences (Figure 8). All showed absorbance bands for

hydrogen-bonded OH at about 3400 cm–1 [Ertel and Hedges 1983; Goh and Stevenson 1971; Stevenson

and Goh 1971]; aromatic C–H stretch as a shoulder between 3100 and 3000 cm–1 [Silverstein et al. 1991];

aliphatic C–H stretch between 3000 and 2850 cm–1 [Ertel and Hedges 1983; Goh and Stevenson 1971;

Stevenson and Goh 1971; Rubinsztain et al. 1984]; C=O stretch of aldehydes, ketones, and acids as a

shoulder at 1710 cm–1 [Ertel and Hedges 1983; Rubinsztain et al. 1984; Kononova 1966]; amide linkages

of proteins at 1650 cm–1 [Ertel and Hedges 1983; Goh and Stevenson 1971; Stevenson and Goh 1971];

C–O stretch of ethers, esters, and phenols at 1230 cm–1 [Ertel and Hedges 1983; Goh and Stevenson 1971;

Stevenson and Goh 1971]; COO– stretching at 1532 and 1386 cm–1 [Stevenson and Goh 1971]; C–O

stretch of carbohydrates at 1040 cm–1 [Ertel and Hedges 1983; Goh and Stevenson 1971; Stevenson and

Goh 1971]; and possible NH3

+

 torsional vibrations at 530 and 470 cm–1 [Silverstein et al. 1991]. The

Beaufort Sea 5A humic acid showed stronger absorption bands associated with ethers, esters, and phenols

than the other natural and synthetic humic acids, including a sharp phenolic and alcoholic OH stretch at

3622 cm–1 [Silverstein et al. 1991]. The Lower Cook Inlet samples all showed pronounced absorption

bands associated with amide linkages, while the same bands in other humic acids were less prominent.

The synthetic humic acid, 1:10–24, produced a more defined peak in the aromatic C–H stretching band

than any of the other humic acids.
20

200 ppm 100 ppm 0 ppm

Figure 6. Cross-polarized magic angle spinning/13C nuclear resonance spectroscopy. The figure compares

spectral bands of a natural humic acid (TB2, top) and a melanoidin (1:10–24, bottom).

Port Valdez

Lower Cook Inlet

Beaufort Sea

Synthetic

0.05

N/C Ratio Humic Acid

0 0.1 0.15 0.2 0.25 0.3

0

5

10

15

20

25

Fraction Carboxyl and Amide C

Figure 7. Increasing nitrogen to carbon ratios of humic acid were associated with

an increase in the fraction of carbon atoms associated with oxygen and

nitrogen bonds in natural and synthetic humic acids.
21

4000 3500 3000 2500 2000 1500 1000 500

120

110

100

90

80

70

60

40

50

30

20

10

% Transmittance

4000 3500 3000 2500 2000 1500 1000 500

Wave numbers (cm–1)

105

95

85

75

65

55

45

35

25

% Transmittance

Figure 8. Fourier transform infrared spectroscopy. The figure compares spectral bands of a natural humic

acid (TB2, top) and a melanoidin (1:10–24, bottom).

The humic acids of this study, both natural and synthetic, had major 13CNMR peaks corresponding to

terminal methyl groups (~24 ppm), methylene carbon (~31 ppm), aliphatic esters and ethers, methoxyl

and ethoxyl carbons (~55 ppm), ring carbons of polysaccharides and ether bonded aliphatic carbons

(~70 ppm), carbon singly bonded to two oxygen atoms and anomeric carbon in polysaccharides, acetal

or ketal groups (~100 ppm), unsubstituted and alkyl substituted aromatic carbons (~129 ppm) [cf. Ikan

et al. 1992], and mostly carboxyl with some esters and amide carbon (~173 ppm) [Malcolm 1990].

TARO ranged from values of less than 100 μg PAH kg–1 sediment to larger values, such as

708 μg PAH kg–1 sediment at the Homer Boat Harbor (HBH) in Lower Cook Inlet, a site with substantial

anthropogenic contaminants. The high Beaufort Sea sample (5A), with 632 μg PAH kg–1 sediment,

contained mainly natural TARO derived from tundra vegetation on the nearby coast [Naidu et al. 2001].

Environmental TARO correlated neither with sediment organic carbon content nor with humic acid N/C

ratios (r

2 = 0.00, t = 0.385). In fact, there was no trend for environmental TARO concentrations with any

of the humic acid properties assessed in this study.

Adsorption measurements

While all of the humic acids investigated showed strong adsorption of phenanthrene, the synthetic

humic acids’ partition coefficients averaged about 50% less than those for the natural humic acids. The

montmorillonite clay alone adsorbed phenanthrene at a much reduced level, with the average partition
22

coefficients for the montmorillonite clay being three times smaller than those for the humic acid coated

clay with the smallest partition coefficients.

Neither aromaticity (r

2 = 0.15, t = 1.121) nor aliphaticity (r2 = 0.03, t = 0.456) appeared to control

measured partition coefficients in the laboratory (Figures 9a and 9b). An increase in humic acid nonpolar

carbons (i.e., those only involved in aliphatic or aromatic bonds) weakly correlated with an increase in the

partition coefficient (r

2 = 0.38, t = 2.091, p = 0.10) (Figure 9c). However, partition coefficients decreased

as nitrogen increased relative to carbon (Figure 10a); this trend was mainly due to the difference between

natural and synthetic humic acid properties. The combined effects of nitrogen and oxygen, measured as

the percent carboxyl and amide carbon, correlated with a decrease in the partition coefficient (r

2 = 0.78,

t = 4.982) (Figure 10b). The natural humic acids showed this trend clearly, and it was also seen in the

comparison of natural and synthetic humic acids (Figure 6). Partition coefficients (Kp; not normalized to

organic carbon content) did not correlate with the amount of organic carbon coated on the clay particles

(mg OC g–1 clay) used in the laboratory experiments (r

2 = 0.03, t = 0.429).

Port Valdez

Lower Cook Inlet

Beaufort Sea

Synthetic

15

% Aromaticity (carboxyl-free)

10 20 25 30 35

0

5000

KOC (mL g–1)

10000

15000

20000

25000

30000

Figure 9a. Partition coefficients (KOC) and percent aromaticity of humic acids

were not significantly correlated.
23

Port Valdez

Lower Cook Inlet

Beaufort Sea

Synthetic

30

% Aliphaticity

25 35 40 45

0

5000

KOC (mL g–1)

10000

15000

20000

25000

30000

Figure 9b. Partition coefficients (KOC) and percent aliphaticity of humic acids

had no significant correlation.

Port Valdez

Lower Cook Inlet

Beaufort Sea

Synthetic

50

Fraction Nonpolar C

45 55 60 65

0

5000

KOC (mL g–1)

10000

15000

20000

25000

30000

Figure 9c. Partition coefficients (KOC) increased with the fraction of nonpolar

carbons (the sum of those carbons involved with aromatic and aliphatic

bonds).
24

Port Valdez

Lower Cook Inlet

Beaufort Sea

Synthetic

0.05

N/C Ratios Humic Acid

0.00 0.10 0.15 0.25 0.30

0

5000

KOC (mL g–1)

10000

15000

20000

25000

30000

0.20

Figure 10a. Partition coefficients (KOC) decreased with increasing nitrogen to

carbon (N/C) ratios of humic acid.

Port Valdez

Lower Cook Inlet

Beaufort Sea

Synthetic

12

Fraction Carboxyl and Amide C

10 14 16 20 22

0

5000

KOC (mL g–1)

10000

15000

20000

25000

30000

18

Figure 10b. Partition coefficients (KOC) decreased as the number of carbons in

O and N bonds (fraction carboxyl and amide C) increased.
25

Discussion

The organic matter examined in earlier studies on hydrophobic organic compound adsorption to humic

acids was isolated from a variety of sources (e.g., river dissolved organic matter [Chiou et al. 1987], river

sediment [Means et al. 1980; Karickhoff et al. 1979], estuarine sediment [Gundersen et al. 1997], coastal

marine sediment [Maruya et al. 1996]), but was not usually very diverse in its properties within each

study. This project sought to broaden the range of organic matter properties examined within a single

investigation. Although this study used marine humic acids possessing a wide variety of properties,

the characteristics were within the range of those seen in previous studies. The melanoidins, used as

synthetic, model humic acids, were also similar to the natural marine derived humic acids in many

respects, but had greater average N/C, a greater fraction of amide and carbonyl carbon, and negligible

TARO. These synthetic humic acids are condensation products of reducing sugars and amino acids

[Maillard 1913]. One hypothesis on the origins of naturally formed humic acids in the marine environment

is that they are produced by the same types of browning reactions [Ertel and Hedges 1983; Enders and

Theis 1938; Hodge 1953], based upon chemical similarities between melanoidins and natural marine

humic acids [Gagosian and Stuermer 1977; Nissenbaum and Kaplan 1972; Hedges and Parker 1976].

Sediment isotope ratios tend to reflect the sources of organic carbon, namely photosynthetic organisms

[Nissenbaum and Kaplan 1972 and references therein; Ikan et al. 1990]. Several investigations have used

stable isotopic ratios of carbon to help determine the source material of marine humic acids [Gagosian

and Stuermer 1977; Nissenbaum and Kaplan 1972; Malcolm 1990; Ikan et al. 1992; Ikan et al. 1990].

Biomolecules formed in the marine environment usually have isotopic signatures around –20 to –23‰.

Terrestrially derived biomolecules from C3 plants have δ13C values around –25 to –27‰. Literature

values for marine humic acids range between –19 and –23‰ [Gagosian and Stuermer 1977; Nissenbaum

and Kaplan 1972; Malcolm 1990; Ikan et al. 1992; Ishiwatari 1992]. Humic acids from coastal and littoral

sediments have a larger range, –19 to –27‰, indicating mixed marine and terrestrial inputs [Nissenbaum

and Kaplan 1972]. Nitrogen isotopic signatures in the marine environment are usually in the range of

6–10‰, reflecting the sources of inorganic nitrogen in the form of ammonium and nitrate in the ocean,

while terrestrial plant δ15N values are closer to that of atmospheric nitrogen (0‰).

N/C ratios can be used in conjunction with isotopic studies to make a clearer distinction between marine

and terrestrial humic acids. Organic matter of most Recent soils and sediments has N/C ratios of about

0.083 to 0.1 [Stevenson 1994]. A similar range was found for extracted humic acids from most marine

sediments [Ertel and Hedges 1983; Ikan et al. 1992; Ishiwatari 1992; Rashid et al. 1972; Gauthier et al.

1987], while some authors have reported lower marine humic acid N/C values [Nissenbaum and Kaplan

1972; Sastre et al. 1994]. Terrestrial plants contain nitrogen deficient lignin and cellulose, and have low

N/C [Goodell 1972], while marine phytoplankton is richer in protein, leading to higher N/C (~0.167)

[Muller 1977]. N/C ratios of the synthetic humic acids were significantly higher than those previously

reported in the literature [Ertel and Hedges 1983; Rubinsztain et al. 1984; Ikan et al. 1992], perhaps due

to the use of a protein (bovine casein) as opposed to an amino acid reactant. Synthetic humic acid N/C is

very dependent upon the reactants and the amino acid/carbon ratio of starting materials [Rubinsztain et al.

1984].

The values of δ13C, δ15N and N/C, shown in Table 6, indicate that Lower Cook Inlet sediment humic

acids are mostly of marine origin, even with the surrounding steeply sloping, forested coast. Like Lower

Cook Inlet, Port Valdez is also a fjordal environment. Port Valdez humic acids, however, show greater

terrestrial input. Beaufort Sea humic acids have a marine δ15N signature and N/C similar to Lower

Cook Inlet. The δ13C, however, is relatively light. Zooplankton in the Beaufort Sea have light δ13C,

indicating that marine primary producers there are isotopically light [Schell et al. 1998]. Hence isotopic

and N/C ratio data suggest that Beaufort Sea humic substances are mainly of marine origin. However,

spectroscopic data indicate a significant terrigenous component.
26

Intercomparison of FTIR spectra is limited to qualitative identification of functional groups. The bands

result from the absorption of energies required for vibrational modes of functional groups and provide

no quantitative means of assessing the number or distribution of chemical bonds within a sample. To

overcome this limitation, the use of peak ratios has been explored [Johnston et al. 1994], but was not

successful as a method in this study. Both natural and synthetic humic acid FTIR spectra of this study

were similar to those previously reported in the literature [Ertel and Hedges 1983; Goh and Stevenson

1971; Hedges 1978]. The natural humic acids from all sites showed strong adsorption bands characteristic

of marine derived humic acids [Ertel and Hedges 1983]. A strong phenolic signal in the Beaufort Sea 5A

sample suggests lignin-derived structures.

Similar to the humic acids of this study, Hatcher et al. [1980] reported major 13CNMR peaks in extracted

marine sediment humic acids at 30 ppm, 130 ppm, and 175 ppm, and weak or absent peaks (compared

to terrestrial humic acids) at 75 ppm and 150 ppm. No statistically significant differences in 13CNMR

spectra could be found when comparing the integrated peak areas of the humic acids from Port Valdez

and Lower Cook Inlet, when examined as separate geographical groups. The Beaufort Sea spectra (5A

and 6G) were of very low resolution even after 30,500 and 49,800 scans, respectively. Low resolution for

marine humic acids at up to 182,000 scans has been previously reported [Malcolm 1990]. The Beaufort

Sea 5A humic acid was found to be paramagnetic, probably due to iron content, which interferes with

NMR spectral acquisition.

The synthetic humic acids displayed a larger and sharper 13CNMR peak in the carboxylic region and

smaller integrated aliphatic areas than the natural humic acids. Rubinsztain et al. [1984] showed these

trends to be indicative of sugar rich synthetic humic acids. Gauthier et al. [1987] questioned whether

amide carbon was insignificant in the 160–190 ppm region for marine humic acids. However, the high

N/C ratios of the humic acids of this study, coupled with the fact that the 13CNMR carboxylic region

(160–190 ppm) also includes amide carbons, and the strong presence of amide groups, as determined by

FTIR, indicate that amide carbon is probably responsible for a larger portion of the “carboxylic” 13CNMR

signal than originally assumed.

The remainder of the discussion addresses relationships among humic acid properties and phenanthrene

adsorption. Simple linear correlation coefficients are reported, since the sample size is insufficient for

multivariate techniques. The results of this study suggest that several humic acid properties influence

aromatic hydrocarbon adsorption, but don’t allow a quantitative assessment of the contribution of each

factor. The results do indicate which properties should be the focus of further study.

Table 6. N/C ratios and carbon and nitrogen stable isotope values versus PDB and air,

respectively. n/a – not available

Sample Humic Acid

δ

13C

Humic Acid

δ

15N

Humic Acid

N/C

Humic Acid

N/C

PV25 humic acid –27.57 1.27 0.0657 0.1036

PV82 humic acid –24.37 4.35 0.0878 0.0987

HBH humic acid –21.04 8.56 0.0879 0.0966

JB2 humic acid –17.38 6.95 0.1067 0.1142

TB2 humic acid –18.21 5.72 0.1297 0.1497

5A humic acid –25.94 5.72 0.1109 n/a
27

No significant correlations were found between E4/E6 and aromaticity, any of the other humic acid

characteristics examined, or adsorption of phenanthrene. The E4/E6 ratio has been used to measure the

degree of condensation of aromatic groups [Ertel and Hedges 1983], with smaller ratios of E4/E6 being

found with humic acids of higher aromaticity [Kononova 1966]. However, most studies investigating

E4/E6 as a proxy for aromaticity of humic acids have been unable to support its use [Chin et al. 1994;

Nissenbaum and Kaplan 1972; Johnston et al. 1994; Chen et al. 1977], and the more specific 13CNMR

data presented here provide clear evidence that this measure should not be used.

Environmental TARO were unrelated to sediment TOC, N/C and aromaticity of humic acids, or other

organic matter properties. All environmental TARO values, ranging from 10.8 to 914.4 μg PAH kg–1

sediment, were much less than the sediments’ capacities for PAH adsorption. After 14 days of laboratory

adsorption, the adsorbed quantities for phenanthrene were between 46 and 81 mg kg–1 sediment (46,000

and 81,000 μg kg–1 sediment). There were large variations in TARO from sediments of similar origin

with the same percent organic carbon; for example, TARO values ranged from 36.3 to 517 μg PAH kg–1

sediment with 0.7% OC. Variations in TARO were clearly due to variations in anthropogenic and natural

sources of PAH at the different sampling locations, rather than variations in sediment sorptive properties.

Partition coefficients measured in the laboratory are a better method of determining sorptive properties

of organic matter than bulk TARO measurements, since dissolved PAH concentrations are controlled.

In this study, extracted humic acids were coated onto a uniform clay mineral before use in phenanthrene

adsorption experiments. This procedure reduced variability due to the mineral phase or particle size that

can be present with natural sediments. Clay with no added organic matter showed much less phenanthrene

adsorption than the clay with added humic acids, and so the humic acids were clearly the major adsorbent.

Partition coefficients of phenanthrene determined by this research (14,000 ± 6000 mL g

–1 C) are consistent

within the range found by others. For example, Kopinke et al. [1995] determined phenanthrene KOC

for coal wastewater sediment to be 15,000 mL g–1 C. Karickhoff et al. [1979] found KOC to be

23,000 mL g–1 C for coarse silt fractions of two ponds of northern Georgia. This is comparable to the

partition coefficient of 19,800 ± 954 mL g–1 C found for Boston Harbor sediments by Chin and Gschwend

[1992]. Partition coefficients for intact Jakolof Bay sediments were about 50,000 ± 20,000 mL g–1 C

[Henrichs et al. 1997]; this is larger but of similar magnitude to values for the extracted humic substances

reported here.

Although a positive correlation between organic matter concentration and partition coefficients was found

in several earlier studies [Means et al. 1980; Schlautman and Morgan 1993; Chiou et al. 1987; Karickhoff

et al. 1979; Gundersen et al. 1997; Maruya et al. 1996], it was absent here, probably because of the wide

range of organic matter properties present. This correlation is more likely where organic matter properties

are fairly uniform over the range of organic matter concentrations. Samples used in previous studies were

usually from small geographical regions and characterization of the organic matter was limited. For

example, Karickhoff et al. [1979] sampled two ponds and one small river in northern Georgia. Although

the ponds were associated with different terrestrial vegetation (grass watershed vs. wooded watershed),

other aspects of their environments were similar (lignin, pH, temperature, and weathering regime) and the

structural properties of the organic matter were not characterized. Gundersen et al. [1997] limited their

study to a single estuary in Virginia. Sediment grain size varied, but organic matter differences were

not characterized. Maruya et al. [1996] sampled a very small section of a marsh on San Francisco Bay.

Characterization of their samples was limited to PAH, grain size analysis, and total organic carbon. While

Means et al. [1980] did cover a larger region, by investigating sediments and soils in the Great Lakes

area of the United States, the sampling still took place within an area of similar weathering and parent

materials. Their soils and sediments showed a range in carbon and nitrogen values, but no other

characterizations of the organic matter were done. The results of the present study show that sediment
28

TOC alone is not necessarily the best indicator of sorptive properties. More information on organic matter

properties, selected based on understanding mechanisms of adsorption, is needed.

Hydrophobic adsorption to humic acids is thought to be a pH-independent mechanism of binding between

aliphatic chains and aromatic moieties with non-ionic HOC [Senesi 1992]. A number of laboratory

studies have shown a positive correlation between humic substance aromaticities and measurements of

PAH adsorption such as partition coefficients [Gauthier et al. 1987; Chin et al. 1997; Maruya et al. 1996],

PAH solubility [Uhle et al. 1999], and Freundlich exponents [Xing 2001]. Aromaticity, however, did not

predict partition coefficients in this study. The sum of both the aromatic and aliphatic fractions (as the

nonpolar fraction) predicted partition coefficients better than either alone. The data suggest that

phenanthrene adsorbs equally well to aliphatic and aromatic parts (hydrophobic parts) of the humic

molecule.

Partition coefficients for phenanthrene decreased with increasing polarity of the humic acid and increased

with increasing nonpolar fractions [cf. Kile et al. 1999]. Humic acid polarity, although not measured

directly, was inferred from increasing N/C, increasing fraction of carboxyl and amide carbon, and

decreasing fraction of nonpolar carbon (the sum of aliphatic and aromatic carbons by 13CNMR). The

polarity of the overall structure should increase based on the non-symmetrical addition of nucleophilic

atoms such as nitrogen and oxygen. Similar to the results of this study, Rutherford et al. [1992] showed

a correlation between partition coefficients of aromatic hydrocarbons and polar-to-nonpolar group ratios

(quantified as (O+N)/C) of soil organic matter. More recently, Kile et al. [1999] demonstrated that

polarity (determined by 13CNMR) was a significant factor in the binding of carbon tetrachloride to soil

and sediment organic matter. This is not surprising, following the concept that “like dissolves like”. It

has been proposed that humic acids, being amphiphilic, form regions similar to detergent micelles. These

“pseudomicelles” are considered to have a hydrophobic interior into which nonpolar compounds can

partition [Ragel et al. 1997 and references therein].

The relationship between KOC and the fraction of carboxyl and amide carbon is tighter than the

relationship of KOC and nonpolar carbon. Carboxyl and amide carbons constitute only 10–20% of the

carbon, and it is not obvious why small changes in this fraction have a strong effect on KOC. One

possibility is that when adsorbed on the clay, the polar functional groups of the humic acid are positioned

outward, so that they have an especially strong influence on the particle surface properties.

The sediments with the highest partition coefficients had humic acid structural characteristics indicating

mixed terrestrial and marine organic matter input (PV25, PV82, and 5A). While it has been suggested

that terrestrial humic acids bind aromatic hydrocarbons better due to their lignin precursors, and therefore

greater aromaticity, it was shown that aromaticity is not a predictor in this study. Rather, the data suggest

that the precursors to humic acids in these mixed environments are lower in polarity, that is, have less

carboxylic and amide carbon and lower N/C, contributing to the low polarity of these humic acids and

greater affinity for phenanthrene. However, another characteristic shared by natural humic acids with

greater partition coefficients was origin in subtidal sediments. The intertidal sediments of Lower Cook

Inlet probably contained more living organisms and fresh detritus than the subtidal sediments of Port

Valdez and the Beaufort Sea, and this could have contributed to their greater content of amide and

carboxylic acid functional groups.

Overall, the results of this work indicate that sediment organic content alone is not a reliable predictor

of phenanthrene adsorption when a broad range of sediment organic matter types are examined. The

differences in adsorption appear to be best explained by a measure of polar functional group content

(fraction carboxyl and amide carbon), but the 13CNMR data required for this parameter are costly and not

readily available in most laboratories. A simple N/C ratio captures much of the variability in humic acid

properties in this sample set, and may be a better candidate for routine assessments of sediment adsorptive

properties.
29

Chapter 3. Phenanthrene Adsorption to Mineral-Bound Humic Acid: Kinetics

and Influence of Previous Phenanthrene Adsorption2

Abstract

This study was comprised of two related components. One was an investigation of the kinetics of

phenanthrene adsorption to mineral-bound humic acids. The other focused on the effects previous

phenanthrene contamination of sediments had on subsequent phenanthrene adsorption. Natural humic

acids were isolated from coastal Alaska sediments, and represented a wide range of organic matter

properties. Several synthetic melanoidins were used as humic acid analogs. These humic acids or

melanoidins were coated onto a standard clay, and adsorption of phenanthrene was measured using a

radiotracer. Adsorption isotherms were linear up to a concentration of 5500 μg phenanthrene per gram

organic carbon. Isotherm slopes, interpretable as organic carbon normalized partition coefficients, KOC,

and initial adsorption rates were negatively correlated to the polarity of the organic matter. Partition

coefficients were found to be independent of initial phenanthrene concentration, indicating that the

binding sites were unlimited and uniform in strength. Initial rates were rapid, but the approach to steady

state was slow, requiring approximately one week, consistent with a slow diffusion within the sorbent.

Finally, the comparison of rates and partition coefficients for pristine sediments with those of sediments

already having adsorbed phenanthrene showed that previous phenanthrene adsorption had no effect.

Introduction

Sediments are a major reservoir of persistent environmental contamination by hydrophobic substances

such as polycyclic aromatic hydrocarbons (PAH), polychlorinated biphenyls (PCB) and pesticides

[Chin and Gschwend 1992; Hatzinger and Alexander 1995] because of the low aqueous solubility and

high surface affinity of these compounds. Adsorption has been investigated as a key process leading to

sediment contamination and its persistence in the environment. For uncharged organic pollutants, organic

matter is considered the primary sorbent component [Karickhoff and Morris 1985]. The adsorption

partition coefficient, Kp, increases with increasing organic content of sediments and soils [Means et al.

1980; Schwartzenbach and Westall 1981; Shaw and Terschak 1998]. The adsorption of neutral

hydrophobic organic compounds (HOC) to organic matter is generally considered to involve rapid

van der Waals type interactions [Brusseau and Rao 1991], although mechanisms for the binding of

PAH to organic matter have not been completely elucidated [Ragle et al. 1997].

Organic matter properties influence adsorption. Adsorption of various aromatic hydrocarbons has

correlated positively with the nonpolar–polar functional group ratio (Rutherford et al. [1992], Garbarini

and Lion [1986], and see Chapter 2) and aromaticity [Gauthier et al. 1987]. Adsorption of PAH to humic

acids is thought to result, at least in part, from weak dipole interactions between aromatic rings of the

PAH and those of humic acids, particularly humic acids derived from lignins and related materials of

woody terrigenous plants [Chin et al. 1997; Chin et al. 1994].

The sediment or soil organic matter, of which humic acid is a part, is a flexible, cross-linked, branched,

amorphous, polyelectrolytic, polymeric substance within which organic pollutants can diffuse [Brusseau

et al. 1991]. Schnitzer and Khan [1972] theorize that humic acid is an open structure with hydrophobic

cavities. The organic matter is porous as are clay mineral surfaces, but while the pores of sediment

mineral particles are fixed and rigid, the organic matter pores are dynamic and ephemeral [Brusseau et al.

1991]. The size, shape, and hydrophobicity of the pores within organic matter are sensitive to variations

 

2 Terschak, J.A., and S.M. Henrichs. Phenanthrene Adsorption to Mineral-Bound Humic Acid: Kinetics and

Influence of Previous Phenanthrene Adsorption. In preparation for submission to Environmental Science and

Technology.
30

in solution chemistry, and, therefore, changes in the ability to bind HOC can occur [Schlautman and

Morgan 1993]. The hydrophobicity of PAH suggests that PAH association is governed by diffusion into

the nonpolar environments of organic matter [Ragle et al. 1997; Gauthier et al. 1987; Schlautman and

Morgan 1993].

Investigations into the adsorption kinetics of HOC to natural organic matter have shown a rapid initial

uptake of the HOC followed by a slow approach to steady state. Wu and Gschwend’s [1986] adsorption

kinetics data of four congeners of chlorobenzene (pentachlorobenzene, 1,2,3,4-tetrachlorobenzene,

1,2,4-trichlorobenzene, 1,4-dichlorobenzene) to North River sediments showed that adsorption processes

were complete at times longer than two days. Their results for pentachlorobenzene adsorption to Charles

River sediments reached steady state around 16 hours. Chin and Gschwend [1992] examined pyrene

adsorption to Fort Point Channel sediments and found that it reached steady state sometime after

24 hours. The results from these studies indicate that HOC adsorption to natural organic matter generally

takes a minimum of one to two days to reach steady state.

Preliminary studies of this laboratory on PAH adsorption to natural humic acids found that previous

exposure to PAH may alter subsequent PAH binding (unpublished data). Hypothetically, the nonpolar

environments described above could increase in hydrophobic character when PAH are bound to the

organic matter, and further PAH adsorption could be enhanced by dipole–dipole interactions as proposed

for PAH–lignin interactions. Existing PAH contamination enhancing further PAH adsorption has been

observed by Boyd and Sun [1990], who reported that pentachlorophenol (PCP) adsorption to petroleum

contaminated environmental soils was greater than that estimated via modeled octanol–water partition

coefficients and soil organic content. In contrast, PCP adsorption to a Capac soil that was assumed to be

uncontaminated was consistent with the model, so they concluded that the petroleum contamination was

enhancing PCP adsorption. The effect of prior HOC contamination was also used to explain high KOC in

environmental samples by Chin and Gschwend [1992] and Kile et al. [1995].

On the other hand, partition coefficients would be expected to decrease with increasing PAH-loading

if there was very strong adsorption to a limited number of sites, with further adsorption having lower

binding energies, or if saturation of binding sites occurred. Several adsorption sites of varying strength

is one explanation for the common observation that PAH adsorption is partly irreversible [Hatzinger and

Alexander 1995; Yonge et al. 1985; Kan et al. 1994]. Other hypothesized mechanisms of irreversible

adsorption include slow diffusion into tortuous mineral particle micropores [Brusseau et al. 1991; Wu

and Gschwend 1986; Steinberg et al. 1987] and diffusion into the organic matrix [Karickhoff and Morris

1985; Brusseau and Rao 1991; Brusseau et al. 1991; Schlautman and Morgan 1993; Kan et al. 1994].

Intraorganic matter diffusion is thought to be the most likely mechanism for HOC adsorption [Brusseau

and Rao 1991; Brusseau et al. 1991]. Adamson [1990] suggests that irreversible adsorption is probably

due to a mechanical or structural rearrangement of the adsorbent, that is, the matrix from which

desorption takes place is different from that during adsorption.

This research was conducted as part of an examination of the role humic acids play in the adsorption and

desorption of PAH by the sediments of coastal marine environments. The adsorbents investigated were

extracted humic acids bound to a standard clay mineral, in order to eliminate variability due to the mineral

phase. This study addresses the rates and mechanisms of phenanthrene adsorption to humic acid and

melanoidin-coated clay, and is the first to specifically test the effects of previous PAH contamination on

subsequent binding. Melanoidins were used as model marine humic acids because of their lack of initial

PAH contamination relative to their natural counterparts. The melanoidins were compared with natural

humic acids isolated from coastal Alaska sediments. These organic substances represent a wide range of

properties potentially affecting associations with PAH.
31

Experimental Section

Substrate preparation

A montmorillonite standard (A.P.I. # 26, 49 E 2600 from Clay Spur, Wyoming; Ward’s Natural Science

Establishment, Inc.) was ground for 2 minutes using a shatter box equipped with carbide rings. The

resulting powder was passed through a 270 mesh (53 μm) sieve using a Ro-Tap® apparatus. Organic

material was removed from the clay by oxidizing it with 30% hydrogen peroxide. A slurry of clay and

H2O2 was prepared in an open Pyrex container. The clay was mixed for 12 hours at room temperature.

Low heat was then applied, bringing the slurry to 35–40 °C. Fresh H2O2 was added as needed. Stirring

was continuous until no evolution of CO2 gas was evident. The slurry was then centrifuged at 14,000 g

for 2 hours and the clear supernatant was discarded. A wash of organic-free, glass-distilled water was

used and the clay was again centrifuged as above. A small sample of the organic-free clay was submitted

for total organic carbon analysis and the remaining material was stored under ultra high purity nitrogen

gas (UHP-N2) gas at 5–10 °C.

Sediment subsamples were analyzed for size distribution of particles using the sieve-pipette method

[Jackson 1965]. The sediment was divided into six fractions: fine sand > 63 μm, very fine sand

(63–50 μm), coarse silt (50–20 μm), medium silt (20–5 μm), fine silt (5–2 μm), and clay (< 2 μm).

Fine sand and larger particles were removed from the sediment by sieve. In a 1-L graduated cylinder,

a suspension of the material that passed through the sieve was made in 1 L of distilled water and allowed

to settle for a prescribed time based on Stokes’ Law [Jackson 1965]. Particle size distribution was

calculated as a percentage of the recovered material (99.99% recovery).

Humic acid sources

Natural humic acids were taken from sediment samples collected at three separate coastal locations in

Alaska (Table 7). Intertidal sediment samples from Lower Cook Inlet were collected in July 1995, July

1996, and May 1999. Sampling occurred during spring tides below the mean lower-low water level.

Subtidal sediment samples were collected from the Port Valdez region in August 1996 using a Haps

corer and from the Beaufort Sea region during September 1997 using a Kynar-coated van Veen grab

sampler. Any overlying debris was removed and then sediment was collected from the oxic layer (upper

2 cm) with metal implements that had been heated to redness before use. All samples were stored until

needed at –50 °C in pre-combusted glass jars.

Table 7. Site locations of sediments discussed in this study; n/a – not applicable, as Lower

Cook Inlet samples were intertidal and collected during low tide.

Location Station Latitude (N) Longitude (W) Depth (m)

Port Valdez PV25 61° 05.5′ 146° 23.3′ 73

Port Valdez PV82 61° 05.4′ 146° 22.3′ 79

Lower Cook Inlet HBH 59° 36.3′ 151° 25.0′ n/a

Lower Cook Inlet JB2 59° 27.1′ 151° 29.3′ n/a

Lower Cook Inlet TB2 59° 24.8′ 151° 17.0′ n/a

Beaufort Sea 5A 70° 29.7′ 148° 46.0′ 11.4
32

Humic acid was isolated using the alkali (0.5 M NaOH) extraction procedure of Anderson and Schoenau

[1983]. After extraction from approximately 30 g of wet sediment, humic acid was precipitated by the

addition of 6 M HCl until a pH of 1.5 was obtained. The precipitated extract was allowed to stand for

15 minutes and was then centrifuged for 30 minutes at 14,000 g. The precipitated humic acid was

lyophilized at –85 °C for 24 hours. The resulting dry humic acid was stored at room temperature in a

glass vial under N2(g).

Artificial humic acids were prepared in the laboratory by refluxing various ratios of glucose and bovine

casein in buffered solutions for 24 to 48 hours as described by Yamamoto and Ishiwatari [1989] The

buffer was composed of 125 ml 0.1 M KH2PO4, 75 ml 0.1 M NaOH and 50 ml of glass-distilled water

to give a final pH of 7.0. After the prescribed reaction time, the mixture was allowed to cool with stirring,

then transferred to 250 ml Nalgene® centrifuge bottles. The artificial humic acid was isolated using the

extraction procedure above.

Substrate coating

Humic acids were coated to the montmorillonite substrate for use in the adsorption/desorption

experiments. Approximately 1 g of lyophilized humic acid was dissolved in 100 mL of organic-free

water and brought to a pH of 10 with sodium hydroxide. The solution was mixed under an atmosphere

of UHP-N2 for 3 to 7 days with adjustments made daily to maintain the pH. Once dissolved, the humic

acid solution was added to the washed clay and mixed for 24 hours. The suspension was transferred to

a centrifuge bottle containing the crystallized salts resulting from the evaporation of an equal amount of

artificial seawater and mixed for 48 hours. The suspension was then centrifuged at 14,000 g for 2 hours

and the supernatant was discarded. A sample of the humic acid coated clay was dried and submitted for

TOC analysis to determine the extent of humic acid adhesion. After flushing the headspace with UHP-N2

gas, the humic acid coated clay was stored in a refrigerator at 5–10°C.

Total organic carbon measurements

Total organic carbon analyses of the clay substrates were performed by Donald Schell’s laboratory at the

University of Alaska Fairbanks using a Delta Plus mass spectrometer equipped with a Finnigan MAT

Conflo II interface and a Carlo Erba elemental analyzer.

Adsorption experiments

All glassware and Teflon® liners were treated to remove any hydrocarbon contamination before use.

Glassware was baked in a muffle furnace at 450 °C for at least 8 hours. Teflon® cap liners were soaked

in chromic acid for 20 minutes and then rinsed 3 times with clean organic-free water.

Artificial seawater was prepared from organic-free water that had been glass distilled over a saturated

solution of potassium permanganate. To each liter of water the following salts were added: 23.260 g

NaCl, 10.636 g MgCl2 •6H2O, 3.918 g anhydrous Na2SO4, 1.102 g CaCl2, 0.664 g KCl, 0.192 g NaHCO3,

0.096 g KBr, and 0.026 g H3BO3 [Lyman and Fleming 1940]. Also, 0.500 g HgCl2 was added as an

antibiologic agent.

A stock phenanthrene solution was prepared by dissolving 0.0060 g of phenanthrene in a 100-mL

volumetric flask with acetonitrile. This solution was then used to prepare the various phenanthrene

solutions required for the adsorption experiments. After evaporation of the acetonitrile, aliquots of the

phenanthrene stock solution were diluted with the required amount of artificial seawater to make final

concentrations of 50, 300, and 700 μg L–1. Radiolabeled phenanthrene solutions were prepared using a

similar technique. However, 10% of the phenanthrene was replaced with [9-14C] phenanthrene (5–15 mCi

mmol–1; Sigma Chemical Co.) dissolved in acetonitrile. All solutions were prepared 24 hours in advance
33

and stored in a refrigerator (5–10 °C). Previous experiments within this laboratory have shown that for

concentrations of phenanthrene ranging from 10 to 750 μg L–1, there was no loss of hydrocarbon due to

evaporation or adhesion to the walls of the container [Henrichs et al. 1997]. In addition, all experiments

had control vials containing no sediment.

Adsorption experiments began by weighing 0.1g of wet, humic-coated clay into 2-dram vials with

Teflon® lined caps. Five milliliters of artificial seawater solutions containing either phenanthrene (for

adsorption experiments with PAH-loaded sediments) or radiolabeled phenanthrene (for adsorption

experiments with pristine sediments) at several concentrations (50, 300, or 700 μg L–1) were added to

the vials. A 15-second agitation on a vortex mixer thoroughly mixed the clay and solution. The vials

were then placed on a table shaker at 150–200 rpm. After adsorption for 1 h, 1 d, 3 d, 7 d, or 14 d, the

vials were centrifuged at 2400 g for 30 minutes and a 1.0 mL aliquot of each of the radiolabeled reaction

solutions was removed and scintillation counted to determine phenanthrene adsorption to pristine

sediments. The remaining supernatant, including that from all vials containing phenanthrene without

radiolabel, was discarded. The second adsorption segment of the experiment began with the replacement

of the nonlabeled supernatant with fresh radiolabeled 300 μg L–1 phenanthrene solution to determine the

adsorption of phenanthrene to sediments previously loaded with phenanthrene. The vials were vortex

mixed as above and returned to the table shaker at 150–200 rpm. The second adsorption was carried

out for 1, 3, 7, 14, 30, 60, or 90 days. After the adsorption time was completed, a group of vials was

centrifuged as above and a 1.0 mL aliquot was taken from each supernatant and scintillation counted.

Upon the completion of the second adsorption, the supernatant was discarded and the sediments were

placed in a 65 °C oven until dry. Sediment dry weights were measured and pore water volumes were

calculated from the weight loss upon drying.

The calculation to determine dissolved concentrations of phenanthrene remaining in solution after

prescribed reaction times with pristine sediments is shown in Equation 3–1, where [diss] is the amount

of phenanthrene remaining in solution in μg L–1, dpm is the disintegrations per minute of the radiolabeled

phenanthrene solutions by scintillation counting, and [phen]i is the initial phenanthrene concentration.

[] [ ] diss dpm

dpm phen dissolved

control

i =

As shown in Equation 3–2, the adsorption of phenanthrene to pristine sediments was calculated as the

amount of the phenanthrene removed from the initial solution as a function of dissolved phenanthrene

activity, where [ads] is the amount of phenanthrene adsorbed to the sediment in μg•g OC–1, a is the

solution volume, b is dry weight of the clay substrate, and OC is the fraction of organic carbon in grams

of carbon per 1 g of clay.

[] [ ] ± ads dpm dpm

dpm phen a

b

OC control dissolved

control

i = ⎛

⎝

⎞

⎠

− 1

The adsorption of phenanthrene to PAH-loaded sediments was calculated assuming reversible adsorption

using Equations 3–1 and 3–2 for dissolved and adsorbed concentration, respectively. However, [phen]i

becomes the total amount of phenanthrene in the system, expressed as [phen]t (Equation 3–3), that is, the

second phenanthrene solution (300 μg L–1) together with the amount of phenanthrene adsorbed in the first

step ([ads]1).

(3–2)

(3–1)
34

[ ] [] phen g

L

ads OC b

a t = + ⎛

⎝

⎞

⎠ 300 1

μ

In both situations (phenanthrene adsorption to pristine sediments and adsorption to PAH-loaded

sediments), the ratio of adsorbed to dissolved phenanthrene concentrations was reported as a partition

coefficient normalized to organic carbon content (KOC), as shown in Equation 3–4.

K dpm dpm

dpm

OC a

b OC

control dissolved

dissolved

= ⎛

⎝

⎞

⎠

− ±1

Results

The particle size distribution for the montmorillonite clay used in the adsorption experiments is shown

in Figure 11. The particles were predominantly clay, followed by lesser quantities of medium silt, fine

silt, and coarse silt. The organic carbon content of the clay after hydrogen peroxide treatment was

≤ 0.095 wt %.

coarse silt

medium silt

fine silt

clay

> 20 μm

20–5 μm

5–2 μm

91.57% < 20 μm

1.85% 4.19%

2.39%

Figure 11. Montmorillonite particle size distribution.

Analyses were performed to confirm that humic acid was, indeed, bound to the clay after the coating

process and that humic acid was not removed from the clay during the course of the adsorption

experiments. Initial concentrations of organic matter were similar to those found naturally in marine

sediments. The organic carbon contents of the clays after the adsorption experiments were not statistically

different from those measured before the experiments.

Fourteen-day adsorption isotherms were found to be linear up to 5500 μg phenanthrene•g OC–1

(r

2 > 0.97) (Figure 12). Intercepts of the y-axis were not significantly different from 0 μg phenanthrene•

gram sediment–1, so linear regressions were forced through the origin. Slopes (interpretable as partition

coefficients, KOC) were found to range from 5872 to 17,699 mL•g OC–1. Slopes associated with natural

humic acids (m ≥ 10,750 mL•g OC–1) were consistently larger than those for synthetic humic acids

(m ≤ 8370 mL•g OC–1).

(3–3)

(3–4)
35

1:10–24

1:10–48

10:1–48

5A

HBH

JB2

TB2

1:10–24

1:10–48

10:1–48

5A

HBH

JB2

TB2

0.05

[diss]1 (μg mL–1)

0.00 0.10 0.15 0.25 0.30

0

1000

[ads]1 (μg•g OC–1)

2000

3000

4000

5000

6000

0.20 0.35 0.40

7000

Figure 12. Linear adsorption of phenanthrene to pristine sediments (without Port Valdez

samples). Initial adsorption time was 14 d. Initial phenanthrene concentrations

were 50, 300, and 700 μg L–1.

A comparison between phenanthrene adsorption to pristine sediment ([ads]1) and adsorption to PAHloaded sediment ([ads]2) can be seen in Figure 13. Adsorption coefficients for pristine sediments

measured at 1 hour, and at 1, 3, 7, and 14 days for initial phenanthrene concentrations of 50, 300, and

700 μg L–1, range from 99.5 to 5722.4 μg•g OC–1. Corresponding adsorption coefficients for PAH-loaded

sediments after 14 days at 300 μg L–1 phenanthrene concentration, were larger, especially at lower values

of [ads]1, ranging from 786.3 to 5292.9 μg•g OC–1.

Initial rates of phenanthrene adsorption to pristine sediments (Rate1) were calculated for 1 hour and

1 day of reaction time (Tables 8a and 8b). Figure 14 shows the initial rate to be directly proportional to

the initial dissolved concentration of phenanthrene. Overall, the rates ranged from 122 μg•g OC–1 •d–1

for a synthetic humic acid, 1:10–24, at 50 μg L–1 phenanthrene to 4656 μg•g OC–1 •d–1 for a natural

humic acid, 5A, at 700 μg L–1 phenanthrene. Mineral particles coated with natural humic acids generally

displayed faster initial rates of adsorption than did synthetic humic acids (PV82 > JB2 > PV25 > 5A >

TB2 > 1:10–48 > HBH > 10:1–48 > 1:10–24), corresponding to their greater KOC. Initial rates, evaluated

after 1 day and at 300 μg L–1 phenanthrene concentration, were found to be significantly negatively

correlated with proxies for humic acid polarity (N/C: r

2 = 0.54, t = 2.867, n = 9; fraction carboxyl and

amide carbons: r

2 = 0.59, t = 3.174, n = 9) (see Chapter 2). The rate of phenanthrene adsorption at 50,

300, and 700 μg L–1 to sediment with previously adsorbed phenanthrene (14 day reaction time) was

greater (Figure 15) than that to pristine sediment, as was seen for [ads]2, above.
36

1:10–24

1:10–48

10:1–48

5A

HBH

JB2

TB2

PV25

PV82

1000

[ads]1 (μg• g OC–1)

0 2000 3000

0

[ads]2 (μg•g OC–1)

1000

2000

3000

4000

5000

4000 5000 6000

6000

Completely Reversible Adsorption (HBH)

Partially Reversible Adsorption (HBH)

Partially Reversible Adsorption (10:1–48)

Completely Reversible Adsorption (10:1–48)

Figure 13. Phenanthrene adsorption to pristine sediments ([ads]1, contact time of 1 h, and 1, 3,

7, and 14 d) and to PAH-loaded sediments ([ads]2, reaction time of 14 d).

1:10–24

1:10–48

10:1–48

5A

HBH

JB2

TB2

1:10–24

1:10–48

10:1–48

5A

HBH

JB2

TB2

100

[phenanthrene]initial (μg L–1)

0 200 300 500 600

0

Initial Rate (μg• g OC–1 •d–1)

1500

2000

3500

4000

4500

400 700 800

5000

2500

3000

500

1000

Figure 14. Initial rate of phenanthrene adsorption as a function of initial phenanthrene concentration

to pristine sediment (without Port Valdez samples). Reaction time was 1 day.
37

Table 8a. Initial rates of phenanthrene adsorption at various concentrations based on

1 hour initial time increments. Initial rates at 1 hour were determined only

for those sites listed, since adsorption was not measured at 1 hour for the

others. Concentrations are for micrograms of phenanthrene per liter of

solution at the start of the reaction. Rates of adsorption are on a per gram

of organic carbon basis (g OC).

Sample Initial Rate 50 μg L–1

(μg• g OC–1 •d–1)

Initial Rate 300 μg L–1

(μg• g OC–1 •d–1)

Initial Rate 700 μg L–1

(μg• g OC–1 •d–1)

1:10–24 99.5 461.4 1542.7

10:1–48 127.3 805.7 1874.4

JB2 306.9 1543.8 3599.3

Table 8b. Initial rates of phenanthrene adsorption at various concentrations based

on 1 day initial time increments. Concentrations are for micrograms of

phenanthrene per liter of solution at the start of the reaction. Rates of

adsorption are on a per gram of organic carbon basis (g OC). n/d – not

determined

Sample Initial Rate 50 μg L–1

(μg• g OC–1 •d–1)

Initial Rate 300 μg L–1

(μg• g OC–1 •d–1)

Initial Rate 700 μg L–1

(μg• g OC–1 •d–1)

1:10–24 121.6 706.3 1512.5

10:1–48 131.0 838.1 1739.6

1:10–48 188.4 1173.3 2593.1

HBH 177.6 1156.1 2711.6

TB2 343.4 1665.9 n/d

PV25 n/d 2361.4 n/d

JB2 358.5 2386.2 4102.0

PV82 n/d 2386.6 n/d

5A 334.9 1845.2 4655.7
38

1:10–24

1:10–48

10:1–48

5A

HBH

JB2

TB2

1000

Initial Rate to Pristine Sediments (μg• g OC–1 •d–1)

0 2000 3000

0

1000

2000

3000

4000

5000

4000 5000

6000

6000

Initial Rate to PAH-Loaded Sediment (μg• g OC–1 •d–1) Figure 15. Comparison of initial rates of phenanthrene adsorption to pristine

sediments and to PAH-loaded sediments. Reaction times were 1 day. The

solid line describes completely reversible adsorption, and the dashed line

describes partially reversible adsorption.

Organic carbon-normalized partition coefficients (KOC) did not vary significantly for the different

initial phenanthrene concentrations (50, 300, and 700 μg L–1). Eighty percent (60 of 75) of the pairings

were statistically the same at p = 0.05. In cases where KOC differed, there was no consistent pattern with

increasing concentration.

Figure 16a shows the variation in time of partition coefficients for both pristine and PAH-loaded

sediments of sample 5A. This sample is typical in that 7 samples of 9 (5A, HBH, 10:1–48, 1:10–48, TB2,

PV25, PV82) display the same trend, with KOC2 being greater than KOC1 at early time points. In 4 of the 7

cases (5A, HBH, 10:1–48, PV82), the partition coefficients become approximately equal at 7 to 14 days.

Figure 16b, for the artificial humic acid 1:10–24, shows results contrary to most of the samples, with KOC1

> KOC2 initially, and is presented for contrast and comparison. In 6 of 9 cases (5A, HBH, 10:1–48, JB2,

1:10–24, PV82), there are large changes in KOC1 within the first 7 days of adsorption, but most reactions

reach steady state by 1 week. A few (TB2, 1:10–48, PV25) appear to reach steady state after only 1 day,

and their partition coefficients did not vary significantly from that point on.
39

KOC1

KOC2

20

14C-Phenanthrene Contact Time (days)

0 40 60

0

5000

10000

15000

20000

25000

80 100

KOC (mL• g OC–1) Figure 16a. Partition coefficients for phenanthrene adsorption to JB2. KOC1 and KOC2

describe partition coefficients for pristine and PAH-loaded sediments,

respectively. KOC2 was determined for sediments after 14 d of PAH-loading.

20

14C-Phenanthrene Contact Time (days)

0 40 60

0

2000

6000

8000

10000

12000

80 100

KOC (mL• g OC–1)

4000

KOC1

KOC2

Figure 16b. Partition coefficients for phenanthrene adsorption to 1:10–48. KOC1 and KOC2

describe partition coefficients for pristine and PAH-loaded sediments,

respectively. KOC2 was determined for sediments after 14 d of PAH-loading.
40

Discussion

The particle sizes of the ground montmorillonite were mostly clay, giving a large surface area-to-volume

ratio. Montmorillonite was chosen as the standard material matrix because of its uniformity and its ability

to adsorb humic acid. By removing the variability due to different substrate (i.e., natural sediment)

properties, the intrinsic phenanthrene-adsorptive characteristics of the humic acid, alone, could be

elucidated. The analyses of clays after the humic adsorption step indicated the carbon contents were

comparable to those of normal marine sediments. Further, the humic clay association was stable under

the experimental conditions employed, since percent OC did not change over time. Hence, the sorbents

used in this study were good analogs of natural sediments, but controlled the variability of the mineral

phase present in natural sediments and soils.

Reaction time was a minor factor in controlling the amount of phenanthrene adsorbed by pristine

sediment on an organic carbon content normalized basis, while the initial concentrations of phenanthrene

in solution (50, 300, or 700 μg L–1) were the primary controlling factor. Linear adsorption was found

for both the natural and synthetic humic acids of this study. Chiou et al. [1998] also found phenanthrene

adsorption, up to 600 mg kg–1 substrate, to be linear in soils from Oregon and Illinois and oxic sediments

of the Lower Mississippi River and Massachusetts Bay. Similarly, Kan et al. [1994] found linear

adsorption for phenanthrene in sediments from a small river flood plain in Oklahoma. If phenanthrene

adsorption sites were limited in number and the adsorbed phenanthrene saturated these sites within the

experiments’ concentration ranges, higher concentrations of phenanthrene would have lower partition

coefficients. A similar result would be expected if there were a limited number of strong sites, even

though weaker adsorption sites were numerous. However, KOC varied little with concentration, leading

to the inference that the organic matrix contained an unlimited number of sites with uniform strength.

The slopes of the individual linear adsorption isotherms were related to the properties of the humic acids

discussed in Chapter 2. Generally, lower KOC were found for humic acids with high N/C and fraction

of carboxyl and amide functional groups. Rutherford et al. [1992] found similar relationships between

adsorption isotherm slopes of benzene and carbon tetrachloride and the properties of organic matter

extracted from natural soils. The phenanthrene KOC values for natural humic acids in this study are of the

same magnitude as those found by Chiou et al. [1998] using various North America sediment sources

(21,000 to 44,000 mL•g OC–1) and those found by Kan et al. [1994] using Oklahoma flood plain sediment

(12,000 mL•g OC–1).

In Figure 13, which compares adsorption properties of pristine sediments to those of sediments that

were pre-loaded with phenanthrene, two pairs of theoretical lines are presented. The lines shown are for

the HBH humic acid sample (see Chapter 2), chosen because it had the median adsorption of natural

sediments in terms of partition coefficient as well as rate constant (expressed as L•g OC–1 •d–1), and for

sample 10:1–48, which was typical of the melanoidins. Both lines in each pair assume that KOC1 = KOC2.

One line (solid), for complete reversibility, is based on the premise that phenanthrene adsorbed to the

sediment is freely able to dissociate from the organic matrix, limited only by the rate of diffusion. The

other line (dashed), describing partial reversibility, is based on the premise that only a portion of the

phenanthrene adsorbed to the sediment is free to redissolve, while 20% of the phenanthrene is irreversibly

bound. That phenanthrene could either be trapped within the organic matrix, due to changes in the

conformational structure of the matrix, or be associated with sites having unusually large binding energies

for phenanthrene. However, the linear adsorption isotherms indicate that the second possibility is

unlikely. In either the wholly reversible or partly reversible case, some phenanthrene desorbed from the

organic matrix and was added to the total dissolved phenanthrene pool in the second experiment, in which

an aliquot of 300 μg L–1 14C-phenanthrene was added to the PAH-loaded sediments. This, in turn, caused

the adsorbed phenanthrene concentration after the second phenanthrene addition to be greater, even in the

absence of any change in KOC or any irreversible adsorption. For total phenanthrene adsorbed to the
41

sediments on a per unit organic carbon basis, the influence of reversibility can be seen by comparing

the theoretical lines to the observed data for HBH. Most of the humic acids were similar, in that the

theoretical line describing complete reversibility fitted the data better than the line for partial reversibility.

In this study, the possibility was considered that the greater [ads]1 of humic acids could be due, in part,

to their previous exposure to environmental PAH. Even though the extraction of humic acids from natural

sediments was a fairly vigorous process, natural PAH may have been retained by the humic acid and

carried over into the laboratory controlled experiments. The environmental phenanthrene could affect

[ads]1 much like [ads]1 influenced [ads]2. Total aromatic content (TARO) of the natural sediments (see

Chapter 2) was weakly correlated with [ads]1 (p = 0.10, n = 92, α = 1.665, t = 0.17779). However, the

concentrations of TARO in natural sediments were small, and should not influence [ads]1 because

the potential desorbed phenanthrene concentration was much less than concentrations used in the

experiments. The effect seen may, in fact, be due to just the greater partition coefficient and, hence,

a greater affinity for environmental PAH.

The initial rates of PAH adsorption in this study were rapid, with a slow approach to steady state

occurring after 24 hours (Figures 14 and 16a). The linear relationship between phenanthrene

concentration and initial adsorption rate (Figure 14) for each humic acid is consistent with a diffusion

process. Adsorption was nearly complete only after three days in most cases, as opposed to time scales

on the order of minutes or hours, which would be expected of a process occurring solely at the particle

surface. Similar patterns of nonionic organic compound adsorption have been observed by others [Wu

and Gschwend 1986; Leenheer and Ahlrichs 1971; Karickhoff 1980]. The slow approach to steady state

is indicative of diffusion of HOC into a three-dimensional organic matrix. Kan et al. [1994] reported

phenanthrene adsorption to Oklahoma flood plain sediment reaching equilibrium in one to four days,

while Allen-King et al. [1995] reported perchlorethene adsorption to sediments of Ontario, Canada,

reached equilibrium by three days. Wu and Gschwend [1986] described a diffusion model in which initial

rapid HOC diffusion into pore fluids of the interstitial spaces between particle aggregates is followed by

much slower microscale partitioning into the organic matter. It is interesting that, in this study, the pattern

was seen in the sediments coated with reconstituted humic acids, and argues that the phenomenon is a

general characteristic of a three-dimensional structure that humic acids acquire when adsorbing to clay

minerals. It is not only true of associations that form in the environment.

In Figure 15, initial rates of adsorption of phenanthrene to sediments pre-loaded with PAH are compared

to the initial rates of phenanthrene adsorption to pristine sediments. On the figure, two theoretical lines

are plotted; one represents complete reversibility of phenanthrene adsorption (solid), and the other

represents partially reversible adsorption of phenanthrene (dashed). The partially reversible line is based

on an average of desorption partition coefficients for HBH, as was done earlier for the [ads2] calculations.

As illustrated by the HBH example, most data points fall near, but slightly above, theoretical lines

describing completely reversible adsorption. This indicates that a small fraction of the initial

phenanthrene adsorption was not reversible under the conditions of these experiments.

The points in Figure 16a also show evidence of slow adsorption. As seen, KOC2 was greater than KOC1

initially for most of the humic acids, but the two became similar later. This is expected, since the initially

adsorbed phenanthrene takes some time to exchange with the solution during desorption from the organic

matter, according to its diffusion rate. In the two cases where KOC1 was initially greater than KOC2

(samples JB2 and 1:10–24), an examination of their desorption from pristine sediments over 14 days

(Figure 16b) shows that it took much longer to achieve steady state (possibly greater than 14 days) than

for the other humics. The diffusion of adsorbed phenanthrene back into the solution phase would likewise

be slow. So, at the initial time points of the investigation into adsorption to PAH-loaded sediments, the

previously adsorbed phenanthrene contributes less to the available phenanthrene pool in solution as it

desorbs, leading to smaller apparent KOC2.
42

The range of organic matter properties and loadings examined in this study spans a large part

of the natural variability seen in the environment. The rates of adsorption, ranging from 100 to

4700 μg•g OC–1 •d–1 are not different from those found for intact sediments [Henrichs et al. 1997]. From

this, it follows that the conclusions drawn from this data may be applied to natural sediments and other

environmental samples. The slow approach to steady state indicates that the rate of adsorption is primarily

controlled by diffusion within the organic matrix, and investigations into PAH adsorption may need to

last at least seven days for steady state to be reached. Once within the organic matrix, phenanthrene

binding is almost entirely reversible, but this too is a slow process requiring one to more than 14 days

for completion.

Finally, the earlier reports [Chin and Gschwend 1992; Boyd and Sun 1990; Kile et al. 1995] that

subsequent adsorption is enhanced in previously contaminated sediments is not supported by the findings

of this study. Except for the synthetic humic acid 1:10–48, the adsorption of phenanthrene to humic acids

previously contaminated with PAH was not affected by the adsorbed PAH already present in the organic

matrix.
43

Chapter 4. Observed Desorption Kinetics of Phenanthrene from Mineral-Bound

Humic Acids: Consequences of Conformational Changes 3

Abstract

This study addresses the rates and mechanisms of phenanthrene desorption from marine humic acids

and melanoidins coated onto clay. The humic acids were extracted from Alaska coastal sediments and

the melanoidins synthesized from glucose and bovine casein. These organic substances represent

a wide range of properties potentially affecting associations with polycyclic aromatic hydrocarbons.

Montmorillonite was used as a common mineral substrate in order to eliminate variability due to the

mineral phase. The desorption of phenanthrene was measured using a radiotracer. The extent of

desorption was less for both increasing phenanthrene concentrations used in the initial adsorption phase

of the experiment and shorter adsorption reaction time. Desorption steady state was reached after three

to seven days for all of the humic acids studied in this experiment, reflecting slow diffusion processes

within the organic matrix. Initial rates were found to be directly proportional to the initial concentration

of phenanthrene used during the adsorption phase, and decreased with the carbon content of the humic

acid coated clay. Desorption was not related to humic acid structural characteristics such as aliphaticity,

aromaticity, percent carboxylic, amidic, or carbonyl carbons measured by 13C nuclear magnetic

resonance spectroscopy. Decreased desorption with increasing adsorbed phenanthrene concentration

was consistent with the interpretation that conformational changes to the humic acid structure occurred

during the adsorption and diffusion of phenanthrene into the organic matrix.

Introduction

Sediments are a major reservoir of persistent environmental contamination by hydrophobic substances

such as polycyclic aromatic hydrocarbons (PAH), polychlorinated biphenyls (PCB) and pesticides [Chin

and Gschwend 1992; Hatzinger and Alexander 1995], because of the low aqueous solubility and high

surface affinity of these compounds. Adsorption has been investigated as a key process leading to

persistent sediment contamination in the environment. It is believed that hydrophobic organic compounds

(HOC) can either become entrapped within the solid micropores or between the silicate layers of

montmorillonite [Lahlou and Ortega-Calvo 1999] or partition into the organic matter [Karickhoff and

Morris 1985; Brusseau et al. 1991]. For uncharged organic pollutants, organic matter is considered the

primary sorbent component [Karickhoff and Morris 1985].

Desorption processes have a significant bearing on the redistribution of hydrocarbon pollutants and their

uptake by marine organisms. They have, however, received less attention than adsorption processes [Kan

et al. 1994]. While theories exist to explain adsorption of hydrophobic organic compounds to organic

matter [Karickhoff and Morris 1985; Brusseau et al. 1991], understanding of desorption presents many

challenges that are only just beginning to be explored. The reversibility of HOC adsorption is the subject

of much debate in the literature, with reports of adsorption/desorption hysteresis [Kan et al. 1994 and

references therein] due to the sequestration of HOC into a non-labile fraction. There are also reports that

adsorption of HOC is completely reversible [Wu and Gschwend 1986; Borglin et al. 1996], with the

appearance of hysteresis actually being an artifact of slow sorption kinetics [Kan et al. 1994 and

references therein].

Studies have attributed slow sorption kinetics to the complex pore geometry of particle matrices

[Pavlostathis and Mathavan 1992]. The sediment or soil organic matter, of which humic acid is a part,

is a flexible, cross-linked, branched, amorphous (noncrystalline), polyelectrolytic, polymeric substance

 

3 Terschak, J.A., and S.M. Henrichs. Observed Desorption Kinetics of Phenanthrene from Mineral-Bound Humic

Acids: Consequences of Conformational Changes. In preparation for submission to Environmental Science and

Technology.
44

within which organic pollutants can diffuse [Brusseau and Rao 1991]. Humic acid has been described as

an open structure with hydrophobic cavities [Schnitzer and Khan 1972]. The organic matter is porous, and

in contrast to the fixed, rigid pores of sediment mineral particles, the organic matter pores are dynamic

and ephemeral [Brusseau and Rao 1991]. The size, shape, and hydrophobicity of the pores within the

organic matter are sensitive to variations in solution chemistry (e.g., ionic strength), and, therefore,

changes in the ability to bind HOC can occur [Schlautman and Morgan 1993]. The hydrophobicity of

PAH suggests that PAH association is governed by diffusion into nonpolar environments of the organic

matter [Schlautman and Morgan 1993; Ragle et al. 1997; Gauthier et al. 1987].

Investigations into the adsorption kinetics of HOC to natural organic matter have shown a rapid initial

uptake of the HOC followed by a slow approach to steady state [Chin and Gschwend 1992; Wu and

Gschwend 1986]. Slow adsorption kinetics have been shown to be correlated with humic acids possessing

more rigid structural components [Schlebaum et al. 1998]. This leads to the conclusion that the labile

fraction of humic acids is composed of relatively open humic structures while the non-labile fraction is

composed of more condensed, rigid humic structures [Schlebaum et al. 1998 and references therein].

This chapter addresses the rates and mechanisms of phenanthrene desorption from humic acids and

melanoidin-coated clay. This study was conducted as part of an examination of the role humic acids play

in the adsorption and desorption of PAH by sediments of coastal marine environments. The sorbents

investigated were humic acids that were extracted from Alaska coastal sediments, then bound to a

standard clay mineral, in order to eliminate variability due to the mineral phase. Melanoidins were used

as model marine humic acids because of their lack of initial PAH contamination (see Chapter 2), and

because they extend the range of structural properties within this set of humic acids. These organic

substances represent a wide range of properties potentially affecting associations with PAH.

Experimental Section

Substrate preparation

A montmorillonite standard (A.P.I. # 26, 49 E 2600 from Clay Spur, Wyoming; Ward’s Natural Science

Establishment, Inc.) was ground for 2 minutes using a shatter box equipped with carbide rings. The

resulting powder was passed through a 270 mesh (53 μm) sieve using a Ro-Tap® apparatus. Organic

material was removed from the clay by oxidizing it with 30% hydrogen peroxide. A slurry of clay and

H2O2 was prepared in an open Pyrex container. The clay was mixed for 12 hours at room temperature.

Low heat was then applied, bringing the slurry to 35–40 °C. Fresh H2O2 was added as needed. Stirring

was continuous until no evolution of CO2 gas was evident. The slurry was then centrifuged at 14,000 g

for 2 hours and the clear supernatant was discarded. A wash of organic-free, glass-distilled water was

used and the clay was again centrifuged as above. A small sample of the organic-free clay was submitted

for TOC analysis and the remaining material was stored under ultra high purity nitrogen gas (UHP-N2)

gas at 5–10 °C.

Sediment subsamples were analyzed for size distribution of particles using the sieve-pipette method

[Jackson 1965]. The sediment was divided into six fractions: fine sand > 63 μm, very fine sand

(63–50 μm), coarse silt (50–20 μm), medium silt (20–5 μm), fine silt (5–2 μm), and clay (< 2 μm).

Fine sand and larger particles were removed from the sediment by sieve. In a 1-L graduated cylinder,

a suspension of the material that passed through the sieve was made in 1 L of distilled water and allowed

to settle for a prescribed time based on Stokes’ Law [Jackson 1965]. Particle size distribution was

calculated as a percentage of the recovered material (see Chapter 3, Figure 11).
45

Humic acid sources

Natural humic acids were taken from sediment samples collected at three separate coastal locations in

Alaska (see Chapter 3, Table 7). Intertidal sediment samples from Lower Cook Inlet were collected in

July 1995, July 1996, and May 1999. Sampling occurred during spring tides below the mean lower-low

water level. Subtidal sediment samples were collected from the Port Valdez region in August 1996 using

a Haps corer and from the Beaufort Sea region during September 1997 using a Kynar-coated van Veen

grab sampler. Any overlying debris was removed and then sediment was collected from the oxic layer

(upper 2 cm) with metal implements that had been heated to redness before use. All samples were stored

until needed at –50 °C in pre-combusted glass jars.

Humic acid was isolated using the alkali (0.5 M NaOH) extraction procedure of Anderson and Schoenau

[1983]. After extraction from approximately 30 g of wet sediment, humic acid was precipitated by the

addition of 6 M HCl until a pH of 1.5 was obtained. The precipitated extract was allowed to stand for

15 minutes and was then centrifuged for 30 minutes at 14,000 g. The precipitated humic acid was

lyophilized at –85 o

C for 24 hours. The resulting dry humic acid was stored at room temperature in a

glass vial under N2(g).

Artificial humic acids were prepared in the laboratory by refluxing various ratios of glucose and bovine

casein in buffered solutions for 24 to 48 hours as described by Yamamoto and Ishiwatari [1989] The

buffer was composed of 125 ml 0.1 M KH2PO4, 75 ml 0.1 M NaOH and 50 ml of glass-distilled water

to give a final pH of 7.0. After the prescribed reaction time, the mixture was allowed to cool with stirring,

then transferred to 250-ml Nalgene® centrifuge bottles. The artificial humic acid was isolated using the

extraction procedure above.

Substrate coating

Humic acids were coated to the montmorillonite substrate for use in the adsorption/desorption

experiments. Approximately 1 g of lyophilized humic acid was dissolved in 100 mL of organic-free

water and brought to a pH of 10 with sodium hydroxide. The solution was mixed under an atmosphere

of UHP-N2 for 3 to 7 days with adjustments made daily to maintain the pH. Once dissolved, the humic

acid solution was added to the washed clay and mixed for 24 hours. The suspension was transferred to

a centrifuge bottle containing the crystallized salts resulting from the evaporation of an equal amount of

artificial seawater and mixed for 48 hours. The suspension was then centrifuged at 14,000 g for 2 hours

and the supernatant was discarded. A sample of the humic acid coated clay was dried and submitted for

TOC analysis to determine the extent of humic acid adhesion. After flushing the headspace with UHP-N2

gas, the humic acid coated clay was stored in a refrigerator at 5–10°C.

Total organic carbon measurements

Total organic carbon analyses of the clay substrates were performed by Donald Schell’s laboratory at the

University of Alaska Fairbanks using a Delta Plus mass spectrometer equipped with a Finnigan MAT

Conflo II interface and a Carlo Erba elemental analyzer.

Desorption experiments

All glassware and Teflon® liners were treated to remove any hydrocarbon contamination before use.

Glassware was baked in a muffle furnace at 450 °C for at least 8 hours. Teflon® cap liners were soaked

in chromic acid for 20 minutes and then rinsed 3 times with clean organic-free water.

Artificial seawater was prepared from organic-free water that had been glass distilled over a saturated

solution of potassium permanganate. To each liter of water the following salts were added: 23.260 g

NaCl, 10.636 g MgCl2 •6H2O, 3.918 g anhydrous Na2SO4, 1.102 g CaCl2, 0.664 g KCl, 0.192 g NaHCO3,
46

0.096 g KBr, and 0.026 g H3BO3 [Lyman and Fleming 1940]. Also, 0.500 g HgCl2 was added as an

antibiologic agent.

A stock phenanthrene solution was prepared by dissolving 0.0060 g of phenanthrene in a 100-mL

volumetric flask with acetonitrile. This solution was then used to prepare the various phenanthrene

solutions required for the experiments. Aliquots of the phenanthrene stock solution, along with [9-14C]

phenanthrene (5–15 mCi mmol–1; Sigma Chemical Co.) dissolved in acetonitrile were added to vials.

The acetonitrile was evaporated, and the required amount of artificial seawater was added to make final

concentrations of 50, 300, and 700 μg L–1, in which 10% of the final phenanthrene concentration was

radiolabeled. All solutions were prepared 24 hours in advance and stored in a refrigerator (5–10°C).

Previous experiments within this laboratory have shown that for concentrations of phenanthrene ranging

from 10 to 750 μg L–1, there was no loss of hydrocarbon due to evaporation or adhesion to the walls of

the container. In addition, all experiments had control vials containing no sediment.

Experiments began by weighing 0.1 g of wet, humic-coated clay into 2-dram vials with Teflon® lined

caps. Controls were treated exactly as experimental vials from this point forward, but did not have

sediment added. Five milliliters of artificial seawater solutions containing radiolabeled phenanthrene

at several concentrations (50, 300, or 700 μg L–1) were added to the vials. A 15-second agitation on a

vortex mixer thoroughly mixed the clay and solution. The vials were then placed on a table shaker at

150–200 rpm. After adsorption for 1 h, 1 d, 3 d, 7 d, or 14 d, the vials were centrifuged at 2400 g for

30 minutes and a 1.0 mL aliquot of each of the radiolabeled reaction solutions was removed and

scintillation counted to determine phenanthrene adsorption to pristine sediments. The remaining

supernatant was discarded and replaced with a 300 μg L–1 phenanthrene solution, without radiolabel,

to begin the desorption phase of the experiment.

Vials containing the previously adsorbed sediment and the fresh phenanthrene solution were vortex mixed

for 15 seconds and then placed on a table shaker at 150–200 rpm. After desorption for 1, 3, 7, 14, 30, 60,

or 90 days, the vials were centrifuged at 2400 g for 30 minutes and a 1.0 mL aliquot of the supernatant

was removed and scintillation counted to determine the amount of labeled phenanthrene desorbed from

the sediment.

Upon the completion of each desorption experiment, the remaining supernatant was discarded and the

sediments were placed in a 65 °C oven until dry. Sediment dry weights and pore water volumes were

calculated from these dried sediments.

Results

The organic carbon content of the clay after hydrogen peroxide treatment was ≤ 0.095 wt %. Analyses

were performed to confirm that humic acid was, indeed, bound to the clay after the coating process and

that humic acid was not removed from the clay during the course of the adsorption and desorption

experiments. Initial concentrations of organic matter were similar to those found naturally in marine

sediments. The organic carbon contents of the clays after the adsorption and desorption experiments

were not statistically different from those measured before the experiments (data not shown).

Radiolabeled phenanthrene concentrations remained constant over time in the controls, indicating

that there were no significant losses due to adsorption to vessel walls of the Teflon® cap liner, nor to

volatilization. Desorption plots similar to Figures 17a (melanoidins) and 17b (natural humic acids) were

constructed. For each humic acid coated sediment, radiolabeled phenanthrene was allowed to adsorb for

a prescribed number of days (7 days for sample 10:1–48 in Figure 17a), after which the solution was

replaced with non-labeled 300 μg L–1 phenanthrene. The desorption of adsorbed radiolabeled

phenanthrene was measured at 1, 3, 7, 14, 30, 60, and 90 days and reported as a fraction of what was
47

expected based on 100% reversibility, that is, partition coefficients associated with adsorption were

assumed to be the same as those associated with desorption [Wu and Gschwend 1986].

50 μg L–1

300 μg L–1

700 μg L–1

20

Days Desorption

0 40 60

0

0.2

0.8

1

1.4

1.6

80 100

1.2

0.6

Fraction of Expected Desorbed

0.4

Figure 17a. Desorption plot for a synthetic humic acid (10:1–48). Radiolabeled

phenanthrene (50, 300, and 700 μg L–1) was adsorbed for 7d, then replaced

with non-labeled phenanthrene (300 μg L–1). Desorption was measured

after 1, 3, 7, 14, 30, 60, and 90 d.

Figure 17a is typical of these plots, in that desorption decreases with increasing concentrations of

phenanthrene adsorbed to the sediment. In general, desorption required 3 to 7 days to reach steady state.

The observed fraction of expected phenanthrene desorption ranged from 0.6 to 1.6 for all of the humic

acids investigated, with most values lying between 0.8 and 1.2. Note that the expected (calculated)

desorption is much less than 100% of the phenanthrene adsorbed initially, because of the very strong

phenanthrene adsorption. So, fractions of expected phenanthrene adsorption >1 still reflect incomplete

desorption.

A second series of desorption plots (Figures 18a and 18b, melanoidins and natural humic acids,

respectively) was constructed to investigate the influence adsorption time had on desorption. In these

plots, the percent of expected phenanthrene desorption after a fixed time (7 days for sample 10:1–48

shown in Figure 18a) was compared to adsorption times of 1 hour, and 1, 3, 7 and 14 days for a particular

humic acid. While 20 of 44 samples approached steady state after approximately 3 to 7 days of adsorption

time, 24 of 44 continued a trend of increasing desorption with increasing days of adsorption, but to

varying degrees (from 10% to 40% increase). The magnitude of these slow increases in fraction of

expected desorbed phenanthrene was not correlated with any of the humic acid structural properties

studied (see Chapter 2).
48

50 μg L–1

300 μg L–1

20

Days Desorption

0 40 60

0

0.2

0.8

1

1.4

1.6

80 100

1.2

0.6

Fraction of Expected Desorbed

0.4

Figure 17b. Desorption plot for a natural humic acid (TB2). Radiolabeled

phenanthrene (50, 300, and 700 μg L–1) was adsorbed for 14 d, then

replaced with non-labeled phenanthrene (300 μg L–1). Desorption was

measured after 1, 3, 7, 14, 30, 60, and 90 d.

50 μg L–1

300 μg L–1

700 μg L–1

Days Desorption

0

0.2

0.8

1

1.4

1.6

1.2

0.6

Fraction of Expected Desorbed

0.4

0 4 10 14 2 6 8 12 16

Figure 18a. Extent of desorption for a synthetic humic acid (10:1–48). Radiolabeled

phenanthrene (50, 300, 700 μg L–1) was adsorbed for 1 h, 1, 3, 7, and 14 d.

The solution was replaced with non-labeled phenanthrene (300 μg L–1) and

allowed to desorb for 90 d.
49

2

Days Desorption

0 4 10

0

0.2

0.8

1

1.4

1.6

14 16

50 μg L–1

300 μg L–1

1.2

0.6

Fraction of Expected Desorbed

0.4

6 8 12

Figure 18b. Extent of desorption for a natural humic acid (TB2). Radiolabeled

phenanthrene (50, 300, 700 μg L–1) was adsorbed for 1 h, 1, 3, 7, and 14 d.

The solution was replaced with non-labeled phenanthrene (300 μg L–1) and

allowed to desorb for 14 d.

Initial rates of phenanthrene desorption from sediments were calculated for 1 day of desorption after

1 day of adsorption time (Table 9). Figure 19 shows the initial rate to be directly proportional to

the dissolved concentration of phenanthrene during the prior adsorption reaction. Overall, the rates

ranged from 169 μg•g OC–1 •d–1 for a synthetic humic acid, 1:10–24, at 50 μg L–1 phenanthrene to

7800 μg•g OC–1 •d–1 for a natural humic acid, JB2, at 700 μg L–1 phenanthrene. Mineral particles coated

with natural humic acids generally displayed faster initial rates of desorption than did synthetic humic

acids (JB2 > 5A >> 10:1–48 = TB2 > 1:10–48 > HBH > 1:10–24), corresponding to their greater

partition coefficients (see Chapter 3). Initial rates, evaluated after 1 day of desorption and at 300 μg L–1

phenanthrene concentration, were found to be significantly negatively correlated with both the carbon

content of the humic acid coated clay (r

2 = 0.52, t = 1.714, n = 9) and the percent carbon of the humic

acid itself (r

2

 = 0.62, t = 3.409, n = 9).
50

Table 9. Initial rates of phenanthrene desorption after 1 day of adsorption at various

concentrations based on 1 day time increments. Concentrations are for

micrograms of phenanthrene per liter of solution at the start of the adsorption

phase. Rates of desorption are on a per gram of organic carbon basis (g OC).

n/d – not determined

Sample

Initial Desorption Rate

50 μg L–1

(μg• g OC–1 •d–1)

Initial Desorption Rate

300 μg L–1

(μg• g OC–1 •d–1)

Initial Desorption Rate

700 μg L–1

(μg• g OC–1 •d–1)

1:10–24 169 838 2110

1:10–48 317 1210 2940

10:1–48 257 1250 3710

5A 501 2660 6050

JB2 370 3250 7800

HBH 185 974 2530

TB2 441 1410 n/d

PV25 n/d 1090 n/d

PV82 n/d 3540 n/d

1:10–24

1:10–48

10:1–48

5A

HBH

JB2

TB2

PV25

PV82

1:10–24

1:10–48

10:1–48

5A

HBH

JB2

TB2

100

[phenanthrene]initial

0 200 300 500 600

0

Initial Rate (μg phenanthrene •g OC–1 •d–1)

3000

4000

7000

8000

400 700 800

9000

5000

6000

2000

1000

Figure 19. Initial desorption rates after 1 day of desorption time. Radiolabeled phenanthrene

was in contact with the sediments for 1 day before desorption began.
51

Discussion

In previous work, phenanthrene adsorption was found to be dependent upon the degree of nonpolar

character of the humic acids (see Chapter 2), but desorption trends of the current study were not explained

by humic acid structural characteristics such as aliphaticity, aromaticity, or percent carboxylic, amidic, or

carbonyl carbons as measured by 13C nuclear magnetic resonance spectroscopy. Other studies have been

unable to link the observed extent of desorption to additional characteristics of the organic matter such as

cation exchange capacity, organic carbon content, surface area, and solubility [Pavlostathis and Mathavan

1992].

Contact time is a minor factor in controlling the amount of phenanthrene adsorbed to sediments on a

normalized to organic carbon content basis, while the initial concentration of phenanthrene in solution

is the primary factor (see Chapter 3). The extent of phenanthrene desorption was found to decrease for

both increasing initial phenanthrene concentrations used in the adsorption phase of the experiment and

decreased reaction time. The reduction in the proportion of desorbed phenanthrene with increasing

concentrations is consistent with other studies [Schlebaum et al. 1998] which attributed this pattern to

increasing partitioning of phenanthrene into more structurally rigid portions of the humic acid matrix.

However, decreased desorption was also expected with increasing adsorption times, but this was not

observed. Numerous studies in the literature report that decreased reversibility of hydrocarbon adsorption

is a function both of increasing initial concentrations [Borglin et al. 1996; Pavlostathis and Mathavan

1992; Schlebaum et al. 1998] and increasing adsorption times [Karickhoff and Morris 1985; Borglin et al.

1996; Pavlostathis and Mathavan 1992; Schlebaum et al. 1998].

One potential explanation for changes in desorption with phenanthrene concentration is conformational

changes in the humic acid structure. A possible cause of conformational change is the concentration

of salt. The humic substances used in this study were first dissolved in fresh water, but then artificial

seawater salts were added before the material was coated onto humic substances for 48 hours. Subsequent

experiments took place in artificial seawater. The addition of salt causes negative sites on the humic

acid to become neutralized and the intramolecular repulsions to decrease, allowing for the humic acid

molecule to fold and coil upon itself [Ragle et al. 1997]. This conformational change has been observed to

take place within two hours of the addition of salt [Engebretson and von Wandruszka 1998] for dissolved

humic substances, and, unless the process was much slower for the particle associated humics, it should

have been complete long before the experiments began. Large anions such as bromide and chloride

were found to be unable to penetrate the pseudomicellar domains of a folded humic acid in the studies

conducted by Ragle et al. [1997]. If increasing conformational change took place during longer adsorption

times, the more compact structure could have excluded phenanthrene from tortuous paths within the

organic matter, making desorption more likely. However, this salt effect is unlikely to explain most of the

results, because it should have been largely complete before experiments began.

A related interpretation would be that conformational change was caused by mercuric chloride added as

an antibiological agent. The mercuric ion has a radius similar to that of sodium and calcium ions, which

were present in the artificial seawater in greater concentration than the mercuric ion. Any conformation

effect of mercuric ion should have been complete on a time scale similar to that for the other cations. It

could have had a proportionally greater effect because of its strong association with humic acid, but it

constituted a small fraction of total ion concentration (1.24% by wt).

Finally, hydration of the humic acid coating could have changed, most likely beginning at the time

the moist humic acid coated clay was first suspended in the artificial seawater at the beginning of the

adsorption phase of the experiment. Again, this would be expected to occur more quickly than the

adsorption of phenanthrene owing to the smaller size of the water molecule. In addition, conformational

changes due to sea salts, mercuric ion, or water would not be expected to vary with phenanthrene

concentration, but desorption decreased as phenanthrene concentration increased.
52

The decrease in phenanthrene desorption with increasing phenanthrene concentration is contrary to

the pattern expected if equally accessible strong and weak adsorption sites were present in the humic

structure. That situation could explain apparently irreversible adsorption; phenanthrene adsorbed at

stronger sites would appear to be irreversibly adsorbed under conditions that readily desorbed it from

weaker sites. However, more phenanthrene would be adsorbed at the stronger sites at lower phenanthrene

concentrations. This would result in increasing apparent KOC with decreasing concentration, which was

not observed (see Chapter 3). It would also result in decreased desorption at lower concentrations,

opposite to the observed pattern. Decreased desorption with increasing initial dissolved and adsorbed

concentration could be caused by the phenanthrene itself affecting the humic structure, reducing its

porosity. At most, the phenanthrene would comprise only 1% of the mass of the humic substance, so

it could exert an observable effect only if a small proportion of the material was accessible to adsorbing

phenanthrene molecules.

Possible phenanthrene effects on humic acid structure are less helpful in explaining why desorption

increased with increasing initial adsorption time. Since adsorbed phenanthrene concentration increased

with adsorption reaction time, decreased adsorption would be the expectation. Rather, that result can be

best explained by the fact that for short adsorption times, adsorption had not yet reached steady state. For

the longest reaction times, presumably all accessible regions of the humic acid coating have a quantity

of adsorbed phenanthrene described by the steady state partition coefficient. At earlier times, however,

the amount of adsorbed phenanthrene in less accessible regions could be lower. So, redistribution of

phenanthrene during the desorption step could include diffusion into those less accessible regions as

well as desorption into the solution.

Desorption steady state was reached between three and seven days for all of the humic acids studied in

this experiment. These results are comparable to those found in other studies (two to four days) using

radiolabeled phenanthrene desorbing from humic acid coated montmorillonite clays [Lahlou and OrtegaCalvo 1999]. The approach to steady state slowed after relatively fast initial desorption within the first

24 hours. These biphasic patterns of desorption have been observed in the literature, with the fast stage

lasting from a few hours [Karickhoff and Morris 1985; Schlebaum et al. 1998] to one day [Pavlostathis

and Mathavan 1992]. As with slow adsorption, biphasic patterns of desorption are indicative of diffusion

within the organic matrix.

Initial rates of desorption were found to be directly proportional to the initial concentration of

phenanthrene used during the adsorption phase. The pattern in initial desorption rates may simply be

due to concentration effects on diffusion rates, since the concentration of adsorbed phenanthrene was

correlated with initial solution concentration. In a study by Borglin et al. [1996], it was also found that

increasing initial concentrations of hexachlorobenzene led to faster initial desorption rates from natural

sediments. The observation that humic acids with larger organic normalized partition coefficients were

associated with faster initial rates of desorption could also be related to the concentration gradient

driving outward diffusion, since, of course, more phenanthrene associated with humic acids having higher

partition coefficients. However, this result differs from those of previous studies [cf. Borglin, et al. 1996;

Schlebaum et al. 1998], where desorption rate decreased with increasing partition coefficient. Schlebaum

et al. [1998] suggested that larger partition coefficients increase the amount of a hydrocarbon bound to a

non-labile fraction, resulting in significantly lower desorption rates, but points out that this has not been

observed experimentally.

The negative correlation of initial desorption rate with the organic carbon concentration of the coated clay

is also consistent with the interpretation that desorption rate is controlled by slow diffusion within the

organic matrix. In the absence of major differences in the tortuosity of the pores within the structure of

the different humic acids, thicker coatings should lead to longer diffusive paths and slower desorption.
53

The humic acids with lower organic carbon content had higher rates of desorption. Folding and coiling

may be limited due to a decrease in flexibility of the humic acid resulting from fewer long carbon chains.

This would produce a more open structure through which molecules would have a less tortuous path to

follow as they diffuse both in and out of the organic matrix. Alternatively, humic substances with less

carbon per gram generally have more oxygen, which was not measured. Higher oxygen content would

make them more hydrophilic, leading to a more open structure with greater water content. This would

result in faster rates of adsorption as well as desorption.

Slow desorption is clearly a contributing factor in the phenomenon of irreversible adsorption. Desorption

increased for up to a week, and at steady state was near 100% for most of the experiments. However,

shorter reaction times would have led to the conclusion that adsorption was partly irreversible. In some

experiments, more than the expected amount of phenanthrene desorbed. The most likely explanation is

a change in the conformation of the humic acid—due to salt, phenanthrene concentration, or simply time

—that decreased the phenanthrene partition coefficient. Most experiments with greater than expected

desorption were those with the smallest initial phenanthrene concentration in the adsorption reactions,

indicating that phenanthrene concentration is the most likely factor.

While there was little evidence for irreversible adsorption in the results of the experiments described here,

this is a widely observed phenomenon in sediments and soils [Hatzinger and Alexander 1995; Lahlou and

Ortega-Calvo 1999; Kan et al. 1994]. Adsorption that is irreversible over much longer time scales could

occur for materials more complex than the deliberately simplified model sediments prepared for this

work. In particular, the model sediments included only the base extractable portion of the sediment

organic matter. Other sediment organic constituents could exhibit different interactions with

phenanthrene.
54
55

Chapter 5. Conclusion

At present it is impossible to accurately predict the ecotoxicological effects of any specific combination

of pollutant and sedimentary organic material. The goal of the research was to improve our ability to

predict the persistence and fate of polycyclic aromatic hydrocarbons in contaminated sediments, through

elucidation of the role sediment organic matter has in adsorption and desorption processes. The project

investigated the interaction between an aromatic hydrocarbon, phenanthrene, and humic acids obtained

from marine sediments from Lower Cook Inlet, Port Valdez, and the inner Beaufort Sea, Alaska, and

from laboratory synthesized homologs (melanoidins). The work sought understanding of how the

structural characteristics of humic acids influence the rate and extent of phenanthrene adsorption.

The first hypothesis tested was that aromatic hydrocarbon adsorption that was partly irreversible in

experiments lasting a few days would be reversible with longer desorption times. The results described

in Chapter 4, were consistent with this hypothesis in general. Adsorbed phenanthrene was completely

desorbed within one week under most experimental conditions. While the rate of desorption was directly

proportional to the initial concentration of phenanthrene adsorbed to the sediments, the extent to which

the bound phenanthrene desorbed was negatively correlated with the initial phenanthrene concentration.

Desorption was biphasic, with a slow approach to steady state again taking approximately one week.

These facts, when taken together, indicate that desorption is not governed solely by slow diffusion within

the three-dimensional structure of humic acid. In addition, the adsorbed phenanthrene appears to

influence the adsorptive or diffusive properties of the humic acid, leading to the concentration effects

observed.

The results were therefore consistent with the second hypothesis, that interactions of aromatic

hydrocarbons with sediment organic matter are responsible for adsorption that appears to be irreversible,

at least under some conditions. A subhypothesis, that binding to a variety of adsorption sites with

different partition coefficients causes differences in the apparent, experimental partition coefficient,

was rejected. A mixture of strong and weak adsorption sites in the humic acid would lead to decreasing

adsorption partition coefficients with increasing concentration, which was not observed (see Chapter 3).

It would also lead to decreases in percentage of expected desorption with decreasing phenanthrene

concentration adsorbed, opposite to the trend actually found (see Chapter 4).

The third hypothesis was that variations in organic matter properties influence the rate and extent of

adsorption and desorption processes. The tests of this hypothesis were described in Chapter 2. It was

found that the partitioning of phenanthrene to the sediments was not related to the concentration

(g OC•g clay–1) of humic acid coating the mineral substrate. Adsorption, however, was found to be

negatively correlated with the polarity of the humic acid. This indicates that it is the nature of the organic

matter and not the quantity alone that governs the extent of hydrophilic organic carbon adsorption to

sediments. Desorption, on the other hand, was found to be inversely related with the amount of humic

acid coating the mineral substrate (see Chapter 4), probably because thicker coatings result in longer

diffusion paths out of the organic matrix. Structural characteristics of the humic acids were not found to

influence the desorption of phenanthrene. The subhypothesis that there is a direct correlation between the

aromatic character of sediment humic acid and the adsorption partition coefficient of phenanthrene was

rejected (see Chapter 2). There was no relationship between aromaticity and adsorption within our sample

group. The subhypothesis that the synthetic humic substances would show clearer relationships between

organic matter composition and adsorption was rejected; trends for natural sedimentary humic acids and

melanoidins were similar.

Finally, the hypothesis that previous phenanthrene adsorption of sediments affects subsequent adsorption

was rejected, as discussed in Chapter 3. The investigation was comprised of two related components—the

kinetics of phenanthrene adsorption to mineral-bound humic acids and the effects previous phenanthrene

contamination of sediments had on subsequent phenanthrene adsorption. Organic normalized partition
56

coefficients (KOC) were determined by using the slopes of linear adsorption isotherms. These KOC were

found to be negatively correlated with the polarity of the organic matter, as was seen in Chapter 2. Initial

rates of adsorption were also found to be negatively correlated with the polarity of humic acids coating

the mineral substrate. The observation that the initial uptake of phenanthrene was rapid, but the approach

to steady state required approximately one week, was consistent with hydrocarbon diffusion into a threedimensional organic matrix. The initial concentration of phenanthrene in solution had no effect on the

partitioning of the hydrocarbon into the sediment, indicating that binding sites within the organic matter

are both uniform in strength and unlimited. This, again, supports the idea that rates of hydrocarbon

adsorption are controlled by intraorganic matter diffusion. If there existed only a finite number of sites to

which phenanthrene could bind, hydrocarbons from previous contamination would be expected to occupy

these sites, thus hindering additional uptake of phenanthrene. This was not observed experimentally.

The results presented in Chapter 2 showed that humic acids with properties indicating a mixed terrestrial

and marine origin adsorbed phenanthrene to a greater extent than those of a wholly marine origin.

However, this was not attributable to greater aromaticity derived from lignin, but rather to a relative lack

of amide and carboxylic acid functional groups. Another characteristic shared by humic acids exhibiting

stronger PAH adsorption and lower polarity was a subtidal sediment source, while the less adsorptive

Lower Cook Inlet samples were all from intertidal sediments. The intertidal sediments were probably

richer in living organisms and fresh detritus than the subtidal sediments, potentially another factor

contributing to the greater polarity of their humic acids. Although the number of samples studied was not

sufficient to conclusively identify geographic patterns in adsorptive properties of humic acids, the results

do suggest that terrigenous inputs and, possibly, extent of diagenetic alteration of organic matter should

be investigated as potential predictors of adsorptive properties.

The conclusions reached should serve to warn against applying results regarding adsorption of pollutants

by sediments from one study to another. Policies and predictions need to be made on a case-by-case

basis, because of the clear dependence of adsorption/desorption processes on sediment organic matter

structures. Additional work, however, is required to further define the role the organic matrix has on

the processes of adsorption and diffusion. The body of work shows that hydrocarbon associations with

organic matter in marine sediments are controlled by diffusion processes, regulated by the structural

characteristics of the organic matrix. Additional desorption studies that focus on both the initial

hydrocarbon concentrations, and the hydrocarbon concentrations within the organic matrix once

adsorbed, are needed. Supplemental investigations of the adsorption process would be desirable to

complement that research focus. These include investigations of HOC other than phenanthrene and

organic matter of a greater variety of types or origins.

Acknowledgments

This work was funded by the University of Alaska Coastal Marine Institute, Alyeska Pipeline Service

Company, the State of Alaska, and the Alaska Sea Grant College Program. The authors would like to

express their thanks to Thomas Gedris at the Florida State University’s NMR Laboratory for his efforts

in acquiring solid-state humic acid spectra, Joan Braddock, Sathy Naidu, John Goering, and John Kelly

at the University of Alaska Fairbanks for supplying sediment samples, and Joan Braddock, Thomas

Clausen, and Bruce Finney, also at the University of Alaska Fairbanks, for their comments and review

of the manuscript.
57

Study Products

Henrichs, S.M. 2000. Kinetics and mechanisms of slow PAH desorption from Lower Cook Inlet and

Beaufort Sea sediments. University of Alaska Coastal Marine Institute Annual Research Review,

March 2000, Fairbanks, Alaska.

Henrichs, S.M. 2000. Kinetics and mechanisms of slow PAH desorption from Lower Cook Inlet and

Beaufort Sea sediments, p. 76. In University of Alaska Coastal Marine Institute Annual Report No. 6.

OCS Study MMS 2000-046, University of Alaska Fairbanks and USDOI, MMS, Alaska OCS

Region.

Henrichs, S.M., and J.A. Terschak. 2000. Kinetics and mechanisms of slow PAH desorption from Lower

Cook Inlet and Beaufort Sea sediments, p. 42–50. In University of Alaska Coastal Marine Institute

Annual Report No. 7. OCS Study MMS 2000-070, University of Alaska Fairbanks and USDOI,

MMS, Alaska OCS Region.

Henrichs, S.M., and J.A. Terschak 2001. Kinetics and mechanisms of slow PAH desorption from Lower

Cook Inlet and Beaufort Sea sediments. University of Alaska Coastal Marine Institute Annual

Research Review, February 2001, Fairbanks, Alaska.

Henrichs, S.M., and J.A. Terschak 2001. Kinetics and mechanisms of slow PAH desorption from Lower

Cook Inlet and Beaufort Sea sediments. Alaska OCS Region of the Minerals Management Service,

8th Information Transfer Meeting, 3–5 April 2001. Anchorage.

Henrichs, S.M., and J.A. Terschak. 2002. Kinetics and mechanisms of slow PAH desorption from Lower

Cook Inlet and Beaufort Sea sediments, p. 28–38. In University of Alaska Coastal Marine Institute

Annual Report No. 8. OCS Study MMS 2002-001, University of Alaska Fairbanks and USDOI,

MMS, Alaska OCS Region.

Henrichs, S.M., and J.A. Terschak 2002. Kinetics and mechanisms of slow PAH desorption from Lower

Cook Inlet and Beaufort Sea sediments. University of Alaska Coastal Marine Institute Annual

Research Review, February 2002, Fairbanks, Alaska.

Henrichs, S.M., J.A. Terschak and D.G. Shaw. 2003. Kinetics and mechanisms of slow PAH desorption

from Lower Cook Inlet and Beaufort Sea sediments, p. 8. In University of Alaska Coastal Marine

Institute Annual Report No. 9. OCS Study MMS 2003-003, University of Alaska Fairbanks and

USDOI, MMS, Alaska OCS Region.

Terschak, J.A. 2002. Phenanthrene Adsorption and Desorption by Melanoidins and Marine Sediment

Humic Acids. Ph.D. Dissertation, University of Alaska Fairbanks, 179 p.

Terschak, J.A., and S.M. Henrichs. In preparation. Phenanthrene adsorption to mineral-bound humic

acid: Kinetics and influence of previous phenanthrene adsorption.

Terschak, J. A., and S.M. Henrichs. In preparation. Desorption kinetics of phenanthrene from mineralbound humic acids: Consequences of conformational changes.

Terschak, J.A., S.M. Henrichs and D.G. Shaw. 2004. Phenanthrene Adsorption and Desorption by

Melanoidins and Marine Sediment Humic Acids. Final Report. OCS Study MMS 2004-001,

University of Alaska Coastal Marine Institute, University of Alaska Fairbanks and USDOI, MMS,

Alaska OCS Region. 65 p.

Terschak, J.A., S.M. Henrichs, and D.G. Shaw. In preparation. Effects of humic acid properties on

phenanthrene adsorption.
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