5

CHAPTER 2

LITERATURE REVIEW

2.1 SOURCES OF PAHs

The worldwide production of crude oil was more than three billion

tonnes per year, and about half of this quantity is transported by sea

(Harayama et al 1999). This is mainly attributed to the majority of oil fields

being limited to areas such as the Persian Gulf region. All tanker ships take on

ballast water which contaminates the marine environment when it is

subsequently discharged. Disasters such as tanker accidents as exemplified by

that of the T/V Exxon Valdez in Prince William Sound, Alaska and Bahia

Paraiso in Antarctica during 1989, which severely affected the local marine

environment (Harayama et al 1999). On January 4, 1997, heavy fuel oil of

6,200 KL was spilled from the Russian tanker Nakhodka in the sea of Japan

(Koyama et al 2004).

Petroleum hydrocarbons are the most common environmental

pollutants and oil spills pose a great hazard to terrestrial and marine

ecosystems. Petroleum is a viscous liquid mixture that contains thousands of

compounds mainly consisting of carbon and hydrogen. The total influx of oil

into the sea is estimated to be between 1.7 and 8.8 million tonnes (Leahy and

Colwell 1990). During the gulf war in 1991, about 140 million gallons of oil

was spilled due to demolition of oil storage tanks, oil terminals and tankers in

Kuwait.

6

Historically, PAHs were formed during the production of coke for

steel, as well as in the manufacture of coal gas and fuel gas from hard coal,

during which the tar oils were left over as distillation residues. As the toxic

and biocidal effects of tar oils were known since the 18th century, such

distillation residues were used only for the conservation of wood and rope

(Wiesmann 1994). Gaswork sites and impregnation plants used for the

production of railway sleepers are nowadays considered as PAH

contaminated sites (Wiesmann 1994). Off-shore drilling is now a common

method followed to explore new oil resources which adds up another source

for petroleum pollution. However, the largest source of marine contamination

by petroleum seems to be the runoff from land. Annually, more than two

million tonnes of petroleum was estimated to end up in the sea.

Polycyclic aromatic hydrocarbons (PAHs) are introduced into the

environment from fossil fuel, other organic material combustion activities,

accidental spilling of processed hydrocarbons and oils, coal liquefaction and

gasification, organic oil seepage and surface run-off from forest/brush fires

and natural geologic processes (Guerin and Jones 1988; Freeman and Cattell

1990).

McElroy et al (1989) reported that harbour sediments are

commonly contaminated with hydrocarbons from shipping activities, fuel

spills, runoff and inputs from sewage treatment plants. Although the

monoaromatic hydrocarbon components of these wastes are often readily

degraded, PAHs deposit in the bottom sediment due to its less volatile nature

and high affinity towards particulate matter. The petroleum introduced to the

sea seems to be degraded either biologically or abiotically (Readman et al

1992).

7

Possible fates for PAHs released into the environment include

volatilization, photo-oxidation, chemical oxidation, bioaccumulation and

adsorption on to soil particles. The role of PAHs in marine environment is

discussed in section 2.2.

2.2 PAH IN MARINE ENVIRONMENTS

When petroleum is spilled into the sea, it spreads over the surface

of the water. It is subjected to many modifications, and the composition of the

petroleum changes with time. This process is called weathering, and is mainly

due to evaporation of the low-molecular-weight fractions, dissolution of the

water-soluble components, mixing of the oil droplets with seawater,

photochemical oxidation, and biodegradation. Those petroleum components

with a boiling point below 250 ºC are subjected to evaporation. Therefore, the

content of n-alkanes, whose chain length is shorter than C14, is reduced by

weathering. The content of aromatic hydrocarbons within the same boiling

point range is also reduced as they were subjected to both evaporation and

dissolution. The mixing of oil with seawater occurs in several forms

(Figure 2.1). Dispersion of the oil droplets into a water column is induced by

the action of waves, while water-in oil emulsification occurs when the

petroleum contains polar components that act as emulsifiers. A water-in-oil

emulsion containing more than 70% of seawater becomes quite viscous and is

called chocolate mousse from its appearance. After the light fractions have

evaporated, heavy residues of petroleum can aggregate to form tar balls

whose diameter ranges from microscopic size to several tenths of a centimeter

(Tjessen et al 1984).

Under sunlight, petroleum oil discharged into sea was subjected to

photochemical modification. Some reports have suggested the light-induced

polymerization of petroleum components, while others have suggested their

8

photodegradation. An increase in the polar fraction and a decrease in the

aromatic fraction have also been observed (Ehrhardt and Weber 1991).

Aliphatic components do not significantly absorb solar light, and are

photochemically inert in nature. However, they can be degraded by

photosensitized oxidation. The aromatic or polar components in petroleum

and anthraquinone that is present in seawater can provoke the degradation of

n-alkanes into terminal n-alkenes (a carbon double bond at position 1) and

low-molecular-weight carbonyl compounds (Ehrhardt and Weber 1991).

Figure 2.1 Processes affecting the rate of hydrocarbons in marine

environment after oil spill

The sources of PAH in food are mainly from environmental

pollution and from food processing (drying, smoking) and cooking (roasting,

grilling, and frying) (WHO 1998, 2005). PAHs found in coal tar, crude oil,

creosote, and roofing tar, but a few are used in medicines or to make dyes,

plastics, and pesticides act as other sources of marine pollution (ATSDR

1996). Some of the major accounts of oil spills in Indian waters are listed in

Table 2.1.

9

Table 2.1 Oil pollution incidents in Indian waters (1995 to 2005)

Date Quantity Spilled (T) Position Name of the

Vessel 1995March 26 200/Diesel Off Vizag, Andhra Pradesh Dredger Mandovi

-2September 24 -/FO Off Dwarka, Gujarat MC Pearl November 13 Tanker Wash Eliot beach, Madras, Tamil

Nadu Unknown

1996May 21 370 FO Off Hooghly River, West

BengalMV Prem Tista

June 16 120/FO Off Prongs Lighthouse, Maharashtra

MV Tupi Buzios

June 18 132/FO Off Bandra, Maharashtra MV Zhen Don June 18 128/FO Off Karanja, Maharashtra MV Indian

ProsperityJune 23 110/ FO Off Worli, Maharashtra MV Romanska August 16 124/FO Malabar Coast, Kerala MV Al-Hadi 1997January 25 Tanker Wash Kakinada Coast, Andhra

PradeshUnknown

June 19 210/FO Off Prongs Lighthouse, Maharashtra

MV Arcadia Pride

September 14 Naptha, Diesel Petrol

Vizag, Andhra Pradesh HPC Refinery

August 2 70/FO Off Mumbai, Maharashtra MV Sea Empress1998June 1 20/Crude Off Vadinar, Gujarat Vadinar, SBM 1999July 8 15/FO Mul Dwarka, Gujarat MV Pacific

Acadian December 17 1/FO Bombay Harbour,

MaharashtraMV StonewallJackson

10

Table 2.1 (Continued)

Date Quantity Spilled (T) Position Name of the

Vessel 2003April 29 2000 L of

Arab Light Crude Oil

05 miles Off Kochi, Kerala MT BR Ambedkar

May 9 2000 tonnes ofNaptha

Mumbai Harbour (SW of West Colaba Pt.), Maharashtra

MT UPCO_ III

August 10 300 Ton Crude Oil

ONGC Rig (BHN), Maharashtra

URAN Pipe Line

2004February 28 01 Ton Crude

Oil36 inches ONGC Pipe line at MPT Oil Jetty (Tata Jetty - OPL Pir Pau), Maharashtra

During Crude Oiltransfer fromJawahar Dweep toONGC- Trombaythrough 36 InchesPipe

2005June 30 49,537 Ton

Cargo and 640 Ton FO

South backwaters of Vishakhapatnam Port

MV Jinan VRWD - 5

Jul 4 350 cu meter Base Lube Oil

Mumbai Harbour (2.5 cables NW Of Tucker beacon)

Dumb Barge Rajgiri

Jul 25 33 Ton FO 1.2 NM NE of Paget Island (North Andaman)

MV Edna Maria

FO: Fuel Oil, HO: Heavy Oil

Source: Blue Waters (2007).

11

2.3 TOXICITY OF PAHs

Generally PAHs do not directly play the role of carcinogens and

mutagens in humans. They undergo a variety of metabolic reactions in human

cells before causing cancer or mutations (Baird 1995). The first chemical

transformation that occurs in the body is the formation of an epoxide ring

across one C=C bond in the PAH. A fraction of these epoxide molecules

subsequently react with water, to yield two –OH groups on adjacent carbons

forming trans-dihydrodiol. The double bond that remains in the same ring as

the two –OH groups subsequently undergo epoxidation, thereby yielding the

trans-diol epoxide, which is an active carcinogen. By the addition of H+, this

molecule can form a particular stable cation that can bind to molecules such

as DNA, thereby inducing mutations and cancer. The metabolic reactions and

water addition are part of the human body’s attempt to introduce –OH groups

into hydrophobic molecules like PAHs and thereby making them more

capable of becoming water soluble and get eliminated (Baird 1995).

Heitkamp and Cerniglia (1988) reported that US Environmental Protection

agency has identified 16 PAH compounds as priority pollutants and their

levels in industrial effluents required to be monitored (Table 2.2).

Large amount of oily wastewater is generated during oil exploration

and production activities. Produced waters contain a wide range of salinities.

Spilled brine inhibits plant growth, leading to increased erosion and loss of

topsoil and contamination of ground water by both salt and hydrocarbons

(Nicholson and Fathepure 2004).

The water-soluble components of petroleum exert a toxic effect on

marine organisms. In general, aromatic compounds are more toxic than

aliphatic compounds, and smaller molecules are more toxic than larger ones

in the same series. Solar irradiation affects oil toxicity as the surface films

12

become less toxic due to the loss of polycyclic aromatic hydrocarbons, but the

toxicity of the water-soluble fraction increases as its concentration increases

(Nicodem et al 1997).

Table 2.2 Physical and chemical properties of the 16 priority PAH

pollutants (USEPA)

PAHs Molecular Weight

Solubilitya

(mg/L) Carcinogenic

potentialb Naphthalene 128 30 - Acenaphthylene 154 16.1 - Acenaphthene 152 3.47 ± Fluorene 166 1.8 ± Phenanthrene 178 1.29 - Anthracene 178 0.073 - Fluoranthene 202 0.26 - Pyrene 202 0.135 - Benzo[a]anthracene 228 0.014 ++Chrysene 228 0.0006 + Benzo[b]fluoranthene 252 0.0012 ++Benzo[k]fluoranthene 252 0.00055 ++Benzo[a]pyrene 252 0.0038 ++Dibenzo[ah]anthracene 278 0.0005 ++Benzo[ghi]perylene 276 0.00026 ± Indeno[123-cd]pyrene 276 0.062 ++

According to Kästner (2000)a Bhatt (2001) b, Cerniglia (1993) b ++ : sufficient evidence of causal relationship between the tested agent and human

cancer + : limited evidence, causal relationship is likely, but not proven

: inadequate evidence, both negative and positive data available. : sufficient evidence to exclude carcinogenity of the tested agent

13

PAHs are rarely encountered alone in the environment and many

interactions occur within a mixture of PAHs whereby the potency of known

genotoxic and carcinogenic PAHs can be enhanced (Kaiser et al 1997). For

example, 1-nitropyrene, a nitrated PAH, is produced during reactions between

ketones in products of burning automobile fuel and airborne nitrogen oxides

that take place on the surface of hydrocarbon particles in diesel exhaust. In the

Ames Salmonella typhimurium assay, 1-nitropyrene was found to be highly

mutagenic and carcinogenic, whereas the parent compound, pyrene, is non-

carcinogenic and only weakly mutagenic (Pothuluri and Cerniglia 1994).

Samanta et al (2002) studied the fate of PAHs in the environment

and reported that a wide variety of PAHs present in nature is due to

incomplete combustion of organic matters. The PAHs from extraterrestrial

matter are oxidized and reduced owing to prevalent astrophysical conditions

and result in the formation of various organic molecules, which are the basis

of early life on primitive earth. The microorganisms (naturally occurring or

genetically engineered) were able to mineralize toxic PAHs into CO2 and H2O

(Figure 2.2). Thus bioremediation acts as major treatment technology for

removing PAHs from the environment.

14

Figure 2.2 Fate, toxicity and remediation of PAHs from the

environment (Samanta et al 2002)

2.3.1 Toxic Effects of PAHs on Animals

PAHs released into the marine environment tend to adsorb rapidly

on suspended materials and sediments and they are bioavailable to fish and

other marine organisms in the food chain, as waterborne compounds and from

contaminated sediments. Of these three possible routes, uptake of waterborne

PAHs across the gills is considered to be the most significant route, also,

PAHs uptake always depends on their bioavailability as well as the

physiology of the organisms (Meador et al 1995).

Jonsson et al (2004) reported that in vertebrates, the majority of

absorbed PAHs were efficiently biotransformed by enzymes and increase

their water solubility allowing excretion to take place, but invertebrates with

high metabolic capacity accumulate PAHs.

15

Perugini et al (2007) analyzed the presence of PAHs in bivalves

(Mitylus galloprovincialis), cephalopods (Todarodes sagittatus), crustaceans

(Nephrops norvegicus) and fish (Mullus barbatus, Scomber scombrus,

Micromesistius poutassou, Merluccius merluccius) in several pools coming

from the Central Adriatic Sea. Chrysene was detected only in mussels with

very low values (average 0.74 ng/g wet weight). PAHs composition pattern is

dominated by the presence of PAHs with 3-rings (62%) followed from those

with 4-rings (37%) and 5-rings (1%).

The carcinogenic effects of PAHs on mammalian cells are by

consequence of the metabolic activation of diol epoxides, which are highly

reactive molecules that covalently bind to DNA. This activation occurs

mainly in the microsomes of the endoplasmic reticulum and is catalysed by

monooxygenase enzymes associated to cytochrome P-450 (Harvey 1991).

PAHs were also shown to affect the immune system of mammals (White

1986).

Table 2.3 PAH concentrations (LC50) with acute effects on animals

Substances Concentration of PAHs in Animals (mg/L)Napthalene (0.01) 0.11-7.9 Methylnaphthalene 1.0-3.4Dimethylnaphtalenes 0.08-5.1 Trimethylnaphtalenes 0.32-2.0 Acenapthene 0.66 Fluorene 0.3-5.8Phenanthrene 0.03-1.1 0.Methylphenanthrenes 0.3-5.5Anthracene 0.2Fluoranthene 0.024-0.5 Benz(a)anthracene 0.01-1.0 Chrysene > 1.07,12-dimethylbenz(a)anthracene < 0.5Benzo(a)pyrene 0.005- > 1.0

NRC/Canada (1983), Eisler (1987) and Knutzen (1989).

16

2.3.2 Toxic Effects of PAHs on Humans

Inhalation exposure assessment showed that people in industrial

area inhaled a quantity of benzo(a)pyrene, which is equivalent to smoking

7-14 cigarettes/day (Raiyani et al 1993). Naphthalene was found to be a

common micropollutant in potable water which covalently binds to molecules

in liver, kidney and lung tissues, thereby enhancing its toxicity and also

inhibiting mitochondrial respiration. Acute naphthalene poisoning in humans

leads to haemolytic anaemia and nephrotoxicity. In addition to that, dermal

and ophthalmological changes have been observed in workers occupationally

exposed to naphthalene (Goldman et al 2001).

Phenanthrene is known to be a photosensitizer of human skin, a mild

allergen and mutagenic to bacterial systems under specific conditions

(Mastrangela et al 1997). It is a weak inducer of sister chromatid exchanges

during mitosis and a potent inhibitor of gap junctional intercellular

communication (Weis et al 1998). Many PAHs contain a ‘bay-region’formed

by the branching in the benzene ring sequence as well as ‘K-region’, both of

which allow metabolic formation of bay- and K-region epoxides, which are

highly reactive (Samanta et al 2002). Phenanthrene was the smallest PAH to

have a bay-region and a K-region, it was often used as a model substrate for

studies on the metabolism of carcinogenic PAHs (Bucker et al 1979).

Little information is available for other PAHs such as

acenaphthene, fluoranthene and fluorene with respect to their toxicity in

mammals. However, the toxicity of benzo(a)pyrene, benzo(a)anthracene,

benzo(b)fluoranthene, benzo(k)fluranthene, dibenz(a,h)anthracene and

indeno(1,2,3-c,d)pyrene has been studied and there is sufficient experimental

evidence to show that they are carcinogenic (Mastrangela et al 1997; Liu et al

2001 and Sram et al 1999).

17

PAHs have attracted most attention because of the carcinogenic

potential presented by some of them. The 64th Joint FAO/WHO Expert

Committee on Food Additives (JECFA) concluded that 13 PAHs were clearly

carcinogenic and genotoxic namely anthracene, chrysene, fluorene,

fluoranthene phenanthrene, pyrene, acenaphthene, benz(a)anthracene,

benzo(b)xuoranthene, benzo(k)xuoranthene, benzo(a)pyrene, indeno[123-

cd]pyrene and dibenz(a,h)-anthracene (CAC 2005; WHO 2005).

PAHs emissions from motor vehicle caused adverse effects on

humans. Burgaz et al (2002) reported that the traffic police officers are often

exposed to high levels of PAHs in urban streets because of motor vehicle

emission. Significant cytogenetic damage in peripheral lymphocytes due to

PAH exposure has previously been reported for traffic police officers working

in Ankara, Turkey.

Liu et al (2007) also identified the exposure of PAH profiles on

traffic police and found that large daily variations occur both in summer and

winter, because of the changes in the weather conditions, especially wind

speed and relative humidity which tend to disperse and scavenge PAHs in air.

2.3.2.1 Exposure of PAH in humans (ATSDR 1996)

Some of the possible routes through which PAH was exposed to

humans are:

Breathing air containing PAHs in the workplace of coking,

coal-tar, and asphalt production plants; smokehouses; and

municipal trash incineration facilities.

18

Breathing air containing PAHs from cigarette smoke, wood

smoke, vehicle exhausts, asphalt roads, or agricultural burn

smoke.

Coming in contact with air, water, or soil near hazardous

waste sites.

Eating grilled or charred meats; contaminated cereals, flour,

bread, vegetables, fruits, meats; and processed or pickled

foods.

Drinking water or cow’s milk contaminated with

hydrocarbons.

Nursing infants of mothers living near hazardous waste sites

may be exposed to PAHs through their mother's milk.

2.4 PROCESSES INVOLVED IN PAHs REMOVAL

PAHs present in the environment persist for long time. In order to

remove them, several removal methods have been employed.

2.4.1 Dispersants

Dispersants are surface-active agents that reduce interfacial tension

between oil and water in order to enhance the natural process of dispersion by

generating larger numbers of small droplets of oil that are entrained into the

water column by wave energy (NRC 2005). Dispersants are known to

emulsify petroleum by reducing the interfacial tension between petroleum and

water. The small droplets that are formed were dispersed into a water column

to a depth of several meters, preventing wind-induced drift of the oil slick.

The dispersants used in the treatment of petroleum contaminated sites were

19

highly toxic (Tjessen et al 1984); however, less toxic dispersants have

subsequently been developed. Mulkins-Phillips and Stewart (1974) reported

that the oil dispersants should be selected based on three main criteria. They

are: (i) they should be biodegradable; (ii) in the presence of oil, they must not

be preferentially utilized as carbon source; and (iii) they must be nontoxic to

indigenous bacteria. The amounts of modern dispersants necessary for oil

degradation are several-fold smaller than biosurfactant, yet their toxicity

remains a problem (Lessard and Demarco 2000).

2.4.1.1 Chemical Surfactants

Surfactants are amphiphilic molecules having two major

components (moieties): a hydrophilic, or water soluble, head group and a

hydrophobic, or water insoluble, tail group. This dual nature causes

surfactants to adsorb at interfaces thereby reducing the interfacial energies

(Rosen 1989). Surfactants are used to remove PAHs from the contaminated

sites. Jimenez and Bartha (1996) studied the mineralization of pyrene by

Mycobacterium sp. in the presence of Triton X 100 at concentrations below

and above critical micelle concentrations (CMC). Triton X 100 below the

CMC increased the pyrene mineralization rate to 154% and above the CMC

severely inhibited the pyrene mineralization.

Jin et al (2007) investigated the effects of concentration, polar/ionic

head group, and structure of surfactants on the biodegradation of polycyclic

aromatic hydrocarbons (PAHs) in the aqueous phase, as well as their effects

on the bacterial activity. The degradation of 14C-phenanthrene showed either a

decrease or no obvious change with the surfactants present at all tested

concentrations (5–40 mg/L). Thus, the surfactant addition was not beneficial

to remove phenanthrene or other PAH contaminants. This is because

surfactants at higher concentration inhibit the microbial activity; so the

20

preferential utilization of surfactants at low concentration acts as the non-

toxic nutrient resource. Biodegradation of phenanthrene was also influenced

by the surfactant concentration, type of head group and structure. The toxicity

of the surfactants was also studied and ranked as: non-ionic surfactants

(Tween 80, Brij30, 10LE and Brij35) < anionic surfactants- Linear Alkyl

Sulfonate (LAS) < cationic surfactants -Tetra decyl trimethyl ammonium

bromide (TDTMA). For the same head group and similar molecular structure,

the toxicity to the bacteria is due to the chain length, in which the toxicity

becomes lower as the chain length increases.

2.4.1.2 Biosurfactants

Biosurfactants are biomolecules containing both a lipophilic and

hydrophilic moiety. The lipophilic part is the hydrocarbon chain of a fatty

acid or sterol ring. The polar or hydrophilic part is the carboxyl group of fatty

acids or amino acids, the phosphoryl group of phospholipids, hydroxyl group

of saccharides, and peptides. Most of the biosurfactants are produced by

bacteria, yeasts, and fungi during cultivation on various carbon sources

(Healy et al 1996). Biosurfactants act as an important tool for the biotreatment

of hydrocarbon-polluted environments. These compounds increase the

availability of hydrophobic substrates to indigenous degrading

microorganisms. Furthermore, their biodegradability, production from

renewable resources and functionality under extreme conditions are useful

characteristics that offer advantages over chemical surfactants (Banat 1995;

Gutnick and Shabtai 1987; Jain et al 1992).

Vasudevan and Arulazhagan (2007) showed that Sodium Dodecyl

Sulphate and Tween–80 and rhamnolipid biosurfactant produced by

Pseudomonas fluorescens NSI could recover 98% of PAHs from the

contaminated soil.

21

Biosurfactant producing organisms and PAHs utilizing organisms

enhance the recovery and biodegradation of PAHs compounds (Deziel et al

1996). The main applications of biosurfactant are enhanced oil recovery,

removal of organic compounds from soil and formation of emulsions that

facilitate assimilation by microorganisms and application for therapeutic

purposes (Lin 1996). Noordman and Janssen (2002) reported rhamnolipid

biosurfactant produced by P. aeroginosa strain UG2 enhanced fast uptake of

aliphatic hydrocarbon at 73% with 0.2 mg of rhamnolipid/mL.

Van Dyke et al (1993) reported that the rhamnolipids produced by

P. aeroginosa UG2 was capable of recovering 40-78% of PAHs (anthracene,

naphthalene, phenanthrene and fluorene) from soil.

The bacterial strains isolated from marine sediments were capable

of degrading PAHs. Ross et al (2002) reported the ability of sediment bacteria

to utilize polycyclic aromatic hydrocarbons when present as components of

mixtures. Mycobacterium flavescens utilized fluoranthene in the presence of

pyrene whereas the utilization of pyrene was slower in the presence of

fluoranthene than in its absence. He also reported that Rhodococcus sp,

utilized fluoranthene in the presence of anthracene. Daane et al (2001)

isolated Paenibacillus sp from contaminated estuarine sediment and salt

marsh rhizosphere which was capable of degrading naphthalene,

phenanthrene or biphenyl as sole carbon source. Several other bacterial strains

such as Novosphingobium pentaromativorans, Neptunomonas naphthovorans,

Rhodococcus, Acinetobacter and Pseudomonas isolated from marine

sediments were found to be capable of degrading PAHs (Sohn et al 2004),

Hedlund et al 1999, Yu et al 2005).

2.4.2 Direct Photolysis and UV/ H2O2 Oxidation

In Advanced Oxidation Processes (AOPs), by which hydroxyl

radicals are generated in order to destroy the organic contaminants. This is an

22

alternative to biodegradation for removal of PAHs from aqueous solutions

(Miller and Olejnik 2004). Shemer and Lindane (2007) studied the

photodegradation of a mixture of three polycyclic aromatic hydrocarbons-

fluorene (FLU), dibenzofuran (DBF), and dibenzothiophene (DBT) using UV

and UV/H2O2 processes. Toxicity testing using a luminescent inhibition

bioassay was correlated to intermediates generated during UV-based

oxidation reactions. An inhibition of luminescence toxicity assay indicated

formation of toxic intermediates generated during UV-based photolysis and

oxidation reactions. Subsequent oxidative degradation of these by-products

along with the parent compounds resulted in reduced toxicity. From the study

it was also important to note the fact that the DBF and FLU did not show

toxicity after exposure to 1000 mJ/cm2 which does not necessarily mean that

there was no toxic by-products in the irradiated solution. The highly reactive

hydroxyl radicals, generated during AOPs, can lead to complete

mineralization of the pollutant but most typically lead to formation of

products of higher polarity and solubility in water such as phenols, quinones,

and acids (Beltran et al 1996). These metabolites may be far more toxic as

compared to their parent compounds (El-Alawi et al 2002).

2.4.3 Ozonation

Zeng et al (2000) examined pyrene degradation pathways using

ozonation along with biological method in batch and packed column reactors.

After different ozonation times, samples containing reaction intermediates

and byproducts from both reactors were collected, identified for organic

contents, and further biologically inoculated to determine biodegradability.

Ozone (O3) pretreated samples incubated for 5, 10, 15, and 20 days, produced

intermediates such as 4,5-phenanthrenedialdehyde, 2,2',6,6'-

biphenyltetraaldehyde, and long-chain aliphatic hydrocarbons. Further

23

oxidation was carried out via reactions with both O3 and OH- until complete

mineralization.

The integrated treatment of ozone chemical pretreatment and

biological oxidation was reported for the remediation of a phenanthrene-

contaminated soil (Kemenade et al 1995). Few mechanistic details are

available on the aqueous phase degradation of PAHs by O3, and rarely have

intermediates and reaction products been clearly identified. Integrated

chemical and biological processes are potentially more effective than either

one alone (Scott and Ollis 1995; 1996).

2.5 BIOREMEDIATION

Bioremediation is the use of living organisms, primarily

microorganisms, to degrade the environmental contaminants into less toxic

forms. In this process, naturally occurring bacteria and fungi/plants were used

to degrade or detoxify substances hazardous to human health and the

environment. The microorganisms may be indigenous to a contaminated area

or they may be isolated from elsewhere and introduced into the contaminated

site. Toxic compounds are transformed by living organisms through reactions

that take place as a part of their metabolic processes. Bioremediation

techniques are typically more economical than traditional methods such as

incineration; and some pollutants can be treated on site, thus reducing

exposure risks to clean-up personnel, or potentially wider exposure as a result

of transportation accidents. Since bioremediation is based on natural

attenuation, the public considers it more acceptable than other technologies

(Vidali 2001).

When microorganisms are imported to a contaminated site to

enhance degradation, the process is known as bioaugmentation.

24

Biodegradation of a compound is often a result of the actions of multiple

organisms. For bioremediation to be effective, microorganisms must

enzymatically attack the pollutants and convert them to harmless products.

Like other technologies, bioremediation has its limitations. Some

contaminants, such as chlorinated organic or high aromatic hydrocarbons, are

resistant to microbial attack. They are degraded either slowly or not at all.

Hence, it is not easy to predict the rates of clean-up for a bioremediation

exercise; there are no rules to predict if a contaminant can be degraded (Vidali

2001).

As bioremediation can be effective only when environmental

conditions permit microbial growth and activity, its application often involves

the manipulation of environmental parameters to allow microbial growth and

degradation to proceed at a faster rate. Bio-remediation was found to have

cost and technical advantages (Vogel et al 1996). Li et al (2008) inoculated

microbial consortia (bacteria, fungi and bacteria–fungi mixtures) to degrade

PAHs in soil contaminated with oil. The highest PAH removal was observed

in the inoculation with fungal consortia, both in the soil and in the slurry. The

microbial consortia grown on phenanthrene and pyrene efficiently degraded

three to five ring PAHs (anthracene, fluoranthene, and benz(a)anthracene) in

the polluted medium. The study concluded that using microbial consortia

isolated from contaminated soil to remediate the original contaminated soil is

an effective method of bioremediation. Vasudevan and Rajaram (2001)

reported that wheat bran amended soil showed 76% hydrocarbon removal

compared to 66% from oil sludge with inorganic nutrient amendment.

Srikanth et al (2007) also reported that wheat bran enhanced the

bioremediation of anthracene in contaminated soil.

The bacterial strains isolated from marine sediments were capable of

degrading PAHs. Ross et al (2002) reported the ability of sediment bacteria to

25

utilize polycyclic aromatic hydrocarbons (PAHs) when present as components

of mixtures. One strain, identified as Mycobacterium flavescens, which

utilized fluoranthene in the presence of pyrene, although utilization of pyrene

was slower in the presence of fluoranthene than in its absence. The second

strain, a Rhodococcus species, utilized fluoranthene in the presence of

anthracene. Daane et al (2001) isolated Paenibacillus sp from contaminated

estuarine sediment and salt marsh rhizosphere which was capable of

degrading naphthalene, phenanthrene or biphenyl as sole caron source.

Several other bacterial strains such as Novosphingobium pentaromativorans,

Neptunomonas naphthovorans, Rhodococcus, Acinetobacter and

Pseudomonas isolated from marine sediments were found to be capable of

degrading PAHs (Sohn et al 2004), Hedlund et al 1999, Yu et al 2005).

2.6 MICROBIAL DEGRADATION OF PAHs

An understanding of microbial degradation pathways of PAHs is

necessary to control biodegradation and biotechnological systems. The ability

to degrade PAH is not limited to individual species, but occurs in various

groups of micro- organisms and also in thermophilic microorganisms

(Fritkenhauer et al 1996). PAH was metabolised not only by cytochrome

P-450 monooxygenase in mammalian cells, but also by a large number of

enzymes in bacteria, fungi and algae. These microorganisms are able to

oxidize PAH with specific dioxygenases to form cis-dihydrodiols. The

degradation pathway permits complete metabolisation and mineralization of

2- and 3 ring PAH (Figure 2.3) and proceeds via

(1) formation of cis-dihydrodiol

(2) dehydrogenation to form dihydroxy PAH

(3) extradiol ring cleavage

26

(4) release of C3/C2 and C1 compounds (elimination of the first

ring)

(5) decarboxylation of the hydroxyl napthoic acid

(6) extradiol ring cleavage and degradation of the (second) ring

and on further reactions carbon dioxide and water was formed

as end products (Kästner 2000).

Error!

Figure 2.3 Microbial degradation of PAHs (Kästner 2000)

2.6.1 Biodegradation of PAHs by Algae

Algae were found to play a vital role in the degradation of PAHs. In

most studies, degradation of individual model compounds by various algal

species has been reported. Degradation of naphthalene, a major component of

the water-soluble fraction of crude oil, by Oscillatoria sp. (strain JCM) was

O- methyl

27

reported by Narro et al (1992a). The unicellular marine cyanobacterium

Agmenellum quadruplicatum PR-6 metabolized phenanthrene (Narro et al

1992b). Marine cyanobacteria have been shown to oxidize aromatic

hydrocarbons under photoautotrophic growth conditions (Cerniglia et al 1980;

Cerniglia 1984). Oil degradation by Microcoleus chthonoplastes and

Phormidium corium was reported from the Arabian Gulf coasts

(Al-Hasan et al 1994).

Raghukumar et al (2001) used marine cyanobacteria Oscillatoria

salina, Plectonema terebrans and Aphanocapsa sp. which degraded Bombay

High crude oil when grown in filter-sterilized artificial seawater with nutrients

(Nitrogen and phosphate at a ratio of 6:1) and in natural seawater. Around

45–55% of the total fractions of crude oil (containing 50% aliphatics, 31%

waxes and bitumen, 14% aromatics and 5% polar compounds) were removed

in the presence of these cultures within 10 days. Between 50% and 65% of

pure hexadecane and 20% and 90% of aromatic compounds (anthracene and

phenantherene) were removed within 10 days. On the whole, mixed cultures

of the three cyanobacterial species removed over 40% of the crude oil. Hence,

the culture was capable of mitigating the oil polluted seashores, either

individually or in combination without addition of any nutrients.

2.6.2 Biodegradation of PAHs by Fungi

Investigations into the microbial bioconversion of PAHs have

shown that wood-decay fungi causing white-rot are efficient degraders of

these organo-pollutants (Sutherland et al 1995). Generally, ligninolytic

enzymes such as lignin peroxidase (LiP; formerly “ligninase”), MnP, and

laccase, were involved in the degradation of a wide range of organo-pollutants

including polycyclic aromatic hydrocarbons (Hammel et al 1986; Kästner

2000; Pointing 2001). Bumpus et al (1985) first reported that the white rot

28

basidiomycete Phanerochaete chrysosporium partially degraded

benzo(a)pyrene to carbon dioxide by the action of their ligninolytic enzymes

such as lignin peroxidase (LiP) and manganese peroxidase (MnP). Litter

decomposing fungi consist mostly of basidiomycetes, ubiquitously occurring

in forests and grasslands, where they colonize the upper soil and humus layers

(Dix and Webster 1995).

Colombo et al (1996) reported that biodegradation of aliphatic and

aromatic hydrocarbons by natural soil microflora and seven fungal species,

including imperfect strains and higher level ligninolytic species. The natural

microbial soil assemblage isolated from an urban forest area was unable to

significantly degrade crude oil, whereas pure fungal strains effectively

reduced the residues by 26-35% in 90 days. Aspergillus terreus and Fusarium solani isolated from oil polluted areas were more efficient in degrading

aliphatic and aromatic hydrocarbons, respectively.

Giraud et al (2001) studied the role of fungi in treating the water

contaminated by polycyclic aromatic hydrocarbons (PAHs), particularly

fluoranthene in Pilot-scale constructed wetlands. About 40 fungal species

(24 genera) were isolated and identified from a contaminated wetland out of

which 33 species degraded over 70% of fluoranthene and anthracene

efficiently. Species such as Absidia cylindrospora, Cladosporium sphaerospermum, and Ulocladium chartarum are able to degrade the PAH-

model compounds.

2.6.3 Biodegradation of PAHs by Bacteria

In recent years, a variety of bacterial strains have been isolated that

have the ability to utilize PAHs as the sole source of carbon and energy.

These bacteria belong to non halophilic environment as well as halophilic

environment (Table 2.4).

29

Table 2.4 Polycyclic aromatic hydrocarbons degraded by different

species of bacteria (adapted from Cerniglia 1992)

PAH Organisms

Naphthalene Acinetobacter calcoaceticus, Alcaligenes denitrificans, Mycobacterium sp., Pseudomonas sp., P. putida, P. fluorescens, Sphingomonas paucimobilis, Brevundimonas vesicularis, Burkholderia cepacia, Comamonas testosteroni, Rhodococcus sp., Corynebacterium renale, Moraxella sp., Streptomyces sp., B. cereus, P. marginalis, P. stutzeri, P. saccharophila, Neptunomonas naphthovorans, Cycloclasticus sp.

Acenaphthene Beijernickia sp., P. putida, P. fluorescens, Burkholderia cepacia, Pseudomonas sp., Cycloclasticus sp., Neptunomonas naphthovorans, Alcaligenes eutrophus, Alcaligenes paradoxus

Phenanthrene Aeromonas sp., A. faecalis, A. denitrificans, Arthrobacter polychromogenes, Beijernickia sp., Micrococcus sp., Mycobacterium sp., P. putida, Sphingomonas paucimobilis, Rhodococcus sp., Vibrio sp., Nocardia sp., Flavobacterium sp., Streptomyces sp., S. griseus, Acinetobacter sp., P. aeruginosa, P. stutzeri, P. saccharophila, Stenotrophomonas maltophilia, Cycloclasticus sp., P. ¯uorescens, Acinetobacter calcoaceticus, Acidovorax delafieldii, Gordona sp., Sphingomonas sp., Comamonas testosteroni, Cycloclasticus pugetii, Sphingomonasyanoikuyae, Agrobacterium sp., Bacillus sp., Burkholderia sp., Sphingomonas sp., Pseudomonas sp., Rhodotorula glutinis, Nocardioides sp., Flavobacterium gondwanense, Halomonas meridiana

30

Table 2.4 (Continued)

PAH Organisms

Anthracene Beijernickia sp., Mycobacterium sp., P. putida, sp. paucimobilis, Burkholderia cepacia, Rhodococcus sp., Flavobacterium sp., Arthrobacter sp., P. marginalis, Cycloclasticus sp., P. fluorescens, sp. yanoikuyae, Acinetobacter calcoaceticus, Gordona sp., Sphingomonas sp., Comamonas testosteroni, Cycloclasticus pugetii

Fluoranthene A. denitrificans, Mycobacterium sp., P. putida, Sphingomonas paucimobilis, Burkholderia cepacia, Rhodococcus sp., Pseudomonas sp., Stenotrophomonas maltophilia, Acinetobacter calcoaceticus, Acidovorax delafieldii, Gordona sp., Sphingomonas sp., P. saccharophilia, Pasteurella sp.

Pyrene A. denitrificans, Mycobacterium sp., Rhodococcus sp., Sphingomonas paucimobilis, Stenotrophomonas maltophilia, Acinetobacter calcoaceticus, Gordona sp., Sphingomonas sp., P. putida, Bu cepacia, P. saccharophilia

Chrysene Rhodococcus sp., P. marginalis, Sphingomonas paucimobilis, Stenotrophomonas maltophilia, Acinetobacter calcoaceticus, Agrobacterium sp., Bacillus sp., Burkholderia sp., Sphingomonas sp., Pseudomonas sp., P. saccharophilia

Benz[a]anthracene A. denitrificans, Beijernickia sp., P. putida, Sphingomonas paucimobilis, Stenotrophomonas maltophilia, Agrobacterium sp., Bacillus sp.,Burkholderia sp., Sphingomonas sp., Pseudomonas sp., P. saccharophilia

Dibenz[a,h] anthracene

Sphingomonas paucimobilis, Stenotrophomonas maltophilia

31

2.6.3.1 Biodegradation by Non-halophilic bacteria

The bacteria which do not require salt (NaCl) for its growth are

known as non-halophilic bacteria. Carter et al (2000) identified Rhodococcus

opacus, Thalassolituus oleivorans, Chromohalobacter salinarum and

Sphingomonas sp. as PAH degrading bacteria in freshwater environment.

Srikanth et al (2005) showed that when pyrene supplied as microcrystals than

crystals, it increased overall pyrene mineralisation by Gordona BP9 from

53 - 58% at mineralization rates of 160 ng/mL/h and 166 ng/mL/h. Limited

number of PAH degrading bacteria have been identified in freshwater

systems. Chang et al (2000) reported the freshwater culture was mostly

dominate - -proteobacteria. PAHs present in soil

exhibit toxic activity towards different plants, microorganisms and

invertebrates. Microorganisms, being in intimate contact with the soil

environment, are considered to be the best indicators of soil pollution. In

general, they are very sensitive to low concentrations of contaminants and

rapidly respond to soil perturbation. Bioremediation process is an effective

way to decontaminate PAHs-contaminated soils.

During the last few decades, a variety of bacteria capable of

degrading PAHs, particularly low-molecular weight compounds, were

discovered. Most of these bacteria belong to the genera Agmenellum,

Aeromonas, Alcaligenes, Acinetobacter, Bacillus, Berjerinckia, Burkholderia,

Corynebacterium, Cyclotrophicus, Flavobacterium, Micrococcus, Moraxella,

Mycobacterium, Nocardioides, Pseudomonas, Lutibacterium, Rhodococcus,

Streptomyces, Sphingomonas, Stenotrophomonas, and Vibrio (Kim et al 2005;

Juhasz et al 2000; Daane et al 2002; Van Hamme et al 2003). Moreover, some

studies have shown that bacteria such as Mycobacterium, Rhodococcus,

Alcaligenes, Pseudomonas and Sphingomonas were able to grow on the four-

ring PAHs (Boldrin et al 1993; Kästner et al 1994; Dagher et al 1997).

32

2.6.3.2 Biodegradation by Halophilic bacteria

Microorganisms requiring salt for growth are referred to as

“halophiles” (Salt Lover). Microorganisms that are able to grow in the

absence as well as in the presence of salt are designated halotolerant species.

Extreme halophiles require generally at least 1 M NaCl (approx. 6% w/v) for

growth, and grow optimally at NaCl concentrations above 3 M (Kushner

1978; Grant et al 1998) (Table 2.5).

Halophiles are found in industrial plants that produce salt by

evaporation of seawater and salted proteinaceous materials such as salted fish.

In solar lanterns, gram positive aerobic heterotrophs are not as common as

gram negative bacteria but similar species of the genera e.g., Marinococcus,

Sporosarcina, Salinococcus and Bacillus have been recovered from saline

soils and salterns (Smith 2000).

Table 2.5 Classification of microorganisms according to salt resistance

(Kushner 1978; 1993)

Category Salt concentration (M)

Range Optimum

Nonhalophile 0-0.1 <0.2

Slight halophile 0.2-2.0 0.2-0.5

Moderate halophile 0.4-3.5 0.5-2.0

Borderline extreme halophile 1.4-4.0 2.0-3.0

Extreme halophile 2.0-5.2 >3.0

Halotolerant 0->1.0 <0.2

Haloversatile 0->3.0 0.2-0.5

33

Halobacteria (or haloarchaea) is the common name applied to

members of the Class Halobacteria (Order Halobacteriales), and consists of

extremely halophilic Archaea. Halobacteriales includes about 15 genera

namely Haloarcula, Halobacterium, Halobaculum, Halococcus, Haloferax,

Halorubrum, Halogeometricum, Halorhabdus, Haloterrigena, Natrialba,

Natrinema, Natronobacterium, Natronococcus, Natronomonas,

Natronorubrum (Smith 2001). Most genera are found in normal salt lakes and

salterns, whereas the natronobacteria are found in highly alkaline soda lakes

(e.g. Wadi Natrun in Egypt, pH around 11). Some halobacteria have been

isolated from beach sand (Natrialba), and salty soils (Halobacterium

distributum). Most laboratory studies have used members of only three

genera, Halobacterium, Haloferax, and Haloarcula (Smith 2000).

A range of organic pollutants have been shown to be mineralized

or transformed by microorganisms able to grow in the presence of salt

(Oren et al 1992; Margesin and Schinner 2001). Eubacteria are more

promising degraders than archaea as they have a much greater metabolic

diversity. Their intracellular salt concentration is low, though their enzymes

involved in biodegradation are conventional (i.e. not salt-requiring) enzymes

similar to those of non-halophiles (Oren et al 1992). However, halophilic

archaea maintain an osmotic balance with the hypersaline environment by

accumulating high salt concentrations, which requires adaptation of the

intracellular enzymes to varying salt concentrations (Oren et al 1992).

Diverse petroleum-degrading bacterial strains inhabit in marine

environments. They have often been isolated as degraders of alkanes or

aromatic compounds such as toluene, naphthalene and phenanthrene. Several

marine bacteria capable of degrading petroleum hydrocarbons have been

newly isolated, which includes Pesudomonas, Flavobacterium, Marinobacter

and Paenibacillus sp (Gauthier et al 1992, Daane et al 2001). Although

34

terrestrial PAHs degrading bacteria such as Pseudomonas sp have been found

in marine environment, most recent studies have indicated that these are

obligately marine PAHs degradation genera. Work done by members of

Gieselbrecht et al (1998) who have isolated numerically important PAHs

degrading bacteria from Gulf of Mexico has resulted in the discovery of the

genus Cycloclasticus, Neptunomonas (Hedlund et al 1999). All these bacteria

require sodium for good growth and therefore can be considered obligatory

marine organisms.

2.6.3.3 Moderately halophilic bacteria

Moderately halophilic bacteria constitute a group able to grow in a

wide range of saline environments. Several studies on their molecular

adaptation to media with different salt concentrations has increased scientific

interest in these bacteria (Nieto and Vargas 2002). Garcia et al (2004) isolated

a moderate halophile (Halomonas organivorans sp. nov.,) able to degrade a

wide range of aromatic compounds (benzoic acid, p-hydroxybenzoic acid,

cinnamic acid, salicylic acid, phenylacetic acid, phenylpropionic acid, phenol,

p-coumaric acid, ferulic acid and p-aminosalicylic acid), used for the

decontamination of polluted saline habitats.

Halophiles grow by using a limited number of organic compounds

including the aromatic hydrocarbons biphenyl, naphthalene, phenanthrene and

toluene as sole carbon sources. They grow poorly on media containing no

aromatic compounds and require at least 10% salinity for growth

(Gieselbrecht et al 1998). An unidentified halophilic archaeon degraded PAHs

(acenaphthene, phenanthrene, anthracene; at a concentration of 500 mg/L

each) as well as saturated hydrocarbons (C14, C16, C18, C21, pristane) in a

medium prepared with natural hypersaline water from a salt marsh (21% w/v

NaCl). No growth on hydrocarbons occurred below 11% w/v NaCl (Bertrand

35

et al 1990). Four bacterial strains, belonging to the genera Micrococcus,

Pseudomonas and Alcaligenes tolerating 7.5% w/v NaCl, could grow on 0.1%

naphthalene and anthracene (Ashok et al 1995).

Zhuang et al (2002) isolated Bacillus naphthovorans MN-003 from

tropical marine sediments in Singapore contaminated with marine fuel oil.

The strain was able to grow between 0.28 to 7.00% NaCl concentrations but

the optimum salinity was found to be from 1.75 to 3.50%. The strain MN-003

showed no growth with phenanthrene and anthracene, but utilized benzene,

toluene, xylene isomers and diesel oil as sole carbon sources.

Daane et al (2001) reported the presence of PAHs degrading

bacteria such as Paenibacillus sp. PR-P1 and Arthroacter sp. PR-P3 from

antartic lakes which could utilize naphthalene, phenanthrene, or biphenyl as

the sole source of carbon and energy in salt marsh plant systems.

Hedlund and Staley (2001) studied PAH degradation using Vibrio

cyclotrophicus isolated from creosote contaminated marine sediments which

degraded phenanthrene as sole carbon source. The organism also utilized

several two and three ring PAHs such as naphthalene and

2-methylphenanthrene as substrates. Engelhardt et al (2001) isolated novel

hydrocarbon degrading gram positive bacterium from inter-tidal beach

sediments which tolerated salt up to 3.3%. The strain was capable of

degrading n-alkanes in crude oil from C11 to C33, but was unable to degrade

aromatic hydrocarbons. Tam et al (2002) reported 90% of phenanthrene and

fluorene degradation in 7 days using bacterial consortium enriched with Sai

Keng and Ho Chung sediments at 2% salinity.

Many hydrocarbonoclastic bacteria have been isolated. Some

examples are Vibrio, Pseudoalteromonas, Marinomonas and Halomonas,

36

which are capable of degrading phenanthrene or chrysene (Melcher et al

2002). Some hydrocarbon-degrading bacteria isolated from marine

environments have been classified into several genera that include terrestrial

hydrocarbon degrading bacteria namely, naphthalene-degrading

Staphylococcus and Micrococcus (Zhuang et al 2003), 2-methylphenanthrene-

degrading Sphingomonas (Gilewicz et al 1997) and alkane-degrading

Geobacillus (Maugeri et al 2002).

2.6.3.4 Extremely halophilic bacteria

Field experiments have shown that oil oxidizing microflora were

widely distributed in polluted sites of Estonia. Oil oxidising organisms were

isolated from stratal waters with salinities of up to 272 mg/L. Only single

oxidising microbial cells were found in stratal waters of production wells. Oil

oxidising eubacteria were found to be active in media with salinities up to

15% sodium chloride while the extremely halophilic oil oxidising

archaebacteria were active at salinities up to 32%. Sodium chloride was

isolated from oil samples of Bondyzhskoye oil field. The archaebacteria also

possessed high oil emulsifying activity (Rozkov et al 1998).

Diaz et al (2002) reported that halotolerant bacterial consortium

isolated from Colombian mangrove sediment was able to treat various

hydrocarbons immobilized onto polypropylene fibres. A wide range of

salinity was used in this study (0 to 180 g/L). In the study, free cells degraded

4 to 49% of PAHs, while the immobilized cells degraded 26.8 to 65% of

PAHs (Phenanthrene and Napthalene) respectively.

Zhao et al (2006) used two different bacterial consortia in two

reactors to analyse the PAH degradation efficiency. Reactor A with B350 and

Reactor B with B350M of bacterial consortia immobilized with biological

37

aerated filter (BAF). B350 and B350M were able to degrade 90% and 84% of

PAHs in oil field wastewater in 120 days. The consortia was able to treat

PAH contaminated oil field wastewater with high salinity and without

additional nitrogen and phosphate. The bacterial consortium in reactor B

(B350M) was found to be more effective in degrading the PAHs and grew

well especially on chrysene. Results obtained using DGGE analysis showed

that 20 different bacteria in reactor B and 13 different bacteria in reactor A

were present respectively.

2.7 FACTORS INFLUENZING POLYCYCLIC AROMATIC

HYDROCARBON DEGRADATION IN MARINE

ENVIRONMENT

The fate of petroleum hydrocarbons in the environment is largely

determined by abiotic factors which influence biodegradation of the oils.

Factors which influence rates of microbial growth and enzymatic activities

affect the rates of petroleum hydrocarbon biodegradation. The persistence of

petroleum pollutants depends on the quantity and quality of the hydrocarbon

mixture and on the properties of the affected ecosystem (Atlas 1981).

2.7.1 Salinity

One of the first experiments on the effects of salt on biological

wastewater treatment was conducted by Zobell et al (1937). In this research,

dilutions of water collected from the Great Salt Lake in Utah were used to

prepare agar plates which were inoculated with organisms from the domestic

sewage, soil and other sources. None of the sewage organisms and less than

1% of the soil bacteria survived in full strength lake water (at 28% salt conc.).

Higher survival rates were recorded in dilutions containing even less salt.

However, only 7-18% of the sewage bacteria grew on medium containing

38

2.8 to 7% salt. Analogous studies conducted with seawater showed similar

results (Zobell 1946). Only 13% of the sewage organisms grew on nutrient

medium made with full strength seawater (approximately 3.5% salt). Similar

results were obtained with soil organisms. A slight stimulatory affect was

observed when soil and sewage organisms were grown in medium containing

10% seawater (0.35% salt).

Ward and Brock (1978) examined hydrocarbon biodegradation in

hypersaline environments. Hydrocarbons were added to natural samples

having various salinities (from 3.3 to 28.4%) from salt evaporation ponds of

Great Salt Lake, Utah. Rates of metabolism of these compounds decreased as

salinity increased. Similar results were obtained by Diaz et al (2002); Oren

et al (1992); Rambeloarisoa et al (1984); Mille et al (1991) and Bertrand

et al (1993). Thus, salinity plays major role in the degradation of PAHs.

2.7.2 Nutrients

In a natural marine environment, the amount of nutrients, especially

those of nitrogen and phosphorus, are insufficient to support the microbial

requirements for growth, after a sudden increase in the hydrocarbon level

associated with an oil spill. Therefore, nitrogen and phosphorus as nutrients

may be added to the contaminated environment to stimulate the growth of

hydrocarbon degrading microorganisms, thereby increasing the rate of

biodegradation of polluting hydrocarbons (Harayama et al 2004; Rodriguez-

Valera et al 1981).

In general, assessment of nutrient requirements become more

complex as the salt concentration of the medium is increased. The nutritional

characteristics of halophilic bacteria are inherently difficult to assess because

the salt content of the growth medium affects nutrient requirements

39

(Rodriguez-Valera 1988; Hochstein 1987). For example, complex medium

containing peptone and casamino acids could support the growth of Vibrio

costicola, a moderate halophile, over a wider range of salt concentrations than

defined medium containing only inorganic salts and glucose (Forsyth and

Kushner 1970).

The microbial degradation of petroleum hydrocarbon pollutants in

open systems, such as lakes, oceans, and wastelands, is limited by a utilizable

source of nitrogen and phosphorus (Atlas and Bartha 1972b; Rosenberg et al

1998). Since petroleum contains only traces of nitrogen, the required nitrogen

must come from the surrounding environment. To overcome the nitrogen

limitation for petroleum degradation in open systems, Atlas and Bartha (1973)

studied the effectiveness of several oleophilic nitrogen compounds with low

C/N ratios. Subsequently, an oleophilic fertilizer (Inipol EAP 22) was used in

the bioremediation of polluted shorelines after the Exxon Valdez spill (Atlas

and Bartha 1973). The addition of nitrogen and phosphate was proved to be

an effective bioremediation treatment on several shorelines (Swannell et al

1996; Swannell et al 1999; Venosa et al 1996). Koren et al (2003) reported

that in a simulated open system, uric acid as nitrogen source bound to crude

oil potentially enhanced the bacterial growth and petroleum biodegradation.

As a result of nutrient deficiency, many halophiles have not

developed complete biosynthetic pathways and have complex nutrient

requirements. Although certain organisms can grow on defined medium

containing only glucose, salts and ammonia, many halophiles require growth

factors such as amino acids or vitamins. These growth factors have

historically been supplied using rich medium formulations that contain yeast

extract and protein hydrolysates (Kushner and Kamekura 1988). The

historical use of complex media and the difficulties associated with assessing

specific nutrient requirements have led to a shortage of information on the

40

requirements for halophile growth. In addition, data on the biodegradation of

environmental compounds by halophilic bacteria is limited. Further use of

halophilic bacteria in waste treatment processes, additional data on the

nutritional requirements and the ability of these organisms to degrade

common pollutants would be necessary (Oren et al 1992).

2.7.3 pH

Seawater pH is about 8. It was reported by Zaidi et al (1988), that

any fluctuation in pH higher than 8.0 may potentially slow down the

degradation of PAHs in the sea waters. A neutral to slightly alkaline pH was

best for the growth of nonalkaliphilic halobacteria (e.g. pH 7.2 to 7.5). For

alkaliphilic halobacteria, the optimum pH ranged between 8.5 to 9.5 (Smith

2001).

2.7.4 Temperature

Despite living in natural waters all around the world, most PAH

degrading strains grow best between 30

thermophilic (e.g. 50oC for Halorubrum saccharovorum) (Smith 2001). Even

the Antarctic isolate, Halorubrum lacusprofundi, grows optimally at about

30 an conveniently be

used for both E. coli and most halobacterial strains. The only problem is the

drying out of agar plates over longer incubation periods which can be avoided

by use of plastic wrappers or plastic containers (Smith 2001).

2.7.5 Oxygen

Oxygen supply is a problem at high salt solutions (oxygen

solubility decreases with increasing salt), especially at raised temperature

41

(Smith 2001). Sediment tilling has been evaluated as a bioremediation

treatment to increase the penetration depth of oxygen and nutrient

supplements. Use of chemical oxidants such as hydrogen, calcium and

magnesium peroxides to alleviate oxygen deficiency within sediments has

also been considered. While commercial forms of these products have been

used in terrestrial environments for groundwater remediation, their

application in the marine environment has not been addressed in detail

(Marine Bioremediation Technologies Screening Matrix and Reference Guide

2002).

2.8 POTENTIAL APPLICATIONS OF HALOPHILES

Halophilic bacteria have the potential for exciting and promising

applications. They are as follows:

2.8.1 Bioremediation of Marine Oil Pollution

The potential of bioremediation to treat oil contaminated shorelines

has been established. Bacterial community structure changes in response to

oil spills and subsequent bioremediation treatments, and members of the

alkane-degrading genus Alcanivorax become dominant (Kasai et al 2002).

Hydrocarbon degrading halophilic bacterial consortia isolated from crude oil

and mangrove sediments are capable of treating oily wastes over such a wide

range of salinity (Diaz et al 2000).

A halophilic archaea (strain EH4) was found to be capable of

degrading a wide range of n-alkanes and aromatic hydrocarbons in the

presence of high salt (Bertrand et al 1990; Oren et al 1992, Ward and Brock

1978). Marinobacter hydrocarbonoclasticus degraded a variety of aliphatic

and aromatic hydrocarbons (Gauthier et al 1992). In addition, bacteria isolated

42

from salt-impacted material degraded polycyclic aromatic hydrocarbons

(PAHs) (Plotnikova et al 2001).

2.8.2 Treatment of Saline Wastewater

Halophilic bacteria were mainly used in the biological treatment of

hypersaline industrial wastes. Woolard and Irvine (1995) investigated the

waste treatment potential of halophiles, which were able to degrade phenol in

simulated oil field produced water. Unlike many halophilic species, the

organisms isolated in their study did not have complex nutrient requirements.

Sustained phenol degradation occurred in simple medium containing salt

(15%), ammonia, phosphorus and iron.

Naturally occurring hydrocarbon degrading bacteria in marine

environments are usually found in low numbers. However, pollution by

petroleum hydrocarbons may stimulate the growth of such organisms and

cause changes in the structure of bacterial communities in the contaminated

area (Oren et al 1992). Identification of the key organisms that play roles in

pollutant biodegradation is important for understanding, evaluating and

developing in situ bioremediation strategies. For this reason, huge efforts have

been made to characterize bacterial communities, to identify responsible

degraders, and to elucidate the catalytic potential of these degraders. Kargi

and Ugyur (1997) reported the inclusion of Halobium and Halobacter sp.

along with activated sludge culture resulted in significant COD removal

efficiency at high salt concentrations such as 5%.

Kargi and Dincer (1996a,b; 1997) studied the effect of salt

concentration on the aerobic biological treatment of a synthetic saline effluent

using a fedbatch biological reactor. The synthetic effluent was made up of

diluted molasses, urea, KH2PO4 and NaCl up to a concentration of 50 g/L and

43

characterised by a COD:N:P ratio of 100:10:1. The treatment process used

activated sludge. Kargi and Dincer (1997) observed that the effluent COD

removal efficiency fell from 85% to 59% when salinity increased from 0 to

5%. Dincer and Kargi (2001) also reported that aerobic rotating discs were

able to purify a synthetic effluent under conditions of increasing salinity

(0–10%) with more than 80% of COD removal efficiency at a salt

concentration lower than 50 g /L.

Kargi et al (2000) were able to successfully treat an effluent

generated by the pickling industry using activated sludge enriched in

Halobacter halobium, exceeding 95% of COD removal. The same technique

(inoculation of the halotolerant bacteria Staphylococcus sp. and Bacillus

cereus) applied to another agro-industrial hypersaline effluent (15% of NaCl)

generated by the production of plum pickles achieved COD removal

efficiency of 90% in a sequencing batch reactor (Kubo et al 2001). Lefebvre

and Moletta (2006) reported that even though biological treatment of

carbonaceous, nitrogenous and phosphorus pollution has proved to be feasible

at high salt concentrations, the performance obtained depends on a proper

adaptation of the biomass or the use of halophilic organisms.

2.8.3 Other Potential Applications

The halophiles produce compounds of industrial interest (enzymes,

polymers and Bio makers etc.), which possess useful properties, a few of

which are listed:

Fermented foods: In Sauerkraut (pickled cabbage)

manufacturing processes, optimum salt concentrations range

for Leuconostac mesentroideis is from 2% to 3%. High salt

44

concentrations are associated with development of pink

sauerkraut (Thongthai and Suntinanalert 1991).

Biomarkers: In petroleum exploration, Mycobacterium sp.

were used as bio markers which forms carotenoids and lipids

during hydrocarbon mineralization in the petroleum industry

(Chosson et al 1991).

Enzymes: A number of extracellular and intracellular

enzymes (Amylase produced by Halobacillus sp. strain MA2)

were isolated, characterized and screened for the production of

bioactive compounds like antibiotics (Koyama et al 1994).

Halophiles are also used in recovery of hypersaline waste brines

derived from the olive oil industry and leather or in curing process. Many

halophiles produce orange or pink colonies probably due to production of

carotenoids as a protective mechanism against photooxidation process.

Carotenoids have a major application in the food industry as additives in

health food products (Margesin and Schinner 2001; Rodriguez-Valera 1992).