CHAPTER 2 LITERATURE REVIEW - Shodhgangashodhganga.inflibnet.ac.in/bitstream/10603/26696/7/07... ·...
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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.
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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).
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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
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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.
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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
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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).
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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
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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
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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.
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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.
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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).
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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).
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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.
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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
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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
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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.
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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
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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
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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.
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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
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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
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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).