Soil and aquatic ecosystems contamination by petroleum hydrocarbons (so called TPHs) is a serious global issue. These contaminants originate from the distillates of crude oil such as diesel, gasoline, lubricating oil, and other typical petroleum hydrocarbons. They are highly toxic, mutagenic and carcinogenic in nature, for that they have received much attention globally. As molecular weight increases, the toxicity of petroleum hydrocarbons is also increased. Low molecular weight cyclic alkanes are more toxic to aquatic organisms than aliphatic and aromatic hydrocarbons of same molecular weight.
In terrestrial environment, the scenario is little bit different, where aromatic hydrocarbons are more toxic than aliphatic compounds.
Even lower aromatic compounds other than PAHs are toxic. Importantly, toxicity of petrocarbon in an organism is directly proportional to its bioavailability. Hydrophilic petrocarbons are more bioavailable than the hydrophobic and/or bound petrocarbons. The bioavailable pollutant is highly accessible and to be adsorbed by an organism, and cause sublethal or lethal effects by interacting with specific sites/receptors in the organisms.
During toxicity development, petrocarbons usually disrupts the cell membrane, which results in the fluctuations in membrane fluidity, integrity, and functioning in the target organisms. In aquatic system, non-bioavailable and/or hydrophobic petroleum hydrocarbons become bioavailable to several benthic organisms (e.g. invertebrates, fish, deposited fish eggs etc.) by adsorbing to particulates and sediments.
Certain aquatic invertebrates (e.g. mussels, oysters, crabs, cockles etc.) experience the toxicity of petroleum hydrocarbons by the ingestion of suspended oil droplets/oil bound particulates. Human beings suffer potential health disorders upon exposure to petroleum compound via inhalation, ingestion and dermal contact.
Any effect that affect a population without causing mortality is called sublethal effect, which are usually involved in the development of lesions, developmental defects, changes in molecular functions, behavior changes in feeding and breeding.
Whereas lethal effects occur in aquatic environment due to short term exposure to oil spills, where they disrupt central nervous system through partitioning into cell membranes and nerve tissues. Fatality caused by petroleum hydrocarbons is broadly termed as narcosis. The main conclusion is that the accumulation and persistence of petroleum hydrocarbons in the environment can bring forth harmful effects in both terrestrial and aquatic ecosystems. In this angle, the present chapter has been designed to address several issues concerning TPHs and their ecological impacts on microorganisms, plants and animals (invertebrates and vertebrates) of both terrestrial and aquatic system. Comment by Naga Raju Maddela: Sir, can we mention like this please?
Petroleum hydrocarbons (PHs), upon their discharge into environment, can pose risks to human health, ecosystems and ground water. Foremost components (more than 50%) responsible for the pollution are mineral oil, polycyclic aromatic hydrocarbons (PAHs) or volatile aromatic hydrocarbons (i.e. benzene, toluene, ethylbenzene and xylenes – BTEX); which are wide spread in gasoline, diesel, and lubricants. The physico-chemical and toxicological properties of TPHs of above listed components are varied from one another. Usually, low MW (molecular weight) components have greater toxicological effect, which is attributed to their higher solubility and bioavailability.
For example, lighter volatile hydrocarbons (C5 to C10) are having higher human and ecotoxicological risks. For this reason, practically TPHs are not useful proxy for the assessment of potential environmental risks, rather fractionation of TPHs is more reliable and widely considered to quantitatively determine the PHs effects on the ecosystem. Nevertheless, complex composition of PHs and existence of multiple analytical techniques for TPHs collectively make the human health risk assessments difficult and complicate.
There is a severe pressure on the environmental components and other receptive systems from accidental and incidental discharge of petroleum products, which is mainly due to increasing dependency of countries economy on the PHs exploration and extraction. Toxic organic substances can enter the coastal environments or discharge on to the fertile soils by one of several routes such as leachate and seepage during operations, extraction, transportation, distribution storage, and refining. When the human activities are involved in the above routes, oil spillage can be minimized for some extent with proper monitoring of oil infrastructure with “state-of-the-act” technology.
But the complete elimination of spillage is only possible when an effective regulatory oversight is in place besides latest technologies for detection of oil spills are put in practice. In fact, now a days, oil spill damages can be evaluated very well by using some advanced methodologies such as basic oil spill prediction models, methodology alongside baseline, or near real time data etc. There are other routes that can be controlled to reduce spills are leaks from wellhead, pipelines, overflows and dumping of slurries in the environment.
There is a limited focus on soil quality criteria for ecological receptors when compared to aquatic or human receptors, because the protection of human health is the primary basis of most of guidelines. Additionally, there is no proper integration in the soil quality standards framed by many Member States for ecologically based threshold soil concentrations for many pollutants. For instance, in Europe, guidelines for environmental protection are framed by European Commission Technical Guidance Document on Risk Assessment. In general, ecotoxicological data of the most sensitive species combined with an assessment factor or species sensitive distribution (SSDs) curves are used to derive lower HC (hazardous concentration) values (i.e. HC5).
On the other hand, geometric mean of ecotoxicological data of several species or SSDs are considered to derive higher HC values (i.e. HC50). For certain PHs (e.g. PAHs), due to lack of data, deterministic approach is used to frame the guidelines (Cachada et al., 2016). Still HC values are varied between the countries, which is attributed to many factors such as, differences in the scientific methodology adapted, endpoint of protection, land use considered, political reasons etc. In Dutch, deterministic approach is used to frame the guidelines of environmental risk limits for direct terrestrial ecotoxicity correspond to the MPC (maximum permissible concentration) and SRC (serious risk concentration). MPC values will protect all species in the ecosystem from adverse effects, whereas SRC values indicate the concentration at which soil functions or species are seriously affected. NOEC (no observed effect concentration), and LOEC (lowest observed effect concentration) are the key factors considered by Danish EPA (environmental protection agency) to establishes ecotoxicological soil quality criteria for soil dwelling organisms, plants, microorganisms and microbiological processes, with the primary aim of protecting ecosystem’s function and structure (DEPA, 2002).
According to Spanish guidelines, generic reference values are the basis for the protection of ecosystems for individual compounds (MP, 2005), where values are usually calculated based on PNEC (predicted no effect concentration) values for different environmental components such as water column, soil and wildlife food. Similarly, ecological risk-based numerical standards have been established in Canada by CCME (Canadian Council of Ministers of the Environment). The main aim of CCME is to protect key sustainable ecological receptors with in the four defined land use categories – agricultural, residential/park-land, commercial, and industrial (CCME, 2008). Generally, these values are lower for two land use categories, and higher for later two. For example, SQGE (soil quality guidelines for environmental health protection) values (µg kg-1) of pyrene (a PAH) as per the CCME guidelines are 100, 10000, 100000, and 100000 for agricultural, residential/parkland, commercial, and industrial land use categories, respectively. The main intention of SQG is to ensure and protect biological importance from adverse effects of pollutant based on toxicological data associated with acute, sub-chronic, and chronic responses in the representative species. To sum up, for the different land uses and receptors considered, the site-specific risk assessment is advised. At the higher steps, one must consider the available fraction of contaminants, instead of following deterministic approaches.
Terrestrial (microbes, plants, invertebrates and vertebrates)
Oil spilled on terrestrial environments, though it is volatilized and biodegraded, a significant fraction is infiltered vertically into the subsurface area of soil column, penetrate the soil micropores and remains in the soil matrix for years. On the other hand, the remaining refractory fractions are potentially mutagenic, carcinogenic, and able to bioaccumulate, which are chemically high MW hydrocarbons, resins and polar fractions. Thus, oil spills concerning terrestrial ecosystem are known to be acute, and the recovery of these systems is relatively slow. Both human and wildlife experience the effects of soil pollution either by unintentional or intentional exposure to polluted soil. In this section, how lower and higher organisms are affected by petroleum hydrocarbons upon land spillage is described in depth.
Biogeochemical cycles on Earth are meticulously regulated by soil microflora and there by this microflora play a key role in soil ecosystems. At the same time, soil microbes are the first sufferers if any disturbance caused by soil pollution, this led to changes in metabolic activities of soil microbes, subsequent changes in soil microbial community composition and diversity (Sun et al., 2016). Despite pollutants decrease the diversity of soil microflora, there is an enrichment of pollutant-tolerant species which may in turn affect the overall ecosystem functions and balance of soil microorganisms. For example, crude oil polluted soil had shown nearly 100 times fewer heterotrophic and cultivable bacteria than the unpolluted soil (Maddela et al., 2015), this can be linked to toxic properties and/or concentrations of petroleum hydrocarbons, also, these compounds are usually not a good source of carbon and energy for the most soil microflora. Similar negative effects of petroleum hydrocarbons on soil microflora have been identified in many terrestrial ecosystems. In gasoline-polluted sandy soils, the microbial biomass has been badly affected.
Likewise, the microbial biomass carbon (MBC) is being lowered the toxic effects resulting from the crude oil contaminations in soil. In order to understand more about the negative impacts of petrocarbons on soil microflora, evaluation of soil microbial properties including soil enzyme activities is widely considered. Many PAHs affect the soil enzymes, especially soil urease and dehydrogenase are appearing to be more sensitive to pollution and are used to determine the influence of various pollutants on the microbiological quality of soil. Thus, contamination with petrocarbons have a profound effect on soil fauna. For example, the activities of urease and dehydrogenase in oil contaminated soils have been decreased by ~50% compared with control (Klamerus-Iwan et al., 2015). Similarly, gasoline had shown significant inhibitory effect on the hydrolase activities concerning N, P, and C cycles (urease, protease, phosphatase and β-glucosidase) in clay and sandy soils. Especially light hydrocarbons are more toxic to microorganisms than heavy hydrocarbons.
Soil habitat structure become unfavorable for the survival of microflora after oil spill. There is soil particle smothering and blocking air diffusion into soil pores are the major cause of developing anaerobic conditions at the spilled site, with subsequent effects on soil microbial communities. Even water resistance of soil aggregates is modified by the presence of oily substances rich in aromatic hydrocarbons. In addition to these, crude oil polluted soils are hydrophobic in nature, which is also not much supportive and favorable for microbial growth. Another considerable factor affecting the intensity of influence of petroleum hydrocarbons on soil microflora is the organic matter of that soil. Since the organic matter is capable of adsorbing petroleum hydrocarbons, subsequently the bioavailability of crude oil components to microflora is decreased and their negative effects on soil microflora are minimized. PAHs with more than 4 benzene rings are strongly adsorbed, and thus poorly bioavailable. Similarly, higher the cation exchange capacity and clay content in soil lowers the pollutant impact on microflora.
The study of abiotic stress responses in plants against environmental pollution by petroleum products has become ever more important in many sectors like agriculture, forest management, and ecosystem restoration strategies. Seed germination of several species has been affected negatively by several volatile and hydrophilic aromatics (e.g. benzene, indene, naphthalene, styrene, toluene, and xylene isomers) (Henner et al., 1999). Such results haven’t been changed even when experiments were conducted under different settings such as open versus closed lids, pure PAHs versus composite-contaminated soil, contaminated soil versus contaminated soil leachate, and freshly excavated versus aged vegetated soils. Even though the results are inconsistent regarding the impact of PHs on seed germination, light fractions of (nC9−nC14) were found 20 times more toxic than the heavy fractions (ChaIneau et al., 1995).
It is also important to note that the phytotoxicity of a pollutant is changed with the time of plant growth. For example, volatile branched cyclohexanes of diesel causes detrimental effects on germination and emergence (MacKinnon and Duncan, 2013), but phytotoxicity in early stages is short-lived as germination improves with time. In a pot experiment, seed emergence of the 11 species was not affected significantly by 1% diesel- or crude oil-contaminated soil (Shahsavari et al., 2013). In contrast to these results, crude oil did not show any impact on seed emergence over diesel-contaminated soils, this is attributed to toxic nature of diesel fuel. Diesel oil is rich in aliphatic hydrocarbons (80-90%) than aromatic hydrocarbons (10-20%), as discussed before, petroleum hydrocarbons with higher concentrations of volatile and aliphatic hydrocarbons are more toxic in nature.
In certain instances, soil contamination with aged and low-levels of petroleum hydrocarbons is favorable for the seed germination, which is termed as ‘hormesis’ (Ma et al., 2010). The actual mechanism behind the hormesis is that the slight level of stress optimizes the physiological process in the plant led to give a head start to developing embryo and often results in higher germination counts. Enhanced crop growth and early flowering are also witnessed by hormesis in oil-contaminated soils. Likewise, low concentrations of hydrocarbons are to be considered as phytohormones, such as auxins. If we look at more details about how low concentrations of petroleum hydrocarbons favors the plant growth, there are several possible reasons, (i) pollutants can kill the microorganisms there by enhancing the plant-available nutrients in soil, (ii) compounds present in the oil can promote the plant growth, (iii) biological nitrogen fixation may be increased with the presence of oil in the soil (Baker, 1971).
Generally, freshly-contaminated soils are more phytotoxic than the aged-soils (soils rested with contamination for some time), which might be due to loss of volatile fractions as well as reduced bioavailability of the pollutant. Bioavailability is defined as the fraction of contaminant that can be readily transformed or assimilated by microorganisms. Once oil spillage occurs, some compounds of oil are adsorbed by organic matter and colloids present in the soil, therefore, small fraction of spillage is bioavailable (Tao et al., 2009). Sometimes, natural weathering of parent compound produces by products, which are more phytotoxic and can induce plant stress. For example, certain photo-induced PAHs (e.g. anthracene, benzo(α)pyrene and fluoranthene) are more toxic and water soluble than the parent compounds, which are also responsible for poor root-water retention.
Another considerable impact of oil spillage in soil is oxidative stress which is occurred when air-filled pore spaces are decreased by hydrocarbons. Oxidative stress causes considerable metabolic changes in the plant tissues, for example, plants produce high concentration of reactive oxygen species (ROS) in their tissues. There are different kinds of ROS, such as superoxide (O−₂), peroxide (H2O2), hydroxyl radicle (•OH), and singlet oxygen (1[O2]), all are able to cause cell damage. Several PAHs and their derivatives are known to bioaccumulate and produce ROS in plant tissues. The extent of plant cell’s response to oxidative stress is usually evaluated based on the concentration of hydrogen peroxide, cellular DNA damage and enzyme activities. Under stressed conditions, hydrogen peroxide is normally stored in peroxisomes and diffuses out due to loss of membrane integrity, and ultimately leads to the damage of cell metabolism (Chen et al., 2013).
Practically it has been observed that increased diesel stress led to increased hydrogen peroxide accumulation. In Arabidopsis thaliana, phenanthrene-induced stress caused cell death in shoots by an increased production of hydrogen peroxide (Alkio et al., 2005). Nevertheless, persistent hydrophobic organic films are witnessed on the surface of seeds growing in soils polluted with petroleum hydrocarbons, such hydrophobic films greatly obstruct the gaseous exchange in the embryo and eventually killing it. On the other hand, oxygen depletion around the seed in polluted soil is also caused by oil-degrading microorganisms, since these microorganisms able to utilize petroleum hydrocarbons as an additional carbon source for their growth and metabolism. Similarly, ability of soil to provide moisture and nutrients to plants is also adversely affected by pollution caused by petroleum hydrocarbons. Since these compounds are in hydrophobic nature, distribution of water in soil layers is adversely affected, which elicits in the uneven distribution of water at oil-polluted sites, subsequently this situation may lead to water deficiency.
As well, petroleum hydrocarbons are found to cause several negative effects on plant, such as shorter roots in seedlings, inhibition of rood fresh weight accumulation, mechanical disruption to the fragile cellular membranes of new roots, diminish the capacity to retain water by roots, disruption of hormonal or metabolic system in the cells, slower expansion of cotyledons, significant reduction in the plant-available forms of nutrients (e.g. phosphorous, potassium nitrate), increased concentrations of certain toxic elements (e.g. manganese and sulfur), negative impacts on plant carbon exchange, decreased respiration and photosynthesis (in mangroves only, but no reports available for terrestrial plants), decreased enzymatic activities (e.g. starch-degrading enzymes, total amylase, and starch phosphorylase) etc. The most potent damage by petroleum hydrocarbons to seeds is penetrative damage, which is caused by the movement of pollutant into the outer fruit structure and internal embryonic tissues. Micropylar end (dicots), the coleorhizal end (monocots) or even injured tissue are the major routes for the penetration of oil into the seed, and the viability of the embryo may be lost in sometimes by these penetrations.
Sand beaches are highly dynamic habitats which are not exempted from the impacts of oil spills. In this section, major oil spills and their impact on invertebrate communities in sand beaches is discussed in depth. According to available literature, abundance and species diversity of invertebrate community has experienced a measurable reduction due to mortality and oil fouling. After Amoco Cadiz oil spill in 1978, Brittany (France), there was an extensive mortality observed in microbenthic community (bivalves, crustaceans, the heart urchin, polychaetes) in low-energy sand beaches. Whereas tolerant polychaetes have been completely decimated. Regarding of meiobenthic community, the abundance and diversity of nematodes has been decreased by 58% (Boucher, 1980).
The devasting effects of oil spill were lasting 2 and 3 years for nematodes and copepods, respectively, though complete recovery of meiobenthic community took 5 years after oil spill. Similarly, macrobenthos community has also been badly affected by oil spill, where the population of crustaceans, Donax, and polychaetes were declined by 89, 89 and 51%, respectively. In the incidence of Prestige oil spill (2002) along the coast of Galicia, Spain, up to 67% of species richness of macrobenthos has been decreased (de la Huz et al., 2005). Certain meiofauna and macrofauna have been depressed, with the impacts lasting over a year after oil spill. There was a complete elimination of rare species including mollusks, whereas two mollusk species (Donax trunculus and Angulus tenuis) and a commonly found nemertean species, Psammamphiporus elongates have been disappeared for 1-2 years after oil spill. High concentrations of PAHs in spill oil has shown significant adverse effects on the abundance of oligochaete.
One of the largest oil spills in the history is Gulf War oil spill (1991), caused massive ecological effects on sand beaches. The impacts include prevention of crab resettlement due to hardening of heavily oiled sediments, lowered macrofauna diversity, reduced habitat availability to fauna, reduced oxygen supply to under sediment layers etc. Especially, in top 60 cm layer, spill oil has caused adverse chronic effects to benthic organisms (Bejarano and Michel, 2010). Transient effects of oil spill on the density of dominant meiofauna taxa (harpacticoid copepods and nematodes) has been observed in 1994 M/V Sea Transporter oil spill. The total abundance and specie density of macrobrenthos (including polychaete, amphipids, mysid, isopd etc) have been declined by 30% and 10%, respectively after 2009 T/V Pacific Adventurer spill.
Likewise, other investigations on oiled beaches have also found decreased harpacticoid and oligochaetes densities over reference sites. In fact, at the spill site, 89 µg of TPH per g soil had shown lowest density of macrobenthos (including microalgae, protozoans, juvenile forms of meiobenthos with < 10 individuals/3.3 cm2). On the other hand, low concentrations of hydrocarbons (600 m) through trophic transfer (Quintana-Rizzo et al., 2015). This section provides detailed information about the impact of aquatic spillage on microorganisms, plants, invertebrates and vertebrates.
In coastal sediments, there was a significant community shift in the benthic microbial eukaryotes upon oil spillage. Diverse assemblages of Metazoa were predominant before oil spill, however, after the spill, they were replaced by diverse communities of hydrocarbon degrading fungal species. Being lipophilic compounds, petroleum hydrocarbons tends to reside in the hydrophobic area which is located in between cytoplasmic membrane monolayer, subsequently it causes fluid fluctuations, changes in the protein structure with decreased enzyme activities in the bacterial cell. By passive transport, spill-oil compounds can interact with the bacterial cells. There are convincing evidence for explaining the impact of petroleum hydrocarbons on at the polluted sites.
For instance, the microbial community composition and diversity was affected greatly by diesel contamination (Sutton et al., 2013), where it was found that the phyla such as Acidobacteria, Actinobacteria, Chloroflexi, Euryarchaeota (Archaea), Firmicutes, and Proteobacteria were significantly higher in abundance, irrespective of soil type. According to a study conducted in Algerian oilfields (Lenchi et al., 2013), it was found that α, β, γ proteobacteria and several archeal classes (e.g. Methanobacteria, Methanomicrobia, Halobacteria, and Thermoprotei) were higher in injection waters than production waters. Importantly, aqueous and oil phases of spilled oil also have significant impact on microbial community. For example, there was a higher number of bacterial diversities of genus Arcobacter in aqueous phase of water-flooded petroleum reservoir, however oil phase has showed higher diversities in Pseudomonas and Sphingomonas (Wang et al., 2014). Likewise, β-proteobacteria was dominant in the hydrocarbon contaminated groundwater samples, with high prevalence of resistance genes/genes products which is a valid evidence for the microbial adaptation to hydrocarbon contamination.
Against stress and adverse conditions provoked by oil-spill, bacterial cells have several protective and adaptive mechanisms, which have been less explored in the literature. Some bacteria possess a cellular metabolic mechanism for utilizing petroleum compounds at the spilled site as their source of carbon and energy. On other hand, certain bacterial species produce thick biofilms around their cell surface, with subsequent alterations in the hydrophobic environment in their cell surface. The biofilm is a strong physical barrier and prevents the diffusion of petrocarbons compounds into the bacterial cell. Certain bacteria show tolerance to high concentrations of aromatic hydrocarbons by excreting petroleum hydrocarbons via. active solvent efflux pumps (Torres et al., 2011). Bacteria cells can also exhibit another interesting phenomenon called ‘chemotaxis’ through the help of chemoreceptors. This phenomenon greatly helps in controlling the spatial position of bacterial cells with response to presence and absence of petroleum hydrocarbons in the surrounding environment.
Future research focusing on identification of chemoreceptors by genomics and proteomics, and how petroleum hydrocarbons induce signal transduction in chemoreceptors will give more insights into understanding the interactions of petroleum compounds with bacterial cell and their subsequent metabolism by bacteria. It is also important to note that the ‘gene-amplification’ is one of the consequences in the bacteria exposed to oil spills, so detection of catabolic genes is helpful in finding novel bacteria with effective catabolic features for petroleum hydrocarbon metabolism. In order to assess the impact of petroleum hydrocarbons on microorganisms exist at the spilled site, commonly used methods are isolation and identification. However, these methods are of greatest limitation that they are unique to cultivable microflora.
To overcome this problem, metagenomics, proteomics, metabolomics approaches are widely used now a days, which provides comprehensive overview on several issues include but not limited to microbial community structure, accountability of functional genes of adaptability, catabolic functions, heavy metal resistance, nutrient metabolism etc. In diverse deep-sea environments of the Gulf of Mexico, there is a rapid response of cold adapted microorganisms to spill oil within hours to days, which is attributed to an outstanding metabolic potential of in situ community of rare microflora. The rare microflora was the active in shifting the microbial community which led to the development of dominant hydrocarbon oxidizers concurrent impact on Gulf’s particulate carbon cycle as well as planktonic food web (Crespo-Medina et al., 2014).
As a part of remediation of spill site, dispersants are usually sprayed on the surface of oil slick which results in the breakdown of oil into small droplets that are readily mix with the water. But laboratory experiments suggest that the dispersants play crucial role in the shaping and altering the microbial response to petroleum hydrocarbons. Dispersants release break-down products from oil are believed to be highly toxic to marine or aquatic ecosystem. Usually dispersed oil is more harmful to marine flora and fauna than untreated oil, which is attribute to the toxicity of break-down products alone, or their combination with oil droplets or dispersant chemical.
Heavy oiled marsh areas showed considerable negative effects on marsh vegetation (Zengel et al., 2015). Especially there is a poor maintenance in their root growth by marsh plants at the spilled sites, which consequently affects the oil biodegradation process in affected marshes. In deepwater horizon oil spill, nearly 109 hectares (1.09 km2) of seagrass beds were destroyed, indicates the sensitivity of seagrass to the petroleum hydrocarbons. Oils are varying in their toxicity according to their chemical nature. Once they penetrate a plant, it may travel through intercellular spaces and vascular system. Petrocarbons can enter the cells by damaging the cell membrane, simultaneously there is a leakage cell contents due to cell membrane damage. By blocking the stomata and intercellular spaces, oils can reduce the transpiration rate in the plants.
There is also disruption of chloroplast membranes by petroleum hydrocarbons concurrent reduction of photosynthesis. Plants exposed to oils may show various symptoms including oil-trapping ability, yellowing and death of oiled leaves, reduction of seedlings etc. If the pollution is chronic type, there may be complete elimination of vegetation. Aquatic plants are key in the functioning of ecosystem due to their unique features such as oxygen production, carbon sequestration and having base position in aquatic food chains. The intertidal and subtidal habitats are rich in plants and are often exposed to spill oils. It has been known for at least 65 years that petroleum hydrocarbons (crude and refined oils) are phytotoxic. Oils can show many sublethal effects on enzyme systems, photosynthesis, respiration, protein and nucleic acid synthesis.
Since the toxicity assessment in aquatic plants at in vivo is extremely difficult, different laboratory-based tests have been introduced to understand more about the toxic effects of oils. In these tests, plants have been exposed to different crude oils, dispersants, which have been applied to different media such as water, substrate, and foliage. When wetland plats were exposed to crude or refined oils through soil or foliar application, the most toxic levels were found to be in the range of 1 L/m2 and 24 L/m2 (Lewis and Pryor, 2013), and lowest toxic levels found for freshwater plant Pistia stratiotes exposed to Urucu crude oil for 14 weeks.
Similarly, the effect levels of Louisiana crude oil on Marsh species (Spartina spp.) of U.S Gulf of Mexico were in between 0.28 L/m2 and 8 L/m2. In saltmarsh plants, recovery has been reported which is attributed to the regeneration of plant tissues from roots and rhizomes, but these recovery times are vary from 6 months to 20 years. The toxicity of dispersants to aquatic plants have also been tested by using in vitro experiments. For example, freshwater Sagittaria lancifolia was not affected by 4000 ppm or less of dispersant namely JD-2000, where principal criteria considered were photosynthetic activity, survival and aboveground biomass. The recoverable dose and LC50 (6w) values of JD-2000 to Sagittaria lancifolia were 16,000 ppm and 20,000 ppm, respectively.
On the other hand, mangroves are considered as most sensitive to petrocarbons than all other aquatic plat systems (e.g. mangrove-associated plants, seagrasses, salt marshes). Once mangroves are affected by oils, it takes nearly 50 years to recover from the adverse effects. At the same time, upon exposure to oil, symptoms may be appeared within days or take several years. According to the environmental reports, 13 of the 70 mangrove species have been affected by oils and dispersants. According to laboratory experiments, it is known that seedlings are more sensitive to oils than juvenile and mature mangroves.
Also, highest toxicity of has been observed when oil was applied to the canopy of trees rather than base oiling, roots and leaves. In mangroves exposed to lubricating oil, stem biomass was less likely to be affected than leaf, root and total biomass. In Rhizophora mangle, stem height was found to be more sensitive after acute exposure to Bonny light oil, however in chronic exposure, leaf length was found to be most sensitive. There was a wide difference in the LC50 values of lab and field level investigations.