The physical–biological processes of petroleum hydrocarbons in seawater/sediments after an oil spill

Abstract

The adsorption and desorption behaviors of dissolved petroleum hydrocarbons (DPHs) in a seawater–sediment system were investigated. Tidal flat sediment was used as the adsorbent, and crude oil was used as the adsorbate. The processes of adsorption and desorption at low concentration (<14.3 mg L⁻¹) were described by the first-order kinetics model. The rate of desorption was slower than that of adsorption, and about 49% of the DPHs remained on the sediment. Therefore the potential risk of pollution would exist for a long time. The adsorption isotherms could be better fitted to the linear isotherm model than the Freundlich and Langmuir models. The adsorption process is a physical adsorption, because |ΔH| was 39.0 kJ mol⁻¹ which is less than 42.0 kJ mol⁻¹. The change in n-alkanes in the process was more obvious than the aromatics; the weathering loss rate was 25.56%, the emulsification loss rate of the dispersant was 0.65% and the microbial degradation rate was 15.46%. The results showed the degradation processes of petroleum hydrocarbons in tidal flats.

---

1. Introduction

Petroleum hydrocarbons are a complex mixture of components, which include hydrocarbons (such as alkanes, cycloalkanes and aromatics) and nonhydrocarbon components (such as oxygen containing compounds, sulfur containing compounds, nitrogen containing compounds, resins and asphaltenes).¹ Many of these components have teratogenic, carcinogenic and mutagenic effects.² With the development of the petroleum industry, marine oil spill accidents have happened frequently in recent years. When crude oil is spilled into the ocean, it is subject to several weathering processes, which include evaporation, dissolution, emulsification, oxidation, and biodegradation.^3–5 However, the petroleum hydrocarbon pollution is not only confined to the ocean, but is also extended to the beach due to tides and waves.⁶

Petroleum hydrocarbons have a very devastating effect on marine ecosystems. DPHs would certainly be harmful to the marine environment, especially the near shore and estuarine areas.⁷ The particle size of the sediment is fine. Once petroleum-contaminated seawater touches the tidal flat sediment, DPHs will be removed from the aqueous phase. Although the oil concentration in the seawater is decreased, the pollution is not exactly eliminated. The study of the adsorption and desorption behaviors can provide a theoretical basis for the migration of DPHs in the marine environment.⁸

Many studies have simulated the coastal environment. Nine wave tanks were used to investigate the behavior of chemically dispersed oil and whole oil on a near-shore environment, studying the effects of wave action on the biodegradation of Iranian light crude oil.⁹ Santas used two mesocosm experiments and each experiment was performed in three tanks simulating beach and open-water conditions.¹⁰ Govarthanan conducted oil remediation studies using the SKK11 bacteria.¹¹ Gertler investigated the fate of oil-degrading bacteria for a bioremediation strategy in the simulated coastal environment.¹²

In this research, the adsorption and desorption behaviors of DPHs in the seawater–sediment system were investigated using tidal flat sediment as the adsorbent and DPHs as the adsorbate. The thermodynamic and kinetic laws of DPHs adsorbed on the sediment were studied systematically, and the mechanism of adsorption was explored. Bioremediation agents were employed to deal with coastal weathered spilled oil adsorbed on the rocks of the artificial beach, and the petroleum hydrocarbons of the seawater, intertidal and subtidal zones.

2. Materials and methods

2.1. Crude oil sample

A crude oil sample was obtained from Shengli oilfield (Shandong, China). The properties of the oil sample were as follows: the viscosity was 22.2 mPa s at 50 °C, the freezing point was 23.0 °C and the density was 0.8552 g cm⁻³.

2.2. Seawater and sediment sample

A seawater sample (salinity 32‰ and pH 7.95) was collected from the Shilaoren bathing beach, Qingdao, China (36°5′34.61′′N, 120°28′13.40′′W). The sample was filtered with a 0.45 μm polycarbonate filter¹³ and sterilized using an autoclave. The aim of the pretreatment was to remove large particles and microorganisms. A sediment sample was also obtained from the tidal flats of the Shilaoren bathing beach. The gravel and the remains of organisms were eliminated. The sediment sample was dewatered, and sieved using a mesh (224 μm) before the experiment.¹⁴

2.3. Bacteria and dispersant

Strain S-1 isolated from crude oil contaminated sediment in our previous study was used for the biodegradation experiment.^15,16 Sequence analysis of the 16S rDNA gene, BLAST sequence comparison and phylogenetic analysis confirmed that strain S-1 was affiliated with Bacillus cereus (KF033125).¹⁵ Strain S-1 can catabolize crude oil as a carbon and energy source, and it prefers alkanes to aromatics. In this study it was used for the biodegradation experiment.

The commercial chemical dispersant GM-2 was employed as a surface active agent to increase the concentration of petroleum hydrocarbons in the water (purchased from Qingdao Guangming Environmental Technology Co., Ltd). Its biodegradation rate was BOD5/COD (%) > 30, pH 7–7.5, and it was stored at normal temperature. GM-2 is an approved product by the Maritime Bureau, Ministry of Transportation. The main purpose of GM-2 was to disperse the crude oil into tiny particles in the waves or additional force, which can no longer rise when an oil spill occurs, and to accelerate the process of microbial degradation of crude oil in the water.

2.4. Batch static equilibrium experiments

The seawater/crude oil system was stirred at 100 rpm without phase separation and oil–water emulsions. The concentration of DPHs in the seawater was measured until it was constant, and a petroleum hydrocarbon saturated solution was obtained. The oil in the aqueous phase was extracted with petroleum ether and analyzed by UV spectrophotometry.^17,18 The adsorption and desorption were studied using sediment as the adsorbent and DPHs as the adsorbate by static immersion. The adsorption experiments were conducted in 250 mL conical flasks. 0.3 g of sediment was transferred into a flask before the addition of 100 mL of saturated petroleum hydrocarbon solution. Two control experiments were conducted: the first one (sediment and clean seawater) was to eliminate the influence of dissolved substances and the second (only saturated petroleum hydrocarbon solution) was to remove the influence of evaporation. These conical flasks were placed in a reciprocal shaker at 25 °C and 120 rpm, and were sampled at 10, 20, and 30 min and 1, 2, 4, 6, 12, and 24 h to determine the time-variable concentration of the DPHs. The amount of adsorption Q_A (mg g⁻¹) could be measured by the concentration of the petroleum hydrocarbons.

To investigate the effect of temperature and evaluate the kinetic activation energy and the thermodynamic parameters of the process, some experiments were conducted at 293 K, 298 K and 303 K, and the sediment concentrations in some experiments were 3, 5, and 10 g L⁻¹.

After the adsorption equilibrium was established, the supernatant was replaced with 100 mL of clean seawater. Then the conical flasks with contaminated sediment samples and clean seawater were used in the desorption experiments. The experimental conditions were the same as for the adsorption experiment. They were sampled at 0.5, 1, 2, 6, 12, 24, 48, and 54 h to determine the concentration of DPHs in order to calculate the amount of desorption Q_D (mg g⁻¹).

The oil in the aqueous phase was extracted with petroleum ether and analyzed by UV spectrophotometry. All trials were carried out in triplicate to ensure the accuracy of the experiments.

2.5. Column simulation experiment of crude oil migration

The indoor simulation device consisted of a cylindrical separating funnel filled with sediment collected from different depths. A certain amount of crude oil was set on the surface of the sediment. The tidal flats simulation experiment column was connected to a peristaltic pump, eluting with artificial seawater for a certain time. The concentration of petroleum hydrocarbons was determined from the collected eluent every day. The experimental period was 15 days and it was leached for 12 hours every day.

The oil in the eluent was extracted with petroleum ether, and ultrasonic extraction was used to extract the oil content of the sediment. All the extracts were analyzed using UV spectrophotometry.

2.6. Simulation experiment of oil spill in tidal flats

Two tanks (108 × 42 × 48 cm) were designed to simulate the remediation after an oil spill in tidal flats. The sediment in both tanks was at a depth of 18 cm, and its slope was about 9.46°. About 12.6 L of artificial seawater was poured into the tanks.

In stage I, 50 g of crude oil was dissolved in petroleum ether, then dumped into tanks A and B to form an oil film at the surface of the artificial seawater. A week later the petroleum ether was evaporated completely and the crude oil was weathered. The oscillator was activated and the motor frequency was maintained at 0.4. In stage II, about 15 g (30% of the crude oil) of the GM-2 dispersant was poured into tank A at day 15. In stage III, 380 mL (3% of the artificial seawater) of strain S-1 enrichment culture was added to tank A and 380 mL of artificial seawater was added to tank B at day 29. The temperature ranged from 15 to 20 °C.

The oil in the aqueous phase was extracted with petroleum ether, and ultrasonic extraction was used to extract the oil content in the sediment. All the extracts were analyzed using UV spectrophotometry. At the end of each stage about 1 g of the crude oil sampled from tank A and B was stored. The samples were analyzed by GC-MS using a Shimadzu (Kyoto, Japan) GC-MS-QP2010. Helium was employed as a carrier gas. The details of the experimental conditions and quality control were reported previously by Sun.¹⁹

3. Results and discussion

3.1. Adsorption and desorption kinetics of DPHs

The solubility of the petroleum hydrocarbons was 14.3 mg L⁻¹ (eventually), so we investigated low concentrations (<14.3 mg L⁻¹) of DPHs only. According to the results of the control experiments the influence of DPH evaporation could be neglected, and the amount of petroleum hydrocarbons released from the sediment was also disregarded. Fig. 1a shows the adsorption kinetics curve of the sediment sample. Q_A increased over time until an adsorption equilibrium was established, and settled at a stable value after about 320 min. A hypothesis was proposed that the adsorption rate was proportional to the difference between the amount of equilibrium adsorption and instantaneous adsorption. The adsorption rate could be expressed as:

dQ_A/dT = A(L_A − Q_A) (1)

Fig. 1 Adsorption (a) and desorption (b) kinetics of DPHs on sediment.

In equation (1), Q_A (mg g⁻¹) is the amount of instantaneous adsorption, A (min⁻¹) is the adsorption rate constant, and L_A (mg g⁻¹) is the amount of adsorption at equilibrium.

The integral transform of equation (1) is:

Q_A = L_A(1 − e^−AT) (2)

The experimental data of adsorption were fitted with equation (2). The correlation coefficient was about 0.965, so equation (2) could be satisfied to describe the adsorption process. The amount of equilibrium adsorption was 0.3955 mg g⁻¹ by the fitting curve analysis. Reaching 80% of the adsorption capacity required about 69 min. Therefore, once the DPHs in the seawater touched the sediment, they would soon be adsorbed. Although the DPH concentration in the aqueous phase is decreased, the potential risk of pollution still exists.

Fig. 1b illustrates the desorption kinetics curve of the sediment sample. According to the similar method of adsorption kinetics, the process of desorption was fitted with equation (2). The correlation coefficient was about 0.967. The amount of equilibrium desorption was 0.2011 mg g⁻¹. In view of the different units of the adsorption and desorption rate constants, the desorption rate was slower than the adsorption rate. The DPHs adsorbed by the sediment would be released slowly with continuous scouring with clean seawater. Compared with the adsorption process, the process of desorption would be time-consuming, so the pollution threat to a bathing beach would exist for a long period of time.

The amount of equilibrium adsorption and desorption (0.3955 and 0.2011 mg g⁻¹, respectively) demonstrates that the DPHs adsorbed on the sediment would not be released completely. About 49% of the total adsorbed amount would remain on the sediment. Thus we should not forget the pollution adsorbed on sediment after an oil spill or the pollution will be consistently present.

3.2. Adsorption isotherms of DPHs

Fig. 2a and b illustrate the relationship between C_e and Q_e. The standard deviation (SD) of the triplicate sample was not more than 2%. In this study three adsorption models were used to fit the experimental data. The first model was the Langmuir isotherm model Q_e = Q_mC_e/(C_e + A) (where Q_m is the saturated adsorption capacity, and A is a constant associated with the adsorption energy), the second was the Freundlich isotherm model Q_e = K_FC_e^1/n (where K_F, and 1/n are empirical coefficients for each adsorbent pair at a given temperature), and the last was the linear isotherm model Q_e = K_PC_e (where K_P is the linear distribution coefficient or Henry adsorption constant).^20,21 The adsorption isotherm parameters and the correlation coefficients are summarized in Table 1.

Fig. 2 Adsorption isotherms of DPHs at different sediment concentrations (a) and different temperatures (b).

Table 1 Adsorption isotherm models of DPHs on sediment

Effect factors	Parameters	Langmuir isotherm model			Freundlich isotherm model			Linear isotherm model
		Q^0 (mg L^−1)	A (mg L^−1)	R^2	KF (mg^1−n g^−1 L^n)	n^−1	R^2	KP (L g^−1)	R^2
Sediment content (g L^−1)	3	−1.58	−11.8	0.983	0.145	1.13	0.976	161	0.993
	5	−0.886	−11.5	0.975	0.085	1.08	0.926	93.0	0.992
	10	−0.571	−13.6	0.974	0.046	1.09	0.967	48.9	0.992
Temperature (°C)	20	−2.40	−14.4	0.982	0.179	1.08	0.968	190	0.994
	25	−1.58	−11.8	0.983	0.145	1.13	0.976	161	0.993
	30	−3.27	−31.7	0.987	0.107	1.05	0.975	112	0.996

---

In Table 1 the correlation coefficients of the linear, Langmuir and Freundlich isotherm models were in descending order. The adsorption of DPHs on sediment at low concentration fitted the linear isotherm model better than the Freundlich isotherm model. The Langmuir isotherm model had been traditionally used to quantify and contrast the performance of different bio-sorbents.²² The values of saturated adsorption capacity fitted by the Langmuir isotherm model were negative. They were not consistent with the experimental results, so the adsorption of DPHs on sediment did not belong in this category.

The Freundlich isotherm model was used to describe non-uniform and multiple substance adsorption which was not restricted to the formation of a monolayer.²³ The value of 1/n is a measure of adsorption intensity or surface heterogeneity, ranging between 0 and 1. The adsorption intensity or surface heterogeneity will become more heterogeneous as its value gets closer to zero.^24,25 But the values of 1/n in our paper were slightly larger than 1 according to the curve fitting analysis. This means that the adsorption was closer to a linear adsorption.

The correlation coefficients fitted by the linear isotherm model were more than 0.99 at low concentration. The linear adsorption isotherm indicated that the attractive force between the adsorbate and the adsorbent was invariant, and the distribution of adsorbate was irrelevant to its initial concentration. This is usually applied to a situation where the surface coverage is far from saturation. In this study very fine sediment was used which had many adsorption sites, and the DPH concentration in seawater was pretty low. Therefore the adsorption of DPHs on the sediment can be described by the linear adsorption model.

3.3. Effect factors and thermodynamic analysis of adsorption

The linear model used to interpret the adsorption equilibrium has been widely accepted due to its lack of mathematical complexity.²⁶ Table 1 shows that the values of K_P (L g⁻¹) were 161, 93.6 and 48.9 when the sediment concentrations were 3, 5 and 10 g L⁻¹, respectively. K_P decreased with increasing sediment content. This means that 3 g L⁻¹ sediment was able to provide sufficient adsorption sites, and a higher sediment content did not contribute to the adsorption process.

At the same time, the values of K_P (L g⁻¹) were 190, 161 and 112 when the temperature was 20, 25 and 30 °C, respectively. K_P decreased with increasing temperature. This meant that the petroleum hydrocarbons were more likely to exist in the aqueous phase rather than being adsorbed by sediment with an increase of temperature. Therefore, if marine oil spill accidents happened in the summer, the degree of beach pollution would be lower. For example, temperature inversely affects crude oil uptake due to the exothermic nature of the sorption.²⁷

The adsorption isotherms at different temperatures were conducted for the thermodynamic analysis. The van’t Hoff equation^28,29 is as follows:

dln K_P/dT = ΔH/RT² (3)

In equation (3), K_P is the distribution coefficient, ΔH is the heat of adsorption, T stands for temperature, and ΔH is thought to be constant in a narrow temperature range. The integral transform of equation (3) was taken for linear regression analysis:

ln K_P = −A/T + B (4)

In equation (4), A = ΔH/R. According to the above method, the ΔH of DPHs on sediment could be obtained from the values of K_P and T in Table 1. ΔH = −39.0 kJ mol⁻¹. The ΔH shows that the adsorption process was exothermic, and increasing temperature was not conducive to the adsorption. The adsorption can be attributed to physical adsorption according to |ΔH| < 42 kJ mol⁻¹.³⁰

3.4. Crude oil migration result

Fig. 3a shows the concentration changes of the hydrocarbon component in the leaching liquid, which rose sharply at first. Because the water constantly scoured the sediment, at the beginning the soluble components in the crude oil spread to the sediment, where they were partly adsorbed onto the sediment and partly removed along with the water; the concentration in the leaching liquid reached 142.1 mg L⁻¹ at day 5. The concentration of oil in the leaching liquid constantly decreased from the sixth day, and it was only 16.8 mg L⁻¹ at the end of the experiment. This is because as time goes on, the diminishing soluble components in the crude oil leads to the oil concentration in the leaching liquid medium declining. At the same time, the continuous leaching made the adsorbed hydrocarbon component start to dissolve, and eventually a balance of adsorption and desorption was reached.

Fig. 3 Changes in oil concentration in effluent seawater (a) and vertical distribution of crude oil in the sediment (b).

The Fig. 3b shows that the hydrocarbon component of the oil content at 3 cm can be up to 500 mg g⁻¹ after water leaching. Because the sediment blocks the infiltration of the hydrocarbon component,³¹ the adsorbed hydrocarbon component decreased gradually with increasing sediment depth; the hydrocarbon component of the oil content was 18.69 mg g⁻¹ at 9 cm deep, and it gradually stabilized. Just as the finding that the spilled hydrocarbons had remained predominantly in the organic surface horizons of the soil where the spillage occurred,³² thus the concentration of the hydrocarbon component of oil on the beach sediment in the vertical direction was reduced constantly and the oil pollution was mainly concentrated in the surface layer of beach sediment in the marine offshore after the oil spill. This is same as with Bejarano’s investigation that tidal flats had more high-risk samples at the surface than at the bottom.⁶ Therefore the remediation process of surface sea sand within 10 cm should be strengthened in oil spill pollution areas.

3.5. The oil content changes in seawater

The changes in petroleum hydrocarbon content in the seawater phase are shown in Fig. 4.

Fig. 4 Oil concentration changes in the aqueous phase.

In stage I, the first 14 days, the petroleum hydrocarbon content in the seawater phase of tanks A and B presented an increasing tendency. The maximum DPH concentration was 3.107 mg L⁻¹. In this stage the crude oil was dissolved under the external force, but the function of the oscillator was weaker than the stirrer, and the concentration was lower than in Section 2.4.

In stage II, from day 15 to 28, the oil concentration in tank A at 15 d reached the highest value of 4.875 mg L⁻¹, which was mainly due to the scatter and emulsification of dispersant GM-2 which increased the solubility of the petroleum hydrocarbons. However, the emulsified crude oil–seawater system cannot be stable long-term. The increase of hydrocarbons can be due to the presence of oil droplets in the emulsion or the increased dissolution of hydrocarbons from the surfaces of the numerous droplets (surface area effects).³³ This was consistent with Zhao’s study in which the dispersant was much more effective in solubilizing more hydrophobic PAHs.³⁴

In stage III, from day 29 to 48, the oil concentration in tank A decreased while in tank B it increased. At 48 d, the oil concentration in tank A had reduced to 2.28 mg L⁻¹, which means that strain S-1 catabolized the crude oil as a carbon and energy source so it reduced the DPH content.^35,36 The degradation removal percentage of the DPHs was about 33.72%, including physical and chemical degradation such as weathering and catabolizing.

3.6. The oil content changes in sediment

The petroleum hydrocarbon content in the sedimentary phase decreased with depth. The oil content in the surface sediment at 0–1.7 cm was greater than in the sediment at 1.7–3.3 cm and 3.3–5.0 cm. This result was similar to the migration result in Section 3.4; crude oil entered the surface sediment through many paths, and then entered the deep layers as the result of hydrodynamics and infiltration, but the surface had a higher concentration than the other depths at all times. The DPHs and suspended oil were in direct contact with the interface of seawater and sediment, leading to the DPHs being consistently adsorbed onto the sediment.

Fig. 5a shows that the oil content of intertidal sediments in the experimental group presented a downward trend after it rose; the change in the diagram is the same as in Fig. 4. In the presence of dispersant the strong sorption of the dispersant on sediments and the weak solubilization of phenanthrene enhanced its uptake.³⁷ Similarly the dispersant can enhance the sorption of crude oil. In this study because the emulsification of GM-2 caused the DPHs to increase, the crude oil in the sediment peaked at 21 d, which is consistent with the dispersant enhancing the sorption extent.³⁵ Dispersants make crude oil content decrease, but at the same time, they may increase oil pollution in tidal flats. Fig. 5b shows that the oil content of intertidal sediments in the control group also fell after rising, but the change lagged in the experimental group. This means that the dispersant can accelerate the dissolution of crude oil but cannot change its properties. Here the oil content of the surface sediment achieved its maximum.

Fig. 5 Oil concentration changes of the intertidal (a and b) and supratidal (c and d) sediments in the experimental (a and c) tanks and in the control (b and d) tanks.

Fig. 5c shows that the oil content of the supratidal zone (0–1.7 cm) increases gradually as time goes on in the experimental group. Because fine sediment particles accumulated at the supratidal zone due to scouring by the waves, the small sized particles can easily adsorb the oil,³⁸ so petroleum hydrocarbons accumulated on the sediment. At 28 d, the oil content of the sediment at 1.7–3.3 cm is slightly greater than at 0–1.7 cm. This may be due to the lateral migration of petroleum hydrocarbons at the interface between the high water line and the supratidal zone. As shown in Fig. 5d, in tank B the oil content of the supratidal zone presented a rising trend, and also reached a maximum at 33 d.

3.7. Hydrocarbon components change in seawater

Fig. 6 shows the change in hydrocarbon components in the seawater; Tables 2 and 3 show the loss rate of n-alkanes and aromatics. We can see that the loss of hydrocarbon components C₁₀−C₂₆ was bigger than other n-alkanes at 14 d in Fig. 6a, which indicated that natural weathering has little effect on heavy hydrocarbon components after C₂₆ because long chain alkanes have a complex structure and are difficult to evaporate but the short chain alkanes evaporate easily. Tables 2 and 3 show that the natural weathering total loss rate of n-alkanes was 26.56%. We can speculate that the loss of n-alkanes was not only because of their low boiling point but also because of their large solubility that caused n-alkane migration into the surrounding environment. Fig. 6b describes several typical aromatics content changes. Naphthalene was strongly influenced by natural weathering; the loss rate of naphthalene was the biggest and reached 71.89% but the loss rate of the other aromatics was low. This matches the finding that removal was more effective for short-chain alkanes and low ring-number PAHs.³⁹ All in all the n-alkanes were removed more easily than aromatic hydrocarbons, and the light components of the hydrocarbons were greatly influenced by weathering. The initial crude oil diagnosis ratio parameter of pristane/phytane (Pr/Ph) was 1.35 (>1); the data shows that the environment of the crude oil was a brackish water oxidation environment, and the crude oil belongs to the mature crude oil category.

Fig. 6 Concentration changes of n-alkanes (a, c, and e) and aromatics (b, d, and f) at 14 d (a and b), 28 d (c and d) and 48 d (e and f).

Table 2 The related parameters of petroleum hydrocarbons

Number	group	n-Alkanes (μg g^−1)	Aromatic (μg g^−1)	Biomaker (μg g^−1)	Pr/Ph
1	0 d experimental group	501 793	53 510	39 376	1.35
2	14 d experimental group	368 512	106 602	42 587	1.32
3	28 d experimental group	307 481	68 007	33 053	1.33
4	28 d control group	309 890	42 970	42 121	1.29
5	48 d experimental group	247 835	59 103	48 790	1.12
6	48 d control group	297 792	85 611	34 448	1.29

---

Table 3 The loss ratios of petroleum hydrocarbons in the inoculated tank^a

Test stage	n-Alkanes		Aromatic
	Total loss rate (%)	Net loss rate (%)	Total loss rate (%)	Net loss rate (%)
a Total loss rate is the hydrocarbon component total loss rate in the experimental group; the net loss rate deducts the weathering loss in the control group; “—” represent negative value.
I	26.56	26.56	—	—
II	16.56	0.65	36.20	—
III	19.40	15.46	13.09	—

---

Fig. 6c shows that the C₂₂–C₃₇ components of n-alkanes in tank A experienced greater degradation than those of tank B, but other hydrocarbons showed the opposite phenomenon. This showed that GM-2 had an impact on high molecular weight hydrocarbons; it can dissolve more C₂₂–C₃₇ components of n-alkanes in the water and accelerate their degradation. Fig. 4 also shows an increasing trend for stage II in tank A. Combined with Fig. 4 we know that the dispersant lowered the crude oil–seawater interfacial tension and oil slicks were broken up into fine droplets and dispersed into the water column, thereby the degradation was higher than in tank B.⁴⁰ We can see from Table 3 that excluding the natural weathering loss rate in tank B, the loss rate of n-alkanes was 0.65% in tank A. In this stage GM-2 was immediately directionally adsorbed on the oil–water interface; the hydrophilic group was dispersed into the water and the hydrophobic group dispersed into the oil, which made the oil film emulsify into small droplets and turn into the water under the action of an external force. So the crude oil was degraded more easily in tank A than in tank B. Sauret also investigated the changes of hydrocarbon composition in crude oil. Our study was the same, with the investigation into adding emulsifier to result in a more rapid and more efficient removal of hydrocarbons.⁴¹ Fig. 6d shows that the dispersants had little effect on the aromatic hydrocarbon component of oil. This is because on the one hand the structure of the aromatics was more complex and stable than the n-alkanes, but on the other hand the aromatics were more difficult to dissolve than the n-alkanes.

Fig. 6e shows that the oil content of C₁₁–C₂₆ in tank A was smaller than tank B, which suggested that short chain alkanes were more likely to be used by strain S-1 than long chain alkanes. This illustrated that strain S-1 has a special hydrocarbon oxidase which can degrade the oil content of C₁₁–C₂₆ whereas the long chain alkanes were difficult to degrade because of their rigid structure. The total biodegradation rate of n-alkanes reached 15.46% and the biodegradation rate of C₁₁–C₁₄ could reach 81.24%–100% when excluding the influence of weathering. This was due to the effect of the dispersant; when used in the right amount and circumstances, the application of chemical dispersants may promote the biodegradation of oil in the water column and eliminate or reduce the impact of oil on sensitive shorelines and habitats significantly.⁴² We used the dispersant before adding strain S-1 and the dispersant made the crude oil dissolve in the seawater so the strain S-1 could degrade it easily and the biodegradation rate of the alkanes was high. We can see from Fig. 6f that strain S-1 had no obvious effect on typical PAHs. This is mainly because the selected strain S-1 is a degradation bacterium for alkanes, but its utilization for aromatic hydrocarbons is poor.

4. Conclusion

The thermodynamic parameters and adsorption kinetics of DPHs at low concentration were studied systematically. At low concentration the adsorption and desorption processes were fitted to the first-order kinetics model, the linear isotherm model could be used to describe the adsorption isotherms well, and the adsorption belonged to the physical adsorption type according to the thermodynamic analysis. Tidal flats remediation should focus on the sediment surface at a depth of 0–10 cm. The DPH concentration in seawater was about 4.875 mg L⁻¹. The oil content of the intertidal zone is higher than the supratidal zone. The total loss ratio of n-alkanes in the three stages was 25.56%, 16.56% and 19.40%, the loss rate caused by the dispersant was 0.65%, the microbial degradation rate was 15.46%, and crude oil was more sensitive to natural weathering and microbial degradation than the dispersant after an oil spill. Future research should focus on the effect of other chemicals on adsorption such as surface active agents and microbial agents, because these chemicals may affect the concentration of crude oil, potentially resulting in a different law of adsorption and desorption.

Acknowledgements

This research was founded by grants from “The National Natural Science Foundation of China” (41376084); the “Program for New Century Excellent Talents in University” (NCET-11-0464); the “Program for Innovative Research Team in University” (IRT1289); “the Open Foundation of Key Laboratory of Marine Spill Oil Identification and Damage Assessment Technology of SOA” (201402). This is MCTL Contribution No. 98.

References

1. R. B. Long and J. G. Speight, Oil Gas Sci. Technol., 1989, 44(2), 205–217.
2. G. Karavalakis, G. Fontaras, D. Ampatzoglou, M. Kousoulidou, S. Stournas, Z. Samaras and E. Bakeas, Environ. Pollut., 2010, 158(5), 1584–1594.
3. P. J. Brandvik and L. G. Faksness, Cold Reg. Sci. Technol., 2009, 55(1), 160–166.
4. P. F. Kingston, Spill Sci. Technol. Bull., 2002, 7(1), 53–61.
5. J. M. Neff, S. Ostazeski, W. Gardiner and I. Stejskal, Environ. Toxicol. Chem., 2000, 19(7), 1809–1821.
6. A. C. Bejarano and J. Michel, Environ. Pollut., 2010, 158(5), 1561–1569.
7. L. Shen and R. Jaffé, Mar. Environ. Res., 2000, 49(3), 217–231.
8. A. P. Rowland, D. K. Lindley, G. H. Hall, M. J. Rossall, D. R. Wilson, D. G. Benham, A. F. Daniels and R. E. Daniels, Environ. Pollut., 2000, 109(1), 109–118.
9. C. A. Page, J. S. Bonner, P. L. Sumner and R. L. Autenrieth, Mar. Chem., 2000, 70(1), 79–87.
10. R. Santas, A. Korda, A. Tenente, K. Buchholz and P. H. Santas, Mar. Pollut. Bull., 1999, 38(1), 44–48.
11. M. Govarthanan, S. H. Park, Y. J. Park, H. Myung, R. R. Krishnamurthy, S. H. Lee, L. Nanh, K. K. Seralathan and B. T. Oh, RSC Adv., 2015, 5(67), 54564–54570.
12. C. Gertler, G. Gerdts, K. N. Timmis and P. N. Golyshin, FEMS Microbiol. Ecol., 2009, 69(2), 288–300.
13. A. Kapoor, T. Viraraghavan and D. R. Cullimore, Bioresour. Technol., 1999, 70(1), 95–104.
14. J. A. Ferretti, D. F. Calesso and T. R. Hermon, Environ. Toxicol. Chem., 2000, 19(8), 1935–1941.
15. Y. R. Pi, N. N. Xu, M. T. Bao, Y. M. Li and D. Lv, Environ. Sci.: Processes Impacts, 2015, 17, 877–885.
16. H. Y. Gong, Y. M. Li, M. T. Bao, D. Lv and Z. N. Wang, RSC Adv., 2015, 5, 37640–37647.
17. X. Y. Cao, G. P. Yang, S. W. Wei and H. Han, Mar. Pollut. Bull., 2011, 62, 741–746.
18. SBQTS (State Bureau of Quality and Technical Supervision of the People’s Republic of China), The Specification for Marine Monitoring, Part 4: Seawater analysis (Chinese GB 17378.4–1998), Beijing, China, 1998.
19. P. Y. Sun, M. T. Bao, G. Li, X. Wang, Y. Zhao, Q. Zhou and L. Cao, J. Chromatogr. A, 2009, 1216(5), 830–836.
20. H. Yu, G. H. Huang, C. J. An and J. Wei, J. Hazard. Mater., 2011, 190(1), 883–890.
21. X. T. Guo, J. H. Ge, C. Yang, R. R. Wu, Z. Dang and S. M. Liu, RSC Adv., 2015, 5(72), 58865–58872.
22. K. Y. Foo and B. H. Hameed, Chem. Eng. J., 2010, 156(1), 2–10.
23. K. V. Kumar, M. M. Castro, M. Martinez-Escandell, M. Molina-Sabio and F. Rodriguez-Reinoso, J. Phys. Chem. C, 2010, 114(32), 13759–13765.
24. F. Haghseresht and G. Q. Lu, Energy Fuels, 1998, 12(6), 1100–1107.
25. F. Qin, B. Wen, X. Q. Shan, Y. N. Xie, T. Liu, S. Z. Zhang and S. U. Khan, Environ. Pollut., 2006, 144, 669–680.
26. T. C. Voice and W. J. Weber, Water Res., 1983, 17(10), 1433–1441.
27. L. Sørensen, A. G. Melbye and A. M. Booth, Mar. Pollut. Bull., 2014, 78(1), 146–152.
28. W. Y. Chen, H. M. Huang, C. C. Lin, F. Y. Lin and Y. C. Chan, Langmuir, 2003, 19(22), 9395–9403.
29. V. C. Srivastava, M. M. Swamy, I. D. Mall, B. Prasad and I. M. Mishra, Colloids Surf., A, 2006, 272(1), 89–104.
30. F. Pan, J. Z. Ma, Y. Q. Wang, Y. L. Zhang, L. H. Chen and W. M. Edmunds, Environ. Monit. Assess., 2013, 185(10), 8023–8034.
31. Y. Qin, Y. Xiong, C. L. Kang, J. R. Wang, H. F. Liu, C. He and X. Y. Lin, Adv. Mater. Res., 2012, 383, 3939–3943.
32. S. Cram, C. Siebe, R. Ortiz-Salinas and A. Herre, Soil Sediment Contam., 2004, 13(5), 439–458.
33. S. D. Ramachandran, P. V. Hodson, C. W. Khan and K. Lee, Ecotoxicol. Environ. Saf., 2004, 59(3), 300–308.
34. X. Zhao, Y. Y. Gong, S. E. O’Reilly and D. Y. Zhao, Mar. Pollut. Bull., 2015, 92(1), 160–169.
35. B. Khazra, S. M. Mousavi, S. Mehrabi, M. Hashemi and S. A. Shojaosadati, RSC Adv., 2015, 5(42), 33414–33422.
36. Q. X. Lin, I. A. Mendelssohn, K. Carney, S. M. Miles, N. P. Bryner and W. D. Walton, Environ. Sci. Technol., 2005, 39(6), 1855–1860.
37. L. Ferreira, E. Rosales, M. Á. Sanromán and M. M. Pazos, RSC Adv., 2015, 5(46), 36665–36672.
38. Y. Y. Gong, X. Zhao, S. E. O’Reilly, T. Qian and D. Zhao, Environ. Pollut., 2014, 185, 240–249.
39. X. Y. Cao, H. Han, G. P. Yang, H. B. Ding and H. H. Zhang, Water, Air, Soil Pollut., 2013, 224(7), 1–8.
40. National Research Council, http://www.nap.edu/openbook.php?record_id=11283, accessed July 2005.
41. C. Sauret, U. Christaki, P. Moutsaki, I. Hatzianestis, A. Gogou and J. F. Ghiglione, Mar. Environ. Res., 2012, 79, 70–78.
42. R. R. Lessard and G. de Marco, Spill Sci. Technol. Bull., 2000, 6(1), 59–68.

---