55 xenobiotic organic compounds in Tripoli Landfill-Lebanon Leachate and their Fluxes to the Abou Ali River and Mediterranean Sea

doi:10.21203/rs.3.rs-1840336/v1

Pollution generated from landfill solid wastes constitute one of the major threat to the environment. The landfill leachate contains various toxic pollutants, making it the most dangerous issue of the landfills. Monitoring the xenobiotic organic concentrations in landfill leachate is an important step to evaluate the environmental impacts. This work aims to monitor the concentration of organic pollutants such as polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs), phthalic acid esters (PAEs), and bisphenols (BPs) in leachate samples collected from the old and new units of Tripoli landfill-Lebanon. The organic pollutants were extracted by using the solid-phase extraction and quantified by using GC-MS/MS. The results showed high level of PAEs, BPs, PCBs and PAHs in the leachate samples. The fluxes of pollutants to the Abou Ali River and Mediterranean Sea have been detected at 0.23 kg, 0.01 kg, 116.85 kg, 15.93 kg, and 7.58 kg for Σ₁₆PAHs, Σ₂₈PCBs, Σ₆PAEs, Σ₄BPs, and 4-NP respectively.

Landfill

leachate

persistent organic pollutants

xenobiotic organic compounds

PAHs

PCBs

Landfills are the most used method in waste management in developing countries. The landfill waste is exposed to biological and physicochemical transformations leading to highly polluted effluent called leachate. The leachate characteristics are influenced by several factors like waste composition, landfill age, landfill structure, compaction technique, and weather conditions (Naveen et al., 2017). The leachate pollutants are classified into four types: dissolved organic matter, inorganic macro compounds, heavy metals and xenobiotic organic compounds. These last include polychlorobiphenyls (PCB), polycyclic aromatic hydrocarbons (PAHs), bisphenols (BPs), phthalic acid esters (PAEs), pharmaceutical residues and pesticides (Gallen et al., 2017, Kalčíková et al., 2011). PAHs can be generated from anthropogenic and pyrogenic sources. While PCBs have anthropogenic origin; they are used in many industries and commercial applications such as electrical, and heat transfer equipment, dyes, and carbonless copy paper (USEPA, 2019). Municipal dumpsites constitute an important source of PAHs and PCBs into marine environment via leachate (Merhaby et al., 2019). PAHs and PCBs are listed as persistent organic pollutants (POP) in the Stockholm convention (United Nation Enviromental Programme, 2019) and as priority pollutants due to their toxicity, mutagenicity, and carcinogenicity effect (USEPA, 2014). While PAEs and BPs are endocrine disrupting chemicals, and their environmental behavior has attracted considerable attention due to their potential impact on ecosystem and on public health. PAEs present in many materials or products including PVC and building materials, personal-care products, medical devices, children’s toys, food containers and others. While BPs are used in the production of polycarbonates, epoxy resins, thermal papers, cans, and plastics industry, the automotive industry, glasses and others. PAEs and BPs are classified as endocrine disruptors and as priority polluans (USEPA, 2014). They are ubiquitous in the environment including in fresh and marine waters, soil, sediments, manure, compost, effluents, waste dump water, and discharge percolates (Idowu et al., 2019, Net et al., 2015, Sghaier et al., 2017).

Open dumps and uncontrolled landfills are dominant in Lebanon especially in big cities like Beirut and Tripoli (Idowu et al., 2019). The inappropriate strategy of waste in Lebanon is characterized by the absence of waste sorting and recycling policy (Halwani et al., 2020). Due to the dumping of many hazard waste into landfills, such as unused pharmaceutical product, medical waste, personal care product, paints, pesticides, plastics, electronic materials, and nylon, the landfill leachate constitute an important source of organic pollutants like PAHs, PCB, PAEs and phenols (Crawford & Quinn, 2017). The landfills constitute a high risk to the environment due to the various pollutants in leachate. Landfill’s leachate are frequently rejected into the aquatic system without any treatment especially in developing countries. Thus toxic compounds can affect fauna, flora, human health and also aquatic system (Singh et al., 2016, Qi et al., 2018). Several studies have focused on the physicochemical characteristics of landfill leachate (Naveen et al., 2017). However, there is a lack of data on the state of xenobiotic organic compounds release from landfill into leachate and into the aquatic environment. This paper aims to evaluate the level of 55 xenobiotic compound including 16 PAHs, 28 PCBs, 6 PAEs, 4 BPs, and 4-nonyl phenol in Tripoli landfill leachate, North of Lebanon where no previous studies have investigated, and to assess the quality and quantity of the pollutant’s flux to the Abou Ali river and the Mediterranean Sea nearby. This last could affect the marine fauna, and thus human health via the food chain and touristic activities.

2.1 Chemicals and materials

C-18 cartridge (200mg/6ml), PAEs and bisphenols were purchased from Sigma-Aldrich (USA). PAHs standard were purchased from Restek (Bellefonte, USA). Naphthalene-d8, Acenaphtene-d10, Phenanthrene-d10, Pyrene-d10 and Perylene-d12 used as internal standard for PAHs quantification were obtained from LGC-promochem (UK). PCBs standard was purchased from Accustandard Inc. (USA). TCN, CB112 and OCN used as internal standards for PCBs quantification were obtained from Dr. Ehrenstorfer (Augsburg, Germany).

Targeted compounds: Four families of micro-pollutants including 16 PAHs, 28 PCBs, 6 PAEs and 5 phenolic compounds were studied. (i) 16 PAHs: Naphthalene (Nap), Acenaphtylene (Acy), Acenaphtene (Ace), Fluorene (Flu), Phenanthrene (Phen), Anthracene (Ant), Pyrene (Pyr), Fluoranthene (Flt), Benzo[a]anthracene (B(a)A), Chrysen (Cry), Benzo[b]fluoranthene (B(b)F), Benzo[K]fluoranthene (B(k)F), Benzo[a]Pyrene (B(a)P), Benzo[ghi]perylene (B(g,h,i)P), Dibenzo[ah]anthracene (DB(ah)A) and Indeno[1,2,3-cd]pyrene (IP). (ii) 28 PCBs: CB8, CB18, CB28, CB44, CB52, CB66, CB101, CB77, CB81, CB114, CB105, CB123, CB128, CB126, CB138, CB187, CB118, CB167, CB157, CB156, CB169, CB170, CB180, CB153, CB189, CB195, CB206, CB209. (iii) 6 PAEs: Dimethyl phthalate (DMP), diethyl phthalate (DEP), di-n-butyl phthalate (DNBP), butyl benzyl phthalate (BBP), bis-2-ethylhexyl phthalate (DEHP) and di-n-octyl phthalate (DNOP). (iv) 5 BPs: Bisphenol C (BPC), bisphenol E (BPE), bisphenol F (BPF), bisphenol G (BPG), and 4-nonylphenol (4-NP). Chemical formula, qualification and quantification ions, retention time (RT), internal standard for the quantification of each compound and number of chlorinated atom or cycle of PCBs, PAHs, PAEs and BPs are presented in the Table S1 et S2 in supplemental information.

2.2 Study area

Tripoli is the largest city in northern Lebanon with 450000–550000 inhabitants, and is characterized by a Mediterranean climate, cold and wet during the winter. Tripoli landfill, located in Tripoli along the coastline of the Mediterranean Sea and adjacent to the Abou Ali River, covers an area of 60000 m². Tripoli landfile was used as an uncontrolled dumpsite for over 20 years till 1999 when the site was rehabilitated and operated as a controlled landfill receiving the waste from Al Fayhaa Union municipalities, with an average of 14000 tons/month. Tripoli landfill received various types of wastes without sorting process where the organic wastes represents 64% of the total volume of wastes, followed by papers and cardboards (15%), plastics (10%), glass (5%) and metal wastes (2%). Collected wastes are landfilled by spreading and compression in layers of 50 cm, then covering by inert soil. The landfill body features a heterogeneous physical structure, which is constituted of two sections, the lower section (old – unit A), and the upper section (new – unit B). These units present different conditions like height, degree of compaction. Age of waste in Unit A is about 30 years and lower than 10 years in unit B. The flow rate of leachate reach 3 m³/d in unit A and 40 m³/d in unit B. The generated leachate is drained out by a pipes system and periphery ditches and then passes into the estuary of Abou Ali River, which leads to the Mediterranean Sea. The continuous accumulation of rubbish has transformed the site into a mountain of trash with about 45 m levels making Tripoli’s dumpsite one of the highest landfills in Lebanon. The photo 1 presente the sampling points.

2.3 Sample collection

The samples were collected in duplicate from the output leachate pipe of both sections every 2 months from September 2017 to September 2018. Six sampling campaigns were performed in the new section of landfill, and four sampling campaigns were performed in the old section due to the drying of leachate during summer. The drying contribute to the sorting of organic matter from landfilled waste and/or the deviation of leachate flow rate caused by the damage of periphery wall in section A (Halwani et al. 2020). Samples collected in pre-cleaned glass bottles were directly transported to the laboratory. The location of sampling points of leachate are provided in the photo 1. The leachate generated by Tripoli landfill draining directly to the Abou Ali River by a pipes and then reach the Mediterranean Sea.

2.4 Sample extraction

The samples were directly filtered with pre-calcinated 0.47 µm Whatman GMF. The target compounds were extracted by using solid-phase extraction (SPE) with C-18 cartridge. Firstly, each sample was spiked with the internal standards. Briefly, the cartridges were conditioned by 9 ml of acetonitrile, 9 ml of 2-propanol, and then 12 ml of a mixture of Milli-Q/2-propanol (85/15, v/v) acidified to pH 2.5 with sulfuric acid. After conditioning step, 100 ml of leachate samples were passed through the cartridges with a flow rate of 1 drop/second. The cartridges were washed with 30 ml of a mixture of Milli-Q/2-propanol (85/15, v/v) acidified to pH 2.5, then dried with N₂. The targeted compound was eluted with a 12 ml of a mixture of hexane/acetone/2-propanol (90/5/5, v/v/v) followed by 3 ml of dichloromethane. The extract was then concentrated to a volume of 200 µl by N₂ and kept at -18°C until GC-MS analysis.

2.5 GC-MS analysis

Each family of organic compound was analyzed separately, using an Agilent 7890B GC, equipped with ZB-XLB-HT a Zebron capillary column (30 m x 0.25 mm i.d.x0.25 µm film thickness), and coupled with a TQ Agilent series 7000/7010 MS. Helium was used as carrier gas at a flow rate of 1 ml/min. Samples were analyzed in the splitless mode at 280°C. The oven temperature was programmed as follows: from 48°C (1.5 min) to 170°C at 13°C/min, then ramped at 4°C/min to 230°C, and then at 3°C/min (32 min) to 280°C.

2.6. Quantitative analysis and quality control

The concentration of each compound was determined according to a six-point internal calibration methods. Internal standards were added for each calibration point in order to better fit to the properties of each targeted compounds. To minimize the error of quantification, the procedural blank for the entire analytical procedure was performed in triplicate together with each environmental sample batch. Our results showed that the procedural blanks were for PCBs, PAHs, and BPs were free from any targeted PCBs, PAHs, and BPs. However signification level higher than LOQ of PAEs were sometimes detected in procedural blanks, especially for DNBP and DEHP. The concentration of DNBP and DEHP measured in blanks could reach up to 4% of the concentration measured in leachate sample. Consequently, the results of DNBP and DEHP were corrected by subtracting the values found in procedural blanks. The precision and accuracy of the method were confirmed by duplicate analysis of the samples. The precision of the technique, specified by the standard deviation, was found between 1 and 5%.

3.1. PAHs concentration and composition profile

Among the 16 PAHs, 12 of them were detected and their concentrations are presented in Table 1. In unit A, Nap showed the highest level and followed by Acy, Pyr. In unit B, similarly, Nap was detected dominant and it followed by Phen and Flu (Fig. 2). However, B(b)F, B(a)P, DB(ah)A, and B(ghi)P were bellow LOQ in both units. The Σ₁₆PAHs were found to be much higher in leachate from unit B than in unit A. The Σ₁₆PAHs ranged from 0.1 to 5.9 µg.l^− 1, average of 1.8 µg.l^− 1 in unit A and from 0.8 to 29.6 µg.l^− 1 (average of 9.2 µg.l^− 1) in unit B (Table 1).

Table 1

Individual concentration of PAHs in units A and B from Sept-2017 to Sept-2018.
Sampling date	Sep 2017		Dec 2017		Feb 2018		Apr 2018		Jun 2018	Sep 2018
PAHs ng.l^− 1	A	B	A	B	A	B	A	B	B	B
Nap*	260 ± 10	153 ± 7	5751 ± 12	12221 ± 14	8 ± 0.1	5345 ± 11	154 ± 7	1867 ± 11	754 ± 12	2946 ± 12
Acy	357 ± 13	503 ± 14	-	-	9 ± 0.3	358 ± 10	19 ± 2	22 ± 1	845 ± 2	18 ± 1
Ace	-	-	13 ± 1	38 ± 2	5 ± 0.1	172 ± 6	3 ± 0.3	8 ± 1	246 ± 4	50 ± 2
Flu	3 ± 0.1	11 ± 1	5 ± 0.4	67 ± 3	8 ± 0.4	2720 ± 9	21 ± 2	48 ± 3	84 ± 1	203 ± 3
Phen	16 ± 1.6	125 ± 10	-	-	15 ± 2	18843 ± 8	-	-	352 ± 5	1111 ± 10
Ant	1 ± 0.01	3 ± 0.2	59 ± 6	169 ± 6	5 ± 0.1	432 ± 4	24 ± 2	84 ± 4	414 ± 7	370 ± 9
Pyr	-	-	81 ± 7	1488 ± 13	3 ± 0.1	263 ± 2	22 ± 1	108 ± 4	232 ± 4	98 ± 1
Flt	9 ± 0.7	53 ± 4	-	-	-	-	-	-	-	-
B(a)A*	-	-	-	-	11 ± 1	670 ± 6	44 ± 3	527 ± 6	334 ± 6	8 ± 1
Chry*	-	-	5 ± 0.3	12 ± 1	0.1 0.01	429 ± 5	5 ± 0.4	37 ± 1	21 ± 1	10 ± 1
B(b)F*	-	-	-	-	-	-	-	-	-	-
B(k)F*	-	-	10 ± 1	54 ± 4	13 ± 1	40 ± 1	49 ± 3	49 ± 1	80 ± 2	107 ± 2
B(a)P*	-	-	-	-	-	-	-	-	-	-
DB(ah)A*	-	-	-	-	-	-	-	-	-	-
B(g,h,i)P	-	-	-	-	-	-	-	-	-	-
I(cd)P*	-	-	-	101 ± 6	14 ± 1	349 ± 3	34 ± 2	61 ± 3	34 ± 1	121 ± 3
Min	-	-	-	-	-	-	-	-	-	-
Max	357	503	5751	12221	15	18843	154	1867	754	2946
Σ ₁₆ PAHs	647	849	5931	14150	90	29620	375	2810	2636	5041
Σ ₈ Carcinogenic PAHs	261	153	5774	12389	45	6833	286	2542	1224	3192
BaA/BaA + Chry	-	-	-	-	1	1	1	1	1	-
Phe/Ant	14	39	-	-	3	44	-	-	1	3
Ant/Ant + Phe	-	-	1	1	-	-	1	1	-	-
“-” : < LOQ; n = 4 replicate; Human carcinogen compounds IARC (2002).*

The differences between these two units can be attributed to the type, age, degradation state of waste and leachate flow rate. In comparison, our results were much higher than which found by Borjac et al. (2019) and Kalmykova et al. (2013) in south Lebanon (0.19–1.1 µg.l^− 1) and in leachate of Sweden at 0.6 µg.l^− 1. However, our results were much lower than which reported by Jiries et al. (2005) in Jordan where the Σ₁₆PAH ranged at 7.1–12.6 mg.l^− 1 with an average of 9.1 mg.l^− 1. 8 PAHs are classified as probably (B(a)A, B(a)P, DB(ah) A) or possibly carcinogenic (Nap, Chry, B(b)F, B(k)F, I(cd)P) to humans by the (USEPA 1993 and IARC 2002). The highest Σ₈PAHs cancinogenic were determined in unit B at 4.4 µg.l^− 1 of average (0.2–13.8 µg.l^− 1). While the average in unit A were determined at 1.6 µg.l^− 1 and the Fig. 1 shows the composition profile of PAHs.

3.1.1 PAHs source distribution

The PAHs source can be identify by various ratios. Among the ratios, B(a)A/B(a)A + Chry (BaA/228) ratio is commonly used. The ratio < 0.20 indicate petroleum origin and > 0.35 indicate combustion source (Yunker et al. 2002, Christensen & Bzdusek 2005). From Table 1, the B(a)A/(B(a)A + Chry) ratio was calculated. The ratio show the mixt of combustion and petrogenic source of PAHs pollution. Moreover, Phe/Ant ratio can be used to distinguish the petrogenic and combustion sources. When Phe/Ant ratio < 10 indicates pyrolytic sources and > 10 indicates petrogenic sources (Maliszewska-Kordybach et al. 2008). Our results showed the ratio of 0-13.6, where 75% of Phe/Ant ratio < 10 and 25% >10. Thus, the combustion is the dominant source in the unit A. Similarly for unit B, the Phe/Ant ratio were from 0 to 43.6, where 67% <10 and 33% >10. The results indicate the combustion as dominant source of PAHs pollution. Besides, Ant/Ant + Phe ratio < 0.1 indicates a petrogenic source and > 0.1 indicates a pyrogenic source (Bucheli et al. 2004). From the data in Table 1, the Ant/Ant + Phe ratio can be calculated where the values ranged from 0 to 1 for both unit. For the unit A, 25% of Ant/Ant + Phe ratio < 0.1 and 75% >0.1, while in the unit B, 33% < 0.1 and 67% of Ant/Ant + Phe ratio > 0.1. Ant/Ant + Phe ratio confirm the combustion process as the major origin of PAHs in our leachate. The cross plots of BaA/BaA + Ch and Ant/Ant + Phe in the two units indicated that PAHs originated mainly from combustion sources (Fig. 2). Our results show that the PAHs in leachate samples of Tripoli landfill were predominantly pyrogenic sources. These results may be due to the historical state of waste incineration and it is compatible with research established by Borjac et al. (2019) in leachate in south Lebanon (Borjac et al., 2019, Zhang et al., 2013, Halwani et al., 2020). In addition, the atmospheric deposition can be a source of PAHs from the incineration of tire and electric cable near the landfill and the incineration of fossil fuel in the power plants and road traffic (Lebanese Ministry of Environment, 2006).

3.2. PCBs composition profile

Among the 28 PCBs, 15 were detected and the level of Σ₂₈PCBs were in the range of 30–400 ng.l^− 1 with an average of 200 ng.l^− 1 in unit A, and 90-2000 ng.l^− 1 with an average of 500 ng.l^− 1 in unit B. The individual concentration of PCB is presented in the table 2.

Table 2: The concentration of PCBs in units A and B leachate from Sept-2017 to Sept-2018.

“-” : <LOQ; N.A.: Not analysed

Our results are in the same rang of which found by Herbert et al. (2006) in leachate of Portugal landfill (700–2090 ng.l^− 1). However, our results were much lower than which founded by Yusoff et al. (2013) in leachate of Malaysians landfill where 8.9 mg.l^− 1 of Σ₂₈PCBs were found. PCB 28, 52, 101, 118, 138, 153 and 180 were classified as PCB-indicators (PCBi) which are frequently detected in the environment. Also 12 PCBs were classified as PCBs-dioxin-like (PCBs-DL: 77, 81, 105, 114, 118, 123, 126, 156, 157, 167, 169, and 189) known to be toxic to humans and persistent in the environment (Kimbrough et al., 2010, Tanabe & Minh, 2010). The composition profile of PCB in leachate of unit A and unit B is presented in the Fig. 3.

In unit A, the Σ₇PCBi were 4-324 ng.l^− 1 with an average of 126 ng.l^− 1 and represents 45% of the Σ₂₈PCBs. For unit B, the average Σ₇PCBi was 47 ng.l^− 1 (8-148 ng.l^− 1) and represents 12% of Σ₂₈PCBs. The Σ₇PCBi detected in this study are much lower than those founded in the leachate of Portuguese landfills at 713–2098 ng.l^− 1 (Herbert et al., 2006) and Malaysian landfill at 770,000 ng.l^− 1 (Yusoff et al. 2013). While the Σ₁₂PCB-DL was quantified at 18.8 ng.l^− 1 of average (3.7–38 ng.l^− 1) and represents 14% of Σ₂₈PCBs in unit A. In unit B, Σ₁₂PCB-DL was 240 ng.l^− 1 of average (25-1105 ng.l^− 1) and represents 37%. These finding is similar to which found by Herbert et al. (2006) and Yusoff et al. (2013). The Σ₇PCBi and Σ₁₂PCB-DL represent 59% and 50% of Σ₂₈PCBs for unit A and B respectively. The present of PCBs in Tripoli landfill leachate might be related to the dumping of PCB-containing waste like e-waste, ink, capacitors, and transformers, atmospheric deposition from the incineration of the tire and electric cable in the study area, and from the power plants. The PCBs in unit A lower than in unit B can be due to the age and the type of waste.

3.3. PAEs concentration profile

Five PAEs among the six were detected and the result showed in Fig. 4. The individual concentration of PAEs is presented in the Table 3. The Σ₆PAEs varied from 571 to 8031 µg.l^− 1 with an average of 2673 µg.l^− 1 in unit A. Similar value was found in unit B with the Σ₆PAEs of 399–4892 µg.l^− 1 and average of 2891 µg.l^− 1.

Our result is higher than which found in China at 50–62 µg.l^− 1 (He et al., 2009) and in Sweden landfill leachate (1.2–46.8 µg.l^− 1) (Kalmykova et al. 2013, Cecilia & Junestedt 2008). High levels of Σ₆PAEs might be attributed to the absence of sorting of recyclable materials and the landfilling of various hazardous wastes containing PAEs. Indeed, PAEs have been used in a very broad range of applications, and their content can be up to 10 − 60% by weight of final product (Net et al., 2015). PAEs have been used in many materials or products including PVC products, building materials including in paint, adhesive, wall covering, in personal-care products, medical devices, packaging, printing inks and coatings, pharmaceuticals and food products, textiles and others. These products and materials have been frequently used in everyday life of Tripoli inhabitants and then reached the Tripoli landfill as solid waste. It is interesting to note that PAEs are not chemically but only physically bound to the polymeric matrix. Thus, they can be easily released from the waste containing PAEs to the leachate (Net et al. 2015). DEHP was the most detected followed by DEP, DnBP, BBP and DMP with concentration of 1624, 1010, 35, 3, and 1 µg.l^− 1 in unit A, and 2187, 645, 57, 2, and 1 µg.l^− 1 in unit B respectively. DnOP is under LOQ for both units. Our results were similar to previous studied with references herein (He et al., 2009, Kalmykova et al., 2013, Cecilia & Junestedt, 2008).

High level of DEHP could be due to the wide and abundant in material and products used in everyday life. DEHP could be persistence in Tripoli landfill since was detected at high concentration even in the old unit (A) after more than 40 years of waste storage. All types of rubbish including PVC products have been dumped into Tripoli landfill without sorting, which can be the important source of DEHP. Indeed, DEHP is abundantly used as PCV plasticizers (Kalmykova et al., 2013, Net et al., 2015). Also, other types of waste might be the sources of PAEs in Tripoli landfill like fibers, lubricants, cosmetics, and pharmaceutical residue (Peijnenburg, 2008). Figure 4 shows the PAEs composition profile in leachate of unit A and unit B from Sept-2017 to Sept-2018.

Table 3: The concentration of PAEs and BPs in leachate from Sept-2017 to Sept-2018.

3.4 Phenolics and its derivatives composition profile

The results of phenolic compounds are shown in Table 3 and in Fig. 5. The 4-nonylphenol and all BPs, except for BPC, were found in leachate samples in both units. Lower concentrations of the Σ₄BPs were found in unit A at 277–422 µg.l^− 1 (340 µg.l^− 1 of average) versus 359–596 µg.l^− 1 (419 µg.l^− 1) in unit B. The Σ₄BPs are in the same range of which found in the Japanese landfills (0.009–3600 µg.l^− 1, average of 230 µg.l^− 1) and higher than which reported for the leachate in south Lebanon (Borjac et al. 2019, Kurata, et al. 2008). A minor deviation of BPs concentration was observed between the sampling period for both units, which might be due to the seasonal variation (Urase & Miyashita 2003, Asakura, et al. 2004). Difference concentration of BPs was observed between units A and B, which might be due to the stability of BPs with time. The levels of BPs in solid waste landfill leachates depend on the types of waste. The BPs in leachate can be due to the landfilling of many wastes contain BPs like food contact materials (Liao & Kannan 2013, Česen et al. 2016), digital media, electronic equipment, medical instruments, pipes and toys (Chen et al. 2016), thermal paper (Skledar & Mašič 2016), dyes, leather tanning products (Cao et al. 2012).

The 4-nonylphenol were detected at 173 µg.l^− 1 (140–200 µg.l^− 1) in unit A and 188 µg.l^− 1 (155–276 µg.l^− 1) in unit B. The detected values are much higher than which found in Japanese landfills at 0.027-6.4 µg.l^− 1 (Kurata, et al. 2008), and in Sweden 0.1–7.3 µg.l^− 1 (Kalmykova et al. 2013). The high levels of 4-nonylphenol could be explained by the disposal of detergent containers, paints, personal care products, plastics, incombustible, and incineration residues (Kurata, et al. 2008). The PBs in Tripoli landfill leachates are thought to be important pollutants and pose a high risk to the surrounding environment and human.

The annual rainfall and the surface of landfill were determined at 700 mm/year and 60 000 m² respectively. If the total among of rainfall on the landfill surface solubilize pollutants containing in the landfill and then, they throw in the Abou Ali River and Mediterranean Sea, the annual volume of leachate throw out is approximately 42000 m³/year. A rough estimation of the pollutants inputs of the Tripoli landfill to the Abou Ali River and then to the Mediterranean Sea can be estimated by the contaminant flux by multiplying the concentrations measured with the corresponding annual volume of leachate. The annual fluxes of xenobiotic organic pollutants throw out from the Tripoli landfill to the Abou Ali River and Mediterranean Sea were calculated at 0.23 kg, 0.01 kg, 116.85 kg, 15.93 kg, and 7.58 kg for Σ₁₆PAHs, Σ₂₈PCBs, Σ₆PAEs, Σ₄BPs, and 4-NP respectively. These assessment give an image about the pollution effect of Tripoli landfill on the Abou Ali estuary and Mediterranean Sea.

Due to the absence of waste management strategy in Lebanon, environmental pollution generated from uncontrolled landfills become more complicated especially in big cities like Tripoli. The concentration of fifty-five organic pollutants including 16 PAHs, 28 PCBs, 6 PAEs, 4 BPs, and 4-nonylphenol in the leachate of Tripoli landfill have been investigated. The concentration of studied pollutants in leachate collected from old unit is lower than which collected from new unit, which may due to the age and the transformation of waste in each unit. The concentration of carcinogenic POPs were detected the high level in leachate and thus may cause high toxicity. There are no landfills for chemical and hazardous waste in Lebanon, the results of the present study reflect the historical use of chemicals and provide additional insight into the nature of wastes eliminated at Tripoli landfill. The disposable of waste containing POPs in a municipal solid waste landfill harm environmental and human health. It is clear that stopping the dumping of hazardous waste into Tripoli landfill is one of the solution to prevent the damage. The treatment of the leachate become the necessity to reduce the impacts on the surrounding environment and particularly the Abou Ali River and coastal area nearby. That’s summons attention from the national authorities to stop the landfilling of hazardous waste in a municipal solid waste landfill and to stop the drainage of leachate to the water resources and also the citizens to stop any touristic activities and reap seafood in the study area.

Acknowledgements

The authors are grateful for the technical support provided by CPER CLIMBIO project. We acknowledge the financial support from the “Agence Universitaire de la Francophonie (AUF)” of Tripoli, North Libanon, which provided a PhD scholarship for Ahmad Moustafa.

Funding

Ahmad Moustafa is financially supported by a PhD fellowship from the Agence Universitaire de la Francophonie (AUF)” of Tripoli, North Libanon.

Conflicts of interests

The authors declare no conflict of interest. The funders had no role in the writing of the manuscript, or in the decision and use of data.

Availability of data

The authors declare that all relevant data supporting the findings of this study are included in this article.

Author’s contributions

Moustafa A and Net S conducted the experiments and analyzed the data. Hamzeh, Baroudi M and Ouddane B assisted in the experiments and discussed the results. Moustafa A, and Net S wrote the manuscript and drew the graphs. All the co-authors revised and approved the final manuscript.

Asakura, H., Matsuto, T., Tanaka, N. (2004) Behavior of Endocrine-Disrupting Chemicals in Leachate from MSW Landfill Sites in Japan. Waste Management 24(6), 613–622.
Borjac, J., El Jomaa, M., Kawach, R., Youssef, L., Blake, D.A. (2019) Heavy Metals and Organic Compounds Contamination in Leachates Collected from Deir Kanoun Ras El Ain Dump and Its Adjacent Canal in South Lebanon. Heliyon 5(8), e02212.
Cao, G., Lu, J., Wang, G. (2012) Photolysis Kinetics and Influencing Factors of Bisphenol S in Aqueous Solutions. Journal of Environmental Sciences 24(5), 846–851.
Cecilia, B.Ö., Junestedt, C. (2008) Chemical Characterization of Landfill Leachates – 400 Parameters and Compounds. Waste Management 28(10), 1876–1891.
Česen, M., Lambropoulou, D., Laimou-Geraniou, M., Kosjek, T., Blaznik, U., Heath, D., Heath, E. (2016) Determination of Bisphenols and Related Compounds in Honey and Their Migration from Selected Food Contact Materials. Journal of Agricultural and Food Chemistry 64 (46), 8866–8875.
Chen, D., Kannan, K., Tan, H., Zheng, Z., Feng, Y.-L., Wu, Y., Widelka, M. (2016) Bisphenol Analogues Other than BPA: Environmental Occurrence, Human Exposure, and Toxicity-a Review. Environmental Science & Technology 50(11), 5438–5453.
Christensen, E.R., Bzdusek, P.A. (2005) PAHs in Sediments of the Black River and the Ashtabula River, Ohio: Source Apportionment by Factor Analysis. Water Research 39(4): 511–524.
Crawford, C.B., Quinn, B. (2017) The Interactions of Microplastics and Chemical Pollutants. In: Microplastic Pollutants. Elsevier Science, 131–157.
Gallen, C., Drage, D., Eaglesham, G., Grant, S., Bowman, M., Mueller, J.F. (2017) Australia-Wide Assessment of Perfluoroalkyl Substances (PFASs) in Landfill Leachates. Journal of Hazardous Materials 331, 132–141.
Halwani, J., Halwani, B., Amine, H., Mohammad, B.K. (2020) Waste Management in Lebanon-Tripoli Case Study. In Waste Management in MENA Regions. Springer, 223–239.
He, P.-J., Zheng, Z., Zhang, H., Shao, L.-M., Tang, Q.-Y. (2009) PAEs and BPA Removal in Landfill Leachate with Fenton Process and Its Relationship with Leachate DOM Composition. Science of the Total Environment 407(17), 4928–4933.
Herbert, P., Silva, A.L., João, M.J. (2006) Determination of Semi-Volatile Priority Pollutants in Landfill Leachates and Sediments Using Microwave-Assisted Headspace Solid-Phase Microextraction 324–331.
Idowu, I.A., Atherton, W., Hashim, K., Kot, P., Alkhaddar, R., Alo, B.I., Shaw, A. (2019) An Analyses of the Status of Landfill Classification Systems in Developing Countries: Sub Saharan Africa Landfill Experiences. Waste Management 87: 761–771.
International Agency for Research on Cancer (2002) IARC Handbooks of Cancer Prevention. IARC publications.
Jennings, A.A. and Zijian, L. (2015) Residential Surface Soil Guidance Values Applied Worldwide to the Original 2001 Stockholm Convention POP Pesticides. Journal of Environmental Management 160, 16–29.
Jiao, H., Wang, Q., Zhao, N., Jin, B., Zhuang, X., Bai, Z. (2017) Distributions and Sources of Polycyclic Aromatic Hydrocarbons (PAHs) in Soils around a Chemical Plant in Shanxi, China. International Journal of Environmental Research and Public Health 14(10), 1198.
Jiries, A., Rimawi, O., Lintelmann, J., Batarseh, M. (2005) Polycyclic Aromatic Hydrocarbons (PAH) in Top Soil, Leachate and Groundwater from Ruseifa Solid Waste Landfill, Jordan. International Journal of Environment and Pollution 23(2), 179–188.
Kalčíková, G., Vávrová, M., Zagorc-Končan, J., Gotvajn, A.Ž. (2011) Seasonal Variations in Municipal Landfill Leachate Quality. Management of Environmental Quality: An International Journal 22(5), 612–619.
Kalmykova, Y., Björklund, K., Strömvall, A.-M. Blom, L. (2013) Partitioning of Polycyclic Aromatic Hydrocarbons, Alkylphenols, Bisphenol A and Phthalates in Landfill Leachates and Stormwater. Water Research 47(3), 1317–1328.
Kiddee, P., Naidu, R., Wong, M.H., Hearn, L., Muller, J. F. (2014) Field Investigation of the Quality of Fresh and Aged Leachates from Selected Landfills Receiving E-Waste in an Arid Climate. Waste Management 34(11), 2292–2304.
Kjeldsen, P., Barlaz, M.A., Rooker, A.P., Baun, A., Ledin, A., Christensen, T.H. (2002) Present and Long-Term Composition of MSW Landfill Leachate: A Review. Critical Reviews in Environmental Science and Technology 32(4), 297–336.
Kurata, Y., Ono, Y., Ono, Y., (2008) Occurrence of Phenols in Leachates from Municipal Solid Waste Landfill Sites in Japan. Journal of Material Cycles and Waste Management 10(2), 144–152.
Lebanese Ministry of Environment (2006) National Implementation Plans for the Management of Persistent Organic Pollutants. Available at: https://www.informea.org/en/action-plan/republic-lebanon-ministry-environment-national-implementation-plans-management. (accessed 20 November 2019).
Liao, C., and Kannan, K. (2013) Concentrations and Profiles of Bisphenol A and Other Bisphenol Analogues in Foodstuffs from the United States and Their Implications for Human Exposure. Journal of Agricultural and Food Chemistry 61(19), 4655–5662.
Maliszewska-Kordybach, B., Smreczak, B., Klimkowicz-Pawlas, A., Terelak, H. (2008) Monitoring of the Total Content of Polycyclic Aromatic Hydrocarbons (PAHs) in Arable Soils in Poland. Chemosphere 73(8), 1284–1291.
Matsumoto, M., Hirata-Koizumi, M., Ema, M. (2008) Potential Adverse Effects of Phthalic Acid Esters on Human Health: A Review of Recent Studies on Reproduction. Regulatory Toxicology and Pharmacology 50(1), 37–49.
Merhaby, D., Rabodonirina, S., Net, S., Ouddane, B., Halwani, J. (2019) Overview of Sediments Pollution by PAHs and PCBs in Mediterranean Basin: Transport, Fate, Occurrence, and Distribution. Marine Pollution Bulletin 149, 110646.
Naveen, B.P., Mahapatra, D.M., Sitharam, T.G., Sivapullaiah, P.V., Ramachandra, T.V. (2017) Physico-Chemical and Biological Characterization of Urban Municipal Landfill Leachate. Environmental Pollution 220, 1–12.
Net, S., Sempéré, R., Delmont, A., Paluselli, A., Ouddane, B. (2015) Occurrence, Fate, Behavior and Ecotoxicological State of Phthalates in Different Environmental Matrices. Environmental Science & Technology 49(7), 4019–4035.
Peijnenburg, W.J.G.M. (2008) Ecotoxicology/Phthalates. In: Encyclopedia of Ecology. Elsevier, 2733–2738.
Qi, C., Huang, J., Wang, B., Deng, S., Wang, Y., Yu, G. (2018) Contaminants of Emerging Concern in Landfill Leachate in China: A Review. Emerging Contaminants 4(1), 1–10.
Sghaier, R.B., Tlili, I., El Atrache, L.L., Net, S., Ghorbel-Abid, I., Ouddane, B., Hassan-Chehimi, D.B., Trabelsi-Ayadi, M. (2017) A Combination of Factorial Design, off-Line SPE and GC–MS Method for Quantifying Seven Endocrine Disrupting Compounds in Water. IJER 11(5–6), 613–624.
Shukla, S.K., Mangwani, N., Rao, T.S., Das, S. (2014) Biofilm-Mediated Bioremediation of Polycyclic Aromatic Hydrocarbons. In: Microbial Biodegradation and Bioremediation. Elsevier, 203–232.
Singh, S., Raju, N.J., Gossel, W., Wycisk, P. (2016) Assessment of Pollution Potential of Leachate from the Municipal Solid Waste Disposal Site and Its Impact on Groundwater Quality, Varanasi Environs, India. Arabian Journal of Geosciences 9(2), 131.
Skledar, D.G., Mašič, L.P. (2016) Bisphenol A and Its Analogs: Do Their Metabolites Have Endocrine Activity. Environmental Toxicology and Pharmacology 47, 182–199.
Tanabe, S. and Minh, T.B. (2010) Dioxins and Organohalogen Contaminants in the Asia-Pacific Region. Ecotoxicology 19(3), 463–478.
United Nation Enviromental Programme (2019) All POPs Listed in the Stockholm Convention 2019. Accessed on 20 November 2020 at: http://chm.pops.int/TheConvention/ThePOPs/AllPOPs/tabid/2509/Default.aspx.
United States Enviromental Protection Agency (2014) Priority Pollutant List. Available at: https://www.epa.gov/sites/production/files/2015-09/documents/priority-pollutant-list-epa.pdf. accessed (20 November 2020).
Urase, T. and Miyashita, K.-I. (2003) Factors Affecting the Concentration of Bisphenol A in Leachates from Solid Waste Disposal Sites and Its Fate in Treatment Processes. Journal of Material Cycles and Waste Management 5(1), 77–82.
Yusoff, I., Alias, Y., Yusof, M., Aqeel, M. (2013) Assessment of Pollutants Migration at Ampar Tenang Landfill Site, Selangor, Malaysia. ScienceAsia 39, 392–409.
Zhang, Q.Q., Tian, B.H., Zhang, X., Ghulam, A., Fang, C.R., He, R. (2013) Investigation on Characteristics of Leachate and Concentrated Leachate in Three Landfill Leachate Treatment Plants. Waste Management 33(11), 2277–2286.

Photo 1 is available in the Supplementary Files section.

Supplemental Tables S1 and S2 are not available with this version.

No competing interests reported.

Photo1.jpg
Photo 1: Photo of sampling site and sampling points.

55 xenobiotic organic compounds in Tripoli Landfill-Lebanon Leachate and their Fluxes to the Abou Ali River and Mediterranean Sea

Status:

Version 1

Abstract

Figures

1. Introduction

2. Materials And Methods

2.1 Chemicals and materials

2.2 Study area

2.3 Sample collection

2.4 Sample extraction

2.5 GC-MS analysis

2.6. Quantitative analysis and quality control

3. Results And Discussion

3.1. PAHs concentration and composition profile

3.1.1 PAHs source distribution

3.2. PCBs composition profile

3.3. PAEs concentration profile

3.4 Phenolics and its derivatives composition profile

4. Fluxes And Potential Sources To The Abou Ali River And Mediterranean Sea

Conclusion

Declarations

References

Photo1

Supplemental Information

Additional Declarations

Supplementary Files

Status:

Version 1