55 xenobiotic organic compounds in Tripoli landfill-Lebanon leachate and their fluxes to the Abou Ali River and Mediterranean Sea

Abstract

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 characterize the seasonal variation of 55 xenobiotic organic compounds including polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs), phthalic acid esters (PAEs) and bisphenols (BPs) in the leachate from municipal solid waste landfill of Tripoli, Lebanon. And also, the quantity of the pollutant’s flux to the Abou Ali River and the Mediterranean Sea nearby has been estimated. 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.

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Environ Monit Assess (2022) 194:856
https://doi.org/10.1007/s10661-022-10522-w
55 xenobiotic organic compounds inTripoli
landfill‑Lebanon leachate andtheir fluxes totheAbou Ali
River andMediterranean Sea
AhmadMoustafa· MariamHamzeh·
MoomenBaroudi· BaghdadOuddane ·
SopheakNet
Received: 8 July 2022 / Accepted: 23 September 2022
© The Author(s), under exclusive licence to Springer Nature Switzerland AG 2022
of pollutants to the Abou Ali River and Mediter-
ranean 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.
Keywords Landfill· Leachate· Persistent organic
pollutants· Xenobiotic organic compounds· PAHs·
PCBs
Introduction
Landfills are the most used method in waste manage-
ment in developing countries. The landfill waste is
exposed to biological and physicochemical transfor-
mations leading to highly polluted effluent called lea-
chate. The leachate characteristics are influenced by
several factors like waste composition, landfill age,
landfill structure, compaction technique, and weather
Abstract Pollution generated from landfill solid
wastes constitute one of the major threat to the envi-
ronment. The landfill leachate contains various toxic
pollutants, making it the most dangerous issue of the
landfills. Monitoring the xenobiotic organic concen-
trations in landfill leachate is an important step to
evaluate the environmental impacts. This work aims
to characterize the seasonal variation of 55 xenobiotic
organic compounds including polycyclic aromatic
hydrocarbons (PAHs), polychlorinated biphenyls
(PCBs), phthalic acid esters (PAEs) and bisphenols
(BPs) in the leachate from municipal solid waste
landfill of Tripoli, Lebanon. And also, the quantity
of the pollutant’s flux to the Abou Ali River and the
Mediterranean Sea nearby has been estimated. 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
Supplementary Information The online version
contains supplementary material available at https:// doi.
org/10.1007/s10661-022-10522-w.
A.Moustafa· B.Ouddane· S.Net(*)
CNRS, LASIRE UMR 8516, Equipe Physico-chimie de
l’Environnement, Univ. Lille, F-59000Lille, France
e-mail: sopheak.net@univ-lille.fr
A. Moustafa
e-mail: Ah.MOUSTAFA@outlook.com
B. Ouddane
e-mail: baghdad.ouddane@univ-lille.fr
A.Moustafa
Department ofEnvironmental Biotechnology,
Biotechnology Laboratory, Doctoral School-AZM Center
forResearch inBiotechnology andIts Application,
Lebanese University, Tripoli, Lebanon
A.Moustafa· M.Hamzeh· M.Baroudi
Department ofHealth andEnvironment, Laboratory
ofSciences andWater Environment, Faculty ofPublic
Health Section III, Lebanese University, Tripoli, Lebanon
e-mail: mariamhamze85@hotmail.com
M. Baroudi
e-mail: mbaroudi@ul.edu.lb

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conditions (a. The leachate pollutants are classified
into four types: dissolved organic matter, inorganic
macrocompounds, heavy metals, and xenobiotic
organic compounds. These last include polychloro-
biphenyls (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 gener-
ated 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 car-
bonless copy paper (USEPA, 2014). Municipal dump-
sites 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 prior-
ity pollutants due to their toxicity, mutagenicity, and
carcinogenicity effect (USEPA, 2014), while PAEs and
BPs are endocrine disrupting chemicals, and their envi-
ronmental behavior has attracted considerable attention
due to their potential impact on ecosystem and on pub-
lic health. PAEs are present in many materials or prod-
ucts including PVC and building materials, personal
care products, medical devices, children’s toys, and
food containers (Net etal., 2015b), while BPs are used
in the production of polycarbonates, epoxy resins, ther-
mal papers, cans, and plastics industry, the automotive
industry, glasses, and others. PAEs and BPs are classi-
fied as endocrine disruptors and as priority pollutants
(USEPA, 2014). PAEs and BPs 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., 2015a,b; Paluselli etal., 2018a,b; Ben Sghaier
etal., 2017a,b; Xing etal., 2022).
Open dumps and uncontrolled landfills are domi-
nant 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, med-
ical waste, personal care product, paints, pesticides,
plastics, electronic materials, and nylon, the landfill
leachate constitutes an important source of organic pol-
lutants like PAHs, PCB, PAEs, and phenols (Crawford
& Quinn, 2017). The landfills constitute a high risk to
the environment due to the various pollutants in lea-
chate. Landfill’s leachate is 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 sys-
tem (Qi etal., 2018; Singh etal., 2016). 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 com-
pounds release from landfill into leachate and into the
aquatic environment. This paper focused on the evalua-
tion 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. The objective
was to characterize the seasonal variation of these 55
xenobiotic organic compounds in the leachate from
municipal solid waste landfill of Tripoli, Lebanon, and
to assess 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.
Materials andmethods
Chemicals and materials
C-18 cartridge (200mg/6ml), the standards of PAEs,
bisphenols, and 4-nonyl phenol were purchased from
Sigma-Aldrich (USA). PAHs standard were pur-
chased 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-Pro-
mochem (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 are as follows: Naphthalene
(Nap), Acenaphthylene (Acy), Acenaphthene (Ace),
Fluorene (Flu), Phenanthrene (Phen), Anthracene
(Ant), Pyrene (Pyr), Fluoranthene (Flt), Benzo[a]

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anthracene (B(a)A), Chrysene (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 (D(ah)A), and
Indeno[1,2,3-cd]pyrene (IP). (ii) 28 PCBs are as
follows: 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, and CB209. (iii) 6 PAEs are
as follows: Dimethyl phthalate (DMP), diethyl phtha-
late (DEP), di-n-butyl phthalate (DNBP), butyl benzyl
phthalate (BBP), bis-2-ethylhexyl phthalate (DEHP),
and di-n-octyl phthalate (DNOP). (iv) 5 BPs are as
follows: bisphenol C (BPC), bisphenol E (BPE), bis-
phenol F (BPF), bisphenol G (BPG), and 4-nonyl-
phenol (4-NP). Chemical formula, the abbreviation,
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, S2, and S3 in supplemental information.
Study area
The Tripoli landfill is located in Tripoli, on the
coastal area of the Mediterranean Sea and at the level
of the Abu Ali estuary with the coordinates 34° 27′
19.49″ N and 35° 50′ 26.91″ E. Tripoli is the largest
city in northern Lebanon with 450000–550000 inhab-
itants and is characterized by a Mediterranean cli-
mate, cold and wet during the winter. Tripoli landfill,
located in Tripoli along the coastline of the Mediter-
ranean Sea and adjacent to the Abou Ali River, cov-
ers an area of 60000 m2. Tripoli landfill was used
as an uncontrolled dumpsite for over 20years until
1999 when the site was rehabilitated and operated as
a controlled landfill receiving the waste from Al Fay-
haa 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 and then covered 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 con-
ditions like height and degree of compaction. Age
of waste in unit A is about 30years and lower than
10years in unit B. The flow rates 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 periph-
ery 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. Figure1 presents the sampling
points.
Sample collection
The samples were collected in duplicate from the out-
put leachate pipe of both sections every 2months from
September 2017 to September 2018. Six sampling cam-
paigns 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 coordinates of the two sampling points
of leachate are 34° 27′ 21.26″ N and 35° 50′ 29.6″ E
for point A and 34° 27′ 19.22″ N and 35° 50′ 25.46″ N
for point B, and they are depicted in the Fig.1. The lea-
chate was generated by Tripoli landfill draining directly
to the Abou Ali River by pipes and then reached the
Mediterranean Sea without any treatment.
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-pro-
panol, and then 12ml of a mixture of Milli-Q/2-pro-
panol (85/15, v/v) acidified to pH 2.5 with sulfuric
acid. After conditioning step, 100ml of leachate sam-
ples 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,

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v/v) acidified to pH 2.5 and then dried with N2. The
targeted compound was eluted with a 12ml of a mix-
ture 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 (Fig.2).
Fig. 1 Photo of sampling site and sampling points from September 2017 to September 2018, for every 2months
Fig. 2 PAHs distribution
prole for the leachate of
units A and B
0
20
40
60
80
1
00
A B A B A B A B B B
Sep 2017 Dec 2017 Feb 2018 Apr 2018 Jun
2018
Sep
2018
%
Ring 2 Ring 3 Ring 4 Ring 5 Ring 6

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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 × 0.25mm i.d. × 0.25-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.
Quantitative analysis andquality control
The concentration of each compound was determined
according to 6-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 quanti-
fication, 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 and were free from any targeted PCBs, PAHs,
and BPs. However signification level higher than
LOQ of PAEs was sometimes detected in procedural
blanks, especially for DNBP and DEHP. The concen-
tration of DNBP and DEHP measured in blanks could
reach up to 4% of the concentration measured in lea-
chate sample. Consequently, the results of DNBP and
DEHP were corrected by subtracting the values found
in procedural blanks as recommended by Net etal.
(2015c). The precision and accuracy of the method
were confirmed by duplicate analysis of the sam-
ples. The precision of the technique, specified by the
standard deviation, was found between 1 and 5%.
Results anddiscussion
PAHs concentration and composition profile
Among the 16 PAHs, 12 of them were detected,
and their concentrations are presented in Table1.
High heterogeneity of the number of detected com-
pound and the concentration level of each sampling
point and each campaign were found. In unit A, Nap
showed the highest level and followed by Acy and
Pyr. In unit B, similarly, Nap was detected dominant,
and it followed by Phen and Flu (Fig.3). Our results
showed the dominant of low molecular weight PAH
with 2 and 3 aromatic rings (Nap, Ace, Phen, and
Flu). High molecular weights with 5–6 aromatic ring
including B(b)F, B(a)P, D(ah)A, and B(ghi)P were
bellow LOQ in both units. However, the Σ16PAHs in
leachate from unit B were found to be 5 times higher
than which found 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).
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) in south Lebanon (0.19–1.1µg l−1) and
Kalmykova etal. (2013) 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; IARC,2002). The
highest Σ8PAHs carcinogenic 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
Fig.2 shows the composition profile of PAHs.
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 indicates
petroleum origin and > 0.35 indicates 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 shows the mixture of com-
bustion 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the pyrolytic sources and > 10 indi-
cates thepetrogenic sources (Maliszewska-Kordybach
etal., 2008). Our results showed the ratio of 0–13.6,
where 75% of Phe/Ant ratio is < 10 and 25% > 10.
Thus, the combustion is the dominant source in the
unit A. Similarly for unit B, the Phe/Ant ratio was

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Table 1 Individual concentration of PAHs in units A and B from September 2017 to September 2018
“-” : < LOQ; n = 4 replicate
* Human carcinogen compounds IARC (2002)
Sampling date Sep 2017 Dec 2017 Feb 2018 Apr 2018 Jun 2018 Sep 2018
PAHs ng·l−1A B A B A B A B B B
Nap* 260 ± 10 153 ± 7 5751 ± 12 12,221 ± 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 18,843 ± 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 0.01 429 ± 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
Σ8 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 - -

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from 0 to 43.6, where 67% is < 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 is < 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
confirms the combustion process as the major ori-
gin of PAHs in our leachate. The cross plots of BaA/
BaA + Ch and Ant/Ant + Phe in the two units indi-
cated that PAHs originated mainly from combustion
sources (Fig.3). Our results show that the PAHs in
leachate samples of Tripoli landfill were predomi-
nantly 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; Halwani etal., 2020; Zhang etal., 2013). 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).
PCBs composition profile
Among the 28 PCBs, 15 were detected, and the level
of Σ28PCBs was 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 individ-
ual concentration of PCB is presented in Table2.
Our results are in the same range 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 environ-
ment. 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; Merhaby etal., 2019).
The composition profile of PCB in leachate of unit A
and unit B is presented in Fig.4.
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 units 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;
Fig. 3 Cross plot of Ant/
Ant + Phe versus BaA/
BaA + Ch for leachate sam-
ples during the study period

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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.
PAEs concentration profile
Five PAEs among the six were detected, and the
result is shown in Fig.5. The individual concentration
of PAEs is presented in Table3. PAEs were detected
at the highest level compared to PAHs and PCBs
in the leachate obtained from points A and B. The
Σ6PAEs varied from 571 and which can be up 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.
PAEs in the leachate of Tripoli landfill is
higher than which found in China at 50–62µg l−1
Table 2 The concentration of PCBs in units A and B leachate from September 2017 to September 2018
“-” : < LOQ; N.A.not analyzed
Sampling date Sep 2017 Dec 2017 Feb 2018 Apr 2018 Jun 2018 Sep 2018
PCBs ngl−1A B A B A B A B B B
CB8 40 ± 3 40 ± 1.4 60 ± 2.8 163 ± 7.1 5 ± 0.3 80 ± 4.2 15 ± 1.4 116 ± 7.2 840 ± 11.3 44 ± 3.5
CB18 - - - - - - - - - -
CB28 374 ± 11 18 ± 1.1 28 ± 0.7 15 ± 1.4 3 ± 0.2 2 ± 0.1 1 ± 0.1 14 ± 0.9 28 ± 1.4 6 ± 0.6
CB44 4 ± 0.4 2 ± 0.1 2 ± 0.1 6 ± 0.6 7 ± 0.4 6 ± 0.5 6 ± 0.4 7 ± 0.6 38 ± 1.8 6 ± 0.6
CB52 - - - - - - - - - -
CB66 - - - - - - - - - -
CB101 - - 63 ± 2.5 18 ± 1.6 20 ± 1.4 16 ± 1 - - 3 ± 0.3 -
CB77 5 ± 0.4 N.A 19 ± 1.3 61 ± 3.5 3 ± 0.1 40 ± 1.4 1 ± 0.03 42 ± 1.7 44 ± 1.4 43 ± 3
CB81 2 ± 0.1 2 ± 0.1 1 ± 0.1 2 ± 0.2 - 1 ± 0.04 1 ± 0.03 6 ± 0.3 926 ± 11 4 ± 0.4
CB114 - - - - - - - - - -
CB105 - - - - - - - - - -
CB123 - - 6 ± 0.2 4 ± 0.3 - - - 3 ± 0.1 9 ± 0.7 -
CB128 1 ± 0.1 1 ± 0.1 1 ± 0.1 1 ± 0.1 - 1 ± 0.03 - 1 ± 0.01 2 ± 0.1 -
CB126 1 ± 0.03 2 ± 0.2 2 ± 0.1 4 ± 0.4 - 2 ± 0.1 - 13 ± 1.3 3 ± 0.3 2 ± 0.1
CB138 - - - - - - - - - -
CB187 7 ± 0.7 9 ± 0.9 9 ± 0.8 12 ± 1.1 4 ± 0.3 3 ± 0.2 2 ± 0.1 7 ± 0.4 3 ± 0.2 8 ± 0.7
CB118 N.A N.A 9 ± 0.6 5 ± 0.2 1 ± 0.01 1 ± 0.1 2 ± 0.03 35 ± 1.4 117 ± 7.1 1 ± 0.1
CB167 - - - - - - - - - -
CB157 2 ± 0.1 1 ± 0.04 1 ± 0.1 1 ± 0.03 - - - 2 ± 0.1 1 ± 0.04 -
CB156 - - - - - - - - - -
CB169 12 ± 1.1 17 ± 1.4 1 ± 0.1 1 ± 0.1 8 ± 0.8 25 ± 1.4 - 17 ± 1.1 4 ± 0.03 -
CB170 - - - - - - - - - -
CB180 - 1 ± 0.1 - - - 1 ± 0.1 - 1 ± 0.03 - -
CB153 - - - - - - - - - -
CB189 - - - - - - - - - -
CB195 1 ± 0.05 1 ± 0.1 - - - 2 ± 0.1 1 ± 0.03 1 ± 0.04 2 ± 0.1 -
CB206 - - - - - - - - - -
CB209 - - - - - - - - - -
Min - - - - - - - - - -
Max 374 39 63 163 20 79 15 116 926 44
Σ7PCBi 374 19 101 38 25 19 4 49 148 8
Σ12PCB‑DL 21 25 38 77 13 69 4 117 1105 50
Σ28PCB 446 96 202 293 52 178 30 261 2020 115

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(He etal., 2009) and in Sweden landfill leachate
(1.2–46.8µg l−1) (Kalmykova etal., 2013; Öman &
Junestedt, 2008). High levels of Σ6PAEs might be
attributed to the abundance of plastics material and
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.,
2015b). PAEs have been used in many materials or
products including PVC products, building materi-
als including paint, adhesive, wall covering, personal
Fig. 4 PCB composition
prole in leachate of unit A
and unit B from September
2017 to September 2018 in
ng l−1
0
50
100
150
200
250
300
350
400
450
A B A B A B A B B B
Sep 2017 Dec 2017 Feb 2018 Apr 2018 Jun 2018 Sep
2018
Σ 7 PCBi Σ 12 PCB-DL Σ other PCB
1100
1400
1700
2000
2300
ng·l-1
0
500
1000
1500
2000
2500
3000
3500
4000
4500
5000
5500
6000
6500
7000
7500
8000
A B A B A B A B B B
Sep 2017 Dec 2017 Feb 2018 Apr 2018 Jun 2018 Sep 2018
µg·l-1
DMP DEP DnBP BBP DEHP
Fig. 5 PAEs composition prole in leachate of unit A and unit B from September 2017 to September 2018 in µg l−1

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Table 3 The concentration of PAEs and BPs in leachate from September 2017 to September 2018
Sampling
date
Sep 2017 Dec 2017 Feb 2018 Apr 2018 Jun 2018 Sep 2018
PAEs µgl−1A B A B A B A B B B
DMP 0.1 ± 0.01 0.1 ± 0.007 1 ± 0.1 0.1 ± 0.01 0.5 ± 0.02 0.3 ± 0.01 0.1 ± 0.01 0.001 ± 0.0001 3 ± 0.2 1 ± 0.02
DEP 71 ± 1.8 47 ± 1.4 70 ± 6.5 66 ± 2.8 3388 ± 11.3 692 ± 16 510 ± 9.9 184 ± 10.6 1477 ± 8 1403 ± 15.5
DnBP 3 ± 0.2 4 ± 0.01 19 ± 1.4 19 ± 1.6 115 ± 2.8 155 ± 2.8 5 ± 0.4 44 ± 1.4 48.9 ± 2.8 72 ± 3
BBP 0.2 ± 0.01 0.03 ± 0.001 10 ± 0.4 5 ± 0.4 3 ± 0.1 3 ± 0.1 0.1 ± 0.01 1 ± 0.01 1 ± 0.1 1 ± 0.05
DEHP 497 ± 14.1 1985 ± 15.5 1043 ± 16.9 4110 ± 19.8 4525 ± 17 4042 ± 20 433 ± 18.5 170 ± 2.8 1809 ± 17 1005 ± 18.4
DNOP - - - - - - - - - -
Min - - - - - - - - - -
Max 497 1985 1043 4110 4525 4042 510 184 1809 1403
Σ 6 PAE 571 2036 1143 4200 8031 4892 948 399 3340 2481
BPs µg l−1A B A B A B A B B B
BPF 199 ± 10 268 ± 14.1 175 ± 11.3 170 ± 5.6 140 ± 3.5 132 ± 3.5 175 ± 8.4 154 ± 5.7 122 ± 4.3 156 ± 5
BPE 200 ± 9.4 281 ± 18.4 175 ± 9.1 173 ± 7.7 140 ± 2.8 163 ± 5.7 98 ± 4.5 175 ± 8.2 153 ± 6.3 162 ± 3.6
BPC - - - - - - - - - -
BPG 23 ± 1.2 47 ± 1.4 27 ± 1.2 89 ± 2.1 1 ± 0.08 64 ± 1.3 3 ± 0.12 51 ± 2 107 ± 1.5 45 ± 0.8
4‑NP 200 ± 11.3 276 ± 16.9 175 ± 9.8 173 ± 7 140 ± 4.2 173 ± 7 175 ± 7.3 175 ± 8.4 155 ± 5.7 175 ± 7.9
Min - - - - - - - - - -
Max 200 281 175 173 140 173 175 175 155 175
Average 124 174 111 121 84 106 90 111 107 108
Σ 4 BPs 422 596 378 433 282 359 277 380 382 363

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care products, medical devices, packaging, print-
ing inks and coatings, pharmaceuticals and food
products, and textiles. These products and materials
have been frequently used in everyday life of Trip-
oli 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 poly-
meric matrix. Thus, they can be easily released from
the waste containing PAEs to the leachate (Net etal.,
2015b). 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; Öman & 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 land-
fill since it 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 prod-
ucts have been dumped into Tripoli landfill without
sorting, which can be the important source of DEHP.
Indeed, DEHP is abundantly used as PCV plasticiz-
ers (Kalmykova etal., 2013; Net etal., 2015b). Also,
other types of waste might be the sources of PAEs in
Tripoli landfill like fibers, lubricants, cosmetics, and
pharmaceutical residue (Peijnenburg, 2008). Figure5
shows the PAEs composition profile in leachate of
unit A and unit B from September 2017 to September
2018.
Phenolics and its derivatives composition profile
The results of phenolic compounds are shown in
Table3 and in Fig.6. 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, aver-
age 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 con-
centration was observed between the sampling period
0
150
300
450
600
750
900
A B A B A B A B B B
Sep 2017 Dec 2017 Feb 2018 Apr 2018 Jun 2018 Sep 2018
µg·l-1
4-NP BPF BPE BPC BPG
Fig. 6 Phenolic compound prole in leachate of unit A and unit B from September 2017 to September 2018 in µg l−1

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for both units, which might be due to the seasonal
variation (Asakura etal., 2004; Urase & Miyashita,
2003). Different concentrations of BPs were 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 containing BPs like food contact materi-
als (Česen etal., 2016; Liao & Kannan, 2013), digital
media, electronic equipment, medical instruments,
pipes and toys (Chen etal., 2016), thermal paper
(Skledar & Mašič, 2016), dyes, and 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 Swe-
den at 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 incinera-
tion residues (Kurata etal., 2008). The PBs in Tripoli
landfill leachates are thought to be important pollut-
ants and pose a high risk to the surrounding environ-
ment and human. For example, BPA can cause a pro-
found effect on organism at low centration. Indeed,
BPA can impact on the endocrine and reproductive
systems, on the nervous systemand has genotoxicity
caracter and many others (Xing etal., 2022).
Fluxes andpotential sources totheAbou Ali River
andMediterranean Sea
The annual rainfall and the surface of landfill were
determined at 700mm/year and 60000 m2 respec-
tively. If the total among of rainfall on the landfill
surface solubilize pollutants containing in the land-
fill 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 pollutant inputs of the Tripoli land-
fill to the Abou Ali River and then to the Mediter-
ranean 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
thrown 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 assessments givean imageabout
the pollution effect of Tripoli landfill on the Abou Ali
estuary andMediterraneanSea. In view of this result,
the flow of PAEs, BPs, and 4-NP were significant
and can strongly impact the ecosystem functioning
due to their potential impact on the endocrine system
of wild flora and fauna (Net etal., 2015b; Xing etal.,
2022). Moreover, a recent study conducted by our
group showed high contamination level of physico-
chemical parameter and metallic pollutants in Tripoli
leachate. The annual fluxes of TDS, COD, chloride,
BOD, TKN, ammonium, TSS, VSS, sulfates, total
phosphorus, and nitrates were 711 tons, 577 tons,
253 tons, 207 tons, 137 tons, 96 tons, 58, 38 tons, 25
tons, 7 tons, and 3 tons, respectively. And, the annual
fluxes of the Σ21metals achieved 217 tons (Moustafa
etal., 2022).
Conclusions
Due to the absence of waste management strategy in
Lebanon, environmental pollution generated from
uncontrolled landfills become more complicated to
manage especially in big cities like Tripoli. The con-
centration of fifty-five organic pollutants including 16
PAHs, 28 PCBs, 6 PAEs, 4 BPs, and 4-nonylphenol in
the leachate of Tripoli landfill has been investigated.
The concentration of the selected xenobiotic organic
pollutants in leachate collected from old unit is lower
than which collected from new unit, which may be due
to the age and the transformation of waste in each unit.
The concentration of carcinogenic POPs detected the
high level in leachate and thus may cause high toxic-
ity. Among the selected compounds, the highest levels
were determined for PAEs and followed by phenolic
compounds (PBs and 4-nonyl phenol). It is worth to
note that PAEs and phenolic compounds are known
as endocrine disruptor compounds. The annual fluxes
of Σ6PAEs, Σ4BPs, and 4-NP were estimated, respec-
tively, at 116.85kg, 15.93, and 7.58kg to the Abou
Ali River and Mediterranean Sea. The release of these
endocrine disruptor compounds may impact on the
ecosystem functioning.
There are no landfills for chemical and hazardous
waste in Lebanon, and the results of the present study

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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 harms envi-
ronmental and human health. It is clear that stopping
the dumping of hazardous waste into Tripoli landfill
is one of the solutions to prevent the damage. The
treatment of the leachate becomes the necessity to
reduce the impacts on the surrounding environment
and particularly the Abou Ali River and coastal area
nearby. That 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 techni-
cal support provided by CPER CLIMBIO project. We acknowl-
edge the nancial support from the “Agence Universitaire de la
Francophonie (AUF)” of Tripoli, North Lebanon, which pro-
vided a PhD scholarship for Ahmad Moustafa.
Author contribution 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 nal
manuscript.
Funding Ahmad Moustafa is nancially supported by a PhD
fellowship from the Agence Universitaire de la Francophonie
(AUF)” of Tripoli, North Lebanon.
Data availability The authors declare that all relevant data
supporting the ndings of this study are included in this article.
Declarations
Conflict of interest The authors declare no competing inter-
ests.
Disclaimer The funders had no role in the writing of the man-
uscript or in the decision and use of data.
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