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Metal dissolution from end-of-life solar photovoltaics in real landfill
leachate versus synthetic solutions: One-year study
Preeti Nain, Arun Kumar
⇑
Department of Civil Engineering, Indian Institute of Technology, New Delhi, India
article info
Article history:
Received 30 March 2020
Revised 3 July 2020
Accepted 4 July 2020
Keywords:
End-of-life
Solar panel
Photovoltaic
Metal leaching
MSW landfill leachate
abstract
To investigate the after end-of-life concerns of solar panels, four commercially available photovoltaics
(reduced to 15!15 cm
2
size) in broken and unbroken conditions were exposed to three synthetic solu-
tions of pH 4, 7, 10 and one real municipal solid waste landfill leachate for one year. Metals leaching,
encapsulant degradation and release, probability of leached metals exceeding their surface water limits,
and change in pollution index of leachate after dumping of solar panels were investigated. Rainwater
simulating solution was found to be predominant for metal release from silicon-based photovoltaics,
with silver, lead and chromium being released up to 683.26 mg/L (26.9%), 23.37 mg/L (17.6%), and
14.96 mg/L (13.05%), respectively. Copper indium gallium (de) selenide (CIGS) photovoltaic was found
to be least vulnerable in various conditions with negligible release of indium, molybdenum, selenium
and gallium with values ranging between 0.2 and 1mg/L (0.30%-0.74%). In contrast, minimal metals were
released in real landfill leachate compared to other leaching solutions for all photovoltaics. Positive cor-
relation was observed between encapsulant release and metal dissolution with a maximum encapsulant
release in silicon-based photovoltaics in rainwater conditions. The calcualtion of values of probability of
exceedance of leached metals to their respective surface water limits for aluminium (multi- and mono-
crystalline-silicon), silver (amorphous photovoltaic) and indium (CIGS) indicated maximum value to be
92.31%. The regression analysis indicated that conditions of the modules and pH of the leaching solution
play significant roles in the metal leaching. The increase in leachate contamination potential after one-
year of photovoltaics dumping was found to be 12.02%, 10.90%, 15.26%, 54.19% for amorphous, CIGS,
mono and multi crystalline-silicon photovoltaics, respectively. Overall, the maximum metal release
observed in the present study is 30% of the initial amount under the most stressful conditions, which sug-
gests that short-term leaching studies with millimeter sized sample pieces do not represent the realistic
dumping scenarios.
!2020 Elsevier Ltd. All rights reserved.
1. Introduction
In the recent years, solar photovoltaic (PV) industry has grown
tremendously among all the renewable energy sectors because of
declining cost of manufacturing materials and increasing energy
demand. This promising technology raises the concern of unin-
tended outcome of the huge going-to-be generated end-of-life
(EoL) solar PV waste stream once their operational life ends
(Mahmoudi et al., 2019). The working life of a solar panel is
approximately 25 to 30 years suggesting the generation of enor-
mous waste in next few years which could have potential to
impact environment and human health (Nain and Kumar. 2020).
Solar panels consist of both precious and carcinogenic metals (cad-
mium, chromium, lead, silver, selenium and tellurium) that require
recycling and EoL management to recover rare valuable metals and
prevent environmental pollution (Sica et al., 2018). Despite the
usage of hazardous metals in solar panels, there is limited litera-
ture available on (i) their impacts on environment and fate in real-
istic environmental settings (i.e., landfill conditions) (Cyrs et al.,
2014), and (ii) potential risks towards environment and human
health (Mahmoudi et al., 2019). In specific, there is potential for
metal release and environmental contamination on dumping of
EoL solar panels (Brown et al., 2018).
At present, China and India have the highest installed solar PV
capacity, while regulations for managing EoL waste and recycling
are still non-existent there (Ding et al., 2016) and PV waste han-
dling comes under e-waste guidelines mostly (Komoto, 2016).
Developed countries, such as Germany has revised waste electrical
and electronic equipment’s regulations incorporating obligations
https://doi.org/10.1016/j.wasman.2020.07.004
0956-053X/!2020 Elsevier Ltd. All rights reserved.
⇑
Corresponding author.
E-mail address: arunku@civil.iitd.ac.in (A. Kumar).
Waste Management 114 (2020) 351–361
Contents lists available at ScienceDirect
Waste Management
journal homepage: www.elsevier.com/locate/wasman
on producers for EoL solar waste treatment. Japan’s environment
ministry require manufacturers to participate in recycling along
with venturing with Hamada company. American solar company,
First Solar has setup recycling facilities in United States, Germany
and Malaysia (Xu et al., 2018). The details on regulations can be
obtained from various published studies (Jia and Fang, 2016;
Wild-Scholten et al., 2005). Detailed PV leaching studies under
realistic landfill conditions are crucial to understand the metal dis-
solution from this waste stream. The present aggressive and short-
period standard waste characterization tests (i.e., Waste extraction
tests, WET and Toxicity characterization leaching procedure, TCLP)
investigated in previous studies (Brown et al., 2018; Collins and
Anctil, 2017) might not represent accurately the complex reaction
and conditions prevalent in a real landfill. Microbial processes can
impact the fate of PVs in municipal solid waste (MSW) landfills as
they can do biotransformation and biomagnification of nano-
metallic species (Liu et al., 2015).
Commercial available photovoltaics can be classified into crys-
talline silicon (c-Si) and thin-films. The former has the advantage
of high efficiency while the latter are more flexible and cost-
effective. Both technologies are safe during the use, as all the met-
als are well encapsulated providing protection against moisture
(Rebecca Brun et al., 2016). Most of the solar PV studies in litera-
ture focused on recycling and efficiency aspect of PVs (Gangwar
et al., 2019; Deng et al., 2019). However, after operational life,
these panels could be dumped in open environment resulting in
prolonged exposure to water. A recent study investigated the prob-
ability of material release from EoL solar PVs on the basis of survey
of stakeholders of PV industry (Nain and Kumar, 2020). With
respect to amorphous-Si (a-Si) studies, limited literature is avail-
able probably due to less amount of semiconductor material and
small market share (Savvilotidou et al., 2017; Tammaro et al.,
2016; Zeng et al., 2015). Crystalline-Si modules have lower poten-
tial to release materials compared to thin-film modules (Tammaro
et al., 2016). Thin-films PVs are evolving rapidly and entering into
market due to recent advancements in nanomaterial- based quan-
tum dots (Brown et al., 2018). Further, more rare and valuable met-
als like gallium, indium and silver are used in thin-film PVs, and
thus, detailed investigation is required in order to understand their
leaching behaviour.
In low-pH condition, more than 15% of lead can be released
from c-Si PVs in 56 days (Zapf-Gottwick et al., 2015). In similar
conditions, high concentrations of cadmium, molybdenum and
selenium are reported to be released from copper indium gallium
(de) selenide (CIGS) PVs (Zimmermann et al., 2013). Ramos-Ruiz
et al. (2017) reported the 73% of cadmium and 21% of tellurium
release in column simulating acidic landfill conditions. Chakankar
et al. (2019) investigated the use of microorganisms for metals
recovery from spent PVs and recovered Al up to 89% and Te up
to 100%. Further, the most comprehensive work done so far was
by Nover et al. (2017), who reported the release of 1.4% of lead
from 5!5 cm
2
c-Si panel pieces and 62% of cadmium from cad-
mium telluride (CdTe) pieces in low pH conditions after 360 days.
Whereas the present study employs bigger size module pieces
treated with real MSW landfill leachate in order to investigate
the contamination potential of solar PV waste under landfill
conditions.
To address any concerns related to EoL disposal of solar PVs, this
study investigated the magnitude of heavy metal leaching from
existing PV technologies for one year in three different synthetic
water-based leaching solutions and one real MSW landfill leachate.
Previous short-term leaching studies carried out in the past inves-
tigated the stability of PV modules and metal leaching behaviour in
extremely harsh conditions (Zapf-Gottwick et al., 2015;Rebecca
Brun et al., 2016), which are by no means representative of field
conditions. Also, in contrast to previous studies, the present study
investigated the fate of second generation thin-film modules in
realistic environmental conditions, i.e., without shaking, uncon-
trolled temperature, and large size panel fragments, and real
MSW leachate, which is essential considering the utilisation of
valuable metals along with few toxic elements (Tammaro et al.,
2016). Encapsulant degradation and release was quantified in
order to understand its effect on metal leaching mechanism. Lea-
ched metal values were compared with surface water regulatory
limits to examine the potential concerns towards environmental
contamination. Overall, the findings will provide a comprehensive
understanding on (i) the fate of two solar PVs generation in simu-
lated solutions and real municipal landfill leachate, and further, (ii)
pollution potential after their disposal in terms of leachate pollu-
tion index (LPI).
2. Materials and methods
The overview of methodology followed in the present study is
shown in Fig. 1.
2.1. Sample preparation and metal content determination
Four commercially available solar panels were purchased:
Mono-crystalline silicon (Mono-Si), Multi-crystalline silicon
(Multi-Si), Amorphous PV (a-Si) and CIGS (New Delhi, India) (spec-
ifications are shown in Table S1). The layered structure of a typical
c-Si PV consists of following components (mass %): frame (16.9%),
cover glass (67.8%), anti-reflective coating (<1%), encapsulant
(3.4%), front contact (0.4%), absorber (6.1%), back contact (0.4%),
encapsulant (3.4%), backsheet (0.8%) and junction box (Mani
et al., 2020). Further, the detailed descriptions about various layers
and compositions of PVs can be found in literature (Lasnier, 2017).
Panels have been primarily dismantled manually by removing
embedded cables and junction-box (whenever present). Further,
outer aluminium frames were removed considering the easily
recyclability of frames and supporting structure at present, thus,
assuming that panels will be dumped without frames (García-
Valverde et al., 2009). The purchased solar panels with size from
30!30 cm
2
to 167!60 cm
2
were reduced to 15!15 cm
2
pieces
using metal sheet cutting machine and glass cutter. In order to
simulate realistic leaching scenario, actual panels with all the lay-
ers, i.e., including glass and encapsulation layer were used.
For total metal content determination, the 15!15 cm
2
size
pieces were reduced to 1–2 cm
2
in size using micro-shear wire cut-
ter, which was further reduced manually to mm size using cast
iron mortar-pestle and stainless steel mixer-grinder for achieving
homogenised samples. The resulted samples were then sieved
manually using a standard 1 mm ASTM sieve, and the fragments
that passed through were used for metal analysis. For heavy metal
analysis, standard method APHA (2010) was adopted (Federation,
2005). The digested samples were then cooled down to room tem-
perature; filtered with <0.22 mm Millipore filter papers after dilu-
tion and analysed with ICP-MS (Agilent 7900). All analyses for
leached metals were conducted in triplicates.
2.2. Leaching solutions
Four different leaching solutions were used: Rainwater, RW (pH
4); Groundwater, GW (pH7); Sea water, SW (pH10); MSW landfill
leachate, ML (pH 7.6). Leaching solutions were first prepared and
stored overnight in serum bottles with screw caps to make sure
that they reached equilibrium before use. The RW, GW, SW leach-
ing solutions (chemical composition information is given in
Table S2) were made using Milli-Q water except ML (leachate col-
lection and characterisation information is given in SI) which was a
352 P. Nain, A. Kumar / Waste Management 114 (2020) 351–361
real leachate taken from Okhla landfill in Tughlakabad, New Delhi,
India.
2.3. One-year leaching experiment
All leaching experiments were carried out in closed borosil glass
trays (40.4!25.7!6.1 cm
3
) with leaching solution of 2000 ml and
one 15!15 cm
2
module piece without agitation at room tempera-
ture. The solid- to- liquid ratios (PV sample: leaching solution) for
a-PV, mono-Si, multi-Si and CIGS samples used were approxi-
mately 1:13, 1:11, 1:11 and 1:3, respectively. Two sets of experi-
mental setups were used:
i. The first setup include broken pieces. These pieces were pre-
pared by applying impact on the front glass surface to
induce breakage that would most likely take place when
modules are dumped into landfill. This condition was used
to simulate the scenario of dumping of broken panels due
to breakage during transportation, installation or operation.
The impact was given by a fixed pressure of 250kN using an
impact machine.
ii. The second setup includes the intact PV pieces without any
glass cracking. This condition was used to simulate the sce-
nario of direct dumping of PV panels in the environment
after completion of their operational life. For the CIGS mod-
ules, only second condition was used as breakage was not
possible due to absence of glass and the module being the
flexible panel. The glass and waterjet cutter were used for
cutting the modules for definite shape and sharp edges with-
out any glass breakage.
Monthly samples of 50 ml were withdrawn for metal and car-
bon release analysis. To keep the leaching solution’s initial volume
constant at 2000 ml, volume was corrected with respective solu-
tion after each sampling. These corrections were included during
data analysis which considers the leached metal and carbon con-
tent loss in the solution due to sampling. For quantifying the bind-
ing agent (i.e., encapsulant or ethylene vinyl acetate, EVA) release,
which is mainly formed of carbon, monthly samples were analysed
for total carbon release using total organic carbon (TOC) analyser.
The experiment was performed without agitation to simulate real-
istic dumping scenarios conditions prevailing in an actual landfill.
2.4. Solar PV waste-associated LPI
In order to examine the change in existing pollution potential of
a matrix if dumped with solar PVs, the solar PV waste-associated
leachate pollution index (SPW-LPI) is calculated. LPI represents
the contamination potential of a landfill leachate with summariz-
ing the complex leachate pollution data and, thus, facilitating its
communication to general public, field professionals and policy
makers (Kumar & Allapat, 2004). In the present study, LPI values
for two leaching solutions, ML (real MSW leachate) and RW (rain-
water) were calculated before (LPI) and after 12 months of solar
panels dumping (LPI’) as per Kumar and Alappat (2005) (Equation
(1)). Further, the change in the leachate pollution potential was
calculated as
D
LPI = LPI’–LPI. Out of 18 parameters used for the cal-
culation of LPI, phenolic compounds and cyanide were not
considered.
SPW "LPI ¼P
m
i¼1
w
i
p
i
P
m
i¼1
w
i
ð1Þ
where, LPI: weighted additive leachate pollution index, w
i
: weight
for the i
th
pollutant variable, p
i
: sub index score of the i
th
leachate
pollutant variable, m: number of leachate pollutant variables used
in calculating LPI P
m
i¼1
w
i
= 1. If m < 18, thenP
m
i¼1
w
i
< 1.
2.5. Mathematical methods
Analysis of variance (ANOVA) for various output parameters
was conducted using Minitab version 18 software (Windows 10)
and significance value was tested for a 95% confidence level test
(i.e.,
a
= 0.05). t-tests were conducted using Minitab 18, to see if
mean values of two different samples differ statistically.
Fig. 1. Overview of methodology for present study.
P. Nain, A. Kumar / Waste Management 114 (2020) 351–361 353
The fittings of linear, exponential and sigmoidal models were
investigated, and sigmoidal model was found to be the best fit
empirically for maximum cases:
yt
ðÞ¼y
max
þa"y
max
1þð
x
i
Þ
K
hi
m
ð2Þ
concentration of dissolved element over time, y(t) [mg/L]; maximal
leached element concentration, y
max
[mg/L]; minimal leached
concentration, a[mg/L]; hill slope, K [d
-1
]; inflection point i,
[(y
max
-a)/2]; asymmetrical factor, m. y
max
is constrained by
maximum element concentration leached after one year. The rate
constants (K
d
) of models with coefficient of correlation, R
2
>0.95
are only considered.
In addition, probability of exceedance (PoE, %) of leached metals
to their respective standard surface water concentration values is
calculated as per Equation (3):
PoE %ðÞ¼ N"R
m
Nþ1
!"
!100 ð3Þ
where, N: total number of observations (N = 12); R
m
: the rank of
metal value exceeding the limit when arranged in ascending order.
3. Results and discussion
3.1. Metal dissolution as a function of time
In the past, most of the leaching and recycling studies have been
mostly done for crystalline (mono- and multi-Si) solar PVs
(Gangwar et al., 2019; Nain and Kumar, 2019) and thus, knowledge
gap exists for studies focusing on a-Si modules. The initial metal
concentrations of four PVs are summarized in Table S1 which
shows that all PVs consist of general elements and few rare ele-
ments like In and Te. Precious metal, Ag was found to be present
in high concentration in amorphous PV. Elemental content analysis
shows that all the selected heavy metals except Hg and Ti were
found to be present in the studied PVs. Further, As, In and Cd were
found to be present in low amounts in silicon PV as compared to
CIGS.
Amorphous thin-film modules do not contain much hazardous
metals, however, small amounts of Pb is used in active layer for
electrical connections. The absolute metal concentrations released
from a-Si PVs in four leaching solutions (primary y-axis) along with
the percent of eluted metal with respect to initial mass after
365 days (secondary y-axis) are shown in Fig. 2. It also includes
the best- fitted model (i.e., sigmoidal mode) (Table S3). Metals with
environmental relevance and high leachability are discussed here
and the information about the remaining leached metals are pre-
sented in the Table S4. With respect to initial content of metals
of a-Si PV, Ag leached maximum with 683.26 mg/L (26.9%) fol-
lowed by Pb, 23.37 mg/L (17.6%); Cu, 483.24 mg/L (13.7%) and least
by Cr, 14.96 mg/L (13.05%). In terms of mg per gm of PV, stated
metals leached 5.0, 3.6, 0.174 and 0.112 respectively. Initially,
the leaching rates of Pb and Ag were found to be almost same,
though, with increase in time, extent of leaching was found to be
higher in RW than in ML. However, extent of leaching gradually
diminished in the last three months of study. In RW after 365 days,
the maximum Pb release from broken panel reached approxi-
mately 17.6% whereas the maximum Pb release for unbroken panel
reached 12.5%. It can be seen from Figure S1 that there is visible
metal layer dissolution happened at the corner of an amorphous-
PV after 12 months of exposure to RW. Semiconductor film was
found to be highly dissolved from the extremely dissected sections
than from the intact corner which was undisturbed during the cut-
ting process. In case of GW conditions, the amounts of leached Pb
were found to be 12.9% and 3.05% for broken and unbroken a-Si PV,
respectively. ANOVA analysis for Pb leaching indicated that mod-
ule condition (i.e., broken or unbroken) and leaching solution type
(i.e., RW, GW, SW, ML) had significant effects on metal dissolution
rate constant (p <0.05). However, the p-value for factor (interaction
of module condition*leaching solution type) was found to be>0.05
(Pb: 0.128; Ag: 0.683; Cu: 0.673) suggesting the insignificant effect
of interaction of module condition and leaching solution type on
metal release. For Ag, Cu and Cr, the solution type was found to
affect the metal dissolution significantly whereas the p-value for
module condition was found to be greater than 0.05, showing
insignificant effect on metal release. In RW-broken conditions,
other metals-of-concern such as As and Cd showed maximum
0
2
4
6
8
10
12
14
16
18
0
4
8
12
16
20
24
30 60 90 120 150 180 210 240 270 300 330 360 390
% leached of ini!al concentra!on
Pb release (mg/l)
Time (days)
Pb
B-RW UB-RW B-GW UB-GW B-SW UB-SW B-ML UB-ML
0
3
6
9
12
15
18
21
24
27
0
100
200
300
400
500
600
700
30 60 90 120 150 180 210 240 270 300 330 360 390
% leached of ini!al concentra!on
Ag release (mg/l)
Time (days)
Ag
B-RW B-GW B-SW B-ML UB-RW UB-GW UB-SW UB-ML
0
2
4
6
8
10
12
14
0
50
100
150
200
250
300
350
400
450
500
30 60 90 120 150 180 210 240 270 300 330 360 390
% leached of ini!al concentra!on
Cu release (mg/l)
Time (days)
Cu
B-RW B-GW B-SW B-ML UB-RW UB-GW UB-SW UB-ML
0
1
2
3
4
5
6
7
8
9
10
11
12
13
0
3
6
9
12
15
30 60 90 120 150 180 210 240 270 300 330 360 390
% of ini!al concentra!o n leached
Cr release (mg/l)
Time (days)
Cr
B-RW B-GW B-SW B-ML
UB-RW UB-GW UB-SW UB-ML
Fig. 2. Amorphous photovoltaics one year leaching kinetics for lead, silver, copper,
and chromium into RW, pH 4 (square); GW, pH7 (daimond); SW, pH10 (circle) and
ML, MSW leachate (triangle); Two modules conditions: Broken, B (unfilled symbols and
dashed lines) and UBroken, UB (filled symbols and solid lines).
354 P. Nain, A. Kumar / Waste Management 114 (2020) 351–361
release of 0.003 mg/L (1%) and 0.001 mg/L (0.125%) respectively
(Table S4). Ga and In were not detected in all the leaching scenarios
after one-year, whereas Se and Zn showed release of 0.02 mg/L
(1.25%) and 0.194 mg/L (2.84%), respectively.
Metal releases in SW and ML conditions were found to be min-
imal (<1%), which is in accordance with previous studies reporting
decrease in leaching rate with an increase in pH of leaching solution
(Zapf-Gottwick et al., 2015). Alkaline behaviour facilitates the sur-
face adsorption and solids formation with major minerals and
therefore, reducing the release of anions (Cornelis et al., 2008). Also,
high pH favours pozzolanic reaction that results in formation of
lead silicate and consequently no observed release of lead (Moon
and Dermatas, 2006). Though, the pH of leachate is sub-neutral
(7.5–7.8), the high background concentrations of metals, organic
or inorganic compounds and dissolved solids in leachate have been
reported to affect metal dissolution (Karnchanawong and
Limpiteeprakan, 2009). Cr released from a-Si PV after 7 months of
leaching was found to be 2.45 mg/L (2.13%). Cr concentration was
found to exceed the Indian Standard Waste (ISW) disposal thresh-
old (2 mg/L), and the WHO DW limit (0.05 mg/L) after the 60 days
of leaching period (Organisation, 2008). After 365 days, the Cr
release was found to be 14.96 mg/L (13.05%, broken-RW) and
11.04 mg/L (9.63%, broken-GW) whereas negligible Cr release
(<1%) was obtained in SW and ML conditions. It might be due to
high content of Ca (296 mg/gm of PV) which can form Ca-Cr
III
com-
pounds in alkaline system and thus, reduces the solubility of Cr
regardless of its amphoteric behaviour (Glasser, 1997). The maxi-
mum eluted Ag from a-Si was found to be higher in broken-RW
683.26 mg/L (26.9%) than minimum value of 84.69 mg/L (3.34%)
in unbroken-ML. The results obtained in this work are well consis-
tent with findings of previous leaching studies on a-PV. For exam-
ple, Cu in this study leached up to ~14%, which was found to be
similar to the values reported by Nover et al. (2017).
Being the second-generation PV, CIGS requires less semiconduc-
tor material and lately, various studies have investigated the
effects of different solutions on the their degradation
(Zimmermann et al., 2013; Rebecca Brun et al., 2016). For the CIGS
modules, among all metals, maximum amount of leaching was
observed for Cu metal (328.55 mg/L; 8.04%) at pH7 conditions
(Figure S2). Similar to the results of Zimmermann et al. (2013)
study, Se and Ga exhibited the least amount of leaching with con-
centrations ranged between 0.2 and 1 mg/L (0.30% "0.74%), while
the other elements exhibited high concentrations. Neutral pH
enhanced the release of soluble Ga and Se species. At pH 7, the
measured concentrations of Ga and Se at the end of one year were
observed to be 165.8 mg/L (0.28%) and 895.6 mg/L (0.74%), respec-
tively. These concentrations were found to be nearly 25-fold lower
than Pb content recorded in RW treatment for a-PV. In case of CIGS
PV, the observed small release of metals could be possible due to
low solid -to- liquid ratio as compared to other silicon PVs. Further,
in contrast to previous studies, the low release of Pb, Ag, and Se
was observed, which might be due to considerable differences in
experimental setups used (Gu et al., 2018; Nover et al., 2017).
Metals such as Al, Cu, Fe, Ga, Ni and Se leached 0.168, 10.58,
1.072, 0.0054, 0.062 and 0.028 mg/gm of CIGS, respectively. The
release of In was smaller in comparison to that of Cu from CIGS.
However, it gradually became steady in ML after 300 days of leach-
ing. This could be attributed to the effect of decrease in surface area
of exposed copper with time because of the formation of copper
chlorides (CuCl) on the PV surface. Other metals such as Al, Fe,
Ni, Pb, Cd and Cr leached 5.23 mg/L (15.89%), 33.22 mg/L (5.28%),
1.9 mg/L (0.46%), 0.008 mg/L (2.80%) 0.745 mg/L (5.60%) and
0.11 mg/L (0.15%) after 12 months (Table S4). The comparison of
metal dissolution data in ML conditions with those in the GW
and SW conditions (with similar neutral to alkaline pH) shows that
the metal leaching was significantly affected by the presence of
organic and additional inorganic compounds in the landfill lea-
chate. ANOVA analysis proves this hypothesis indicating that the
different solutions with different modules (broken vs unbroken)
affect the metal dissolution rate. Investigations suggest that
release quantities of Se, Ga and In are very minimal on disposal
of CIGS under various wetting conditions. As per the standard haz-
ardous waste characterisation tests by the USA and California Envi-
ronment Protection Agency (EPA), a solid waste can be considered
toxic if the concentration of Se dissolved in the TCLP or WET tests
exceeds 1000 mg/l. Given the results of this study, CIGS could be
classified as non-toxic in terms of Se as per the conditions used
in this study. Also, Ga leached in minimal amount with 166 mg/l
(0.28%) after 365 days and is not regulated because it is considered
as non-toxic generally (Ramos-Ruiz et al., 2018). Passivation of the
bound Ga in the PV by gallium oxide layer could possibly be the
reason for a low Ga dissolution rate. The same observations were
recorded for In from CIGS PV, though with slightly higher dissolu-
tion rates (at neutral pH conditions). For example, the In release
rate under alkaline and neutral pH conditions were found to be
two times greater than that under the acidic condition.
In mono-silicon modules, Si and Al have the major share in terms
of mass (wt. %) followed by Cu, Fe and Zn as per characterisation
reported in literature (Matsubara et al., 2018) and conducted in the
present study (Table S1). Under the acidic condition (RW), the maxi-
mum metal dissolution was observed for Ni followedby Pb, Al and Cu
with approx. 68.15 mg/L (29.8%), 29.01 mg/L (17.9%), 233.4 mg/L
(13.9%) and 210.28 mg/L (12.9%) of the initial metal content, respec-
tively. Other metals, such as Cd, Cr, As and Se showed a maximum
release of 0.007 mg/L (2.3%), 2.11 mg/L (3.00%), 0.003 mg/L (1%)
and 0.005 mg/L (0.36%) in broken-RW condition, respectively
(Table S4). In terms of mg per grams of mono-Si, Ni, Pb, Al and Cu
showed release of 0.648, 0.276, 2.22, 2.00, respectively, whereas Cd,
Cr, As and Se leached<0.2mg/gm of multi-Si. In case of MSW leachate,
leaching of various metals from mono-Si was observed up to 4.19%
(17.08 mg/L of Al forbroken-ML), which was found to be greater than
those from other PVs (Fig. 3). Pb release was observed to vary from
2.84 mg/L (1.75%, unbroken-GW) to 6.19 mg/L (3.82%, broken-SW)
in neutral to alkaline conditions. Pb was found to release up to
29.01 mg/L (17.92%, broken-RW) in acidic condition. It might be
due to the effect of high pH that might have resulted in the release
of oxyanions. With respect to module condition, significant effect
was observed in case of acidic leaching condition than in other solu-
tion types. Maximum K
d
was observed for acidic and leachate condi-
tions for all metals, though, no relative effect of module breakage
condition on K
d
was seen (Table S3). Raising the pH had significant
effect on the leaching extent as metal release was observed to
decrease significantly with increasing pH. The fact that encapsulant
release was found to be maximum among silicon-based PVs (dis-
cussed later), it might have resulted in high exposure of bound-
metals in semiconductor layer and anti-reflection coating of silicon
nitride to extraction solutions. The observed metal release from
mono-Si modules in this work was found to be smaller than that of
previous studies (Tammaro et al., 2016; Zimmermann et al., 2013).
This difference in extent of leachingmight be due to use of large sam-
ple size and absence of stressful conditions in literature-reported
studies. This shows that leaching studies which use small panel size
(in mms) in high acidic solutions, are by no means representative of
dumping of panels in real conditions.
For multi-Si modules, the metal release trend was alike to those
observed for mono-silicon PV. The metal dissolution for
broken-RW was up to 5.10% (5.10% for Ni, 4.97% for Al, 1.85% for
Cu and 1.67% for Pb) during the initial phase of the incubation till
150
th
day. The extent of metal dissolution was found to increase
exponentially for next 5 months and then it was found to be
become stable till the 365
th
day. The experimental data were best
fitted to the sigmoidal model (Figure S3). The highest concentra-
P. Nain, A. Kumar / Waste Management 114 (2020) 351–361 355
tions of soluble Pb, Cu, Al and Ni released were observed to be
31.27 mg/L (24.43%), 584.59 mg/L (16.95%), 417.29 mg/L (21.29%)
and 119.15 mg/L (25.96%), respectively, after one year of incuba-
tion. With respect to PV mass, Pb, Cu, Al and Ni released were
found to be 0.29, 5.42, 3.68 and 1.104 mg per grams of multi-Si,
respectively. It could be seen from Figure S3 that metal dissolution
for multi-Si is greater than that for CIGS PV and significantly higher
in RW solution. In contrast, metal leaching did not reach a plateau
in the solutions with alkaline conditions even after 365 days of
incubation, which may be attributed to the effect of high Cl content
or other compounds in leachate (Kim et al., 2011). Previous metal
leaching studies have reported that leaching kinetics are controlled
by mass transfer diffusion or a metal deposit on product surface
(Mishra et al., 2008). Since the dissolution of CuCl is low, it forms
a precipitated layer on the copper surface which inhibits the leach-
ing (Kim et al., 2011). Also, other compounds, such as bicarbonate,
mineral salts, volatile fatty acids (VFA) and other degradable
organic compounds can affect the interaction and metal dissolu-
tion (Kjeldsen et al., 2002). Landfill leachate contains different
types of organic and inorganic contaminants (Kjeldsen et al.,
2002; Naveen et al., 2017; Somani et al., 2018) that could possibly
act on PV materials, resulting in release of hazardous metals. More-
over, the metal leaching from PVs can be affected by presence of
other pollutants, such as metallic oxy-hydroxides, which can act
as adsorbents of inorganic species (Dixit and Hering, 2003); or cal-
cium and sulfide ions that could result in formation of insoluble
complexes (Zhang et al., 2017). Other metals, such as Cr, Se and
Zn showed a maximum release of 1.63 mg/L (1.96%), 0.007 mg/L
(1.4%) and 0.27 mg/L (1.79%), respectively. As, Cd and In were
not detected in the study period. The leached heavy metals from
multi-Si modules in ML solution appears to be relatively small to
raise any concerns related to EoL disposal at present. Significant
release of metal particulates from various PVs was observed under
the acidic rainwater conditions considered in the present study
(Figs. 2 and 3). However, it is challenging to approximate the actual
metal release that could possibly happen in actual landfill waste
under normal rainwater conditions. Among the heavy metals
released, release of Pb and Cd was concerning because of their car-
cinogenic nature. However, the high Pb release may be accredited
to the older generation soldering materials used in silicon-based
modules that could alleviate these concerns. It could be expected
that newer and more sophisticated soldering materials in the third
generation of solar PVs would considerably reduce their use in PV
and subsequenlty their release from PV. Further, recent studies
have successfully recovered metals from e-waste using lixiviants
(Jafari et al., 2019), fluid extraction (Ding et al., 2019), microorgan-
ismand by various other methods, which can be utilized for recov-
ering leached metals from EoL solar waste.
3.2. Probability of exceedance of leached metals to surface water limit
To put leached metal concentrations in perspective, PoE (%) was
estimated to surface water limit (Oklahoma’s Water Quality Stan-
dards; USEPA effluent discharge standards), irrespective of the sce-
nario considered. Among various PVs, maximum exceedance was
shown by multi-Si. In case of metals, Al (multi-Si & mono-Si), Ag
(a-PV) and In (CIGS) have shown maximum exceedance up to
92.31%. For a-PV, metals leached in RW and GW have higher prob-
abilities of exceeding the standards but least chance of exceedance
in ML (Fig. 4). After 1 month itself, leached Pb, Ag and Cu exceeded
their surface water limits with 92.31% of exceedance. Similar find-
ings were observed for Al released from mono- and multi-Si PVs.
Leached In and Cu from CIGS exceeded after the first and the sec-
ond month of leaching, respectively. Pb showed the minimum
exceedance in all PVs, with least in unbroken-ML leaching sce-
nario. In case of leaching solution, metals exceeded minimum in
SW and ML solutions from their respective surface water limits.
However, overall there was slight difference in PoE of metals for
broken and unbroken conditions. For rainwater solution, after
2nd or 3rd month of study period, metals exceeded their respective
surface water thresholds for both, broken and unbroken panel con-
ditions. These findings suggest that dumping of EoL solar PVs in
rainwater environment can violate the surface water standards.
3.3. Encapsulant release with time
Encapsulant plays a crucial role in the degradation of a PV (Brecl
et al., 2018). A number of studies have shown the encapsulant dis-
solution, mostly in crystalline silicon PVs using organic solvents
0
2
4
6
8
10
12
14
16
18
0
3
6
9
12
15
18
21
24
27
30
30 60 90 120 150 180 210 240 270 300 330 360 390
% leached of ini!al concentra!on
Pb release (mg/l)
Time (days)
Pb
B-RW B-GW B-SW B-ML UB-RW UB-GW UB-SW UB-ML
0
0.9
1.8
2.7
3.6
4.5
5.4
6.3
7.2
0
27
54
81
108
135
162
189
216
30 60 90 120 150 180 210 240 270 300 330 360 390
% leached of ini!al concentra!on
Cu release (mg/l)
Time (days)
Cu
B-RW B-GW B-SW B-ML UB-RW UB-GW UB-SW UB-ML
0
2
4
6
8
10
12
14
0
40
80
120
160
200
240
30 60 90 120 150 180 210 240 270 300 330 360 390
% leached of ini!al concentra!on
Al release (mg/l)
Time (days)
Al
B-RW B-GW B-SW B-ML UB-RW UB-GW UB-SW UB-ML
0
5
10
15
20
25
30
0
10
20
30
40
50
60
70
30 60 90 120 150 180 210 240 270 300 330 360 390
% leached of ini!al concentra!on
Ni release (mg/l)
Time (days)
Ni
B-RW B-GW B-SW B-ML UB-RW UB-GW UB-SW UB-ML
Fig. 3. Mono-silicon PV one year leaching kinetics of lead, copper, aluminium, and
nickel into RW, pH 4 (square); GW, pH7 (daimond); SW, pH10 (circle) and ML, MSW
leachate (triangle) Two modules conditions: Broken, B (unfilled symbols and dashed
lines) and Unbroken, UB (filled symbols and solid lines).
356 P. Nain, A. Kumar / Waste Management 114 (2020) 351–361
(Gangwar et al., 2019) and ultrasonic irritation (Xu et al., 2018).
Several experimental attempts have been done for delamination
using chemicals; recycling using etching process; EVA decomposi-
tion using thermal treatment; back cover removal using solvents,
most of which focused on aspects of recycling (Gu et al., 2018;
Yousef et al., 2019). The inner metal layer is well protected by
outer glass and encapsulant layer. But metal layer or coating can
be exposed to environment if glass or encapsulant damages during
transportation and assembly process (Su et al., 2019). A recent
study by Savvilotidou et al. (2017) showed the EVA dissolution
using various acids or acid mixtures along with different tempera-
tures. Another study by Su et al. (2019) showed the increase in
metal release from buried thin-film PV in soil with increase in con-
tact time and acid concentration (Su et al., 2019). However, the
effect of encapsulant on metal release and its quantification had
not been primarily focused on in literature. Thus, in order to under-
stand metal leaching mechanism, release of binding agent (in
terms of total carbon released) with time was quantified in this
study. Studies have shown that water penetration through the
laminated edges and back sheet of module can result in encapsu-
lant degradation and delamination (Gxasheka et al., 2005; Su
et al., 2019). Also, thin-film PV modules are more prone to water
penetration than silicon-based modules (Parida et al., 2011). Due
to delamination and encapsulant degradation, moisture can get
to the cells and interact with metals present in semiconductor
layer. Most common encapsulant or binding agent used in PVs
for adhering various layers is EVA (Chandel et al., 2015) with
molecular formula C
2
H
4
(H
2
C = CH
2
)(Osayemwenre & Meyer.,
2014). Therefore, the carbon released was hypothesised with EVA
degradation and total carbon (i.e., EVA release) released was quan-
tified monthly in various leaching scenarios. It is important to note
that carbon release is assumed to only represent EVA for approxi-
mation purposes. In reality, the released organic compound (which
is represented as total carbon) might be a combination of EVA and
different polymers. For different EVA-based encapsulated solar PV,
EVA can have different monomer composition, varied peroxide
concentrations and nature of UV used for stabilization (Adothu
et al., 2020). With advancement in PV materials research, new
polyolefin-based encapsulants and ionomers thermoplastics are
being used as an alternative to EVA and can behave differently in
various environment settings (Tracy et al, 2020). The investigation
of possible contributions of different polymers and EVA in the
observed release of carbon is beyond the scope of the present
investigation. This study assumed EVA to be the sole contributor
of the observed carbon as an approximation to illustrate the need
for linking encapsulant with metals release quantities. This study
acknowledges the need for performing detailed characterization
of released organic compounds. The schematic for metal release
after encapsulant degradation is shown in Fig. 5.
Among the PVs studied, the highest carbon release was
observed in multi and mono-Si in broken-RW with K
d
values equal
to 2.7 and 3.8, respectively. The fitting of sigmoidal model for total
carbon release as a function of time with 5 parameters gave the
best fit with R
2
>0.98 and adjusted R
2
>0.95 with respect to other
models (Fig. 6, Table S4). This may be due to two layers of encap-
sulant used for adhesion of glass and back sheet to semiconductor
film. Dissolution rate decreased with increase in pH of solution,
however, it was found to be minimum for ML despite of having
mildly alkaline conditions. Similar behaviour was observed for
thin-film PVs, amorphous and CIGS with minimal carbon release
in ML, and this might be due to the presence of bicarbonate and
mineral salts in landfill leachate (Kjeldsen et al., 2002). Also, the
EVA dissolution strongly depends on the solvent type.
Amorphous PV showed maximum K
d
of 5.02 (broken-RW) and
minimum of 2.47 (broken-SW) among all the PVs followed by CIGS
with maximum value of K
d
equals to 4.38 (unbroken-RW) and
minimum value of K
d
equals to 1.29 (unbroken-ML) (Table S3).
For leachate, total carbon released was minimal, with K
d
observed
between 1.1 and 3.9. The low encapsulant release resulted in lesser
layer delamination and reduced accessibility of leaching solution
to semiconductor layer. Thus, this could explain the observation
of low metal release in MSW leachate in various PVs. Further, the
2-way ANOVA analysis of carbon release with regards to module
condition (broken and unbroken) and leaching solution (RW, GW,
SW and ML) at
a
=0.05 (95% confidence test) shows the significant
effect of individual parameters, i.e., module condition and leaching
Fig. 4. Probability of Exceedance, PoE (%) of metals to standard surface water limit (mg/L, Ag: 0.5; Al: 0.75; Cu: 3; Cr: 0.5; Ni: 1; Pb: 1; Se: 0.5) in different leaching scenarios.
P. Nain, A. Kumar / Waste Management 114 (2020) 351–361 357
solution on carbon release. However, the effect of interaction of
these factors (i.e., factors: module condition*leaching solution) do
not significantly affect (p >0.05) the encapsulant release in all
PVs. The clear visible difference in carbon release with time can
be seen in Fig. 7, which shows the images after first and twelfth
months of leaching from CIGS panel. The released encapsulant
was jelly-like material adhering onto the edges of sample.
The main effect of EVA degradation is the corrosion of conduc-
tion lines and metallic components. This is well established in lit-
erature, where several studies have proved significant effect of EVA
degradation on metallic components of a module (Brecl et al.,
2018;Oreski et al., 2017). The similar inference was also derived
from the analysis of correlation between EVA and metal release
in present study. High values of spearman correlation coefficient
Fig. 5. Schematic of hypothesized metal leaching from semiconductor film.
Fig. 6. Carbon release in ML:pH7.6 (triangle); RW: pH4 (square); GW: pH7 (diamond); SW: pH10 (circle) for photovoltaics in Broken, B (unfilled symbols and dashed lines)
and Unbroken, UB (filled symbols and solid lines) conditions.
358 P. Nain, A. Kumar / Waste Management 114 (2020) 351–361
(rho) between binding agent release and metal leaching were
observed for all PVs (Figure S4). The EVA release observation also
implicitly indicates the variation of permeability of leaching solu-
tion within the EVA film. Studies have shown that permeability
of leaching solution in the EVA film, which depends upon diffusion
and solubility coefficient (Kim and Han, 2013). More work is
required to understand this aspect in detail. Further, the age of
EVA is another crucial factor which affects degradation. The aged
EVA degrade fastly with a degradation rate of 0.2% per day (Brecl
et al., 2018). Present study provided a preliminary investigation
on relationship between EVA and metal release. In this regard, fur-
ther work is required to understand the roles of various aspects
such as interaction of other reactive peroxides in modules with
metallic layer and contribution in EVA quantification; effect of
temperature and agitation on EVA stability or moisture ingress;
leaching solution with different composition.
3.4. Solar PV waste-associated LPI (SPW-LPI)
LPI has been used in several studies in the past (Lothe and
Sinha, 2017; Naveen et al., 2017). The LPI of leachate used in the
present study at the time of sampling was calculated to be 24.75.
After 12 months of leaching from dumped broken solar PVs in
MSW leachate, the LPI values were calculated to be 30.81, 30.50,
Table 1
Percentage change in leachate pollution potential after one year of solar photovoltaics disposal.
PV type Leaching solution T
0
Amorphous CIGS Mono-Si Multi-Si
Parameter* Wi Value Pi WiPi Value Pi WiPi Value Pi WiPi Value Pi WiPi Value Pi WiPi
MSW landfill leachate as leaching solution
BOD
5
0.06 2,189.3 ± 120.7 40 2.44 1563.58 35 2.14 1268.44 35 2.14 1369.16 35 2.14 1068.30 35 2.14
COD 0.06 47,562.8 ± 678.40 90 5.58 53686.4 87 5.39 58168.6 88 5.46 57364.6 87 5.39 50356.3 85 5.27
Chlorides 0.05 8,116.2 ± 831.5 100 4.9 8936.42 80 3.92 8236.30 80 3.92 6845.60 80 3.92 6934.30 80 3.92
NH
3
N 0.05 31.5 ± 1.37 10 0.51 12.36 5 0.26 6.14 5 0.26 9.54 5 0.26 10.66 5 0.26
pH 0.06 7.67 ± 1.40 5 0.28 7.80 5 0.28 7.79 5 0.28 7.64 5 0.28 7.89 5 0.28
TDS 0.05 18,232.5 ± 0.12 65 3.25 28246.3 6 3.00 26856.4 60 3.00 20634.3 55 2.75 1956.60 54 2.70
TCB 0.05 40,002 65 3.38 400.00 6 3.12 500.00 60 3.12 500.00 60 3.12 400.00 60 3.12
TKN 0.05 260 ± 3.50 7 0.37 190.30 5 0.27 164.60 5 0.27 165.36 5 0.27 139.43 5 0.27
As 0.06 0.04 ± 0.036 5 0.31 0.04 5 0.31 0.05 5 0.31 0.16 5 0.31 0.40 5 0.31
Cu 0.05 0.15 ± 0.032 5 0.25 65.40 100 5.00 98.25 100 5.00 29.31 100 5.00 209.60 100 5.00
Cr 0.06 0.185 ± 0.03 5 0.32 0.59 5 0.32 0.35 5 0.32 0.47 5 0.32 0.53 5 0.32
Fe 0.05 22.9 ± 0.031 5 0.23 25.43 5 0.23 14.36 5 0.23 19.61 5 0.23 26.65 5 0.23
Pb 0.06 0.004 ± 0.027 5 0.32 1.40 10 0.63 0.06 5 0.32 1.28 5 0.32 13.10 90 5.67
Ni 0.05 0.227 ± 0.03 5 0.26 0.26 5 0.26 0.35 5 0.26 5.27 30 1.56 52.54 100 5.20
Zn 0.06 1.281 ± 0.031 5 0.28 1.26 5 0.28 1.36 5 0.28 1.29 5 0.28 2.65 5 0.28
Total 0.82 22.66 25.38 25.13 26.12 34.9
LPI’ 24.75 30.81 30.50 31.70 42.4
% change in LPI 12.02 10.90 15.26 54.2
Rainwater as leaching solution
BOD
5
0.06 0 0 0 439 10 0.61 526 10 0.61 250.00 10 0.61 160.00 10 0.61
COD 0.06 0 0 0 653 5 0.31 1036 5 0.31 569.00 5 0.31 403.00 5 0.31
Chlorides 0.05 0 0 0 35 5 0.25 102 5 0.25 22.00 5 0.25 14.00 5 0.25
NH
3
N 0.05 0 0 0 0.65 5 0.26 1.38 5 0.26 2.30 5 0.26 3.00 5 0.26
pH 0.06 4.11 ± 0.2 60 3.3 4.50 60 3.30 4.23 60 3.30 4.13 60 3.30 4.32 60 3.30
TDS 0.05 0 0 0 1693 5 0.25 1602 5 0.25 1069.00 5 0.25 1260.00 5 0.25
TC 0.05 0 0 0 100 5 0.26 0.0 5 0.26 100.00 5 0.26 200.00 5 0.26
TKN 0.05 0 0 0 30 5 0.27 11.0 5 0.27 20.30 5 0.27 15.05 5 0.27
As 0.06 0 0 0 1.62 10 0.61 0.02 5 0.31 0.08 5 0.31 0.01 5 0.31
Cu 0.05 0 0 0 483.2 100 5.00 222.6 100 5.00 210.30 100 5.00 584.32 100 5.00
Cr 0.06 0 0 0 0.01 5 0.32 0.99 5 0.32 0.02 5 0.32 0.01 5 0.32
Fe 0.05 0 0 0 55.13 5 0.23 18.13 5 0.23 11.66 5 0.23 9.66 5 0.23
Pb 0.06 0 0 0 23.37 100 6.30 0.16 5 0.32 29.02 100 6.30 31.27 100 6.30
Ni 0.05 0 0 0 0.15 5 0.26 2.41 5 0.26 68.10 100 5.20 119.24 100 5.20
Zn 0.06 0 0 0 1.54 5 0.28 1.29 5 0.28 2.70 10 0.56 2.31 5 0.28
Total 0.82 3.3 18.49 12.20 23.41 23.13
LPI’ 4.001 22.44 14.81 28.40 28.06
% change in LPI 459.6 269.2 608.3 599.9
Note: LPI- Leachate pollution index, TDS- Total dissolved solids, BOD- Biological oxygen demand, COD- Chemical oxygen demand, NH
3
N Ammonical nitrogen, TC- Total
coliform, TKN- Total Kjeldhal Nitrogen,. *All leachate values are in mg/L except TC.
Fig. 7. Encapsulant release from CIGS panel in leaching solution at T
0
months (left side) and T
12
months (right side).
P. Nain, A. Kumar / Waste Management 114 (2020) 351–361 359
31.70, 42.40 for amorphous, CIGS, mono-Si and multi-Si PVs,
respectively (Table 1). The increase in leachate contamination
potential of the leachate after one-year study was observed to be
12.02%, 10.90%, 15.26% and 54.19%, respectively for the above
PVs. It is evident that dumping of EoL solar panels can double
the contamination potential of a site. The increase in LPI of landfill
on disposal of solar PVs has also been proved in an earlier work by
Nain & Kumar (2019). Whereas in case of acidic rainwater condi-
tions, the contamination potential of the solution changes up to
600% if dumped with solar PVs. The increase in LPI is mainly due
to the leached heavy metals such as Pb or Cr from PVs as changes
in values of other parameters are negligible. Such significant
increase in the LPI indicates a cause of concern and the importance
of understanding the impact of EoL solar PVs on environment if
dumped without any treatment for preventing leaching of metals
from EVs. The results from present study shows that the more real-
istic studies in environmently- relevant conditions are required for
estimating pollution potential of solar PV waste. There have been
several studies on fate of e-waste in landfills, which suggest that
disposal of e-waste in environment could cause serious environ-
ment contamination and health concern. However, limited infor-
mation is available on fate of EoL solar PVs in environment.
Further, present work utilises only one module of one brand for
each technology, which is a limitation. Though, the use of modules
from four different technologies gives an overview on preliminary
fate of metals from different modules under various scenarios.
Thus, more research is required investigating their fate in environ-
ment and efforts are required for guiding regulatory bodies in pol-
icy formulation and implementation of safe management practices.
3.5. Conclusions
This study analysed metal release from broken and unbroken
solar panels in four different leaching solutions (pH4, pH7, pH10,
and real MSW leachate) for one year. With exceptions to findings
for few metals in rainwater solutions, the result of leaching from
commercially available PVs was minimal. Thus, the high acidic
conditions and short-term reactions in TCLP and WET tests could
poorly represent the metal release from encapsulated materials
like solar modules. In this regard, the results from present study
demonstrate the fate of solar panels in landfill leachate. For initial
months after disposal, the PV cells and films maintain their integ-
rity because of encapsulation. Given results, disposal of EoL solar
PVs in MSW landfills could result in less leaching from semicon-
ductor layers. However, these findings do not imply that PV dis-
posal in landfills are environmentally safe. Though, the observed
less metal release in MSW leachate suggests the need of further
investigation on different aspects, such as, effect of microbial activ-
ity, temperature variations and composition of MSW waste on
metal solubilization from PVs.
Metal leaching kinetics from this study show that the slow
leaching behavior is more relevant in PVs since the maximum
metal release observed was found to be 68.15 mg/L (29.8%) for
Ni (mono-Si) followed by 683.26 mg/L (26.9%) for Ag (a-Si),
31.27 mg/L (24.43%) for Pb (multi-Si) and 417.29 mg/L (21.29%)
for Al (multi-Si) under most stressful condition. This clearly shows
that the short-term leaching studies with millimeter-sized sample
pieces do not represent the realistic dumping scenarios with big
sample size in absence of agitation. Therefore, experimental condi-
tions in future studies are required to represent more
environmentally- relevant scenarios, such as dumping of EoL solar
panels in anaerobic conditions of MSW landfills. The findings also
suggest that encapsulant degradation plays a critical role and have
positive correlation with metal dissolution. With respect to previ-
ous studies, the low metal release was observed, which might be
due to considerable differences in sample size, leaching solutions
and experimental setups. The regression analysis indicated that
conditions of the modules and pH of the leaching solution also play
significant roles on metal dissolution. Similarly, the probability of
exceedance of leached metals to standard limits depends upon
the mass of PVs exposed per volume of water. Al (multi-Si and
mono-Si), Ag (a-Si) and In (CIGS) have shown the exceedance of
92.31%. In terms of pollution potential, the increase in leachate
contamination after oneyear of PV dumping was found to be
12.02%, 10.90%, 15.26%, 54.19% for amorphous, CIGS, mono-Si
and multi-Si PVs, respectively.
Overall, the outcomes from the present oneyear study provide
an insight on the fate of four PV types in various environmental
scenarios and could be beneficial in decision making-process
towards EoL solar PVs management. Lead, cadmium, and other
potentially leachable heavy metal releases should be monitored
to ensure that the EoL solar PV disposal do not violate the regula-
tory limits. Since the solar PV industry is young, positive steps such
as recycling, regulations for EoL management, use of panels with
less hazardous metals and take-back policies could be enforced
to maintain the green credibility and sustainability of the product
and moreover, to avoid an additional e-waste issue.
Declaration of Competing Interest
The authors declare that they have no known competing finan-
cial interests or personal relationships that could have appeared
to influence the work reported in this paper.
Acknowledgements
The authors would also like to thank Indian Institute of Tech-
nology (IIT Delhi), India and University Grant Commission (UGC
award Sr. No. 24004768 Ref. No: 3509/NET-JULY2016) for provid-
ing financial support for this study.
Appendix A. Supplementary data
Supplementary data to this article can be found online at
https://doi.org/10.1016/j.wasman.2020.07.004.
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